United States                              EPA Document# 740‐R1‐4002

Environmental Protection Agency        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

June 2014

   
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

DEGREASING 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
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‐1 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‐1 ENVIRONMENTAL FATE ..................................................................................................................................157

C‐1‐1 Fate in Air........................................................................................................................................157

C‐1‐2 Fate in Water..................................................................................................................................157

C‐1‐3 Fate in Soil, Sediment, and Groundwater .......................................................................................158

C‐1‐4 Bioconcentration.............................................................................................................................159

C‐1‐5 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
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

I‐1 DEFAULT PARAMETERS USED IN CEM FOR EMISSION AND HOUSEHOLD CHARACTERISTICS............................................171

I‐1‐1 Air Exchange Rate ...........................................................................................................................171

I‐1‐2 Overspray fraction ..........................................................................................................................173

I‐1‐3 Emission Rate..................................................................................................................................173

I‐1‐4 House Volume and movement within the home.............................................................................175

I‐1‐5 Inhalation rate and body weight ....................................................................................................176

I‐2 CONSUMER BEHAVIOR PATTERNS .....................................................................................................................176

I‐2‐1 Westat data for solvent type cleaning fluids..................................................................................177

I‐2‐2 Westat data for brake cleaners/quieters........................................................................................178

I‐2‐3 Summary of Westat data................................................................................................................178

I‐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‐1 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‐1 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
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 (μg/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
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 Assessment77

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‐1.  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‐1.  Environmental Fate Characteristics of TCE........................................................................160

Table G‐1.  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‐1.  Model Inputs for Dry Cleaning Facilities.....................................................................169

Table H‐2.  Potential Workplace TCE Inhalation Exposures and Number of Workers Exposed...170
Table I‐1.  Summary of CEM Parameters for Estimation of TCE Indoor Air Concentrations........172

Table I‐2.  Comparison of Westat Survey Data and EPA’s TCE Degreaser Simulation Values......177

Table J‐1.  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
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‐1.  Illustration of an Imperfectly Mixed Room:  Near‐Field/Far‐Field Approximation of a

Solvent Cleaning Facility ............................................................................................................. 164

Figure I‐1.  Screen capture of Summary of Recommended Values for Residential Building

Parameters from the Exposure Factors Handbook (EPA, 2011a)............................................... 173

Figure I‐2.  Screen capture of E‐FAST equations for estimation of emission rate...................... 174

Figure I‐3.  Screen Capture of MCCEM Model Evaluation Efforts.............................................. 179

Figure K‐1.  Study Quality Considerations for Epidemiological Studies ..................................... 185

Figure K‐2.  Study Quality Considerations for Animal Studies.................................................... 186
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)
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
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.  

Private consultant

Jeffrey Driver, Ph.D.

Risksciences.net, LLC

Montserrat Fuentes, Ph.D.

North Carolina State University

Kathleen M. Gilbert, PhD.

University of Arkansas for Medical Sciences

Michael Jayjock, Ph.D.

LifeLine Group

Ronald L. Melnick, Ph.D.

Ron Melnick Consulting, LLC.  

Kenneth M. Portier, Ph.D.

American Cancer Society

Barry Ryan, Ph.D.

Emory University

Calvin Willhite, Ph.D.

Risk Sciences International

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/
Page 12 of 212

GLOSSARY OF TERMS AND ABBREVIATIONS

Ɛ0    Vacuum permittivity

μg/m3

    Microgram(s) per cubic meter

AC    Acute concentration

ADC     Average daily concentration

ADR    Acute dose rate  

ADRpot    Potential acute dose rate

AEGL    Acute exposure guideline level

AER    Air exchange rate

AT    Averaging time

Atm    Atmosphere

ATSDR   Agency for Toxic Substances and Disease Registry

BAF    Bioaccumulation factor

BCF    Bioconcentration factor

BMD    Benchmark dose

BMDL    Benchmark dose, lower confidence limit(s)

BLS    Bureau of Labor Statistics

BOD    Biochemical oxygen demand

BW    Body weight

C    Contaminant concentration

Cair    Air concentration

°C    Degree Celsius

CFF    Average far field concentration

CFF,TWA    Time weighted average far field concentration

CNF    Average near field concentration

CNF,TWA   Time weighted average near field concentration

Cp pot     Modeled peak concentration

CASRN   Chemical abstracts service registry number

CBI    Confidential business information

CCD    Chemical Control Division

CCRIS     Chemical Carcinogenesis Research Information System

CDR    Chemical data report

CEM    Consumer exposure module

CEPA      Canadian Environmental Protection Act  

CERCLA Comprehensive Environmental Response, Compensation, and Liability Act

CFC‐12   Dichlorodifluoromethane, also called Freon‐12

CH    Chloral hydrate

CI    Confidence interval

cm    Centimeter(s)

cm3

    Cubic meter(s)

CNS    Central nervous system

CO2    Carbon dioxide
Page 13 of 212

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

FMO3    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 cancer value

HEC    Human equivalent concentration
Page 14 of 212

HEC50    Human equivalent concentration at the 50th percentile

HEC95    Human equivalent concentration at the 95th percentile

HEC99    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    Interleukin‐2

IEMB     Indoor Environmental Management Branch

IRIS    Integrated Risk Information System

IUR    Inhalation unit risk

k    Emission rate

Kow    Octanol:water 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 15 of 212

MOEacute Margin of exposure for acute exposures

MOEchronic Margin of exposure for chronic exposures

MOU    Memorandum of understanding

MSDS(s) Material safety data sheet(s)

MW Molecular weight

NAcDCVC N‐Acetyl‐S‐(1,2‐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
Page 16 of 212

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
Page 17 of 212

ݒேி    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 18 of 212

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 (lbs) 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.
Page 19 of 212

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.  
Page 20 of 212

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 (NIOSH)(NIOSH, 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
Page 21 of 212

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 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 μg/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
Page 22 of 212

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 HEC50 was 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 HEC99 was 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 HEC50 and 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.
Page 23 of 212

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

MOE of 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
Page 24 of 212

bystander exposure worst case scenarios, and the spot cleaning worker and bystander

worst case scenarios with no LEV.
Page 25 of 212

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 26 of 212

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 lbs/yr (Glauser and  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 ASSESSMENT

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.   

                                                      

3 This source material includes data or information derived from IHS Products provided to the U.S. EPA. IHS

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 IHS is not liable to you or any third party for your use of

and/or reliance on the IHS data and information contained in this document, and any such use shall be at your

own risk.
Page 27 of 212

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)

                                                      

4 For the purpose of this risk assessment, “we” refers to U.S. EPA/OPPT.
Page 28 of 212

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?

Spotting agent 10‐100

Workers and

bystanders at dry

cleaning facilities  

(all adults1

)

Yes – potential for high content,

potential for frequent use, high

potential for human exposure (i.e.,

chronic exposures)

Solvent degreaser >90

Consumer users

(adults >16 yrs old 1

)

and bystanders 2

(all ages)

Yes – high content, low frequency of

use, high potential for human

exposure (i.e., acute exposures)

Plastic clear

protective coating

spray (hobbyists;

arts/crafts)

20‐30

Yes – low content, but possibly largest

use of consumer products (i.e., acute

exposures)

Film cleaner

(hobbyists) >90

Consumer users

(adults >16 yrs old 1

)

and bystanders 2

  

(all ages)

No – high content, low frequency of

use (use of negatives/cameras with

film is assumed to be negligible); low

potential for human exposure

Toner aide

(home office) 15‐20 No – low content, low frequency ofuse, low potential for human exposure

Mirror edge sealant

(hobbyist/home

maintenance)

20‐30 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 29 of 212

1.3.3 Identification of Human Populations Exposed During TCE Uses

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.

1.3.4 Risk Evaluation of TCE’s Human Health Hazards

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”. 6

   NAS report, “Assessing the human health risks of trichloroethylene: Key scientific issues (2006)”:

http://www.nap.edu/catalog.php?record_id=11707   7

   EPA’s SAB peer review report for the 2009 EPA’s Draft Assessment entitled “Toxicological Review of

Trichloroethylene”:

http://yosemite.epa.gov/sab/sabproduct.nsf/c91996cd39a82f648525742400690127/B73D5D39A8F184BD8525

7817004A1988/$File/EPA‐SAB‐11‐002‐unsigned.pdf
Page 30 of 212

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.
Page 31 of 212

2 HUMAN HEALTH RISK ASSESSMENT

2.1 PHYSICAL‐CHEMICAL PROPERTIES

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

octanol:water 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 Value

CASRN 79‐01‐6

Molecular weight 131.39

Molecular formula C2HCl3

Physical state Colorless liquid

Odor Sweet, pleasant, resembles chloroform

Density 1.46 g/cm3 @ 20 °C a

Flash point 90 °C (closed cup) a

Auto flammability 410 °C b

Viscosity 0.53 mPa∙s @ 25 °C c

Refractive index 1.4775d

Dielectric constant 3.4 ε0 @ 16 °C c

Melting point ‐84.7 °C (measured)

Boiling point 87.2 °C (measured)

Vapor pressure 73.46 mm Hg at 25 °C (measured) e

Dissociation constant (pKa) Not applicable

Henry’s law constant 9.85×10‐3 atm‐m3/mole (measured)

Water solubility 1,280 mg/L at 25 °C (measured) f

Octanol/water partition

coefficient (Kow)

2.61(measured) g

Sources:

a ECB (2000) b SRC (2012) c

Weast and Selby (1966)

d

O'Neil et al. (2001) e

Daubert and Danner (1989)

f Horvath et al. (1999) g Hansch et al. (1995)
Page 32 of 212

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 μg/m3 (i.e., which is equivalent to 5.6 x 10‐5 parts per million or ppm).  
Page 33 of 212

Table 2‐2.  TCE Ambient Air Monitoring Data (μg/m3) a

Year Number of

Monitors

Number of

States Mean StandardDeviation Median Range

1999 162 20 0.30 0.53 0.16 0.01‐4.38

2000 187 28 0.34 0.75 0.16 0.01‐7.39

2001 204 31 0.25 0.92 0.13 0.01‐12.90

2002 259 41 0.37 1.26 0.13 0.01‐18.44

2003 248 41 0.35 0.64 0.16 0.02‐6.92

2004 256 37 0.32 0.75 0.13 0.00‐5.78

2005 313 38 0.43 1.05 0.14 0.00‐6.64

2006 258 37 0.23 0.55 0.13 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 DEGREASING

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 34 of 212

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
Page 35 of 212

1. 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.

2. 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 of 2X.

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
Page 36 of 212

of TCE emissions came from NPS (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

(lbs/yr)

Point source

(Large facility)

180

(degreasing units) 1,480,000

Nonpoint source

(Small facility) 1,779 a 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 2011 TRI 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  

(lbs/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 Total Annual Emissions

(lbs/yr)

Stack air emissions 850,000

Fugitive air emissions 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.

                                                      

8 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 37 of 212

Table 2‐6.  EPA/OPPT’s Estimated Total Annual Air Emissions of TCE from Small

Commercial Degreasing Facilities

Type of Emission Total Annual Emissions

(lbs/yr)

Air emission 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  )  =  2,230,800 lbs

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
Page 38 of 212

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 Number of Units Total Annual Air

Emissions (lbs/yr)a

Open top vapor degreasing

(batch vapor) 116 890,000

Cold solvent cleaning

(batch cold) 13 140,000

Conveyorized vapor degreasing

(in‐line) 11 120,000

General degreasing units

(unknown) 40 330,000

Source:  EPA (2008)  

Notes:

a  The total is equal to 1,480,000 lbs 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:
Page 39 of 212

TCE Emission Rate

ൌ ൬Annual	TCE	air	emissions

Number	of	small	facilities൰∗൬ 1Operating	days	per	yr൰

∗ ൬ 1

Operating	hrs	per	day൰ ∗ 					൬ 1	hr

60	min൰

ൌ ൬2,230,800	pounds	per	yr

1,746	small	facilities ൰∗൬ 454	g

pound൰∗൬ 1260	days	per	yr൰

∗ ൬ 1

2	hrs	per	day൰ ∗ 					൬ 1	hr

60	min൰

ൌ 19	 ୥	୭୤	୘େ୉

୫୧୬ per	facility  

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 et al., 1989).

Based on another source, operating emissions from batch cleaning machines can range

from 5 to 10 g of TCE 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 a

5 to 10 b

16 to 50 c

Sources:

a

  Wadden et al. (1989) b   EPA (2001a) c   EPA (2004b)
Page 40 of 212

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 of small facilities =  1,746

2. Number of 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

(Jayjock et al., 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%.
Page 41 of 212

Table 2‐9.  Summary of Potential Workplace TCE Inhalation Exposures at Small

Commercial Degreasing Facilities based on the NF/FF Model

Population

Exposed

Exposure Type

(Duration)

Estimated  

TCE Air

Concentration

(ppm)

With LEV4

Estimated  

TCE Air

Concentration (ppm)

No LEV

Estimated

Number

of People

Exposed

Workers1

Inhalation

exposure  

(8‐hr TWA3

)

0.3

(low‐end estimate)

20   

(upper‐end estimate)

3  

(low‐end estimate)  

197

(upper‐end estimate)

8,730

Occupational

bystanders2

Inhalation

exposure  

(8‐hr TWA)

0.04  

(low‐end estimate)

17  

(upper‐end estimate)

0.4  

(low‐end estimate)  

172

(upper‐end estimate)

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

                                                      

9 OSHA monitoring data and analysis supporting Figure 2‐2 can be found in the supplementary file “OSHA IMIS TCE

SAMPLES_062314v1.xlsx”
Page 42 of 212

chronic workplace exposures is provided in Table 2‐10. Acute workplace exposures were

estimated as follows:

AC ൌ

C ∗ ED

AT

where:

AC       =  acute concentration (24 hr TWA in ppm)

C       =  contaminant concentration in air (8 hr TWA in ppm; from Table 2‐9)

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 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 × 365 days/yr × 24 hrs/day; where LT = lifetime; LT = 40

yrs for non‐cancer risks; LT=70 yrs for cancer risks)
Page 43 of 212

Figure 2‐2.  Comparison of EPA/OPPT’s Workplace TCE Exposure Estimates with 2003‐2010

OSHA Monitoring Data for Small Commercial Degreasing Facilities

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_062314v1.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).
Page 44 of 212

Table 2‐10.  Summary of Acute and Chronic Workplace TCE Inhalation Exposures at

Small Commercial Degreasing Facilities

Population

Exposed

Workplace Exposure Concentrations

(24‐hr TWA3 in ppm)

Acute

Chronic

Non‐cancer Cancer

With LEV No LEV With LEV No LEV With LEV No LEV

Workers1

0.1  

(low‐end

estimate)

7  

(upper‐end

estimate)

1

(low‐end

estimate)

66

(upper‐end

estimate

0.07  

(low‐end

estimate)

5

(upper‐end

estimate

0.7  

(low‐end

estimate)

47

(upper‐end

estimate

0.04  

(low‐end

estimate)

3

(upper‐end

estimate

0.4  

(low‐end

estimate)

27

(upper‐end

estimate

Occupational

bystanders2

0.01  

(low‐end

estimate)

6

(upper‐end

estimate)

0.1

(low‐end

estimate)

  

57

(upper‐end

estimate)

0.01  

(low‐end

estimate)

4

(upper‐end

estimate)

0.1  

(low‐end

estimate)

41

(upper‐end

estimate)

0.005

(low‐end

estimate)

2

(upper‐end

estimate)

0.05  

(low‐end

estimate)

23

(upper‐end

estimate)

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 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*
Page 45 of 212

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; Luhring and  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.
Page 46 of 212

Figure 2‐3.  Overview of EPA/OPPT’s occupational release and exposure assessment

for spot cleaning operations

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

ൌ ሺ୅୬୬୳ୟ୪	୙ୱୟ୥ୣ	୭୤	ୗ୮୭୲୲୧୬୥	୅୥ୣ୬୲ୱሻ	∗	ሺ୘େ୉	େ୭୬ୡୣ୬୲୰ୟ୲୧୭୬	୧୬	ୗ୮୭୲୲୧୬୥	୅୥ୣ୬୲ሻ

୒୳୫ୠୣ୰	୭୤	୊ୟୡ୧୪୧୲୧ୣୱ ൌ

ሺସଶ,଴଴଴	୥ୟ୪୪୭୬ୱ	୮ୣ୰	୷୰ሻ∗ሺଵ଴%	୲୭	ଵ଴଴%ሻ

ହ,଴଴଴	୤ୟୡ୧୪୧୲୧ୣୱ ൌ 0.84	to	8.4	 ୥ୟ୪୪୭୬ୱ

୷୰ 	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 47 of 212

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 Emission Rate (no LEV)

ൌ ሺTCE	Use	Rateሻ ∗ ቀଷ,଻଼ହ	ୡ୫య

୥ୟ୪୪୭୬ ቁ ∗ ሺTCE	Densityሻ ∗ ቀ ଵ	୷୰

ଶ,଴଼଴	୦୰ୱቁ∗ቀ ଵ	୦୰଺଴	୫୧୬ቁ ൌ

ቀ

଴.଼ସ	୲୭	଼.ସ	୥ୟ୪୪୭୬ୱ

୷୰ି୤ୟୡ୧୪୧୲୷ ቁ∗ቀଷ,଻଼ହ	ୡ୫య

୥ୟ୪୪୭୬ ቁ∗ቀଵ.ସ଺	୥

ୡ୫య ቁ∗ቀ ଵ	୷୰ଶ,଴଼଴	୦୰ୱቁ∗ቀ ଵ	୦୰଺଴	୫୧୬ቁ ൌ

଴.଴ଷ଻	୲୭	଴.ଷ଻	୥	୘େ୉

୫୧୬ 	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

(g TCE/min)

With LEV 1

TCE Emissions Potentially Escaping

into the Workplace

(g TCE/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

ൌ ሺTCE	Use	Rateሻ ∗ ቀ3,785 ܿ݉3

݈݈݃ܽ݋݊ ቁ ∗ ሺTCE Densityሻ ∗ ቀ 1 ݎݕ

ݎ݄	1 ቀ∗ቁݏݎ݄ 2,080

60	݉݅݊ ቁ ∗

ሺ100% െ LEV	Effectivenessሻ ൌ ቀ0.84	݋ݐ	8.4	݈݈݃ܽ݋݊ݏ

ݎݕ െ݂݈ܽܿ݅݅ݕݐ ቁ∗ቀ3,785	ܿ݉3

݈݈݃ܽ݋݊ ቁ∗ቀ1.46	݃

ܿ݉3 ቁ ∗

ݎݕ	1 ቀݎ݄	1 ቀ∗ቁݏݎ݄	2,080

60	݉݅݊ ቁ ∗ ሺ10%ሻ ൌ 0.0037 ݋ݐ 0.037 ݃ ܶܧܥ݉݅݊ per facility  
Page 48 of 212

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

(Jayjock et al., 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
Page 49 of 212

Table 2‐12.  Summary of Potential Workplace TCE Inhalation Exposures from

Spot Cleaning at Dry Cleaning Facilities

Population

Exposed

Type of Exposure

(Exposure

Duration)

Estimated

TCE Air

Concentration

(ppm)

With LEV4

Estimated

TCE Air

Concentration

(ppm)

No LEV

Estimated

Number of

People

Exposed5

Worker1

Inhalation

exposure  

(8‐hr TWA3

)

0.008  

(low‐end

estimate)

2

(upper‐end

estimate)

0.08  

(low‐end

estimate)

19

(upper‐end

estimate)

36,000

Occupational  

bystanders2

Inhalation

exposure  

(8‐hr TWA)

0.0007  

(low‐end

estimate)

2

(upper‐end

estimate)

0.007

(low‐end

estimate)

  

18

(upper‐end

estimate)

252,000

Notes:

1

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

ppm).
Page 50 of 212

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:

AC ൌ

C ∗ ED

AT

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:

LADC ൌ

C ∗ ED ∗ EF ∗ WY

AT

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 × 365 days/yrs × 24 hrs/day; where LT = lifetime; LT = 40

yrs for non‐cancer risks; LT=70 yrs for cancer risks)
Page 51 of 212

Table 2‐13.  Summary of Acute and Chronic Workplace TCE Inhalation Exposures from

Spot Cleaning at Dry Cleaning Facilities

   Population

Exposed

Workplace Exposure Concentrations

(24‐hr TWA3 in ppm)

Acute

Chronic

Non‐cancer Cancer

With LEV 4 No LEV With LEV No LEV With LEV No LEV

Workers1

0.003  

(low‐end

estimate)

1

(upper‐end

estimate)

0.03  

(low‐end

estimate)

6  

(upper‐end

estimate)

0.002  

(low‐end

estimate)

0.5

(upper‐end

estimate)

0.02  

(low‐end

estimate)

5

(upper‐end

estimate)

0.001  

(low‐end

estimate)

0.3

(upper‐end

estimate)

0.01  

(low‐end

estimate)

3

(upper‐end

estimate)

Occupational

bystanders2

0.0002

(low‐end

estimate)  

1

(upper‐end

estimate

0.002

(low‐end

estimate)  

6

(upper‐end

estimate

0.0002

(low‐end

estimate)  

0.5

(upper‐end

estimate

0.002

(low‐end

estimate)

4

(upper‐end

estimate

0.0001

(low‐end

estimate)  

0.3

(upper‐end

estimate

0.001

(low‐end

estimate)  

2

(upper‐end

estimate

Notes:

1

Workers 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*
Page 52 of 212

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.  
Page 53 of 212

  Table 2‐14.  TCE Products in NIH’s Household Products Database

Producta %TCE Content by Weight

(as of February 2014)b

EPA/OPPT determined that these products contain TCE

Sprayway C‐60 063 solvent degreaser >90%

Sprayway C‐60 064 solvent degreaser >90%

Sprayway 201 clear protective coating spray 20‐30%

Sprayway 205 film cleaner >90%

Sprayway 208 toner aide 15‐20%

Sprayway 209 mirror edge sealant 20‐30%

EPA/OPPT determined that these products did not contain TCE

Lectra Clean 02018 heavy duty

electrical parts degreaser

None

Lectra Clean 02120 Lectra Clean II non‐chlorinated

heavy duty electrical parts degreaser

None

Sprayway 073 brake parts cleaner None

EPA/OPPT determined that these products were discontinued

Sprayway 669 gravel guard Product not found on  

manufacturer’s product list

Sprayway 732 industrial cleanup dry cleaner Product not found on

manufacturer’s product list

TrakAuto Trouble Free Rust Buster 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.

b 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.   
Page 54 of 212

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.

                                                      

11 EPA’s E‐FAST2 website:  http://www.epa.gov/oppt/exposure/pubs/efast.htm
Page 55 of 212

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 Scenario1 Population of Interesta

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:

a

   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.   

                                                      

12 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. 13 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”). 14 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”
Page 56 of 212

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
Page 57 of 212

of pure solvents (Chinn, 1981).  E‐FAST2/CEM used the Chinn 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 (CO2) 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.
Page 58 of 212

Table 2‐16. Summary of E‐FAST2/CEM’s Input Parameters and Assumptions for Estimation of Potential Acute 

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

TCE mass in product 24 g per use N/A 11 g per use N/A

To

mu

fra

Consumer Product TCE

Weight Fraction

0.9

Value from MSDS N/A

0.3

High end values

from MSDS

N/A Value 

Duration of event and

occupant behavior pattern

(hrs/event)

1 hr using product

2 hrs total spent in

room of use

24 hrs in home or

outside

0 hr using product

0 hr in room of use

24 hrs in home or

outside

0.5 hr using

product

2 hrs total spent in

room of use

24 hrs in home or

outside

0 hr using product

0 hr in room of use

24 hrs in home or

outside

The be

user an

on a da

home w

the ho

not allo

presen

during 

Frequency of use

(events/day) 1

It was a

exposu

and se

events 

bystan

primar

of TCE 

residua

metaboexpect

occurs 

EPA/OP

that th

substa

upon p

life.

Exposure duration (days) 1
Page 59 of 212

Table 2‐16. Summary of E‐FAST2/CEM’s Input Parameters and Assumptions for Estimation of Potential Acute 

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

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/Youngadults: 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 gro

EPA’s E

Handb

Age gro

showed 

concen

the sam

adult in

consum

catego

groups 

include 

based o

adults 

users o

EPA/OP

that th

by teen

even yo

particu

art/cra

coating 

not sur

behavi

create 

pattern

groups

concen

2‐17 ar

in this 
Page 60 of 212

Table 2‐16. Summary of E‐FAST2/CEM’s Input Parameters and Assumptions for Estimation of Potential Acute 

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

Air exchange rate (air

exchanges

per hr)

0.45

Recom

value oexchan

within 

2011a)

Portion of aerosol

in air, also called overspray

(unitless)

0.01

This va

percenproduc

therefoavailab

inhalaton prof

(EPA, 2

treated 

TCE as 

over th

Whole House Volume (m3) 369

Value o

(1997). EFH pro

(492 m

older v

provide estima

369 m3

median 

update

althoug

to 492 
Page 61 of 212

Table 2‐16. Summary of E‐FAST2/CEM’s Input Parameters and Assumptions for Estimation of Potential Acute 

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

Zone 1 Volume, m3

(user location) 20 N/A 20 N/A The vo

assigne

which w

hobby/volume 

m

3

. The 

entire d

outside

2 hours 

hours e

outside

Zone 2 Volume, m3 (the

rest of the house) 349

TCE Emission rate constant

(hr‐1) 101.06

Estima

algorith

model 

1981). the mo

vapor p

emissio

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 

short‐t

light ac

After u

data ob

2011a)
Page 62 of 212

Table 2‐16. Summary of E‐FAST2/CEM’s Input Parameters and Assumptions for Estimation of Potential Acute 

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

Body weight (kg) 80 Age group‐specific value 80 Age group‐specific value

Mean 

for 

bas

Body 

gr

Suppl

and w
Page 63 of 212

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 Day a) from the

Residential Indoor Use of Solvent Degreasers or Clear Protecting Coating Sprays

Clear Protective

Coating Spray User

(ppm) b

Clear Protective Coating

Spray Bystander c

(ppm)

Solvent Degreaser

User (ppm) b

Solvent Degreaser Bystander

(ppm) c

0.4 0.1 2 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.  
Page 64 of 212

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.

2.5.4.2 Indoor Air Monitoring of TCE

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 (<MDL 0.91 μg/m3

or 0.00017 ppm) to 37.4 μg/m3 (0.007 ppm) (Lindstrom and  Pleil, 1996).   

The TCE IRIS assessment included several studies that measured indoor levels of TCE, but none

with monitoring data related to the use of consumer products containing TCE (EPA, 2011e).  

EPA/OPPT also performed a literature search to determine if indoor monitoring data were

available for products using TCE, but none were found.  

A supplementary document entitled “Literature Review of Measured TCE Concentrations in

Indoor Air” (November 2013) contains additional information about the indoor monitoring data

studies reviewed for this assessment.
Page 65 of 212

2.6 HAZARD/DOSE‐RESPONSE ASSESSMENT

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

                                                      

15   NAS report, “Assessing the human health risks of trichloroethylene: Key scientific issues (2006)”:

http://www.nap.edu/catalog.php?record_id=11707   16   EPA’s SAB peer review report for the 2009 EPA’s Draft Assessment entitled “Toxicological Review of

Trichloroethylene”:

http://yosemite.epa.gov/sab/sabproduct.nsf/c91996cd39a82f648525742400690127/B73D5D39A8F184BD8525

7817004A1988/$File/EPA‐SAB‐11‐002‐unsigned.pdf
Page 66 of 212

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 Assessment

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 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 μg/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. 

                                                      

17  The dose‐response analysis for cancer endpoints can be found in Chapter 5 of the TCE IRIS assessment EPA

(2011e).
Page 67 of 212

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 of once‐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) ÷ (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 68 of 212

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).
Page 69 of 212

Table 2‐18. Lowest PBPK‐derived HECs for different effects domains based on analysis in TCE IRIS assessment

Exposure

Duration for Risk

Analysis Target Organ/ System Species Route of Exposure Range of Doses or Concentrations1 Duration POD Type2 Effect HEC50 (ppm) 3 HEC95 (ppm) 

CHRONIC 

Liver

Mouse (male) Inhalation 37 to 3,600 ppm Continuous and

intermittent

exposures,

variable time

periods for

30‐120 days

BMDL10 =

21.6 ppm

Increased

liver/body weight

ratio

25 12 

Kidney

Rat

(female)

Oral

(gavage)

500 to 1,000

mg/kg‐bw/day

5 days/week for

104 weeks

BMDL05 =

9.45 mg/kg‐

bw/day

Toxic nephropathy 0.042 0.0085 

Nervous

System

Rat

(male)

Inhalation 50 to 300 ppm 8 hrs/day,

5 days/weeks

for 6 weeks

LOAEL = 12

ppm

Significant

decreases in

wakefulness

13 6.4 

Immune System

Mouse

(female)

Oral

(drinking

water)

0.001 to 14 ppm

(0.001 to 14

mg/kg‐bw/day)

27‐30 weeks

LOAEL =

0.35 mg/kg‐

bw/day

Decrease in thymus

weight and thymus

cellularity

0.092 0.045 

Mouse

(female)

Oral

(drinking

water)

0.001 to 14 ppm

(0.001 to 14

mg/kg‐bw/day)

27‐30 weeks LOAEL =

0.35 mg/kg‐

bw/day

Autoimmunity

(increased anti‐

dsDNA and ssDNA

antibodies)

0.092 0.045 

Reproductive System

Human

(male)

Inhalation 29.6 ppm

(mean exposure)

Measured

values after an

8‐hour work

shift; mean 5.1

years on job

BMDL10 =

1.4 ppm

Decreased normal

sperm morphology

and

hyperzoospermia

1.4 0.7 
Page 70 of 212

Table 2‐18. Lowest PBPK‐derived HECs for different effects domains based on analysis in TCE IRIS assessment

Exposure

Duration for Risk

Analysis Target Organ/ System Species Route of Exposure Range of Doses or Concentrations1 Duration POD Type2 Effect HEC50 (ppm) 3 HEC95 (ppm) 

ACUTE OR 

CHRONIC Developmental effects

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 

Notes: 1 Control concentrations were not included in the table, but discussed in the study summaries in section 2.6.2.

2 POD type can be NOAEL, LOAEL, or BMDL; the IRIS program adjusted all values to continuous exposure.3 1 ppm = 5.374 mg/m34 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 

toxicity, whereas a total UF of 30 was used for the autoimmune effects. The TCE IRIS assessment allocated different LOAEL‐to‐NOAEL u

based on the severity of the effects, which resulted in different total UF (EPA, 2011e).
Page 71 of 212

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# NON‐CANCER IRIS APPROACH

NON‐CANCER OPPT

APPROACH

Considered (+) / Not

Considered (―)  

1 Evaluation of all studies (inhalation and oral) that provided (1) non‐cancer

adverse health effects and (2) quantitative dose‐response data +

2 Identification of the Points of Departure (POD) based on applied dose +

3 Adjustment of each POD by endpoint/study‐specific uncertainty factors

(UFs) ― 

4 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

― 

5 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

+

6 Estimation of interspecies and within‐human pharmacokinetic variability

for the internal dose PODs by PBPK modeling +

7 Adjustment of each internal dose POD by endpoint/study‐specific UFs 2  ―

8 Derivation of PBPK model‐based candidate RfCs (p‐cRfC) for each

candidate critical effect   ― 

9 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

― 

10 Evaluation of the most sensitive cRfCs and set final RfC value.  ― 

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.
Page 72 of 212

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 73 of 212

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 GSH Conjugation Metabolites

Chloral  

(metabolized to TCOHa

)

DCVGe

(metabolized to DCVCf

isomers)

Trichloroethylene oxide

(re‐arranged to DCACb

)

Trichloroethanol or TCOH  

(metabolized to TCOGc

)

Trichloroacetic acid or TCA  

(may lead to DCAd

)

Abbreviations: a TCOH = trichloroethanol; b 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).
Page 74 of 212

Table 2‐21.  Common Metabolites of TCE and Related Compounds

                      Parent  → 

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    X   

Chloral X    X      

Chloral hydrate (CH) X    X        

Monochloroacetic acid X X X X X X

Dichloroacetic acid (DCA) X X X     X

Trichloroacetic acid (TCA) X X X X    

Trichloroethanol (TCOH) X X X X    

Trichloroethanol‐

glucuronide

X X X X    

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‐(1,2‐

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‐(1,2‐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,

NAcDCVC) (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 β‐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 (FMO3) 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‐glutamyl transpeptidase (i.e., GGT, an enzyme

involved in DCVC production) activity in kidneys seems to be highest in rats, then humans, and
Page 75 of 212

then mice (EPA, 2011e). Furthermore, species‐dependent enzymatic activities have been

reported for the ‐lyase and FMO3 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 CO2, 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 CO2 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).   
Page 76 of 212

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

for TTC, 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; Prout et al., 1985).  

2.6.1.4 PBPK Modeling Approach Supporting the TCE IRIS Assessment

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 below to

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

Chiu et al. (2009).   
Page 77 of 212

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

Explanation of What the  

Dose‐Metric Identifier Represents

ABioactDCVCBW3/4 Amount of DCVC bioactivated in the kidney per unit body weight

ABioactDCVCKid Amount of DCVC bioactivated in the kidney per unit kidney mass

AMetGSHBBW3/4 Amount of TCE conjugated with GSH

AMetLiv1BW3/4 Amount of TCE oxidized in liver

AMetLivOtherBW3/4 Amount of TCE oxidized to metabolites other than TCA or TCOH per unit body weight

AMetLivOtherLiv Amount of TCE oxidized to metabolites other than TCA or TCOH per unit liver weight

AMetLngBW3/4 Amount of TCE oxidized in respiratory tract (per unit body weight)

AMetLngResp Amount of TCE oxidized in respiratory tract per unit respiratory tract tissue

AUCCBld Area under the curve of venous blood concentration of TCE

AUCCTCOH Area under the curve of blood concentration of TCOH

AUCLivTCA Area under the curve of the liver concentration of TCA

TotMetabBW3/4 Total amount of TCE metabolized per unit body weight

TotOxMetabBW3/4 Total amount of TCE oxidized per unit body weight

TotTCAInBW Total amount of TCA produced

Table 2‐23.  Dose‐Metrics for Cancer and Non‐Cancer Endpoints Used in the OPPT Assessment Non-Cancer Endpoints

Target Organ/ System Reference Preferred

Dose‐Metric 1

Liver Kjellstrand et al. (1983) AMetLiv1BW3/4

Kidney NTP (1988) ABioactDCVCBW3/4

Nervous System Arito et al. (1994) TotMetabBW3/4

Immune System Keil et al. (2009) TotMetabBW3/4

Reproductive System Chia et al. (1996) TotMetabBW3/4

Developmental effects 2 Johnson et al. (2003) TotOxMetabBW3/4    

Cancer 

Carcinogenic effects in

various organs

Charbotel et al. (2006)(kidney

cancer) and human epidemiological

studies for NHL and liver cancer

ABioactDCVCBW3/4

(IUR for kidney cancer)

AMetLiv1BW3/4 (IUR for liver cancer)

TotMetabBW3/4 (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.
Page 78 of 212

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

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 79 of 212

Figure 2‐5.  Example of HEC99 Estimation through Interpecies, Intraspecies and Route‐to‐

Route Extrapolation from a Rodent Study LOAEL/NOAEL

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

HEC99 estimates, respectively. The HEC95 and HEC99 were 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 HEC50 and HEC95 values. However, risk estimates based on the HEC50 and HEC95 values are

also included.   
Page 80 of 212

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 HEC95/99 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 HEC Equivalent to the Benchmark Dose Lower Confidence Limit

for Cardiac Malformations (g/m3

) a

Median estimate Upper 99th percentile estimate

Chronic (steady‐state) 62 20

9 months (40 weeks) 62 20

3 weeks 63 20

1 day 71 21

Notes:

a   The internal dose metric of μg oxidized per day per [kg body weight]3⁄4 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.  
Page 81 of 212

Table 2‐25.  Comparison of TCE Human Equivalent Concentrations (HECs) Under

Intermittent (Occupational) Exposure for Different Exposure Durations

Duration HEC Equivalent to the Benchmark Dose Lower Confidence Limit

for Cardiac Malformations (g/m3

) a

Median estimate Upper 99th percentile estimate

Chronic (steady‐state) 62 20

9 months (40 weeks) 62 20

3 weeks 63 20

1 day 68 21

Notes:

a   The internal dose metric of μg oxidized per day per [kg body weight]3⁄4 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 μg/m3

, to which a duration adjustment of 8/24 was applied to derive the reported value of 68

μg/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.   

2.6.2.1 Genetic Toxicity of TCE and its metabolites

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).     
Page 82 of 212

2.6.2.2 Human Toxicity Following Acute Exposure to TCE

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 et al. (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.

2.6.2.3 Toxicity Following Repeated Exposures to TCE		

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
Page 83 of 212

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 (Including 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 et al.

(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 et al. (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 (BMDL10) 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 BMDL10 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 BMDL10 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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2.6.2.3.2 Kidney Toxicity (Including Cancer)

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 β‐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. BMDL10 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 BMDL05 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 BMDL10 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 BMDL10 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 × time).  

Using historical monitoring data, mean exposures were calculated as 704 ppm × 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#1 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
Page 89 of 212

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).   

2.6.2.3.4 Immunotoxicity (Including Cancer)

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 (Iavicoli 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 following 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‐1 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 BMDL1SD

27 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.
Page 91 of 212

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).   

2.6.2.3.5 Reproductive Toxicity

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 for TCE‐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
Page 92 of 212

sperm morphology and hyperzoospermia. A BMDL10 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‐1 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,
Page 93 of 212

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‐1 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‐1 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
Page 94 of 212

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)]. 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), Ruckart et al. (2013)]. 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
Page 95 of 212

relative risk estimate of 6.2 (95% CI: 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% CI: 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 (CI) of 1.07 to 1.73), small for gestational age [RR of 1.23; 95

percent (CI) of 1.03 to 1.48], and cardiac defects (RR of 2.15; 95 percent CI 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 Th1 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 et al., 1987; Bove, 1996; Bove et al., 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).
Page 96 of 212

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 PFC response was seen in male pups after 3 and 8

weeks of exposure, whereas female pups showed the suppression of PFC 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 PFC 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 BMDL01 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 liƩers)

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).
Page 97 of 212

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 BMDL1SD

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 for TCE‐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:

1_Blossom_NObject_SD.xlsm and 2_Blossom_NMouse_SD.xlsm).
Page 98 of 212

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 BMDL01 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

et al., 2000; Bross et al., 1983; Drake, V. et al., 2006; Drake, V. J. et al., 2006; Loeber et al., 1988;

Mishima et al., 2006; Rufer et al., 2008) and oral developmental toxicity studies in rats (Dawson

et al., 1990, 1993; Johnson et al., 2005; Johnson, 2014; Johnson et al., 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 et al., 1992; Johnson et al., 1998a, 1998b; Smith et al., 1989, 1992). For example, the

Johnson et al. and Smith et al. studies reported increased incidences of cardiac malformation

following gestational TCA exposures (Johnson et al., 1998a, 1998b; Smith et al., 1989).  

Similarly, pregnant rats exhibited increased incidence of cardiac defects following DCA exposure

during pregnancy (Epstein et al., 1992; Smith et al., 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 RfD/RfC derivation with the critical effect of

cardiac malformations.
Page 99 of 212

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 et al., 2008; Collier et al., 2003; Selmin et al., 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 HEC99 and 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.

2.7 HUMAN HEALTH RISK CHARACTERIZATION		

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.  
Page 100 of 212

Table 2‐26.  Summary of Hazard Information Used in the Risk Evaluation of Acute and Chronic Scenarios

Exposure

Duration for

Risk Analysis Target Organ/

System Species Route ofExposure

POD Type Effect HEC99

(ppm)

Total

Uncertainty

Factor (UF) for

Benchmark

MOE

Reference

CHRONIC 

Liver Mouse

(male)

Inhalation BMDL10 = 21.6

ppm

Increased liver/body

weight ratio 9.1

Total UF=10 Kjellstrand

et al. (1983)

Kidney Rat

(female)

Oral (gavage) BMDL05 = 9.45

mg/kg‐bw/day Toxic nephropathy 0.0056 Total UF=10

NTP (1988)

Nervous

System

Rat

(male)

Inhalation LOAEL = 12

ppm

Significant decreases in

wakefulness 4.8

Total UF=300 Arito et al.

(1994)

Immune

System

Mouse

(female)

Oral (drinking

water) LOAEL = 0.35

mg/kg‐bw/day

Decrease in thymus

weight and thymus

cellularity

0.033

Total UF=100 Keil et al.

(2009)

Mouse

(female)

Oral (drinking

water) LOAEL = 0.35

mg/kg‐bw/day

Autoimmunity

(increased anti‐dsDNA

and ssDNA antibodies)

0.033

Total UF=30 Keil, D. E. et

al. (2009)

Reproductive

System

Human

(male)

Inhalation

BMDL10 = 1.4

ppm

Decreased normal

sperm morphology and

hyperzoospermia

0.5

Total UF=30

Chia et al.

(1996)

ACUTE OR 

CHRONIC 

Developmental

effects

Rat

(female)

Oral

(drinking

water)

BMDL01 =

0.0207 mg/kg‐

bw/day Heart malformations 0.0037

Total UF=10

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).
Page 101 of 212

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

(4)

Residential use of

arts/crafts clear

protective coating

spray

Population of Interest and

Exposure Scenario:

Users

Adult pregnant1

worker (>16 years

old) exposed to TCE

for a single 2‐hr

exposure during an

8‐hr workday 2,3, 4

.  

Adult pregnant1

worker (>16 years

old) exposed to TCE

for a single 8‐hr

exposure 2,3.   

Adult pregnant 1

consumers (>16 yrs

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.   

Adult pregnant 1

consumers (>16 yrs

old) exposed to TCE

for a single 0.5‐hr

exposure when

using product once

per week 2,3,4.

Residential

exposure model

provided the 24‐hr

acute exposure

estimate.   

Population of Interest and

Exposure Scenario:

Bystander

Adult pregnant women 1 (>16 years old)

exposed to TCE indirectly by being in the

same building.  

Adult pregnant bystander1

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 et al., 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=1) x (UFA= 3) x (UFH=3) x (UFL=1) 6 =  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
Page 102 of 212

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:

1

  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.
Page 103 of 212

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 × 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 104 of 212

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 HEC50/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 105 of 212

Table 2‐31. Acute Non‐Cancer Risk Estimates for Commercial Use of Degreaser Product at Small Shops 

Effects: Congenital Heart Malformations, Johnson et al., 2003)

Lowest PBPK‐

derived

HECs (ppm)

of the

developmental

toxicity health

domain

WORKER NON‐CANCER MOEs BYSTANDER NON‐CANCER MOEs

With LEV—

Low‐end

exposure

estimate

With LEV—

Upper‐end

exposure

estimate

No LEV—

Low‐end

exposure

estimate

No LEV—

Upper‐end

exposure

estimate

With LEV—

Low‐end

exposure

estimate

With LEV—

Upper‐end

exposure

estimate

No LEV—

Low‐end

exposure

estimate

No LEV

Upper‐e

exposu

estima

HEC50= 0.012 0.12 0.0018 0.012 0.00018 0.9 0.002 0.09 0.0002

HEC95= 0.0051 0.051 0.0008 0.0051 0.00008 0.38 0.0009 0.038 0.0000HEC99= 0.0037 0.037 0.0006 0.0037 0.00006 0.28 0.0007 0.028 0.0000

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 Faci

(Developmental Effects: Congenital Heart Malformations, Johnson et al., 2003)

Lowest PBPK‐

derived

HECs (ppm)

of the

developmental

toxicity health

domain

WORKER NON‐CANCER MOEs BYSTANDER NON‐CANCER MOEs

With LEV—

Low‐end

exposure

estimate

With LEV—

Upper‐end

exposure

estimate

No LEV—

Low‐end

exposure

estimate

No LEV—

Upper‐end

exposure

estimate

With LEV—

Low‐end

exposure

estimate

With LEV—

Upper‐end

exposure

estimate

No LEV—

Low‐end

exposure

estimate

No LEV

Upper‐e

exposu

estima

HEC50= 0.012 4.5 0.018 0.45 0.0019 51.4 0.018 5.1 0.002

HEC95= 0.0051 1.9 0.0077 0.19 0.00081 21.9 0.0077 2.2 0.0008HEC99= 0.0037 1.4 0.0056 0.14 0.00058 15.9 0.0056 1.6 0.0006

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 106 of 212

Table 2‐33. Acute Non‐Cancer Risk Estimates for Residential Uses of TCE‐containing degreasers and art/

(Developmental Effects: Congenital Heart Malformations, Johnson et al., 2003)

Lowest PBPK‐

derived

HECs (ppm)

of the

developmental

toxicity health

domain

RESIDENTIAL USE OF DEGREASER PRODUCT

MOEs 1

RESIDENTIAL USE OF ARTS/CRAFTS CLEAR

PROTECTIVE COATING SPRAY

MOEs 1 

USER 2 BYSTANDER 3 USER 2 BYSTANDER3

HEC50= 0.012 0.0060 0.015 0.03 0.12 

HEC95= 0.0051 0.0026 0.0064 0.013 0.051 HEC99= 0.0037 0.0019 0.0046 0.0093 0.037 

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 

adults, but teenagers and even children may be users or be in the same room with the user while engaging in arts and crafts pr

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 107 of 212

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 1x10‐4 and 1x10‐6 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. 1x10‐6:   the probability of 1 chance in 1 million of an individual developing cancer

2. 1x10‐5:   the probability of 1 chance in 100,000 of an individual developing cancer, which

is equivalent to 10  cancer cases in 1 million

3. 1x10‐4

:   the probability of 1 chance in 10,000 of an individual developing cancer, which

is equivalent to 100 cancer cases in 1 million

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 1x10‐5

and 1x10‐6 (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 1x10‐4 (i.e., “Spot cleaner + LEV—Typical Exposure”;

“Bystander + LEV—Typical Exposure”; and “Bystander No LEV—Typical Exposure”) (Figure 2‐7).
Page 108 of 212

Figure 2‐6.  Cancer Risk Estimates for Commercial Use of Degreaser Product at Small Shops

Figure 2‐7.  Cancer Risk Estimates for Commercial Use of Spotting Agent at Dry Cleaning

Facilities

0.1 1 10 100 1000 10000 1000001000000

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

Number of Human Cancer Cases Per Million

1x10‐4 1x10‐5 1x10‐6

0.1 1 10 100 1000 10000 1000001000000

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

Number of Human Cancer Cases Per Million

1x10‐4 1x10‐5 1x10‐6
Page 109 of 212

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<benchmark MOE) for the following chronic non‐cancer effects: developmental

toxicity, kidney and immunotoxicity. Also, all of the worker scenarios showed risks for

reproductive effects and neurotoxicity. These risks were observed in the worker/bystander

scenarios regardless of the type of TCE exposure (typical vs. worst case) and the availability of

local exhaust ventilation (with LEV vs. no LEV). Moreover, risks for reproductive effects and

neurotoxicity were also found in the degreaser bystander scenarios, but only for worst case

exposure conditions and typical exposure circumstances with no local exhaust ventilation.  

Furthermore, risks were also found for liver effects, but only for those degreaser

worker/bystander scenarios representing worst case exposure conditions irrespective of the

presence of room ventilation (Table 2‐34).

The non‐cancer chronic risk profile was slightly different for the TCE use in spot cleaners

because the TCE exposure concentrations in dry cleaning facilities were 10 to 60‐fold lower

than those estimated in degreaser small shops. Risks for developmental effects, kidney toxicity

and immunotoxic effects were generally observed for all of the spot cleaning worker/bystander

exposure scenarios (i.e., typical vs worst case; typical exposure vs worst case) with two

exceptions. Non‐risk findings were reported for few spot cleaning worker/bystander scenarios

under LEV and typical exposure conditions that used the HEC50th percentile for the MOE

calculations. Also, the spot cleaning bystander exposure scenario did not report potential non‐

                                                      

30 Risk estimates were calculated for all of the studies per health effects domain that the EPA’s IRIS program

identified as suitable for dose‐response analysis.  These risk estimates are found in the supplemental Excel file

“TCE OPPT Risk Estimates_061814”.
Page 110 of 212

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 10-6 to 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 111 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 112 of 212

Table 2‐34. Chronic Non‐Cancer Risk Estimates for Commercial Use of Degreaser Product at Small Shops

Health Effect

Domain and Study

Lowest HEC

(ppm) of each

health effects

domain

WORKER NON‐CANCER MOEs 1 BYSTANDER NON‐CANCER M

With LEV‐‐ 

Low‐end

exposure

estimate

No LEV‐‐ 

Low‐end

exposure

estimate

With LEV‐‐ 

Upper‐end

exposure

estimate

No LEV‐‐ 

Upper‐end

exposure

estimate

With LEV‐‐ 

Low‐end

exposure

estimate

No LEV‐‐ 

Low‐end

exposure

estimate

With LEV‐

Upper‐end

exposure

estimate

DEVELOPMENTAL

TOXICITY

(Johnson et al.,

2003)

HEC50= 0.012 0.17 0.017 0.0025 0.0003 1.3 0.13 0.0030 

HEC95= 0.0051 0.072 0.0072 0.0011 0.0001 0.54 0.05 0.0013 HEC99= 0.0037 0.052 0.0052 0.0008 0.0001 0.40 0.04 0.0009 

KIDNEY

(NTP, 1998)

HEC50= 0.042 0.59 0.059 0.0088 0.0009 4.4 0.4 0.01 

HEC95= 0.0085 0.11 0.012 0.0018 0.0002 0.9 0.09 0.0021 HEC99= 0.0056 0.079 0.0079 0.0012 0.0001 0.6 0.06 0.0014 

IMMUNOTOXICITY

Keil et al., 2009

(Decrease in thymus

weight and thymus

cellularity)

HEC50= 0.092 1.3 0.1 0.019 0.0020 9.7 0.97 0.023 

HEC95= 0.045 0.6 0.06 0.0095 0.0010 4.7 0.47 0.011 HEC99= 0.033 0.5 0.05 0.0069 0.0007 3.5 0.35 0.0082 

IMMUNOTOXICITY

Keil et al., 2009

(Autoimmunity)

HEC50= 0.092 1.3 0.13 0.019 0.0020 9.7 0.97 0.022 

HEC95= 0.045 0.6 0.063 0.0095 0.0010 4.7 0.47 0.011 HEC99= 0.033 0.5 0.046 0.0069 0.0007 3.5 0.35 0.0082 

REPRODUCTIVE

TOXICITY

(Chia et al. 1996)

HEC50= 1.4 19.7 2.0 0.3 0.03 147 15 0.35 

HEC95= 0.7 9.8 1.0 0.15 0.015 74 7 0.17 HEC99= 0.5 7.0 0.7 0.11 0.011 53 5 0.12 

NEUROTOXICITY

(Arito et al., 1994)

HEC50= 13 183 18 2.7 0.28 1369 137 3.2 

HEC95= 6.4 90 9 1.3 0.14 674 67 1.6 HEC99= 4.8 67 7 1.0 0.10 505 51 1.2 

LIVER

(Kjellstrand et al.

1983)

HEC50= 25 351 35 5.3 0.53 2632 263 6 

HEC95= 12 168 17 2.5 0.26 1263 126 3 

HEC99= 9.1 128 13 1.9 0.19 958 96 2 

Notes: (1) MOEs below benchmark MOE indicating risk are denoted in bold text. They indicate potential health risks. (2) Exposure es

are found in Table 2‐10.
Page 113 of 212

Table 2‐35. Chronic Non‐Cancer Risk Estimates for Commercial Use of Spotting Agent at Dry Cleaning Facilitie

Health Effect

Domain and Study

Lowest HEC

(ppm) of each

health effects

domain

WORKER NON‐CANCER MOEs 1 BYSTANDER NON‐CANCER M

With LEV‐‐ 

Low‐end

exposure

estimate

No LEV‐‐ 

Low‐end

exposure

estimate

With LEV‐‐ 

Upper‐end

exposure

estimate

No LEV‐‐ 

Upper‐end

exposure

estimate

With LEV‐‐ 

Low‐end

exposure

estimate

No LEV‐‐ 

Low‐end

exposure

estimate

With LEV‐

Upper‐end

exposure

estimate

DEVELOPMENTAL

TOXICITY

(Johnson et al.,

2003)

HEC50= 0.012 6.3 0.0253 0.63 0.0027 72 0.0253 7.2 

HEC95= 0.0051 2.7 0.0107 0.27 0.0011 31 0.0107 3.1 HEC99= 0.0037 1.9 0.0078 0.19 0.0008 22 0.0078 2.2 

KIDNEY

(NTP, 1998)

HEC50= 0.042 22 0.088 2.2 0.0093 253 0.0884 25 

HEC95= 0.0085 4.5 0.018 0.45 0.0019 51 0.0179 5.1 HEC99= 0.0056 2.9 0.012 0.29 0.0012 34 0.0118 3.4 

IMMUNOTOXICITY

Keil et al., 2009

(Decrease in thymus

weight and thymus

cellularity)

HEC50= 0.092 48 0.194 4.8 0.020 554 0.194 55 

HEC95= 0.045 24 0.095 2.4 0.010 271 0.095 27 HEC99= 0.033 17 0.069 1.7 0.007 199 0.069 20 

IMMUNOTOXICITY

Keil et al., 2009

(Autoimmunity)

HEC50= 0.092 48 0.194 4.8 0.020 554 0.194 55 

HEC95= 0.045 24 0.095 2.4 0.010 271 0.095 27 HEC99= 0.033 17 0.069 1.7 0.007 199 0.069 20 

REPRODUCTIVE

TOXICITY

(Chia et al. 1996)

HEC50= 1.4 737 2.9 74 0.31 8423 2.9 842 

HEC95= 0.7 369 1.5 37 0.16 4212 1.5 421 HEC99= 0.5 263 1.1 26 0.11 3008 1.1 301 

NEUROTOXICITY

(Arito et al., 1994)

HEC50= 13 6844 27 684 2.9 78214 27 7821 

HEC95= 6.4 3369 13 337 1.4 38505 13 3851 HEC99= 4.8 2527 10 253 1.1 28879 10 2888 

LIVER

(Kjellstrand et al.

1983)

HEC50= 25 13161 53 1316 5.5 150412 53 15041 

HEC95= 12 6317 25 632 2.7 72198 25 7220 

HEC99= 9.1 4791 19 479 2.0 54750 19 5475 

Notes: (1) MOEs below benchmark MOE indicating risk are denoted in bold text. They indicate potential health risks. (2) Exposure est

found in Table 2‐13.
Page 114 of 212

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 (intra‐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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2.8.2 Uncertainties in the Hazard and Dose‐Response Assessments

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 × 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 × 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), “[c]onvergence of the

MCMC samples for model parameters was good for mice, and adequate for rats and

humans. Evaluation of posterior parameter distributions suggest[ed] 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 Uncertainties in the Risk Assessment

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 UFA 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 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 UFA was 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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2.9 CONCLUSIONS OF THE HUMAN HEALTH RISK ASSESSMENT

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

MOE of 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.
Page 127 of 212

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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,

2011e).

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.

ATSDR (Agency for Toxic Substances and Disease Registry). 1997. Toxicological Profile for

Trichloroethylene. Division of Toxicology, Toxicology Information Branch, Atlanta, GA.

http://www.atsdr.cdc.gov/toxprofiles/tp19.pdf.

ATSDR (Agency for Toxic Substances and Disease Registry). 1998. Volatile Organic Compounds

in Drinking Water and Adverse Pregnancy Outcomes. United States Marine Corps Base,

Camp Lejeune, NC. PB98‐156540. Atlanta, GA. http://www.atsdr.cdc.gov/hs/lejeune/.

(as cited in EPA, 2011e).

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Page 133 of 212

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Page 135 of 212

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Page 136 of 212

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Page 137 of 212

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Hazardous Air Pollutants: Halogenated Solvent Cleaning. 72 Federal Register 85 (May 3,

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Trichloroethylene on the Exploratory and Locomotor Activity of Rats Exposed During

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591‐600.

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Page 151 of 212

APPENDICES

Appendix A PRODUCTION VOLUME AND INVENTORY

UPDATE RULE DATA

In 2011, global consumption of TCE was 945 million lbs with expected growth at about 1.5

percent annually over the next five years. The corresponding U.S. consumption was 255 million

lbs (Glauser and  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 lbs 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‐1.  National Chemical Information for TCE from 2012 CDR

Production Volume (aggregate) 224.7 million lbs

Maximum Concentration (at manufacture or import site) >90%

Physical form(s) Liquid

Number of reasonably likely to be exposed industrial manufacturing, processing,

and use workers (aggregated)

>1,000

Was industrial processing or use information reported? Yes

Was commercial or consumer use information reported? Yes

                                                      

31 This source material includes data or information derived from IHS Products provided to the U.S. EPA.  IHS

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 IHS is not liable to you or any third party for your use of

and/or reliance on the IHS data and information contained in this document, and any such use shall be at your

own risk.
Page 152 of 212

Table A‐2.  Summary of Industrial TCE Uses from 2012 CDR

Industrial Sector

(Based on NAICS) Industrial Function Type of Processing

All Other Basic Organic

Chemical Manufacturing Intermediates Processing as a reactant

Industrial Gas Manufacturing Functional fluids

(closed systems) Processing as a reactant

Wholesale and Retail Trade Solvents

(for cleaning or degreasing) Processing as a reactant

Agriculture, Forestry, Fishing

and Hunting

Agricultural chemicals

(non‐pesticidal) Processing as a reactant

All Other Basic Organic

Chemical Manufacturing

Solvents

(for cleaning or degreasing) Processing‐repackaging

All Other Chemical Product and

Preparation Manufacturing

Not Known or Reasonably

Ascertainable Use‐non‐incorporative activities

All Other Chemical Product and

Preparation Manufacturing

Solvents

(which become part of product

formulation or mixture)

Processing‐incorporation into

formulation, mixture, or reaction

product

All Other Basic Organic Chemical

Manufacturing

Not Known or Reasonably

Ascertainable Processing‐repackaging

Primary Metal Manufacturing Solvents

(for cleaning or degreasing)

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
Page 153 of 212

Appendix B REGULATORY HISTORY OF TCE AT THE USEPA

AND RELATED ACTIONS		

B‐1 Regulatory History of TCE at the U.S. EPA

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 μg/L maximum daily average and 26 μg/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 μg/L for drinking

water and 370 μg/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 154 of 212

EPA/OSWER set a reportable quantity of 100 lbs (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 (MOU) 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 MOU 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 155 of 212

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 1B), 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 156 of 212

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 157 of 212

Appendix C ENVIRONMENTAL FATE OF TCE		

C‐1 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‐1‐1 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‐1‐2 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 158 of 212

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) (Dilling et al., 1975).  

Based on these studies, biodegradation and hydrolysis in surface waters are not expected to be

important environmental fate processes.

C‐1‐3 Fate in Soil, Sediment, and Groundwater

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; Mu and  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 159 of 212

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‐1‐4 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‐1‐5 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 160 of 212

Table C‐1.  Environmental Fate Characteristics of TCEa

Property Value

CASRN 79‐01‐6

Photodegradation half‐life 13.2 days (estimated)

Hydrolysis half‐life Does not hydrolyze under environmental conditionsb

Biodegradation 19% after 28 days (not readily biodegradable)b;

4% after 28 days (not inherently biodegradable)b

;

100% after 2 days (anaerobic conditions using mixed march cultures)b

;

2.4% after 14 days (not readily biodegradable)c

Bioconcentration 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)a

Log Koc 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)

Fugacity (Level III Model)b

Air (%)

Water (%)

Soil (%)

Sediment (%)

35.4

54.2

10.1

0.3

Persistenced P2 (moderate)

Bioaccumulationd B1 (low)

Sources:

a

  EPA (2012b) b

  ECB (2000) c

  NITE (2012) d  EPA (1999)
Page 161 of 212

Appendix D NAICS CODES FOR TCE DEGREASING		

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‐1).   

Table D‐1.  TCE Used as a Degreaser Primarily in These Industries (USDOC, 2008)

NAICS Codes

33121 321113 332112 332721 332999 334417 335929 339114

33272 323116 332116 332722 333132 334419 335999 339992

33341 325188 332117 332811 333298 334513 336321 339995

33422 325998 332211 332812 333311 334515 336340 339999

33512 326299 332212 332813 333415 335121 336411 488111

33531 331111 332311 332912 333921 335211 336413 493110

33634 331210 332313 332913 333994 335312 336414 811310

33641 331419 332431 332919 333999 335313 336510 928110

33999 331421 332510 332994 334413 335911 337125   

314999 332111 332618 332996 334414 335921 337127   

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 162 of 212

Appendix E ESTIMATION OF TCE EMISSION RATE AT SMALL DEGREASIN
Page 163 of 212

Appendix F ESTIMATION OF TCE EXPOSURES AT SMALL DEGREASING FA
Page 164 of 212

Appendix G CALCULATION OF TCE EXPOSURES AT SMALL

DEGREASING FACILITIES

In Figure G‐1, 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‐1.

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.

   

Near‐Field Mass Balance

ࡲࡺࢂ

ࡲࡺ࡯ࢊ

[1‐G [ࡳ ൅ ࡲࡺࡽࡲࡺ࡯ െ ࡲࡺࡽࡲࡲ࡯ ൌ ࢚ࢊ
Page 165 of 212

Far‐Field Mass Balance

ࡲࡲࢂ

ࡲࡲ࡯ࢊ

[2‐G [ࡲࡲࡽࡲࡲ࡯ െ ࡲࡺࡽࡲி࡯ െ ࡲࡺࡽࡲࡺ࡯ ൌ ࢚ࢊ

Where:

ܸேி   =  near‐field volume

ܸிி =  far‐field volume

ܳேி =  near‐field ventilation rate

ܳிி =  far‐field ventilation rate

ܥேி =  average near‐field concentration

ܥிி =  average far‐field concentration

ܩ=   average generation rate

time   =  ݐ

At steady‐state, Equations [G‐1] and [G‐2] can be reduced to the following:

ࡳ ൌ ࡲࡺ࡯

ࡲࡺࡽ െ૚ ቀ ࡲࡺࡽ

ࡲࡲࡽ ൅ ࡲࡺࡽ

ቁ [G‐3]

ࡲࡺࡽࡲࡺ࡯ ൌ ࡲࡲ࡯

ࡲࡲࡽ ൅ ࡲࡺࡽ

[G‐4]

Equations [G‐1] and [G‐2] can be solved for the time‐varying concentrations in the near‐field

and far‐field (Keil, C. B. et al., 2009).   

ܥேி ൌ ܩ൫݇ଵ ൅ ݇ଶ݁ఒభ௧ െ ݇ଷ݁ఒమ௧൯ [G‐5]

ܥிி ൌ ܩ ൬ 1

ܳிி

൅ ݇ସ݁ఒభ௧ െ ݇ହ݁ఒమ௧

൰ [G‐6]

Where:

݇ଵ ൌ 1

ቀ ܳேி

ܳேி ൅ ܳிி

ቁ ܳிி

[G‐7]

݇ଶ ൌ ܳேிܳிி ൅ ߣଶܸேிሺܳேி ൅ ܳிிሻ

ܳேிܳிிܸேிሺߣଵ െ ߣଶሻ [G‐8]
Page 166 of 212

݇ଷ ൌ ܳேிܳிி ൅ ߣଵܸேிሺܳேி ൅ ܳிிሻ

ܳேிܳிிܸேிሺߣଵ െ ߣଶሻ [G‐9]

݇ସ ൌ ൬

ߣଵܸேி ൅ ܳேி

ܳேி

൰ ݇ଶ [G‐10]

݇ହ ൌ ൬

ߣଶܸேி ൅ ܳேி

ܳேி

൰ ݇ଷ [G‐11]

ߣଵ ൌ 0.5 ቎െ ቆܳேிܸிி ൅ ܸேிሺܳேி ൅ ܳிிሻ

ܸேிܸிி ቇ

൅ ඨቆ

ܳேிܸிி ൅ ܸேிሺܳேி ൅ ܳிிሻ

ܸேிܸிி ቇ

ଶ

െ 4 ൬

ܳேிܳிி

ܸேிܸிி

൰ ቏

[G‐12]

ߣଶ ൌ 0.5 ቎െ ቆܳேிܸிி ൅ ܸேிሺܳேி ൅ ܳிிሻ

ܸேிܸிி ቇ

െ ඨቆ

ܳேிܸிி ൅ ܸேிሺܳேி ൅ ܳிிሻ

ܸேிܸிி ቇ

ଶ

െ 4 ൬

ܳேிܳிி

ܸேிܸிி

൰ ቏

[G‐13]

Time‐weighted‐average (TWA) concentrations in the near‐field and far‐field can be calculated

as follows:

ܥேி,்ௐ஺ ൌ ׬ ܥேி

௧మ

ݐ݀ ௧భ

௧మ ݐ݀ ׬

௧భ

ൌ ׬ ܩ൫݇ଵ ൅ ݇ଶ݁ఒభ௧ െ ݇ଷ݁ఒమ௧൯ ௧మ

ݐ݀ ௧భ

ሺݐଶ െ ݐଵሻ

ൌ

൭ܩ ൬݇ଵݐଶ ൅ ݇ଶ݁ఒభ௧మ

ଵߣ

െ ݇ଷ݁ఒమ௧మ

ߣଶ ൰ െ ܩ ൬݇ଵݐଵ ൅ ݇ଶ݁ఒభ௧భ

ଵߣ

െ ݇ଷ݁ఒమ௧భ

ߣଶ ൰൱

ሺݐଶ െ ݐଵሻ

[G‐14]
Page 167 of 212

ܥிி,்ௐ஺ ൌ ׬ ܥிி

௧మ

ݐ݀ ௧భ

௧మ ݐ݀ ׬

௧భ

ൌ

1 ቀ ܩ ׬

ܳிி

൅ ݇ସ݁ఒభ௧ െ ݇ହ݁ఒమ௧ቁ ௧మݐ݀ ௧భ

ሺݐଶ െ ݐଵሻ

ൌ

൭ܩ ൬ ݐଶ

ܳிி

൅ ݇ସ݁ఒభ௧మ

ଵߣ

െ ݇ହ݁ఒమ௧మ

ଶߣ

൰ െ ܩ ൬ ݐଵ

ܳிி

൅ ݇ସ݁ఒభ௧భ

ଵߣ

െ ݇ହ݁ఒమ௧భ

ଶߣ

൰൱

ሺݐଶ െ ݐଵሻ

[G‐15]

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, ܨܵܣ, 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‐1, the Near‐Field floor would not be available for mass transfer.  

Thus, ܨܵܣ would be less than the actual surface area of the Near‐Field:

ܨܵܣ ൌ 2ሺܮேி ∗ ܪேிሻ ൅ 2ሺܹேி ∗ ܪேிሻ ൅ ሺܮேி ∗ ܹேிሻ [G‐16]

Where:  ܮேி, ܹேி, ܪேி are the length, width, and height of the Near‐Field, respectively.

If the Near‐Field indoor wind speed, ݒேி, is known and the area for mass transfer into and from

the Near‐Field is equal, then the Near‐Field ventilation rate, ܳேி, is given by:

  

ܳேி ൌ 1

ேிݒ ∗ ܣܵܨ ∗ 2

[G‐17]

If the Far‐Field volume, ܸிி, and the air exchange rate, ܧܣܴ, are known, then the Far‐Field

ventilation rate, ܳிி, is given by:

ܳிி ൌ ܸிி ∗ ܧܣܴ] G‐18]

Based on the model inputs in Table G‐1, 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 168 of 212

Table G‐1.  Model Inputs for Small / Industrial Commercial Degreasing Facilities

Parameter Units Parameter Values Comments

ܸிி

ftଷ

(mଷ)

10,593 to 70,620

(300 to 2,000)

Values supported by von Grote et al. (2003)

ܧܣܴ 1/hr 2 to 15

Values supported by von Grote et al. (2003)  

and EPA (2013b)

ேிݒ

cm / s

(m/hr)

10

(360)

Value is ~50th percentile supported by

Baldwin and Maynard (1998)

ேிܮ

ft

(m)

10

(3.05)

Assumes volatile source is centered in the

near‐field and worker activities are within 5

feet of the emitting source

ܹேி

ft

(m)

10

(3.05)

ேிܪ

ft

(m)

6

(1.83)

Adequate height to capture a typical

worker’s breathing zone

ftଶ ܣܵܨ

(mଶሻ

340

(32) Equation [G‐16]

ݐଵ hr 0 Starting time for Equations [G‐14] and [G‐15]

ݐଶ hr 2

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)

ܩ g / min

(mg/hr)

5 to 50

(3E+5 to 3E+6)

No local exhaust ventilation (LEV; see Table

2‐8 in section 2.3.2 in the main document

0.5 to 5

(3E+4 to 3E+5)

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

Potential Workplace TCE Inhalation Exposures

(8‐hr TWA)

Near‐Field Far‐Field

LEV

(ppm)

No LEV

(ppm)

Number of

Workers

LEV

(ppm)

No LEV

(ppm)

Number of

Workers

Small commercial

degreasing facility

0.3  

(low‐end

estimate)  

20

(upper‐end

estimate)

3  

(low‐end

estimate)

197

(upper‐end

estimate)

8,730

0.04  

(low‐end

estimate)  

17

(upper‐end

estimate)

0.4  

(low‐end

estimate)  

172

(upper‐end

estimate)

20,952
Page 169 of 212

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‐1, 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‐1.  Model Inputs for Dry Cleaning Facilities

Parameter Units Parameter Values Comments

ܸிி

ftଷ

(mଷ)

7,062 to 70,620

(200 to 3,000)

Values supported by von Grote et al. (2006)

ܧܣܴ 1/hr 1 to 19

Values supported by von Grote et al. (2006) and EPA

(2013b)

ேிݒ

cm / s

(m/hr)

10

(360)

Values is ~50th percentile supported by  

Baldwin and Maynard (1998)

ேிܮ

ft

(m)

10

(3.05)

Assumes volatile source is centered in the near‐field and

worker activities are within 5 feet of the emitting source

ܹேி

ft

(m)

10

(3.05)

ேிܪ

ft

(m)

6

(1.83)

Adequate height to capture a typical worker’s breathing

zone

ftଶ ܣܵܨ

(mଶሻ

340

(32) See Equation [G‐16] in Appendix G

ݐଵ hr 0 Starting time for Equations [G‐14] and [G‐15] (see

equations in Appendix G)

ݐଶ hr 8

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)

ܩ g / min

(mg/hr)

0.037 to 0.37

(2.22E+3 to 2.22E+4)

No local exhaust ventilation (LEV); see Table 2‐11 in

section 2.4.2 in the main document

0.0037 to 0.037

(222 to 2.22E+3 )

With LEV; TCE emissions reduced by 90% (Wadden et

al., 1989)
Page 170 of 212

Table H‐2.  Potential Workplace TCE Inhalation Exposures and Number of Workers Exposed

Type of Facility

Potential Workplace TCE Inhalation Exposures

(8‐hr TWA)

Near‐Field Far‐Field

With LEV

(ppm)

No LEV

(ppm)

Number of

Workers

With LEV

(ppm)

No LEV

(ppm)

Number of

Workers

Dry Cleaning

Facility

0.008  

(low‐end

estimate)  

2

(upper‐end

estimate)

0.08  

(low‐end

estimate)

19

(upper‐end

estimate)

36,000

0.0007  

(low‐end

estimate)  

2

(upper‐end

estimate)

0.007  

(low‐end

estimate)  

18

(upper‐end

estimate)

252,000
Page 171 of 212

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.

I‐1 Default parameters used in CEM for emission and household

characteristics

Table I‐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 I‐1) (EPA,

1997, 2011a).
Page 172 of 212

Table I‐1.  Summary of CEM Parameters for Estimation of TCE Indoor Air Concentrations

S. No Modeling Input Value Justification/Source

1

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)

2

Overspray

fraction

(unitless)

0.01 Selection based on professional judgment (US EPA

2007a).

3 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.

4 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.

5 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.  

6 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)

7

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 173 of 212

Figure I‐1.  Screen capture of Summary of Recommended Values for Residential Building Parameters

from the Exposure Factors Handbook (EPA, 2011a)

I‐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.

I‐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 I‐2).
Page 174 of 212

Figure I‐2.  Screen capture of E‐FAST equations for estimation of emission rate

Using Equation 3‐42 to calculate EvapTime:  

ଵସହ 	 ൌ ܶ݅݉݁݌ܽݒܧ

ሺଵଷଵ.ସ	ൈ	଻ଷ.ସ଺ሻబ.వఱరల

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):

݇ ൌ	 ୪୬	ሺଵ଴ሻ

ଵ.ଷ଺

Hence, Emission Rate (k) = 101.06 hrs‐1 or 1.68 min‐1
Page 175 of 212

Because Chinn’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.  

I‐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 176 of 212

I‐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.

I‐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 I‐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 177 of 212

Table I‐2.  Comparison of Westat Survey Data and EPA’s TCE Degreaser Simulation Values

Westat data for solvent type

cleaning fluids or degreasers

EPA TCE Degreaser

Simulation values

Mean Median

(50th%) 90th%

Time spent using product 29.48 min 15.0 min 60.0 min 60 min

Time spent in room

after use

33.29 min 3.0 min 60.0  min 60 min

Amount of product

per event

9.45 oz

(268 g)

3.3 oz

(94 g)

16.0 oz

(454 g)

0.85 oz

(24 g)

Household location

Room of use

Garage 12.2%

Outside 28.0%

Inside room 49.1 %

Utility room

Westat data for brake cleaners/quieters Simulation values

Mean Median

(50th%) 90th%

Time spent using product 23.38 min 15.0 min 49.5 min 60 min

Time spent in room

after use

10.27 min 0.0 min 30.0 min 60 min

Amount of product

per event

6.26 oz

(177 g)

4.0 oz

(113 g)

12.0 oz

(340 g)

0.85 oz

(24 g)

Household location

Room of use

Garage 17.7%

Outside 77.1 %

Inside room 5.2 %

Utility room

I‐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 I‐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 178 of 212

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.

I‐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.

I‐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.

  

I‐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 179 of 212

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 I‐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 I‐3.  Screen Capture of MCCEM Model Evaluation Efforts
Page 180 of 212

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‐1). 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‐1.  Estimated TCE Potential Acute Dose Rates from the Residential Indoor Use of

Solvent Degreasers or Clear Protecting Coating Sprays

Age

(yrs)

Clear Protective Coating

Spray User  

ADRpot (mg/kg‐bw/day)

Clear Protective Coating

Spray Bystander  

ADRpot (mg/kg‐bw/day)

Solvent  

Degreaser User  

ADRpot

(mg/kg‐bw/day)

Solvent  

Degreaser Bystander

ADRpot (mg/kg‐

bw/day)

<1

NA

0.5

NA

3

1‐2 0.4 3

3‐5 0.4 2

6‐10 0.3 2

11‐15 0.2 1

16‐20 0.5 0.2 3 1

21‐78 0.4 0.1 3 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 x FQ x DEv x ED)

BW x AT

where:

ADRpot = 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 181 of 212

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:

   

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)]

Cair =                        (0.45 mg/kg‐bw/day)  (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

.
Page 182 of 212

Table J‐2.  Estimated TCE Inhalation Calculated Concentration in Air (Over Course of Day) from Use of

Two Consumer Products Indoors at Residences

Ages

Clear Protective Coating

Spray User  

ADRpot (ppm)

Clear Protective Coating

Spray Bystander

ADRpot (ppm)

Solvent

Degreaser User

ADRpot (ppm)

Solvent Degreaser

Bystander

ADRpot (ppm)

<1 NA 0.1 NA 0.8

1‐2 NA 0.1 NA 0.8

3‐5 NA 0.1 NA 0.8

6‐10 NA 0.1 NA 0.8

11‐15 NA 0.1 NA 0.8

16‐20 0.4 0.1 2 0.8

>21 0.4 0.1 2 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‐1 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 183 of 212

Appendix K RISK ASSESSMENT GUIDELINES, LITERATURE

SEARCH STRATEGY, STUDY SELECTION AND

DATA QUALITY CRITERIA FOR THE TCE IRIS

TOXICOLOGICAL REVIEW AND OPPT STUDY

REVIEW

K‐1 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 184 of 212

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‐1).  

•       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.
Page 185 of 212

Figure K‐1.  Study Quality Considerations for Epidemiological Studies
Page 186 of 212

Figure K‐ 2.  Study Quality Considerations for Animal Studies
Page 187 of 212

Appendix L LIST OF ORAL AND INHALATION STUDIES SUITABLE FOR NO

DOSE‐RESPONSE ANALYSIS IN THE TCE IRIS ASSESSMENT

L‐1 Liver Effects

Target

Organ Species Route of Exposure Range of Doses or Concentrations 1 Duration POD Type 2 Effect HEC50 (ppm) HEC95 (ppm) HEC99 (ppm) Uncer

Factors 

Liver

Mouse

(male) Inhalation 37 to 3,600 ppm

Continuous and

intermittent

exposures,

variable time

periods for

30‐120 days

BMDL10= 21.6

ppm

Increased

liver/body weight

ratio

25 12 9.1

UFS=1; 

UFH=3; 

Total 

Mouse

(male)

Oral

(gavage)

100 to 3,200

mg/kg‐bw/day 6 weeks BMDL10= 82 mg/kg‐bw/day 32 15 11 UFS=1; 

UFH=3; 

Total 

Rat

(female) Inhalation 100 to 1,000 ppm

6 hr/day, 5

days/week for

4 weeks

BMDL10= 25

ppm

53 24 19

UFS=1; 

UFH=3; 

Total 

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.
Page 188 of 212

L‐2 Kidney Effects

Target

Organ/

System

Species Route of Exposure Range of Doses or Concentrations 1 Duration POD Type 2 Effect HEC50 (ppm) HEC95 (ppm) HEC99 (ppm) Uncer

Factors 

Kidney

Rat

(female)

Oral

(gavage)

500 to 1,000

mg/kg‐bw/day

5 days/week for

104 weeks

BMDL05 = 9.45

mg/kg‐bw/day Toxic nephropathy 0.042 0.0085 0.0056

UFS=1; 

UFH=3; 

Total 

Rat

(female) Inhalation 100 to 1,000 ppm

6 hr/day, 5

days/week for 4

weeks

BMDL10= 15.7

ppm

Increased kidney

weight/body

weight ratio

0.099 0.020 0.013

UFS=1; 

UFH=3; 

Total 

Rat

(male)

Oral

(gavage)

50 to 250

mg/kg‐bw/day

4‐5 days/week

for 52 weeks

BMDL10 = 34

mg/kg‐bw/day

Pathology changes

in renal tubule 0.19 0.038 0.025

UFS=1; 

UFH=3; 

Total 

Rat

(male) Inhalation 100 to 600 ppm

7 hrs/day, 5

days/week for 2

years

BMDL10= 40.2

ppm

Pathology changes

in renal tubule 0.28 0.057 0.038

UFS=1; 

UFH=3; 

Total 

Mouse

(male) Inhalation 37 to 3,600 ppm

Continuous and

intermittent

exposures for

30‐120 days

BMDL10 = 34.7

ppm

Increased kidney

weight/body

weight ratio

0.88 0.18 0.12

UFS=1; 

UFH=3; 

Total 

Mouse

(female)

Oral

(gavage)

869 to 1,739

mg/kg‐bw/day

5 days/week,

TWA during

exposure period

(78 weeks),

animals

observed for 90

weeks

LOAEL = 620

mg/kg‐bw/day

Pathology changes

in renal tubule

(toxic nephrosis)

3.9 0.77 0.5

UFS=1; 

UFH=3; 

Total U

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.
Page 189 of 212

L‐3 Neurotoxicity

Target

Organ/

System

Species Route of Exposure Range of Doses or Concentrations 1 Duration POD Type 2 Effect HEC50 (ppm) HEC95 (ppm) HEC99 (ppm) Uncer

Factors 

Nervous

system

Rat

(male) Inhalation 50 to 300 ppm

8 hrs/day, 5

days/weeks for 6

weeks

LOAEL = 12 ppm

Significant

decreases in

wakefulness

13 6.4 4.8

UFS=3; 

UFH=3; 

Total U

Human

(both

sexes)

Inhalation

704 ppm × years of

exposure

(mean cumulative

exposure)

Mean of 16

years

LOAEL = 14 ppm

Trigeminal nerve

effects (increased

latency in

masseter reflex)

14 7.0 5.3

UFS=1; 

UFH=3; 

Total 

Rat

(male)

Oral

(drinking

water)

24 to 47

mg/kg‐bw/day 8 weeks LOAEL = 47 mg/kg‐bw/day Cognitive effects (demyelination of hippocampus) 18 9.2 7.1 UFS=10; 

UFH=3; 

Total U

Rat

(male)

Oral

(gavage)

1,000

mg/kg‐bw/day

5 days/week for

6 weeks

LOAEL = 710

mg/kg‐bw/day

Degeneration of

dopamine‐

containing

neurons in

substantia nigra

126 62 47

UFS=10; 

UFH=3; 

Total U

Rat

(female) Inhalation 300 ppm 24 hrs/day for 24 days LOAEL = 300 ppm Decreased

regeneration of

sciatic nerve

274 127 93

UFS=10; 

UFH=3; 

Total U

Mouse

(male) Inhalation 150 to 300 ppm 24 hrs/day for 24 days LOAEL = 150 ppm 378 163 120 UFS=10; 

UFH=3; 

Total U

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.
Page 190 of 212

L‐4 Immunotoxicity

Target

Organ/

System

Species Route of Exposure Range of Doses or Concentrations 1 Duration POD Type 2 Effect HEC50 (ppm) HEC95 (ppm) HEC99 (ppm) Uncer

Factors 

Immune

system

Mouse

(female)

Oral

(drinking

water)

1.4 to 14 ppm

(0.35 to 3.5

mg/kg‐bw/day)

27‐30 weeks LOAEL = 0.35 mg/kg‐bw/day

Decrease in

thymus weight

and thymus

cellularity

0.092 0.045 0.033

UFS=1; 

UFH=3; 

Total U

Mouse

(female)

Oral

(drinking

water)

1.4 to 14 ppm

(0.35 to 3.5

mg/kg‐bw/day)

27‐30 weeks LOAEL = 0.35 mg/kg‐bw/day

Autoimmunity

(increased anti‐

dsDNA and ssDNA

antibodies)

0.092 0.045 0.033

UFS=1; 

UFH=3; 

Total U

Mouse

(female)

Oral

(drinking

water)

18 to 660

mg/kg‐bw/day

16 or 24 weeks

(4 or 6 months)

LOAEL = 18

mg/kg‐bw/day

Immuno‐

suppression 4.8 2.4 1.7

UFS=1; 

UFH=3; 

Total U

Rat

(female) Inhalation 100 to 1,000 ppm

6 hrs/day, 5

days/week for 4

weeks

BMDL1SD= 24.9

ppm

Immuno‐

suppression 29 15 11

UFS=10; 

UFH=3; 

Total U

Mouse

(males;

auto‐

immune

prone

strain)

Inhalation 500 to 2,000 ppm

4 hrs/day, 6

days/week for 8

weeks

LOAEL = 70 ppm

Autoimmunity

(changes in

immunoreactive

organs)

97 48 37

UFS=10; 

UFH=1; 

Total U

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.4 Two different effects were reported by Keil et al, (2009): decreased thymic weight and cellularity and autoimmunity. A total UF of 100 was used 

whereas a total UF of 30 was used for the autoimmune effects. The TCE IRIS assessment allocated different LOAEL‐to‐NOAEL uncertainty factors 

of the effects, which resulted in different total UF (EPA, 2011e).
Page 191 of 212

L‐5 Reproductive Toxicity

Target

Organ/

System

Species Route of Exposure Range of Doses or Concentrations 1 Duration POD Type 2 Effect HEC50 (ppm) HEC95 (ppm) HEC99 (ppm) Unc

Facto

Reproductive

system

Human

(male) Inhalation 29.6 ppm (mean exposure)

Measured values

after an 8‐hr

work shift; mean

5.1 years on the

job

BMDL10 = 1.4

ppm

Decreased normal

sperm

morphology and

hyperzoospermia

1.4 0.7 0.5

UFS=1

UFH=

Tota

Rat

(male)

Oral

(drinking

water)

143 to 270

mg/kg‐bw/day 14 days LOAEL = 141 mg/kg‐bw/day Decreased in vitro fertilization 16 11 9.3 UFS=1

UFH=3

Total 

Rat

(male) Inhalation 376 ppm

4 hrs/day,

5 days/week,

2‐10 weeks

exposed,

2‐8 weeks

unexposed

LOAEL = 45 ppm

Sperm effects and

male reproductive

tract effects

32 16 13

UFS=1

UFH=3

Total 

4 hrs/day,

5 days/week for

12 or 24 weeks

Rat

(female

damns)

Oral

(gavage)

10.1 to 1,125

mg/kg‐bw/day

9 days (during

gestational days

6 to 15)

LOAEL = 475

mg/kg‐bw/day

Delayed

parturition 98 48 37

UFS=

UFH=3

Tota

Mouse

(male) Inhalation 1,000 ppm

6 hrs/day, 5

days/week for

19 days over 4

weeks

LOAEL = 180 ppm

Effects on

epididymis

epithelium

190 91 67

UFS=1

UFH=3

Total 

6 hrs/day, 5

days/week for

1‐4 weeks 

Mouse

(male) Inhalation 1,000 ppm

6 hrs/day, 5

days/week for

6 weeks

LOAEL = 180 ppm

Sperm effects

(decreased in vitro

sperm‐oocyte

binding and in vivo

fertilization)

190 91 67

UFS=1

UFH=3

Total 
Page 192 of 212

Target

Organ/

System

Species Route of Exposure Range of Doses or Concentrations 1 Duration POD Type 2 Effect HEC50 (ppm) HEC95 (ppm) HEC99 (ppm) Unc

Facto

Reproductive

system

Rat

(male)

Oral

(drinking

water)

72 to 389

mg/kg‐bw/day

Breeders

exposed 1 week

premating and

then for 13

weeks/ Pregnant

females exposed

throughout

gestation (i.e.,

18 weeks total)

LOAEL = 389

mg/kg‐bw/day

Decreased mating

(both sexes

exposed)

204 97 71

UFS=

UFH=3

U

Tota

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.
Page 193 of 212

L‐6 Developmental Toxicity

Target Organ/

System Species Route of Exposure Range of Doses or Concentrations 1 Duration POD Type 2 Effect HEC50 (ppm) HEC95 (ppm) HEC99 (ppm) Un

Fact

Developmental

effects

PRE‐ AND POSTNATAL MORTALITY

Rat

(female) Inhalation 1. ppm

4 hrs/day,

gestational days

8 to 21 LOAEL =17 ppm Increased resorptions 16 8 6.2 UFS=

UFH=

Tot

Rat

(female)

Oral

(gavage)

10.1 to 1,125

mg/kg‐bw/day

Gestational days

6 to 15

BMDL01= 32.2

mg/kg‐bw/day

Increased

resorptions 57 29 23

UFS=

UFH

Tot

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=

UFH=

Tot

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=

UFH

Tot

DEVELOPMENTAL NEUROTOXICITY

Rat

(male

pups)

Oral

(gavage)

50 to 290

mg/kg‐bw/day

Postnatal days

10 to 16

LOAEL = 50

mg/kg‐bw/day

Decreased rearing

activity 8 4 3

UFS=

UFH=

Tot

Rat

(female)

Oral

(drinking

water)

45 to 140 mg/kg‐

bw/day

Dams and pups

exposed from 14

days prior to

mating until end

of lactation

LOAEL = 45

mg/kg‐bw/day

Increased

exploratory

behavior in male

pups (offspring)

22 11 8.4

UFS=

UFH=

Tot

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.
Page 194 of 212

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‐1.  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‐1.  Neurotoxicological Endpoints Selected for Dose‐Response Modeling

                      (Blossom et al., 2013)

TCE in drinking water Sample

size1

Figure 7A,  

novel object (inverted

wire cup)

Figure 7B,  

novel mouse, placed under

inverted wire cup

mg/ml mg/kg‐day Mean SD Mean SD

0 0 9 5.27 5.78 5.32 3.96

0.01 2 8 4.36 5.16 7.85 6.88

0.1 28 8 10.54 4.30 10.19 7.46

Notes:

1

   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: 1_Blossom_NObject_SD.xlsm and 2_Blossom_NMouse_SD.xlsm).  The

BMDL1SD ranged from 14‐20 mg/kg/day depending upon the neurobehavioral endpoint

(Endpoint A or Endpoint B).  
Page 195 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 BMD1SDa

(mg/kg‐d)

BMDL1SDa

(mg/kg‐d) Basis for Model Selection

p‐value AIC

Endpoint A  

(novel object)

       Only the linear model provided an

adequate fit.   

Linear b 0.577 110.05 23.7 14.3

Endpoint B  

(novel mouse)

       Only the linear model provided an

adequate fit.    Linear c 0.428 119.67 41.9 19.8

Notes:

a

   BMR is 1 SD change from the control mean            b   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‐1.  Linear Dose‐Response Model for Endpoint A:  Number of Times Entered Central

Zone With Novel Object

 2

 4

 6

 8

 10

 12

 14

 16

 0 5 10 15 20 25 30 35 40

Mean Response

dose

Linear Model, with BMR of 1 Std. Dev. for the BMD and 0.95 Lower Confidence Limit for the BMDL

09:14 02/21 2014

BMDL BMD

 

Linear
Page 196 of 212

Figure M‐2.  Linear Dose‐Response Model for Endpoint B:  Number of Times Entered Central

Zone with Novel Mouse

*Note that BMD lies above highest observed response.

 0

 2

 4

 6

 8

 10

 12

 14

 0 5 10 15 20 25

Mean Response

dose

Linear Model, with BMR of 1 Std. Dev. for the BMD and 0.95 Lower Confidence Limit for the BMDL

09:08 02/21 2014

BMDL BMD

 

Linear
Page 197 of 212

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 198 of 212

Key Factor a,b Type of

evidence to

consider

Data Evidence for stronger

weight of association

Evidence for weaker

weight of association

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 functionalalterations.

 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 et al., 1993;

Johnson et al., 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 et al., 1998a). Gavage

administration of TCE metabolites

(DCA and TCA) on GD 6‐15 (Smith et

al., 1989, 1992) or of DCA duringdiscrete windows of time within GD 6‐

15 (Epstein et al., 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 et al., 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 et al., 2006) or to rats

and mice on GD 6‐15 (Schwetz et al.,1975) did not result in treatment‐related cardiac defects.
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

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 et al., 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).

Exposureoccurs beforeoutcomesonset

Epi  Four cohort or case‐control studies

consider temporality (Forand et al.,

2012; Goldberg et al., 1990; Ruckart etal., 2013; Yauck et al., 2004). Threestudies observe an association

between the TCE exposure surrogate

and major cardiac defects (Forand et

al., 2012; Goldberg et al., 1990; Yaucket al., 2004). An association withconotruncal defects, specifically,

observed in Forand et al. (2012).

 Temporality was not considered in

Bove (1996); Bove et al. (1995);

Goldberg et al. (1990); and Lagakos etal. (1986)

Strength ofassociation

Study quality,including study

strengths and

limitations

Tox  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

 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 useof tap water as the vehicle for some of
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

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 bythe use of historical controls that did

not demonstrate a temporal shift incardiac 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

et al. (2003) there was a highlysignificant 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

(Fisher et al., 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 et al. (1993) indicated

that levels of TCE in dose formulations

were tested by gas chromatography,

the analytical findings were not

reported. Johnson et al. (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 et al. (1993) and Johnson

et al. (2003) studies estimated dosesbased on the average water

consumption. This method does not

provide precise information tocalculate TCE dose because variability

in drinking water consumption among

dams is not characterized.

 The dose selection for Johnson et al.

(2003) resulted in a NOAEL that isapproximately 700‐fold lower than thenext highest dose.

 Some studies that did not identify

treatment‐related cardiac defectsfollowing developmental exposures to

TCE (e.g.,Carney et al., 2006; Fisher et
Page 201 of 212

Key Factor a,b Type of

evidence to

consider

Data Evidence for stronger

weight of association

Evidence for weaker

weight of association

al., 2001; Schwetz et al., 1975) werewell‐conducted and adequately‐

reported GLP and/or guideline studies

with no substantive limitationsidentified.

 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 theassessment team, yet found no

treatment‐related cardiac defects.

Magnitude ofthe effectmeasure

Epi  Increased risk estimates between all or

major cardiac defects ranged from

1.24 (95% CI: 0.75, 1.94) to 2.40 (95%

CI: 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 anincreased risk estimate of 6.2 (95% CI:

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.
Page 202 of 212

Key Factor a,b Type of

evidence to

consider

Data Evidence for stronger

weight of association

Evidence for weaker

weight of association

Variabilityanalysis

Sources of

within‐ andcross‐study

variability that

contribute touncertainty

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 Johnsonet al. (2003). Fetal cardiac evaluationmethodology, 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 ofmalformations.

 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 cardiacdefects 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 definitivelylink 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 et al. (2001) did not. In the

Fisher et al. (2001) study, care wastaken to ensure that the same cardiac

evaluation methods were used as in

the Dawson et al. (1993) and Johnson

et al. (2003) studies, including fetalevaluation with knowledge oftreatment 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 andcontrol for Johnson et al. (2003) andDawson et al. (1993) studies.

 The Johnson et al. (2003) and Dawson
Page 203 of 212

Key Factor a,b Type of

evidence to

consider

Data Evidence for stronger

weight of association

Evidence for weaker

weight of association

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.

et al. (1993) studies did not calculate

variability in TCE dose by measuring

individual dam water consumption.

Sources of

within‐ andcross‐study

variability that

contribute touncertainty

Epi  NE (not considered in Hill analysis)  NE (not considered in Hill analysis) 

Uncertaintyanalysis

Missinginformation ordata gaps,within and

across studies

Tox  For the studies conducted by the UAZ

laboratory that identified cardiacdefects following exposures to TCE,

DCA, or TCA (Dawson et al., 1993;

Johnson et al., 1998a; Johnson et al.,2003), detailed descriptions ofevaluation 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, wereprovided to the EPA for Dawson et al.

(1993) and Johnson et al. (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 studydetails, 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 et al., 2006; Fisher

et al., 2001) adequate descriptions ofstudy methodology and summary data
Page 204 of 212

Key Factor a,b Type of

evidence to

consider

Data Evidence for stronger

weight of association

Evidence for weaker

weight of association

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.

Missinginformation ordata gaps,within and

across studies

Epi  NE (not considered in Hill analysis)  NE (not considered in Hill analysis)

Qualitativedose‐response

Associationbetweenexposure/doseand degree ofeffect

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 et al., 1989, 1992).

 The incidence of cardiovascular effects

increased as a function of dose in

Johnson et al. (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‐1 mice, chicken embryos, and

zebra fish (Dawson et al., 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 et al., 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 et al. (2003)

could only be fit to a model withelimination of the high dose data from

the analysis. The lowest dose tested

had a zero response for cardiac

defects, below the historical controlincidence. 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.
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

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.

Exposure‐responsegradient:Associationbetweenexposure/doseand degree ofeffect

Epi  NE  Goldberg et al. (1990) and Lagakos et

al. (1986) examined exposure‐response; none observed.

Experimentalevidence

Hypothesistesting:manipulationof exposure

scenario withresultingalterations inresponse

Tox  A study by Epstein et al. (1992)

administered the metabolite DCA to

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.

 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 TCEvia 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 et al., 1996),

that identified cardiac defects, the acid

nature of DCA and TCA may have

impacted dysmorphogenesis.
Page 206 of 212

Key Factor a,b Type of

evidence to

consider

Data Evidence for stronger

weight of association

Evidence for weaker

weight of association

Association not

observed once

exposureceases

Epi  NE  No differences between observed and

expected numbers of cardiac defect

cases once wells were closed in

contaminated area (Goldberg et al.,

1990).

Reproducibility[Consistency]

Reproducibility:Corroborationacross 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 identifiedtreatment and dose‐related cardiac

malformations (Dawson et al., 1993;

Johnson et al., 1998a; Johnson et al.,2003). A preliminary screening studythat 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 theresearch program.

 Studies on TCE metabolites (TCA and

TCA) conducted in other laboratories

(Epstein et al., 1992; Smith et al.,

1989, 1992) identified cardiac defectssimilar 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

 Studies conducted in other

laboratories than UAZ and that

administered TCE by gavage or

inhalation (Carney et al., 2006; Fisher

et al., 2001; Schwetz et al., 1975) didnot identify statistically significant

increases in cardiac defects. Fisher et

al. (2001) used the same cardiacevaluation methods as the UAZ lab.
Page 207 of 212

Key Factor a,b Type of

evidence to

consider

Data Evidence for stronger

weight of association

Evidence for weaker

weight of association

al., 1996; Mishima et al., 2006) withTCE and/or metabolite exposures.

Zebrafish studies also demonstrated

evidence of alterations in cardiac

development (Hassoun et al., 2005;

Williams et al., 2006).

Consistency:Associationobserved in

differentpopulations,places, time

andcircumstances.

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). Twoother studies of weaker designs were

of different populations and carried

out in two different locations in the

United States, and provide supporting

evidence (Goldberg et al., 1990; Yauck

et al., 2004).

 Lagakos et al. (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.

Biologicalplausibility

Observedoutcome can

be attributed

to toxic insult

given the

known science

Tox  Avian in ovo studies and

atrioventricular cell culture studiessupport 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 mechanisticdata are needed to support the

hypothesized AOP.

 There are insufficient mechanistic data

to characterize additional potential

MOAs other than that hypothesized in

the AOP.
Page 208 of 212

Key Factor a,b Type of

evidence to

consider

Data Evidence for stronger

weight of association

Evidence for weaker

weight of association

model for TCE‐induced congenital

heart defects. In this construct, the

vulnerable period is defined by

endocardial morphogenesis.Endothelial–mesenchyme transition is

disrupted in the area of theatrioventricular canal, leading toseptal 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.

Observedassociationplausible given

the known

science

Epi  NE  NE 

Alternative ormultipleexplanations

Other possibleexplanationsfor observed

outcome after

the exposure of

interest

Tox  Given the presumed contribution of

both environmental exposures andgenetic predisposition in humancongenital heart disease (Richards and

Garg, 2010), it is possible that the testsubjects used in the Johnson et al.(2003) study and others conducted inthat laboratory may have been

 There is a possibility that cardiac

defects detected in the Dawson et al.

(1993) study were associated in partwith the use of tap water as a control

vehicle (i.e., possible presence of

contaminants).
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Data Evidence for stronger

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Evidence for weaker

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 inwell‐conducted TCE studies (Carney et

al., 2006; Fisher et al., 2001) ormetabolite studies (Fisher et al., 2001)was due to confounding variables such

as differences in strain/source of

animal model, route of exposure,

toxicokinetics, vehicle [e.g., soybeanoil in Fisher et al. (2001)], or

differences in cardiac evaluationmethods.

 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.

Other possibleexplanationsfor observed

outcome after

the exposure of

interest (not

considered in

Epi  Potential maternal risk factors were

adjusted in statistical analysis in

Forand et al. (2012) and Yauck et al.

(2004) or were not found in statisticalanalyses to influence observed

association by +15% (Bove, 1996;

Bove et al., 1995).

 Potential for confounding from

another exposure given the poor

exposure definition in Yauck et al.

(2004). The positive association inGoldberg et al. (1990) may result fromlikely selection biases in controls.
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Data Evidence for stronger

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weight of association

Hill analysis)

Specificity Single cause

and effectrelationshipresulting from

exposure totest 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 et al. (2003)

reported that TCE exposure via

drinking water to pregnant rats did not

result in maternal toxicity. Carney et

al. (2006) reported minimal decreasesin 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 et al. (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

et al., 2001; Schwetz et al., 1975) didnot identify cardiac defects. Fisher etal. (2001) used the same cardiacevaluation methods as the UAZ lab.

 The cardiac defects detected in the

Dawson et al. (1993) study may havebeen related to the use of tap water as

a vehicle (i.e., possible contaminants).
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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

Single causeand effectrelationshipresulting from

exposure totest substance

Epi  NE  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 cardiacdefects in infants have not been well

studied.

Coherence Summary:

Extent to which

data are similar

in outcome and

exposureacrossdatabase

Tox  Multiple studies were conducted at

UAZ (Dawson et al., 1993; Johnson et

al., 1998a; Johnson et al., 2003), inwhich 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. Theoutcomes 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.

 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).

Cause andeffectinterpretationshould not

conflict with

the generally

known facts of

the natural

Epi  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).

 NE
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Data Evidence for stronger

weight of association

Evidence for weaker

weight of association

history andbiology of thedisease

NE = No relevant evidence.HH = Hamburger‐Hamilton stages of chick development (Hamburger and Hamilton, 1951).Tox = Animal toxicology studies; Epi = Epidemiological studiesKey Factor references:

a EPA (2006b)

b Hill (1965)