United States
Environmental Protection Agency
EPA Document** 740-R1-4002
June 2014
Office of Chemical Safety and
Pollution Prevention
TSCA Work Plan Chemical Risk Assessment
Trichloroethylene:
Degreasing, Spot Cleaning and Arts & Crafts Uses
CASRN: 79-01-6
Cl
Cl
H
Cl
June 2014
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TABLE OF CONTENTS
TABLE OF CONTENTS 2
LIST OF TABLES 5
LIST OF FIGURES 8
AUTHORS/CONTRIBUTORS/ACKNOWLEDGEMENTS/REVIEWERS 9
GLOSSARY OF TERMS AND ABBREVIATIONS 12
EXECUTIVE SUMMARY 18
1 PURPOSE, BACKGROUND, AND SCOPE 25
1.1 PURPOSE AND AUDIENCE 25
1.2 BACKGROUND 25
1.2.1 Rationale for Selecting TCE for Risk Assessment 25
1.2.2 Overview of TCE Uses, Production Volume and EPA's Regulatory History 25
1.3 SCOPE OF THE ASSESSMENT 26
1.3.1 Selection of TCE Uses 26
1.3.2 Selection of Exposure Pathway 28
1.3.3 Identification of Human Populations Exposed During TCE Uses 29
1.3.4 Risk Evaluation of TCE's Human Health Hazards 29
1.3.5 Why Environmental Risks Were Not Evaluated For Selected TCE Uses 30
2 HUMAN HEALTH RISK ASSESSMENT 31
2.1 PHYSICAL-CHEMICAL PROPERTIES 31
2.2 OVERVIEW OF ENVIRONMENTAL FATE AND RELEASES OF TCE 32
2.2.1 Ambient Air Concentrations of TCE 32
2.3 ENVIRONMENTAL RELEASES AND OCCUPATIONAL EXPOSURE ESTIMATES FOR SMALL COMMERCIAL
DECREASING OPERATIONS 33
2.3.1 What is a Small Operation? 33
2.3.2 A Brief Summary of Solvent Cleaning 33
2.3.3 EPA/OPPT's Release and Exposure Assessment for Degreasing Operations 34
2.4 ENVIRONMENTAL RELEASES AND OCCUPATIONAL EXPOSURE ESTIMATES FOR SPOT CLEANING AT DRY
CLEANING OPERATIONS 45
2.4.1 A Brief Summary of Spot Cleaning at Dry Cleaning Operations 45
2.4.2 EPA/OPPT's Release and Exposure Assessment for Spot Cleaning Operations 45
2.5 CONSUMER EXPOSURES - DEGREASER AND ARTS/CRAFTS USES OF TCE 52
2.5.1 TCE Uses Targeted for Consumer Exposure Assessment 52
2.5.2 Overview of the E-FAST2/CEM Model 53
2.5.3 Consumer Model Scenarios and Input Parameters for Indoor Exposure to Specific TCE Uses 55
2.5.4 Consumer Model Results 63
2.5.4.1 Sensitivity of Model Parameters 64
2.5.4.2 Indoor Air Monitoring of TCE 64
2.6 HAZARD/DOSE-RESPONSE ASSESSMENT 65
2.6.1 Approach and Methodology 65
2.6.1.1 Selection of TCE IRIS Assessment as the Source Document for the TCE TSCA Assessment 65
2.6.1.2 Aspects of the TCE IRIS Assessment that Were Adopted in the OPPT Risk Assessment 66
2.6.1.2.1 Carcinogenic Hazard and Dose-Response Assessment 66
2.6.1.2.2 Non-Cancer Hazard and Dose-Response Assessment 67
2.6.1.3 Absorption, Distribution, Metabolism, and Excretion 72
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2.6.1.4 PBPK Modeling Approach Supporting the TCE IRIS Assessment 76
2.6.2 Human Health Hazard Summary 81
2.6.2.1 Genetic Toxicity of TCE and its metabolites 81
2.6.2.2 Human Toxicity Following Acute Exposure to TCE 82
2.6.2.3 Toxicity Following Repeated Exposures to TCE 82
2.6.2.3.1 Liver Toxicity (Including Cancer) 83
2.6.2.3.2 Kidney Toxicity (Including Cancer) 85
2.6.2.3.3 Neurotoxicity 87
2.6.2.3.4 Immunotoxicity (Including Cancer) 89
2.6.2.3.5 Reproductive Toxicity 91
2.6.2.3.6 Developmental Toxicity 93
2.6.2.4 Summary of Hazard Studies Used to Evaluate Acute and Chronic Exposures 99
2.7 HUMAN HEALTH RISK CHARACTERIZATION 99
2.7.1 Risk Estimation Approach for Acute and Repeated Exposures 101
2.7.2 Acute Non-Cancer Risk Estimates for Inhalation Exposures to TCE 104
2.7.3 Chronic Non-Cancer and Cancer Risk Estimates for Inhalation Exposures to TCE 107
2.7.3.1 Cancer Risks for Occupational Scenarios 107
2.7.3.2 Chronic Non-Cancer Risks for Occupational Scenarios 109
2.7.4 Human Health Risk Characterization Summary 110
2.8 DISCUSSION OF KEY SOURCES OF UNCERTAINTY AND DATA LIMITATIONS 114
2.8.1 Uncertainties in the Occupational and Consumer Exposure Assessments 114
2.8.1.1 Small Commercial Degreasing Operations 114
2.8.1.2 Spot Cleaning at Dry Cleaning Facilities 116
2.8.1.3 Degreaser and Arts/Crafts Uses in Residential Settings 117
2.8.2 Uncertainties in the Hazard and Dose-Response Assessments 118
2.8.2.1 Uncertainties in the Cancer Hazard/Dose-Response Assessments 118
2.8.2.2 Uncertainties in the Non-Cancer Hazard/Dose-Response Assessments 119
2.8.3 Uncertainties in the Risk Assessment 123
2.9 CONCLUSIONS OF THE HUMAN HEALTH RISK ASSESSMENT 125
3 REFERENCES 127
APPENDICES 151
Appendix A PRODUCTION VOLUME AND INVENTORY UPDATE RULE DATA 151
Appendix B REGULATORY HISTORY OF TCE AT THE USEPA AND RELATED ACTIONS 153
B-l REGULATORY HISTORY OF TCE AT THE U.S. EPA 153
B-2 OTHER REGULATORY ACTIONS IN THE U.S. AND ABROAD 154
Appendix C ENVIRONMENTAL FATE OF TCE 157
C-l ENVIRONMENTAL FATE 157
C-l-1 Fate in Air 157
C-l-2 Fate in Water 157
C-l-3 Fate in Soil, Sediment, and Groundwater 158
C-l-4 Bioconcentration 159
C-l-5 Conclusions on Environmental Fate 159
Appendix D NAICS CODES FOR TCE DEGREASING 161
Appendix E ESTIMATION OF TCE EMISSION RATE AT SMALL DEGREASING FACILITIES 162
Appendix F ESTIMATION OF TCE EXPOSURES AT SMALL DEGREASING FACILITIES 163
Appendix G CALCULATION OF TCE EXPOSURES AT SMALL DEGREASING FACILITIES 164
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Appendix H CALCULATION OF TCE EXPOSURES FROM SPOT CLEANING AT DRY CLEANING
FACILITIES 169
Appendix I EFAST2/CEM INDOOR MODELING, CONSUMER BEHAVIOR PARAMETERS AND MODEL
COMPARISONS 171
1-1 DEFAULT PARAMETERS USED IN CEM FOR EMISSION AND HOUSEHOLD CHARACTERISTICS 171
1-1-1 Air Exchange Rate 171
1-1-2 Overspray fraction 173
1-1-3 Emission Rate 173
1-1-4 House Volume and movement within the home 175
1-1-5 Inhalation rate and body weight 176
1-2 CONSUMER BEHAVIOR PATTERNS 176
1-2-1 Westat data for solvent type cleaning fluids 177
1-2-2 Westat data for brake cleaners/quieters 178
1-2-3 Summary of Westat data 178
1-3 COMPARISON OF EFAST CEM WITH MCCEM 178
Appendix J CONVERTING E-FAST ADRs TO AIR CONCENTRATIONS 180
Appendix K RISK ASSESSMENT GUIDELINES, LITERATURE SEARCH STRATEGY, STUDY SELECTION
AND DATA QUALITY CRITERIA FOR THE TCE IRIS TOXICOLOGICAL REVIEW AND OPPT
STUDY REVIEW 183
K-l RISK ASSESSMENT GUIDELINES 183
K-2 LITERATURE SEARCH STRATEGY 184
K-3 STUDY SELECTION AND DATA QUALITY CRITERIA WHICH SERVED As THE BASIS FOR THIS ASSESSMENT 184
Appendix L LIST OF ORAL AND INHALATION STUDIES SUITABLE FOR NON-CANCER DOSE-
RESPONSE ANALYSIS IN THE TCE IRIS ASSESSMENT 187
L-l LIVER EFFECTS 187
L-2 KIDNEY EFFECTS 188
L-3 NEUROTOXICITY 189
L-4 IMMUNOTOXICITY 190
L-5 REPRODUCTIVE TOXICITY 191
L-6 DEVELOPMENTAL TOXICITY 193
Appendix M BENCHMARK DOSE ANALYSIS OF BLOSSOM ET AL. 2013 194
Appendix N Weight-of-Evidence Analysis for Fetal Cardiac Malformations Following TCE Exposure
197
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LIST OF TABLES
Table 1-1. Primary Uses of TCE and Determination of Inclusion in this Risk Assessment 27
Table 2-1. Physical-Chemical Properties of TCE 31
Table 2-2. TCE Ambient Air Monitoring Data (u.g/m3) 33
Table 2-3. Number of TCE Emission Sources and Corresponding Total Annual Air Emissions of
TCE as reported in the 2008 NEI (EPA, 2008) 36
Table 2-4. Total Annual Air Emissions of TCE as Reported in the 2008 TRI 36
Table 2-5. Total Annual Air Emissions of TCE as Reported in the 2011 TRI 36
Table 2-6. EPA/OPPT's Estimated Total Annual Air Emissions of TCE from Small Commercial
Degreasing Facilities 37
Table 2-7. Breakdown of Degreasing Machine Type based on NEI Data for Point Sources 38
Table 2-8. TCE Emissions Potentially Escaping into the Workplace at Small Commercial
Degreasing Facilities (Calculated and Reported Values) 39
Table 2-9. Summary of Potential Workplace TCE Inhalation Exposures at Small Commercial
Degreasing Facilities based on the NF/FF Model 41
Table 2-10. Summary of Acute and Chronic Workplace TCE Inhalation Exposures at Small
Commercial Degreasing Facilities 44
Table 2-11. TCE Emissions Potentially Escaping into the Workplace from Spot Cleaning 47
Table 2-12. Summary of Potential Workplace TCE Inhalation Exposures from Spot Cleaning at
Dry Cleaning Facilities 49
Table 2-13. Summary of Acute and Chronic Workplace TCE Inhalation Exposures from Spot
Cleaning at Dry Cleaning Facilities 51
Table 2-14. TCE Products in NIH's Household Products Database 53
Table 2-15. Consumer Model Scenarios and Populations of Interest 55
Table 2-16. Summary of E-FAST2/CEM's Input Parameters and Assumptions for Estimation of
Potential Acute Dose Rates for Specific TCE Uses 58
Table 2-17. Estimated TCE Air Concentrations (Time Averaged Over a Day) from the Residential
Indoor Use of Solvent Degreasers or Clear Protecting Coating Sprays 63
Table 2-18. Lowest PBPK-derived HECs for different effects domains based on analysis in TCE
IRIS assessment 69
Table 2-19. Steps of the Non-Cancer TCE IRIS Process that Were Considered in OPPT's Non-
Cancer Hazard/Dose-Response Approach 71
Table 2-20. TCE Metabolites Identified by Pathway 73
Table 2-21. Common Metabolites of TCE and Related Compounds 74
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Table 2-22. List of All of the PBPK-Modeled Dose Metrics Used in the TCE IRIS Assessment 77
Table 2-23. Dose-Metrics for Cancer and Non-Cancer Endpoints Used in the OPPT Assessment 77
Table 2-24. Comparison of TCE Human Equivalent Concentrations (HECs) Under Constant,
Continuous Exposure for Different Exposure Durations 80
Table 2-25. Comparison of TCE Human Equivalent Concentrations (HECs) Under Intermittent
(Occupational) Exposure for Different Exposure Durations 81
Table 2-26. Summary of Hazard Information Used in the Risk Evaluation of Acute and Chronic
Scenarios 100
Table 2-27. Use Scenarios, Populations of Interest and Toxicological Endpoints for Assessing
Acute Risks to TCE-containing Degreasers, Spotting Agents and Arts/Crafts Products 101
Table 2-28. Use Scenarios, Populations of Interest and Toxicological Endpoints for Assessing
Chronic Risks to TCE-containing Degreasers and Spotting Agents 102
Table 2-29. Equation to Calculate Non-Cancer Acute or Chronic Risks Using Margin of Exposures
103
Table 2-30. Equation to Calculate Cancer Risks 103
Table 2-31. Acute Non-Cancer Risk Estimates for Commercial Use of Degreaser Product at Small
Shops (Developmental Effects: Congenital Heart Malformations, Johnson et al., 2003) 105
Table 2-32. Acute Non-Cancer Risk Estimates for Commercial Use of Spotting Agent at Dry
Cleaning Facilities (Developmental Effects: Congenital Heart Malformations, Johnson et al.,
2003) 105
Table 2-33. Acute Non-Cancer Risk Estimates for Residential Uses of TCE-containing degreasers
and art/crafts products (Developmental Effects: Congenital Heart Malformations, Johnson et al.,
2003) 106
Table 2-34. Chronic Non-Cancer Risk Estimates for Commercial Use of Degreaser Product at
Small Shops 112
Table 2-35. Chronic Non-Cancer Risk Estimates for Commercial Use of Spotting Agent at Dry
Cleaning Facilities 113
Table A-l. National Chemical Information for TCE from 2012 CDR 151
Table A-2. Summary of Industrial TCE Uses from 2012 CDR 152
Table A-3. TCE Commercial/Consumer Use Category Summary 152
Table C-l. Environmental Fate Characteristics of TCE 160
Table G-l. Model Inputs for Small / Industrial Commercial Degreasing Facilities 168
Table G-2. Potential Workplace TCE Inhalation Exposures and Number of Workers Exposed 168
Table H-l. Model Inputs for Dry Cleaning Facilities 169
Table H-2. Potential Workplace TCE Inhalation Exposures and Number of Workers Exposed...170
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Table 1-1. Summary of CEM Parameters for Estimation of TCE Indoor Air Concentrations 172
Table 1-2. Comparison of Westat Survey Data and EPA'sTCE Degreaser Simulation Values 177
Table J-l. Estimated TCE Potential Acute Dose Rates from the Residential Indoor Use of Solvent
Degreasers or Clear Protecting Coating Sprays 180
Table J-2. Estimated TCE Inhalation Calculated Concentration in Air (Over Course of Day) from
Use of Two Consumer Products Indoors at Residences 182
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LIST OF FIGURES
Figure 2-1. Overview of EPA/OPPT's occupational release and exposure assessment for
degreasing operations 34
Figure 2-2. Comparison of EPA/OPPT's Workplace TCE Exposure Estimates with 2003-2010
OSHA Monitoring Data for Small Commercial Degreasing Facilities 43
Figure 2-3. Overview of EPA/OPPT's occupational release and exposure assessment for spot
cleaning operations 46
Figure 2-4. Dose-Response Analyses of Rodent Non-Cancer Effects Using the Rodent and
Human PBPK Models 78
Figure 2-5. Example of HEC99 Estimation through Interpecies, Intraspecies and Route-to-Route
Extrapolation from a Rodent Study LOAEL/NOAEL 79
Figure 2-6. Cancer Risk Estimates for Commercial Use of Degreaser Product at Small Shops...108
Figure 2-7. Cancer Risk Estimates for Commercial Use of Spotting Agent at Dry Cleaning
Facilities 108
Figure G-l. Illustration of an Imperfectly Mixed Room: Near-Field/Far-Field Approximation of a
Solvent Cleaning Facility 164
Figure 1-1. Screen capture of Summary of Recommended Values for Residential Building
Parameters from the Exposure Factors Handbook (EPA, 2011a) 173
Figure 1-2. Screen capture of E-FAST equations for estimation of emission rate 174
Figure 1-3. Screen Capture of MCCEM Model Evaluation Efforts 179
Figure K-l. Study Quality Considerations for Epidemiological Studies 185
Figure K-2. Study Quality Considerations for Animal Studies 186
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AUTHORS/CONTRIBUTORS/ACKNOWLEDGEMENTS/REVIEWERS
This report was developed by the United States Environmental Protection Agency (U.S. EPA),
Office of Chemical Safety and Pollution Prevention (OCSPP), Office of Pollution Prevention and
Toxics (OPPT). The work plan risk assessment for trichloroethylene (TCE) was prepared based
on existing data and any additional information received during the public comment period and
peer review process. Mention of trade names does not constitute endorsement by the US EPA.
EPA Assessment Team
Lead: Iris A. Camacho-Ramos, OPPT/Risk Assessment Division (RAD)
Team Members:
Marcus Aguilar, OPPT/Chemical Control Division (CCD)
Stanley Barone Jr., OPPT/RAD
Rehan Choudhary, OPPT/RAD
Cathy Fehrenbacher, OPPT/RAD
Kathy Hart, OPPT/ Economics, Exposure and Technology Division (EETD)
David Lynch, OPPT/RAD
Nicholas Nairn-Birch/CCD
Laura Nielsen, OPPT/EETD
Nhan Nguyen, OPPT/RAD
Justin Roberts, OPPT/ EETD
Yvette M. Selby - Mohamadu, OPPT/RAD
David Tobias, OPPT/RAD
Acknowledgements
Development of Draft TCE Risk Assessment
Lynn Delpire (OPPT/RAD; retired)
Louis Scarano (OPPT/RAD)
Benchmark Dose Analysis of Blossom et al. study (Appendix M)
John Fox [U.S. EPA's Office of Research and Development (ORD), National Center for
Environmental Assessment (NCEA)]
TCE PBPK Modeling
Weihsueh Chiu (ORD/NCEA)
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Weight-of-Evidence Analysis for Fetal Cardiac Malformation (Appendix N)
Barbara Abbott (U.S. EPA/ORD/National Health and Environmental Effects Research Laboratory, NHEERL)
Xabier Arzuaga (ORD/NCEA)
Weihsueh Chiu (ORD/NCEA)
Susan Euling (ORD/NCEA)
John Fox (ORD/NCEA)
Karen Hogan (ORD/NCEA)
Andrew Hotchkiss (ORD/NCEA)
Sid Hunter (ORD/NHEERL)
Jennifer Jinot (ORD/NCEA)
Thomas Knudsen (U.S. EPA/ORD/National Center for Computational Toxicology, NCCT)
Susan Makris (ORD/NCEA)
Michael Narotsky (ORD/NHEERL)
Christina Powers (ORD/NCEA)
Cheryl Siegel Scott (ORD/NCEA)
Jamie Strong (ORD/NCEA)
Also, portions of this document were prepared for the U.S. EPA by the Syracuse Research
Corporation and Versar.
EPA Internal Peer Reviewers
Office of Children's Health Protection
Greg Miller
Office of Research and Development/
National Center for Environmental
Assessment
David Bussard
Weihsueh Chiu
John Fox
Jennifer Jinot
Susan Makris
Office of Research and Development/
National Exposure Research Laboratory
Peter Egehhy
Office of Policy
Sharon Cooperstein
Office of Solid Waste and Emergency
Response
Kathleen Raffaele
Stiven Foster
David Cooper
Rich Kapuscinski
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External Peer Review
EPA/OPPT contracted with The Scientific Consulting Group, Inc. (SCG) to convene a panel of ad
hoc reviewers to conduct an independent external peer review for the EPA's draft work plan
risk assessment for TCE. As an influential scientific product, the draft risk assessment was peer
reviewed in accordance with US EPA's peer review guidance. The peer review panel performed
its functions by web conference and teleconference on July 9, July 17 and August 21, 2013. The
panel consisted of the following individuals:
Penelope A. Fenner-Crisp (Chair), Ph.D. Ronald L Melnick, Ph.D.
Private consultant Ron Melnick Consulting, LLC.
Jeffrey Driver, Ph.D. Kenneth M. Portier, Ph.D.
Risksciences.net, LLC American Cancer Society
Montserrat Fuentes, Ph.D. Barry Ryan, Ph.D.
North Carolina State University Emory University
Kathleen M. Gilbert, PhD. Calvin Willhite, Ph.D.
University of Arkansas for Medical Sciences Risk Sciences International
MichaelJayjock, Ph.D.
LifeLine Group
Please visit the EPA/OPPT's Work Plan Chemicals web page for additional information on the
TCE's peer review process, including the peer review report: http://www.scgcorp.com/tcl2013/
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GLOSSARY OF TERMS AND ABBREVIATIONS
l-ig/m3
AC
ADC
ADR
ADRpot
AEGL
AER
AT
Atm
ATSDR
BAF
BCF
BMD
BMDL
BLS
BOD
BW
C
Cair
°C
CFF
CFFJWA
CNF
CNFJWA
Cppot
CASRN
CBI
CCD
CCRIS
CDR
CEM
CEPA
CERCLA
CFC-12
CH
Cl
cm
cm3
CNS
C02
Vacuum permittivity
Microgram(s) per cubic meter
Acute concentration
Average daily concentration
Acute dose rate
Potential acute dose rate
Acute exposure guideline level
Air exchange rate
Averaging time
Atmosphere
Agency for Toxic Substances and Disease Registry
Bioaccumulation factor
Bioconcentration factor
Benchmark dose
Benchmark dose, lower confidence limit(s)
Bureau of Labor Statistics
Biochemical oxygen demand
Body weight
Contaminant concentration
Air concentration
Degree Celsius
Average far field concentration
Time weighted average far field concentration
Average near field concentration
Time weighted average near field concentration
Modeled peak concentration
Chemical abstracts service registry number
Confidential business information
Chemical Control Division
Chemical Carcinogenesis Research Information System
Chemical data report
Consumer exposure module
Canadian Environmental Protection Act
Comprehensive Environmental Response, Compensation, and Liability Act
Dichlorodifluoromethane, also called Freon-12
Chloral hydrate
Confidence interval
Centimeter(s)
Cubic meter(s)
Central nervous system
Carbon dioxide
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cRfC Candidate reference concentration
CYP Cytochrome P450
DCA Dichloroacetic acid
DCAC Dichloroacetyl chloride
DCVC S-Dichlorovinyl-L-cysteine (collectively, the 1,2- and 2,2- isomers)
DCVG S-Dichlorovinyl-glutathione (collectively, the 1,2- and 2,2- isomers)
DCVT Dichlorovinyl thiol
DEv Duration of an event
DIY Do-it-yourself
DNA Deoxyribonucleic acid
DART/ETIC Developmental and Reproductive Toxicology/Environmental Teratology
Information Center
DOS Disk operating system
ECA Enforceable consent agreement
ED Exposure duration
EETD Economics, Exposure and Technology Division
EF Exposure frequency
E-FAST2 Exposure and Fate Assessment Screening Tool version 2
EFH Exposure factors handbook
EMIC Environmental Mutagens Information Center
EMICBACK Environmental Mutagen Information Center Backfile databases
EPA Environmental Protection Agency
ESRD End-stage renal disease
EU European Union
EvapTime Evaporation time
FF Far field
FM03 Flavin-containing monooxygenase
FQ Frequency of product use
FSA Free surface area
ft Foot/feet
ft2 Square foot/feet
ft3 Cubic foot/feet
g Gram(s)
g/cm3 Grams per cubic centimeters
g/L Grams per liter
G Average generation rate
GD Gestational day
GENE-TOX Genetic Toxicology Data Bank
GGT Gamma glutamyl transpeptidase
GSH Glutathione (reduced)
HNF Near field height
HAPs Hazardous air pollutants
HCV Human cancervalue
HEC Human equivalent concentration
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HEC5o Human equivalent concentration at the 50th percentile
HEC95 Human equivalent concentration at the 95th percentile
HECgg Human equivalent concentration at the 99th percentile
hr(s) Hour(s)
HSDB Hazardous Substances Data Bank
HSIA Halogenated Solvents Industry Alliance, Inc.
HVICL High Volume Industrial Chemicals List
IA Indoor air
IARC International Agency for Research on Cancer
id POD Internal dose point of departure
InhR Inhalation rate
IgA Immunoglobulin A
IL-2 lnterleukin-2
IEMB Indoor Environmental Management Branch
IRIS Integrated Risk Information System
IUR Inhalation unit risk
k Emission rate
Kow Octanohwater partition coefficient
kg Kilogram(s)
Koc Soil organic carbon partition coefficient
L Liter(s)
lb(s) Pound(s)
LNF Near field length
LADC Lifetime average daily concentration
LADD Lifetime average daily dose
LEV Local exhaust ventilation
LT Lifetime
LOAEL Lowest-observed-adverse-effect level
m Meter(s)
m2 Square meter(s)
m3 Cubic meter(s)
m3/hr Cubic meter(s) per hour
MCL Maximum contaminant level
MCLG Maximum contaminant level goal
mg Milligram(s)
mg/kg-bw/day Milligram(s) per kilogram body weight per day
mg/L Milligram(s) per liter
mg/m3 Milligram(s) per cubic meter
mg/mL Milligram(s) per milliliter
min Minute(s)
MITI Ministry of International Trade and Industry
Mlbs Million of pounds
mm Hg Millimeters of mercury
MOE Margin of exposure
Page 14 of 212
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MOEacute Margin of exposure for acute exposures
MOEchronic Margin of exposure for chronic exposures
MOD Memorandum of understanding
MSDS(s) Material safety data sheet(s)
MW Molecular weight
NAcDCVC N-Acetyl-S-(l,2-dichlorovinyl)-L-cysteine or N-acetyl-S (2,2 dichlorovinyl)-L-
cysteine
NAICS North American Industry Classification System
NAPL Nonaqueous phase liquid
NAS National Academies
NCEA National Center for Environmental Assessment
NCI National Cancer Institute
NEI National Emissions Inventory
NESHAP National Emissions Standards for Hazardous Air Pollutants
NF Near field
NF/FF Near field/far field
NHANES National Health and Nutrition Examination Survey
NHL Non-Hodgkins lymphoma
NICNAS National Industrial Chemicals Notification and Assessment Scheme
NIH National Institutes of Health
NIOSH National Institute for Occupational Safety and Health
NIST National Institute of Standards and Technology
nm Nanometer(s)
NOAEL No-observed-adverse-effect level
NOES National Occupational Exposure Survey
NOHSC National Occupational Health and Safety Commission
NPI National Pollutant Inventory
NPL National Priority List
NPS Nonpoint source
NTP National Toxicology Program
OAR Office of Air and Radiation
OCSPP Office of Chemical Safety and Pollution Prevention
OECD Organisation for Economic Co-operation and Development
OPPT Office of Pollution Prevention and Toxics
OR Odds ratio
OSHA Occupational Safety and Health Administration
OSWER Office of Solid Waste and Emergency Response
OW Office of Water
oz Ounce(s)
PA Personal air
PBPK Physiologically-based pharmacokinetic
p-cRFCs PBPK model-based candidate RfCs
PEC Priority Existing Chemical
PFC Plaque-forming cell
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PND Postnatal day
POD Point of departure
POTW Publicly owned treatment works
ppb Parts per billion
ppm Parts per million
PS Point Source
PSL Priority Substances List (PSL1)
PVC Polyvinyl chloride
QFF Far field ventilation rate
QNF Near field ventilation rate
QA Quality assurance
QC Quality control
RAD Risk Assessment Division
RCRA Resource Conservation and Recovery Act
RfC Reference concentration
RfD Reference dose
RR Rate ratio
RRm Summary relative risk
RTECS Registry of Toxic Effects of Chemical Substances
s Second(s)
SAB Science Advisory Board
SARA Superfund Amendments and Reauthorization Act
SCG Scientific Consulting Group, Inc.
SD Standard deviation
t Time
TCA Trichloroacetic acid
TCE Trichloroethylene
TCOG Trichloroethanol, glucuronide conjugate
TCOH Trichloroethanol
TOXLINE Toxicology Literature Online
TRI Toxics Release Inventory
TTC Total trichloro compounds
TSCA Toxic Substances Control Act
TSCATS Toxic Substance Control Act Test Submission Database
TWA Time-weighted average
UF Uncertainty factor
UFS Subchronic to chronic uncertainty factor
UFA Interspecies uncertainty factor
UFH Intraspecies uncertainty factor
UFL LOAEL to NOAEL uncertainty factor
UFD Database uncertainty factor
US or U.S. United States
U.S. EPA or EPA United States Environmental Protection Agency
VFF Far field volume
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VNF Indoor wind speed
VNF Near field volume
VCCEP Voluntary Children's Chemical Evaluation Program
VOC Volatile organic compound
VP Vapor pressure
WNF Near field width
WY Working years
Yr(s) Year(s)
Page 17 of 212
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EXECUTIVE SUMMARY
The United States Environmental Protection Agency (U.S. EPA), Office of Pollution Prevention
and Toxics (OPPT), identified and chose trichloroethylene (TCE) for risk evaluation as part of its
Existing Chemicals Management Program under the Toxics Substances Control Act (TSCA).
TCE is a volatile organic compound (VOC) that is classified as a human carcinogen. Its
consumption in the U.S. is 255 million pounds (Ibs) per year. TCE is widely used in industrial and
commercial processes, and also has some limited uses in consumer products.
Main Conclusions of this Risk Assessment
This risk assessment identifies cancer risk concerns and short-term and long-term non-cancer
risks for workers and occupational bystanders at small commercial degreasing facilities and dry
cleaning facilities that use TCE-based solvents and spotting agents, respectively.
EPA/OPPT also identifies short-term non-cancer risks for consumers and residential bystanders
from the use of TCE-containing solvent degreasers and spray-applied protective coatings.
The Focus of this Risk Assessment
This assessment characterizes human health risks from inhalation exposures to TCE for the
following uses:
1. Commercial use of TCE as a solvent degreaser
2. Consumer use of TCE as a solvent degreaser
3. Consumer use of TCE as a spray-applied protective coating for arts and crafts
4. Commercial use of TCE as a spotting agent at dry-cleaning facilities
EPA/OPPT selected these uses because they were expected to make frequent use of TCE in high
concentrations and/or pose high potential for human exposure. Additional information is
provided in the risk assessment regarding the criteria for inclusion of uses and the various
assumptions in applying these criteria.
The main route of exposure for TCE is believed to be inhalation for the uses identified in this
assessment. EPA/OPPT recognizes that highly volatile compounds such as TCE may also be
absorbed through the skin. However, based on the physical-chemical properties of TCE and the
scenarios described in this assessment, EPA/OPPT believes that inhalation is the main exposure
pathway for this risk assessment. Recent modeled and experimental work supports this
assumption that inhalation is the predominant exposure pathway (see Section 1.3.2). This
assessment may underestimate total exposures resulting from the uses of TCE due to this
assumption.
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This risk assessment does not include an assessment of environmental effects. Based on TCE's
moderate persistence, low bioaccumulation, and low hazard for aquatic toxicity, potential
environmental impacts are judged to be low for the environmental releases associated to the
TSCA uses under the scope of this risk assessment. That judgment should not be
misinterpreted as saying that the fate and transport properties of TCE suggest that water and
soil contamination is likely low or do not pose an environmental concern. In fact, EPA's Office of
Solid Waste and Emergency Response and the EPA Regions are addressing TCE contamination
in groundwater and contaminated soils at large number of sites. While the primary concern
with this contamination has been human health, there is potential for TCE exposures to
ecological receptors in some cases.
Human Populations Targeted in This Assessment
EPA/OPPT assessed acute and chronic risks for workers at small degreasing facilities and dry
cleaning facilities that may use TCE as a solvent degreaser or spotting agent, respectively.
EPA/OPPT assumes that workers at these small degreasing and dry cleaning facilities would be
adults of both sexes (>16 and older, including pregnant women) based upon occupational work
permits, although exposures to younger individuals may be possible in occupational settings.
Risks are also estimated for occupational bystanders, who are assumed to be workers in the
vicinity of the degreasing and spotting operations, but not actually performing the operation.
EPA/OPPT also examined acute risks for consumer exposures in residential settings. EPA/OPPT
assumes that consumers would be individuals that intermittently use TCE in and around their
homes, whereas bystanders would be individuals physically close to the use activity but not
using the product. EPA/OPPT assumes that consumer users would generally be adults of both
sexes (>16 and older, including pregnant women), although exposures to teenagers and even
younger individuals may be possible in residential settings. However, risk estimates are focused
on the most susceptible life stage, which are pregnant women and their developing fetus. This
focus is supported by the hazard findings in the TCE IRIS assessment, which conclude that
developmental toxicity is the most sensitive health effect associated to TCE exposure (EPA,
2011e).
Workplace Exposures at Commercial Degreasing and Dry Cleaning Facilities
In order to estimate TCE emissions in the workplace, EPA/OPPT used readily available
information from the National Emissions Inventory (NEI), the Toxics Release Inventory (TRI),
and a study on the use of spotting chemicals prepared for the California EPA and EPA Region 9
(CalEPA/EPA, 2007). To estimate workplace exposures, these emission estimates were
incorporated into a Near Field/Far Field (NF/FF) mass balance model.
It is important to note that the NF/FF model has been extensively peer-reviewed, is extensively
used, and results of the model have been compared with measured data. The comparison
indicated that model and measured values agreed to within a factor of about three (Jayjock et
al., 2011). In estimating workplace exposures, EPA/OPPT assessed various exposure scenarios.
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For example, engineering controls such as local exhaust ventilation (LEV) were taken into
account.
Although relevant exposure monitoring data are limited, EPA/OPPT did identify monitoring data
from the Occupational Safety and Health Administration (OSHA)(Coble, 2013) and site-specific
data from the National Institute for Occupational Safety and Health (NIOSHHNIOSH. 1997b).
The exposure estimates (with and without LEV) were of the same order of magnitude as
measured values:
1. For commercial degreasing facilities, EPA's exposure estimate ranged from 0.04 to 197 parts
per million (ppm); measured data from OSHA ranged from 0.06 to 380 ppm.
2. For dry cleaning facilities, EPA's site-specific exposure estimate ranged from 0.8 to 2.1 ppm;
measured data reported by NIOSH ranged from 2.37 to 3.11 ppm.
Consumer Exposures from Solvent Degreasing and Spray-Applied Coatings
EPA/OPPT used the Exposure and Fate Assessment Screening Tool Version 2 (E-FAST2)
/Consumer Exposure Module (CEM) (EPA, 2007b) to estimate TCE exposures for the consumer
use scenarios. This modeling approach was selected because emissions and monitoring data
were not available for the TCE uses under consideration.
The model used a two-zone representation of a house to calculate the TCE exposure levels for
consumers and bystanders. The modeling approach integrated assumptions and input
parameters about the chemical emission rate over time, the volume of the house and the room
of use, the air exchange rate and interzonal airflow rate. The model also considered the
exposed individual's locations, body weights and inhalation rates during and after the product
use (EPA. 2007b).
The high-end inhalation exposure estimates for the consumer scenarios were as follows:
1. 0.4 ppm for users of TCE-containing clear protective coating sprays
2. 0.1 ppm for bystanders of TCE-containing clear protective coating sprays
3. 2 ppm for users of TCE-containing solvent degreasers
4. 0.8 ppm for bystanders of TCE-containing solvent degreasers
Characterization of Hazards and Risks to Human Health
The assessment uses the hazard and dose-response information published in the final
toxicological review that the U.S. EPA's Integrated Risk Information System (IRIS) published in
2011 (EPA, 2011e). The TCE IRIS assessment used a weight-of-evidence approach, the latest
scientific information and physiologically-based pharmacokinetic (PBPK) modeling to develop
hazard and dose-response assessments for TCE's carcinogenic and non-carcinogenic health
effects resulting from lifetime inhalation and oral exposures. In addition to relying on the latest
scientific information, the TCE IRIS assessment underwent several levels of peer review
including agency review, science consultation on the draft assessment with other federal
agencies and the Executive Office of the President, public comment, external peer review by
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the EPA's Science Advisory Board (SAB) in 2002, scientific consultation by the U.S. National
Academy of Sciences (NAS) in 2006, external peer review of the revised draft assessment by the
EPA's Science Advisory Board (SAB) in January 2011, followed by final internal agency review
and EPA-led science discussion on the final draft.
TCE's Carcinogenic Hazards and Risks:
TCE is carcinogenic to humans by all routes of exposure as documented in the TCE IRIS
assessment. This conclusion is based on strong cancer epidemiological data that reported an
association between TCE exposure and the onset of various cancers, primarily in the kidney,
liver and the immune system (i.e., non-Hodgkin lymphoma or NHL) (EPA, 2011e). Further
support for TCE's carcinogenic characterization comes from (1) positive results in multiple
rodent cancer bioassays in rats and mice of both sexes, (2) similar toxicokinetics between
rodents and humans, (3) mechanistic data supporting a mutagenic mode of action for kidney
tumors, and (4) the lack of mechanistic data supporting the conclusion that any of the mode(s)
of action for TCE-induced rodent tumors are irrelevant to humans (EPA, 2011e). Additional
support comes from the recent evaluation of TCE's carcinogenic effects by the International
Agency for Research on Cancer (IARC). IARC (2014) classifies TCE as carcinogenic to humans
(Group 1).
EPA/OPPT used the inhalation unit risk (IUR) of 2 x 10"2 per ppm (4 x 10"6 per u,g/m3) reported in
the TCE IRIS assessment to estimate excess cancer risks for the occupational scenarios. The IUR
is the estimated upper bound excess lifetime cancer risk resulting from continuous exposure to
an airborne agent at 1 u.g/m3 (EPA, 2011b). The IUR for TCE is based on human kidney cancer
risks and adjusted for potential risks for NHL and liver cancer based on human epidemiological
data (EPA, 2011e). There is high confidence in the IUR because it is based on good quality
human data and it is similar to unit risk estimates derived from multiple rodent bioassays (EPA,
2011e). Moreover, there is sufficient weight of evidence to conclude that TCE operates through
a mutagenic mode of action for kidney tumors, which supports the linear extrapolation
approach (EPA, 2011e).
TCE's Non-Carcinogenic Hazards and Risks:
TCE exposure is associated with a range of non-cancer health effects in humans and animals.
Non-cancer risks for the various exposure scenarios were evaluated using the dose-response
information reported in the TCE IRIS assessment (EPA, 2011e). The TCE IRIS assessment used
physiologically-based pharmacokinetic (PBPK) modeling to estimate hazard values (i.e., human
equivalent concentrations or HECs) indicative of adverse health effects representing six health
effects domains: kidney, liver, immunotoxicity, neurotoxicity, reproductive toxicity, and
developmental toxicity.
Different health endpoints were used to evaluate risks based on the expected durations of
exposure in the scenarios considered in this assessment. For instance, both acute and chronic
health effects endpoints were used for the occupational scenarios (i.e., small commercial
degreasers and spot cleaning workers and bystanders). In that case, a variety of health effects
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endpoints were used to evaluate repeated (chronic) exposures to TCE (i.e., toxicity to the liver,
kidney, nervous system, immune system, the reproductive system, and developmental toxicity).
For the consumer use scenarios, developmental toxicity endpoints were used to assess risks for
acute exposures.
EPA/OPPT used developmental endpoints for the acute risk assessment based on U.S. EPA's
policy that a single exposure of a chemical within a critical window of fetal development may
produce adverse developmental effects (EPA, 1991). Particularly, this assessment used the
PBPK-derived HECs reported for developmental animal studies reporting fetal cardiac defects.
TCE-induced fetal cardiac malformations are biologically plausible based on the weight of
evidence analysis presented in the TCE IRIS assessment, which considered human and animal
findings as well as mechanistic data.
These hazard values were expressed as HECs at the 50th, 95th or 99th percentile of the combined
uncertainty and variability distribution of human internal doses, as estimated by the TCE PBPK
model (EPA, 2011e). The HEC95 and HEC99 were defined as the concentrations of TCE in air for
which there is 95% and 99% likelihood, respectively, that a randomly selected individual would
have an internal dose less than or equal to the internal dose of the hazard value. On the other
hand, the HEC50was defined as the concentration of TCE in air for which there is a 50%
likelihood that a randomly selected individual would have an internal dose less than or equal to
the internal dose of the hazard value (EPA, 2011e). The TCE IRIS assessment preferred the HEC99
for the non-cancer dose-response derivations because the HEC99was interpreted to be
protective for a sensitive individual. EPA/OPPT supported the interpretation of the HEC99 as
expressed in the TCE IRIS assessment. Hence, HEC99-based risk estimates are favored in this
assessment over those estimated from the HEC50and HEC95 values, but risk estimates for all of
the HEC percentiles were presented to provide a sense of the variability in the risk estimates.
EPA/OPPT used margin of exposures (MOEs) to estimate non-cancer risks based on (1) the
lowest PBPK-derived HECs within each health effects domain reported in the TCE IRIS
assessment; (2) the same endpoint/study-specific uncertainty factors (UFs) that the IRIS
program applied to the PBPK-derived HECs; and (3) the exposure estimates calculated for the
TCE uses examined in this risk assessment. MOEs allowed us to have a better picture of the
non-cancer risk profile of TCE by presenting a range of risk estimates for different non-cancer
health effects for different exposure scenarios.
Uncertainties of this Risk Assessment
As with any risk assessment, there are uncertainties that need to be considered when
interpreting the results. Assumptions were used in estimating the occupational and consumer
exposure scenarios covered in this assessment. In addition, there are uncertainties in the
hazard/dose-response and risk characterization assessments. EPA/OPPT discusses these
uncertainties qualitatively and recognizes that they may under- or over-estimate actual risks.
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The Results of this Risk Assessment
Size of the Exposed Population
• Approximately 30,000 workers and occupational bystanders at small commercial degreasing
operations.
• Approximately 300,000 workers and occupational bystanders at dry cleaning operations.
• No data were available to estimate the number of consumers and bystanders exposed to
TCE during the use of degreasers and arts/crafts clear protective coating spray.
Cancer Risks:
• There are cancer risk concerns for users and bystanders occupationally exposed to TCE
when using TCE-containing vapor degreasers and spot cleaners in small commercial shops
and dry cleaning facilities, respectively.
• Many of the commercial vapor degreasing and spot cleaning exposure scenarios exceed the
excess lifetime cancer risk probabilities of 1 chance in 10,000, 100,000 or 1 million (i.e.,
target cancer risks of 10"4, 10"5 and 10"6, respectively) of an individual developing cancer.
• The occupational exposures to commercial degreasers show the greatest cancer risk when
compared to the spot cleaning exposure scenarios.
Acute Non-Cancer Risks:
• There are acute non-cancer risks for developmental effects (i.e., cardiac defects) for most
occupational and residential exposure scenarios (i.e., MOEs were below the benchmark
MOEof 10).
• The commercial vapor degreasing and consumer spray degreasing exposure scenarios show
greater acute risks for developmental effects than those reported for the spot cleaning
exposure scenarios.
Chronic Non-Cancer Risks:
• There are chronic non-cancer risks for a range of human health effects in both the
occupational degreaser and spot cleaning exposure scenarios (i.e., MOEs were below the
benchmark MOE).
• The greatest concern is for developmental effects (i.e., fetal cardiac defects), followed by
kidney effects and then immunotoxicity, with an overall higher chronic risk for the
degreaser exposure scenarios. In general, this concern is present for lower and upper-end
exposures and in the presence or absence of room ventilation (LEV vs. no LEV).
• There are chronic risks for reproductive effects and neurotoxicity for degreaser worker
exposure scenarios and most of the degreaser bystander exposure scenarios. However, the
risks concern for these effects are reported for fewer spot cleaning worker and bystander
scenarios, and are generally attributed to exposure conditions without room ventilation.
• There are chronic risks for liver effects although the risks were less prominent than those
reported for other health effects. These risks were found only in the degreaser worker and
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bystander exposure worst case scenarios, and the spot cleaning worker and bystander
worst case scenarios with no LEV.
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1 PURPOSE, BACKGROUND, AND SCOPE
1.1 PURPOSE AND AUDIENCE
The purpose of TSCA Work Plan chemical risk assessments, such as this one developed for TCE,
is to assess the potential risks of chemicals under the EPA's Existing Chemicals Program. If risks
are found in the risk assessments, the information will be used to inform risk management
strategies to reduce identified risks.
The target audience for this risk assessment is the risk assessment community, including U.S.
EPA's risk assessors and risk managers, as well as U.S. stakeholders that are interested in risk
assessment issues related to TCE. 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 chemical risks of TCE. The risk assessment may also help those
interested in reducing risk in the identified use areas of commercial and consumer solvent
degreasing and arts and crafts products.
1.2 BACKGROUND
1.2.1 Rationale for Selecting TCE for Risk Assessment
TCE is a liquid VOC that ranked high for human health hazards and exposure potential when
using OPPT's work plan chemical prioritization criteria (EPA, 2012d)1. TCE is classified as a
human carcinogen and is widely used in industrial and commercial processes as well as in some
consumer products. Moreover, these TCE uses may pose an inhalation hazard as evaporation of
TCE readily occurs due to its high vapor pressure. TCE is ubiquitously present in the
environment with levels detected in drinking water, indoor environments, surface water,
ambient air, groundwater, and soil.
Given these concerns, TCE was identified and chosen for risk evaluation as part of EPA/OPPT's
Existing Chemicals Management Program under TSCA2.
1.2.2 Overview of TCE Uses, Production Volume and EPA's Regulatory
History
TCE has historically had a wide range of uses drawn from various markets, including
intermediate chemicals (for refrigerant and polyvinyl chloride [PVC] manufacture), industrial
and commercial solvents, Pharmaceuticals, insecticides, fumigant, textiles (processing and
1 EPA's TSCA Work Plan Chemicals: Methods Document:
http://www.epa.gov/oppt/existingchemicals/pubs/wpmethods.pdf
2 EPA's TSCA Work Plan Chemicals website: http://www.epa.gov/oppt/existingchemicals/pubs/workplans.html
Page 25 of 212
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flame retardants), adhesives, and paints (as diluent) (Ash and Ash, 2009). However, as of 2011,
most U.S. consumption is attributable to two specific uses: 83.6 percent of total TCE production
volume is used as an intermediate for manufacturing the refrigerant (closed system) HFC-134a
(a major alternative to CFC-12), and 14.7 percent is used as a solvent for metals degreasing
solvent; the remaining 1.7 percent is attributed to "other uses." In total, U.S. TCE consumption
is 255 million Ibs/yr (Glauserand Funda, 2012)3. More information on production volumes can
be found in Appendix A.
The U.S. EPA regulates TCE through various environmental regulations given TCE's human
health hazards and exposure potential. The latter is informed by TCE's high production volume,
physical-chemical properties, uses and environmental releases. Appendix B summarizes the U.S.
EPA's regulatory history of TCE, including those U.S. EPA offices that manage the
implementation of the existing TCE regulations.
Under the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA)
and the Superfund Amendments and Reauthorization Act (SARA), the U.S. EPA and the Agency
for Toxic Substances and Disease Registry (ATSDR) have established a prioritized list of
substances most commonly found at facilities on the National Priority List (NPL) (ATSDR, 2011).
The listing is based on the frequency of occurrence, toxicity, and potential for human exposure.
TCE was ranked 16th out of 847 candidate substances for the 2011 ranking; only the top 275 are
considered to be on the list.
1.3 SCOPE OF THE ASSESS^
1.3.1 Selection of TCE Uses
This assessment characterizes inhalation exposures to TCE from the following uses:
1. Commercial use of TCE as a solvent degreaser
2. Consumer use of TCE as a solvent degreaser
3. Consumer use of TCE as a spray-applied protective coating for arts and crafts
4. Commercial use of TCE as a spotting agent at dry-cleaning facilities
Table 1-1 lists the primary uses of TCE, indicates whether a use was considered in this
assessment, and also presents the rationale for why a use was included or excluded from
further consideration. The criteria for inclusion were: high concentration of TCE, frequent use
of TCE, and high potential for human exposure.
This source material includes data or information derived from IMS Products provided to the U.S. EPA. IMS
products have been provided to the U.S. EPA for its internal use and in the context of a license agreement. By
receiving and accessing this material, you agree that IMS is not liable to you or any third party for your use of
and/or reliance on the IMS data and information contained in this document, and any such use shall be at your
own risk.
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EPA/OPPT made various assumptions in developing the use-related criteria. For instance, when
using TCE as an intermediate for refrigerant manufacturing, the frequency of use is likely to be
high as is the concentration of TCE. However, the process is expected to take place in a closed
system (low potential for human exposure) and thus this use is not considered in this
assessment. Also, although the use of TCE as a solvent degreaser at large commercial/industrial
operations is expected to be frequent and the concentration of TCE high, human exposures in
these settings are expected to be monitored and controlled by Occupational Safety and Health
Administration (OSHA); thus, this use is also not considered in this assessment.
Toner aides and mirror edge sealants were dropped from further consideration because these
uses reported a low TCE content and we4 assumed less frequent use when compared to the
consumer use of degreasers and arts/crafts products. Although film cleaners had a high TCE
content, we did not evaluate this use because the exposure frequency is expected to be low
since the U.S. consumer population predominantly uses digital cameras, and even for film use,
only people performing home development of camera film would be exposed to these
products. We cannot rule out that frequent exposure to TCE could occur in a small population
of amateur photographers that use film cleaning products. In addition, professional
photographers' use of products containing TCE while developing film would not be assessed as
a consumer use.
Table 1-1. Primary Uses of TCE and Determination of Inclusion in this Risk Assessment
Use
Category
Typical
Percent
TCE by
Weight
Population
Exposed
To Be Considered in
this Assessment?
Intermediate
in the manufacturing
of refrigerant
>99
Workers and
bystanders in the
refrigerant
manufacturing
process (all adults1)
No - high content, frequent use, low
potential for human exposure (the use
of TCE as an intermediate is expected
to take place in a closed system)
Solvent degreaser
>90
Workers and
bystanders in large
commercial/industrial
settings (all adults1)
No- high content, frequent use, low
potential for human exposure
(exposures at large
commercial/industrial operations are
expected to be monitored and
controlled by OSHA)
Solvent degreaser
>90
Workers and
bystanders in small
commercial settings
(all adults1)
Yes - high content, potential for
frequent use, high potential for
human exposure (i.e., chronic
exposures)
1 For the purpose of this risk assessment, "we" refers to U.S. EPA/OPPT.
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Table 1-1. Primary Uses of TCE and Determination of Inclusion in this Risk Assessment
Use
Category
Spotting agent
Solvent degreaser
Plastic clear
protective coating
spray (hobbyists;
arts/crafts)
Film cleaner
(hobbyists)
Toner aide
(home office)
Mirror edge sealant
(hobbyist/home
maintenance)
Typical
Percent
TCE by
Weight
10-100
>90
20-30
>90
15-20
20-30
Population
Exposed
Workers and
bystanders at dry
cleaning facilities
(all adults1)
Consumer users
(adults >16 yrs old *)
and bystanders 2
(all ages)
Consumer users
(adults >16 yrs old *)
and bystanders 2
(all ages)
To Be Considered in
this Assessment?
Yes - potential for high content,
potential for frequent use, high
potential for human exposure (i.e.,
chronic exposures)
Yes - high content, low frequency of
use, high potential for human
exposure (i.e., acute exposures)
Yes - low content, but possibly largest
use of consumer products (i.e., acute
exposures)
No - high content, low frequency of
use (use of negatives/cameras with
film is assumed to be negligible); low
potential for human exposure
No - low content, low frequency of
use, low potential for human exposure
No - low content, low frequency of
use, low potential for human exposure
Notes:
1= "adults" include individuals of both sexes, including pregnant women (>16 yrs of age)
2= "bystanders" include individuals of both sexes, including children and pregnant women
1.3.2 Selection of Exposure Pathway
This risk assessment assumed that TCE is primarily absorbed through the respiratory tract
because of TCE's high vapor pressure. EPA/OPPT recognizes that highly volatile compounds
such as TCE may also be absorbed through the skin. However, based on the physical-chemical
properties of TCE and the scenarios described in this assessment, EPA/OPPT believes that
inhalation is the main exposure pathway for this risk assessment. This assessment may
underestimate total exposures resulting from the uses of TCE due to this assumption.
Recent modeled and experimental work supports the assumption that inhalation is the
predominant exposure pathway. The dermal model described by Tibaldi et al. (2014) estimates
that about 1% of TCE on the skin will be absorbed into the epidermis with the other 99%
evaporating. Also, an experimental comparison of dermal to vapor exposure found that TCE and
hexane had the least dermal absorption amongst a set of volatile solvents. The ratio of dermal
to respiratory intake was found to be 0.1 % for TCE (Kezic et al., 2000).
Page 28 of 212
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J±^L^
EPA/OPPT assessed risks for workers at small degreasing facilities and dry cleaning facilities that
may use TCE as a solvent degreaser or spotting agent, respectively, at some point as a part of
their daily activities. We assumed that workers at these small degreasing facilities and dry
cleaning facilities would be adults of both sexes (>16 and older), although exposures to younger
individuals may be possible in occupational settings. EPA/OPPT also assumed that pregnant
women may be part of the workforce. Risks were also estimated for bystanders (i.e., workers in
the vicinity of degreasing and spotting operations, but not actually performing the operation).
This assessment also examined consumer exposures in residential settings. Consumers were
individuals that used TCE in and around their homes (Table 1-1), whereas bystanders were
individuals that did not use the product but were physically close to the use activity. We
assumed that consumer users would be adults of both sexes (>16 and older, including pregnant
women), although exposures to younger individuals may be possible in residential settings.
J^S-^^RiskEvaluatio^
Risks for the various exposure scenarios were evaluated using the dose-response information
that was reported in the final TCE health assessment prepared by the U.S. EPA's Integrated Risk
Information System (IRIS) (EPA, 2011e)5. The final TCE IRIS assessment characterizes TCE as
carcinogenic to humans and identifies non-cancer hazards associated with TCE exposure. IARC
(2014) has also classified TCE as carcinogenic to humans (Group 1).
The TCE IRIS assessment underwent several levels of peer review including agency review,
science consultation on the draft assessment with other federal agencies and the Executive
Office of the President, public comment, external peer review by the EPA's Science Advisory
Board (SAB) in 2002, scientific consultation by the U.S. National Academy of Sciences (NAS) in
2006 (NRC, 2006)6, external peer review of the revised draft assessment by the EPA's Science
Advisory Board (SAB) in January 2011 (EPA, 2011c)7, followed by final internal agency review
and EPA-led science discussion on the final draft.
In light of this, EPA/OPPT decided to use the TCE IRIS assessment as the preferred data source
for TCE's human health toxicity information, rather than developing a new hazard and dose-
response assessment. See Chapter 2 for more information about the hazard/dose-response
approach for cancer and non-cancer health endpoints.
5 From now on, the EPA's IRIS Toxicological Review of TCE will be referred as the "TCE IRIS assessment".
NAS report, "Assessing the human health risks of trichloroethylene: Key scientific issues (2006)":
http://www.nap.edu/catalog.php7record id=11707
EPA's SAB peer review report for the 2009 EPA's Draft Assessment entitled 'Toxicological Review of
Trichloroethylene":
http://vosemite.epa.gov/sab/sabproduct.nsf/c91996cd39a82f648525742400690127/B73D5D39A8F184BD8525
7817004A1988/$File/EPA-SAB-ll-002-unsigned.pdf
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1.3.5 Why Environmental Risks Were Not Evaluated For Selected TCE
Uses
EPA/OPPT did not assess the risks of environmental effects related to the manufacture and use
of TCE for the TSCA uses under consideration in this assessment. The decision of not conducting
an environmental assessment was based on TCE's moderate persistence, low bioaccumulation,
and low hazard for aquatic toxicity. This information supported a low concern for potential
environmental impacts related to the TCE releases associated to the TSCA uses under the scope
of this risk assessment.
EPA/OPPT also assumed that very low concentrations of TCE would be present in surface water.
In making this assumption, EPA/OPPT evaluated the TCE releases to water and wastewater
treatment reported to the Toxics Release Inventory (TRI) as well as the fate of TCE in
wastewater treatment.
Total releases of TCE to water for all industries reporting to TRI from 1988 to 2012 ranged from
a high of 15,849 pounds in 1989 and trended downward to a low of 67 pounds in 2012. Total
transfers of TCE to publicly owned treatment works (POTWs) for all industries reporting to TRI
from 1988 to 2012 ranged from a high of 78,921 pounds in 1996 and trended downward to a
low of 100 pounds in 2012. Also, TCE concentrations are reduced through volatilization when
entering surface waters from POTWs. The full environmental fate assessment is located in
Appendix C.
Information on the aquatic toxicity of TCE suggested no immediate concern for potential
environmental effects for the TSCA uses under consideration. The European Union (EU)'s TCE
risk assessment also concluded that there were no concerns for environmental effects on
aquatic organisms, including benthic organisms, terrestrial organisms, and the atmosphere
(ECB, 2004). The EU conclusions were based on the production and use of TCE, including
releases to wastewater treatment plants, to air from all uses, and from dichloroacetic acid (i.e.,
photodegradation product of TCE).
The absence of an environmental risk assessment of the TCE TSCA uses should not be construed
as saying that the fate and transport properties of TCE suggest that water and soil
contamination is likely low or do not pose an environmental concern. In fact, EPA's Office of
Solid Waste and Emergency Response and the EPA Regions are addressing TCE contamination
in groundwater and contaminated soils at large number of sites. While the primary concern
with this contamination has been human health, there is potential for TCE exposures to
ecological receptors in some cases.
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2 HUMAN HEALTH RISK ASSESSMENT
2.1
TCE is a colorless liquid with a pleasant, sweet odor resembling that of chloroform. It is
considered a VOC because of its moderate boiling point, 87.2 °C, and high vapor pressure, 73.46
mm Hg at 25 °C. TCE is moderately water soluble (1.280 g/L at 25 °C), and has a log
octanohwater partition coefficient (Kow) of 2.42. The density of TCE, 1.46 g per cubic meter
(cm3) at 20 °C, is greater than that of water. Table 2-1 lists the chemical/physical properties of
TCE.
Table 2-1. Physical-Chemical Properties of TCE
Property
CASRN
Molecular weight
Molecular formula
Physical state
Odor
Density
Flash point
Autoflammability
Viscosity
Refractive index
Dielectric constant
Melting point
Boiling point
Vapor pressure
Dissociation constant (pKa)
Henry's law constant
Water solubility
Octanol/water partition
coefficient (Kow)
Value
79-01-6
131.39
C2HCI3
Colorless liquid
Sweet, pleasant, resembles chloroform
1.46 g/cm3 @ 20 °C a
90 °C (closed cup) a
410 °Cb
0.53mPa-s@25°Cc
1.4775d
3.4 e0@ 16°CC
-84.7 °C (measured)
87.2 °C (measured)
73.46 mm Hg at 25 °C (measured) e
Not applicable
9.85xlO~3atm-m3/mole (measured)
1,280 mg/L at 25 °C (measured) f
2.61(measured) g
Sources:
a ECB (2000) d O'Neil et al. (2001) f Horvath et al. (1999)
b SRC (2012) e Daubert and Danner (1989) g Hansch et al. (1995)
°Weast and Selbv (1966)
Page 31 of 212
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2.2 OVERVIEW OF ENVIRONMENTAL FATE AND RELEASES OF TCE
Knowledge of the environmental fate (transport and transformation) of a compound is
important to understanding its potential impact on specific environmental media (e.g., water,
sediment, and soil) and exposures to target organisms of concern.
TCE is a volatile liquid with high vapor pressure, moderate water solubility, and high mobility in
soil. TCE is slowly degraded by sunlight and reactants when released to the atmosphere.
Volatilization and microbial biodegradation influence the fate of TCE when released to water,
sediment or soil. The biodegradation of TCE in the environment is dependent on a variety of
factors and so a wide range of degradation rates have been reported (ranging from days to
years). TCE is not expected to bioconcentrate in aquatic organisms due to measured
bioconcentration factors of less than 1000. More information on TCE's environmental fate is in
Appendix C.
The manufacture, processing and use of TCE can result in TCE releases to air, water, sediment,
and soil. Most reported environmental releases of TCE are to air with much lower releases to
landfills and very little released to surface water. The Toxics Release Inventory database
reported some of the highest fugitive and point source air releases reported for TCE when used
as a degreaser (i.e., roughly 15 percent of the production/importation volume in the U.S.).
Disposal of TCE wastes could be an environmental concern because TCE has moderate
persistence under certain environmental conditions, is volatile and water soluble, and has high
mobility in soil and groundwater. TCE may enter publicly owned treatment works (POTWs),
which will likely result in releases to surface waters and air. More information on TCE's
environmental fate is in Appendix C.
2.2.1 Ambient Air Concentrations of TCE
Table 2-2 lists a summary of the ambient air monitoring data for TCE (i.e., measured data) in the
U.S. from 1999 to 2006 as reported in EPA (2011e). These data suggest that TCE levels have
remained fairly constant in ambient air for the U.S. since 1999, with an approximate mean
value of 0.3 u.g/m3 (i.e., which is equivalent to 5.6 x 10"5 parts per million or ppm).
Page 32 of 212
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Table 2-2. TCE Ambient Air Monitoring Data (|ig/m3) a
Year
1999
2000
2001
2002
2003
2004
2005
2006
Number of
Monitors
162
187
204
259
248
256
313
258
Number of
States
20
28
31
41
41
37
38
37
Mean
0.30
0.34
0.25
0.37
0.35
0.32
0.43
0.23
Standard
Deviation
0.53
0.75
0.92
1.26
0.64
0.75
1.05
0.55
Median
0.16
0.16
0.13
0.13
0.16
0.13
0.14
0.13
Range
0.01-4.38
0.01-7.39
0.01-12.90
0.01-18.44
0.02-6.92
0.00-5.78
0.00-6.64
0.03-7.73
a The U.S. EPA's Air Quality System database at the AirData Web site: http://www.epa.gov/airdata/ (as
summarized in EPA(2011e). Note that the data were not from a statistically based survey and cannot be
assumed to provide nationally representative values.
2.3 ENVIRONMENTAL RELEASES AND OCCUPATIONAL EXPOSURE
ESTIMATES FOR SMALL COMMERCIAL DECREASING
OPERATIONS
2.3.1 What is a Small Operation?
There is no standard or universal definition for the term "small shop". The various meanings of
this term can depend upon the industry sector (e.g., metal finishing, furniture repair, foam
production, chemical manufacturing) or governmental jurisdiction (e.g. OSHA, EPA, other
countries). For the purpose of risk assessment of work plan chemicals, EPA generally refers to
entities, businesses, operators, plants, sites, facilities, or shops interchangeably and considers a
number of factors to categorize these as small. The factors that have been usually considered
include revenue; capacity, throughput, or production or use rate of materials, or number of
employees.
For this risk assessment, EPA/OPPT has determined that more research will be required to
determine which factors will best define small shops for the industries that do vapor
degreasing. However, EPA/OPPT's interest in small shops for this assessment is due to the
possibility that these shops may have fewer resources or less expertise and awareness of
hazards, exposures, or controls as compared to large shops.
2.3.2 A Brief Summary of Solvent Cleaning
Solvent cleaning (degreasing) is widely used to remove grease, oils, waxes, carbon deposits,
fluxes, and tars from metal, glass, or plastic surfaces (EPA, 2006d, 2007a).
There are two general types of degreasing machines: batch and in-line. Batch cleaning
machines are the most common type, while in-line cleaners are typically used in large-scale
Page 33 of 212
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industrial operations (EPA, 2006d). The size of a degreasing machine is defined by the area of its
solvent-to-air interface.
Emissions from degreasing machines typically result from (EPA, 2006d):
1. evaporation of the solvent from the solvent-to-air interface
2. "carry out" of excess solvent on cleaned parts
3. evaporative losses of the solvent during filling and draining of the degreasing machine
2.3.3 EPA/OPPT's Release and Exposure Assessment for Degreasing
Operations
EPA/OPPT used readily available information from the National Emissions Inventory (NEI) and
the Toxics Release Inventory (TRI) to estimate releases of and exposures to TCE from degreasing
operations. EPA/OPPT used several steps in order to produce these estimates. An overview of
these steps is shown in Figure 2-1. Further elaboration of these steps has been structured to
follow the sequence in this figure.
Figure 2-1. Overview of EPA/OPPT's occupational release and exposure assessment for
degreasing operations
Identify relevant industries by North American
Industrial Classification System codes (NAICS
codes)
Estimate air releases of TCE based on the National
Emissions Inventory (NEI) and Toxics Release
Inventory (TRI):
1. Compare NEI and TRI air releases for the same year (2008)
a. NEI air emissions are approximately 2 times (2X)
higher than those reported in TRI
b. Use data from 2011 TRI but scale air releases by 2X;
this captures downward trends in usage {if any) but
also accounts for NEI and TRI discrepancy
®
Estimate facility and operating parameters:
1. Use data from the 2008 NEI to identify/estimate:
a. Number of point/nonpoint sources
b. Degreasing units per facility
2. Use EPA's risk assessment for the halogenated solvent
cleaning source category to estimate number of facilities.
3. Use EPA's draft generic scenario on vapor degreasing to
estimate days and hours of operation per year.
Estimate TCE emissions
potentially escaping into the
workplace
Estimate potential workplace
exposures
1. Use a Near-Field/Far-Field
mass balance model
2. Estimate exposures for workers
and occupational non-users
Page 34 of 212
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5. Identification of relevant industries by North American Industrial Classification System
(NAICS) codes: The degreasing process is used in many industries, both large and small.
Based on a review of EPA's 2008 NEI, 78 different NAICS codes were identified. These NAICS
codes are listed in Appendix D.
6. Estimation of air releases of TCE based on NEI and TRI data: As background, the NEI is a
comprehensive and detailed estimate of air emissions of both criteria and hazardous air
pollutants (HAPs) from various air emissions sources. The NEI is prepared every three years
by the U.S. EPA (EPA, 2008). The TRI is a database that contains detailed information on
environmental releases and transfers of certain listed toxic chemicals from industrial
facilities. The TRI is maintained by the U.S. EPA and updated annually (EPA, 2012c).
In 2008, for the NAICS codes listed in Appendix D, total NEI air emissions of TCE were
approximately two times (2X) greater than those reported in TRI. Whereas the NEI is the
U.S. EPA's primary emissions inventory for HAPs and criteria pollutants, the TRI is another
inventory that may be considered. The TRI provides releases to other environmental media
besides air (e.g., land and water). However, the TRI may exclude releases from small-scale
operations, which are the intended focus of this risk assessment (EPA, 2004a, 2011d).
In this assessment, EPA/OPPT estimated TCE air emissions based on the
2011 TRI. EPA/OPPT elected to use more recent data from the 2011 TRI to
account for downward trends in usage/release of TCE. In order to account
for the discrepancy observed between the 2008 NEI and 2008 TRI, TCE air
emissions from the 2011 TRI were scaled up by a factor of2X.
In NEI, a point source (PS) is a stationary emission source (i.e., sources that remain in one
place). A large facility that houses an industrial process is an example of a point source (EPA,
2004a). A nonpoint source (NPS) refers to a smaller and more diffuse emission source. A
variety of sources are categorized as NPS, including small commercial operations (EPA,
2004a).
Point sources usually include large industrial facilities but they can also include small
commercial facilities, which have traditionally been classified as NPS (EPA, 2008). However,
the choice of whether small commercial facilities are classified as PS or NPS is determined
by the appropriate State, Local, or Tribal air agency (EPA, 2008).
In this assessment, EPA/OPPT assumed that point sources were
representative of large industrial facilities, while nonpoint sources were
assumed to be representative of small commercial facilities.
Data from the 2008 NEI can be used to identify the number of TCE emission sources (Table
2-3) (EPA, 2008). As can be seen from Table 2-3, sixty six percent (66%) of TCE emissions in
2008 were from NPS, while approximately 34 percent were from PS. Thus, nearly two-thirds
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of TCE emissions came from NFS (small facilities). See Appendix E for additional details on
how these data were obtained.
Table 2-3. Number of TCE Emission Sources and Corresponding Total Annual Air
Emissions of TCE as reported in the 2008 NEI (EPA, 2008)
Type of Emission
Source
Number of Sources
Total Annual Air Emissions
(Ibs/yr)
Point source
(Large facility)
180
(degreasing units)
1,480,000
Nonpoint source
(Small facility)
1,779'
2,860,000
Notes:
a Nonpoint sources are aggregated and reported at the county level. Thus, the number of
nonpoint sources (as reported in NEI) will not necessarily correspond to the number of
degreasing units.
Based on the NAICS codes listed in Appendix D, the 2008 and 2011TRI can be queried to
identify stack and fugitive air emissions of TCE (Tables 2-4 and 2-5). Thus, NEI and TRI data
from the same year (2008) can be compared. In 2008, total TCE air emissions in NEI were
approximately two times (2X) greater than those reported in the 2008 TRI8.
Table 2-4. Total Annual Air Emissions of TCE as Reported in the 2008 TRI
(EPA, 2012c)
Type of Emission
Total Annual Air Emissions
(Ibs/yr)
Stack air emissions
1,320,000
Fugitive air emissions
1,230,000
Table 2-5. Total Annual Air Emissions of TCE as Reported in the 2011 TRI
(EPA, 2012c)
Type of Emission
Stack air emissions
Fugitive air emissions
Total Annual Emissions
(Ibs/yr)
850,000
840,000
From the information described above and presented in Tables 2-3 through 2-5, releases of
TCE can be estimated by facility type (large or small). EPA/OPPT's estimate for TCE air
emissions from small commercial degreasing facilities is shown in Table 2-6.
1 Derived by adding the total 2008 NEI emissions (1,480,000 + 2,860,000 = 4,340,000) and dividing this by the total
2008 TRI emissions (1,320,000 + 1,230,000 = 2,550,000), which results in 4,340,000/2,550,000 = 2.
Page 36 of 212
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Table 2-6. EPA/OPPT's Estimated Total Annual Air Emissions of TCE from Small
Commercial Degreasing Facilities
Type of Emission
Air emission
Total Annual Emissions
(Ibs/yr)
2,230,800
1. In this assessment, EPA/OPPT estimated annual air emissions of TCE
from small commercial degreasing facilities as follows:
"Total Annual Emissions'^ 66% * 2 * "Total 2011 TRI TCE Air Emissions'^
= 0.66 *2* (840,000 + 850,000 ) = 2,230,800Ibs
1. TRI is expected to under report air emissions by a factor of 2. Thus, TCE
air emissions from the 2011 TRI were scaled up by a factor of 2.
2. In 2008, small facilities accounted for sixty six percent (66%) of all TCE air
emissions (see Table 2-3).
3. Estimation of facility and operating parameters: For the purposes of this assessment,
small commercial degreasing processes were assumed to operate 260 days per year (yr) for
2 hours (hrs) per day (EPA, 2001a).
Based on NEI data for point source emissions, 154 facilities and 180 degreasing units
reported emissions of TCE in 2008. This translated into about 1.2 degreasing units per
facility. Since these facilities were point sources, they would be considered large industrial
facilities rather than small commercial operations.
In this assessment, EPA/OPPT assumed one degreasing unit per facility for
small commercial operations. This is because smaller facilities were
expected to have less degreasing units per facility than larger one, which
were estimated to have 1.2 degreasing units per facility.
Based on EPA's 2006 risk assessment for the halogenated solvent cleaning source category,
the total number of degreasing facilities was expected to be approximately 1,900 (EPA,
2006d). Thus, the number of small commercial facilities was approximated to be 1,746
(1,900 total facilities minus 154 large industrial facilities). The value of 1,746 total facilities
was used to estimate the number of workers and occupational bystanders potentially
exposed to TCE at small degreasing facilities.
NEI data for point source emissions can also be used to identify the types of degreasing
machines that were being used at large industrial facilities in 2008 (see breakdown by
machine type in Table 2-7). Based on this breakdown, EPA/OPPT estimated that
approximately 90 percent (i.e., 116 out of 129) of batch degreasing machines at large
industrial facilities would likely be open top vapor degreasing machines, while 10 percent
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would be cold solvent degreasing machines. General degreasing units were not included in
this calculation because they could not be categorized; their type was unknown.
Table 2-7. Breakdown of Degreasing Machine Type based on NEI Data for Point
Sources
Type of Degreasing Machine
Open top vapor degreasing
(batch vapor)
Cold solvent cleaning
(batch cold)
Conveyorized vapor degreasing
(in-line)
General degreasing units
(unknown)
Number of Units
116
13
11
40
Total Annual Air
Emissions (lbs/yr)a
890,000
140,000
120,000
330,000
Source: EPA (2008)
Notes:
3 The total is equal to 1,480,000 Ibs as reported for point sources in Table 2-3.
In this assessment, EPA/OPPT assumed that small facilities would use open
top vapor degreasing machines. This is because small commercial facilities
were expected to be comprised entirely of batch cleaning units (EPA, 2006d)
and because open top vapor degreasing machines seem to be most
prevalent (see Table 2-7).
4. Estimation of TCE emissions potentially escaping into the workplace
EPA/OPPT's calculated estimate: Assuming that TCE emissions occur only during the hours
of operation, EPA/OPPT estimated an average emission rate of 19 grams (g) of TCE per
minute (min) from an open top vapor degreasing machine. Since a small facility was
assumed to have only one degreasing unit, this corresponded to 19 g of TCE per min
potentially escaping into the workplace at a small degreasing facility (Table 2-8). This
emission rate was estimated as follows:
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TCE Emission Rate
/Annual TCE air emissionsx / 1
*
VNumber of small facilities/ VOperating days per yr
VOperating hrs per day/ V60 min
/2,230,800 pounds per yr\ /454g\
— I *
1,746 small facilities / Vpound/ V260 days per yr
2 hrs per day
60 min
/
10 gofTCE f ....
19 2 — . — per facility
mm
Estimates reported in the literature: Depending on workplace controls (e.g., local exhaust
ventilation; LEV), average TCE emissions escaping into the workplace from open top
degreasers (i.e., the kind representative of small commercial degreasing operations) can
range from 2.57 to 27.29 g of TCE per min (Wadden etal., 1989).
Based on another source, operating emissions from batch cleaning machines can range
from 5 to 10 gofTCE per min (EPA, 2001a). In addition, EPA's overall emission limit for
implementing the National Emissions Standards for Hazardous Air Pollutants (NESHAP) is
150 kilograms (kg) per square meter (m2) per month (EPA, 2004b). This translates into an
emission rate ranging from 16 to 50 g of TCE per min (Table 2-8). Please also refer to
Appendix E for additional details regarding these emission rates.
In comparing EPA/OPPT's estimated TCE emission rate (based on NEI and
TRI data) with values reported in the literature, it is evident that these
estimates overlap and are of the same order of magnitude. Please refer to
Table 2-8 for a summary of these emission rates.
Table 2-8. TCE Emissions Potentially Escaping into the Workplace at Small Commercial
Degreasing Facilities (Calculated and Reported Values)
Type of Facility
EPA/OPPT's Calculated
Estimate (g TCE/min)
Estimates Reported
in the Literature (g TCE/in)
Small degreasing facility
19
2.57 to 27.29 '
5 to 10 b
16 to 50 c
Sources:
3 Wadden etal. (1989)
EPA (2001a)
EPA(2004b)
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5. Estimation of potential workplace exposures
Workers and occupational bystanders potentially exposed: The National Occupational
Exposure Survey (NOES), conducted by the National Institute for Occupational Safety and
Health (NIOSH) from 1981 to 1983, estimated that 401,000 workers employed at 23,225
plant sites were potentially exposed to TCE in the U.S. This translates into about 17 workers
per facility (ATSDR, 1997). EPA/OPPT assumed that this estimate represented both workers
and occupational bystanders at degreasing facilities. EPA/OPPT estimated the average
number of workers directly involved with solvent cleaning operations as five workers per
facility (EPA, 2001a). The number of occupational bystanders potentially exposed was
estimated as 12 per site (i.e., 17 - 5 = 12). Please refer to Appendix F for more information.
In this assessment, EPA/OPPT estimated the population of workers and
occupational bystanders potentially exposed to TCE at small degreasing facilities
as follows:
1. Number ofsmall facilities = 1,746
2. Numberof Workers Potentially Exposed = (5 workers/facility)* 1,779 facilities = 8,730
3. Number of Occupational Bystanders Potentially Exposed
= (12 occupational bystanders per facility) *1,746 facilities = 20,952
EPA/OPPT's estimated workplace exposures: To estimate workplace exposures, the TCE
emissions from Table 2-8 were incorporated into a Two-Zone Near Field/Far Field (NF/FF)
mass balance model. The NF/FF model has been extensively peer-reviewed, it is extensively
used, and results of the model have been compared with measured data. The comparison
indicated that the model and measured values agreed to within a factor of about three
(Javiocketal.. 2011).
Appendix G presents the rationale and calculations for estimating inhalation exposures for
workers and occupational bystanders at small degreasing facilities. An overview is also
provided in Appendix F. Table 2-9 shows a summary of EPA/OPPT's workplace exposure
estimates for workers and occupational bystanders at small commercial degreasing
facilities. Exposure estimates were calculated for work conditions with and without local
exhaust ventilation (LEV). In this assessment, EPA/OPPT assumed an effectiveness of 90%
for LEV (Wadden et al., 1989). Thus, LEV could reduce emissions escaping into the
workplace by 90%.
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Table 2-9. Summary of Potential Workplace TCE Inhalation Exposures at Small
Commercial Degreasing Facilities based on the NF/FF Model
Population
Exposed
Workers1
Occupational
bystanders2
Exposure Type
(Duration)
Inhalation
exposure
(8-hr TWA3)
Inhalation
exposure
(8-hr TWA)
Estimated
TCE Air
Concentration
(ppm)
With LEV4
0.3
(low-end estimate)
20
(upper-end estimate)
0.04
(low-end estimate)
17
(upper-end estimate)
Estimated
TCE Air
Concentration (ppm)
No LEV
3
(low-end estimate)
197
(upper-end estimate)
0.4
(low-end estimate)
172
(upper-end estimate)
Estimated
Number
of People
Exposed
8,730
20,952
Notes:
1 Workers are directly involved with degreasing operations.
2 Occupational bystanders have the potential to be exposed to TCE but they are not directly involved
with degreasing operations
3 TWA = Time Weighted Average (NOTE: since degreasing facilities are expected to operate 2 hrs per
day, to estimate 8-hr TWA exposures, EPA/OPPT assumed no exposure for 6 hrs per day).
4 LEV = local exhaust ventilation
Workplace exposures based on monitoring data: Although a nationally representative
sample of relevant exposure monitoring data were not available, monitoring data from
OSHA were identified (Coble, 2013)9. These data from 2003 to 2010 were specific to TCE.
Also, the data represented time-weighted average (TWA) personal breathing zone
measurements relevant to the NAICS codes listed in Appendix D. In Figure 2-2, EPA/OPPT's
workplace exposure estimates (Table 2-9) are compared with OSHA monitoring data.
It is evident that estimated and measured exposures are of the same order of
magnitude; EPA/OPPT's estimated exposure range captures approximately 95%
of OSHA field measurements. These results substantiate the suitability of using
the NF/FF mass balance model for the purposes of estimating potential
workplace exposures to TCE at small commercial degreasing facilities.
Acute and chronic workplace exposure estimates: In this assessment, EPA/OPPT assessed
acute and chronic risks for workers and occupational bystanders. A summary of acute and
OSHA monitoring data and analysis supporting Figure 2-2 can be found in the supplementary file "OSHA IMIS TCE
SAMPLES 062314vl.xlsx"
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chronic workplace exposures is provided in Table 2-10. Acute workplace exposures were
estimated as follows:
AC =
C*ED
AT
where:
AC
C
ED
AT
= acute concentration (24 hrTWA in ppm)
= contaminant concentration in air (8 hr TWA in ppm; from Table 2-9)
= exposure duration (8 hrs/day)
= averaging time (24 hrs/day)
EPA/OPPT used the average daily concentration (ADC) and lifetime average daily concentration
(LADC) to estimate workplace exposures for non-cancer and cancer risks, respectively. These
exposures were estimated as follows:
ADC or LADC =
C*ED*EF*WY
AT
where:
ADC = average daily concentration (24-hr TWA in ppm) used for chronic non-cancer
risk calculations
LADC = lifetime average daily concentration (24-hr TWA in ppm) used for chronic
cancer risk calculations
C = contaminant concentration in air (8-hr TWA in ppm; from Table 2-9)
ED = exposure duration (8 hrs/day)
EF = exposure frequency (260 days/yr)
WY = working yrs per lifetime (40 yrs)
AT = averaging time (LT x 365 days/yr x 24 hrs/day; where LT = lifetime; LT = 40
yrs for non-cancer risks; LT=70 yrs for cancer risks)
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Figure 2-2. Comparison of EPA/OPPT's Workplace TCE Exposure Estimates with 2003-2010
OSHA Monitoring Data for Small Commercial Degreasing Facilities
OSHA Monitoring Data
(39 samples)
380 ppm High End
EPA Estimates
170 ppm
~95thpercentile
45 ppm average
17 ppm median
0.06 ppm Low End
197 ppm High End
i 0.04 ppm Low End
OSHA sent EPA/OPPT potentially relevant field measurements for TCE. The OSHA
monitoring data and EPA/OPPT's assessment of it are included in the supplementary file
"OSHA IMIS TCE SAMPLES_062314vl.xlsx". In brief, EPA/OPPT filtered OSHA's data to
arrive at 39 relevant field measurements collected between 2003 and 2010. The data
were filtered by (1) the NAICS codes listed in Appendix D, (2) Sample Type (personal), and
(3) Exposure Type (full shift time weighted average).
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Table 2-10. Summary of Acute and Chronic Workplace TCE Inhalation Exposures at
Small Commercial Degreasing Facilities
Population
Exposed
Workers1
Occupational
bystanders2
Workplace Exposure Concentrations
(24-hr TWA3 in ppm)
Acute
With LEV
0.1
(low-end
estimate)
7
(upper-end
estimate)
0.01
(low-end
estimate)
6
(upper-end
estimate)
No LEV
1
(low-end
estimate)
66
(upper-end
estimate
0.1
(low-end
estimate)
57
(upper-end
estimate)
Chronic
Non-cancer
With LEV
0.07
(low-end
estimate)
5
(upper-end
estimate
0.01
(low-end
estimate)
4
(upper-end
estimate)
No LEV
0.7
(low-end
estimate)
47
(upper-end
estimate
0.1
(low-end
estimate)
41
(upper-end
estimate)
Cancer
With LEV
0.04
(low-end
estimate)
3
(upper-end
estimate
0.005
(low-end
estimate)
2
(upper-end
estimate)
No LEV
0.4
(low-end
estimate)
27
(upper-end
estimate
0.05
(low-end
estimate)
23
(upper-end
estimate)
Notes:
Workers are directly involved with degreasing operations.
2 Occupational bystanders have the potential to be exposed to TCE but they are not directly involved
with degreasing operations
3 TWA = Time weighted average
4 LEV = local exhaust ventilation
The exposure estimates in Table 2-10 are used for deriving MOEs in Tables 2-30 and 2-33, and Figure 2-6*
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2.4 ENVIRONMENTAL RELEASES AND OCCUPATIONAL EXPOSURE
ESTIMATES FOR SPOT CLEANING AT DRY CLEANING
OPERATIONS
2.4.1 A Brief Summary of Spot Cleaning at Dry Cleaning Operations
The dry cleaning industry provides garment cleaning services and consists of the following three
basic functions: cleaning, drying, and finishing (EPA, 1995a; Luhringand Marks, 2000; NIOSH,
1997a). Releases from dry cleaning operations are primarily to air, water, and solid waste (EPA,
1995a).
1. Prior to being machine washed (cleaning), garments are typically pre-treated for stains (spot
cleaning).
2. Next, garments are dried using a combination of aeration, heat and tumbling.
3. If applicable, garments are post-treated for any remaining stains (spot cleaning), and then
they are pressed (finishing).
2.4.2 EPA/OPPT's Release and Exposure Assessment for Spot Cleaning
Operations
EPA/OPPT used readily available information from a 2007 study on spotting chemicals
(CalEPA/EPA, 2007), prepared for the California EPA and U.S. EPA Region 9, to estimate releases
of and exposures to TCE from spot cleaning operations at dry cleaning facilities. EPA/OPPT used
several steps in order to produce these estimates. An overview of these steps is shown in Figure
2-3; further elaboration of these steps has been structured to follow the sequence in this figure.
1. Identification of relevant industries: Based on information provided to EPA/OPPT during
the review and comment period for this TSCA Work Plan Chemical risk assessment, the dry
cleaning industry was identified as using TCE in spot cleaning processes. In order to assess
TCE-based spot cleaning processes, EPA/OPPT used data from a study that were specifically
focused on spotting chemicals (CalEPA/EPA, 2007). This study focused on dry cleaning
facilities in the state of California.
In this assessment, EPA/OPPT assumed that dry cleaning facilities in the state of
California were representative of dry cleaning facilities in the United States.
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Figure 2-3. Overview of EPA/OPPT's occupational release and exposure assessment
for spot cleaning operations
Identify relevant industries:
i. The use of TCE as a spot cleaner was identified
for the dry cleaning industry.
2. Dry cleaning facilities in the state of California
were assumed to be representative of dry
cleaning facilities in the United States.
Estimate how much TCE is used for spot
cleaning at a typical dry cleaning facility:
1. Data from a 2007 study specific to spot
cleaning at dry cleaning facilities in the state of
California was used to estimate how much TCE
is used for spot cleaning on a per facility basis.
Estimate operating parameters:
1. Data from the U.S. Bureau of Labor Statistics
(BLS) was used to estimate days and hours of
operation per year.
Estimate TCE emissions potentially
escaping into the workplace
Estimate potential workplace
exposures
1. Use a Near-Field / Far-Field mass
balance model
2. Estimate exposures for workers and
occupational non-users
2. Estimation of how much TCE is used for spot cleaning at a typical dry cleaning facility: In
2007, the number of textile cleaning facilities in the state of California was estimated to be
about 5,000 (CalEPA/EPA, 2007). Further, in consultation with cleaning facilities and their
suppliers, it was estimated that about 42,000 gallons per year of TCE-based spotting agents
were sold in the state of California (CalEPA/EPA, 2007). Based on an examination of several
material safety data sheets (MSDS), the concentration of TCE in spotting agents was
observed to vary from 10 percent to 100 percent (CalEPA/EPA, 2007). In this assessment,
EPA/OPPT estimated that a typical dry cleaning facility used 0.84 to 8.4 gallons per yr of TCE
for spot cleaning operations. This was estimated as follows:
TCE Use Rate
(Annual Usage of Spotting Agents) * (TCE Concentration in Spotting Agent)
Number of Facilities
(42,000 gallons per yr)*(10% to 100%)
5,000 facilities
= 0.84 to 8.4
per facility
3. Estimation of operating parameters: EPA/OPPT estimated that dry cleaning facilities
operated 260 days per year; 8 hrs a day for a total of 2,080 hrs per year (BLS, 2012). In
addition, EPA/OPPT assumed that spot cleaning operations take place over the course of an
8 hr work day.
Page 46 of 212
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4. Estimation of TCE emissions potentially escaping into the workplace: Chemical emissions
escaping into the workplace can be controlled and thus reduced through the use of
engineering controls. For example, tables used to perform spot cleaning (spotting tables)
can be equipped with LEV and are commercially available (NIOSH, 1997b). In this
assessment, EPA/OPPT assumed an effectiveness of 90% for LEV (Wadden et al., 1989).
Thus, LEV could reduce emissions escaping into the workplace by 90%.
In this assessment, EPA/OPPT assumed that the entire amount of spotting agent
used at a dry cleaning facility was available for evaporation and thus could be
emitted into the workplace. EPA/OPPT assessed scenarios with and without LEV.
A summary of TCE emissions potentially escaping into the workplace from spot cleaning is
provided in Table 2-11. These emission rates were estimated as follows:
TCE Emission Rate (with LEV)
= (TCE Use Rate) * (3'785cm3] * (TCE Density) * ( lyr ) * (-^-) *
v J \ gallon ) ^ JJ \2,080hrsJ \60minJ
r-irmn/ T nr rff <-• ^ f 0.84 to 8.4 gallons \ /3,785 cm3\ /1.46 g\
(100% - LEV Effectiveness) = * * —f- *
\ yr—facility } \ gallon } \ cm6 }
lyr
yr —facility
0.0037 to 0.037 g TCE
mm
gallon
per facility
TCE Emission Rate (no LEV)
* (TCE Density) *
^ JJ
2,080
= (TCE Use Rate) *
^ J
/0.84to8.4gallonsX # /3.78S cm^ # /i.46g\ # / lyr \ # /_ihi_\ =
V yr-facility / V gallon / V cm3 / \2,080 hrs/ \60 min/
60 min
yr-facility
0.037 toO.37 g TCE
mm
gallon
per facility
Table 2-11. TCE Emissions Potentially Escaping into the Workplace from Spot Cleaning
Type of Facility
TCE Emissions Potentially Escaping
into the Workplace
(gTCE/min)
With LEV1
TCE Emissions Potentially Escaping
into the Workplace
(gTCE/min)
No LEV
Dry Cleaning Facility
0.0037 (low-end estimate)
0.037 (upper-end estimate)
0.037 (low-end estimate)
0.37 (upper-end estimate)
1 LEV = local exhaust ventilation
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5. Estimation of potential workplace exposures
Workers and occupational bystander potentially exposed: There are approximately 36,000
dry cleaning facilities (NAICS code 8123000) in the U.S. with a total of about 300,000
workers (USCB, 2011); or about 8 workers per dry cleaning facility. This estimate includes
both, workers and occupational bystanders at dry cleaning facilities. Approximately 10% of
employees at a dry cleaning facility can be in close proximity to the spot cleaning process
(NIOSH. 1997a).
In this assessment, EPA/OPPT assumed one (1) worker per dry cleaning facility
was directly involved with spot cleaning operations while the number of
occupational bystanders was estimated as seven (7) per dry cleaning facility.
EPA/OPPT's estimated workplace exposures: To estimate workplace exposures, the
emissions from Table 2-11 were incorporated into the NF/FF mass balance model. As
indicated previously, the NF/FF model has been extensively peer-reviewed, it is extensively
used, and results of the model have been compared with measured data. The comparison
indicated that the model and measured values agreed to within a factor of about three
(Javiocketal.. 2011).
Appendix H presents the rationale and calculations for estimating inhalation exposures for
workers and bystanders from spot cleaning at dry cleaning facilities. Table 2-12 shows a
summary of EPA/OPPT's workplace exposure estimates for workers and occupational
bystanders at dry cleaning facilities. Exposure estimates were calculated for work conditions
with and without LEV.
Workplace exposures based on monitoring data: Although relevant exposure monitoring
data were limited, EPA/OPPT did identify a study specific to spot cleaning with TCE (NIOSH,
1997a). In this study, TWA exposures to TCE during spot cleaning (with no LEV) ranged from
2.37 to 3.11 ppm, which is within EPA/OPPT's estimated range of 0.08 to 19 ppm (Table 2-
12).
The facility in this study had a floor area of approximately 8,500 square feet (ft2). The room
height was not specified. It was assumed to be 10 feet (ft), giving a room volume of 85,000
cubic feet (ft3) [(or about 2,400 cubic meter (m3)]. On an average day, workers at this facility
used approximately 6 ounces (170 g) of Picrin (NIOSH, 1997a), the contents of which are
approximately 100% TCE (CalEPA/EPA, 2007). Assuming an 8-hr work day, this would
correspond to a Picrin use rate of about 0.35 g of TCE per min. Based on these site-specific
parameters (e.g., room volume, Picrin use rate, no LEV), the NF/FF model estimated that
the TWA TCE worker exposures could range from 0.8 to 2.1 ppm, while measured values
reported in the study ranged from 2.37 to 3.11 ppm
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Table 2-12. Summary of Potential Workplace TCE Inhalation Exposures from
Spot Cleaning at Dry Cleaning Facilities
Population
Exposed
Worker1
Occupational
bystanders2
Type of Exposure
(Exposure
Duration)
Inhalation
exposure
(8-hr TWA3)
Inhalation
exposure
(8-hr TWA)
Estimated
TCE Air
Concentration
(ppm)
With LEV4
0.008
(low-end
estimate)
2
(upper-end
estimate)
0.0007
(low-end
estimate)
2
(upper-end
estimate)
Estimated
TCE Air
Concentration
(ppm)
No LEV
0.08
(low-end
estimate)
19
(upper-end
estimate)
0.007
(low-end
estimate)
18
(upper-end
estimate)
Estimated
Number of
People
Exposed5
36,000
252,000
Notes:
Workers are directly using TCE-based spot cleaners.
2 Occupational bystanders have the potential to be exposed to TCE but they are not directly
involved with spot cleaning.
3 TWA = Time weighted average
4 LEV = local exhaust ventilation
5 The number of people does not sum up to 300,000 due to rounding error.
Since a study specific to spot cleaning with TCE was identified, site-specific
parameters from this study were incorporated into the NF/FF model to obtain
site-specific model estimates of worker exposure. Model estimates (0.8 to 2.1
ppm) were of the same order of magnitude as measured values (2.37 to 3.11
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Acute and chronic workplace exposure estimates: In this assessment, EPA/OPPT assessed
acute and chronic risks for workers and occupational bystanders from spot cleaning at dry
cleaning facilities. A summary of acute and chronic workplace exposures is provided in Table
2-13. Acute workplace exposures were estimated as follows:
C*ED
where:
AC = acute concentration (24-hr TWA in ppm)
C = contaminant concentration in air (8-hr TWA in ppm; from Table 2-12)
ED = exposure duration (8 hrs/day)
AT = averaging time (24 hrs/day)
EPA/OPPT used the average daily concentration (ADC) and lifetime average daily
concentration (LADC) to estimate workplace exposures for non-cancer and cancer chronic
risks, respectively. These exposures were estimated as follows:
C * ED * EF * WY
LADC =
where:
LADC = lifetime average daily concentration (24-hr TWA in ppm)
C = contaminant concentration in air (8-hr TWA in ppm; from Table 2-12)
ED = exposure duration (8 hrs/day)
EF = exposure frequency (260 days/yr)
WY = working years per lifetime (40 yrs)
AT = averaging time (LT x 365 days/yrs x 24 hrs/day; where LT = lifetime; LT = 40
yrs for non-cancer risks; LT=70 yrs for cancer risks)
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Table 2-13. Summary of Acute and Chronic Workplace TCE Inhalation Exposures from
Spot Cleaning at Dry Cleaning Facilities
Population
Exposed
Workers1
Occupational
bystanders2
Workplace Exposure Concentrations
(24-hr TWA3 in ppm)
ArutP
r^tw ic
With LEV 4
0.003
(low-end
estimate)
1
(upper-end
estimate)
0.0002
(low-end
estimate)
(upper-end
estimate
No LEV
0.03
(low-end
estimate)
6
(upper-end
estimate)
0.002
(low-end
estimate)
(upper-end
estimate
Chronic
Non-cancer
With LEV
0.002
(low-end
estimate)
0.5
(upper-end
estimate)
0.0002
(low-end
estimate)
Or
.5
(upper-end
estimate
No LEV
0.02
(low-end
estimate)
5
(upper-end
estimate)
0.002
(low-end
estimate)
(upper-end
estimate
Cancer
With LEV
0.001
(low-end
estimate)
0.3
(upper-end
estimate)
0.0001
(low-end
estimate)
OT
.3
(upper-end
estimate
No LEV
0.01
(low-end
estimate)
3
(upper-end
estimate)
0.001
(low-end
estimate)
(upper-end
estimate
Notes:
1Workers are directly involved with spot cleaning operations.
2 Occupational bystanders have the potential to be exposed to TCE but they are not directly involved with
spot cleaning operations
3 TWA = Time weighted average
4 LEV = local exhaust ventilation
*The exposure estimates in Table 2-13 are used for deriving MOEs in Tables 2-31 and 2-34, and Figure 2-7*
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2.5 CONSUMER EXPOSURES - DEGREASER AND ARTS/CRAFTS USES OF
TCE
2.5.1 TCE Uses Targeted for Consumer Exposure Assessment
Among the many uses of TCE in consumer products, EPA/OPPT selected those products used as
degreasers and arts/crafts products for further risk evaluation. The decision of targeting the
assessment to specific consumer products considered (1) consumer product information reported in
the National Institutes of Health's (NIH) Household Products Database (DHHS. 2012). (2) information
reported in Material Safety Data Sheets (MSDS), and (3) product information on the manufacturer's
website.
EPA/OPPT searched the NIH Household Products Database, which links over 13,000 consumer brands
to health effects reported in MSDS documents (DHHS, 2012). The database also allows scientists and
consumers to research products based on chemical ingredients. Our search found three companies
manufacturing 12 consumer products containing TCE (Table 2-14)10.
EPA/OPPT further researched the products, including a review of the MSDS for each product and
inspection of the manufacturer's website. EPA/OPPT confirmed the presence of TCE in six out of the 12
products listed by the NIH Household Products Database (Table 2-14). These six products were
degreaser and arts and crafts aerosol spray products that may be used by consumers at home. Three
products were absent from the product manufacturer's website and discontinued. The remaining three
products in Table 2-14 did not contain TCE and were reformulated by using other chemical alternatives
such as tetrachloroethylene, hydrotreated light distillate, dipropylene glycol n-propyl ether,
dipropylene glycol methyl ether acetate, aliphatic petroleum solvent, or acetone. There may be other
consumer products currently present in the U.S. market that were not reported in the NIH Household
Products Database.
After peer review EPA/OPPT became aware of more products that contain TCE that were not included
in the Households product database. Another product is sold by Sprayway and is called "No Fray
Spray". It can be used to reduce fraying of fabric in sewing and craft projects. It has the same
percentage of TCE as the clear protective coating spray (20-30% TCE) listed in the table above and it
may result in similar exposures for users.
EPA/OPPT also searched the internet for information about TCE-containing spot cleaners used for
consumer uses and identified several spot cleaners for fabrics marketed to U.S. consumers. According
to their MSDS, most of these spot cleaners did not contain TCE and the other spot cleaner products did
not report the list of ingredients. Thus, we could not preclude consumer exposures if some of these
spot cleaners may contain TCE as a main ingredient. Also, OPPT/EPA could not rule out that consumers
are using professional-grade spot cleaners.
10 See supplementary document entitled: "Supplemental Product Information for the TSCA Workplan Chemical Risk
Assessment of TCE (External Review Draft)" for further details on the results of the Household Products Database
retrieval and for the Material Safety Data Sheets and other information retrieved from company websites.
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Table 2-14. TCE Products in NIH's Household Products Database
Product3
%TCE Content by Weight
(as of February 2014)b
EPA/OPPT determined that these products contain TCE
Sprayway C-60 063 solvent degreaser
Sprayway C-60 064 solvent degreaser
Sprayway 201 clear protective coating spray
Sprayway 205 film cleaner
Sprayway 208 toner aide
Sprayway 209 mirror edge sealant
>90%
>90%
20-30%
>90%
15-20%
20-30%
EPA/OPPT determined that these products did not contain TCE
Lectra Clean 02018 heavy duty
electrical parts degreaser
Lectra Clean 02120 Lectra Clean II non-chlorinated
heavy duty electrical parts degreaser
Sprayway 073 brake parts cleaner
None
None
None
EPA/OPPT determined that these products were discontinued
Sprayway 669 gravel guard
Sprayway 732 industrial cleanup dry cleaner
TrakAuto Trouble Free Rust Buster
Product not found on
manufacturer's product list
Product not found on
manufacturer's product list
Company no longer in business
Notes:
a EPA/OPPT searched the NIH Household Products Database in March 2012 and February 2014. Both
searches reported the same list of consumer products containing TCE.However, one product (Sprayway
Plastic Spray Clear Fixative No. 201) was categorized for arts/crafts uses in March 2012, but appears to be
used as a home office product in February 2014. For the purposes of this assessment, EPA/OPPT relied on
the searches obtained in March 2012.
Percent TCE according to Material Safety Data Sheets retrieved from company websites in March, 2012
(See attached document titled "Supplemental Product Information for the TSCA Workplan Chemical Risk
Assessment of TCE (External Review Draft).pdf").
2.5.2 Overview of the E-FAST2/CEM Model
The Exposure and Fate Assessment Screening Tool Version 2 (E-FAST2) Consumer Exposure Module
(CEM) (EPA, 2007b) was selected for the consumer exposure modeling because it is the appropriate
model to use due to the lack of available emissions and monitoring data for the TCE uses under
consideration. Moreover, EPA/OPPT did not have the input parameter data required to run more
complex indoor air models for the consumer products under the scope of this assessment.
E-FAST2/CEM uses high-end input parameters/assumptions to generate conservative, upper-bound
inhalation exposure estimates for aerosol spray products. The advantages of E-FAST2/CEM are the
following:
1. CEM model was peer-reviewed in 1999.
2. Accommodates the inputs available for the products containing TCE in the indoor air model.
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3. Uses the same calculation engine to compute indoor air concentrations from a source as the Multi-
Chamber Concentration and Exposure Model (MCCEM), but it does not require measured emission
values (e.g. chamber studies).
The model used a two-zone representation of a house to calculate the potential acute dose rate
(mg/kg-bw/day) of TCE for consumers and bystanders. Zone 1 represents the area where the consumer
is using the product, whereas Zone 2 represents the remainder of the house. Zone 2 was used for
modeling passive exposure to house residents (bystanders), such as children, adults, pregnant women
and the elderly.
The general steps of the calculation engine within the CEM model included:
1. introduction of the chemical (i.e., TCE) into the room of use,
2. transfer of the chemical to the rest of the house due to exchange of air between the different
rooms,
3. exchange of the house air with outdoor air and,
4. summing of the exposure doses as the modeled occupant moves about the house
The chemical of concern (i.e., TCE) entered the room air through two pathways: (1) overspray of the
product and (2) evaporation from a thin film. One percent (1%) of the product was assumed to become
instantly aerosolized (i.e. product overspray) and was available in the room air for inhalation.
The CEM model used data from the evaporation of a chemical film to calculate the rate of the mass
evaporating from the application surface covered during product use (Chinn, 1981). The model
assumed that air exchanged from the room of use (Zone 1) and the rest of the house (zone 2)
according to interzonal flow. The model also allowed air exchange from the house (Zone 1 & 2) with
the outdoor air.
EPA/OPPT used the default activity pattern in CEM based on the occupant being present in the home
for most of the day. As the occupants moved around the house in the model, their exposure to the
calculated air concentrations were summed to form a potential 24-hr dose.
The potential inhalation acute dose rates (ADRpot) were computed iteratively by calculating the peak
concentrations for each simulated 10-second interval and then summing the doses over 24 hrs. These
calculations took into consideration the chemical emission rate over time, the volume of the house and
the zone of use, the air exchange rate and interzonal airflow rate, the exposed individual's locations,
body weights and inhalation rates during and after the product use (EPA, 2007b). The reader is
referred to EPA's E-FAST2 website11 and Appendix I to obtain additional information about the model,
including the model documentation and algorithms used.
EPA's E-FAST2 website: http://www.epa.gov/oppt/exposure/pubs/efast.htm
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2.5.3 Consumer Model Scenarios and Input Parameters for Indoor Exposure
to Specific TCE Uses
Table 2-15 describes the four acute inhalation indoor scenarios and populations of interest that
EPA/OPPT evaluated in the consumer exposure assessment. As indicated in section 1.3.2, EPA/OPPT
believes that inhalation is the main exposure pathway. Ingestion exposure of TCE from the use of these
consumer products appears to be unlikely given the way they are used (i.e., sprayed onto artwork).
Table 2-15. Consumer Model Scenarios and Populations of Interest
Acute Inhalation Indoor Scenario
Population of Interest3
Consumer-degreaser use
Adult consumers >16 yrs old
Most sensitive population of concern: Pregnant
women (fetus)
Bystander to consumer-degreaser use
Individuals of all ages
Most sensitive population of concern:
Pregnant women (fetus) and children
Consumer-clear protecting coating spray use
Adult consumers >16 yrs old
Most sensitive population of concern: Pregnant
women (fetus)
Bystander to consumer-clear protecting coating
spray use
Individuals of all ages
Most sensitive population of concern:
Pregnant women (fetus) and children
Notes:
3 EPA/OPPT believes that the users of these products are generally adults, but young teenagers and even
younger children may be users or be in the same room with the user while engaging in arts and crafts projects
or degreasing. Since there are not survey data for consumer behavior patterns or a way to create varying
behavior patterns for different age groups, the indoor air concentrations shown in table 2-17 could be extended
to all users.
To estimate exposures to these products, numerous input parameters are required to generate a single
exposure estimate. These parameters include the characteristics of the house, the behavior of the
consumer and the emission rate of the chemical into the room of use. In the absence of measured
values for many of the needed inputs, the E-FAST2/CEM modeling for TCE used a combination of upper
percentile and mean or median input parameters and assumptions in the calculation of potential
exposure for the user and bystanders. This approach produced high-end acute inhalation estimates
instead of central tendency exposures12 that were hypothetical and intended to be conservative. The
input parameters and assumptions are summarized in Table 2-16 and explained more fully in the
Supplemental Information13'14 and Appendix I.
" High-end exposures represent values above a mean or median and may include the high end of an exposure distribution.
A central tendency value represents some measure of the center of a distribution, such as an average or mean or median.
' See attached document titled "Supplemental Information on E-FAST2 CEM Outputs (Degreaser Use) TSCA Work Plan
Chemical Risk Assessment of TCE (External Review Draft).docx").
1 See attached document titled "Supplemental Information on E-FAST2 CEM Outputs (Clear Protective Coating Spray) TSCA
Work Plan Chemical Risk Assessment of TCE (External Review Draft).docx"
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Consumer behavior pattern parameters in CEM include the mass of product used, the duration of use
and the frequency of use.The default values in CEM for these consumer behavior parameters are set to
high end values.The other parameters (e.g. house volume) in CEM are set to mean or median values
from the literature.The default consumer behavior patterns from CEM were not used in this risk
assessment. This combination of high end and mean or median values is intended to produce a high
end acute inhalation exposure estimate.
EPA/OPPT did not locate consumer product survey data for use patterns for the two consumer uses.
Instead professional judgment was used for these values and was based on the MSDS and product
descriptions.
To determine the appropriateness of the consumer behavior pattern parameters chosen in this risk
assessment, EPA/OPPT examined the consumer categories available in the Westat survey (EPA, 1987a).
The Westat survey contacted thousands of Americans to gather information on consumer behavior
patterns related to product categories that may contain halogenated solvents. The Westat survey data
were not completely aligned with the description of the products that we used in this consumer
exposure assessment (EPA, 1987a). However the data provided some indication that the values that
EPA/OPPT used were below mean or median for the mass, but above the mean for the time spent in
the room of use (Appendix I).
The input parameters for household characteristics (e.g., house volume) were all set to mean or
median values based on data found in the available literature. Likewise, the user's body weight and
inhalation rate were set to either the mean or the median values for the simulations used in this
assessment.
The air exchange rate in the room of use is not reflective of open windows or the use of an exhaust fan.
While it is possible that some users may employ these exposure reduction techniques inside their
homes, the goal of the consumer exposure assessment was to provide an acute exposure estimate for
ventilation conditions representing average household air exchange rates. Moreover, residential users
would not necessarily have the type of indoor exposure reduction tools/equipment (e.g., gloves,
exhaust ventilation) that workers likely have at occupational settings. Consumers would not necessarily
be as aware of potential chemical hazards as workers and would not have a standard operating
procedure in place to assure that they use exposure reduction techniques each time they use a
product.
In this assessment it was assumed that there was no pre-existing concentration of TCE in the home
before product use began. The outdoor air was also assumed to be free of TCE, meaning that the air
exchange rate described the intake of air with no TCE contamination.
The products were assumed to be sprayed on varying surfaces, e.g. a metallic surface for the degreaser
or canvas for the spray fixative. On these surfaces, a thin film of the product was assumed to build up,
which then evaporates and contributes to the air concentration of the chemical in the room.
We relied on modeled emission rates because data from chamber studies were not available. To
generate emission rates, E-FAST2/CEM used empirical data from studies assessing the emission rates
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of pure solvents (Chirm, 1981). E-FAST2/CEM used the Chirm study as surrogate data to calculate the
rate of evaporation of TCE from this unknown surface to the air in the home.
These solvent studies supported the use of an exponentially decaying emission rate for TCE from the
application surface based on vapor pressure and molecular weight (Chinn, 1981), the equations using
the Chinn method are in Appendix I. The solvent degreaser application should be well modeled by the
Chinn study since the degreaser product was over 90% TCE. Also, carbon dioxide (C02) would be the
only other component in the product and it would not be expected to be deposited on the surface or
cause significant changes in the TCE emission rate from the surface.
On the other hand, the spray fixative product is a more complicated mixture, and the interaction of
these chemicals could alter the evaporation rate of TCE. This introduces uncertainty into the
assessment but EPA/OPPT could not find a better data set available to model the emission rates.
Within the current exposure assessment, the 24-hr exposure was not strongly dependent on the
emission rate due to the amount of time the product user spends in the room of use (see Appendix I
for details).
CEM has certain restrictions on the age that is assumed for simulated users, which in turn sets limits
for the dose rates generated for different age groups. However, these restrictions should not be
interpreted as suggesting that younger users would not be exposed. EPA/OPPT believes that the users
of these products are generally adults, but teenagers and even younger children may be users or be in
the same room with the user while engaging in arts and crafts projects or degreasing. Since there are
not survey data for consumer behavior patterns or a way to create varying behavior patterns for
different age groups, the indoor air concentrations shown in table 2-17 could be extended to all users.
Lastly, a chronic consumer exposure assessment was not performed because the frequency of product
used was considered to be too low to create chronic risk concerns. Although CEM model results given
in the supplemental information included chronic exposure estimates, they were not used in this
assessment.
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Table 2-16. Summary of E-FAST2/CEM's Input Parameters and Assumptions for Estimation of Potential Acute Dose Rates for
Specific TCE Uses
Input Parameters and
Assumptions
TCE mass in product
Consumer Product TCE
Weight Fraction
Duration of event and
occupant behavior pattern
(hrs/event)
Frequency of use
(events/day)
Exposure duration (days)
Consumer-
degreaser use
24 g per use
0.9
Value from MSDS
1 hr using product
2 hrs total spent in
room of use
24 hrs in home or
outside
Bystander to
consumer-degreaser
use
N/A
N/A
0 hr using product
0 hr in room of use
24 hrs in home or
outside
Consumer-clear
protecting coating
spray use
11 g per use
0.3
High end values
from MSDS
0.5 hr using
product
2 hrs total spent in
room of use
24 hrs in home or
outside
Bystander to
consumer-clear
protecting
coating spray use
N/A
N/A
0 hr using product
0 hr in room of use
24 hrs in home or
outside
1
1
Comments
Total mass of product
multiplied by weight
fraction of TCE
Value reported in the MSDS.
The behavior patterns for the
user and bystander are based
on a day spent mostly at
home with 3 hrs outside of
the home. Also, model does
not allow bystanders to be
present in the room of use
during use of the product.
It was assumed that these
exposures represent unique
and separate acute exposure
events for both users and
bystanders. This was based
primarily on the short half-life
of TCE in humans. Some
residual of TCE (or
metabolite[s]) would be
expected if another exposure
occurs the next day. However,
EPA/OPPT did not assume
that the residual would be
substantial or build up based
upon pharmacokinetic half-
life.
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Table 2-16. Summary of E-FAST2/CEM's Input Parameters and Assumptions for Estimation of Potential Acute Dose Rates for
Specific TCE Uses
Input Parameters and
Assumptions
Consumer-
degreaser use
Bystander to
consumer-degreaser
use
Consumer-clear
protecting coating
spray use
Bystander to
consumer-clear
protecting
coating spray use
Comments
Age groups
(yrs)
Teenager/Young
adults: 16-20 yrs
Adults: 21-78 yrs
Any age
(<1 yr; 1-2 yr; 3-5 yr;
6-10 yr; 11-15 yr; 16-
20 yrs; and 21-78 yr)
Teenager/Young
adults: 16-20 yrs
Adults: 21-78 yrs
Any age
(<1 yr; 1-2 yr; 3-5 yr; 6-
10 yr; 11-15 yr; 16-20
yrs; and 21-78 yr)
Age groups obtained from
EPA's Exposure Factors
Handbook (EFH) (EPA, 2011a).
Age group-specific simulations
showed that the indoor air
concentrations of TCE were
the same for teenager and
adult individuals in each
consumer and bystander use
categories. Although the age
groups for users did not
include individuals <16 yrs
based on the assumption that
adults would be the primary
users of these products,
EPA/OPPT cannot rule out
that these products are used
by teenagers (<16 yrs) and
even younger children,
particularly in the case of the
art/crafts clear protecting
coating spray. Since there are
not survey data for consumer
behavior patterns or a way to
create varying behavior
patterns for different age
groups, the indoor air
concentrations shown in table
2-17 are extended to all users
in this risk assessment.
Page 59 of 212
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Table 2-16. Summary of E-FAST2/CEM's Input Parameters and Assumptions for Estimation of Potential Acute Dose Rates for
Specific TCE Uses
Input Parameters and
Assumptions
Consumer-
degreaser use
Bystander to
consumer-degreaser
use
Consumer-clear
protecting coating
spray use
Bystander to
consumer-clear
protecting
coating spray use
Comments
Air exchange rate (air
exchanges
per hr)
0.45
Recommended 50 percentile
value of residential air
exchange rate for all regions
within the United States (EPA,
2011a).
Portion of aerosol
in air, also called overspray
(unitless)
0.01
This value assumed that one
percent of the sprayed
product is aerosolized and
therefore immediately
available for uptake by
inhalation. Selection based
on professional judgment
(EPA, 2007b). The model
treated aerosolized portion of
TCE as a constant emitter
over the duration of use.
Whole House Volume (m
369
Value obtained from EPA
(1997). Although the updated
EFH provides a larger number
(492 m3)(EPA, 2011a), the
older value was retained to
provide more conservative
estimates. The smaller value,
369 m3, is also close to the
median value from the
updated EFH (EPA, 2011a),
although mean has increased
to 492 m3.
Page 60 of 212
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Table 2-16. Summary of E-FAST2/CEM's Input Parameters and Assumptions for Estimation of Potential Acute Dose Rates for
Specific TCE Uses
Input Parameters and
Assumptions
Consumer-
degreaser use
Bystander to
consumer-degreaser
use
Consumer-clear
protecting coating
spray use
Bystander to
consumer-clear
protecting
coating spray use
Comments
Zone 1 Volume, m
(user location)
20
N/A
20
N/A
Zone 2 Volume, m (the
rest of the house)
349
The volume of 20 m was
assigned to a utility room,
which was the proxy for a
hobby/craft room. The total
volume of the house is 369
m3. The bystanders spend the
entire day in either Zone 2 or
outside, while the users spend
2 hours in Zone 1 and 22
hours either in Zone 2 or
outside.
TCE Emission rate constant
(hr'1)
101.06
Estimated using Chinn's
algorithm based on E-FAST
model documentation (Chinn,
1981). This algorithm utilizes
the molecular weight and
vapor pressure to estimate
emission rates.
Inhalation rate (m3/hr)
0.74- During use
0.61-After use
Different for each
age group as
specified in EPA,
2011
0.74- During use
0.61-After use
Different for each age
group as specified in
EPA, 2011
During use value is based on
short-term exposure during
light activity level.
After use value is based on
data obtained from EFH (EPA,
2011a).
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Table 2-16. Summary of E-FAST2/CEM's Input Parameters and Assumptions for Estimation of Potential Acute Dose Rates for
Specific TCE Uses
Input Parameters and
Assumptions
Body weight (kg)
Consumer-
degreaser use
80
Bystander to
consumer-degreaser
use
Age group-specific
value
Consumer-clear
protecting coating
spray use
80
Bystander to
consumer-clear
protecting
coating spray use
Age group-specific
value
Comments
Mean value of body weights
for all adults (>21 years)
based on the EFH (EPA,
2011a).
Body weights for different
groupings are in the
Supplementary Information
and were based on the EFH
(EPA, 2011a).
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2.5.4 Consumer Model Results
CEM calculated air concentrations over the course of the simulation for the room of use and
the rest of the house (zone 1 and zone 2). These concentrations were converted to acute dose
rates (ADRs) using the body weight and respiration rate for each age group. The varying weight
and respiration rates of the different age groups resulted in different doses; younger age
groups had a higher ratio of inhalation rate to body mass creating a larger dose for a given air
concentration of a chemical. However, the same air concentrations were used to generate the
doses for each age group within the model's calculation engine. The normal output files for
CEM did not include the air concentrations for the different parts of the house, only the doses
were included.
Table 2-17 presents the results of the conversion from potential acute dose rates (mg/kg-
bw/day) to indoor air concentrations (ppm) for the user and bystander for the two product
scenarios. As noted in Section 2.5.3, the indoor air concentrations shown in Table 2-17 could be
applied to users of different age groups, particularly in the case of the art/crafts clear
protecting coating spray. Although adults are generally the users of these products, EPA/OPPT
cannot rule out scenarios where teenagers and even younger children may be users or be in the
same room with the user while engaging in arts and crafts projects or degreasing.
Table 2-17. Estimated TCE Air Concentrations (Time Averaged Over a Daya) from the
Residential Indoor Use of Solvent Degreasers or Clear Protecting Coating Sprays
Clear Protective
Coating Spray User
(ppm)b
Clear Protective Coating
Spray Bystanderc
(ppm)
Solvent Degreaser
User (ppm)b
Solvent Degreaser Bystander
(ppm)c
0.4
0.1
0.8
Notes:
a See Appendix I for details about the model inputs and Appendix J for the method used to convert acute dose rates
(ADRs) into air concentration of TCE (ppm).
b Air concentrations for the user categories could be extended to different age groups. EPA/OPPT believes that the
users of these products are generally adults, but teenagers and even younger children may be users or be in the
same room with the user while engaging in arts and crafts projects or degreasing.
c All age categories (<1 yrs; 1-2 yrs; 3-5 yrs; 6-10 yrs; 11-15 yrs; 16-20 yrs; and >21 yrs)
The model output sheet reported the peak concentration of TCE, but this air concentration was
not used in the risk assessment. The peak concentration was the highest concentration among
all of the 10-second time intervals that CEM simulated within a 24-hr period. The peak
concentration may only exist in the room of use for a short duration and was not considered a
good indicator of what the concentration of TCE would be for longer time periods. Thus, we did
not use the peak concentration in the risk assessment because it was not representative of a
24-hr exposure.
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2.5.4.1 Sensitivity of Model Parameters
There is no available data to refine the consumer behavior patterns used in the consumer
exposure assessment (i.e., mass of product used, time spent in room of use). Thus, a sensitivity
analysis was not conducted because it would not provide a significant improvement to the
quality of the exposure assessment.
EPA/OPPT relied on professional judgment from product descriptions as the main source of
information for setting the mass of product used and time spent in room of use. Based on our
past experience with the CEM model, it is likely that the mass used and time spent in the room
of use are particularly sensitive model parameters in the inhalation exposure simulations.
With no refinement of consumer behavior pattern data, a sensitivity analysis may erroneously
find certain input parameters to be unimportant based on current assumptions. For example,
the relative contribution of the evaporation rate to the model output variability and uncertainty
may seem unimportant based on the assumption that the user remains in the room of use for 2
hrs. This assumption may allow a wide range of evaporation rates that would result in all of the
TCE entering the room air while the user is in the room.
Measured consumer behavior pattern data could change inputs, like mass of product used and
time spent in the room of use, such that a current sensitivity analysis would lead to
inappropriate conclusions. The sensitivity of the estimated exposure concentrations over the
day of use to the evaporation rate of the chemical is an example of that concern.
J2J5/^^
TCE can be released to indoor air from the use of consumer products that contain it, as well as
from vapor intrusion and volatilization from contaminated ground water (EPA, 2011e). Where
indoor air sources are present, it is likely that indoor levels will be higher than outdoor levels
(EPA. 2011e).
Testing and monitoring has been used to evaluate potential exposures arising from VOCs
entering the indoor environment through contaminated water. Specifically, a study by EPA
researchers measured TCE concentrations from the method detection limit (
-------
2.6.1 Approach and Methodology
2.6.1.1 Selection of TCE IRIS Assessment as the Source Document for the TCE TSCA
Assessment
EPA/OPPT's work plan risk assessment for TCE is based on the hazard and dose-response
information published in the toxicological review that the U.S. EPA's Integrated Risk
Information System (IRIS) published in 2011. EPA/OPPT used the TCE IRIS assessment as the
preferred data source for toxicity information, rather than developing a new toxicological
assessment. The TCE IRIS assessment used a weight-of-evidence approach, the latest scientific
information and physiologically-based pharmacokinetic (PBPK) modeling to develop hazard and
dose-response assessments for TCE's carcinogenic and non-carcinogenic health effects resulting
from lifetime oral or inhalation exposure.
Development of TCE's hazard and dose-response assessments considered the principles set
forth by the various risk assessment guidelines issued by the National Research Council and the
U.S. EPA. Primary, peer-reviewed literature identified through December 2010 was included
where that literature was determined to be critical to the assessment (EPA, 2011e). Appendix K
provides the guidelines that were considered when developing the TCE IRIS assessment. Some
of these guidelines discussed the type of considerations that should be made when evaluating
the quality of toxicity data and study reliability. Also Appendix K describes the study selection
and data quality criteria that the U.S. EPA's IRIS program and OPPT used to evaluate the hazard
data.
The TCE IRIS assessment underwent several levels of peer review including agency review,
science consultation on the draft assessment with other federal agencies and the Executive
Office of the President, public comment, external peer review by the EPA's Science Advisory
Board (SAB) in 2002, scientific consultation by the U.S. National Academy of Sciences (NAS) in
2006 (NRC, 2006)15, external peer review of the revised draft assessment by the EPA's Science
Advisory Board (SAB) in January 2011 (EPA, 2011c)16, followed by final internal agency review
and EPA-led science discussion on the final draft.
Furthermore, EPA/OPPT consulted the EPA's Guidelines for Developmental Toxicity Risk
Assessment when making the decision to use developmental toxicity studies in the acute risk
assessment of commercial and consumer uses of degreasers and arts/crafts spray fixatives
containing TCE (EPA, 1991). The basis for this decision relies on the presumption and EPA's
NAS report, "Assessing the human health risks of trichloroethylene: Key scientific issues (2006)":
http://www.nap.edu/catalog.php7record id=11707
EPA's SAB peer review report for the 2009 EPA's Draft Assessment entitled 'Toxicological Review of
Trichloroethylene":
http://vosemite.epa.gov/sab/sabproduct.nsf/c91996cd39a82f648525742400690127/B73D5D39A8F184BD8525
7817004A1988/$File/EPA-SAB-ll-002-unsigned.pdf
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policy that a single exposure at a critical window of fetal development may produce adverse
developmental effects (EPA, 1991). In other words, a single dose in a developmental toxicity
study may be enough to cause developmental effects.
2.6.1.2 Aspects of the TCE IRIS Assessment that Were Adopted in the OPPT Risk
Assessment
2.6.1.2.1 Carcinogenic Hazard and Dose-Response Assessmen t
TCE is carcinogenic to humans by all routes of exposures as documented in the TCE IRIS
assessment. This conclusion is based on convincing evidence of a causal association between
TCE exposure in humans and kidney cancer (EPA, 2011e). The human epidemiological evidence
is strong for kidney and more limited for non-Hodgkin lymphoma (NHL) and liver cancer, and
less convincing for biliary tract cancer. Less human evidence is found for an association
between TCE exposure and other types of cancer, including bladder, esophageal, prostate,
cervical, breast, and childhood leukemia (EPA, 2011e). Further support for TCE's carcinogenic
characterization comes from positive results in multiple rodent cancer bioassays in rats and
mice of both sexes, similar toxicokinetics between rodents and humans, mechanistic data
supporting a mutagenic mode of action for kidney tumors, and the lack of mechanistic data
supporting the conclusion that any of the mode(s) of action for TCE-induced rodent tumors are
irrelevant to humans (EPA, 2011e).
The cancer dose-response analysis used linear-dose extrapolation to derive an inhalation unit
risk (IUR) of 2 x 10~2 per ppm (4.1 x 10~6 per u,g/m3) for various cancers17. The IUR for TCE was
based on human kidney cancer risks that were reported by Charbotel et al. (2006) and adjusted
for potential risk for NHL and liver cancer based on human epidemiological data (EPA, 2011e).
The IUR is defined as the upper-bound excess lifetime cancer risk estimated to result from
continuous exposure to an agent at a concentration of 1 u.g/m3 in air (EPA, 2011b). The IUR was
used in the EPA/OPPT risk assessment to estimate excess cancer risks for the inhalation
occupational exposures scenarios. There is high confidence in the IUR because it was based on
good quality human data, and was similar to unit risk estimates derived from multiple rodent
bioassays (EPA, 2011e). Moreover, there was sufficient weight of evidence to conclude that
TCE operates through a mutagenic mode of action for kidney tumors (EPA, 2011e).
EPA/OPPT decided not to use the IUR to calculate the theoretical cancer risk associated with a
single (acute) exposure to solvent degreasers, arts/crafts fixative and spotting agent products
containing TCE. NRC (2001) published methodology for extrapolating cancer risks from chronic
to short-term exposures to mutagenic carcinogens. These methods were published with the
caveat that extrapolation of lifetime theoretical excess cancer risks to single exposures has
great uncertainties.
The dose-response analysis for cancer endpoints can be found in Chapter 5 of the TCE IRIS assessment EPA
Page 66 of 212
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As NRC (2001) explains, "There are no adopted state or federal regulatory methodologies for
deriving short-term exposure standards for workplace or ambient air based on carcinogenic risk,
because nearly all carcinogenicity studies in animals and retrospective epidemiologic studies
have entailed high-dose, long-term exposures. As a result, there is uncertainty regarding the
extrapolation from continuous lifetime studies in animals to the case ofonce-in-a-lifetime
human exposures. This is particularly problematical, because the specific biologic mechanisms
at the molecular, cellular, and tissue levels leading to cancer are often exceedingly diverse,
complex, or not known. It is also possible that the mechanisms of injury of brief, high-dose
exposures will often differ from those following long-term exposures. To date, U.S. federal
regulatory agencies have not established regulatory standards based on, or applicable to, less
than lifetime exposures to carcinogenic substances (NRC, 2001)." Thus, the final EPA/OPPT work
plan risk assessment for TCE does not estimate excess cancer risks for acute exposures because
the relationship between a single short-term exposure to TCE and the induction of cancer in
humans has not been established in the current scientific literature.
2.6.1.2.2 Non-Cancer Hazard and Dose-Response Assessment
EPA/OPPT used margin of exposures (MOEs)18 to estimate non-cancer risks based on the
following:
1. the lowest PBPK-derived human equivalent concentrations (HECs) within each health
effects domain reported in the TCE IRIS assessment;
2. the same endpoint/study-specific uncertainty factors (UFs) that the IRIS program applied to
the PBPK-derived HECs; and
3. the exposure estimates calculated for the TCE uses examined in this risk assessment (see
Environmental Releases and Exposure Summary).
MOEs allow for the presentation of a range of risk estimates rather than a single risk estimate
based on the inhalation reference concentration (RfC). Given the different exposure scenarios
considered (both acute and chronic for small commercial degreasers, and just acute for the two
consumer exposure scenarios), different endpoints were used based on the expected exposure
durations. For non-cancer effects, risks to developmental effects were evaluated for acute
(short-term) exposures, whereas risks to other adverse effects (toxicity to the liver, kidney,
nervous system, immune system, and the reproductive system) were evaluated for repeated
(chronic) exposures to TCE.
Table 2-18 lists the studies and corresponding HECs and UFs that EPA/OPPT used in the work
plan risk assessment for TCE. Key studies in Table 2-18 are briefly described in the Human
Health Hazard Summary along with other toxicity and epidemiological studies, with detailed
descriptions provided in the TCE IRIS assessment (EPA, 2011e). Appendix L contains the
complete list of oral and inhalation non-cancer studies within each health effects domain that
the TCE IRIS assessment considered suitable for dose-response analysis.
18 Margin of Exposure (MOE) = (Non-cancer hazard value, POD) 4- (Human Exposure). See equation in Table 2-29.
The benchmark MOE is used to interpret the MOEs and consists of the total UF set by the IRIS program for each
study in Table 2-18. See section 2.7.1 for an explanation of the benchmark MOE.
Page 67 of 212
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The TCE IRIS assessment used a series of steps to generate the PBPK-derived HECs that OPPT
used in its risk assessment (Table 2-19). EPA/OPPT did not rely on those steps that were
associated with the derivation of candidate RfCs (cRfC) or final RfC value (Table 2-19). Below is a
brief discussion of those steps that EPA/OPPT adopted in this risk assessment.
The non-cancer dose-response analysis in the TCE IRIS assessment commenced with the review
and selection of high quality epidemiological and toxicity studies that reported both adverse
non-cancer health effects and quantitative dose-response data19 (Table 2-19). Subsequently,
point of departures (PODs)20 were identified for those studies that had suitable data for dose-
response analysis. PODs can be a NOAEL21 or LOAEL22 for an observed incidence, or change in
level of response, or the lower confidence limit on the dose at the benchmark dose (BMD)23.
The dose-response assessment was organized in five health effects domains: (1) neurotoxicity;
(2) systemic (body weight) and organ toxicity (liver and kidney effects); (3) immunotoxicity; (4)
reproductive; and (5) developmental effects. PBPK modeling was used to estimate internal dose
PODs (idPOD) and subsequently HECs based on the oral and inhalation PODs identified in earlier
steps. The PBPK modeling integrated internal dose-metrics based on TCE's mode of action and
the role of different TCE metabolites in toxicity (EPA, 2011e). Note that the effects within the
same health effect domain were generally assumed to have the same relevant internal dose-
metrics (EPA. 2011e).
19 Non-cancer toxicological and epidemiological studies are described in Chapter 4 of the TCE IRIS assessment.
20 A point of departure (POD) is a dose or concentration that can be considered to be in the range of observed
responses, without significant extrapolation. A POD is used to mark the beginning of extrapolation to determine
risk associated with lower environmentally relevant human exposures EPA(2011b).
21 NOAEL=No-observed-adverse-effect level
22 LOAEL= Lowest-observed-adverse-effect level
23 The benchmark dose (BMD) is a dose or concentration that produces a predetermined change in response rate
of an adverse effect (called the benchmark response or BMR) compared to background EPA(2011b).
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Table 2-18. Lowest PBPK-derived HECs for different effects domains based on analysis in TCE IRIS assessment
Exposure
Duration for Risk
Analysis
CHRONIC
Target Organ/
System
Liver
Kidney
Nervous
System
Immune System
Reproductive System
Species
Mouse (male)
Rat
(female)
Rat
(male)
Mouse
(female)
Mouse
(female)
Human
(male)
Route of
Exposure
Inhalation
Oral
(gavage)
Inhalation
Oral
(drinking
water)
Oral
(drinking
water)
Inhalation
Range of Doses
or
Concentrations1
37 to 3,600 ppm
500 to 1,000
mg/kg-bw/day
50 to 300 ppm
0.001 to 14 ppm
(0.001 to 14
mg/kg-bw/day)
0.001 to 14 ppm
(0.001 to 14
mg/kg-bw/day)
29.6 ppm
(mean exposure)
Duration
Continuous and
intermittent
exposures,
variable time
periods for
30-120 days
5 days/week for
104 weeks
8 hrs/day,
5 days/weeks
for 6 weeks
27-30 weeks
27-30 weeks
Measured
values after an
8-hour work
shift; mean 5.1
years on job
POD Type2
BMDL10 -
21.6 ppm
BMDL05 =
9.45 mg/kg-
bw/day
LOAEL = 12
ppm
LOAEL =
0.35 mg/kg-
bw/day
LOAEL =
0.35 mg/kg-
bw/day
BMDL10 =
1.4 ppm
Effect
Increased
liver/body weight
ratio
Toxic nephropathy
Significant
decreases in
wakefulness
Decrease in thymus
weight and thymus
cellularity
Autoimmunity
(increased anti-
dsDNAandssDNA
antibodies)
Decreased normal
sperm morphology
and
hyperzoospermia
HEC50
(ppm)3
25
0.042
13
0.092
0.092
1.4
HEC95
(ppm)
12
0.0085
6.4
0.045
0.045
0.7
HEC99
(ppm)
9.1
0.0056
4.8
0.033
0.033
0.5
Uncertainty
Factors (UFs)
for
Benchmark
MOE4
UFS=1; UFA= 3;
UFH=3; UFL=1;
Total UF-10
UFS=1; UFA= 3;
UFH=3; UFL=1;
Total UF=10
UFS=3; UFA= 3;
UFH=3;
UFL=10;
Total UF=300
UFS=1; UFA= 3;
UFH=3;
UFL=10;
Total UF=1005
UFS=1; UFA= 3;
UFH=3; UFL=3;
UFS=10; UFA=
1;
UFH=3; UFL=1;
Reference
Kiellstrand
etal.
(1983)
NTP (1988)
Arito et al.
(1994)
Keiletal.
(2009)
Keil, D. E.
etal.
(2009)
Chia et al.
(1996)
Page 69 of 212
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Table 2-18. Lowest PBPK-derived HECs for different effects domains based on analysis in TCE IRIS assessment
Exposure
Duration for Risk
Analysis
ACUTE OR
CHRONIC
Target Organ/
System
Developmental
effects
Species
Rat
(female)
Route of
Exposure
Oral
(drinking
water)
Range of Doses
or
Concentrations1
2. 5 to 1,100 ppm
(2. 5 to 1,100
mg/kg-bw/day)
Duration
22 days
throughout
gestation
(gestational
days 0 to 22)
POD Type2
BMDL01 =
0.0207
mg/kg-
bw/day
Effect
Heart
malformations
HEC50
(ppm)3
0.012
HEC95
(ppm)
0.0051
HEC99
(ppm)
0.0037
Uncertainty
Factors (UFs)
for
Benchmark
MOE4
UFS=1; UFA= 3;
UFH=3; UFL=1;
Total UF=10
Reference
Johnson et
al. (2003)
Notes: Control concentrations were not included in the table, but discussed in the study summaries in section 2.6.2.
POD type can be NOAEL, LOAEL, or BMDL; the IRIS program adjusted all values to continuous exposure.
3 1 ppm = 5.374 mg/m3
4 UFs=subchronic to chronic UF; UFA=interspecies UF; UFH=intraspecies UF; UFL=LOAEL to NOAEL UF. UF values were those used in the
TCE IRIS assessment (EPA, 2011e).
5 Two different effects were reported by Keil et al, (2009): decreased thymic weight and cellularity and autoimmunity. A total UF of 100 was used for the thymus
toxicity, whereas a total UF of 30 was used for the autoimmune effects. The TCE IRIS assessment allocated different LOAEL-to-NOAEL uncertainty factors (UFL)
based on the severity of the effects, which resulted in different total UF (EPA. 2011e).
Page 70 of 212
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Table 2-19. Steps of the Non-Cancer TCE IRIS Process that Were Considered in OPPT's Non-
Cancer Hazard/Dose-Response Approach 1
Step#
1
2
3
4
5
6
7
8
9
10
NON-CANCER IRIS APPROACH
Evaluation of all studies (inhalation and oral) that provided (1) non-cancer
adverse health effects and (2) quantitative dose-response data
Identification of the Points of Departure (POD) based on applied dose
Adjustment of each POD by endpoint/study-specific uncertainty factors
(UFs)
Derivation of candidate reference concentrations (cRfCs) and selection of
the lowest values within each health endpoint domain taking into
account the confidence of the estimate
Application of PBPK modeling to estimate internal dose PODs (i.e., HECs)
for those candidate critical effects selected from inhalation and oral
studies based on applied dose
Estimation of interspecies and within-human pharmacokinetic variability
for the internal dose PODs by PBPK modeling
Adjustment of each internal dose POD by endpoint/study-specific UFs 2
Derivation of PBPK model-based candidate RfCs (p-cRfC) for each
candidate critical effect
Characterization of the uncertainties of the cRfCs derived by applied dose
and PBPK modeling (p-cRfC) by using quantitative uncertainty analyses of
pharmacokinetic uncertainty and variability from the Bayesian population
analysis 3
Evaluation of the most sensitive cRfCs and set final RfC value.
NON-CANCER OPPT
APPROACH
Considered (+) /Not
Considered (—)
+
+
—
—
+
+
—
—
—
—
Notes:
1 Table 2-19 is an adaptation of Figure 5-1 (Chapter 5) in the TCE IRIS assessment (EPA, 2011e).
2 Endpoint/study-specific UFs were used as benchmark MOEs in the OPPT's TCE risk assessment.
3 EPA/OPPT consulted the characterization of the uncertainties in the TCE IRIS assessment and those aspects
related to the uncertainties of the individual studies and the generation of PBPK-derived HECs were discussed in
the OPPT's TCE risk assessment.
Furthermore, the PBPK model was used to estimate the interspecies and within-human
pharmacokinetic variability (or just within-human variability for the human-based PODs)
corresponding to each idPOD for each candidate critical effect. The results of this calculation
were 50th, 95th and 99th percentile HEC estimates for POD analyzed within each health effects
domain. Also, the PBPK model integrated a Bayesian population analysis to characterize
pharmacokinetic uncertainties and variability (EPA, 2011e). More information on the PBPK
modeling is provided in the next section below, including how the HECs were calculated.
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EPA/OPPT used the endpoint/study-specific UFs from the TCE IRIS assessment as the
benchmark MOEs for the risk calculations. These UFs were applied to the PBPK-derived PODs to
account for (1) variation in susceptibility among the members of the human population (i.e.,
inter-individual or intraspecies variability); (2) uncertainty in extrapolating animal data to
humans (i.e., interspecies uncertainty); (3) uncertainty in extrapolating from data obtained in a
study with less-than-lifetime exposure (i.e., extrapolating from subchronic to chronic exposure);
and (4) uncertainty in extrapolating from a LOAEL rather than from a NOAEL (EPA, 2011b).
2.6.1.3 Absorption, Distribution, Metabolism, and Excretion
TCE is fat soluble (lipophilic) and easily crosses biological membranes. Though there are
quantitative differences across species and routes, TCE is readily absorbed into the body
following oral, dermal, or inhalation exposure. Because of its lipophilicity, TCE can cross the
placenta and also passes into breast milk (EPA, 2011e).
Absorption following inhalation of TCE is rapid and the inhaled absorbed dose is proportional to
the exposure concentration, duration of exposure, and lung ventilation rate. Likewise, TCE is
rapidly absorbed from the gastrointestinal tract into the systemic circulation (i.e., blood)
following oral ingestion. Oral absorption of TCE has been shown to be influenced by dose of the
chemical, the dosing vehicle and stomach contents. Absorbed TCE is first transported to the
liver where it is metabolized for eventual elimination (i.e., "first-pass effect") (EPA, 2011e).
Rapid absorption through the skin has been shown by both vapor and liquid TCE contact with
the skin. EPA (2011e) summarized several volunteer studies in which both TCE liquid and vapors
were shown to be absorbed in humans via the dermal route. Following exposures of between
20 and 30 minutes, absorption was rapid, with peak TCE levels in expired air occurring within 15
minutes (liquid) and 30 minutes (vapor).
Regardless of the route of exposure, TCE is widely distributed throughout the body. TCE levels
can be found in many different human and rodent tissues including: brain, muscle, heart,
kidney, lung, liver, and adipose tissues. It can also be found in human maternal and fetal blood
and in the breast milk of lactating women (EPA, 2011e).
Distribution of TCE to the body is determined by how much TCE is absorbed and eliminated by
the lungs. The blood-to-air partition coefficient is used to quantify the resulting concentration
in blood leaving the lungs at equilibrium with alveolar air. Partition coefficients have been
measured in vitro and range between 8.1 and 11.7 in humans and between 13.3 and 25.8 in
rodents (EPA, 2011e). Once in the alveolar blood, the solubility of TCE in blood is the major
determining factor in its distribution to the heart and other parts of the body and is measured
by the blood-to-tissue partition coefficient24. Other factors influencing its distribution are age-
24 This represents the ratio of the concentration of TCE in blood to the concentration of TCE in tissue. When the
ratio is much less than 1, more TCE would be found in tissue rather than in the circulating blood.
Page 72 of 212
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dependencies (i.e., largely based on anatomical and physiological parameters such as metabolic
and ventilation rates) and TCE binding to tissues/cellular components.
The metabolism of TCE has been extensively studied in humans and rodents (EPA, 2011e).
Animals and humans metabolize TCE to metabolites to varying degrees. These metabolites are
known to play a key role in causing TCE-associated toxic effects. TCE metabolites are known to
target the liver and kidney. The two major metabolic pathways are (1) oxidative metabolism via
the cytochrome P450 (CYP) mixed function oxidase system and (2) glutathione (GSH)
conjugation followed by further biotransformations and processing with other enzymes (EPA,
2011e). The liver is the major tissue for the oxidative and GSH conjugation metabolic pathways.
Both pathways are saturable, and above the saturable concentration/dose, TCE is excreted
unchanged in expired air. Table 2-20 presents the important metabolites formed following both
the CYP (oxidation) and GSH (conjugation) pathways in humans and animals. The amount and
types of metabolites formed are important for understanding the toxicity of TCE in both
animals and humans.
These major TCE metabolites as well as a number of minor metabolites are also observed in the
metabolic pathway of TCE-related compounds (Table 2-21). This may be important in
determining exposures because people may be co-exposed to many of these solvents at the
same time [e.g., dichloroacetic acid (DCA) as disinfection by-products of chlorination of drinking
water supplies] (Johnson et al., 1998a). Concomitant exposures to TCE and its related
compounds can affect TCE's metabolism and increase toxicity by generating higher internal
metabolite concentrations than those resulting from TCE exposure only (EPA, 2011e).
Table 2-20. TCE Metabolites Identified by Pathway
Oxidative Metabolites
Chloral
(metabolized to TCOH")
Trichloroethylene oxide
(re-arranged to DCAC?)
Trichloroethanol or TCOH
(metabolized to TCOGC)
Trichloroacetic acid orTCA
(may lead to DCAd)
GSH Conjugation Metabolites
DCVGe
(metabolized to DCVCf isomers)
Abbreviations: a TCOH = trichloroethanol; DCAC= dichloroacetyl chloride; c TCOG= trichloroethanol,
glucuronide conjugate; d DCA=dichloroacetic acid; e DCVG= S-dichlorovinyl-glutathione (collectively, the 1,2-
and 2,2- isomers); f DCVC= S-dichlorovinyl-L-cysteine (collectively, the 1,2- and 2,2- isomers)
A review of in vitro metabolism data in the liver suggested that rodents (i.e., especially mice)
have greater capacity to metabolize TCE via the oxidation pathway (EPA, 2011e). The in vitro
data have also reported modest sex- and age-dependent differences in the oxidative TCE
metabolism in humans and animals. Significant variability may exist in human susceptibility to
TCE toxicity given the existence of CYP isoforms and the variability in CYP-mediated TCE
oxidation (EPA. 2011e).
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Table 2-21. Common Metabolites of TCE and Related Compounds
Metabolites
Tetrachloro-
ethylene
1,1,2,2,-
Tetrachloro-
ethane
TCE
1,1,1-
Trichloro-
ethane
1,2,-
Dichloro-
ethylene
1,1-
Dichloro-
ethane
Oxalic acid
X
X
Chloral
Chloral hydrate (CH)
Monochloroacetic acid
Dichloroacetic acid (DCA)
Trichloroacetic acid (TCA)
Trichloroethanol (TCOH)
Trichloroethanol-
glucuronide
Note: Adapted from Table 2-1 in EPA(2011e)
Conjugation is a process that generally leads to detoxification. However, this is not the case for
TCE and many other halogenated alkanes and alkenes because they are biotransformed into
reactive metabolites. The eventual metabolite(s) of concern for TCE are formed several steps
from the initial GSH conjugate formed in the liver, which ultimately results in toxicity or
carcinogenicity in the kidney (EPA, 2011e). The conjugation of TCE to GSH produces S-(l,2-
dichlorovinyl) glutathione or its isomer S-(2,2-dichlorovinyl) glutathione (collectively, S-
dichlorovinyl-glutathione or DCVG) (EPA, 2011e). Metabolic enzymes then convert DCVG into
two cysteine conjugate isomers: S-(l,2-dichlorovinyl cysteine) [1,2-DCVC] or S-(2,2-
dichlorovinyl cysteine) [2,2-DCVC] (collectively, DCVC). N-acetylation then transforms DCVC into
N-acetyl-S (1,2 dichlorovinyl)-L-cysteine or N-acetyl-S (2,2 dichlorovinyl)-L-cysteine (collectively,
N Ac DCVC) (EPA. 2011e).
There are various theories about how DCVC is toxic to the kidney and, to a lesser extent, to the
liver. One theory states that a 0-lyase enzyme catalyzes the breakdown of 1,2-DCVC to S
dichlorovinyl thiol (DCVT), an unstable intermediate that rearranges to other metabolites
(enethiols) that form covalent bonds with cellular nucleophiles and results in toxicity (EPA,
2011e). Another theory is that there is a kidney enzyme (L-alpha-hydroxy [L-amino] acid
oxidase) that can form intermediates and keto acid analogues that decompose to DCVT. In rat
kidney homogenates, this enzyme appeared to be responsible for up to 35 percent of the GSH
pathway but is not found in humans (EPA, 2011e). A third theory suggests involvement of
sulfoxidation of either the DCVC or NAcDCVC by flavin-containing monooxygenase (FM03) and
CYP3A enzymes, respectively (EPA, 2011e).
In contrast to the CYP oxidation pathway, there appear to be sex and species differences in TCE
metabolism via the GSH pathway (EPA, 2011e). Animal data show that rates of TCE GSH
conjugation in male rats/mice are higher than females. According to some in vitro data, the
rates of DCVG production in liver/kidney cytosol are highest in humans, followed by mice, and
then rats. In vitro data also suggest that y-glutamyl transpeptidase (i.e., GGT, an enzyme
involved in DCVC production) activity in kidneys seems to be highest in rats, then humans, and
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then mice (EPA, 2011e). Furthermore, species-dependent enzymatic activities have been
reported for the |3-lyase and FM03 enzymes (EPA, 2011e).
Thus, the key in evaluating the TCE metabolism data is to determine the relative roles of CYP
and GSH pathways. It appears that, in rodents and humans, the oxidation pathway is clearly
more dominant than the GSH pathway. EPA (2011e) suggests that the GSH pathway may play a
larger role in humans than it does in rodents, but there is substantial uncertainty based on the
available data. In fact, Jollow et al. (2009), using essentially the same data, suggested that
rodents have a higher capacity to conjugate TCE with GSH and are thus more susceptible to
kidney toxicity/cancer compared with humans.
The majority of TCE absorbed into the body is eliminated by the metabolic pathways discussed
above. With the exception of unchanged TCE and C02, which are excreted by exhalation, most
TCE metabolites (i.e., TCA, TCOH, GSH metabolites) are primarily excreted in urine and feces.
Elimination of TCE metabolites can also occur through the sweat and saliva, but these excretion
routes are likely to be relatively minor (EPA, 2011e).
Half-lives are useful indicators for the bioaccumulation potential of the chemical. The excretion
of unchanged TCE in exhaled air was studied by Sato et al. (1977) who exposed four male
volunteers to 100 ppm TCE for 4 hrs. Sato et al. (1977) reported three first-order phases of
pulmonary excretion in the first 10 hrs after cessation of exposure, with fitted half-times of
pulmonary elimination of 0.04, 0.67, and 5.6 hrs, respectively (EPA, 2011e). Opdam (1989)
reported terminal half-lives of 8-44 hrs at rest following exposure of both female and male
volunteers to 6-38 ppm TCE for 0.5-1 hr exposures with up to 20-310 hrs of subsequent
monitoring of alveolar air.
Another human toxicokinetic study was conducted by Chiu et al. (2007), in which male human
volunteers were exposed to 1 ppm TCE for 6 hrs with alveolar air collected during exposure and
up to 6 days post-exposure. Chiu et al. (2007) reported pulmonary terminal half-lives of 14-23
hrs (EPA, 2011e). The long terminal half-times suggest that the lungs require considerable time
to completely eliminate TCE, primarily due to high partitioning to adipose tissues (EPA, 2011e).
As for rodent data, rats and mice exposed to TCE by gavage showed unchanged TCE and C02 as
exhalation excretion products (EPA, 2011e).
Various laboratories have studied the urinary elimination kinetics of TCE and its major
metabolites in humans and rodents. The current practice is to measure urinary oxidative
metabolites, including total trichloro compounds (TTC), because urinary levels of unchanged
TCE have been at or below detection limits. Ikeda and Imamura (1973) measured various
metabolites (i.e., TTC, TCOH, TCA) in human volunteers for three post-exposure days in five
exposure groups (no concentrations provided in the study). The elimination half-lives for TTC
were 26.1-48.8 hrs in males and 50.7 hrs in females. The elimination half-lives for TCOH were
15.3 hrs and 52.7 hrs for males and females, respectively. The elimination half-lives for TCA
were 39.7 hrs and 57.6 hrs for males and females, respectively (EPA, 2011e).
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Ikeda (1977) evaluated the urinary elimination of TCE and its metabolites at the occupational
setting. Female and male workers were intermittently exposed to 50 ppm and 200 ppm TCE,
respectively at the work place (exposure duration not reported). Urinary elimination half-lives
forTTC, TCOH, and TCA were 26.1,15.3, and 39.7 hrs in males, respectively, and 50.7, 42.7 and
57.6 hrs in females, respectively (EPA, 2011e).
Animal studies have shown that rodents exhibit faster urinary elimination kinetics than
humans. For instance, Ikeda and Imamura (1973) evaluated the urinary elimination of TTCs
following inhalation exposure to TCE for 8 hrs to 50, 100, or 250 ppm TCE. A second experiment
exposed rats to 1.47 g/kg TCE by intraperitoneal injection. The urinary elimination half-lives of
TTCs were 14.3-15.6 hrs for female rats and 15.5-16.6 hrs for male rats; the route of
administration did not appear to influence half-life value (EPA, 2011e). Oral rodent studies
reported urinary elimination of radiolabeled TCE within 1 or 2 days after exposure (Dekant et
al.. 1984: Green and Prout. 1985: Proutetal.. 1985).
^6.1^4^
Given the complicated metabolic profile of TCE, understanding the relationship between the
external dose/concentration (i.e., exposure) and internal dose at the target organ of interest is
critical to quantifying potential risk(s) because internal dose is more closely associated with
toxicity at the target tissue (EPA, 2006a). Predictions of internal dose in chemical risk
assessments are achieved by employing PBPK modeling.
PBPK models use a series of mathematical representations to describe the absorption,
distribution, metabolism and excretion of a chemical and its metabolites. Because PBPK
modeling assumes that the toxic effects in the target tissue are closely related to the internal
dose of the biologically active form of the chemical, knowledge about the chemical's mode of
action guides the selection of the appropriate dose metric25. Traditional risk estimates based on
applied dose carry higher uncertainties than those based on PBPK-derived internal dose
metrics. This reduction in uncertainty and the versatility of PBPK approaches have resulted in a
growing interest to use these models in risk assessment products (EPA, 2006a).
U.S. EPA developed a comprehensive Bayesian PBPK model-based analysis of TCE and its
metabolites in mice, rats and humans (EPA, 2011e)26. This model is briefly discussed beloi
provide clarity on how the PBPK modeling was used to estimate the PBPK-derived HECs.
Physiological, chemical, in vitro and in vivo data were considered when building the PBPK
model, including many studies in animals and humans that quantified TCE levels in various
25 Dose metric is defined as the target tissue dose that is closely related to ensuing adverse responses. Dose
metrics used for risk assessment applications should reflect the biologically active form of chemical, its level and
duration on internal exposure, as well as intensity EPA (2006a).
26 Refer to the TCE IRIS assessment to obtain a summary of the history of the TCE PBPK models that have been built
over the years as well as detailed information on the updated model reported in both Evans et al. (2009) and
Chiuetal. (2009).
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tissues following oral and inhalation exposures. Some of these studies provided key
data/parameters for the calibration of the PBPK model used in the IRIS assessment (EPA,
2011e). All of this information was used to build a model that was able to predict different
dose-metrics as measures of potential TCE toxicity. Each dose-metric was developed to
evaluate a different metabolic pathway/target organ effect based on the dose-response
analysis and understanding of metabolism (Tables 2-22 and 2-23).
Table 2-22. List of All of the PBPK-Modeled Dose Metrics Used in the TCE IRIS Assessment
Dose-Metric
Identifier
ABioactDCVCBW374
ABioactDCVCKid
AMetGSHBBW374
AMetLivlBW374
AMetLivOtherBW374
AMetLivOtherLiv
AMetLngBW374
AMetLngResp
AUCCBId
AUCCTCOH
AUCLivTCA
TotMetabBW374
TotOxMetabBW374
TotTCAInBW
Explanation of What the
Dose-Metric Identifier Represents
Amount of DCVC bioactivated in the kidney per unit body weight
Amount of DCVC bioactivated in the kidney per unit kidney mass
Amount of TCE conjugated with GSH
Amount of TCE oxidized in liver
Amount of TCE oxidized to metabolites other than TCAorTCOH per unit body weight
Amount of TCE oxidized to metabolites other than TCAorTCOH per unit liver weight
Amount of TCE oxidized in respiratory tract (per unit body weight)
Amount of TCE oxidized in respiratory tract per unit respiratory tract tissue
Area under the curve of venous blood concentration of TCE
Area under the curve of blood concentration of TCOH
Area under the curve of the liver concentration of TCA
Total amount of TCE metabolized per unit body weight
Total amount of TCE oxidized per unit body weight
Total amount of TCA produced
Table 2-23. Dose-Metrics for Cancer and Non-Cancer Endpoints Used in the OPPT Assessment
Non-Cancer
Endpoints
Cancer
Target Organ/ System
Liver
Kidney
Nervous System
Immune System
Reproductive System
Developmental effects 2
Carcinogenic effects in
various organs
Reference
Kjellstrandetal. (1983)
NTP(1988)
Aritoetal. (1994)
Keiletal. (2009)
Chiaetal.(1996)
Johnson etal. (2003)
Charbotel et al. (2006)(kidney
cancer) and human epidemiological
studies for NHL and liver cancer
Preferred
Dose-Metric 1
AMetLivlBW374
ABioactDCVCBW374
TotMetabBW374
TotMetabBW374
TotMetabBW374
TotOxMetabBW374
ABioactDCVCBW374 (ID R for kidney cancer)
AMetLivlBW374 (ID R for liver cancer)
TotMetabBW374 (IUR for NHL)
Notes:
1 The non-cancer dose-metrics correspond to those studies resulting in the lowest PBPK-derived HECs as listed in Table 2-19
based on the analysis presented in the TCE IRIS assessment (EPA, 2011e).
2 The maternal dose metric was used as surrogate for fetal exposure.
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For developmental toxicity endpoints, the TCE PBPK model did not incorporate a pregnancy
model to estimate the internal dose of TCE in the developing fetus. In this case, the maternal
dose-metric was used as the surrogate measure of target tissue dose in the developing fetus. A
complete description of the TCE PBPK model, including the rationale for parameter choices in
animals and humans, choice of dose metric, and experimental information used to calibrate
and optimize the model is found in the TCE IRIS assessment (EPA, 2011e).
As shown in Figures 2-4 and 2-5, several steps were needed to derive the PBPK-derived HECs
used in this assessment. First, the rodent PBPK model was run to estimate rodent idPODs for
the applied dose PODs (i.e., LOAEL, NOAEL, or BMDL) that were identified in the TCE IRIS
assessment. Separately, the human PBPK model was run for a range of continuous exposures
from 0.1 to 2,000 ppm or 0.1 to 2,000 mg/kg-bw/day to establish the relationship between
human exposure air levels and internal dose for the same dose-metric evaluated in the rodent
PBPK model. This relationship was used to derive HECs corresponding to the idPOD by
interpolation (EPA, 2011e).
Figure 2-4. Dose-Response Analyses of Rodent Non-Cancer Effects Using the Rodent and
Human PBPK Models
Fixed
0.1 to 20OO ppm TCE in
air or 0.1 to 2000
mg/kg-bw/day
continuous exposure
Distribution {separate
uncertainty and variability)
Human internal dose
at 50th percentile as
function of applied
dose
Human internal dose
at 95th percentile as
function of applied
dose
Human internal dose
at 99th percentile as
function of applied
dose
HEC50
HEC^
MEG,,
Notes: Figure adapted from Figure 5-2 (Chapter 5, TCE IRIS assessment) (EPA, 2011e). Square nodes indicate point
values, circle nodes indicate distributions and the inverted triangle indicates a (deterministic) functional
relationship.
Page 78 of 212
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Figure 2-5. Example of HEC99 Estimation through Interpecies, Intraspecies and Route-to-
Route Extrapolation from a Rodent Study LOAEL/NOAEL
Rodent internal
dose
Human internal
dose4
Study dose groups
Uncertainty &
variability
distribution
Human inhalation
'exposure (ppm)
^HEcJj
Figure adapted from Figure 5-3 (Chapter 5,
TCE IRIS assessment)
Notes: When using benchmark dose estimates,
the idPOD is the modeled BMDL in
internal dose units.
The rodent population model was designed to characterize study-to-study variation and used
median values of dose-metrics to generate idPODs. The rodent PBPK model did not characterize
variation within studies and assumed that the rodent idPODs were for pharmacokinetically
identical animals. The basis of that assumption was that animals with the same
sex/species/strain combination were considered pharmacokinetically identical and represented
by the group average. In practice, the use of median or mean internal doses for rodents did not
make much difference except when the uncertainty in the rodent dose-metric was high (EPA,
2011e).
On the other hand, the human population model characterizes toxicokinetic uncertainty and
individual-to-individual variation and used median, 95th and 99th percentile values of dose-
metrics to general human idPODs. The 50th, 95th or 95th percentile of the combined uncertainty
and variability distribution of human internal doses was used to derive the HEC50, HEC95 or
HECgg estimates, respectively. The HEC95and HECggwere interpreted as being the
concentrations of TCE in air for which there is 95% and 99% likelihood, respectively, that a
randomly selected individual will have an internal dose less than or equal to the idPOD derived
from the rodent study. The HEC50 was interpreted as being the concentration of TCE in air for
which there is 50% likelihood that a randomly selected individual will have an internal dose less
than or equal to the idPOD from the rodent study. The TCE IRIS assessment preferred the HEC99
for the non-cancer dose-response analysis because the HEC99 was interpreted to be protective
for a sensitive individual (EPA, 2011e).
EPA/OPPT supported the interpretation of the HEC99 as expressed in the TCE IRIS assessment.
Hence, HEC99-based risk estimates are favored in this assessment over those estimated from
the HEC50and HEC95 values. However, risk estimates based on the HEC50 and HEC95 values are
also included.
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Presenting risks for the HEC50, HEC95 and HEC99 values is intended to provide a sense of the
difference between the median, the 95% and 99% confidence bound for the combined
uncertainty and variability. Calculations of HEC50/95 and HEC50/99 ratios generally showed a 2-3
fold difference for the various studies identified in Table 2-18. The exception was the study
reporting kidney effects (NTP, 1988) that showed high HEC50/95 and HEC50/99 ratios (i.e., 5-9-
fold) due to uncertainties in the rodent internal dose estimates (EPA, 2011e). In contrast, HEC95
values were closely similar to HEC99 values with HECgs/gg ratios showing a 1.3-1.5 fold
difference.
In this assessment, it was assumed there was no substantial buildup of TCE in the body
between exposure events due to TCE's short biological half-life in humans (~51 hrs). This
assumption was supported by PBPK simulations presented in Tables 2-24 and 2-25. In these
simulations, the TCE PBPK model was used to estimate HECs at the 50th and 99th percentile for
the cardiac malformation endpoint under continuous or intermittent exposure to TCE for
different exposure durations (i.e., 1 day, 3 weeks, or 9 months). The 1-day HEC at the 50th and
99th percentile did not show significant variation when compared to the HECs for the other
exposure durations (i.e., 3 weeks and 9 months) for the continuous or intermittent exposures
(Tables 2-24 and 2-25). Thus, the results from the PBPK simulations showed that the
assumption (i.e., no substantial buildup of TCE in the body between exposure events) is
reasonable to use in the EPA/OPPT's risk assessment.
Table 2-24. Comparison of TCE Human Equivalent Concentrations (HECs) Under Constant,
Continuous Exposure for Different Exposure Durations
Duration
Chronic (steady-state)
9 months (40 weeks)
3 weeks
Iday
HEC Equivalent to the Benchmark Dose Lower Confidence Limit
for Cardiac Malformations (ug/m3) a
Median estimate
62
62
63
71
Upper 99th percentile estimate
20
20
20
21
Notes:
The internal dose metric of u.g oxidized per day per [kg body weight]54 was selected as the basis for interspecies,
intraspecies, and route-to-route extrapolation for the cardiac malformations endpoint in the 2011 TCE IRIS
assessment (EPA, 2011e). The PBPK model was parameterized for human females. HECs are rounded to two
significant figures.
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Table 2-25. Comparison of TCE Human Equivalent Concentrations (HECs) Under
Intermittent (Occupational) Exposure for Different Exposure Durations
Duration
Chronic (steady-state)
9 months (40 weeks)
3 weeks
Iday
HEC Equivalent to the Benchmark Dose Lower Confidence Limit
for Cardiac Malformations (ug/m3) a
Median estimate
62
62
63
68
Upper 99th percentile estimate
20
20
20
21
Notes:
a The internal dose metric of u.g oxidized per day per [kg body weight]54 was selected as the basis for interspecies,
intraspecies, and route-to-route extrapolation for the cardiac malformations endpoint in the 2011 TCE IRIS
assessment (EPA, 2011e). The PBPK model was parameterized for human females. HECs are rounded to two
significant figures. Note that standard duration adjustments have been applied to the chronic HECs calculated
from the PBPK model. Specifically, for chronic, nine month, and three week durations, an adjustment for 8/24
hours per day and 5/7 days per week has been applied; for the one day duration, an adjustment of 8/24 hours
per day has been applied. For instance, the median HEC for the 1 day duration was calculated as a single, 8 hour
exposure at 203 u.g/m3, to which a duration adjustment of 8/24 was applied to derive the reported value of 68
M-g/m3.
2.6.2 Human Health Hazard Summary
This section summarizes both cancer and non-cancer hazard information for TCE. The
information was largely taken from the U.S. EPA's TCE IRIS assessment (EPA, 2011e). Regarding
the non-cancer hazard studies, the emphasis is on the repeated-dose oral and inhalation
studies that the TCE IRIS assessment identified as the most appropriate to be carried forward
for the dose-response assessment.
^6,2.1^
The TCE IRIS assessment evaluated data on TCE and its metabolites in a variety of in vitro and in
vivo test systems. TCE does not appear to be a direct-acting mutagen, but has the potential to
bind or induce damage to the structure of DNA or chromosomes. In bacterial test systems, TCE
did not induce mutations unless there was metabolic activation (i.e., the presence of
metabolizing enzymes) (EPA, 2011e).
It is thought that TCE metabolites may be responsible for TCE genotoxicity. TCA, an oxidative
metabolite of TCE, exhibited little, if any, genotoxic activity in vitro. Other TCE metabolites
(DCA, chloral hydrate, DCVG, and particularly DCVC) induced mutations without metabolic
activation (EPA, 2011e). Despite these positive results, uncertainties with regard to the
characterization of TCE genotoxicity remain because not all of TCE metabolites have been
sufficiently tested in the standard genotoxicity screening battery (EPA, 2011e).
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JZ^.2^
The interim acute exposure guideline levels (AEGLs) document for TCE was consulted and used
in this assessment to briefly summarize the acute toxicity data (NAC, 2008). Note that the
EPA/OPPT risk assessment used the developmental studies, but not the acute toxicity studies
described below, for assessing acute risks for reasons explained in the Approach and
Methodology section of the Hazard/Dose-Response assessment. The next section ("Toxicity
Following Repeated Exposures to TCE (Including Cancer)") describes the developmental toxicity
studies that were used for the acute scenarios evaluated in this assessment.
In humans, TCE odors can be detected at concentrations of >50 ppm (HSDB, 1992). It was once
commonly used as an anesthetic agent with concentrations ranging from 5,000 to 15,000 ppm
for light anesthetic use and from 3,500 to 5,000 ppm for use as an analgesic (Parfitt et al.,
1999).
Information on the toxicity of TCE in humans comes from either case reports in the
medical/occupational literature or human inhalation studies. Lethality data in humans have
been reported following accidental exposure to TCE. However, there is insufficient information
about the exposure characterization of these incidents (NAC, 2008).
Human inhalation studies have shown that acute exposure to TCE results in irritation and
central nervous system (CNS) effects in humans. Mild subjective symptoms and nose and throat
irritation were reported by human volunteers exposed to 200 ppm TCE for 7 hrs/day on the first
day of exposure during a 5-day exposure regimen. The study also reported minimal CNS
depression following TCE exposure (Stewart et al., 1970).
CNS depression and effects on neurobehavioral functions were seen in human volunteers
exposed to 1,000 ppm TCE for a 2-hr period (Ferguson and Vernon, 1970; Vernon and
Ferguson, 1969). In the same studies, volunteers were also exposed to 100 or 300 ppm TCE for
2 hrs. Some subjects had similar CNS effects at the middle concentration (300 ppm), with no
such effects observed at the 100 ppm concentration (NAC, 2008). Ettema etal. (1975) also
observed slight to marginal neurobehavioral effects after exposure to 300 ppm TCE for 2.5 hrs.
Cardiac arrhythmias have been reported in humans exposed to high concentration of TCE (NAC,
2008). Several animal studies have reported neurobehavioral effects and the potential for
inducing cardiac sensitization following acute inhalation exposure to TCE. These studies
reporting cardiac effects in humans and animals are not discussed in this assessment and the
reader is referred to the AEGL document (NAC, 2008) to obtain the study descriptions.
^2^6,23^
The studies and corresponding PODs briefly discussed below were those used to evaluate the
risks of acute and repeated (chronic) exposures to TCE-containing degreasers and arts/crafts
products.
EPA/OPPT relied on the systematic study review that the IRIS program conducted as part of the
process of developing the TCE IRIS assessment. Briefly, the IRIS program critically reviewed the
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publicly available animal and human studies that reported adverse cancer and non-cancer
health effects following TCE exposure. In addition, other relevant data such as mechanistic
data, in vitro data, or toxicokinetic data, were reviewed. The review process started with the
identification of primary, peer reviewed literature through January 2011. Multi-disciplinary
teams conducted a systematic review of the study quality of the identified studies using the
principles set forth by the various risk assessment guidelines issued by the National Research
Council and the U.S. EPA (Appendix K) as well as their professional judgment. Data were then
synthesized and integrated to reach hazard conclusions about the biological plausibility of the
relationship between TCE exposure and a particular health effect. Those studies that had
quantitative dose-response information were carried forward into the cancer and non-cancer
dose-response assessments. During the assessment of dose-response data, the IRIS program
adjusted all of the POD values to continuous exposure (i.e., 24 hr/7-day exposures). This
process has been documented in the TCE IRIS assessment, particularly in Chapters 4 and 5 (EPA,
2011e).
2.6.2.3.1 Liver Toxicity (Ineluding Cancer)
Animals and humans exposed to TCE consistently experience liver toxicity. Specific effects
include the following structural changes: increased liver weight, increase in deoxyribonucleic
acid (DNA) synthesis (transient), enlarged hepatocytes, enlarged nuclei, and peroxisome
proliferation. In addition, U.S. EPA concluded that TCE exposure causes liver tumors in mice but
not rats and there is "...minimal support for association between TCE exposure and liver and
gallbladder/biliary cancer" (EPA, 2011e, page 4-238).
For both cancer and non-cancer effects on the liver, the role of metabolites is important but not
well understood. Many investigators have dosed animals with TCE, as well as with many of its
metabolites to determine the role and potency of each in terms of target organ toxicity. It
appears that the oxidation pathway is important for the development of liver toxicity, but the
specific role of each metabolite (i.e., that of TCA, DCA, and chloral hydrate), as well as the
parent TCE, is unclear. As for liver cancer, the TCE IRIS assessment concluded that multiple TCE
metabolites (i.e., and thus pathways) likely contribute to TCE-induced liver tumors (EPA,
2011e).
Human Data
Several human studies (including those in TCE degreaser operations) reported an association
between TCE exposure and significant changes in serum liver function tests used in diagnosing
liver disease, or changes in plasma or serum bile acids (see Table 4-57 in EPA, 2011e for a
summary of the human studies). There was also human evidence for hepatitis accompanying
immune-related generalized skin diseases, jaundice, hepatomegaly, hepatosplenomegaly, and
liver failure in TCE-exposed workers (EPA, 2011e). Cohort studies examining cirrhosis mortality
and either TCE exposure or solvent exposure are generally null, but these studies cannot rule
out an association between TCE and liver disorders/toxicity because of the limitations of the
studies (EPA, 2011e). Overall, while some evidence exists of liver toxicity in humans, the data
are inadequate for making conclusions regarding causality (EPA, 2011e).
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The TCE IRIS assessment evaluated cohort, case-control, and community (geographic) studies
reporting liver and biliary tract cancer, primary liver cancer, and gallbladder and extra-hepatic
bile duct cancer (see Table 4-57 in EPA, 2011e for a summary of the human studies). Most of
these studies have small numbers of exposed cases and controls due to the rarity of liver and
biliary tract cancer. The IRIS program conducted meta-analyses of the studies and reported a
small, statistically significant summary relative risk (RRm) for liver and gallbladder/biliary cancer
with overall TCE exposure (EPA, 2011e). However, the meta-analyses reported a lower,
nonstatistically significant RRm for primary liver cancer when using the highest exposure groups
(EPA. 2011e).
Animal Data
The TCE IRIS assessment reviewed many oral and inhalation studies in rats and mice (see Tables
4-58 and 4-59 in EPA, 2011e for a summary of the animal studies). Animals exposed to TCE
reported increased liver weight, a small, transient increase in DNA synthesis, enlarged
hepatocytes, increased size of nuclei of liver cells, and proliferation of peroxisomes (EPA,
2011e).
The IRIS program determined that the studies of Buben and O'Flaherty (1985); Kjellstrand etal.
(1983) and Woolhiser et al. (2006) were suitable for the dose-response assessment of the liver
health effects domain (Appendix L). These three studies reported increased liver/body weight
ratios.
Kjellstrand etal. (1983) exposed NMRI male mice (10-20/group) with up to nine different TCE
concentrations. These concentrations ranged from 37 to 3,600 ppm and included an air control
group. Exposures were conducted for various durations (1, 2, 4, 8, 16, or 24 hrs/day) and for
different time frames (from 30 to 120 days). The IRIS program calculated a benchmark
concentration lower-bound confidence limit of 21.6 ppm based on the 10% benchmark
response (BMDLio) for increased liver/body weight ratios.
In Woolhiser et al. (2006), Sprague-Dawley female rats (16/group) were exposed to TCE via
inhalation at concentrations of 0,100, 300, or 1,000 ppm for 6 hrs/day, 5 days/week for 4
weeks. A BMDLio of 25 ppm was estimated for increased liver/body weight ratio.
Finally, the Buben and O'Flaherty (1985) exposed Swiss-Cox male mice (12-15 group) to TCE by
gavage. Mice were exposed to a range of TCE doses (100 to 3,200 mg/kg-bw/day plus control)
for 5 days/week for 6 weeks. A BMDLio of 82 mg/kg-bw/day was identified as the POD for
increased liver/body weight ratios.
With respect to liver carcinogenicity, TCE and its oxidative metabolites TCA, DCA, and CH are
clearly carcinogenic in mice, with strain and sex differences in potency. Data in other laboratory
animal species are limited; thus, except for DCA which is carcinogenic in rats, inadequate
evidence exists to evaluate the hepatocarcinogenicity of TCE and its metabolites in rats or
hamsters (EPA. 2011e).
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2^6^^^
Studies in both humans and animals have shown changes in the proximate tubules of the
kidney following exposure to TCE. As for cancer, the TCE IRIS assessment concluded that TCE is
"carcinogenic to humans" based on convincing evidence of a causal relationship between TCE
exposure in humans and kidney cancer. A recent review of TCE by the International Agency for
Research on Cancer (IARC) also supported this conclusion (IARC, 2014).
TCE metabolites appear to be the causative agents that induce renal toxicity, including cancer.
DCVC (and to a lesser extent other metabolites) appears to be responsible for kidney damage
and kidney cancer following TCE exposure (EPA, 2011e). Toxicokinetic data suggest that the TCE
metabolites derived from GSH conjugation (in particular DCVC) can be systemically delivered or
formed in the kidney. Moreover, DCVC-treated animals showed the same type of kidney
damage as those treated with TCE (EPA, 2011e).
The toxicokinetic data and the genotoxicity of DCVC further suggest that a mutagenic mode of
action is involved in TCE-induced kidney tumors, although cytotoxicity followed by
compensatory cellular proliferation cannot be ruled out (EPA, 2011e). As for the mutagenic
mode of action, both genetic polymorphisms (GST pathway) and mutations to tumor
suppressor genes have been hypothesized as possible mechanistic key events in the formation
of kidney cancers in humans (EPA, 2011e).
Human Data
Human studies reported increased excretion of urinary proteins among TCE-exposed workers
when compared to unexposed controls. While some of these studies included subjects
previously diagnosed with kidney cancer, other studies report similar results in subjects who
are disease free. Occupational studies showed increased levels of kidney damage (proximal
tubules) in workers exposed to "high" levels of TCE (EPA, 2011e).
Some additional support for TCE-induced nephrotoxicity in humans came from two studies of
end-stage renal disease (ESRD). Radican et al. (2006) observed a greater incidence of ESRD in
TCE-exposed workers as compared to unexposed controls. Jacob et al. (2007) reported a
greater risk for progression from IgA nephropathy or membranous nephropathy
glomerulonephritis to ESRD following TCE exposure.
As stated previously, the TCE IRIS assessment classified TCE as "carcinogenic to humans" based
on convincing evidence of a causal relationship between TCE exposure in humans and kidney
cancer. The carcinogenic classification was based on a review of more than 30 human studies,
including studies in TCE degreasing operations, and meta-analyses of the cohort and case-
control studies. Relative risk estimates for increased kidney cancer were consistent across a
large number of epidemiological studies of different designs and populations from different
countries and industries (EPA, 2011e). This strong consistency of the epidemiologic data on TCE
and kidney cancer argues against chance, bias, and confounding as explanations for the
elevated kidney cancer risks (EPA, 2011e).
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There appears to be greater susceptibility to TCE-induced kidney cancer in those individuals
that carry an active polymorphism in a gene associated with the GST metabolic pathway.
Particularly, the gene is associated with the 0-lyase gene region which is responsible for
converting DCVC to the unstable intermediate DCVT (EPA, 2011e). Also, there are some human
studies suggesting a role for mutations to the tumor suppressor gene, von Hippel Lindau (VHL
gene). This tumor suppressor gene appears to be inactivated in certain TCE-induced kidney
cancers (EPA, 2011e).
Animal Data
In the animal studies, renal toxicity was evident in both rats and mice following inhalation or
gavage exposures. The toxicity included damage to the renal tubules (e.g., both cytomegaly and
karyomegaly). Under chronic gavage exposure scenarios, rodents exhibited almost 100 percent
kidney toxicity induction. Under inhalation exposure scenarios, male rats were more
susceptible than female rats or mice to kidney toxicity. As noted earlier, this toxicity is likely
caused by DCVC formation, with possible roles for TCOH and TCA (EPA, 2011e).
The IRIS program selected five animal studies reporting kidney toxicity for further non-cancer
dose-response analysis (Appendix L). Maltoni and Cotti (1986), NCI (1976) and NTP (1988)
reported histological changes in the kidney, whereas Kjellstrand et al. (1983) and Woolhiser et
al. (2006) reported increased kidney/body weight ratios (EPA, 2011e).
Maltoni and Cotti (1986) exposed Sprague-Dawley male rats (116-124/group) to TCE via
inhalation (0, 100, 300, or 600 ppm) for 7 hrs/day, 5 days/week for 104 weeks (and allowed all
rats to continue unexposed until they died). The investigators also conducted an oral (gavage)
study that dosed rats with a range of TCE doses (50 to 250 mg/kg-bw/day) for 4-5 days/week
for 52 weeks. BMDLio values of 40.2 ppm and 34 mg/kg-bw/day were calculated for the
inhalation and gavage studies, respectively, based on renal tubular pathological changes (EPA,
2011e).
The National Cancer Institute (NCI) conducted a 90-week oral (gavage) study that dosed B6C3F1
female mice (20-50/group) with a range of TCE doses (869 to 1,739 mg/kg-bw/day) for 5
days/week for 78 weeks (NCI, 1976). Mice were then left unexposed for the final 12 weeks of
the study. A LOAEL of 620 mg/kg-bw/day was identified as the POD for toxic nephrosis (EPA,
2011e).
In another oral (gavage) study, the National Toxicology Program exposed Marshall female rats
(44-50/group) to TCE (i.e., 0, 500, or 1,000 mg/kg-bw/day) for 5 days/week for 104 weeks (NTP,
1988). Rats developed toxic nephropathy following TCE exposure. A BMDL0s of 9.45 mg/kg-
bw/day was calculated for the observed kidney effects (EPA, 2011e).
Woolhiser et al. (2006) conducted an inhalation study that exposed Sprague-Dawley female rats
(16/group) to 0, 100, 300 or 1,000 ppm TCE for 6 hrs/day for 5 days/weeks for 4 weeks. At the
end of the study, rats exhibited increased kidney/body weight ratios and a BMDLio of 15.7 ppm
was estimated for these effects (EPA, 2011e).
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Increased kidney/body weight ratios were also seen in Kjellstrand et al. (1983). NMRI male mice
(10-20/group) were exposed to a range of TCE concentrations (37 to 3,600 ppm) for 30 to 120
days on continuous and intermittent exposure regimens. A BMDLio of 34.7 ppm was identified
as the POD for increased kidney/body weight ratios (EPA, 2011e).
Cancer bioassays with TCE in animals (i.e., both gavage and inhalation exposure routes) did not
show increased kidney tumors in mice, hamsters, or female rats, but did show a slight increase
in male rats. Kidney tumors in rats are relatively rare (EPA, 2011e).
2.6.2.3.3 Neurotoxicity
Neurotoxicity has been demonstrated in animal and human studies under both acute and
chronic exposure conditions (EPA, 2011e). Due to the effects on the nervous system, TCE was
initially synthesized for use as an anesthetic in humans in the early part of the 20th century.
These anesthetic-like effects occurred at high concentrations. A brief summary of these effects
is provided below.
Human Data
Evaluation of the human studies has reported the following TCE-induced neurotoxic effects:
alterations in trigeminal nerve and vestibular function, auditory effects, changes in vision,
alterations in cognitive function, changes in psychomotor effects, and neurodevelopmental
outcomes (EPA, 2011e). Section 2.6.2.3.6 (Developmental Toxicity) discusses the
neurodevelopmental outcomes in more detail.
The strongest neurological evidence of human toxicological hazard is for changes in trigeminal
nerve function or morphology and impairment of vestibular function (EPA, 2011e). Fewer and
more limited epidemiological studies are suggestive of TCE exposure being associated with
delayed motor function, and changes in auditory, visual, and cognitive function or performance,
and neurodevelopmental abnormalities (EPA, 2011e).
Multiple epidemiological studies in different populations have reported TCE-induced
abnormalities in trigeminal nerve function in humans (EPA, 2011e). However, two
epidemiological studies did not report an association between TCE exposure and trigeminal
nerve function. These studies generally had study design limitations such as limited statistical
power, missing control groups, and missing methodology for measuring trigeminal nerve
function (EPA. 2011e).
Among the human studies, Ruijten et al. (1991) was the only epidemiological study that the IRIS
program deemed suitable for further evaluation in the TCE's dose-response assessment for
neurotoxicity. Ruijten et al. (1991) evaluated the TCE exposures and possible health effects of
31 male printing workers (mean age: 44 yrs) and 28 unexposed control subjects (mean age: 45
yrs). The exposure duration was expressed as "cumulative exposure" (concentration x time).
Using historical monitoring data, mean exposures were calculated as 704 ppm x number of
years worked, where the mean number of years was 16 (range: 160-2,150 ppm x yr) (EPA,
2011e). The study measured the trigeminal nerve function by using the blink reflex, but no
abnormal findings were observed. However, the study found a statistically significant average
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increase in the latency response time in TCE-exposed workers on the masseter reflex test,
another test commonly used to measure the integrity of the trigeminal nerve. The POD derived
from the dataset was a LOAEL of 14 ppm (EPA, 2011e).
Human studies have consistently reported vestibular system-related symptoms such as
headaches, dizziness, and nausea following TCE exposure. Although these symptoms are
subjective and self-reported, these effects have been reported extensively in human chamber,
occupational, and geographic-based/drinking water studies (EPA, 2011e).
Animal Data
The TCE IRIS assessment reviewed many animal studies reporting a variety of neurotoxic effects
under different exposure conditions. Animal studies have reported the following TCE-induced
neurotoxic effects: morphological changes in the trigeminal nerve, disruption of the auditory
system, visual changes, structural or functional changes in the hippocampus, sleep
disturbances, changes in psychomotor effects, and neurodevelopmental effects (EPA, 2011e).
Only the following four animal studies were suitable for dose-response analysis for the
neurotoxicity endpoint (Appendix L).
Arito et al. (1994) exposed Wistar male rats (5/group) to TCE via inhalation to concentrations of
0, 50, 100, or 300 ppm for 8 hrs/day, 5 days/week for 6 weeks. Exposure to all of the TCE
concentrations significantly decreased the amount of time spent in wakefulness during the
exposure period. Some carry over was observed in the 22 hr-post exposure period, with
significant decreases in wakefulness seen at 100 ppm TCE. Significant changes in wakefulness-
sleep elicited by the long-term exposure appeared at lower exposure levels. The LOAEL for
sleep changes was 12 ppm (i.e., LOAEL, adjusted for continuous exposure)(EPA, 2011e).
Isaacson et al. (1990) dosed weanling Sprague-Dawley male rats (12/dose group) via the oral
route (drinking water) in an experimental protocol for an 8-week period. The control group had
unexposed rats for 8 weeks. The experimental group#l exposed rats to 47 mg/kg-bw/day TCE
for 4 weeks and then no TCE exposure for 4 weeks. The experimental group#2 exposed rats to
47 mg/kg-bw/day TCE for 4 weeks, no TCE exposure for the following 2 weeks, and then 24
mg/kg-bw/day TCE for the final 2 weeks. Rats in group#3 reported a decreased latency to find
the platform in the Morris water maze test. Also, all of the TCE-treated groups exhibited
hippocampal demyelination. The LOAEL for cognitive effects (i.e., demyelination in the
hippocampus) was 47 mg/kg-bw/day (EPA, 2011e).
Gash et al. (2008) evaluated the effects of TCE on dopamine-containing neurons. F344 male rats
(9/group) were exposed by oral gavage to doses of 0 or 1,000 mg/kg-bw/day TCE for 5
days/week for 6 weeks. Exposed rats showed degeneration of dopamine-containing neurons in
the substantia nigra. The LOAEL for loss of dopamine containing neurons was 710 mg/kg-
bw/day (EPA. 2011e).
Kjellstrand et al. (1987) studied the effect of TCE on the regeneration of the sciatic nerve.
Under heavy anesthesia, the sciatic nerve of NMRI male mice and Sprague-Dawley female rats
was artificially crushed to create a lesion. Prior to the lesion, some animals were pre-exposed to
TCE for 20 days and then for an additional 4 days after the lesion. Another set of animals was
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only exposed to TCE for 4 days following the sciatic nerve lesion. The inhalation exposures to
TCE for the 4 or 20-day exposures were as follows: 0, 150, or 300 ppm TCE for 24 hrs/day for
mice; and 0 or 300 ppm TCE for 24 hrs/day for rats. Both mice and rats exhibited inhibition of
the sciatic nerve regeneration. LOAELs of 150 ppm and 300 ppm were identified as the POD for
the decreased regeneration of the sciatic nerve in mice and rats, respectively (EPA, 2011e).
2^^^
Immune-related effects following TCE exposures have been observed in both animal and
human studies. In general, these effects were associated with inducing enhanced immune
responses as opposed to immunosuppressive effects. Of concern are the immune-related and
inflammatory effects reported in TCE-exposed animals and humans. These effects may
influence a variety of other conditions of considerable public health importance, such as cancer
and atherosclerosis (EPA, 2011e).
Human Studies
Studies have reported a relationship between systemic autoimmune diseases, such as
scleroderma, and occupational exposure to TCE. The TCE IRIS assessment performed a meta-
analysis of a number of human studies evaluating a possible connection between scleroderma
and TCE exposure. Results indicated a significant odds ratio (OR) in men, whereas women
showed a lower but not significant OR. These results may not reflect a true gender difference
because the incidence of this disease is very low in men (approximately one per 100,000 per yr)
and somewhat higher in women (approximately one per 10,000 per yr). In addition, these
results may be affected by gender-related differences in exposure prevalence, the reliability of
the exposure assessment, gender-related differences in susceptibility to TCE toxicity or chance
(EPA. 2011e).
There have been a large number of case reports in TCE-exposed workers developing a severe
hypersensitivity skin disorder, distinct from contact dermatitis, and often accompanied by
systemic effects (hepatitis, lymph nodes, and other organs). These effects appeared after
inhalation exposures ranging from <9 to >700 ppm TCE. Similar effects have been observed in
guinea pigs and mice treated with TCE (EPA, 2011e).
Increased levels of human inflammatory cytokines were measured in an occupational study of
degreasers exposed to TCE (lavicoli et al., 2005). Moreover, similar changes in inflammatory
cytokines were seen in infants exposed to TCE via indoor air. These findings were supported by
studies in auto-immune prone mice (described below) in which short exposures to TCE resulted
in increased levels of inflammatory cytokines (EPA, 2011e).
TCE-related immunosuppression has been reported in one study (Lagakos et al., 1986). This
study found an association between TCE exposure and reported history of bacterial or viral
infections.
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Animal Data
Numerous studies have shown increased autoimmune responses in autoimmune-prone mice,
including changes in cytokine levels similar to those reported in human studies, with more
severe effects, including autoimmune hepatitis, inflammatory skin lesions, and alopecia,
manifesting at longer exposure periods (EPA, 2011e). B6C3F1 mice, a strain that lacks
susceptibility to autoimmune disease have reported immunotoxic effects folio wing TCE
exposure (EPA, 2011e). Furthermore, TCE-exposed guinea pigs and mice have developed
hypersensitivity responses pre- and postnatally following TCE exposure via drinking water. In
addition, evidence of localized immunosuppression has also been reported in mice and rats
(EPA. 2011e).
Only the following four animal studies were suitable for the IRIS' non-cancer dose-response
analysis for the immunotoxicity endpoint (Appendix L).
Keil, D. E. et al. (2009) exposed B6C3F1 mice (10/group), a standard test strain not genetically
prone to develop autoimmune disease, to TCE via drinking water for 27 or 30 weeks at
concentrations in water of 0, 1.4, or 14 ppm (0.35 or 3.5 mg/kg-bw/day). The study reported a
significant decrease in thymus weight concentrations and thymic cellularity as well as an
increase in autoantibodies to ssDNA and dsDNA. A LOAEL of 0.35 mg/kg-bw/day was identified
as the POD for the thymic and autoimmune effects (EPA, 2011e).
Kaneko et al. (2000) exposed auto-immune prone mice (5/group) to TCE at concentrations of 0,
500, 1,000, or 2,000 ppm for 4 hrs/day, 6 days/week, for 8 weeks. At concentrations >500 ppm,
mice exhibited dose-related liver inflammation, splenomegaly and hyperplasia of lymphatic
follicles. Immunoblastic cell formation in lymphatic follicles was observed in mice treated with
1,000 ppm TCE. The LOAEL of 70 ppm was identified for these effects (EPA, 2011e).
In Sanders et al. (1982), male and female CD-I mice (7-25/group) were given TCE in drinking
water concentrations of 0, 0.1, 1.0, 2.5, or 5.0 mg/mL (0, 18, 217, 393 or 660 mg/kg-bw/day) for
4 or 6 months. Female mice showed decreased humoral immunity at 2.5 and 5 mg/mL (393 or
660 mg/kg-bw/day), whereas cell-mediated immunity and bone marrow stem cell colonization
decreased at all four concentrations. Male mice were relatively unaffected after both 4 and 6
months of exposure. A LOAEL of 18 mg/kg-bw/day was identified as the POD for
immunosuppressive effects (EPA, 2011e).
Another study that was previously discussed for liver and kidney effects (Woolhiser et al., 2006)
also reported immunosuppressive effects. Sprague-Dawley female rats (16/group) were treated
with 0, 100, 300 or 1,000 ppm TCE for 6 hrs/day, 5 days/week for 4 weeks. Four days prior to
study termination, the rats were immunized with sheep red blood cells (SRBC), and within 24
hrs following the last exposure to TCE, a plaque-forming cell (PFC) assay was conducted to
determine effects on splenic anti-SRBC IgM response. At 1,000 ppm, rats demonstrated a 64%
decrease in the PFC assay response. A BMDLiso27 of 24.9 ppm was identified for this
immunosuppressive effect (EPA, 2011e).
27 BMDL1SD=the lower-bound confidence limit of the benchmark dose where the effect is 1 standard deviation (SD)
from control value.
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One of the three cancers for which the TCE IRIS assessment based its cancer findings was non-
Hodgkin's lymphoma (NHL) (the other two being kidney and liver cancer) (EPA, 2011e). The
human epidemiological studies strongly support a causal relationship between TCE exposure
and NHL. Further support comes from animal studies reporting rates of lymphomas and/or
leukemias following TCE exposure. For a detailed examination of the cancers of the immune
system, please refer to Chapter 4 of the TCE IRIS assessment (EPA, 2011e).
The toxicological literature provides support for male and female reproductive toxicity
following TCE exposure. Both the epidemiological and animal studies provide suggestive, but
limited, evidence of adverse outcomes to female reproductive outcomes. However, much more
extensive evidence exists in support of an association between TCE exposures and male
reproductive toxicity (EPA, 2011e).
The available human data that associate TCE with adverse effects on male reproductive
function are limited in size and provide little quantitative dose data. However, the animal data
provide strong and compelling evidence forTCE-related male reproductive toxicity. Strengths of
the animal database include the presence of both functional and structural outcomes,
similarities in adverse treatment-related effects observed in multiple species, and evidence that
metabolism of TCE in male reproductive tract tissues is associated with adverse effects on
sperm measures in both humans and animals. Additionally, some aspects of a putative mode of
action (e.g., perturbations in testosterone biosynthesis) appear to have some commonalities
between humans and animals (EPA, 2011e).
The effects of TCE on cancers of the reproductive system have been evaluated in males and
females in both epidemiological and experimental animal studies. However, the association
between TCE exposure and these cancers is generally not robust. Please refer to Chapter 4
(section 4.8.2) of the TCE IRIS assessment for more details about the cancer studies in the
reproductive system (EPA, 2011e).
Human Studies
Most human studies support an association between TCE exposure and alterations in sperm
density and quality, as well as changes in sexual drive or function and serum endocrine levels.
Fewer epidemiological studies exist linking decreased incidence of fecundability (time-to-
pregnancy) and menstrual cycle disturbances in women with TCE exposures (EPA, 2011e).
Among the human studies, Chia et al. (1996) was the only epidemiological study that the IRIS
program deemed suitable for further evaluation in the TCE's dose-response assessment for
reproductive toxicity. Chia et al. (1996) examined a cohort of 85 workers in an electronics
factory. The workers provided urine, blood, and sperm samples.The mean urine TCA level was
22.4 mg/g creatinine (range: 0.8-136.4 mg/g creatinine). In addition, 12 workers provided
personal 8-hr air samples, which resulted in a mean TCE exposure of 29.6 ppm (range: 9-131
ppm). There were no controls in the study. Males experienced decreased percentage of normal
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sperm morphology and hyperzoospermia. A BMDLio of 1.4 ppm was identified as the POD for
these effects (EPA. 2011e).
Animal Data
Laboratory animal studies provide evidence for similar effects, particularly for male
reproductive toxicity. These animal studies have reported effects on sperm, libido/copulatory
behavior, and serum hormone levels, although some studies that assessed sperm measures did
not report treatment-related alterations (EPA, 2011e). Additional studies have observed TCE-
related histopathological lesions in the testes or epididymides, altered in vitro sperm-oocyte
binding or in vivo fertilization due to TCE or metabolites, and reduced fertility (EPA, 2011e).
The reduced fertility effects in male rodents were observed in one study and attributed to
systemic toxicity. However, the total database of reproductive studies suggested that TCE does
induce reproductive toxicity independent of systemic effects (EPA, 2011e).
Fewer animal studies are available for the female reproductive toxicity endpoint. While in vitro
oocyte fertilizability has been reported to be reduced as a result of TCE exposure in rats, a
number of other laboratory animal studies did not report adverse effects on female
reproductive function effects (EPA, 2011e).
Only the following eight reproductive animal toxicity studies were suitable for non-cancer dose-
response analysis in the TCE IRIS assessment (Appendix L).
Xu et al. (2004) exposed male CD-I mice (27/group) to TCE at concentration of 0 or 1,000 ppm
for 6 hrs/day, 5 days/week for 6 weeks. Inhalation exposure to TCE did not result in altered
body weight, testis and epididymis weights, sperm count, or sperm morphology or motility.
Percentages of acrosome-intact sperm populations were similar between treated and control
animals. However, decreased in vitro sperm-oocyte binding and reduced in vivo fertilization
were observed TCE-treated male mice. A LOAEL of 180 ppm was identified as the POD for these
effects (EPA. 2011e).
Kumar et al. (2000) and Kumar et al. (2001) exposed male Wistar rats by inhalation at
concentrations of 0 or 376 ppm TCE. Both study protocols exposed rats for 4 hrs/day, 5
days/week, but had variable duration scenarios. For instance, Kumar et al. (2000) treated rats
for the following exposure durations: 2 weeks (to observe the effect on the epididymal sperm
maturation phase), 10 weeks (to observe the effect on the entire spermatogenic cycle), 5 weeks
with 2 weeks of rest (to observe the effect on primary spermatocytes differentiation to sperm),
8 weeks with 5 weeks of rest (to observe effects on an intermediate stage of spermatogenesis),
or 10 weeks with 8 weeks of rest (to observe the effect on spermatogonial differentiation to
sperm). Kumar et al. (2001) exposed rats for either 12 or 24 weeks.
Kumar et al. (2000) reported altered testicular histopathology, increased sperm abnormalities,
and significantly increased pre- and/or postimplantation loss in litters in the groups with 2 or 10
weeks of exposure, or 5 weeks of exposure with 2 of weeks rest. Multiple sperm effects were
observed in Kumar et al. (2001). After 12 weeks of TCE exposure, rats exhibited decreased
number of spermatogenic cells in the seminiferous tubules, fewer spermatids as compared to
controls, and the presence of necrotic spermatogenic cells. Following 24 weeks of exposure,
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male rates showed reduced testes weights and epididymal sperm count and motility, testicular
atrophy, smaller tubules, hyperplastic Leydig cells, and a lack of spermatocytes and spermatids
in the tubules. Testicular marker enzymes were altered at both 12 and 24 weeks of exposure. A
LOAEL of 45 ppm was identified as the POD for the sperm and male reproductive effects
reported in both studies (EPA, 2011e).
Forkert et al. (2002) exposed male CD-I mice (6/group) by inhalation to 0 or 1,000-ppm TCE for
6 hrs/day, 5 days/week for 19 days over a 4-week period. Mice exhibited sloughing of
epididymal epithelial cells following TCE exposure. A LOAEL of 180 ppm was identified as a POD
for the effects in the epididymis epithelium in male rats (EPA, 2011e).
Kan et al. (2007) also provided evidence for the damage to the epididymis epithelium and
sperm. CD-I male mice (4/group) were exposure by inhalation to 0 or 1,000-ppm TCE for 6
hrs/day, 5 days/week for 1 to 4 weeks. As early as 1 week after TCE exposure, exposed mice
showed degeneration and sloughing of epithelial cells. These effects increased in severity at 4
weeks of exposure. A LOAEL of 180 ppm was identified as a POD for the effects in the
epididymis epithelium which is consistent with the results from (EPA, 2011e); Forkert et al.
(2002).
DuTeaux et al. (2004) conducted a drinking water study that treated two strains of male rats
(Sprague-Dawley or Simonson albino; 3/group) with 0, 0.2%, or 0.4% TCE (v/v) (0, 143, or 270
mg/kg-bw/day) in a solution of 3% ethoxylated castor oil for 14 days. These TCE concentrations
were within the range of those reported in formerly contaminated drinking water wells. Cauda
epididymal and vas deferens sperm from treated males were incubated in culture medium with
oviductal cumulus masses from untreated females to assess in vitro fertilization capability.
Results showed a dose-dependent decreased fertilization in both rat strains with a LOAEL of
141 mg/kg-bw/day identified as the POD for the observed effects (EPA, 2011e).
Narotsky et al. (1995) administered TCE to F344 timed-pregnant rats (8-12 dams/group) by
gavage. Dams were exposed to TCE doses of 0, 10.1, 32, 101, 320, 475, 633, 844 or 1125 mg/kg-
bw/day during gestational days (GD) 6 to 15. The study was a prequel to a complicated protocol
with other chemicals in a mixture study. Delayed parturition was observed at >475 mg/kg-
bw/day. The LOAEL for female reproductive effects was 475 mg/kg-bw/day (EPA, 2011e).
Narotsky et al. (1995) exposed F344 male and female rats to TCE via the diet at estimated doses
of 0, 72,186, or 389 mg/kg-bw/day. Male and female animals were treated for one week pre-
mating and then for 13 weeks. Pregnant rats were continued on TCE-treated diet throughout
gestation. Results showed a decrease in mating in both sexes and the LOAEL of 389 mg/kg-
bw/day was used as the POD for consideration in the TCE IRIS' dose-response assessment (EPA,
2011e).
2.6.2.3.6 Developmental Toxicity
An evaluation of the overall weight and strength of the evidence of the human and animal
developmental toxicity data suggests an association between pre- and/or postnatal TCE
exposures and potential developmental adverse outcomes. TCE-induced heart malformations in
animals have been identified as the most sensitive developmental toxicity endpoint in the TCE
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IRIS' dose-response analysis. This information is briefly discussed below, including a summary of
the weight of evidence for the fetal cardiac malformations observed in animal data (EPA,
2011e).
EPA/OPPT relied on the PBPK-derived HECs reported for the developmental animal studies
reporting fetal cardiac defects to estimate acute risks for exposures to TCE-containing
degreasers, dry cleaning spotting agents, and arts/crafts products. This approach is consistent
with EPA's current policy that a single exposure of a chemical at a critical window of fetal
development may produce adverse developmental effects (EPA, 1991).
Human Studies
The TCE IRIS assessment evaluated numerous human studies that examined the possible
association of TCE with various developmental outcomes, including prenatal (e.g., spontaneous
abortion and perinatal death, decreased birth weight, and congenital malformations) and
postnatal (e.g., growth, survival, developmental neurotoxicity, developmental immunotoxicity,
and childhood cancers) effects. Most of these studies represent workplace exposures [e.g.,
Finnish studies of Taskinen et al. (1989) and Taskinen et al. (1994)1. In addition, geographically-
based epidemiological studies have been conducted in various parts of the United States,
including Arizona (Tucson Valley), Colorado (Rocky Mountain Arsenal), Massachusetts, New
York (Endicott), Camp Lejeune, North Carolina and Milwaukee, Wisconsin.
The Endicott, New York, and the Camp Lejeune studies focused on reproductive and
developmental outcomes (ATSDR, 1998, 2006, 2008). The Camp Lejeune studies have been the
subject of an NAS investigation (NRC, 2009). Some of these studies have reported associations
between parental exposure to TCE and spontaneous abortion or perinatal death, and decreased
birth weight. However, other occupational and geographically-based studies have failed to
detect a positive association between TCE exposure and developmental toxicity in humans
(EPA, 2011e). Note that none of these studies were suitable for dose-response analysis in the
TCE IRIS assessment due to study limitations (e.g., small sample size, co-exposures with other
chemicals, lack of exposure levels) (EPA, 2011e).
There have been some epidemiological studies that have consistently reported an increased
incidence of birth defects in TCE-exposed populations. For instance, ATSDR has conducted
studies at Camp Lejeune, North Carolina, where individuals were exposed to VOC-contaminated
drinking water [e.g., ATSDR (1998), Ruckartetal. (2013)1. TCE was one of the main
contaminants found in the drinking water.
Ruckart et al. (2013) recently conducted a case control study to determine if children born from
mothers exposed to contaminated drinking water during pregnancy at Camp Lejeune were
more likely to develop childhood hematopoietic cancers, neural tube defects, or oral clefts. The
study found an association between neural tube defects and TCE exposure above 5 ppb during
the first trimester of pregnancy (i.e., OR of 2.4; 95% confidence limit: 0.6-9.6).
Yauck et al. (2004) conducted a small case-control study of 245 cases and 3,780 controls in
Milwaukee, Wisconsin. The study used a geographic information system to estimate distances
between maternal residences and facilities emitting TCE emissions. The study observed a strong
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relative risk estimate of 6.2 (95% Cl: 2.6, 14.5) for cardiac defects in infants born to mothers
aged 38 years or older after controlling for potential confounding. No association for cardiac
defects was observed among infants of mothers aged less than 38 years (RR = 0.9, 95% Cl: 0.6,
1.2).
Since the publication of the 2011 TCE IRIS assessment, one recent update for the Endicott, NY
community was published by the NY State Health Department that evaluated maternal
exposure to TCE and other VOCs and pregnancy outcome (Forand et al., 2012). The study
evaluated all births recorded in Endicott, NY from either 1978 to 2002 (to assess low birth
weight, pre-term and fetal growth) or from 1983 to 2000 (birth defects). The comparison group
was the rest of NY State except for the city of New York. A large chemical spill occurred in the
town in 1979, and monitoring of the contaminant plume occurred for years. Residents obtained
their drinking water from an uncontaminated water source. However, TCE and other VOCs have
been measured in groundwater, soil, and inside buildings, the latter due largely to vapor
intrusion. The study authors reported significant adjusted rate ratios (RRs) for the TCE-
contaminated area for, among others, the following endpoints: low birth weight (RR of 1.36; 95
percent confidence interval (Cl) of 1.07 to 1.73), small for gestational age [RR of 1.23; 95
percent (Cl) of 1.03 to 1.48], and cardiac defects (RR of 2.15; 95 percent Cl of 1.27 to 3.62).
Other studies have addressed the potential immunological effects in children exposed to TCE.
Lehmann et al. (2001) studied the relationship between indoor VOC exposure and the risk of
atopy in premature neonates and 36-month-old neonates. The study found no association
between TCE exposure and allergic sensitization to egg white and milk, or to cytokine producing
peripheral T-cells. However, Lehmann et al. (2002) reported a significant reduction in Thl IL-2
producing cells in exposed newborns.
As for human developmental neurotoxicity, the available studies collectively suggest that the
developing brain is susceptible to TCE toxicity. These studies have reported an association with
TCE exposure and CNS birth defects and postnatal effects such as delayed newborn reflexes,
impaired learning or memory, aggressive behavior, hearing impairment, speech impairment,
encephalopathy, impaired executive and motor function and attention deficit (ATSDR, 2001;
Bernad etal., 1987; Bove, 1996; Boveetal., 1995; Burg and Gist, 1997; Lagakos et al., 1986;
White et al., 1997). These studies have many limitations; thus the reported associations must
be interpreted with caution (EPA, 2011e).
Leukemia and CNS cancers during childhood have been observed in a number of studies in
children exposed to TCE. However, other studies have not confirmed the increased risk for
childhood leukemia and CNS cancers (EPA, 2011e).
Animal Data
Many of the TCE-related developmental effects reported in humans have been observed in
animal studies: pre- or post-implantation losses, increased resorptions, perinatal death,
decreased birth weight, and congenital anomalies. Some of these effects appear to be strain-
specific. Overall, based on weakly suggestive epidemiologic data and fairly consistent
laboratory animal data, it can be concluded that TCE exposure poses a potential hazard for
prenatal losses and decreased growth or birth weight of offspring effects (EPA, 2011e).
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The TCE IRIS assessment found 5 animal studies that were suitable for non-cancer dose-
response analysis for the following developmental outcomes: pre- and postnatal mortality;
pre-and postnatal growth; developmental neurotoxicity; and congenital heart malformations
(Appendix L).
Although the focus of the discussion below is on these 5 studies and corresponding endpoints,
it is important to mention that developmental immunotoxicity has been shown in TCE-treated
animals. The most sensitive immune system response was reported by Peden-Adams et al.
(2006). In this study, B6C3F1 mice were exposed to TCE via drinking water. Treatment occurred
during mating and through gestation to TCE levels of 0, 1.4, or 14 ppm. After delivery, pups
were further exposed for either 3 or 8 more weeks at the same concentration levels that the
dams received in drinking water. Suppressed RFC response was seen in male pups after 3 and 8
weeks of exposure, whereas female pups showed the suppression of RFC response and delayed
hypersensitivity at 1.4 ppm following 8 weeks. At the higher concentration (14 ppm), both of
these effects were observed again in both males and females following 3 or 8 weeks of
postnatal exposure. A LOAEL of 0.37 mg/kg-bw/day served as a POD for the decreased RFC and
increased delayed hypersensitivity responses (EPA, 2011e).
~ Pre- and Postnatal Mortality and Growth
The following two studies were suitable for non-cancer dose-response analysis for pre- and
postnatal mortality and growth effects. Healy et al. (1982) exposed female Wistar rats (31-32
dams/group) to TCE via inhalation at concentrations of 0 or 100 ppm for 4 hrs/day during GD 8
to 21. Study reported increased resorptions in dams exposed to 100 ppm. After adjusting to a
continuous 24-hr exposure, the LOAEL of 17 ppm was identified and used as the POD in the
dose-response analysis. The same study also reported reduced fetal weight at 100 ppm
(adjusted LOAEL of 17 ppm) (EPA. 2011e).
Narotsky et al. (1995) was the other study in which a POD was identified for mortality in the
developing fetus. This study was mentioned above in the reproductive toxicity section. F344
timed-pregnant rats (8-12 dams/group) were treated with TCE by gavage during GD 6 to 15.
The BMDLoi for resorptions was 32.2 mg/kg-bw/day (EPA, 2011e).
~ Developmental Neurotoxicity
There is evidence of alterations in animal brain development and in behavioral parameters
(e.g., spontaneous motor activity and social behaviors) following TCE exposure during the
development of the nervous system. Among all of the available studies, there were two oral
studies that reported behavioral changes which were used in the dose-response evaluation for
developmental toxicity.
Fredriksson et al. (1993) treated male NMRI mouse pups (12/group, selected from 3-4 litters)
with TCE via gavage (0, 50, or 290 mg/kg-bw/day) during postnatal days (PND) 10 to 16.
Locomotor behavior was evaluated at PND 17 and 60. TCE-treated mice showed decreased
rearing activity at both dose levels on PND 60, but not PND 17, resulting in a LOAEL of 50
mg/kg-bw/day as a POD (EPA. 2011e).
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Taylor et al. (1985) conducted a drinking water study where pregnant Sprague-Dawley rats
were given TCE at concentrations of 0, 312, 625, or 1,250 mg/L (0, 45, 80 or 140 mg/kg-
bw/day). Exposure occurred 14 days prior to breeding and from GD 0 to PND 21. The number of
litters/group was not reported, nor did the study state how many pups per litter were
evaluated for behavioral parameters. Exploratory behavior was measured in the pups on PND
28, 60, and 90, whereas wheel-running, feeding, and drinking behavior were monitored on
PNDs 55-60. TCE-treated male showed increased exploratory behavior on PND 60 and 90 at all
dose levels, with the largest effect observed at the highest dose level. The LOAEL for this effect
was 45 mg/kg-bw/day (EPA, 2011e).
Blossom et al. (2012) and Blossom et al. (2013) found effects of developmental exposure to
TCE via drinking water (0, 2, or 28 mg/kg/day). MRL mice were maternally exposed from birth
through weaning. Subsequently, males only were exposed directly via drinking water from PND
21 (weaning) to PND 42. The study reported alterations in brain neurotrophin expression,
glutathione redox homeostasis, DNA hypomethylation and a number of behavioral parameters,
such as increased motor activity, and novelty/exploratory behavior at the highest dose tested
(28 mg/kg/day). The NOAEL for neurobehavioral impairments was 2 mg/kg/day with a BMDLiso
ranging from 14-20 mg/kg/day depending upon the neurobehavioral endpoint (Appendix M)28.
~ Congenital Heart Defects
In vivo animal studies in rats and chicks have identified an association between TCE exposures
and cardiac defects in the developing embryo and/or fetus. Mechanistic studies have also
examined various aspects of the induction of cardiac malformations. The critical window for
cardiac development is 1-2 weeks for rodents, 1-2 weeks for chickens, and from the 3rd to the
8th week for the human fetus. As discussed above, human studies have also reported increased
risk of cardiac defects following TCE exposure. Taken together, after evaluating both positive
and negative findings, the TCE IRIS assessment concluded that TCE exposure poses a potential
hazard for congenital malformations, including cardiac defects in offspring. This conclusion is
based on the weakly suggestive epidemiological data in combination with the findings of the
animal and mechanistic studies (EPA, 2011e).
The scientific literature also has examples of well-conducted studies in rats, mice, or rabbits
that have failed to provide evidence forTCE-induced cardiac malformations. It is postulated
that the differences in response across studies may be partially attributed to experimental
design differences (EPA, 2011e).
The fetal cardiac defects reported in Dawson et al. (1990), Dawson et al. (1993), Johnson et al.
(2003), Johnson et al. (2005) and Johnson (2014) were identified as the most sensitive endpoint
within the developmental toxicity domain and across all of the health effects domains
evaluated in the TCE IRIS assessment. Johnson et al. (2003) reported data from different
experiments over a several-year period in which pregnant Sprague-Dawley rats (9-13/group; 55
in control group) were exposed to TCE via drinking water at concentrations of 0, 0.00045,
28 EPA conducted BMD analysis using the dose-response data reported in the Blossom et al. study (2013). For
further information on the BMD analysis, please refer to Appendix M and supplemanty files:
l_Blossom_NObject_SD.xlsm and 2_Blossom_NMouse_SD.xlsm).
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0.048, 0.218 or 129 mg/kg-bw/day. Treatment of pregnant rats occurred during the entire
gestational period (i.e., GD 0 to GD22). The study reported a statistically and biologically
significant increase in the formation of heart defects at the 0.048 mg/kg-bw/day dose level at
both the individual fetus level and the litter level. There was a statistically significant increase in
the percentage of abnormal hearts and the percentage of litters with abnormal hearts at 0.048
mg/kg-bw/day and higher dose levels29. A BMDL0i of 0.0207 mg/kg-bw/day was identified as
the POD for heart malformations (EPA, 2011e).
The fetal cardiac findings in the Johnson et al. studies have been controversial due to
limitations in their study design, data reporting issues and inconsistencies with the database.
EPA/OPPT received a number of comments from the public and the peer review panel objecting
to the use of the Johnson et al. studies based on methodological issues and the fact that the
findings have not been replicated in other animal and human studies. A recent erratum
(Johnson, 2014) and subsequent evaluation of the developmental toxicity data reaffirmed that
the Johnson et al. studies are adequate to use in hazard identification and dose-response
assessment (Appendix M). While the Johnson et al. studies have limitations, there is insufficient
reason to dismiss their findings, especially when the findings are analyzed in combination with
the remaining body of human, animal and mechanistic evidence (EPA, 2011e). Appendix N
discusses the weight-of-evidence analysis supporting the association of TCE exposure and fetal
cardiac malformation.
~ Summary of Weight-of-Evidence Analysis for Congenital Heart Defects
TCE exposure has been associated with cardiac malformations in chick embryos studies (Boyer
etal.. 2000: Brossetal.. 1983: Drake. V. et al.. 2006: Drake. V. J. et al.. 2006: Loeberetal.. 1988:
Mishima et al., 2006; Rufer et al., 2008) and oral developmental toxicity studies in rats (Dawson
etal., 1990, 1993; Johnson et al., 2005; Johnson, 2014; Johnson etal., 2003). In addition to the
consistency of the cardiac findings across different species, the incidence of congenital cardiac
malformation has been duplicated in several studies from the same laboratory group and has
been shown to be TCE-related (EPA, 2011e).
TCE metabolites have also induced cardiac defects in developmental oral toxicity studies
(Epstein etal.. 1992: Johnson et al.. 1998a. 1998b: Smith etal.. 1989. 1992). For example, the
Johnson et al. and Smith et al. studies reported increased incidences of cardiac malformation
following gestational TCA exposures (Johnson etal., 1998a, 1998b; Smith etal., 1989).
Similarly, pregnant rats exhibited increased incidence of cardiac defects following DCA exposure
during pregnancy (Epstein etal., 1992; Smith etal., 1992).
A number of studies have been conducted to elucidate the mode of action for TCE-related
cardiac teratogenicity. During early cardiac morphogenesis, outflow tract and atrioventricular
endothelial cells differentiate into mesenchymal cells (EPA, 2011e). These mesenchymal cells
have characteristics of smooth muscle-like myofibroblasts and form endocardial cushion tissue,
which is the primordia of septa and valves in the adult heart (EPA, 2011e). Many of the cardiac
29 The EPA Science Advisory Board EPA(2011c) reviewed these data and suggested that the IRIS program use the
Johnson et al. (2003) study as one of the principal studies for Rf D/RfC derivation with the critical effect of
cardiac malformations.
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defects observed in humans and laboratory species involved septal and valvular structures
(EPA, 2011e). Thus, a major research area has focused on the disruptions in cardiac valve
formation in avian in ovo and in vitro studies following TCE treatment. These mechanistic
studies have revealed TCE's ability to alter the endothelial cushion development, which could
be a possible mode of action underlying the cardiac defects involving septal and valvular
morphogenesis in rodents and chickens (EPA, 2011e). These mechanistic data provide support
to the plausibility of TCE-related cardiac effects in humans (EPA, 2011e).
Other modes of actions may also be involved in the induction of cardiac malformation following
TCE exposure. For example, studies have reported TCE-related alterations in cellular Ca2+ fluxes
during cardiac development (Caldwell etal., 2008; Collier et al., 2003; Selmin etal., 2008).
2.6.2.4 Summary of Hazard Studies Used to Evaluate Acute and Chronic Exposures
Table 2-26 summarizes the hazard studies, health endpoints by target organ/system, PBPK-
derived HEC99and UFs that are relevant for the risk evaluation of acute and chronic exposure
scenarios. Risk estimates were estimated for a range of HECs, but the HEC99 was preferred since
it is considered protective for a sensitive individual as discussed in the TCE IRIS assessment
(EPA, 2011e). Appendix L contains the complete list of oral and inhalation non-cancer studies
within each health effects domain that the TCE IRIS assessment considered suitable for dose-
response analysis.
TCE and its metabolites are associated with adverse effects on cardiac development based on a
weight-of-evidence analysis of developmental studies from rats, humans and chickens.These
adverse cardiac effects are deemed important for acute and chronic risk estimation for the
scenarios and populations addressed in this risk assessment. The rationale for using TCE
associated fetal cardiovascular lesions for acute scenarios is based on the relatively short critical
window of vulnerability in humans, rodent and avian cardiac development.The rationale for
using fetal cardiac effects for chronic risks estimation is also based on the fact that relatively
low dose short term/acute exposures can result on long-term adverse consequences on cardiac
development persisting into adulthood.
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Table 2-26. Summary of Hazard Information Used in the Risk Evaluation of Acute and Chronic Scenarios
Exposure
Duration for
Risk Analysis
CHRONIC
ACUTE OR
CHRONIC
Target Organ/
System
Liver
Kidney
Nervous
System
Immune
System
Reproductive
System
Developmental
effects
Species
Mouse
(male)
Rat
(female)
Rat
(male)
Mouse
(female)
Mouse
(female)
Human
(male)
Rat
(female)
Route of
Exposure
Inhalation
Oral (gavage)
Inhalation
Oral (drinking
water)
Oral (drinking
water)
Inhalation
Oral
(drinking
water)
POD Type
BMDL10 - 21 6
ppm
BMDL05 = 9.45
mg/kg-bw/day
LOAEL - 12
ppm
LOAEL = 0.35
mg/kg-bw/day
LOAEL -0.35
mg/kg-bw/day
BMDL10 - 1 4
ppm
BMDL01 =
0.0207 mg/kg-
bw/day
Effect
Increased liver/body
weight ratio
Toxic nephropathy
Significant decreases in
wakefulness
Decrease in thymus
weight and thymus
cellularity
Autoimmunity
(increased anti-dsDNA
and ssDNA antibodies)
Decreased normal
sperm morphology and
hyperzoospermia
Heart malformations
HEC99
(ppm)
9.1
0.0056
4.8
0.033
0.033
0.5
0.0037
Total
Uncertainty
Factor (UF) for
Benchmark
MOE
Total UF=10
Total UF=10
Total UF=300
Total UF=100
Total UF=30
Total UF=30
Total UF=10
Reference
Kjellstrand
etal. (1983)
NTP(1988)
Arito et al.
(1994)
Keil etal.
(2009)
Keil. D. E. et
al. (2009)
Chiaetal.
(1996)
Johnson et
al. (2003)
Notes:
1 1 ppm = 5.374 mg/m3
2 Two different effects were reported by Keil et al, (2009): decreased thymic weight and cellularity and autoimmunity. A
total UF of 100 was used for the thymus toxicity, whereas a total UF of 30 was used for the autoimmune effects. The
TCE IRIS assessment allocated different LOAEL-to-NOAEL uncertainty factors (UFL) based on the severity of the effects,
which resulted in different total UFs for effects repoted by the same study (EPA, 2011e).
Other adverse non-cancer effects are deemed relevant for risk estimations for other
populations for the chronic scenarios, but may not occur at the lowest exposures found with
cardiac defects and they include kidney toxicity, immunotoxicity, male reproductive toxicity,
neurotoxicity, liver toxicity and cancer of kidney, liver and immune system (NHL).
TCE is carcinogenic to humans. The cancer risk assessment uses the IUR derived in the 2011
TCE IRIS assessment based on human kidney cancer. The weight-of-evidence analysis for the
cancer endpoint was sufficient to conclude that TCE operates through a mutagenic mode of
action for kidney tumors (EPA, 2011e).
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2.7.1 Risk Estimation Approach for Acute and Repeated Exposures
Tables 2-27 and 2-28 show the use scenarios, populations of interest and toxicological
endpoints used in the acute and chronic risk assessment, respectively.
Table 2-27. Use Scenarios, Populations of Interest and Toxicological Endpoints for Assessing
Acute Risks to TCE-containing Degreasers, Spotting Agents and Arts/Crafts Products
Use
Scenarios
Populations
And Toxicological
Approach
(1)
Solvent degreasing
at small commercial
facilities
(2)
Spot cleaning at dry
cleaning facilities
(3)
Residential use of
degreaser product
Adult pregnant
consumers (>16yrs
old) exposed to TCE
for a single 1-hr
exposure when
using product 2X
per month 2'3'4.
Residential
exposure model
provided the 24-hr
acute exposure
estimate.
(4)
Residential use of
arts/crafts clear
protective coating
spray
Population of Interest and
Exposure Scenario:
Users
Adult pregnant
worker (>16 years
old) exposed to TCE
for a single 2-hr
exposure during an
8-hr workday2'3'4.
Adult pregnant
worker (>16 years
old) exposed to TCE
for a single 8-hr
exposure 2'3.
Adult pregnant
consumers (>16 yrs
old) exposed to TCE
for a single 0.5-hr
exposure when
using product once
per week2'3'4.
Residential
exposure model
provided the 24-hr
acute exposure
estimate.
Population of Interest and
Exposure Scenario:
Bystander
Adult pregnant women (>16 years old)
exposed to TCE indirectly by being in the
same building.
Adult pregnant bystander and individuals
of several age groups that are exposed to
indirect TCE exposures by being in the rest
of the house.
Health Effects of Concern,
Concentration and Time
Duration
Non-Cancer Health Effects: Fetal cardiac defects (Johnson et al., 2005; Johnson, 2014;
Johnson etal.,2003)5
1. PBPK-derived Non-Cancer Hazard values or Point of Departures (PODs) (EPA, 2011e):
24-hr HEC50: 0.012 ppm
24-hr HEC95: 0.0051 ppm
24-hr HEC99: 0.0037 ppm
Cancer Health Effects: Acute cancer risks were not estimated. Relationship is not known
between a single short-term exposure to TCE and the induction of cancer in humans.
Uncertainty Factors (UF)
used in Non-Cancer
Margin of Exposure (MOE)
calculations
(UFs=l)x(UFA=3)x(UFH=3)x(UFL=l)b = 10 (EPA, 2011e).
Total UF=Benchmark MOE=10
Notes:
1 The acute risk assessment focused on the most sensitive life stage in humans, which is women of childbearing age and fetus (i.e.,
pregnant worker) due to concerns for developmental effects.
2 Exposure estimate was adjusted to a 24-hr exposure estimate in order to combine it with the 24-hr HECs.
3 It is assumed no substantial buildup of TCE in the body between exposure events due to TCE's short biological half-life (~51 hrs).
4 EPA/OPPT believes that the users of these products are generally adults, but teenagers and even children may be users or be in the
same room with the user while engaging in arts and crafts projects or degreasing.
5 The acute risk assessment focused on developmental toxicity effects as the most sensitive health effect when compared to other
potential acute effects (i.e., neurotoxicity).
6 UFs=subchronic to chronic UF; UFA=interspecies UF; UFH=intraspecies UF; UFL=LOAEL to NOAEL UF
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Table 2-28. Use Scenarios, Populations of Interest and Toxicological Endpoints for Assessing
Chronic Risks to TCE-containing Degreasers and Spotting Agents
Use Scenarios
Populations
And Toxicological
Approach
(1)
Solvent degreasing at small
commercial facilities
(2)
Spot cleaning at dry cleaning
facilities
Population of Interest and
Exposure Scenario:
Users
Adult worker (>16 years old)1
exposed to TCE for 2-hr exposure
during an 8-hr workday for 260 days
per year for 40 working years.
Lifetime average daily concentration
was calculated. Note that the 8-hr
exposure estimate was adjusted to a
24-hr exposure estimate.
Adult worker (>16 years old)1
exposed to TCE for an 8-hr workday
for 260 days per year for
40 working years.
Lifetime average daily concentration
was calculated. Note that the 8-hr
exposure estimate was adjusted to a
24-hr exposure estimate.
Population of Interest and
Exposure Scenario:
Bystander
Adult worker (>16 years old)1 repeatedly exposed to indirect TCE exposures
by being in the same building.
Health Effects of Concern,
Concentration and Time
Duration
Non-Cancer
2. Non-cancer health effects: A range of possible chronic non-cancer
effects in liver, kidney, nervous system, immune system, reproductive
system and developmental effects2
3. PBPK-derived Non-Cancer Hazard values or Point of Departures (PODs):
The lowest POD (i.e., 24-hr HEC50, HEC95 or HEC99 expressed in ppm)
within each health endpoint domain (EPA. 2011e). See Table 2-18.
Cancer
1. Cancer health effects: Possible cancer effects in kidney, liver or non-
Hodgkins lymphoma from chronic exposure.
2. PBPK-derived Cancer Inhalation Unit Risk (IUR): 2 x 10"2 per ppm
(EPA. 2011e).
Uncertainty Factors (UF)
used in Non-Cancer
Margin of Exposure
(MOE) calculations
Study- and endpoint-specific UFs from the TCE IRIS assessment. See Table 2-
18.
Notes:
Adult workers (>16 years old) include both healthy female and male workers.
2 The chronic risk assessment for developmental effects focused on the most sensitive life stage in humans, which
are women of child-bearing age and fetus (i.e., pregnant worker). For other health effects (e.g., liver, kidney, etc.),
healthy female or male workers were assumed to be the population of interest.
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Acute or chronic MOEs (MOEacute or MOEchronic) were used in this assessment to estimate non-
cancer risks (Table 2-29).
Table 2-29. Equation to Calculate Non-Cancer Acute or Chronic Risks Using Margin of
Exposures
MOE acute or chronic = Non-cancer Hazard value (POD)
Human Exposure
MOE =
Hazard value (POD) =
Human Exposure =
Margin of exposure (unitless)
HEC50, HEC95, or HEC99 (ppm) derived from TCE IRIS assessment (EPA, 2011e)
Exposure estimate (in ppm) from occupational or consumer exposure
assessment. ADCs were used for non-cancer chronic risks and acute
concentrations were used for acute risks (see sections 2.3.3 and 2.4.2).
As discussed previously, each non-hazard PBPK-derived POD was adjusted by endpoint/study-
specific UFs as described in the TCE IRIS assessment (EPA, 2011e). These UFs accounted for (1)
the variation in susceptibility among the members of the human population (i.e., inter-
individual or intraspecies variability); (2) the uncertainty in extrapolating animal data to humans
(i.e., interspecies uncertainty); (3) the uncertainty in extrapolating from data obtained in a
study with less-than-lifetime exposure (i.e., extrapolating from subchronic to chronic exposure);
and (4)the uncertainty in extrapolating from a LOAEL rather than from a NOAEL (EPA, 2011e).
The total UF for each non-cancer PBPK-derived POD was the benchmark MOE used to interpret
the MOE risk estimates for each use scenario. The MOE estimate was interpreted as human
health risk if the MOE estimate was less than the benchmark MOE (=total UF). On the other
hand, the MOE estimate indicated negligible concerns for adverse human health effects if the
MOE estimate exceeded the benchmark MOE. Typically, the larger the MOE, the more unlikely
it is that a non-cancer adverse effect would occur.
Cancer risks for repeated exposures to TCE were estimated using the equation in Table 2-30.
Estimates of cancer risks should be interpreted as the incremental probability of an individual
developing cancer over a lifetime as a result of exposure to the potential carcinogen (i.e.,
incremental or excess individual lifetime cancer risk).
Table 2-30. Equation to Calculate Cancer Risks
Risk= Human Exposure x IUR
Risk =
Human exposure =
IUR =
Cancer risk (unitless)
Exposure estimate (LADC in ppm) from occupational exposure assessment
Inhalation unit risk (2 x 10"2 per ppm) (EPA. 2011e)
Page 103 of 212
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2.7.2 Acute Non-Cancer Risk Estimates for Inhalation Exposures to
TCE
The acute inhalation risk assessment used developmental toxicity data to evaluate the acute
risks for the TSCA TCE use scenarios. As indicated previously, EPA's policy supports the use of
developmental studies to evaluate the risks of acute exposures. This policy is based on the
presumption that a single exposure of a chemical at a critical window of fetal development, as
in the case of cardiac development, may produce adverse developmental effects (EPA, 1991).
After evaluating the developmental toxicity literature of TCE, the TCE IRIS assessment
concluded that the fetal heart malformations are the most sensitive developmental toxicity
endpoint associated with TCE exposure (EPA, 2011e). Thus, EPA/OPPT based its acute risk
assessment on the most health protective endpoint (i.e., fetal cardiac malformations; Johnson
et al., 2003) representing the most sensitive human population (i.e., adult women of child-
bearing age and fetus > 16 yrs).
The acute risk assessment used the PBPK-derived hazard values (HEC50, HEC95, or HEC99) from
Johnson et al. (2003) developmental study for each degreaser and spot cleaner use scenario.
Note that the variability among these hazard values is small and no greater than 3-fold (i.e., 2-
fold for HEC5o/HEC95 ratios; 3-fold for HEC50/HEC99 ratios; 1.4-fold for HEC95/HEC99 ratios).
Acute inhalation risks were reported for most occupational and residential exposure scenarios
based on concerns for developmental effects, irrespective of who is using the product (user vs.
bystander), the type of exposure (typical vs. worst case scenario) and the room ventilation
system (LEV vs no LEV). For instance, most of the degreaser and spot cleaner exposure
scenarios and all of the residential use scenarios reported MOE values below the benchmark
MOE of 10 irrespective of the percentile HEC value used to estimate the MOEs. Only one use
scenario reported MOEs above 10, which was the spot cleaner use scenario representing
bystander exposures under typical occupational exposure levels with LEV (i.e., "Bystander +
LEV-Typical Exposure") (Tables 2-31, 2-32 and 2-33).
Page 104 of 212
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Table 2-31. Acute Non-Cancer Risk Estimates for Commercial Use of Degreaser Product at Small Shops (Developmental
Effects: Congenital Heart Malformations, Johnson et al., 2003)
Lowest PBPK-
derived
HECs (ppm)
of the
developmental
toxicity health
domain
HEC50= 0.012
HEC95= 0.0051
HEC99= 0.0037
WORKER NON-CANCER MOEs
With LEV—
Low-end
exposure
estimate
O.12
0.051
0.037
With LEV—
Upper-end
exposure
estimate
O.OO18
0.0008
0.0006
No LEV—
Low-end
exposure
estimate
O.O12
0.0051
0.0037
No LEV—
Upper-end
exposure
estimate
O.OOO18
0.00008
0.00006
BYSTANDER NON-CANCER MOEs
With LEV—
Low-end
exposure
estimate
O.9
0.38
0.28
With LEV—
Upper-end
exposure
estimate
O.OO2
0.0009
0.0007
No LEV—
Low-end
exposure
estimate
O.O9
0.038
0.028
No LEV—
Upper-end
exposure
estimate
O.OOO21
0.000089
0.000065
Total UF or
Benchmark
MOE
10
Notes:
- MOEs below benchmark MOE indicating risk are denoted in bold text. They indicate potential health risks.
Exposure estimates with/without LEV are found in Table 2-10.
Table 2-32. Acute Non-Cancer Risk Estimates for Commercial Use of Spotting Agent at Dry Cleaning Facilities
(Developmental Effects: Congenital Heart Malformations, Johnson et al., 2003)
Lowest PBPK-
derived
HECs (ppm)
of the
developmental
toxicity health
domain
HEC50= 0.012
HEC95= 0.0051
HEC99= 0.0037
WORKER NON-CANCER MOEs
With LEV—
Low-end
exposure
estimate
4.5
1.9
1.4
With LEV—
Upper-end
exposure
estimate
O.O18
0.0077
0.0056
No LEV—
Low-end
exposure
estimate
O.45
0.19
0.14
No LEV—
Upper-end
exposure
estimate
O.OO19
0.00081
0.00058
BYSTANDER NON-CANCER MOEs
With LEV—
Low-end
exposure
estimate
51.4
21.9
15.9
With LEV—
Upper-end
exposure
estimate
O.O18
0.0077
0.0056
No LEV—
Low-end
exposure
estimate
5.1
2.2
1.6
No LEV—
Upper-end
exposure
estimate
O.OO2
0.00085
0.00062
Total UF or
Benchmark
MOE
10
Notes:
- MOEs below benchmark MOE indicating risk are denoted in bold text. They indicate potential health risks.
Exposure estimates with/without LEV are found in Table 2-13.
Page 105 of 212
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Table 2-33. Acute Non-Cancer Risk Estimates for Residential Uses of TCE-containing degreasers and art/crafts products
(Developmental Effects: Congenital Heart Malformations, Johnson et al., 2003)
Lowest PBPK-
derived
HECs (ppm)
of the
developmental
toxicity health
domain
HEC50= 0.012
HEC95= 0.0051
HEC99= 0.0037
RESIDENTIAL USE OF DEGREASER PRODUCT
MOEs1
USER2
O.OO6O
0.0026
0.0019
BYSTANDER3
O.O15
0.0064
0.0046
RESIDENTIAL USE OF ARTS/CRAFTS CLEAR
PROTECTIVE COATING SPRAY
MOEs1
USER 2
O.O3
0.013
0.0093
BYSTANDER3
O.12
0.051
0.037
Total UF or
Benchmark
MOE
10
Notes:
1 MOEs below benchmark MOE indicating risk are denoted in bold text. They indicate potential health risks.
2 MOEs for the user categories could be extended to different age groups. EPA/OPPT believes that the users of these products are generally
adults, but teenagers and even children may be users or be in the same room with the user while engaging in arts and crafts projects or
degreasing.
3 All age categories (<1 yrs; 1-2 yrs; 3-5 yrs; 6-10 yrs; 11-15 yrs; 16-20 yrs ; and >21 yrs)
Page 106 of 212
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2.7.3 Chronic Non-Cancer and Cancer Risk Estimates for Inhalation
Exposures to TCE
Chronic non-cancer and cancer risk estimates for inhalation exposures to TCE were only derived
for occupational scenarios since the exposures for consumer uses were not considered chronic
in nature.
2.7.3.1 Cancer Risks for Occupational Scenarios
Figures 2-6 and 2-7 present the incremental individual lifetime cancer risks for continuous
exposures to TCE occurring during the commercial use of degreaser and spot cleaner products.
The cancer risk estimates were calculated by multiplying the EPA's inhalation unit risk for TCE
(EPA, 2011e) by the exposure estimate (i.e., LADC) for both direct users and bystanders. Cancer
risks were expressed as number of cancer cases per million. Calculations of cancer risks are
provided in the supplemental Excel spreadsheet, TCE OPPT Risk Estimates_061814.xlsx.
It was assumed that the exposure frequency (i.e., the amount of days per year workers or
bystanders are exposed to TCE) was 260 days per year and the occupational exposure duration
was 40 years over a 70-year lifespan. It is recognized that these exposure assumptions are likely
yielding conservative cancer risk estimates, but EPA/OPPT does not have additional information
for further refinement.
EPA typically uses a target cancer risk level between IxlO"4 and IxlO"6 for determining the
acceptability of the cancer risk in a population. Since the target cancer risk level will be
determined during risk management, the occupational cancer risk estimates were compared to
three target levels within EPA's acceptability range. The target levels were:
1. IxlO"6: the probability of 1 chance in 1 million of an individual developing cancer
2. IxlO"5: the probability of 1 chance in 100,000 of an individual developing cancer, which
is equivalent to 10 cancer cases in 1 million
IxlO"4: the probability of 1 chance in 10,00(
is equivalent to 100 cancer cases in 1 million
3. IxlO"4: the probability of 1 chance in 10,000 of an individual developing cancer, which
All of the degreaser exposure scenarios exceeded the three target cancer levels, with the
exception of one of the bystander exposure scenarios (i.e., "Bystander + LEV—Typical
Exposure"). This particular bystander exposure scenario exceeded the target levels at IxlO"5
and IxlO"6 (Figure 2-6).
Likewise, all of the worst case exposures for the spot cleaner scenarios (i.e., both user and
bystander scenarios) and one of the typical exposure scenarios with no LEV (i.e., "Spot Cleaner
No LEV—Typical Exposure") exceeded the three target levels. The remaining spot cleaner
scenarios exceeded the target level of IxlO"4 (i.e., "Spot cleaner + LEV—Typical Exposure";
"Bystander + LEV—Typical Exposure"; and "Bystander No LEV—Typical Exposure") (Figure 2-7).
Page 107 of 212
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Figure 2-6. Cancer Risk Estimates for Commercial Use of Degreaser Product at Small Shops
Number of Human Cancer Cases Per Million
IxlO-6 IxlO-5 IxlO'4
DEGREASER + LEV-Low-end exposure estimate
DEGREASER NO LEV-Low-end exposure estimate
DEGREASER + LEV-Upper-end exposure estimate
DEGREASER NO LEV-Upper-end exposure estimate
BYSTANDER + LEV-Low-end exposure estimate
BYSTANDER NO LEV-Low-end exposure estimate
BYSTANDER + LEV-Upper-end exposure estimate
BYSTANDER NO LEV-Upper-end exposure estimate
10
100 1000 10000 1000001000000
Figure 2-7. Cancer Risk Estimates for Commercial Use of Spotting Agent at Dry Cleaning
Facilities
Number of Human Cancer Cases Per Million
IxlO-6 IxlO-5 IxlO-4
SPOT CLEANER + LEV-Low-end exposure estimate
SPOT CLEANER NO LEV-Low-end exposure estimate
SPOT CLEANER + LEV- Upper-end exposure estimate
SPOT CLEANER NO LEV-Upper-end exposure estimate
BYSTANDER + LEV-Low-end exposure estimate
BYSTANDER NO LEV-Low-end exposure estimate
BYSTANDER + LEV-Upper-end exposure estimate
BYSTANDER NO LEV-Upper-end exposure estimate
0.1
10 100 1000 10000 1000001000000
Page 108 of 212
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For both degreaser and spot cleaner scenarios, the results indicated that the cancer risk was 10
times greater in scenarios with no local exhaust ventilation that those with room ventilation
(+LEV). Moreover, an 8-fold increase in cancer risk was observed in those workers directly
using the TCE-containing product when compared to their corresponding bystanders (Figures 2-
6 and 2-7). Similarly, a 10-fold increase in cancer risk was reported in those workers using
spotting agents containing TCE when compared to their corresponding bystanders indirectly
exposed in the dry cleaning facilities (Figure 2-7).
2.7.3.2 Chronic Non-Cancer Risks for Occupational Scenarios
EPA/OPPT estimated chronic non-cancer risks for the occupational use of TCE-containing
degreasers and spot cleaners. Since TCE exposure has been associated with a variety of health
effects, this assessment estimated human health inhalation chronic risks for developmental
toxicity, kidney toxicity, immunotoxicity, reproductive toxicity, neurotoxicity and liver toxicity.
As previously discussed, the TCE IRIS assessment developed PODs (i.e., HEC50, HEC95 or HEC99)
for multiple studies that were suitable for dose-response analysis (Appendix L). EPA/OPPT used
the lowest POD (or HEC) in each health effects domain for the non-cancer chronic MOE
calculations. The risk estimates for the commercial degreaser and spot cleaner uses are
presented in Tables 2-34 and 2-35, respectively30.
Workers in direct contact with the degreaser products and occupational bystanders reported
risks (i.e., MOEs
-------
cancer chronic risks for the aforementioned health effects domains under LEV and typical
exposure conditions (Table 2-35).
Risks for reproductive effects and neurotoxicity were less consistently seen across the board
with the exception of the spot cleaning worker/bystander exposure scenarios characterizing
typical or worst-case exposure condition under no ventilation system. As for risks for liver
effects, most of the spot cleaning scenarios showed no risk concern with the exception of the
worker/bystander worst case exposure scenarios with no LEV (Table 2-35).
2.7.4 Human Health Risk Characterization Summary
This risk assessment focused on the occupational and consumer uses of TCE-containing
degreasers, spot cleaners and clear protective coating spray in arts/crafts. Specifically, the
following exposure scenarios were evaluated: small commercial degreasing operations; spot
cleaning in dry cleaning facilities; consumer use of an aerosol degreaser; and the consumer use
of a clear protective coating spray in an arts/crafts home setting. The population of interest
consisted of workers and consumers with direct (users) or indirect (bystander) exposure to TCE.
Only the inhalation route of exposure was considered in this risk assessment.
The occupational and consumer exposure assessments generated the TCE exposure levels
required to derive non-cancer and cancer risks. Cancer risks were presented as lifetime risks,
meaning the risk of developing cancer as a result of the occupational exposure over a normal
lifetime of 70 yrs. Lifetime cancer risks from TCE exposure were compared to target risks
ranging from 10"6to 10"4.
Many of the degreaser and spot cleaning exposure scenarios exceeded the target cancer risks
of 10~6,10~5 and 10~4. This analysis resulted in higher modeled incidences of cancer in the small
commercial degreaser than the users of spot cleaners. Thus, the greatest potential for cancer
risk came from the occupational exposures to commercial degreasers. Furthermore, higher
cancer risks resulted from direct use of the degreaser or lack of local exhaust ventilation at the
workplace.
To characterize the risks of adverse health effects other than cancer, MOEs were used to
evaluate non-cancer risks for both acute and chronic exposures using the hazard values
published in the TCE IRIS assessment (EPA, 2011e). Hazard values based on developmental
toxicity (i.e., fetal cardiac defects; Johnson et al., 2003) were used to estimate acute non-cancer
risks for occupational and consumer exposures. On the other hand, the chronic non-cancer risks
for the worker scenarios were evaluated with the hazard values associated with health effects
following long-term exposure to TCE (i.e., developmental toxicity, kidney toxicity,
immunotoxicity, reproductive toxicity, neurotoxicity and liver effects). Note that minimal
variability (i.e., < 3-fold) exist among the acute and chronic non-cancer hazard values (i.e.,
HEC50, HEC95 or HEC99) used in this assessment.
Page 110 of 212
-------
Most occupational and residential exposure scenarios reported acute risks based on concerns
for developmental effects (i.e., cardiac defects) that may occur following a single exposure to
TCE during a critical window of susceptibility. Particularly, the degreaser exposure scenarios
showed greater acute risks than those reported for the spot cleaning exposure scenarios.
There is a concern for a range of human health effects other than cancer that may appear after
chronic exposures to TCE during the occupational use of TCE-containing degreasers and spot
cleaning agents. The greatest concern is for developmental effects (i.e., fetal cardiac defects),
followed by kidney effects and then immunotoxicity, with an overall higher chronic risk for the
degreaser exposure scenarios. In general, the concerns for these three health effects domains
occur regardless of the type of exposure (typical vs worst case) and the availability of room
ventilation (LEV vs no LEV), although there are some exceptions, particularly in the spot
cleaning bystander exposure scenarios.
Potential chronic risks for reproductive effects and neurotoxicity were also observed for
degreaser worker exposure scenarios and most of the degreaser bystander exposure scenarios.
However, the risks concerns for these effects were reported for fewer spot cleaning
worker/bystander scenarios and generally attributed to exposure conditions without room
ventilation.
Concerns for liver effects following chronic exposure to TCE are less prominent than the
concerns for other health effects domains as chronic risks for liver toxicity were not reported
for the majority of the degreaser and spot cleaning worker/bystander scenarios. The exception
was the degreaser worker/bystander exposure worst case scenarios and the spot cleaning
worker/bystander worst case scenarios with no LEV.
Page 111 of 212
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Table 2-34. Chronic Non-Cancer Risk Estimates for Commercial Use of Degreaser Product at Small Shops
Health Effect
Domain and Study
DEVELOPMENTAL
TOXICITY
(Johnson et al.,
2003)
KIDNEY
(NTP, 1998)
IMMUNOTOXICITY
Keiletal., 2009
(Decrease in thymus
weight and thymus
cellularity)
IMMUNOTOXICITY
Keiletal., 2009
(Autoimmunity)
REPRODUCTIVE
TOXICITY
(Chiaetal. 1996)
NEUROTOXICITY
(Arito et al., 1994)
LIVER
(Kjellstrandetal.
1983)
Lowest HEC
(ppm) of each
health effects
domain
HEC50= 0.012
HEC95= 0.0051
HEC99= 0.0037
HEC50= 0.042
HEC95= 0.0085
HEC99= 0.0056
HEC50= 0.092
HEC95= 0.045
HEC99= 0.033
HEC50= 0.092
HEC95= 0.045
HEC99= 0.033
HEC50= 1.4
HEC95= 0.7
HEC99= 0.5
HEC50= 13
HEC95= 6.4
HEC99= 4.8
HEC50= 25
HEC95= 12
HEC99= 9.1
WORKER NON-CANCER MOEs 1
With LEV-
Low-end
exposure
estimate
0.17
O.O72
0.052
O.59
O.11
0.079
1.3
O.6
0.5
1.3
O.6
O.5
19.7
9.8
7.O
183
90
67
351
168
128
No LEV-
Low-end
exposure
estimate
0.017
O.OO72
0.0052
O.O59
O.O12
0.0079
0.1
O.O6
0.05
O.13
O.O63
O.O46
2.O
1.0
O.7
18
9
7
35
17
13
With LEV-
Upper-end
exposure
estimate
0.0025
O.OO11
0.0008
O.OO88
O.OO18
0.0012
0.019
O.OO95
0.0069
O.O19
O.OO95
O.OO69
O.3
0.15
O.11
2.7
1.3
1.O
5.3
2.5
1.9
No LEV-
Upper-end
exposure
estimate
0.0003
O.OOO1
0.0001
O.OOO9
O.OOO2
0.0001
0.0020
O.OO1O
0.0007
O.OO2O
O.OO1O
O.OOO7
O.O3
0.015
O.O11
0.28
0.14
O.1O
0.53
O.26
O.19
BYSTANDER NON-CANCER MOEs 1
With LEV--
Low-end
exposure
estimate
1.3
O.54
0.40
4.4
O.9
0.6
9.7
4.7
3.5
9.7
4.7
3.5
147
74
53
1369
674
505
2632
1263
958
No LEV--
Low-end
exposure
estimate
0.13
O.O5
0.04
O.4
O.O9
0.06
0.97
O.47
0.35
O.97
O.47
O.35
15
7
5
137
67
51
263
126
96
With LEV--
Upper-end
exposure
estimate
0.0030
O.OO13
0.0009
O.O1
O.OO21
0.0014
0.023
O.O11
0.0082
O.O22
O.O11
O.OO82
O.35
0.17
O.12
3.2
1.6
1.2
6
3
2
No LEV--
Upper-end
exposure
estimate
0.0003
O.OOO1
0.0001
O.OO1
O.OOO2
0.0001
0.0023
O.OO11
0.0008
O.OO23
O.OO11
O.OOO8
O.O34
0.017
O.O12
0.32
0.16
O.12
0.61
O.29
O.22
Total UF or
Benchmark
MOE
10
10
100
30
30
300
10
Notes: (1) MOEs below benchmark MOE indicating risk are denoted in bold text. They indicate potential health risks. (2) Exposure estimates with/without LEV
are found in Table 2-10.
Page 112 of 212
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Table 2-35. Chronic Non-Cancer Risk Estimates for Commercial Use of Spotting Agent at Dry Cleaning Facilities
Health Effect
Domain and Study
DEVELOPMENTAL
TOXICITY
(Johnson et al.,
2003)
KIDNEY
(NTP, 1998)
IMMUNOTOXICITY
Keiletal., 2009
(Decrease in thymus
weight and thymus
cellularity)
IMMUNOTOXICITY
Keiletal., 2009
(Autoimmunity)
REPRODUCTIVE
TOXICITY
(Chiaetal. 1996)
NEUROTOXICITY
(Arito et al., 1994)
LIVER
(Kjellstrandetal.
1983)
Lowest HEC
(ppm) of each
health effects
domain
HEC50= 0.012
HEC95= 0.0051
HEC99= 0.0037
HEC50= 0.042
HEC95= 0.0085
HEC99= 0.0056
HEC50= 0.092
HEC95= 0.045
HEC99= 0.033
HEC50= 0.092
HEC95= 0.045
HEC99= 0.033
HEC50= 1.4
HEC95= 0.7
HEC99= 0.5
HEC50= 13
HEC95= 6.4
HEC99= 4.8
HEC50= 25
HEC95= 12
HEC99= 9.1
WORKER NON-CANCER MOEs 1
With LEV-
Low-end
exposure
estimate
6.3
2.7
1.9
22
4.5
2.9
48
24
17
48
24
17
737
369
263
6844
3369
2527
13161
6317
4791
No LEV-
Low-end
exposure
estimate
O.O253
0.0107
O.OO78
O.O88
0.018
0.012
O.194
0.095
O.O69
O.194
O.O95
O.O69
2.9
1.5
1.1
27
13
10
53
25
19
With LEV-
Upper-end
exposure
estimate
O.63
0.27
O.19
2.2
0.45
0.29
4.8
2.4
1.7
4.8
2.4
1.7
74
37
26
684
337
253
1316
632
479
No LEV-
Upper-end
exposure
estimate
O.OO27
0.0011
O.OOO8
O.OO93
0.0019
0.0012
O.O2O
0.010
O.OO7
O.O2O
O.O1O
O.OO7
0.31
O.16
0.11
2.9
1.4
1.1
5.5
2.7
2.0
BYSTANDER NON-CANCER MOEs 1
With LEV--
Low-end
exposure
estimate
72
31
22
253
51
34
554
271
199
554
271
199
8423
4212
3008
78214
38505
28879
150412
72198
54750
No LEV--
Low-end
exposure
estimate
O.O253
0.0107
O.OO78
O.O884
0.0179
0.0118
O.194
0.095
O.O69
O.194
O.O95
O.O69
2.9
1.5
1.1
27
13
10
53
25
19
With LEV--
Upper-end
exposure
estimate
7.2
3.1
2.2
25
5.1
3.4
55
27
2O
55
27
2O
842
421
301
7821
3851
2888
15041
7220
5475
No LEV--
Upper-end
exposure
estimate
O.OO28
0.0012
O.OOO9
O.OO98
0.0020
0.0013
O.O22
0.011
O.OO8
O.O22
O.O11
O.OO8
0.33
O.16
0.12
3.0
1.5
1.1
5.8
2.8
2.1
Total UF or
Benchmark
MOE
10
10
100
30
30
300
10
Notes: (1) MOEs below benchmark MOE indicating risk are denoted in bold text. They indicate potential health risks. (2) Exposure estimates with/without LEV are
found in Table 2-13.
Page 113 of 212
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2.8 DISCUSSION OF KEY SOURCES OF UNCERTAINTY AND DATA
LIMITATIONS
The characterization of variability and uncertainty is fundamental to any risk assessment.
Variability refers to "the true heterogeneity or diversity in characteristics among members of a
population (i.e., inter-individual variability) or for one individual over time (infra-individual
variability)" (EPA, 2001c). The risk assessment was designed to reflect critical sources of
variability to the extent allowed by available methods and data and given the resources and
time available.
On the other hand, uncertainty is "the lack of knowledge about specific variables, parameters,
models, or other factors" (EPA, 2001c) and can be described qualitatively or quantitatively.
Uncertainties in the risk assessment can raise or lower the confidence of the risk estimates. In
this assessment, the uncertainty analysis also included a discussion of data gaps/limitations.
The next section describes the uncertainties and data gaps in the exposure, hazard/dose-
response and risk characterization.
2.8.1 Uncertainties in the Occupational and Consumer Exposure
Assessments
The production volume and release information on TCE are estimates and the actual TCE
production or import data may differ from these estimates. The 2011 production volume and
use data used in this assessment were published recently and are considered reliable. The vast
majority of the TCE used in the U.S. is as an intermediate for the production of a refrigerant
(83.6%) and the second highest use (in terms of production volume) is as a degreaser (~14.7%).
Confidence in the remaining uses (i.e., <2 percent of the production volume) is less certain.
EPA/OPPT expects dermal exposures under some conditions during the occupational and
consumer user of TCE in degreaser and arts/crafts products. However, dermal exposures were
not evaluated in this assessment.
2.8.1.1 Small Commercial Degreasing Operations
Releases of and exposures to TCE can vary from one degreasing facility to the next. EPA/OPPT
attempted to quantify this uncertainty by evaluating multiple scenarios to establish a range of
releases and exposures.
1. Releases: EPA/OPPT used data from NEI and TRI to estimate releases of TCE into the
workplace. EPA/OPPT's estimate was found to be similar in magnitude to release estimates
reported in publications and to estimates based on regulatory emission limits.
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— This indicates that NEI and TRI are adequate data sources for developing order of
magnitude estimates.
— However, rather than just use one release estimate, EPA/OPPT used a range of release
estimates (based on the additional sources identified), thus incorporating this
uncertainty into its assessment.
2. Exposures: EPA/OPPT compared its estimated range of exposures to field measurements
obtained from OSHA. EPA/OPPT's estimates and OSHA values were of the same order of
magnitude (Figure 2-2); EPA/OPPT's exposure range captured approximately 95% of OSHA
field measurements with the higher end values of the monitoring data not covered by EPA
estimates.
— EPA/OPPT used a NF/FF mass balance model to estimate workplace exposures. These
exposure estimates depend on model inputs. Rather than just use a single value for
model inputs, we used ranges for the model inputs.
• For example, based on data in the literature, ranges were used for parameters
such as the room volume, air exchange rate, TCE emissions into the workplace
and effectiveness of engineering controls.
• Based on comments received from external peer review, certain parameter
inputs were deemed adequate and thus were not varied; parameters such as the
indoor air velocity, the size of the near field region, and hours of operation.
3. Population Exposed: EPA/OPPT estimated the number of workers and occupational
bystanders potentially exposed to TCE based on a NIOSH survey from the 1980s and on the
number of small degreasing facilities.
— Regarding the NIOSH survey, EPA/OPPT normalized these data by estimating the
number of workers potentially exposed to TCE on a per facility basis.
• Use of TCE in degreasing has been on the decline, leading to a decline in the
number of facilities.
• By using a more recent estimate for number of facilities, EPA/OPPT's estimate
likely captures this downward trend and provides an adequate order of
magnitude estimate. However, this is our judgment and more recent and
relevant survey data were not identified for the purposes of comparison.
— EPA/OPPT estimated the number of degreasing facilities based on EPA's 2006 risk
assessment for the halogenated solvent cleaning source category. Since use of TCE in
degreasing has been on the decline, this assumption may overestimate the number of
facilities and thus the size of the population exposed.
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2.8.1.2 Spot Cleaning at Dry Cleaning Facilities
Releases of and exposures to TCE can vary from one dry cleaning facility to the next. EPA/OPPT
attempted to quantify this uncertainty by evaluating multiple scenarios to establish a range of
releases and exposures.
For this assessment, EPA relied on a 2007 study specific to spot cleaning at dry cleaning facilities
in the state of California (CalEPA/EPA, 2007). Dry cleaning facilities in California are assumed to
be representative of dry cleaning facilities in the United States. However, this may not be the
case; how this assumption impacts EPA's release and exposure assessment is unclear.
1. Releases: To estimate releases, EPA/OPPT assumed that the entire amount of TCE used for
spot cleaning was available for evaporation and thus could be emitted into the workplace.
The basis for this assumption is that after spot cleaning, garments are usually queued in a
basket prior to the next operation. However, this assumption can overestimate releases of
TCE into the workplace.
2. Exposures: EPA/OPPT compared its exposure estimates to field measurements performed
by NIOSH; these measurements were specific to spot cleaning with TCE. NIOSH
measurements were within EPA/OPPT's estimated exposure range.
— EPA/OPPT used a NF/FF mass balance model to estimate workplace exposures. These
exposure estimates depend on model inputs. We used ranges for the model inputs
rather than using a single value.
• For example, based on data in the literature, ranges were used for parameters
such as the room volume, air exchange rate, TCE emissions into the workplace
and effectiveness of engineering controls.
• Based on comments received from an external peer review, certain parameter
inputs were deemed adequate and thus were not varied; parameters such as the
indoor air velocity and the size of the near field region.
3. Population Exposed: EPA/OPPT estimated the number of workers and occupational
bystanders potentially exposed to TCE based on U.S. Census data. Data on number of
workers were not adjusted to exclude job categories that likely would not be present at dry
cleaning facilities. Thus, EPA/OPPT's estimate likely overestimates the size of the population
exposed.
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2.8.1.3 Degreaser and Arts/Crafts Uses in Residential Settings
Uncertainties in the consumer exposure assessment arise from the following sources:
1. Consumer use information: Although EPA/OPPT found information about TCE products
intended for consumer use, there is some general uncertainty regarding the nature and
extent of the consumer use of TCE for the products under the scope of this assessment.
2. Model assumptions and input parameters: The use patterns assumed for the two
consumer products, including mass of product used per event, duration of event, and
events per year, were hypothetical and not based on consumer product survey data since
this was lacking. Therefore, they are likely the source of the greatest uncertainties/data
gaps in the exposure estimates for the two hobbyist products. However, there is a high
degree of confidence in the consumer product weight fractions identified for the two
consumer products evaluated in this assessment. Also, there is a medium to high degree of
confidence in certain modeling inputs to the CEM model, including vapor pressure,
molecular weight, room volumes, whole house volume, air exchange rate, body weight, and
inhalation rate.
There is no chamber data available for the products modeled in the exposure assessment,
thus CEM calculated the mass of TCE entering the room of use by relying on data from a
paper that studied the emission rates of solvents off a surface (Chinn, 1981). The spray
degreaser results in only TCE being on the surface so it fits well into the Chinn data set,
however the spray fixative product does have other components that may affect the
evaporation rate of TCE. This introduces uncertainty and a further discussion of this issue is
in Appendix I.
3. Conversion of acute dose rates to air concentrations: Because the E-FAST2/CEM model
outputs for exposure to the user and bystander scenarios are reported in mg/kg-bw/day, it
was necessary to convert these values to air concentrations (ppm) in order to perform the
non-cancer and cancer risk assessment. This conversion introduces some uncertainty, but it
is not apparent whether it may over- or under-estimate exposures.
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A*^^
2.8.2.1 Uncertainties in the Cancer Hazard/Dose-Response Assessments
The cancer IUR for TCE was based on human kidney cancer risks reported by Charbotel et al.
(2006) and adjusted for potential risk for NHL and liver cancer based on human epidemiological
data (EPA, 2011e). U.S. EPA has high confidence in the overall derivation of the cancer IUR
because it was based on good quality human data and it was similar to unit risk estimates
derived from multiple rodent bioassays (EPA, 2011e). Moreover, the assumption of linearity in
the relationship between TCE exposure and probability of cancer was based on sufficient
weight of evidence supporting a mutagenic mode of action for at least TCE-induced kidney
tumors (EPA, 2011e). However, there is insufficient information about the operational modes of
actions for the other TCE-induced cancers (e.g., NHL, liver) supporting the default linear
approach to estimate cancer risks in the low-dose region (EPA, 2011e).
Although uncertainties arising from animal to human extrapolation were not present in the IUR
derivations due to use of human data, other sources of uncertainty exist in the cancer dose-
response models and human data used to derive the IUR. These uncertainties are briefly listed
and summarized below from information discussed in the TCE IRIS assessment (EPA, 2011e).
1. A source of uncertainty is the cancer dose-response model used to estimate the POD for the
IUR derivations. A weighted linear regression model was used to fit the epidemiological
data reported by Charbotel et al. (2006). Although a linear model is a good general
approach for human studies with limited data, it cannot be ruled out that other alternate
model would have performed better than the linear model (EPA, 2011e).
2. There is some evidence that exposure misclassification occurred for some of the cases
reported by Charbotel et al. (2006) as a result of retrospectively estimating the cumulative
TCE exposures of those that participated in the case-control study. The inhalation unit risk
could be under- or overestimated depending on the directional bias of the exposure
estimates (EPA. 2011e).
3. Charbotel et al. (2006) accounted for many potential confounding or modifying factors such
as exposure to other solvents, lead, ionizing radiation, cutting fluids and other petroleum
oils, medical history (e.g., kidney stones, infection, chronic dialysis) and lifestyle information
(e.g., smoking, coffee intake). These confounding factors were expected to minimally
impact the IUR, but it is possible that other missing factors could influence the cancer slope
estimate (EPA. 2011e).
4. There are possible uncertainties associated with the inclusion of a lag period in the analysis
of the cancer data. This lag period intended to discount those recent TCE exposures not
likely to contribute to the reported cancer incidence in Charbotel et al. (2006). It seems that
the lag period might not be an important factor in Charbotel et al. (2006) based on
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published (Charbotel et al., 2006) and unpublished analyses (Charbotel et al., 2005) showing
similar results in the absence or presence of a lag period (EPA, 2011e).
5. The cancer IUR derived from the Charbotel et al. (2006) renal cancer data was further
adjusted to account for other types of cancer (i.e., liver cancer and NHL). There are
uncertainties related to the data analysis and assumptions used for the adjustment
calculations (EPA, 2011e). For instance, comparisons of the relative contributions to extra
cancer risk for two different data sets showed that the results were within 25% of the
selected adjustment factor of 4, which shows that the selected factor is reasonable. Also,
there are uncertainties related to the association between TCE exposure and increased risks
of cancer at multiple sites. The human evidence of carcinogenicity from epidemiologic
studies of TCE exposure is strong for NHL, but less convincing than for kidney cancer, and
more limited for liver cancer (EPA, 2011e). Further support for TCE's carcinogenic
characterization comes from positive results in multiple rodent cancer bioassays (EPA,
2011e). Overall, there is sufficient evidence to adjust the IUR for three cancer types.
Alternatively, if the IUR would have been derived for two cancer types (i.e., kidney and
NHL), the cancer IUR estimate would be reduced by 25%.
2.8.2.2 Uncertainties in the Non-Cancer Hazard/Dose-Response Assessments
EPA/OPPT's risk assessment relied on the PBPK-derived hazard values (i.e., HECs) published in
the latest EPA IRIS assessment on TCE (EPA, 2011e). These hazard values were used to estimate
acute and chronic risks to various health effects following TCE exposure related to specific TCE
uses.
The TCE IRIS assessment conducted a comprehensive discussion of the uncertainties inherent to
the data, assumptions and models used to support the derivation of the chronic non-cancer
PODs for different health effects domains. Below is a summary of the major uncertainties
affecting the non-cancer hazard/dose response approach of this assessment. The reader is
referred to the TCE IRIS assessment to obtain details about the non-cancer uncertainty analysis
(EPA. 2011e).
Uncertainties in the acute and chronic hazard values stem from the following sources:
1. Non-cancer hazard values (e.g., NOAELs, LOAELs, BMD): The TCE IRIS assessment
identified PODs from human and animal studies that were suitable for dose-response
analysis. The process of identifying PODs for various health effects domains involved the
evaluation of the strengths and limitations of the data and the weight of evidence for a
particular health effects domain before supporting an association between TCE exposure
and various human health effects. The TCE IRIS assessment issued confidence statements
for the different health effects domains/studies as part of the uncertainty analysis.
However, there are uncertainties about the selected PODs since the values (e.g., NOAEL,
LOAEL or BMD) depend on the current available data and could change as additional studies
are published (EPA. 2011e).
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Also, when selecting a BMD as a POD, the selection of the benchmark dose response (BMR)
(e.g., 1%, 5% or 10% level) directly affects the calculation of the BMD. There are
uncertainties related to the BMRs since their selection depends on scientific judgments on
the statistical and biological characteristics of the dataset and how the BMDs will be finally
used (EPA. 2012a).
In addition, there are uncertainties about the appropriate dose-response model used to
generate the BMDs. However, these uncertainties should be minimal if the chosen model
fits well the observable range of the data, as discussed in the BMDS guidance (EPA, 2012a).
2. Duration adjustment to continuous exposure: Most of the PODs used to derive PBPK-
derived HECs came from studies that did not expose animals or humans to TCE on a
continuous basis. These PODs were then mathematically adjusted to reflect equivalent
continuous exposures (daily doses) over the study exposure period under the assumption
that the effects are related to concentration x time (C x t), independent of the daily (or
weekly) exposure regimen (EPA, 2011e). However, the validity of this assumption is
generally unknown, and, if there are dose-rate effects, the assumption of C x t equivalence
would tend to bias the POD downwards (EPA, 2011e).
3. Extrapolation of repeated dose developmental effects to acute scenarios: There are
uncertainties related to whether developmental effects observed in developmental toxicity
studies may result from a single exposure to TCE. In this assessment, the acute risk
assessment used the hazard value for fetal cardiac defects derived from the Johnson et al.
developmental toxicity studies.
Previously identified uncertainties in the Johnson et al. studies have focused on
methodological and reproducibility issues. The author recently published an errata
(Johnson, 2014) to update the public record regarding Johnson et al. (2003). However, some
questions on that study remain unresolved, i.e., the precise dates that each individual
control animal was on study and the detailed results of analytical chemistry testing for dose
concentration.
Additional possible sources of uncertainty identified for the Johnson et al. studies include
that the research was conducted over a 6-year period, combined control data were used for
comparison to treated groups, and possible imprecision of exposure characterization due to
the use of tap water in the Dawson et al. (1993) study and TCE intake values that were
derived from water consumption measures of group housed animals.
On the other hand, the strengths of the Johnson et al. studies include the examination of
fetal hearts without knowledge of treatment (or control) group, standardized methods of
fetal evaluation, examination of the gross (in situ) and internal structure of the fetal hearts
by a group of 3 senior researchers, confirmation of cardiac anomalies by consensus
agreement, and that the researchers shared individual fetal and litter cardiac abnormality
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data for treated groups with EPA, thereby facilitating independent statistical analysis of the
data.
The potential hazard for congenital malformations is supported by a weight of evidence
analysis of the weakly suggestive epidemiological data in combination with the findings of
the animal and mechanistic studies with TCE and its metabolites (EPA, 2011e). The
robustness of the weight of evidence analysis gives greater confidence to the hazard
conclusions for fetal cardiac defects (Appendix N).
Furthermore, it is unknown if a higher exposure level of TCE may be required to induce fetal
cardiac malformations following a single exposure in light of the short critical window of
vulnerability of cardiac development and the absence of data for a very short window of
exposure. A single exposure to TCE at a critical window of fetal development may produce
adverse developmental effects (EPA, 1991). This was assumed to be a health protective
approach.
4. PBPK model-structure, parameters and model fits: There are uncertainties associated
with the various steps of the model development process of the TCE PBPK model. Most of
the assumptions underlying the PBPK model structure are well established for volatile,
lipophilic chemicals such as TCE. Thus, these assumptions are unlikely to introduce much
bias or inaccuracy in the modeling results. In addition, the model provided reasonable fits to
an extraordinarily large database of in vivo pharmacokinetic data in rodents and humans
(EPA. 2011e).
Moreover, posterior parameter distributions were generated by Markov Chain Monte Carlo
(MCMC) sampling, which employed a hierarchical Bayesian population statistical model and
the available in vivo data (EPA, 2011e). As stated in EPA (2011e), "[Convergence of the
MCMC samples for model parameters was good for mice, and adequate for rats and
humans. Evaluation of posterior parameter distributions suggested] reasonable results in
light of prior expectations and the nature of the available calibration data". Interestingly,
the model predictions in rats and humans were consistent with in vivo data from many
studies that were not used for the calibration process. Furthermore, local sensitivity
analyses were conducted and confirmed that most of the scaling parameters were informed
by at least some of the calibration data, and those that were not, either were informed by
prior data or would not have great impact on dose-metric predictions (EPA, 2011e).
5. PBPK model—dose metrics: Dose-response analysis using PBPK modeling prefers dose-
metrics that are closely associated with one or more key events that lead to the selected
critical effect (i.e., toxic endpoint of concern). Additional uncertainties exist about the
appropriate dose-metric for a particular toxic endpoint, although for some effects, there
was better information about relevant dose-metrics than for others (EPA, 2011e).
Moreover, the dose-metric predictions were evaluated for the degree to which the
simulations have converged to the true posterior distribution, the combined uncertainty
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and population variability, the degree of uncertainty in particular human population
percentiles and the degree to which the model predictions are consistent with in vivo data
(EPA. 2011e).
The analysis showed that the TCE PBPK model appears to be most reliable for the fluxes of
total, oxidative, and hepatic oxidative metabolism. In addition, modest uncertainty was
found for dose-metrics related to blood levels of TCE and oxidative metabolites, TCOH and
TCA. For GSH metabolism, the GSH conjugation predictions had a lower confidence than
those dose-metrics based on the parent compound, total metabolism or oxidative
metabolites (EPA, 2011e). Predictions for other oxidative metabolism and respiratory
oxidative metabolism generally had somewhat more uncertainty than the TCE and
metabolism metrics (EPA, 2011e).
6. PBPK model—population variability: The TCE PBPK model used a Bayesian population
analysis to systematically estimate model parameters and characterize their uncertainty
and variability. Although labor intensive, this approach characterized uncertainty and
variability in a highly transparent and objective manner (EPA, 2011e). The reader is referred
to the TCE IRIS assessment for detailed information about the uncertainties of the PBPK
model, specifically Chapters 3 and 6 as well as Appendix A (EPA, 2011e). Below are the
highlights of the discussion of the uncertainties of the PBPK modeling approach.
The predictions of population variability were based on prior and population distributions.
These selected distributions may introduce inaccuracies in the predictions of population
variability (EPA, 2011e). However, the impact of the chosen distributions was limited to the
human variability related to GSH conjugation, which resulted in changes in the dose-metric
predictions (EPA, 2011e). There are also uncertainties regarding how the PBPK modeling
results address the pharmacodynamic variability of the susceptible human subpopulations
exposed to TCE (EPA. 2011e).
Furthermore, the hierarchical model did not consider certain sources of variability, such as
between-animal variability in rodents and between-occasion variability in humans. Instead,
they were aggregated with other sources of variability in a residual error term (EPA, 2011e).
It seems that this approach did not introduce significant bias in the modeling estimates
because the residuals between predictions and data do not overall appear systematically
high or low (EPA. 2011e).
7. PBPK model—developmental toxicity: The TCE PBPK model does not have a
fetus/gestational compartment. The lack of this compartment introduces uncertainty in the
modeling estimates (i.e., HECs) for developmental effects. Inclusion of a fetus/gestational
compartment would require additional in vivo or in vitro metabolism data to ensure model
identifiability (EPA. 2011e).
8. PBPK model—route-to-route extrapolation: PBPK-derived hazard values were based on
PODs from either inhalation or oral studies. The TCE PBPK model used interspecies and
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route-to-route extrapolation approaches to convert both the inhalation and oral PODs to
human internal doses. Then, the model estimated the human equivalent concentrations
needed to produce the human internal doses. Since the PBPK model was used, the
uncertainties associated with using oral studies for inhalation exposures were minimized.
^2.8.3^JJ^
The non-cancer acute or chronic risks were expressed in terms of MOEs. MOEs are obtained by
comparing the hazard values (i.e., HEC50, HEC95 or HEC99) for various TCE-related health effects
with the exposure concentrations for the specific use scenarios. Given that the MOE is the ratio
of the hazard value divided by the exposure, the confidence in the MOEs is directly dependent
on the uncertainties in the hazard/dose-response and exposure assessments that supported
the hazard and exposure estimates used in the MOE calculations.
The total UF for each non-cancer PBPK-derived POD was the benchmark MOE used to interpret
the MOE risk estimates for each use scenario. The UFs accounted for various endpoint/study-
specific uncertainties in the hazard values, such as:
1. Animal-to-human extrapolation (UFA): The UFA accounts for the uncertainties in
extrapolating from rodents to humans. In the absence of data, the default UFAof 10 is
adopted which breaks down to a factor of 3 for toxicokinetic variability and a factor of 3 for
pharmacodynamic variability. The TCE PBPK model accounted for the interspecies
extrapolation using rodent pharmacokinetic data to estimate internal doses for a particular
dose metric, thus reducing the interspecies toxicokinetic uncertainty to 1. Since the PBPK
model did not address interspecies toxicodynamic differences, the total UFA of 3 was
retained for all of the PBPK-derived HECs, unless a human study was the source of the POD.
In that case, the UFAwas set to 1 (EPA. 2011e).
2. Inter-individual variation (UFH): The UFH accounts for the variation in sensitivity within the
human population. In the absence of data, the default UFH of 10 is adopted which breaks
down to a factor of 3 for toxicokinetic variability and a factor of 3 for pharmacodynamic
variability. The TCE PBPK model reduced the human toxicokinetic variability to 1, but not
the human toxicodynamic variability. Thus, the total UFH was 3 for all of the PBPK-derived
HECs. This is because the PBPK model does not address the uncertainties regarding the
susceptibility of the human subpopulations to TCE exposure and the extent of
pharmacodynamics variability (EPA, 2011e).
3. Database uncertainty factor (UFD): The UFD accounts for deficiencies in the toxicity
database that may result in a lower hazard value. The database for TCE toxicity is extensive
with studies for many different types of effects, including two-generation reproductive
studies, as well as neurological and immunological studies (EPA, 2011e). Thus, a UFD of 1
was retained for all of the PBPK-derived HECs discussed in the OPPT's risk assessment.
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4. Extrapolation from subchronic to chronic (UFS): The UFS accounts for the uncertainty in
extrapolating from a subchronic to a chronic POD. UFS ranging from 3 to 10 were used in
some of the PBPK-derived HECs. Typically, a UFS of 1 was used to extrapolate a POD from a
less-than-chronic developmental toxicity study to a chronic exposure. Developmental PBPK-
derived HECs did not use a higher UFS because the developmental period is recognized as a
susceptible life stage where exposure during certain time windows is more relevant to the
induction of developmental effects than lifetime exposure (EPA, 1991).
5. LOAEL-to-NOAEL extrapolation (UFL): The UFL accounts for the uncertainty in extrapolating
from a LOAEL to a NOAEL. A value of 10 is the standard default UFL value, although lower
values (e.g., 3) can be used if the effect is considered minimally adverse at the LOAEL or is
an early marker for an adverse effect (EPA, 2011e). UFL ranging from 3 to 30 (i.e., 3, 10 or
30) were used in the PBPK-derived HECs. For one of the kidney PODs (NCI, 1976), a UFL
value of 30 was used because the incidence rate for the adverse effect was >90% at the
LOAEL (EPA. 2011e).
Unlike cancer risks, an MOE exceeding the benchmark MOEs is an indicator that there is a
potential risk and cannot be translated to a probability that certain adverse health effects
would occur. Also, those MOEs that exceed but remain close to the benchmark MOE do not
necessarily mean that adverse effects would occur.
The chronic non-cancer risks for the occupational scenarios assumed that the human health
risks are constant for a working lifetime based on the exposure assumptions used in the
occupational exposure assessment. However, the risks could be under- or over-estimated
depending on the variations to the exposure profile of the workers and occupational bystanders
using TCE-containing degreasers and spot cleaners.
Regarding exposure to TCE through the skin, the impact of dermal exposures on human health
risks was not assessed in this assessment for the consumer and occupational scenarios.
Exclusion of dermal exposures is expected to underestimate the risks of the selected TCE uses.
This would likely be an issue of concern in those exposure scenarios that resulted in a "no-risk"
finding, especially those that reported MOEs close to the benchmark MOE, but still above the
benchmark.
As discussed previously, the cancer risk estimates were based on the assumption of linearity in
the relationship between TCE exposure and probability of cancer. Uncertainties are introduced
in the cancer risks when there is limited information justifying the liner cancer dose-response
model when compared to other available models. In the case of TCE, the cancer IUR was based
on reliable data supporting a mutagenic mode of action for at least TCE-induced kidney tumors
(EPA, 2011e). The IUR was adjusted to account for other types of TCE-induced cancers (e.g.,
NHL, liver). There is some uncertainty about the validity of the linear approach for these other
cancers since there was insufficient information about the modes of actions underlying the
onset of these cancers following chronic exposure to TCE (EPA, 2011e).
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EPA/OPPT's risk assessment focused on uses of TCE as a degreaser both in small commercial
settings and consumers, the commercial use of TCE as a spotting agent in dry cleaning shops,
and the consumer use of TCE in a clear protective coating spray by individuals in the arts/crafts
field. In this assessment, EPA/OPPT estimated the size of the population at risk as:
• Approximately 30,000 workers and occupational bystanders at small commercial degreasing
operations.
• Approximately 300,000 workers and occupational bystanders at dry cleaning operations.
• No data were available to estimate the number of consumers and bystanders exposed to
TCE during the use of degreasers and arts/crafts clear protective coating spray.
In summary, the risk assessment showed the following risk findings:
Cancer Risks
1. There are cancer risk concerns for users and bystanders occupationally exposed to TCE
when using TCE-containing degreasers and spot cleaners in small commercial shops and dry
cleaning facilities, respectively.
2. Many of the degreaser and spot cleaning exposure scenarios exceed the target cancer risks
of 10"6,10"5 and 10"4.
3. The occupational exposures to commercial degreasers show the greatest cancer risk when
compared to the spot cleaning exposure scenarios.
Acute Non-Cancer Risks:
1. There are acute non-cancer risks for developmental effects (i.e., cardiac defects) for most
occupational and residential exposure scenarios (i.e., MOEs were below the benchmark
MOEof 10).
2. The degreaser exposure scenarios show greater acute risks for developmental effects than
those reported for the spot cleaning exposure scenarios.
Chronic Non-Cancer Risks:
1. There are chronic non-cancer risks for a range of human health effects in both occupational
degreaser and spot cleaning exposure scenarios (i.e., MOEs were below the benchmark
MOE).
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2. The greatest concern is for developmental effects (i.e., fetal cardiac defects), followed by
kidney effects and then immunotoxicity, with an overall higher chronic risk for the
degreaser exposure scenarios. In general, this concern is irrespective of the type of
exposure (typical vs worst case) and the availability of room ventilation (LEV vs no LEV).
3. There are chronic risks for reproductive effects and neurotoxicity for degreaser worker
exposure scenarios and most of the degreaser bystander exposure scenarios. However, the
risks concerns for these effects are reported for fewer spot cleaning worker/bystander
scenarios and generally attributed to exposure conditions without room ventilation.
4. There are chronic risks for liver effects although the risks are less prominent than those
reported for other health effects domains. These risks are found only in the degreaser
worker/bystander exposure worst case scenarios and the spot cleaning worker/bystander
worst case scenarios with no LEV.
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Level Trichloroethylene on Heart Rate and Wakefulness-Sleep in Freely Moving Rats.
Sangyo Igaku/Japanese Journal of Industrial Health (Japan), 36(1), 1-8. (as cited in EPA,
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Ash, M., and I. Ash. 2009. 1,1,2-Trichloroethylene. In Ash, M., and I. Ash, Handbook of
Industrial Chemical Additives (2nd ed., Vol. 2). Synapse Information Resources.
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ATSDR (Agency for Toxic Substances and Disease Registry). 2001. Final Report: Evaluation of
Priority Health Conditions in a Community with Historical Contamination by
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Exposed to Organic Solvents. Scandinavian Journal of Work, Environment and Health,
15(5), 345-352. (as cited in EPA, 2011e).
Taskinen, H., P. Kyyronen, and K. Hemminki. 1994. Laboratory Work and Pregnancy Outcome.
Journal of Occupational Medicine, 36(3), 311-319. (as cited in EPA, 2011e).
Taylor, D. H., K. E. Lagory, D. J. Zaccaro, R. J. Pfohl, and R. D. Laurie. 1985. Effect of
Trichloroethylene on the Exploratory and Locomotor Activity of Rats Exposed During
Development. Science of the Total Environment, 47, 415-420. (as cited in EPA, 2011e).
Tibaldi, R., W. ten Berge, and D. Drolet. 2014. Dermal Absorption of Chemicals: Estimation by Ih
Skinperm. J Occup Environ Hyg, 11(1), 19-31.
Tikkanen, J., and 0. P. Heinonen. 1988. Cardiovascular Malformations and Organic Solvent
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591-600.
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Leader, and M. G. Wade. 2004. Exposure to Trichloroethylene and Its Metabolites Causes
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Yauck, J. S., M. E. Malloy, K. Blair, P. M. Simpson, and D. G. McCarver. 2004. Proximity of
Residence to Trichloroethylene-Emitting Sites and Increased Risk of Offspring Congenital
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APPENDICES
Appendix A
PRODUCTION VOLUME AND INVENTORY
UPDATE RULE DATA
In 2011, global consumption of TCE was 945 million Ibs with expected growth at about 1.5
percent annually over the next five years. The corresponding U.S. consumption was 255 million
Ibs (Glauserand Funda, 2012)31. here were two U.S. producers for TCE as of 2011: The Dow
Chemical Company in Freeport, TX and PPG Industries, Inc. in Lake Charles, LA (Glauser and
Funda, 2012). Exports of TCE from the U.S. have increased along with similar increases for all
chlorinated solvents (80 percent in 2011, 72 percent in 2010) (ICIS, 2010, 2012).
TCE production volume of 224.7 million Ibs were reported to EPA in the 2012 Chemical Data
Report (CDR). Seven companies reported using TCE in industrial/manufacturing activities: Dow
Chemical, Greenchem, PPG Industries Inc., Shin Etsu, Solvchem Inc., Triinternational Inc., and
Vinmar Overseas LTD, plus two other companies (EPA, 2013a). There were two other
companies that reported to 2012 CDR, but much of this information was claimed confidential
business information (CBI) and cannot be made available to the public. Data in tables A-5 to A-7
were extracted from the 2012 CDR records (EPA, 2013a).
Table A-l. National Chemical Information for TCE from 2012 CDR
Production Volume (aggregate)
Maximum Concentration (at manufacture or import site)
Physical form(s)
Number of reasonably likely to be exposed industrial manufacturing, processing,
and use workers (aggregated)
Was industrial processing or use information reported?
Was commercial or consumer use information reported?
224.7 million Ibs
>90%
Liquid
>1,000
Yes
Yes
This source material includes data or information derived from IMS Products provided to the U.S. EPA. IMS
products have been provided to the U.S. EPA for its internal use and in the context of a license agreement. By
receiving and accessing this material, you agree that IMS is not liable to you or any third party for your use of
and/or reliance on the IMS data and information contained in this document, and any such use shall be at your
own risk.
Page 151 of 212
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Table A-2. Summary of Industrial TCE Uses from 2012 CDR
Industrial Sector
(Based on NAICS)
All Other Basic Organic
Chemical Manufacturing
Industrial Gas Manufacturing
Wholesale and Retail Trade
Agriculture, Forestry, Fishing
and Hunting
All Other Basic Organic
Chemical Manufacturing
All Other Chemical Product and
Preparation Manufacturing
All Other Chemical Product and
Preparation Manufacturing
All Other Basic Organic Chemical
Manufacturing
Primary Metal Manufacturing
Industrial Function
Intermediates
Functional fluids
(closed systems)
Solvents
(for cleaning or degreasing)
Agricultural chemicals
(non-pesticidal)
Solvents
(for cleaning or degreasing)
Not Known or Reasonably
Ascertainable
Solvents
(which become part of product
formulation or mixture)
Not Known or Reasonably
Ascertainable
Solvents
(for cleaning or degreasing)
Type of Processing
Processing as a reactant
Processing as a reactant
Processing as a reactant
Processing as a reactant
Processing-repackaging
Use-non-incorporative activities
Processing-incorporation into
formulation, mixture, or reaction
product
Processing-repackaging
Processing-incorporation into
formulation, mixture, or reaction
product
Table A-3. TCE Commercial/Consumer Use Category Summary
Commercial/Consumer Product
Category
Intended for Commercial and /or
Consumer Uses or Both
Intended for Use in Children's
Products in Related Product
Category
Adhesives and Sealants
Both
No
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Appendix B REGULATORY HISTORY OF TCE AT THE USEPA
AND RELATED ACTIONS
j*:i__^^
The purpose of this section is to provide a brief regulatory history of TCE from the perspective
of the EPA. TCE has been subject to 25 final rules and notices issued by the Agency from 1979
to 2009 that were relevant or significant with regard to TCE. These 25 rules and notices were
promulgated by EPA's Office of Air and Radiation (OAR), the Office of Solid Waste and
Emergency Response (OSWER), the Office of Water (OW) and the Office of Pollution Prevention
and Toxics (OPPT).
EPA/OW initially identified TCE as a "toxic pollutant" in 1979 (EPA. 1979). TCE was classified as
a "priority pollutant" in 1982 and no discharges of TCE were allowed from steam electric power
generating point sources (EPA, 1982). EPA/OW then established a non-enforceable maximum
contaminant level goal (MCLG) of 0 mg/L for TCE in 1985 (EPA, 1985b). Two years later,
EPA/OW set a maximum contaminant level (MCL) of 0.005 mg/L for drinking water (EPA,
1987b) and set an effluent limitation of 69 u.g/L maximum daily average and 26 u.g/L maximum
monthly average for new and existing sources discharging to POTWs from the organic
chemicals, plastics, and synthetic fibers industrial category (EPA, 1987c). The following year,
EPA/OW prohibited injection of TCE into class I underground injection wells (EPA, 1988b). TCE
was identified by EPA/OW as a bioaccumulative chemical of concern pollutant in 1995 for a
final water quality guidance for the great lakes system. This established water quality criteria
for protection of human health by setting a human cancer value (HCV) of 29 u.g/L for drinking
water and 370 u.g/L for non-drinking water for the Great Lakes system (EPA, 1995b). In 1998,
EPA/OW identified TCE as a possible human carcinogen by establishing a national primary
drinking water regulation that specified the following consumer confidence report health effect
language: "some people who drink water containing trichloroethylene in excess of the MCL
[0.005 mg/L] over many years could experience problems with their liver and may have an
increased risk of getting cancer" (EPA, 1998e). EPA/OW identified TCE's major sources in
drinking water originating from "discharge from metal degreasing sites and other factories
(EPA, 1998)." EPA/OW is currently evaluating and revising TCE's MCL based upon analytical
feasibility (EPA. 2010).
EPA/OAR has listed TCE as a HAP from several different industrial emission sources in multiple
rules (EPA. 1985a. 1986a. 1994c. 1994d. 1998d. 2001b. 2002a. 2003. 2004b. 2007c. 2009).
including solvent cleaning operations (EPA, 1994d, 2007c) as well as a "probable or possible
human carcinogen" from operations including printing, coating, and dyeing of fabrics and other
textiles (EPA, 2003). EPA/OAR classified TCE as a group I chemical for emission standards for
equipment leaks in the synthetic organic chemical manufacturing industry (EPA, 1994c). In
addition, EPA/OAR identified TCE as a substitute for two ozone depleting chemicals, methyl
chloroform and CFC-113, for metals, electronics, and precision cleaning, in 2007 (EPA, 2007d).
Page 153 of 212
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EPA/OSWER set a reportable quantity of 100 Ibs (45.4 kg) for releases of TCE from vessels or
facilities in 1989 (EPA, 1989). EPA/OSWER also set a minimum required detection limit for TCE
of 37 mg/kg for hazardous waste combustors in 1998 (EPA, 1998b).
Although EPA/OPPT has only issued two notices relevant to TCE (EPA, 1994e, 2000d), other
voluntary information collection activities for TCE have occurred in the past. These activities
were primarily the result of two separate but related voluntary information collection activities:
data gaps identified by ATSDR (EPA, 1994e) and data gaps identified for pilot chemicals for
EPA's Voluntary Children's Chemical Evaluation Program (VCCEP) (EPA. 2000d).
EPA/OPPT published a notice for voluntary solicitation of testing proposals in order to be
considered for an enforceable consent agreement (ECA) negotiation in 1994 for 12 substances,
including TCE (EPA, 1994e). This notice was based on data gaps identified by ATSDR in
coordination with EPA. After ATSDR updated its data needs for TCE in 1999 (ATSDR. 1999). the
Halogenated Solvents Industry Alliance, Inc. (HSIA) responded with its intent to fulfill four of
seven identified data needs. These four data needs included developmental neurotoxicity,
developmental toxicity, immunotoxicity, and neurotoxicity via the oral route. HSIA entered into
a memorandum of understanding (MOD) with ATSDR in June of 2001 to fulfill these four data
needs.
Over the course of the next several years, from 2001 to 2007, HSIA completed and submitted
two studies to the Agency: a developmental toxicity and an immunotoxicity study via the
inhalation route in rats. These two studies had been planned to be extrapolated to the human
oral route using PBPK modeling. HSIA had also planned to fulfill the data need for neurotoxicity
via the oral route using PBPK modeling of existing published data from the inhalation route. In
addition, HSIA had planned to conduct a developmental neurotoxicity study in rats via the oral
route. HSIA did not fulfill its MOD for these four planned studies due to several factors,
including problems securing an appropriate lab, discontinuation of a strain of rat previously
used in their completed studies, and discrepancies with ATSDR regarding the completeness of
the three aforementioned studies using PBPK modeling.
Since 2008, no further action has been taken by EPA/OPPT with regard to TCE and its existing
data gaps identified by ATSDR.
B-2 Other Regulatory Actions in the U.S. and Abroad
TCE is listed on California's Safer Consumer Products regulations candidate chemicals list and
the Proposition 65 list of chemicals. California also lists TCE as a designated chemical for
biomonitoring because it has the potential to pose higher exposure rates in California in
comparison to other states. Minnesota classifies TCE as a chemical of high concern, while other
states, like Washington and Maine, have considered TCE for similar chemical listings. Several
additional states have various regulatory actions that range from reporting requirements to
contamination limits and use reduction efforts. Some examples include Massachusetts, New
Page 154 of 212
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York, Ohio, Colorado and Michigan as they evaluate and monitor exposure to mitigate health
risks.
TCE is listed in the European Union Authorisation List owing to its classification as carcinogen
(category IB), with a sunset date of April 21, 2014. In 2004, the United Kingdom completed a
Risk Assessment of TCE on behalf of the European Union. In March 2010, France presented a
proposal on the identification of TCE as substance of very high concern because of its
carcinogenic properties. A dossier was circulated to Member States and was made available on
the ECHA website. Comments were received by Member States and interested parties on the
proposal. The dossier was referred to the Member State Committee and on June 2010 the
Member State Committee agreed to identify TCE as substance meeting the criteria of for a
Candidate List of Substance of Very High Concern owing to its classification as carcinogen
(category 2).
In Canada, the first Priority Substances List (PSL1) was published in the Canada Gazette in 1989,
and the assessments of risks to human health or the environment posed by the 44 substances
on the list were completed within the legislated time frame of five years. Options to reduce
exposure for those substances determined to be "toxic" were and are being considered, in
consultation with stakeholders. Canada assessed TCE in 1993 and considered it as a "toxic"
under section 11 of the 1988 Canadian Environmental Protection Act (CEPA 1988). The TCE
assessment concluded that "trichloroethylene occurs at concentrations that may be harmful to
the environment, and that may constitute a danger in Canada to human life or health. It has
been concluded that trichloroethylene occurs at concentrations that do not constitute a danger
to the environment on which human life depends."
In Japan, TCE is consider a Class II substance (Class II Specified Chemical Substances are
substances that may pose a risk of long-term toxicity to humans or to flora and fauna in the
human living environment, and that have been, or in the near future are reasonably likely to be,
found in considerable amounts over a substantially extensive area of the environment). Japan
also controls air emissions and water dischargers containing TCE, as well as aerosol products for
household use and household cleaners containing TCE.
TCE is listed in the Australian National Pollutant Inventory (NPI), a programme run
cooperatively by the Australian, State and Territory governments to monitor common
pollutants and their levels of release to the environment. Reporting obligations may apply to
this chemical. This chemical is included on Australia's High Volume Industrial Chemicals List
(HVICL) because it is manufactured or imported in large quantities (1000 tonnes or
more). Australia requires a secondary notification if significant new information about TCE's
health and/or environmental effects becomes available, for example new data on the
mutagenic or reproductive effects. Notification will also be required if it is used in wool scouring
or any other new use resulting in a significant increase in the quantities imported into
Australia. Australia classifies TCE as a health, physicochemical and/or ecotoxicological hazard,
according to the National Occupational Health and Safety Commission (NOHSC) Approved
Criteria for Classifying Hazardous Substances. TCE has also been reviewed as a part of the
Page 155 of 212
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Australia National Industrial Chemicals Notification and Assessment Scheme (NICNAS) Priority
Existing Chemical (PEC) assessment process. The report is available here:
http://www.nicnas.gov.au/ data/assets/pdf file/0004/4369/PEC 8 Trichloroethylene Full R
eport PDF.pdf
Page 156 of 212
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Appendix C ENVIRONMENTAL FATE OF TCE
C-l Environmental Fate
Knowledge of the environmental fate (transport and transformation) of a compound is
important to understanding its potential impact on specific environmental media (e.g., water,
sediment, and soil) and exposures to target organisms of concern.
TCE is a volatile liquid with high vapor pressure, moderate water solubility, and high mobility in
soil. Most reported environmental releases of TCE are to air with much lower releases to
landfills and very little released to water (see Chapter 2, Section 2.2, Overview of Environmental
Fate and Releases of TCE). If released to air, degradation by sunlight and reactants in the
atmosphere is slow. If released to water, sediment, or soil, the fate of TCE is influenced by
volatilization from the water surface or from moist soil and by microbial biodegradation under
some conditions. The biodegradation of TCE in the environment is dependent on a variety of
factors and so a wide range of degradation rates have been reported (ranging from days to
years). TCE is not expected to bioconcentrate in aquatic organisms due to measured
bioconcentration factors of less than 1000.
C-l-1 Fate in Air
TCE does not absorb light greater than 290 nm very well; therefore, degradation of TCE by
direct exposure to light, if it is released to the atmosphere, is not expected to be an important
fate process (EPA, 1979). TCE is expected to undergo relatively slow atmospheric hydroxy
radical oxidation with an estimated atmospheric half-life of about 13 days (using Version 4.10 of
EpiSuite, EPA, 2012b). Half-life estimates using measured rate data have been reported in the
range of 1 - 11 days using hydroxy radical concentrations expected in relatively polluted and
pristine air, respectively (Howard et al., 1991). Phosgene, dichloroacetyl chloride (DCAC),
chloroform, and formyl chloride can be formed from the reaction of TCE with hydroxyl radicals
(EPA. 1980: Gay et al.. 1976: Kao. 1994).
C-l-2 Fate in Water
Volatilization from water surfaces will be an important fate process based upon TCE's measured
Henry's Law constant. However, its density may cause it to sink in the water column,
potentially increasing the aquatic residence time of TCE. Volatilization half-lives in an
experimental field mesocosm consisting of seawater, planktonic, and microbial communities
ranged from 10.7 to 28 days (Wakeham et al., 1983). In contrast, half-lives of evaporation from
laboratory water surfaces (distilled water) have been reported to be on the order of several
minutes to hours, depending upon the turbulence (Culver et al., 1991; Hutter et al., 1992). TCE
achieved only 19 percent of its theoretical biochemical oxygen demand (BOD) over the course
of a 28-day incubation period using the closed bottle (Organisation for Economic Co-operation
and Development [OECD] 301D) test, and thus is not considered readily biodegradable. It
achieved 2.4 percent of its theoretical BOD using an activated sludge inoculum in the modified
Page 157 of 212
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Ministry of International Trade and Industry test (MITI, OECD 301C) over the course of a 14-day
incubation period. It was not inherently biodegradable in a Zahn-Wellens test (OECD 302B).
These studies suggest that TCE will biodegrade slowly in surface waters. However, slow
photooxidation in water has been reported (half-life of 10.7 months) (Pilling et al., 1975).
Based on these studies, biodegradation and hydrolysis in surface waters are not expected to be
important environmental fate processes.
^C-l-3^
TCE is expected to have high mobility in soil based on measured soil organic carbon partition
coefficients ranging from 72 to 148. Volatilization of TCE from moist soil surfaces is expected to
be an important fate process given its relatively high Henry's Law constant. TCE is expected to
volatilize from dry soil surfaces based upon its high vapor pressure.
Both laboratory tests and field studies in the environment show wide variation in TCE
biodegradation rates. In some cases, laboratory studies have shown rapid biodegradation. TCE
has been shown to biodegrade under aerobic conditions by methanotrophic microbes in the
presence of other substrates and under anaerobic conditions (in suitable reducing
environments) in the presence of other organic matter. Without competent microorganisms
that can degrade TCE and favorable environmental conditions, TCE can persist in the
environment on the order of years.
Aerobic biodegradation of TCE by specialized communities of microorganisms has been
reported (Wackett et al., 1989). Biodegradation of TCE has also been shown to occur under
conditions where additional substrates have been added to the medium (Kao and Prosser,
1999; Muand Scow, 1994; Wilson and Wilson, 1985). Mixed microbial cultures of methane-
utilizing bacteria have been shown to degrade TCE in two days under aerobic conditions (Fogel
et al., 1986). However, there are several factors that can limit the aerobic biodegradation of
TCE, including TCE concentration, pH, and temperature. Toxicity of the degradation products
(e.g., dichloroethylene, vinyl chloride, chloromethane) to the degrading microorganisms may
also reduce the rates of biodegradation of TCE in aerobic soils.
Biodegradation of TCE also occurs under anaerobic conditions. Under these conditions, as
might be seen in flooded soils, sediment, or aquifer environments, TCE is biodegraded via
reductive dechlorination; the extent and rate of degradation are dependent upon the strength
of the reducing environment and other factors (McCarty, 1996). TCE half-lives in the field for
aquifer studies range from 35 days to over six years. Major products of biodegradation of TCE
in groundwater include dichloroethylene, chloromethane, and vinyl chloride (HSDB, 2012).
TCE contamination exists in the subsurface environment as a result of spills and leaking transfer
lines/storage tanks. Because of its density and low Koc, TCE will ultimately move downward in
the soil until an impermeable barrier is reached. This may occur when a TCE spill is of sufficient
magnitude or deep enough in soil for volatilization to be restricted. Once in soil, TCE can
become associated with soil pore water, enter the gas phase because of its Henry's Law
Page 158 of 212
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constant, or exist as a nonaqueous phase liquid (NAPL). It is possible that upward or downward
movement of TCE can occur in each of these three phases, thereby increasing the areal extent
of the original spill. Nonaqueous phase concentrations of TCE which are large enough to
overcome capillary forces will move downward into the aquifer. Once the water table is
penetrated, lateral flow may be mediated by the regional ground-water flow. Due to its high
density, the movement of free-phase TCE is still directed vertically until lower permeability
features are encountered. Once an impermeable layer is encountered, horizontal movement
will occur. Such movement may even be directed against the natural ground-water flow by the
effects of gravity. Since permeability is a function of the liquid as well as the medium, the
vertical movement of TCE through an aquifer is determined by geological properties of the
aquifer material; i.e., granular size of sand or clay lenses. TCE will tend to pool near these
impermeable features. Water passing over and around these pools may solubilize TCE so that it
can be spread throughout the aquifer. This pattern of release and distribution in aquifers and
TCE persistence have led to the widespread detection of TCE in groundwater and drinking water
supplies derived from the contaminated groundwater (EPA, 1992).
C-l-4 Bioconcentration
TCE is not expected to bioconcentrate in fish, with measured bioconcentration factors (BCFs) in
carp ranging from four to 17. TCE's low measured BCF value suggests that bioconcentration in
aquatic organisms is low (NITE, 2012). The estimated upper trophic level bioaccumulation
factor (BAF) for TCE is 24 (EPA, 2012b). Table C-2 provides a summary of the environmental fate
information for TCE.
C-l-5 Conclusions on Environmental Fate
TCE is a volatile liquid and if released to air, will be slowly degraded by atmospheric hydroxy
radicals. If released to water, volatilization to the atmosphere will be an important fate process
and biodegradation will be slow. In soil, TCE does not bind strongly to soil organic matter and if
not biodegraded at an appreciable rate, TCE can migrate through soil to groundwater. Based on
the experimental evidence and environmental fate data available, TCE is expected to have low
bioaccumulation potential in aquatic organisms (bioconcentration/bioaccumulation factor less
than 1000) and moderate persistence in the environment (environmental half-life of greater
than two months but less than six months).
Page 159 of 212
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Table C-l. Environmental Fate Characteristics of TCEa
Property
CASRN
Photodegradation half-life
Hydrolysis half-life
Biodegradation
Bioconcentration
Log Koc
Fugacity (Level III Model)b
Air (%)
Water (%)
Soil (%)
Sediment (%)
Persistenced
Bioaccumulationd
Value
79-01-6
13.2 days (estimated)
Does not hydrolyze under environmental conditions'3
19% after 28 days (not readily biodegradable)13;
4% after 28 days (not inherently biodegradable)13;
100% after 2 days (anaerobic conditions using mixed march cultures)13;
2.4% after 14 days (not readily biodegradable)0
BCF = 4.3-17 (measured in carp at 0.070 mg/L)c;
BCF = 4-16 (measured in carp at 0.007 mg/L)c;
BCF = 17 (measured in freshwater fish at 0.0087 mg/L)c;
BAF = 23.7 (estimated)3
2.17 (measured in silty clay Nebraska loam)c;
1.94 (measured in silty clay Nevada loam)c;
1.86 (measured in a forest soil)c;
1.8 (estimated)
35.4
54.2
10.1
0.3
P2 (moderate)
Bl (low)
Sources:
a EPA(2012b)
b ECB (2000)
c NITE (2012)
d EPA (1999)
Page 160 of 212
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Appendix D NAICS CODES FOR TCE DECREASING
An analysis of the North American Industry Classification System (NAICS) identified 78 different
industries that primarily use TCE as a degreaser (NAICS codes are listed in Table D-l).
Table D-l. TCE Used as a Degreaser Primarily in These Industries (USDOC, 2008) 1
NAICS Codes
33121
33272
33341
33422
33512
33531
33634
33641
33999
314999
321113
323116
325188
325998
326299
331111
331210
331419
331421
332111
332112
332116
332117
332211
332212
332311
332313
332431
332510
332618
332721
332722
332811
332812
332813
332912
332913
332919
332994
332996
332999
333132
333298
333311
333415
333921
333994
333999
334413
334414
334417
334419
334513
334515
335121
335211
335312
335313
335911
335921
335929
335999
336321
336340
336411
336413
336414
336510
337125
337127
339114
339992
339995
339999
488111
493110
811310
928110
Each number listed is a different industry that may be associated with TCE/degreasing
operations. Those interested may go to the following URL and type in a code -
http://www.census.gov/eos/www/naics/ (USDOC, 2008). For example, the following results are
seen when the listed numbers are searched:
33121:
1. 33121: Iron and Steel Pipe and Tube Manufacturing from Purchased Steel
2. 331210: Iron and Steel Pipe and Tube Manufacturing from Purchased Steel
332811 (results below slightly edited for simplicity):
1. Metal Heat Treating - This U.S. industry comprises establishments primarily engaged in heat
treating, such as annealing, tempering, and brazing, and cryogenically treating metals and
metal products for the trade.
2. Establishments primarily engaged in both fabricating and heat treating metal products are
classified in the Manufacturing sector according to the product made.
3. Annealing metals and metal products for the trade
4. Brazing (i.e., hardening) metals and metal products for the trade
5. Burning metals and metal products for the trade
6. Cold treating metals for the trade
7. Cryogenic treating metals for the trade
8. Hardening (i.e., heat treating) metals and metal products for the trade
9. Heat treating metals and metal products for the trade
10. Shot peening metal and metal products for the trade
11. Tempering metals and metal products for the trade
Page 161 of 212
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Appendix E ESTIMATION OF TCE EMISSION RATE AT SMALL DECREASING FACILITIES
i.
2.
3.
4.
NEI Point Sources (PS)
Point sources (PS) assumed to represent large facilities
Data obtained from: 2008neiv2_facility_process.zip
Filter data by 1) pollutant name, 2) relevant NAICS codes, 3)
source classification codes (SCC)for "Solvent-Degreasing"
sectorand4j unittype code(430)
Results for large facilities
a. TCE Air Emissions (1.48 Mlbs per year)
b. NumberofTCE-emittingfacilrties(154basedonnumber
of unique EIS identifiers)
Number of TCE-emittingdegreasing units (180 based on
number of unique EIS Unit Identifiers)
Numberof degreasing units perfacility(180/154 = 1.2)
c.
d.
NEI Nonpoint Sources (NPS)
1. Nonpoint sources (NPS) assumed to represent small facilities
2. Data obtained from: 2008neiv2_nonpoint.zip
3. Filter data by 1) pollutant name and 2) source classification codes (SCC) for
"Solvent-Degreasing" sector
4. Resultsforsmall facilities
a. TCE Air Emissions (2.86 Mlbs per year)
b. Number of NPS (1,779)
c. Based on EPA's 2006 risk assessment forthehalogenated solvent cleaning
source category, total number of facilities expected to be approximately
1,900. NumberofTCE-emittingfacilities = l»900 - 154 = 1,746
d. Assume one (1) degreasing unit per facility; smaller facilities (NPS) are
expected to have less degreasing units per facility than larger (PS) facilities.
TRI
1. Filter TRI data by relevant NAICS codes
2. 2008 TRI: totalTCE air emissions~2.55 Mlbs peryear
3. 2008 NEI: total TCE air emission ~4.34 (1.48 + 2.86) Mlbs
peryear
4. NEI TCE air emissions are about 2 times (2X) those
reported in TRI (4.34/2.55 = 2)
5. Use more recent data from 2011 TRI but adjust TRI TCE air
emissions (1.69 Mlbs peryear) by 2X = 3.38 Mlbs per year
6. Basedon2008 NEI,smallfacilitiesaccountfor66%oftotal
air emissions
a. TotalTCE air emissions from small facilities (66% of
3.38 Mlbs peryear = 2.23 Mlbs peryear)
b. Release per small facility (2.23/1,746 = 1,277 Ibs
peryear)
TCE Emission Rate
1. EPA's draft Generic Scenario on Use of Vapor Degreasers
a. Smallfacilitiesexpectedtooperate260daysperyearfor2 hours
per day (31,200 minutes peryear)
b. SoIvent-to-Air interface can vary from 0.28 to 0.87 mA2
c. Operating TCE emissions can range from 5 to 10 grams per minute.
2. Local exhaustventilation(LEV)canreducepotentialTCEemissionsby
90% (Wadden,etal., 1989). Based on NEI and TRI:
a. Potential TCE emissions escaping into work place with no LEV
(1,277*454/31,200 = 19 grams perminute)
b. Potential TCE emissions escaping into work place with LEV (10% of
19 grams perminute= 1.9 grams perminute)
3. Overall NESHAP emission limit is 150 kg/mA2/month
a. Based on this limit, potential TCE emissions escaping into work place
can vary from 16 to50 grams per minute (2,600 minutes per
month; Solvent-to-Airinterfaceof 0.28 to 0.87 mA2).
Page 162 of 212
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Appendix F ESTIMATION OF TCE EXPOSURES AT SMALL DECREASING FACILITIES
Solvent degreasing facility is partitioned into two
zones: Near-Field / Far-Field
_»
Qff
QHf^
Fai'-Fielcl
Near-Field
G
f
\ olaule Source
QFF
— '»
—
Results of Near-Field / Far-Field approximation
"Workers" are directly involved with degreasing operations; "Occupational
Bystanders" have the potential to be exposed to TCE but they are not directly
involved with degreasing operations; TWA = Time Weighted Average; LEV =
local exhaust ventilation; the OSHA PEL for TCE is 100 ppm (537 mg/mA3);
number of small facilities = 1,746.
Workers
Occupational
Bystanders
Inhalation Expsoyre
(8 hr TWA, ppm)
Small Industrial / Commercial facilities
With LEV
0.3 to 20
0.04 to 17
No LEV
3 to 197
0,4 to 172
Number of Workers
8,730
20,952
Potential worker exposures to TCE
1. Exposure monitoring data for TCE:
a. OSHA monitoring data from 2003 to 2010
was available; data was specific to TCE; this
data was filtered by relevant NAICS codes;
39 personal breathing zone time-weighted
average (TWA} measurements were
available. Exposure values ranged from
0.06 to 380 ppm (average: 45 ppm;
median: 17 ppm; 95th percentile: ~170
ppm)
b. EPA's exposure estimates are similar to
OSHA's data, being of the same order of
magnitude.
2. Number of workers and bystanders potentially
exposed to TCE:
a. Based on (ASTDR, 1997), approximately 17
workers per facility are potentially exposed
to TCE; this estimate is understood to
include both, workers that are and are not
(occupational bystanders) directly involved
with solvent cleaning operations.
b. Based on EPA's draft Generic Scenario on
Use of Vapor Degreasers, the number of
workers expected to be directly involved
with solvent cleaning operations is
estimated to be 5 workers per facility.
c. Occupational bystanders (17 — 5 = 12
occupational bystanders per facility)
Page 163 of 212
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Appendix G
CALCULATION OF TCE EXPOSURES AT SMALL
DECREASING FACILITIES
In Figure G-l, a solvent degreasing facility is partitioned into two zones: the near-field and the
far-field (Keil, C. B. et al., 2009). In occupational settings, this is done because contaminant
levels in the near-field are considered to provide a better representation of a worker's personal
breathing zone than those in the far-field. In other words, potential worker exposures depend
on how close a worker is to the emission source. For this risk assessment, the far-field
exposures represented those exposures received by occupational bystanders. That is to say,
those individuals who would be in the building (and perhaps even the room), but not physically
close to the volatile source as shown in Figure G-l.
Figure G-1. Illustration of an Imperfectly Mixed Room: Near-Field/Far-Field Approximation
of a Solvent Cleaning Facility
Note: Potential Worker Exposures Depend on How Close a Worker is to the Emission (Volatile) Source.
Far Field
Cpp
QFF
QNF
Near-Field
VNF
T
volatile Swire*
QNF
Near-Field Mass Balance
dC
NF
+
[G-l]
Page 164 of 212
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Far-Field Mass Balance
* FF—"71— = ^NpQNF — CppQNF — ^FpQpF [G-2]
Where:
VNF = near-field volume
VFF = far-field volume
QNF = near-field ventilation rate
QPP = far-field ventilation rate
CNF = average near-field concentration
CFp = average far-field concentration
G = average generation rate
t = time
At steady-state, Equations [G-l] and [G-2] can be reduced to the following:
G
-------
fc, =
QNFQFF
QFF)
[G-9]
_ /^Vjyp + QNF\
~ \n /
V V '
[G-10]
[G-ll]
= 0.5
fQNFVFF + VNF(QNF + QFF~)
\
VFF
fQNFVFF + VNF(QNF
[G-12]
A, = 0.5
'QNFV
FF
QFFy
VNFVFF
VNFVFF
[G-13]
Time-weighted-average (TWA) concentrations in the near-field and far-field can be calculated
as follows:
CNF dt
dt
G(kJ2 +
t2\ I /c2e lt:L k3e 2tl\\ [G
-)- ( iti+—^ Al~~Jj
Page 166 of 212
-------
CFF,TWA —
ffG(gUM*"-
[G-15]
+
-ti)
As indicated in section 2.2.2, for the purposes of this assessment, small industrial/commercial
degreasing processes are expected to operate 260 days per year for 2 hrs per day (EPA, 2001a).
In addition, EPA/OPPT assumed that there is no exposure at small industrial/commercial
degreasing facilities for 6 hrs per day. Thus, in order to calculate 8-hr TWA concentrations, the
results from Equations [G-14] and [G-15] were multiplied by a factor of 0.25.
For the purposes of mass transfer from and to the Near-Field, the Free Surface Area, FSA, is
defined to be the surface area that is available for mass transfer and is not necessarily equal to
the surface area of the Near-Field. For instance, if the Near-Field is defined to be a rectangular
region, as illustrated in Figure G-l, the Near-Field floor would not be available for mass transfer.
Thus, FSA would be less than the actual surface area of the Near-Field:
FSA = 2(LNF * tfWF) + 2(WNF * tfWF) + (LNF * WNF} [G-16]
Where: LNF,WNF,HNF are the length, width, and height of the Near-Field, respectively.
If the Near-Field indoor wind speed, vNFl is known and the area for mass transfer into and from
the Near-Field is equal, then the Near-Field ventilation rate, QNF, is given by:
QNF = 2* FSA * VNF
If the Far-Field volume, VFF, and the air exchange rate, AER, are known, then the Far-Field
ventilation rate, QFF, is given by:
QFF = VFF * AER [G-18]
Based on the model inputs in Table G-l, potential workplace TCE inhalation exposure values
can be estimated for workers in the near-field and for bystanders in the far-field (see Table G-
2).
Page 167 of 212
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Table G-l. Model Inputs for Small / Industrial Commercial Degreasing Facilities
Parameter
VFF
AER
VNF
LNF
WNF
HNF
FSA
ti
t2
G
Units
ft3
(m3)
1/hr
cm / s
(m/hr)
ft
(m)
ft
(m)
ft
(m)
ft2
(m2)
hr
hr
g/min
(mg/hr)
Parameter Values
10,593 to 70,620
(300 to 2,000)
2 to 15
10
(360)
10
(3.05)
10
(3.05)
6
(1.83)
340
(32)
0
2
5 to 50
(3E+5to3E+6)
0.5 to 5
(3E+4to3E+5)
Comments
Values supported by von Grote et al. (2003)
Values supported by von Grote et al. (2003)
and EPA(2013b)
Value is ~50th percentile supported by
Baldwin and Maynard (1998)
Assumes volatile source is centered in the
near-field and worker activities are within 5
feet of the emitting source
Adequate height to capture a typical
worker's breathing zone
Equation [G-16]
Starting time for Equations [G-14] and [G-15]
Ending time for Equations [14] and [15]; as
indicated earlier, small commercial/industrial
degreasing processes are expected to
operate for 2 hrs per day (EPA, 2001a)
No local exhaust ventilation (LEV; see Table
2-8 in section 2.3.2 in the main document
With LEV; potential operating TCE emissions
reduced by 90% (Wadden et al.. 1989)
Table G-2. Potential Workplace TCE Inhalation Exposures and Number of Workers Exposed
Type of Facility
Small commercial
degreasing facility
Potential Workplace TCE Inhalation Exposures
(8-hr TWA)
Near-Field
LEV
(ppm)
0.3
(low-end
estimate)
20
(upper-end
estimate)
No LEV
(ppm)
3
(low-end
estimate)
197
(upper-end
estimate)
Number of
Workers
8,730
Far-Field
LEV
(ppm)
0.04
(low-end
estimate)
17
(upper-end
estimate)
No LEV
(ppm)
0.4
(low-end
estimate)
172
(upper-end
estimate)
Number of
Workers
20,952
Page 168 of 212
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Appendix H
CALCULATION OF TCE EXPOSURES FROM SPOT
CLEANING AT DRY CLEANING FACILITIES
For background information on the Near-Field/Far-Field mass balance model, please refer to
Appendix G. Based on the model inputs in Table H-l, potential workplace TCE inhalation
exposures from spot cleaning at dry cleaning facilities were estimated for workers in the near-
field and for occupational bystanders in the far-field (see Table H-2).
Table H-l. Model Inputs for Dry Cleaning Facilities
Parameter
VFF
AER
VNF
LNF
WNF
HNF
FSA
ti
t2
G
Units
ft3
(m3)
1/hr
cm / s
(m/hr)
ft
(m)
ft
(m)
ft
(m)
ft2
(m2)
hr
hr
g/min
(mg/hr)
Parameter Values
7,062 to 70,620
(200 to 3,000)
Itol9
10
(360)
10
(3.05)
10
(3.05)
6
(1.83)
340
(32)
0
8
0.037 to 0.37
(2.22E+3to2.22E+4)
0.0037 to 0.037
(222to2.22E+3)
Comments
Values supported by von Grote et al. (2006)
Values supported by von Grote et al. (2006) and EPA
(2013b)
Values is ~50th percentile supported by
Baldwin and Maynard (1998)
Assumes volatile source is centered in the near-field and
worker activities are within 5 feet of the emitting source
Adequate height to capture a typical worker's breathing
zone
See Equation [G-16] in Appendix G
Starting time for Equations [G-14] and [G-15] (see
equations in Appendix G)
Ending time for Equations [14] and [15] (see equations
in Appendix G); as indicated earlier, dry cleaning
facilities are expected to operate for 8 hrs per day (BLS,
2012)
No local exhaust ventilation (LEV); see Table 2-11 in
section 2.4.2 in the main document
With LEV; TCE emissions reduced by 90% (Wadden et
al., 1989)
Page 169 of 212
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Table H-2. Potential Workplace TCE Inhalation Exposures and Number of Workers Exposed
Type of Facility
Dry Cleaning
Facility
Potential Workplace TCE Inhalation Exposures
(8-hr TWA)
Near-Field
With LEV
(ppm)
0.008
(low-end
estimate)
2
(upper-end
estimate)
No LEV
(ppm)
0.08
(low-end
estimate)
19
(upper-end
estimate)
Number of
Workers
36,000
Far-Field
With LEV
(ppm)
0.0007
(low-end
estimate)
2
(upper-end
estimate)
No LEV
(ppm)
0.007
(low-end
estimate)
18
(upper-end
estimate)
Number of
Workers
252,000
Page 170 of 212
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Appendix I EFAST2/CEM INDOOR MODELING, CONSUMER
BEHAVIOR PARAMETERS AND MODEL
COMPARISONS
The Exposure and Fate Assessment Screening Tool Version 2 (E-FAST2) Consumer Exposure
Module (CEM) performs assessments of exposures to common products to consumers. This
section describes the values that were chosen for the modeling parameters in CEM to provide
more support for the TCE exposure assessment. This material is also described in the E-FAST2
manual available on the EPA's webpage. The first section describes house and emission
properties, the second section describes the consumer use patterns and the final section will
discuss comparisons of CEM with other exposure models.
The default parameters used for household characteristics were all set to mean or median
values based on data found in the available literature and these were used in the TCE
assessment. Consumer behavior patterns were not set to EFAST2's default settings, which were
high -end values. Alternatively, a hypothetical scenario was created for the user of two
products containing TCE. Westat survey data provided some indication that the values used
were below mean or median for the mass but above the mean for the time spent in the room
of use. However, the Westat survey categories were not completely aligned with the
description of the products chosen to model in the exposure assessment section.
1-1 Default parameters used in CEM for emission and household
characteristics
Table 1-1 summarizes the selection and justification of exposure parameters for CEM for the
purposes of estimation of indoor air concentrations of TCE.
I -1 -1 Air Exchange Rate
The air exchange rate used by OPPT for the TCE model runs was the E-FAST/CEM default value of
0.45 air changes per hour (ACH). This choice is consistent with the recommended central value
per the current and prior editions of the Exposure Factors Handbook (EFH) (Figure 1-1) (EPA,
1997. 2011a).
Page 171 of 212
-------
Table 1-1. Summary of CEM Parameters for Estimation of TCE Indoor Air Concentrations
S. No
Modeling Input
Value
Justification/Source
Air exchange
rate (air
exchanges/hr)
0.45
Recommended 50th percentile value of residential air
exchange rate for all regions within the United States
(Koontz and Rector, 1995)
Overspray
fraction
(unitless)
0.01
Selection based on professional judgment (US EPA
2007a).
Whole House
volume (m3)
369
Value was obtained from EPA (1997). Although the
updated EFH (EPA, 2011a) provides a larger number (492
m3), the older value was retained in order to provide
more conservative estimates.
Emission rate
(hrs-1)
101.06
Estimated using Chinn's algorithm (Chinn, 1981) based
on E-FAST model documentation. This algorithm utilizes
molecular weight and vapor pressure to estimate
emission rates.
Inhalation rate
(m3/hr)
0.74- During use
0.61-After use
During product use: 0.74 m3/hr based on short-term
exposure at light activity level (EPA, 2011a)
After product use: 0.611 m3/hr (EPA, 2011a)
Short term inhalation values during light activity (male
and female combined) were taken from the following
age groups and averaged to create an estimate for
inhalation rate during product use. 21 to <31 years; 31
to <41 years; 41 to <51 years; 51 to <61 years; 61 to <71
years; and 71 to <81 years.
Body weight
(kg)
80
Mean value of body weights for all adults (>21 yrs), male
and female combined. Value based on EPA analysis of
NHANES 1999-2006 data (EPA. 2011a)
Interzonal
airflow rate
(m3/hr)
81.73
Air flow rate between the room of use (utility room or
zone 1) and the rest of the house (zone 2 (Koontz and
Rector. 1995)
Page 172 of 212
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Figure 1-1. Screen capture of Summary of Recommended Values for Residential Building Parameters
from the Exposure Factors Handbook (EPA, ZOlla)
Table 19-1. Summary of Recommended Values for Residential Building Parameters
Mean 10th Percentile Source
Volume of Residence3 492 in3 (central estimate)" 154 m3 (lower percentile)'' U.S. EPA 2010 analysis of U.S. DOE,
2008a
Air Exchange Rate 0.45 ACH (central estimate)*1 0.18 ACH (lower percemile)6 Koontz and Rector, 1995
Volumes vary with type of housing. For specific housing type volumes, see Table 19-6.
Mean value presented 111 Table 19-6 recommended for use as a central estimate for all single family homes, including
mobile homes and nmltifamily units.
1 0th p ere entile value from Table 19-8 recommended to be used as a lower percentile estimate-
Median value recommended to be used as a central estimate based across all U.S. census regions (see Table 19-24).
10^ percentile value across all U.S. census regions recommended to be used as a lower percentile value (see
Table 19-24).
ACH = Air changes per hour
1-1-2 Overspray fraction
The selection of a default overspray fraction of 0.01 in CEM was based on professional
judgment (EPA, 2007b). For the scenarios that were modeled, the E-FAST inhalation exposure
metrics were not sensitive to the overspray/aerosolized fraction. For example, for the TCE run
labeled "ID Number: 1 fixative user 21-78" the modeled peak concentration (Cp pot) was 66.97
mg/m3. If this fraction was changed to 0.25, then a nearly identical modeled peak concentration
of 67.04 mg/m3 was obtained. Here we are only using the peak concentration as a model
diagnostic, not as a tool to understand exposures for any time scale longer than 10 seconds.
1-1-3
Emission Rate
The emitted mass was handled in CEM in two ways. When an aerosol product is used, some of
the product does not reach the intended application surface but remains in the air. This
portion, commonly known as the overspray, is assumed to be 1% of the product emitted during
use. This results in the constant emission of TCE to the room air over the duration of use (0.5
hrs or 1 hour depending on the scenario). The remaining fraction (99%) that is striking the
intended application surface forms a film. This film is treated as an incremental source, as
described below (Figure 1-2).
Page 173 of 212
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Figure 1-2. Screen capture of E-FAST equations for estimation of emission rate
For a product that is applied to surface, such as a general purpose cleaner or a latex paint, an
incremental source model is used. This model assumes a constant application rate over the specified
duration of use, each instantaneously applied segment has an emission rate that declines exponentially
over time, at a rate that depends on the chemical's molecular weight (MW) and vapor pressure (VP).
In the case of a general purpose cleaner, the equation for exponentially declining emissions for
each instantaneously applied segment is as follows:
E(t) = E(0) x exp(-fo)
(Eq. 3-41)
where E (t) is the emission rate (mass/time) at time t (in hours), E (0) is the initial emission rate at time 0,
k is a first-order rate constant for the emissions decline (inverse hours), and t is elapsed time (hours). The
value of k is determined from an empirical relationship, developed by Chinn (1981), between the time (in
hours) required for 90 percent of a pure chemical film to evaporate (EvapTime) and the chemical's
molecular weight and vapor pressure:
Evap Time =
145
0 95*5
The value of k is determined from the 90 per cent evaporation time as follows:
EvapTime
(Eq. 3-42)
(Eq. 3-43)
Using Equation 3-42 to calculate EvapTime:
145
EvapTime —
(131.4 X 73.46)0-9546
Where,
Molecular weight (MW) = 131.4 g/mole
Vapor Pressure (VP) = 73.46 torr
Hence, EvapTime = 0.023 hrs or 1.36 min
Using Equation 3-43 to calculate Emission Rate (k):
k _ In(lO)
1.36
Hence, Emission Rate (k) = 101.06 hrs"1 or 1.68 min"1
Page 174 of 212
-------
Because Chirm's algorithm assumes a pure chemical film, it tends to produce a lower-bound
estimate of the evaporation time; thus, overestimates the peak concentration. In products that
are a mixture of chemicals, interaction forces between the different chemicals could alter the
evaporation rate of individual constituents. However, this is only a concern for the spray
fixative since the degreaser use results in only TCE on the surface, which is consistent with
Chinn's study (Chinn. 1981).
In the simulation done for this assessment, the outcome was not expected to be strongly
dependent on the exact value of k due to the long time period the consumer spent in the room
of use after the period of product application. All of the TCE mass was expected to enter the air
before the user leaves the room even if the k value was adjusted to be less conservative.
Currently the evaporation time for 90% of the TCE in the film on the application surface (0.023
hrs or 1. 36 min) was much less than the 2 hrs the user spent in the room of use, about two
orders of magnitude. Even if this value were to increase, due to intermolecular interactions
within a more complicated mixture decreasing the emission rate, it would likely still be less
than the 2 hrs spent in the room of use.
1-1-4 House Volume and movement within the home
CEM currently uses a default house volume (369 m3) that is based on the mean value from the
1997 edition of the EFH (EPA, 1997). The 2011 edition of the EFH recommends a larger value of
492 m3 in recognition of the trend toward larger houses over the time between successive
editions (EPA, 2011a; see Table 19-1 extract on page 1). The 50th% value (median) in the 2011
EFH edition is 395 m3, so the mean volume used from the 1997 EFH is still close to the most
recent median value.
Examination of tables 19-8 and 19-9 of the 2011 EFH indicate that the increased mean house
volume is mostly due to the increase in the volume of the homes at the high end of the
distribution (75th percentile and above), whereas the homes in the lower 50th percentile are
relatively unchanged (EPA, 2011a). The lower value that was used would tend to produce more
conservative exposure estimates, other things being equal, due to the smaller volume available
for dispersal/dilution of the emitted TCE.
However the exposure values for the user could be more impacted by the size of the room
selected during use. The volume assigned to the room of use was 20 m3 for a utility room. The
utility room in this case served as a proxy for a hobby/craft room that might be represented, for
example, by a 9 ft x 10 ft room. With a ceiling height of 8 ft, the room volume would be 720 ft3
or ~ 20.4 m3. This room is zone 1 in the CEM simulations for the two products; zone 2 is the rest
of the house (349 m3). The user and bystander move about the home according to a
hypothetical behavior pattern constructed to represent a day spent mostly indoors. Since the
behavior patterns do not involve the residents entering the hobby room except to use the
product, the user spends the rest of the time either in zone 2 or outside (where there is no
chemical exposure) and the bystander spends the entire 24 hrs either in zone 2 or outside.
Page 175 of 212
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1-1-5 Inhalation rate and body weight
We used the 2011 EFH to obtain the inhalation rate and body weight values for the simulation
(EPA, 2011a). These values were based on the NHANES data (1999-2006) and correspond to the
age groups reported in the 2011 EFH. It is important to note that in the exposure assessment
only the exposure doses will be affected by these parameters. Indoor air concentrations are
determined by the product use patterns, the volume of the room and of the house, and the
physical-chemical properties of TCE. Body weight and inhalation rate do not change the indoor
air concentrations that the model calculates. The indoor air concentrations are the only part of
the exposure assessment that is used in the final risk assessment.
1-2 Consumer behavior patterns
E-FAST2/CEM requires the input of consumer behavior pattern information, including mass of
product used, duration of use, time spent in the room of use and the volume of the room of
use. By default, E-FAST2/CEM uses pre-set, high-end values for a variety of consumer use
scenarios when use information is not available for specific products. Under these conditions,
the model results tend to over predict the exposure.
EPA/OPPT did not have consumer behavior pattern information for the products being
evaluated in this assessment. Instead of using the E-FAST2/CEM's default inputs, EPA/OPPT
used professional judgment to inform the selection of hypothetical input parameters and
assumptions representing the consumers' behavior patterns.
Commenters suggested that EPA/OPPT should look through the Westat survey data (EPA,
1987a)32 to look for information that could be used within the TCE assessment. EPA/OPPT
frequently uses the 1987 Westat survey data for consumer modeling because the report
provides information about a range of product categories. A closer examination of the Westat
data revealed that the Westat survey categories were not well aligned with either of the
products modeled in this assessment.
Table 1-2 provides a summary of the information provided in the Westat survey and how it
compares to the values used in this assessment. The spray degreaser product had some
similarity with two categories from the survey: solvent type cleaning fluids or degreasers and
brake cleaners/quieters. Although the survey values were not used in the TCE's consumer
model assessment, the Westat data supplies an indication that the product masses used for the
degreaser scenario and that the room of use are not unreasonable high (i.e. overly
conservative). The time spent in the room of use is potentially closer to a 90th percentile value.
32 The Westat study was prepared for EPA by Battelle EPA(1987a).
Page 176 of 212
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Table 1-2. Comparison of Westat Survey Data and EPA's TCE Degreaser Simulation Values
Westat data for solvent type
cleaning fluids or degreasers
Time spent using product
Time spent in room
after use
Amount of product
per event
Room of use
Mean
29.48 min
33.29min
9.45 oz
(268 g)
Median
(50th%)
15.0 min
3.0 min
3.3 oz
(94 g)
90th %
60.0 min
60.0 min
16.0 oz
(454 g)
Household location
Garage 12.2%
Outside 28.0%
Inside room 49.1%
Westat data for brake cleaners/quieters
Time spent using product
Time spent in room
after use
Amount of product
per event
Room of use
Mean
23. 38 min
10.27 min
6.26 oz
(177 g)
Median
(50th%)
15.0 min
0.0 min
4.0 oz
(113 g)
90th %
49.5 min
30.0 min
12.0 oz
(340 g)
Household location
Garage 17.7%
Outside 77.1%
Inside room 5.2 %
EPA TCE Degreaser
Simulation values
60 min
60 min
0.85 oz
(24 g)
Utility room
Simulation values
60 min
60 min
0.85 oz
(24 g)
Utility room
1-2-1 Westat data for solvent type cleaning fluids
The description of the Westat data for solvent-type cleaning fluids matched well with the
intended use of the spray degreaser product that EPA/OPPT modeled as seen in Table 1-2.
Unfortunately, a majority of the respondents in the Westat survey stated that they were
describing their use of a liquid product (74.4%), not an aerosol product (25.6 %). EPA/OPPT had
difficulties in making direct comparisons of the survey's masses to those used in the exposure
simulations.However, it is the only currently available data that EPA/OPPT found closely
relating to the products under evaluation.
The mass of product (24 g) that EPA/OPPT used was well below the mean and median values
reported by survey respondents. E-FAST2/CEM is highly sensitive to the mass of product used
suggesting that the EPA/OPPT's simulated exposure would be potentially less conservative (i.e.
show lower air concentrations) if users of the TCE degreaser fit into similar use patterns to the
Westat survey.
Page 177 of 212
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The plurality of Westat-surveyed users (49.1%) also stated that they used the product in a room
inside the house that was not the basement or the living room, making it plausible that the
modeled room have a relatively small volume. This supports the selection of the utility room for
the spray degreaser scenario in this exposure assessment. The time of use selected for the
simulation is at the high end of the survey data (90th%), which would result in more
conservative exposures (i.e. higher exposures) since the selected value was twice the mean
value reported by Westat.
1-2-2 Westat data for brake cleaners/quieters
The description of this product category in the survey matches less well than the previous
survey group. The description for the chosen product in the TCE assessment does not describe
use in an automotive setting but the function and chemical composition of the product may be
similar and a large section of the respondents in the Westat survey for this category stated that
they had used an aerosol product (65.6%). As before the mass used in the TCE simulations was
well below both the mean and median survey values. Since this is an automotive product
category a large fraction of the respondents used the product outdoors or in the garage. If
users are outdoors it would significantly decrease exposures, but it is not clear what fraction of
the chosen product's users take this precaution. The time of use chosen in the simulation was
also larger than the surveyed data for this category.
1-2-3 Summary of Westat data
The survey data provided an indication that the room of use was appropriate and that the mass
used in the simulations was potentially on the low end. The time of use and the subsequent
time spent in the room after use were the only parameters in the simulation that could be
properly described as high-end from comparisons with the survey data. All of these
comparisons suffer from the absence of a Westat survey category that matches well with the
two products selected for the TCE assessment, a solvent spray degreaser and a spray fixative.
1-3 Comparison of EFAST CEM with MCCEM
Peer reviewers recommended that EPA/OPPT attempt to use other models besides CEM, such
as the Multichamber Concentration and Exposure Model (MCCEM) to calculate exposures for
the TCE assessment. Reviewers also were interested if CEM had been validated or compared to
other models that simulate indoor exposures to chemicals. Thus, EPA/OPPT compared the E-
FAST2/CEM model with MCCEM to address peer review comments.
The inhalation exposures determined by E-FAST2/CEM use the same calculation engine as
MCCEM. MCCEM is a peer-reviewed model that EPA/OPPT uses for assessing consumer
inhalation exposure. MCCEM is considered to be a higher tier model than CEM as it is able to
use measured chamber data for mass released due to product use to calculate exposures. An
earlier, DOS-based version of MCCEM was formally presented to the modeling community in a
1991 issue of the Indoor Air journal. In 1998, EPA peer reviewed a Windows-based version of
MCCEM. Unfortunately, chamber data were not available for the TCE-containing products that
Page 178 of 212
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were chosen in this assessment. Thus, EPA/OPPT could not run the MCCEM model for the
characterization of consumer exposures.
MCCEM was the subject of a model evaluation effort for EPA (Figure 1-3) whereby its outputs
were compared with those from two other well-recognized indoor-air models. These two
models were CONTAM developed by National Institute of Standards and Technology (NIST), and
INDOOR developed by EPA's Indoor Environmental Management Branch (IEMB) in Research
Triangle Park, NC. Results from all three of the models were in very close agreement for
cases/scenarios that were included in the evaluation effort. Lastly, a quick comparison of the
two models (EFAST's CEM and MCCEM) yielded almost identical results, for the same set of
input values.
Figure 1-3. Screen Capture of MCCEM Model Evaluation Efforts
Model Evaluation
Available Help Screens
Overview Example House Run Time Emissions Sinks Activities Dose Monte Cado Potions Report Execution Model Evaluation References
The Windows version of MCCEM was patterned after an earlier DOS version that has been described in the peer-reviewed
literature (Koont: and Nagda 1991). To verify that the algorithms in MCCEM for Windows were operating correctly and to
evaluate its accuracy, advantage was taken of the fact that the DOS version already had undergone fairly extensive evaluation.
This evaluation included comparisons with outputs from two other well-recognized indoor-air models - CONTAM (developed by
the National Institute for Standards and Technology) and INDOOR (developed by the USEPA Office of Research and
Development). The data sets used for the evaluation and the comparative results have been described in a report prepared for
the U.S. Environmental Protection Agency (Koontz, Rector, and Nagda 1992).
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Appendix J
CONVERTING E-FAST ADRs TO AIR
CONCENTRATIONS
The Exposure and Fate Assessment Screening Tool Version 2 (E-FAST2) Consumer Exposure
Module (CEM) performs assessments of exposures to common products to consumers. The
exposure values generated using the E-FAST/CEM models are in mg/kg-bw/day (Table J-l). The
only output in the acute exposure scenario expressed as a concentration was the peak
concentration, which represented the maximum concentration in air calculated by the model
during any 10-second time step during (in this case) 24 hrs. This value did not realistically
describe a 24-hr exposure, even as a worst-case scenario (all peak concentrations can be found
in documents cited in the supplementary documents containing the CEM runs.).
Table J-l. Estimated TCE Potential Acute Dose Rates from the Residential Indoor Use of
Solvent Degreasers or Clear Protecting Coating Sprays
Age
(yrs)
<1
1-2
3-5
6-10
11-15
16-20
21-78
Clear Protective Coating
Spray User
ADRp0t (mg/kg-bw/day)
NA
0.5
0.4
Clear Protective Coating
Spray Bystander
ADRp0t (mg/kg-bw/day)
0.5
0.4
0.4
0.3
0.2
0.2
0.1
Solvent
Degreaser User
ADRp0t
(mg/kg-bw/day)
NA
3
3
Solvent
Degreaser Bystander
ADRp0t (mg/kg-
bw/day)
3
3
2
2
1
1
0.8
Notes: ADRpot= potential acute dose rate; NA = not applicable
Thus, to convert the E-FAST CEM outputs from mg/kg-bw/day to ppm, we used the equation for
the potential acute dose rate reported in the E-FAST manual (EPA, 2007b). The general
expression for the potential acute dose rate (ADRpot) is as follows:
ADRpot = (Cair x InhR xFQx DEv x ED)
BWxAT
where:
ADRp0t = potential acute dose rate (mg/kg-bw/day)
Cair = exposure concentration (mg/m3)
InhR = inhalation rate (m3/hr)
FQ = frequency of product use (events/year)
DEv = duration of an event (hour/event)
ED = exposure duration (years of product usage)
BW = body weight (kg)
AT = averaging time (days)
Page 180 of 212
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Rearranging and simplifying this equation to calculate an approximation for Cair over the 24-hr
averaging time for the ADRPOT results in the following equation:
Cair =s ADRsot x B',V
InkR .x 24
This simplification is reasonable since the averaging time for acute exposure is one day (24 hrs).
In both scenarios, the frequency is just once per day. Although the duration of the event for the
two consumer scenarios is either one hour (degreaser) or 0.5 hrs (clear protective coating
spray)33, for the purposes of this exercise and to convert the model output to a more useable
exposure value to compare to the hazard value, there is no correction for this difference. This
assumption is still conservative since the values generated were reasonably high exposures that
probably overestimated the actual exposures.
An example calculation is presented below, since the final value is in mg/m3 and the desired
units will be in ppm. All calculated values are presented in Table J-2.
For the clear protective coating spray use, 21- to 78-yr-old user:
ADRpot = 0.45 mg/kg-bw/day
InhR (during use; 0.5 hrs) = 0.74 m3/hr
InhR (other times; 23.5 hrs) = 0.611 m3/hr
BW = 80 kg [using 2011 Exposure Factors Handbook; EPA (2011a)1
Cair = (0.45 mg/kg-bw/dav) (80 kg)
[0.74 m3/hr x 0.5 hr) + [0.611 m3/hr x 23.5 m3/hr)
= 2.4 mg/m3; converting to ppm34 = 0.446 ppm (rounded to 0.4 ppm to use a single
significant figure given the assumptions in the back-calculation).
33 However, for the user in both scenarios, the inhalation rates were slightly higher during use of the product, as
stipulated in the model outputs. Thus, for the degreaser use, an inhalation rate of 0.74 m3/hr (for 21 to 78 year
olds, 0.72 m3/hr for the 16 to 20 year olds) was used for one hour, and 0.611 m3/hr (for 21 to 78 yr olds, 0.679
m3/hr for the 16 to 20 yr olds) for the remaining 23 hrs. For the clear protective coating spray use, the higher,
user inhalation rate was used for 0.5 hrs, with the "normal" rate used for 23.5 hrs. This correction was not done
for any bystander scenario.
34 1 ppm = 5.374 mg/m3.
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Table J-2. Estimated TCE Inhalation Calculated Concentration in Air (Over Course of Day) from Use of
Two Consumer Products Indoors at Residences
Ages
<1
1-2
3-5
6-10
11-15
16-20
>21
Clear Protective Coating
Spray User
ADRpot (ppm)
NA
NA
NA
NA
NA
0.4
0.4
Clear Protective Coating
Spray Bystander
ADRpot (ppm)
0.1
0.1
0.1
0.1
0.1
0.1
0.1
Solvent
Degreaser User
ADRpot (ppm)
NA
NA
NA
NA
NA
2
2
Solvent Degreaser
Bystander
ADRpot (ppm)
0.8
0.8
0.8
0.8
0.8
0.8
0.8
Notes:
NA = not applicable
1 ppm = 5.374 mg/m3
As seen in the table J-2, each age group is exposed to the same modeled air concentrations and
therefore all age group have the same ADRpot(ppm). The conversion from dose to air
concentrations resulted in 24-hr time averaged indoor air concentrations for TCE (ppm) that
were not sensitive to user specific characteristics such as body weight or respiration rate. This is
why the same value was present throughout each column in Table J-2. Values in Table J-2 were
the only values used in the risk assessment.
The age groups are present in Tables J-l and J-2 and model output sheets in the supplementary
documents, which reflect the output values of the CEM model. Note that CEM also assumes
that the user will be over 16. However, Table J-2 shows that the age groups are irrelevant for
the calculated concentrations of TCE in the air. In section 2.5.4 of the risk assessment, age
groups were collapsed within each user or bystander use scenario. Since there is not
sufficiently refined data to create different consumer behavior patterns for different age
groups, EPA/OPPT assumed that younger users (<16) of degreaser and arts/crafts products
would be exposed to the same concentrations as older users (>16).
Page 182 of 212
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Appendix K RISK ASSESSMENT GUIDELINES, LITERATURE
SEARCH STRATEGY, STUDY SELECTION AND
DATA QUALITY CRITERIA FOR THE TCE IRIS
TOXICOLOGICAL REVIEW AND OPPT STUDY
REVIEW
WC!C!C!C!C!C!C!C!C!C!C!C!C!ro^
K-l Risk Assessment Guidelines
The description below was extracted from the TCE IRIS assessment published in September
2011 (EPA. 2011e. Chapter 1. pages 1-1 to 1-2).
Development of these hazard identification and dose-response assessments for TCE has
followed the general guidelines for risk assessment as set forth by the National Research
Council (NRC, 1983). U.S. Environmental Protection Agency (U.S. EPA) Guidelines and Risk
Assessment Forum Technical Panel Reports that may have been used in the development of
this assessment include the following:
1. Guidelines for the Health Risk Assessment of Chemical Mixtures (EPA, 1986c)
2. Guidelines for Mutagenicity Risk Assessment (EPA, 1986b)
3. Recommendations for and Documentation of Biological Values for Use in Risk Assessment
(EPA. 1988a)
4. Guidelines for Developmental Toxicity Risk Assessment (EPA, 1991)
5. Interim Policy for Particle Size and Limit Concentration Issues in Inhalation Toxicity (EPA,
1994a)
6. Methods for Derivation of Inhalation Reference Concentrations and Application of Inhalation
Dosimetry (EPA. 1994b)
7. Use of the Benchmark Dose Approach in Health Risk Assessment (EPA, 1995c)
8. Guidelines for Reproductive Toxicity Risk Assessment (EPA, 1996)
9. Guidelines for Neurotoxicity Risk Assessment (EPA, 1998a)
10. Science Policy Council Handbook: Risk Characterization (EPA, 2000b)
11. Benchmark Dose Technical Guidance Document (EPA, 2000a)
12. Supplementary Guidance for Conducting Health Risk Assessment of Chemical Mixtures (EPA,
2000c)
13. A Review of the Reference Dose and Reference Concentration Processes (EPA, 2002b)
14. Guidelines for Carcinogen Risk Assessment (EPA, 2005a)
15. Supplemental Guidance for Assessing Susceptibility from Early-Life Exposure to Carcinogens
(EPA. 2005b)
16. Science Policy Council Handbook: Peer Review (EPA, 2006c)
17. A Framework for Assessing Health Risks of Environmental Exposures to Children (EPA,
2006b)
Page 183 of 212
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K-2 Literature Search Strategy
When developing the TCE IRIS assessment, the literature search strategy was based on the
chemical name, Chemical Abstracts Service Registry Number (CASRN), and multiple common
synonyms (EPA, 2011e). Any pertinent scientific information submitted by the public to the IRIS
Submission Desk was also considered in the development of this document.
Primary, peer-reviewed literature identified through December 2010 was included where that
literature was determined to be critical to the assessment. The relevant literature included
publications on TCE which were identified through Toxicology Literature Online (TOXLINE), the
U.S. National Library of Medicine's MEDLINE, the Toxic Substance Control Act Test Submission
Database (TSCATS), the Registry of Toxic Effects of Chemical Substances (RTECS), the Chemical
Carcinogenesis Research Information System (CCRIS), the Developmental and Reproductive
Toxicology/Environmental Teratology Information Center (DART/ETIC), the Environmental
Mutagens Information Center (EMIC) and Environmental Mutagen Information Center Backfile
(EMICBACK) databases, the Hazardous Substances Data Bank (HSDB), the Genetic Toxicology
Data Bank (GENE-TOX), Chemical abstracts, and Current Contents. Other information, including
health assessments developed by other organizations, review articles, and independent
analyses of the health effects data were retrieved and included in the assessment when
appropriate (EPA, 2011e).
K-3 Study Selection and Data Quality Criteria Which Served As the
Basis for This Assessment
• Epidemiology data: Study quality evaluation criteria and a general format for the capture
of epidemiology study data and characterization have previously been developed in the
IRIS program and have been presented in the Guidelines for Developmental Toxicity Risk
Assessment (EPA, 1991). These factors include the study power, potential bias in data
collection, selection bias, measurement biases associated with exposure and outcome,
and consideration of potential confounding and effect modification. This format was used
to summarize study information and observed strengths, biases, and confounding factors
for each study (Figure K-l).
• Animal toxicology data: Study quality evaluation criteria for in vivo, in vitro, and in ovo
toxicology studies were developed.These criteria included considerations described in
(EPA, 1991) and focused on the adequacy of study design and documentation of
information on the test animals (e.g., species, strain, source, sex, age/lifestage/embryonic
stage), environment (e.g., husbandry, culture medium), test substance (e.g.,
identification, purity, analytical confirmation of stability and concentration), treatment
(e.g., dose levels, controls, vehicle, group sizes, duration, route of administration),
endpoints evaluated (e.g., schedule of evaluation, randomization and blinding procedures,
assessment methods), and reporting (quality and completeness) (Figure K-2). Analyses of
the study strengths and limitations were also conducted.
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Figure K-l. Study Quality Considerations for Epidemiological Studies
Feature
Selection
Attrition
Detection
Participants
Comparability
Attrition Rate
Length of
Follow-Up
Exposure
Characteristics
Outcome
Assessment
Confounding
Variables or
Exposure
Statistical Tests
Example Questions
• Were inclusion and exclusion criteria applied consistently
across study groups?
• Are baseline characteristics similar between groups? If not,
did the analysis control for differences?
• Is the comparison group appropriate, including having both
exposed and non-exposed subjects drawn from the same
population:
• Was the attrition rate uniformly low?
• In cohort studies: Dose the length of follow-up differ
between groups?
• In case-control studies: Is the time period between
exposure between exposure/intervention and outcome the
same for cases and controls?
• Was follow-up long enough to assess the outcome of
interest?
• What is the level of exposure misclassification?
• Is there an adequate level of exposure variability to detect
an effect?
• Are there adequate numbers of persons exposed at various
exposure levels to detect a dose-response effect?
• What is the extent of reliance on imputed exposure levels?
• Were the outcome assessors blinded to the exposure or
intervention status of participants?
• Is there confidence that the outcome of interest preceded
exposure?
• Are confounding variables assessed using reliable and
consistent measures?
• Did researchers adjust or control for other exposures or
interventions that are anticipated to bias results?
• Are statistical analyses performed with reliable tests and
implemented consistently?
Page 185 of 212
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Figure K- 2. Study Quality Considerations for Animal Studies
Feature
Example Questions
Exposure
Quality
•Were the exposures well designed and tightly
controlled?
•Was the test article/formulation adequately identified
and characterized? Are co-exposures expected as a
result of test article composition?
• Isthe administration route relevant to human
exposure?
•Are the exposure levels relevant?
• Inhalation exposure: Were analytical concentrations in
the test animals' breathing zone measured and reported
(i.e., not just target or nominal concentrations)?
• Inhalation exposure: For aerosol studies, werethe mass
median aerodynamic diameter and geometric standard
deviation reported?
• Inhalation exposure: Was the chamber type appropriate?
Dynamic chambers should be used; static chambers are not
recommended.
• Inhalation exposure: Were appropriate methods used to generate
the test article and measure the analytical concentration?
• Diet/Water Exposure: Was consumption measured to allow for
accurate dose determinations? Were stability and homogeneity
of the test substance maintained? Was palatability an issue?
•Gavage Exposure: Was an appropriate vehicle used? Are there
any toxicokinetic differences due to bolus dosing? Consider
relevance to human exposures.
Test Animals
•Were the test animals appropriate for evaluation of the
specified effect(s)?
•Were the species, strain, sex, and/or age of the test
animals appropriate for the effect(s) measured?
•Were the control and exposed populations matched in
all aspects other than exposure?
• Were an appropriate number of animals examined, based on
what is known about the particular endpoint(s) in question?
• Were there any notable issues regarding animal housing or food
and water consumption?
Study Design
• Isthe study design appropriate for the effect(s) and
chemical analyzed?
•Were exposure frequency and duration appropriate for
the effect(s) measured?
•Were anticipated confounding factors caused by
selection bias controlled for in the study design (e.g.,
correction for potential litter bias; randomization of
treatment groups)?
•Was the timing of the endpoint evaluation (e.g., latency
from exposure) appropriate?
•Was it a Good Laboratory Practices (GLP) study?
• Was it designed according to established guidelines (e.g., EPA,
OECD)? Was it designed to specifically test the endpoint(s) in
question?
• Did the study design include other experimental procedures (e.g.,
surgery) that may influence the results of the toxicity endpoint(s)
in question? Were they controlled for?
• Was the study design able to detect the most sensitive effects in
the most sensitive population(s)?
• Were multiple exposure groups tested? Was justification for
exposure group spacing given? Was recovery or adaptation
tested?
Toxicity
Endpoints
•Are the protocols used for evaluating a specific
endpoint reliable and the study endpoints chosen
relevant to humans?
•Are the endpoints measured relevant to humans? Do
the endpoints evaluate an adverse effect on the health
outcome in question?
•Werethe outcomes evaluated according to established
protocols? If not, were the approaches biologically
sound? Were any key protocol details omitted?
• Were all necessary control experiments performed to allow for
selective examination of the endpoint in question?
•As appropriate, were steps taken to minimize experimenter bias
(e.g., blinding)?
• Does the methodology employed represent the most appropriate
and discriminating option for the chosen endpoint?
Data
Presentation
and Analysis
•Were statistical methods and presentation of data
sufficient to accurately define the direction and
magnitude of the observed effect(s)?
•Are the statistical methods and comparisons
appropriate?
•Was sufficient sampling performed to detect a
biologically relevant effect (e.g., appropriate number of
slides examined)?
• Does the data present pooled groups that should be displayed
separately (e.g., pooled exposure groups; pooled sexes) and/or
analyzed separately?
• Was an unexpectedly high/low level of within-study variability
and/or variation from historical measures reported or explained?
•As appropriate, were issues such as systemic and maternal
toxicity (e.g., body weight) considered?
Reporting
•Are descriptions of study methods and results for all
endpoints sufficient to allow for study quality
evaluations?
•Were the details of the exposure protocols and
equipment provided?
•Were test animal specifics adequately presented?
•Are the protocols for all study endpoints clearly
described? Is sufficient detail provided to reproduce the
experiment(s)?
•Are the statistical methods applied for data analysis provided and
applied in a transparent manner? Was variability reported?
• Did the study evaluate a unique cohort of animals (i.e., are
multiple studies linked)?
•Are group sizes and results reported quantitatively for each
exposure group, time-point, and endpoint examined?
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Appendix L
LIST OF ORAL AND INHALATION STUDIES SUITABLE FOR NON-CANCER
DOSE-RESPONSE ANALYSIS IN THE TCE IRIS ASSESSMENT
L-l
Liver Effects
Target
Organ
Liver
Species
(male)
Mouse
(male)
Rat
(female)
Route of
Exposure
Inhalation
Oral
(gavage)
Inhalation
Range of Doses or
Concentrations 1
37 to 3,600 ppm
100 to 3,200
mg/kg-bw/day
100 to 1,000 ppm
Duration
Continuous and
intermittent
variable time
periods for
30-120 days
6 weeks
6 hr/day, 5
days/week for
4 weeks
POD Type 2
BMDL10- 21 6
ppm
BMDL10= 82
mg/kg-bw/day
BMDL10- 25
ppm
Effect
Increased
liver/body weight
ratio
HEC50
(ppm)
25
32
53
HEC95
(ppm)
12
15
24
HEC99
(ppm)
9.1
11
19
Uncertainty
Factors (UFs) 3
UFS=1; UFA= 3;
UFH=3; UFL=1;
Total UF- 10
UFS=1; UFA= 3;
UFH-3; UFL-1;
UFS=1; UFA= 3;
Total UF=10
Reference
Kjellstrand et al
(1983)
Buben and
O'Flahertv (1985)
Woolhiser et al.
(2006)
Notes:
i
Controls (or zero dose/concentration) are not presented to reflect the lowest and highest values tested in the studies.
POD type can be NOAEL, LOAEL, or BMDL. The IRIS program adjusted all values to continuous exposure.
UFs=subchronic to chronic UF; UFA=interspecies UF; UFH=intraspecies UF; UFL=LOAELto NOAEL UF.
Page 187 of 212
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L-2 Kidney Effects
Target
Organ/
System
Kidney
Species
Rat
(female)
Rat
(female)
Rat
(male)
Rat
(male)
(male)
Mouse
(female)
Route of
Exposure
Oral
(gavage)
Inhalation
Oral
(gavage)
mnaiation
Oral
(gavage)
Range of Doses or
Concentrations 1
500 to 1,000
mg/kg-bw/day
100 to 1,000 ppm
50 to 250
mg/kg-bw/day
ill to d,bUU ppm
869 to 1,739
mg/kg-bw/day
Duration
5 days/week for
104 weeks
6 hr/day, 5
days/week for 4
weeks
4-5 days/week
for 52 weeks
7 hrs/day, 5
years
Continuous and
exposures for
30-120 days
5 days/week,
TWA during
exposure period
(78 weeks),
animals
observed for 90
weeks
POD Type 2
BMDL05 = 9.45
mg/kg-bw/day
BMDL10- 15.7
ppm
BMDL10-34
mg/kg-bw/day
BMDLlcr 40 2
ppm
BMDLin ~ 34 7
ppm
LOAEL = 620
mg/kg-bw/day
Effect
Toxic nephropathy
Increased kidney
weight/body
weight ratio
Pathology changes
in renal tubule
in renal tubule
Increased kidney
weignt/Dody
weight ratio
Pathology changes
in renal tubule
(toxic nephrosis)
HEC50
(ppm)
0.042
0.099
0.19
.55
3.9
HEC95
(ppm)
0.0085
0.020
0.038
.15
0.77
HEC99
(ppm)
0.0056
0.013
0.025
.12
0.5
Uncertainty
Factors (UFs) 3
UFS=1; UFA= 3;
UFH=3; UFL=1;
Total UF=10
UFS=1; UFA= 3;
UFH-3; UFL-1;
Total UF=10
UFS=1; UFA= 3;
UFH=3; UFL=1;
Total UF=10
UFS=1; UFA= 3;
Total UF=10
UFS=1;UFA=3;
Total UF=10
UFS=1; UFA= 3;
UFH=3; UFL=30;
Total UF=300
Reference
NTP (1988)
Woolhiser et al.
(2006)
Maltoni and Cotti
(1986)
(1986)
Kjellstrand et al
(1983)
NCI (1976)
Notes:
i
Controls (or zero dose/concentration) are not presented to reflect the lowest and highest values tested in the studies.
POD type can be NOAEL, LOAEL, or BMDL. The IRIS program adjusted all values to continuous exposure.
UFs=subchronic to chronic UF; UFA=interspecies UF; UFH=intraspecies UF; UFL=LOAELto NOAEL UF.
Page 188 of 212
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L-3 Neurotoxicity
Target
Organ/
System
system
Species
Rat
(male)
Human
/U.-.4.U
sexes)
Rat
(male)
Rat
(male)
Rat
(female)
Mouse
(male)
Route of
Exposure
Oral
(drinking
water)
Oral
(gavage)
Inhalation
Inhalation
Range of Doses or
Concentrations 1
50 to 300 ppm
704 ppm x years of
exposure
(mean cumulative
exposure)
24 to 47
mg/kg-bw/day
1,000
mg/kg-bw/day
300 ppm
150 to 300 ppm
Duration
8 hrs/day, 5
weeks
Mean of 16
years
8 weeks
5 days/week for
6 weeks
24 hrs/day for
24 days
24 hrs/day for
24 days
POD Type 2
LOAEL - 12 ppm
LOAEL - 47
mg/kg-bw/day
LOAEL = 710
mg/kg-bw/day
LOAEL = 300 ppm
LOAEL = 150 ppm
Effect
Significant
wakefulness
Trigeminal nerve
effects (increased
latency in
masseter reflex)
Cognitive effects
(demyelination of
hippocampus)
Degeneration of
dopamine-
containing
neurons in
substantia nigra
Decreased
sciatic nerve
HEC50
(ppm)
13
18
126
274
378
HEC95
(ppm)
6 4
9.2
62
127
163
HEC99
(ppm)
4 8
7.1
47
93
120
Uncertainty
Factors (UFs) 3
UFS=3; UFA= 3;
UFH-3- UFL-10-
Total UF=300
UFS=1; UFA= 1;
Total UF=10
UFS=10; UFA= 3;
UFH=3; UFL=10;
Total UF=1000
UFS=10; UFA= 3;
UFH=3; UFL=10;
Total UF=1000
UFS=10; UFA= 3;
UFH=3; UFL=10;
Total UF=1000
UFS=10; UFA= 3;
UFH=3; UFL=10;
Total UF=1000
Reference
Arito et al (1994)
Ruijten et al.
(1991)
Isaacson et al.
(1990)
Gash etal. (2008)
Kiellstrand et al.
(1987)
Notes:
Controls (or zero dose/concentration) are not presented to reflect the lowest and highest values tested in the studies.
POD type can be NOAEL, LOAEL, or BMDL. The IRIS program adjusted all values to continuous exposure.
UFs=subchronic to chronic UF; UFA=interspecies UF; UFH=intraspecies UF; UFL=LOAEL to NOAEL UF.
Page 189 of 212
-------
L-4 Immunotoxicity
Target
Organ/
System
Immune
system
Species
Mouse
(female)
Mouse
(female)
Mouse
(female)
Rat
(female)
Mouse
(males;
auto-
immune
prone
strain)
Route of
Exposure
Oral
(drinking
water)
Oral
(drinking
water)
Oral
(drinking
water)
Inhalation
Inhalation
Range of Doses or
Concentrations 1
1.4 to 14 ppm
(0.35 to 3. 5
mg/kg-bw/day)
1.4 to 14 ppm
(0.35 to 3.5
mg/kg-bw/day)
18 to 660
mg/kg-bw/day
100 to 1,000 ppm
500 to 2,000 ppm
Duration
27-30 weeks
27-30 weeks
16 or 24 weeks
(4 or6 months)
6 hrs/day, 5
days/week for 4
weeks
4 hrs/day, 6
days/week for 8
weeks
POD Type 2
LOAEL = 0.35
mg/kg-bw/day
LOAEL = 0.35
mg/kg-bw/day
LOAEL - 18
mg/kg-bw/day
BMDL1SD- 24 9
ppm
Effect
Decrease in
thymus weight
and thymus
cellularity
Autoimmunity
(increased anti-
dsDNAandssDNA
antibodies)
Immuno-
suppression
Immuno-
suppression
Autoimmunity
(changes in
immunoreactive
organs)
HEC50
(ppm)
0.092
0.092
4.8
29
97
HEC95
(ppm)
0.045
0.045
2.4
15
48
HEC99
(ppm)
0.033
0.033
1.7
11
37
Uncertainty
Factors (UFs) 3
UFS=1; UFA= 3;
UFH=3; UFL=10;
Total UF=100 4
UFS=1; UFA= 3;
UFH=3; UFL=3;
Total UF=304
UFS=1; UFA= 3;
Total UF=100
UFS=10; UFA=3;
Total UF=100
UFS=10; UFA= 3;
Total UF=300
Reference
Keil et al. (2009)
Keil et al. (2009)
Sanders et al.
(1982)
Woolhiseret al.
(2006)
Kaneko et al.
(2000)
Notes:
1 Controls (or zero dose/concentration) are not presented to reflect the lowest and highest values tested in the studies.
2 POD type can be NOAEL, LOAEL, or BMDL. The IRIS program adjusted all values to continuous exposure.
3 UFs=subchronic to chronic UF; UFA=interspecies UF; UFH=intraspecies UF; UFL=LOAEL to NOAEL UF
Two different effects were reported by Keil et al, (2009): decreased thymic weight and cellularity and autoimmunity. A total UF of 100 was used for the thymus toxicity,
whereas a total UF of 30 was used for the autoimmune effects. The TCE IRIS assessment allocated different LOAEL-to-NOAEL uncertainty factors (UFL) based on the severity
of the effects, which resulted in different total UF (EPA. 2011e).
Page 190 of 212
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L-5 Reproductive Toxicity
Target
Organ/
System
Reproductive
system
Species
Human
(male)
Rat
(male)
Rat
(male)
Rat
(female
damns)
(male)
Mouse
(male)
Route of
Exposure
Inhalation
Oral
(drinking
water)
Inhalation
Oral
(gavage)
Inhalation
Inhalation
Range of Doses or
Concentrations 1
29.6 ppm
(mean exposure)
143 to 270
mg/kg-bw/day
376 ppm
10.1 to 1,125
mg/kg-bw/day
1,000 ppm
1,000 ppm
Duration
Measured values
after an 8-hr
work shift; mean
5.1 years on the
job
14 days
4 hrs/day,
5 days/week,
2-10 weeks
exposed,
2-8 weeks
unexposed
4 hrs/day,
5 days/week for
12 or 24 weeks
9 days (during
gestational days
6 to 15)
6 hrs/day, 5
days/week for
19 days over 4
weeks
6 hrs/day, 5
days/week for
1-4 weeks
6 hrs/day, 5
days/week for
6 weeks
POD Type 2
BMDL10 = 1.4
ppm
LOAEL - 141
mg/kg-bw/day
LOAEL = 45 ppm
LOAEL - 475
mg/kg-bw/day
LOAEL = 180 ppm
LOAEL - 180 ppm
Effect
Decreased normal
sperm
morphology and
hyperzoospermia
Decreased in vitro
fertilization
Sperm effects and
male reproductive
tract effects
Delayed
parturition
Effects on
epididymis
epithelium
Sperm effects
(decreased in vitro
sperm-oocyte
binding and in vivo
fertilization)
HEC50
(ppm)
1.4
16
32
98
190
190
HEC95
(ppm)
0.7
11
16
48
91
91
HEC99
(ppm)
0.5
9.3
13
37
67
67
Uncertainty
Factors (UFs) 3
UFS=10; UFA= 1;
UFH=3; UFL=1;
Total UF=30
UFS=10; UFA= 3;
UFH=3; UFL=10;
Total UF=1000
UFS-10- UFA- 3-
UFH=3; UFL=10;
Total UF=1000
UFS=1; UFA= 3;
UFH=3; UFL=10;
Total UF=100
UFS=10; UFA= 3;
UFH=3; UFL=10;
Total UF=1000
UFS=10; UFA= 3;
UFH=3; UFL=10;
Total UF=1000
Reference
Chiaetal. (1996)
DuTeaux et al.
(2004)
Kumar et al., 2000
Kumar etal. (2000)
Kumar etal. (2001)
Narotskyet al.
(1995)
Forkert et al.
(2002)
Kan etal. (2007)
Xu et al. (2004)
Page 191 of 212
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Target
Organ/
System
Reproductive
system
Species
Rat
(male)
Route of
Exposure
Oral
(drinking
water)
Range of Doses or
Concentrations 1
72 to 389
mg/kg-bw/day
Duration
Breeders
exposed 1 week
prematingand
then for 13
weeks/ Pregnant
females exposed
throughout
gestation (i.e.,
18 weeks total)
POD Type 2
LOAEL = 389
mg/kg-bw/day
Effect
Decreased mating
(both sexes
exposed)
HEC50
(ppm)
204
HEC95
(ppm)
97
HEC99
(ppm)
71
Uncertainty
Factors (UFs) 3
UFS-1- UFA- 3'
UFH=3; UFL=10;
UFD=1
Total UF=100
Reference
NTP(1986)
Notes:
Controls (or zero dose/concentration) are not presented to reflect the lowest and highest values tested in the studies.
POD type can be NOAEL, LOAEL, or BMDL. The IRIS program adjusted all values to continuous exposure.
UFs=subchronic to chronic UF; UFA=interspecies UF; UFH=intraspecies UF; UFL=LOAEL to NOAEL UF.
Page 192 of 212
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L-6 Developmental Toxicity
Target Organ/
System
Developmental
effects
Species
Route of
Exposure
Range of Doses or
Concentrations 1
Duration
POD Type2
Effect
HEC50
(ppm)
HEC95
(ppm)
HEC99
(ppm)
Uncertainty
Factors (UFs) 3
Reference
PRE- AND POSTNATAL MORTALITY
Rat
(female)
Rat
(female)
Inhalation
Oral
(gavage)
1. ppm
10.1 to 1,125
mg/kg-bw/day
4 hrs/day,
gestational days
8 to 21
Gestational days
6 to 15
LOAEL-17ppm
BMDL01- 32.2
mg/kg-bw/day
Increased
resorptions
Increased
resorptions
16
57
8
29
6.2
23
UFS=1; UFA= 3;
UFH=3; UFL=10;
Total UF=100
UFS=1; UFA= 3;
UFH-3; UFL-1;
Total UF=10
Healvetal.(1982)
Narotsky et al.
(1995)
PRE- AND POSTNATAL GROWTH
Rat
(female)
Inhalation
100 ppm
4 hrs/day,
gestational days
8 to 21
LOAEL - 17 ppm
Decreased fetal
weight and
skeletal effects
16
8
6.2
UFS=1; UFA= 3;
UFH=3; UFL=10;
Total UF=100
Healyetal.(1982)
CONGENITAL DEFECTS
Rat
(female)
Oral
(drinking
water)
2.5 to 1,100 ppm
(2.5 to 1,100
mg/kg-bw/day)
22 days
throughout
gestation
(gestational
days 0 to 22)
BMDL01 - 0 0207
mg/kg-bw/day
Heart
malformations
0.012
0.0051
0.0037
UFS=1; UFA= 3;
UFH=3; UFL=1;
Total UF=10
Johnson et al.
(2003)
DEVELOPMENTAL NEUROTOXICITY
Rat
(male
pups)
Rat
(female)
Oral
(gavage)
Oral
(drinking
water)
50 to 290
mg/kg-bw/day
45 to 140 mg/kg-
bw/day
Postnatal days
10 to 16
Dams and pups
exposed from 14
days prior to
mating until end
of lactation
LOAEL - 50
mg/kg-bw/day
LOAEL = 45
mg/kg-bw/day
Decreased rearing
activity
Increased
exploratory
behavior in male
pups (offspring)
8
22
4
11
3
8.4
UFS=3; UFA= 3;
UFH-3; UFL-10;
Total UF=300
UFS=1; UFA= 3;
UFH=3; UFL=10;
Total UF=100
Fredriksson et al.
(1993)
Taylor etal. (1985)
Notes: Controls (or zero dose/concentration) are not presented to reflect the lowest and highest values tested in the studies.
2 POD type can be NOAEL, LOAEL, or BMDL. The IRIS program adjusted all values to continuous exposure.
3 UFs=subchronic to chronic UF; UFA=interspecies UF; UFH=intraspecies UF; UFL=LOAEL to NOAEL UF.
Page 193 of 212
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Appendix M
BENCHMARK DOSE ANALYSIS OF BLOSSOM ET
AL. 2013
Male MRL +/+ mice were maternally exposed to vehicle control or TCE from birth through
weaning via drinking water (0, 2, or 28 mg/kg/day). At postnatal day (PND) 21, the offspring
were weaned and the male mice were exposed directly to vehicle control or TCE via drinking
water until PND 42. The study reported alterations in brain neurotrophin expression,
glutathione redox homeostasis, DNA hypomethylation and a number of behavioral parameters,
such as increased motor activity, and novelty/exploratory behavior at the highest dose tested
(28 mg/kg/day). The NOAEL for neurobehavioral impairments was 2 mg/kg/day (Blossom et al.,
2013).
EPA conducted benchmark dose (BMD) analysis using the exploratory behavior data reported in
Figure 7A and7B (Blossom et al., 2013). The data are described in Table M-l. The response
(endpoint) is the number of times a test animal entered a central zone in 10 minutes with a
novel object (Figure 7A) or a novel mouse (Figure 7B) at the center of a 1 m2 arena.
Table M-l. Neurotoxicological Endpoints Selected for Dose-Response Modeling
(Blossom et al., 2013)
TCE in drinking water
mg/ml
0
0.01
0.1
mg/kg-day
0
2
28
Sample
size1
9
8
8
Figure 7 A,
novel object (inverted
wire cup)
Mean
5.27
4.36
10.54
SD
5.78
5.16
4.30
Figure 7B,
novel mouse, placed under
inverted wire cup
Mean
5.32
7.85
10.19
SD
3.96
6.88
7.46
Notes:
Number of pups tested (one per dam with different pups tested for each endpoint)
The BMD analysis is summarized in Table M-2 and additional information is provided in the
supplementary files: l_Blossom_NObject_SD.xlsm and 2_Blossom_NMouse_SD.xlsm). The
BMDLiso ranged from 14-20 mg/kg/day depending upon the neurobehavioral endpoint
(Endpoint A or Endpoint B).
Page 194 of 212
-------
Table M-2. Summary of BMD modeling Results for Number of Entries into a Central Zone
With a Novel Object (Endpoint A) or Novel Mouse (Endpoint B) (Blossom et al.,
2013)
Model
Goodness of fit
p-value
AIC
BMD1SDa
(mg/kg-d)
BMDL1SDa
(mg/kg-d)
Basis for Model Selection
Endpoint A
(novel object)
Linear
Only the linear model provided an
adequate fit.
0.577
110.05
23.7
14.3
Endpoint B
(novel mouse)
Linearc
0.428
119.67
41.9
19.8
Only the linear model provided an
adequate fit.
Notes:
a BMR is 1SD change from the control mean
Constant variance models were used (BMDS Test 2, p-value = 0.684), with the selected model in bold. Scaled
residuals for selected model for doses 0, 2, and 28 mg/kg-d were 0.366, -0.418, and 0.0298, respectively.
c Constant variance models were used (BMDS Test 2, p-value = 0.176), with the selected model in bold. Scaled
residuals for selected model for doses 0, 2, and 28 mg/kg-d were -0.519, 0.592, and -0.0423, respectively. The
power model gave an identical fit to these data. Note that the BMD lies above the highest dose.
Figure M-l. Linear Dose-Response Model for Endpoint A: Number of Times Entered Central
Zone With Novel Object
Linear Model, with BMR of 1 Std. Dev. for the BMD and 0.95 Lower Confidence Limit for the BMDL
16
14
12
10
40
09:1402/21 2014
Page 195 of 212
-------
Figure M-2. Linear Dose-Response Model for Endpoint B: Number of Times Entered Central
Zone with Novel Mouse
Linear Model, with BMR of 1 Std. Dev. for the BMD and 0.95 Lower Confidence Limit for the BMDL
I
o
8-
-------
Appendix N Weight-of-Evidence Analysis for Fetal Cardiac
Malformations Following TCE Exposure
Appendix N contains a weight-of-evidence analysis for the association between short-term
exposure to TCE and fetal cardiac defects.
The analysis only addresses the fetal cardiac defects observed following gestational exposures
to TCE and/or its oxidative metabolites dichloroacetic acid (DCA) and trichloroacetic acid (TCA),
and includes updated information that was not part of the 2011 TCE IRIS assessment.
This update includes 1) identification of any new literature, 2) a systematic evaluation of
available data, 3) an evaluation of the weight of evidence for the association of TCE exposures
with cardiac defects, and 4) a transparent presentation of the evaluation.
A systematic literature search was conducted to identify all studies published subsequent to the
final literature search that had been conducted by EPA during completion of the 2011 TCE IRIS
assessment (EPA, 2011e). A total of 1686 unique citations were initially identified from
PubMed, Toxline, and Web of Science (WoS). These citations were screened using the title,
abstract, and/or full text for pertinence to evaluation of the developmental toxicity of TCE, TCA,
and DCA exposure. The literature search identified no new animal toxicology studies of fetal
cardiac defects, one new epidemiology study that assessed the association of TCE or
chlorinated solvent exposures with cardiac defects, and two studies that provided mechanistic
information relevant to alterations of cardiac development following TCE (or metabolite)
exposures.
The analysis does not provide an update on other developmental effects of TCE exposure, i.e.,
ocular malformations, developmental neurotoxicity, and developmental immunotoxicity.
Page 197 of 212
-------
Key Factor
a,b
Type of
evidence to
consider
Data
Evidence for stronger
weight of association
Evidence for weaker
weight of association
Comments
or Null Evidence
Temporality
Timing of
exposures and
response
Tox
Studies in various species in which TCE
(or metabolites DCA or TCA) were
administered during a sensitive period of
in utero cardiac development resulted in
morphological and/or functional
alterations.
• Drinking water administration of TCE
to rats on GD 1-22 resulted in a
statistically significant treatment-
related increase in the incidence of
cardiac defects (Dawson etal., 1993;
Johnson etal.,2003).
• Drinking water administration of TCA
(the TCE oxidative metabolite) to rats
on GD 1-22 resulted in a statistically
significant treatment-related increase
in the incidence of cardiac defects
(Johnson etal., 1998a). Gavage
administration of TCE metabolites
(DCA and TCA) on GD 6-15 (Smith et
al., 1989,1992) or of DCA during
discrete windows of time within GD 6-
15 (Epstein etal., 1992) resulted in
treatment-related increases in the
incidences of cardiac defects.
• Avian in ovo studies that administered
TCE or TCA during the period of
valvuloseptal morphogenesis (e.g., HH
15-20) resulted in altered cardiac
morphology and/or function (Drake, V.
et al., 2006; Drake, V. J. et al., 2006;
Loeber et al., 1988; Rufer et al., 2010).
• A study of DCA exposure to zebra fish
(Hassoun etal., 2005) demonstrated
Some in vivo or in vitro studies rodent
studies in which TCE (or metabolites DCA
or TCA) was administered during a
sensitive period of in utero cardiac
development resulted in no
morphological alterations.
• Gavage administration of TCE or
metabolites (DCA and TCA) to rats on
GD 6-15 did not result in treatment-
related cardiac defects (Fisher et al.,
2001).
• Inhalation exposures of TCE to rats on
GD 6-20 (Carney etal., 2006) or to rats
and mice on GD 6-15 (Schwetz et al.,
1975) did not result in treatment-
related cardiac defects.
Page 198 of 212
-------
Key Factor a'b
Strength of
association
Type of
evidence to
consider
Exposure
occurs before
outcomes
onset
Study quality,
including study
strengths and
limitations
Data
Epi
Tox
Evidence for stronger
weight of association
evidence of a disruption in cardiac
development (pericardial edema and
altered heart rate).
• Mouse whole embryo culture studies
of DCA and TCA administered at the
period of 3-6 somites detected cardiac
defects (Hunter etal., 1996); a chicken
whole embryo culture study of TCE
administered at HH 13-14 detected
alterations in AV cushion (Mishima et
al., 2006).
• Avian atrioventricular canal cell
culture (HH 16) study found evidence
of inhibited endothelial cell separation
and early events of mesenchymal cell
formation in the heart following TCE
exposures (Boyer et al., 2000).
• Four cohort or case-control studies
consider temporality (Forand et al..
2012; Goldberg et al., 1990; Ruckart et
al., 2013; Yauck etal., 2004). Three
studies observe an association
between the TCE exposure surrogate
and major cardiac defects (Forand et
al., 2012; Goldberg et al., 1990; Yauck
etal., 2004). An association with
conotruncal defects, specifically,
observed in Forand et al. (2012).
• For Dawson et al. (1993); Johnson et
al. (1998a); and Johnson et al. (2003),
all of which detected cardiac
malformations, study quality strengths
include randomized assignment to
test group, detailed description of
Evidence for weaker
weight of association
• Temporality was not considered in
Bove (1996); Bove et al. (1995);
Goldberg et al. (1990); and Lagakos et
al. (1986)
• For Johnson et al. (2003) major study
quality limitations include the use of
data pooled from separate study
cohorts conducted over an
approximately 6-year period, the use
of tap water as the vehicle for some of
Comments
or Null Evidence
• The small numbers of
conotruncal heart defects
in Ruckart etal. (2013)
precluded any analysis of
this endpoint and TCE
exposure.
• Some studies that reported
no cardiac defects
following TCE gestational
exposures (Hardin et al..
1981; Healy et al., 1982;
Narotsky etal. ,1995;
Page 199 of 212
-------
Key Factor
a,b
Type of
evidence to
consider
Data
Evidence for stronger
weight of association
Evidence for weaker
weight of association
Comments
or Null Evidence
fetal cardiac dissection and evaluation
procedures, evaluation of fetal hearts
without knowledge of treatment
group, and confirmation of all cardiac
defects by consensus of 3 experts.
Statistical analysis of data from this
study was appropriately conducted by
EPA statisticians using individual fetal
and litter data that were provided by
the study author.
• The power of detection in the Johnson
et al. (2003) study was enhanced by
the use of historical controls that did
not demonstrate a temporal shift in
cardiac defects. A significant dose
related trend in cardiac defects was
observed even without large group
sizes.
• A strong association of exposure to
response was observed at high dose
levels in multiple studies that
identified cardiac defects. In Johnson
etal. (2003) there was a highly
significant positive trend for cardiac
defects.
• Potential confounding factors exist in
studies that did not identify cardiac
defects (e.g., different routes of
exposure, the use of different rodent
strains or suppliers across studies, and
the use of soybean oil as a vehicle in
(Fisheretal., 2001).
control and treated groups (as
reported by Dawson et al. (1993) with
no characterization of possible
contaminants and incomplete
reporting of study methods and
results.
While Dawson etal. (1993) indicated
that levels of TCE in dose formulations
were tested by gas chromatography,
the analytical findings were not
reported. Johnson etal. (2003) did
not report whether dose formulations
were analyzed. Further, levels of TCE
were not assessed in the vehicle
control water; therefore, it is plausible
that TCE contaminated the water and
that doses were actually higher than
measured.
The Dawson etal. (1993) and Johnson
et al. (2003) studies estimated doses
based on the average water
consumption. This method does not
provide precise information to
calculate TCE dose because variability
in drinking water consumption among
dams is not characterized.
The dose selection for Johnson etal.
(2003) resulted in a NOAEL that is
approximately 700-fold lower than the
next highest dose.
Some studies that did not identify
treatment-related cardiac defects
following developmental exposures to
TCE (e.g.,Carney et al., 2006; Fisher et
Narotsky and Kavlock,
1995) or avian in ovo
studies (Brosset al., 1983;
Elovaara et al., 1979) did
not indicate that detailed
evaluation of fetal hearts
was conducted.
A rat whole embryo
culture study of TCE
administered at the period
of 4-7 somites detected no
cardiac defects in a study
bySaillenfaitetal. (1995);
however, the study
methods indicate that
there was no evaluation of
the embryonic heart.
Page 200 of 212
-------
Key Factor
a,b
Type of
evidence to
consider
Data
Evidence for stronger
weight of association
Evidence for weaker
weight of association
Comments
or Null Evidence
al., 2001; Schwetz et al., 1975) were
well-conducted and adequately-
reported GLP and/or guideline studies
with no substantive limitations
identified.
One study (Fisher et al., 2001)
attempted to replicate the methods
used in the Johnson et al. (2003)
study, utilizing the same fetal cardiac
dissection and evaluation techniques,
and including one of Johnson et al.
(2003) study authors in the
assessment team, yet found no
treatment-related cardiac defects.
Magnitude of
the effect
measure
Epi
• Increased risk estimates between all or
major cardiac defects ranged from
1.24 (95% Cl: 0.75,1.94) to 2.40 (95%
Cl: 1.27, 3.62) observed in 3 studies
(Bove, 1996; Bove et al., 1995; Forand
et al., 2012; Goldberg et al., 1990).
Stronger associations, observed with
the TCE exposure surrogate for
conotruncal defects and ventricular
septal defects than for major cardiac
defects, a broader category (Bove,
1996; Bove et al., 1995; Forand et al.,
2012). A fourth study observed an
increased risk estimate of 6.2 (95% Cl:
2.6,14.5) for cardiac defects in infants
of mothers aged >38 years and
maternal residence within 1.32 miles
from at least one TCE emissions source
(Yauck et al., 2004).
No association in Yauck et al. (2004) in
mothers <38 years of age and
maternal residence within 1.32 miles
from at least one TCE emissions
source nor in Lagakos et al. (1986),
which does not observe an association
with cardiac defects. Alternative
reasons such as lower statistical
power may explain these
observations.
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Key Factor
a,b
Type of
evidence to
consider
Data
Evidence for stronger
weight of association
Evidence for weaker
weight of association
Comments
or Null Evidence
Variability
analysis
Sources of
within- and
cross-study
variability that
contribute to
uncertainty
Tox
• Johnson et al. (2003) test subject
source, husbandry, and randomization
procedures were consistent across all
cohorts, i.e., including Dawson et al.
(1993) and metabolite studies Johnson
et al. (2003). Fetal cardiac evaluation
methodology, which included
evaluation without knowledge of
treatment group and confirmation of
all cardiac anomalies by 3 expert
scientists, was also consistently
applied across cohorts and studies
from the UAZ laboratory. This had the
result of reducing intra- and inter-
study variability in the assessment.
• Johnson et al. (2003) reported that
cardiac defect incidences were
consistent across all control cohorts
(55 litters over approximately 6 years).
An EPA review of the available control
data did not observe unusual
heterogeneity in prevalence of
malformations.
• Studies that reported cardiac defects
following administration of
metabolites (DCA and TCA) used
randomized assignment of maternal
animals to test group, thus reducing
intra-study variability.
• Although Dawson et al. (1993) and
Johnson et al. (2003) identified cardiac
defects following exposures to TCE
during development, (Carney et al.,
2006; Fisher et al., 2001; Schwetz et
• The Johnson et al. (2003) study
reported data from several cohorts of
animals, which were on study over a
period of approximately 6 years. The
data included control cohorts, some of
which were concurrent and some that
were non-concurrent to the TCE-
treated groups (Johnson et al., 2005;
Johnson, 2014). Data that definitively
link the individual control litter
response data with each particular
cohort are no longer available for
independent examination.
• Different study outcomes were
observed in studies that had many
similarities in study design and
conduct, i.e., Dawson et al. (1993) and
Johnson et al. (2003) identified
exposure related cardiac defects while
Fisher etal. (2001) did not. In the
Fisher etal. (2001) study, care was
taken to ensure that the same cardiac
evaluation methods were used as in
the Dawson etal. (1993) and Johnson
etal. (2003) studies, including fetal
evaluation with knowledge of
treatment group, and one of the study
authors of Johnson et al. (2003)
participated in the fetal examination.
• The use of soy bean oil in the Fisher et
al. (2001) study vs. water vehicle and
control for Johnson etal. (2003) and
Dawson etal. (1993) studies.
• The Johnson et al. (2003) and Dawson
• Based upon the
toxicokinetic profile of TCE
(EPA, 2011e), it is
considered unlikely that
toxicokinetic factors
contributed significantly to
differences in response
across study protocols.
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Key Factor
a,b
Type of
evidence to
consider
Data
Evidence for stronger
weight of association
Evidence for weaker
weight of association
Comments
or Null Evidence
al., 1975) did not find treatment-
related cardiac abnormalities. This
may be the result of differences in the
study design and assessment
methods. This includes such aspects
as animal strain, age, source, exposure
route and vehicle, duration of
exposure, and cardiac evaluation
methods.
etal. (1993) studies did not calculate
variability in TCE dose by measuring
individual dam water consumption.
Sources of
within- and
cross-study
variability that
contribute to
uncertainty
Epi
NE (not considered in Hill analysis)
NE (not considered in Hill analysis)
• Studies examined different
populations, exposure
levels, gradients, and
media. Additionally,
different sets of strengths
and uncertainties in this
set of studies would
contribute to observed
cross-study variability.
Uncertainty
analysis
Missing
information or
data gaps,
within and
across studies
Tox
• For the studies conducted by the UAZ
laboratory that identified cardiac
defects following exposures to TCE,
DC A, or TCA (Dawson et al., 1993;
Johnson et al., 1998a; Johnson et al.,
2003), detailed descriptions of
evaluation methods for assessment of
cardiovascular effects were provided.
• Individual fetal and litter cardiac
findings data, as well as detailed
information on study conduct and
fetal evaluation methods, were
provided to the EPA for Dawson et al.
(1993) and Johnson etal. (2003).
• The publications for studies conducted
by the UAZ laboratory that identified
cardiac defects following exposures to
TCE, DCA, or TCA (Dawson et al., 1993;
Johnson et al., 1998a; Johnson et al.,
2003) did not report essential study
details, and generally did not include
summaries of maternal data or fetal
data for endpoints other than cardiac
defects.
• For well-conducted studies that did
not detect cardiac defects following
developmental exposures to TCE or
metabolites (Carney etal., 2006; Fisher
et al., 2001) adequate descriptions of
study methodology and summary data
Page 203 of 212
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Key Factor
a,b
Type of
evidence to
consider
Data
Evidence for stronger
weight of association
Evidence for weaker
weight of association
Comments
or Null Evidence
for maternal and fetal findings were
reported.
• Mechanistic data for alterations in
cardiac development are limited and
do not identify initiating events for the
putative AOP.
Missing
information or
data gaps,
within and
across studies
Epi
• NE (not considered in Hill analysis)
NE (not considered in Hill analysis)
Qualitative
dose-response
Association
between
exposure/dose
and degree of
effect
Tox
• Alterations in cardiac development
were observed in multiple studies at
high dose levels following TCE, DCA, or
TCA exposures (Dawson et al., 1993;
Johnson et al., 1998a; Johnson et al.,
2003; Smith etal., 1989,1992).
• The incidence of cardiovascular effects
increased as a function of dose in
Johnson etal. (2003).
• An association between exposure to
TCE (or DCA or TCA) and alterations in
cardiac development was reported in
various animal models, i.e., LE and SD
rats, CD-I mice, chicken embryos, and
zebra fish (Dawson etal., 1993; Drake,
V. et al., 2006; Drake, V. J. et al., 2006;
Hassoun et al., 2005; Johnson et al.,
2003; Smith et al., 1989,1992;
Williams etal., 2006).
• A BMDL for Johnson et al. (2003) was
derived by EPA statisticians from
individual cardiac defect data provided
to EPA. Litter contribution to the
• The dose response for cardiac defects
identified by Johnson etal. (2003)
could only be fit to a model with
elimination of the high dose data from
the analysis. The lowest dose tested
had a zero response for cardiac
defects, below the historical control
incidence. The doses tested were
spaced over several orders of
magnitude, with wide gaps.
• Carney et al. (2006) was the only other
study in the database that evaluated
developmental effects of TCE over
multiple dose levels. In that study, no
fetal toxicity and minimal maternal
toxicity was reported.
• TCE doses tested in
Dawson etal. (1993) and
Johnson etal. (2003)
(drinking water): 2.5 ppb,
250 ppb, 1.5 ppm, or 1100
ppm (0, 0.00045, 0.048,
0.218, or 129 mg/kg-day)
• TCE doses tested Fisher et
al. (2001) (gavage): 500
mg/kg-day
• TCE doses tested in Carney
et al. (2006) (inhalation):
50, 150, or 600 ppm (268.5,
805.5, or 3222 mg/m3)
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Key Factor a'b
Experimental
evidence
Type of
evidence to
consider
Exposure-
response
gradient:
Association
between
exposure/dose
and degree of
effect
Hypothesis
testing:
manipulation
of exposure
scenario with
resulting
alterations in
response
Data
Epi
Tox
Evidence for stronger
weight of association
outcome of interest was incorporated
in the analysis. A significant dose-
response trend was identified,
whether or not the high dose value
was included in the analysis.
• NE
• A study by Epstein et al. (1992)
administered the metabolite DCAto
rats on varied days of gestation and
identified critical windows of exposure
for eliciting cardiac developmental
defects.
• No statistically significant increases in
congenital heart defects were
observed in groups of rats that were
exposed to TCE prior to pregnancy
only Dawson et al. (1993).
• Drake, V. J. et al. (2006) demonstrated
that cardiac defects did not occur in
chick embryos exposed to TCE and
TCA during the period of cardiac
specification (approximately GD 6 in
rats) rather than the period of
valvuloseptal morphogenesis.
Evidence for weaker
weight of association
• Goldberg et al. (1990) and Lagakos et
al. (1986) examined exposure-
response; none observed.
• Studies in rodents that administered
TCE via drinking water detected an
increase in fetuses with cardiac defects
(Dawson et al., 1993; Johnson et al..
2003); studies that administered TCE
via other routes (gavage and
inhalation) were negative for this
response (Carney et al., 2006; Fisher et
al., 2001; Schwetz et al., 1975).
• In a whole embryo culture (WEC) study
of DCA and TCA (Hunter etal., 1996),
that identified cardiac defects, the acid
nature of DCA and TCA may have
impacted dysmorphogenesis.
Comments
or Null Evidence
• Studies that manipulated
the gestational exposure
period were not conducted
with TCE.
Page 205 of 212
-------
Key Factor
a,b
Type of
evidence to
consider
Data
Evidence for stronger
weight of association
Evidence for weaker
weight of association
Comments
or Null Evidence
Association not
observed once
exposure
ceases
Epi
• NE
Reproducibility
[Consistency]
Reproducibility:
Corroboration
across studies,
labs, routes of
exposure,
species, etc.
Tox
• Studies that administered TCE in
drinking water to rats on GD 1-22
were conducted over a period of
approximately 6 years by researchers
at the same academic facility (UAZ,
Tucson) used the same cardiac
evaluation methods and identified
treatment and dose-related cardiac
malformations (Dawson et al., 1993;
Johnson et al., 1998a; Johnson et al.,
2003). A preliminary screening study
that utilized intrauterine
administration of TCE also detected
cardiac defects (Dawson et al., 1990).
The types of cardiac malformations
observed were similar across study
cohorts and treatment groups
throughout the duration of the
research program.
• Studies on TCE metabolites (TCA and
TCA) conducted in other laboratories
(Epstein et al., 1992; Smith et al.,
1989,1992) identified cardiac defects
similar to those observed in the UAZ
studies.
• Cardiac septal anomalies were
observed in avian in ovo studies
(Drake, V. et al., 2006; Rufer et al.,
2010), and in WEC assays (Hunter et
No differences between observed and
expected numbers of cardiac defect
cases once wells were closed in
contaminated area (Goldberg et al.,
1990).
• Studies conducted in other
laboratories than UAZ and that
administered TCE by gavage or
inhalation (Carney et al., 2006; Fisher
etal., 2001; Schwetz et al., 1975) did
not identify statistically significant
increases in cardiac defects. Fisher et
al. (2001) used the same cardiac
evaluation methods as the UAZ lab.
Studies that did not
identify cardiac defects
with TCE and/or metabolite
exposures (Carney etal.,
2006; Fisher et al., 2001;
Schwetz et al., 1975) did
not replicate all aspects of
the Johnson etal. (2003)
study, even though Fisher
et al. (2001) used the same
cardiac evaluation
techniques as Johnson et
al. (2003) andDawson et al.
(1993), and therefore
provide only limited
evidence of lack of
reproducibility.
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-------
Key Factor
a,b
Type of
evidence to
consider
Data
Evidence for stronger
weight of association
Evidence for weaker
weight of association
Comments
or Null Evidence
al., 1996; Mishima et al., 2006) with
TCE and/or metabolite exposures.
Zebrafish studies also demonstrated
evidence of alterations in cardiac
development (Hassoun et al., 2005;
Williams etal., 2006).
Consistency:
Association
observed in
different
populations,
places, time
and
circumstances.
Epi
Association between cardiac defects
and TCE exposure surrogate observed
in four studies. These studies were of
different populations living in
different state (NY, NJ) and covered
slightly different time period (1983-
2000,1985-1988) (Bove, 1996; Bove
et al., 1995; Forand et al., 2012). Two
other studies of weaker designs were
of different populations and carried
out in two different locations in the
United States, and provide supporting
evidence (Goldberg etal., 1990; Yauck
etal., 2004).
Lagakos etal. (1986) compared a
pregnancy receiving contaminated
residential well water to a pregnancy
not receiving residential water from
contaminated wells and does not
observed an association between
cardiac defects and contaminated
drinking water.
Biological
plausibility
Observed
outcome can
be attributed
to toxic insult
given the
known science
Tox
Avian in ovo studies and
atrioventricular cell culture studies
support the biological plausibility of
effects of TCE on cardiac
development, given that early chick
heart development is similar to
mammalian (including human),
particularly regarding the role of the
cardiac cushion in septation (NRC,
2006; Richards and Garg, 2010).
Preliminary exploration of a possible
adverse outcome pathway (AOP) has
resulted in a reasonable conceptual
• A definitive AOP for TCE-induced
cardiac defects, including a putative
initiating event, has not yet been
characterized. Additional mechanistic
data are needed to support the
hypothesized AOP.
• There are insufficient mechanistic data
to characterize additional potential
MOAs other than that hypothesized in
the AOP.
• It is possible that multiple
modes of action are
involved in alterations to
cardiac development.
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Key Factor a'b
Alternative or
multiple
explanations
Type of
evidence to
consider
Observed
association
plausible given
the known
science
Other possible
explanations
for observed
outcome after
the exposure of
interest
Data
Epi
Tox
Evidence for stronger
weight of association
model forTCE-induced congenital
heart defects. In this construct, the
vulnerable period is defined by
endocardial morphogenesis.
Endothelial-mesenchyme transition is
disrupted in the area of the
atrioventricular canal, leading to
septal defects. Possible genetic
contributions to abnormal cardiac
development include disruption of
TGF-beta pathway, endrin pathway,
Notch pathway, VEGF pathway, and
RXR signaling. At a cellular level,
epithelial-mesenchymal transition
may be affected in the endocardium,
at the tissue level, there is altered
cellularity of the endocardial cushion,
and secondary effects such as
dysregulation of cellular Ca2+ fluxes
may result in additional impacts on
the developing heart.
• NE
• Given the presumed contribution of
both environmental exposures and
genetic predisposition in human
congenital heart disease (Richards and
Garg, 2010), it is possible that the test
subjects used in the Johnson et al.
(2003) study and others conducted in
that laboratory may have been
Evidence for weaker
weight of association
• NE
• There is a possibility that cardiac
defects detected in the Dawson et al.
(1993) study were associated in part
with the use of tap water as a control
vehicle (i.e., possible presence of
contaminants).
Comments
or Null Evidence
• In vitro and in vivo animal
studies report cardiac
defects with TCE and TCE-
metabolite exposure.
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Key Factor a'b
Type of
evidence to
consider
Other possible
explanations
for observed
outcome after
the exposure of
interest (not
considered in
Data
Epi
Evidence for stronger
weight of association
particularly susceptible to alterations
in cardiac development.
• Other contributing factors or
confounding factors were not
specifically identified in the evaluated
in-vivo studies.
• It is possible that the absence of
treatment-related cardiac defects in
well-conducted TCE studies (Carney et
al., 2006; Fisher et al., 2001) or
metabolite studies (Fisher etal., 2001)
was due to confounding variables such
as differences in strain/source of
animal model, route of exposure,
toxicokinetics, vehicle [e.g., soybean
oil in Fisher etal. (2001)1, or
differences in cardiac evaluation
methods.
• It is unlikely that the cardiac defects
observed by Johnson et al. (2003)
were an artifact of the evaluation
procedures used, since a study by
Fisher et al. (2001), using the same
fetal cardiac evaluation procedures,
did not identify an association
between TCE exposure and the
incidence of cardiac defects.
• Potential maternal risk factors were
adjusted in statistical analysis in
Forand et al. (2012) and Yauck et al.
(2004) or were not found in statistical
analyses to influence observed
association by +15% (Bove, 1996;
Bove etal., 1995).
Evidence for weaker
weight of association
• Potential for confounding from
another exposure given the poor
exposure definition in Yauck et al.
(2004). The positive association in
Goldberg et al. (1990) may result from
likely selection biases in controls.
Comments
or Null Evidence
Page 209 of 212
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Key Factor
a,b
Type of
evidence to
consider
Data
Evidence for stronger
weight of association
Evidence for weaker
weight of association
Comments
or Null Evidence
Hill analysis)
Specificity
Single cause
and effect
relationship
resulting from
exposure to
test substance
Tox
• Cardiac defects in rats appear to be
attributable to direct chemical
exposure to TCE or metabolites (DCA
or TCA) and are unlikely to be the
result of secondary effect of maternal
toxicity. Johnson etal. (2003)
reported that TCE exposure via
drinking water to pregnant rats did not
result in maternal toxicity. Carney et
al. (2006) reported minimal decreases
in body weight gain in dams, with no
adverse fetal outcomes. In fetuses,
there was no indication of TCE-related
fetal weight deficits, external or
skeletal anomalies, or of soft tissue
alterations other than cardiac defects
in Johnson etal. (2003) nor in any
other study.
• The majority of the cardiac
malformations following TCE
exposures to rats (Dawson et al., 1993;
Johnson et al., 2003) or chicks (Drake,
V. et al., 2006; Rufer et al., 2010)
during sensitive periods of cardiac
development were ventricular septal
defects, valve defects, or outflow tract
abnormalities. Mechanistic data
suggest a common etiology (disruption
of the cardiac cushion formation) for
the observed cardiac defects Boyer et
al. (2000).
• Studies conducted in other
laboratories than UAZ and that
administered TCE by gavage or
inhalation (Carney et al., 2006; Fisher
etal., 2001; Schwetz et al., 1975) did
not identify cardiac defects. Fisher et
al. (2001) used the same cardiac
evaluation methods as the UAZ lab.
• The cardiac defects detected in the
Dawson etal. (1993) study may have
been related to the use of tap water as
a vehicle (i.e., possible contaminants).
Page 210 of 212
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Key Factor a'b
Coherence
Type of
evidence to
consider
Single cause
and effect
relationship
resulting from
exposure to
test substance
Summary:
Extent to which
data are similar
in outcome and
exposure
across
database
Cause and
effect
interpretation
should not
conflict with
the generally
known facts of
the natural
Data
Epi
Tox
Epi
Evidence for stronger
weight of association
• NE
• Multiple studies were conducted at
UAZ (Dawson etal., 1993; Johnson et
al., 1998a; Johnson et al., 2003), in
which rats were administered TCE or
metabolites DCA or TCA in drinking
water on GD 1-22 and for which study
design and cardiac evaluation
methodologies were consistent. The
outcomes of these studies (detection
of cardiac defects, particularly septal
defects, valve abnormalities, and
outflow tract anomalies) are
consistent across these studies.
Additionally, these outcomes are
supported by the results of avian in
ovo and in vitro studies, studies with
TCE metabolites (DCA and TCA) in
rodents, in vitro whole embryo culture
studies, and mechanistic data.
• Associations in epidemiologic studies
of cardiac defects and maternal
occupational exposure to degreasing
solvents or to organic solvents (Gilboa
et al., 2012; Loffredo et al., 1991;
Tikkanen and Heinonen, 1988, 1991).
Evidence for weaker
weight of association
• Specificity not a critical compared to
other Hill aspects since outcomes may
have several risk factors. Maternal
risk factors, specifically chemical risk
factors, associated with cardiac
defects in infants have not been well
studied.
• Developmental toxicity studies with
TCE that were conducted in other
laboratories (Carney et al., 2006; Fisher
et al., 2001; Schwetz et al., 1975)
administered TCE to rats of other
strains or sources, using different
routes of exposure (inhalation or
gavage), administered on different
days of gestation (i.e., not including GD
1-6) than the UAZ studies and did not
identify cardiac defects. No other
study in the TCE database reported
cardiac defects at the low dose levels
reported by Johnson et al. (2003).
• NE
Comments
or Null Evidence
Page 211 of 212
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Key Factor a'b
Type of
evidence to
consider
history and
biology of the
disease
Data
Evidence for stronger
weight of association
Evidence for weaker
weight of association
Comments
or Null Evidence
NE = No relevant evidence.
HH = Hamburger-Hamilton stages of chick development (Hamburger and Hamilton, 1951).
Tox = Animal toxicology studies; Epi = Epidemiological studies
Key Factor references:
a EPA(2006b)
b Hill (1965)
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