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EPA Document# EPA-740-P-23-002
February 2023
Office of Chemical Safety and
Pollution Prevention

x>EPA

United States

Environmental Protection Agency

Draft Proposed Approach for Cumulative Risk Assessment of
High-Priority Phthalates and a Manufacturer-Requested
Phthalate under the Toxic Substances Control Act

February 2023


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l TABLE OF CONTENTS

2	LIST OF TABLES	5

3	LIST OF FIGURES	6

4	LIST OF EQUATIONS	6

5	LIST OF APPENDIX TABLES	7

6	LIST OF APPENDIX FIGURES	7

7	ACKNOWLEDGEMENTS	8

8	ABBREVIATIONS AND ACRONYMS	9

9	EXECUTIVE SUMMARY	11

10	1 BACKGROUND	16

11	1.1 What Is EPA Proposing in this Work?	17

12	2 KEY CONCEPTS AND PROPOSED CONCEPTUAL MODEL	20

13	2.1 Key Concepts	20

14	2.2 Proposed Conceptual Model	21

15	3 CONSIDERATIONS FOR GROUPING PHTHALATES FOR CRA: STEP 1 IN

16	CONCEPTUAL MODEL (FIGURE 2-1)	26

17	3.1 Evidence of Toxicologic Similarity	26

18	3,1.1 Phthalate Syndrome Mode of Action (MOA)	28

19	3,1,2 Key Outcomes for Grouping High-Priority and Manufacturer-Requested Phthalates for

20	CRA	31

21	3.1.2.1 Study Selection Strategy	33

22	3.1.2.2 Availability of Studies to Inform Key Outcomes	34

23	3.1.3 Key Outcomes Data	35

24	3.1.3.1 Fetal Testicular Gene Expression	35

25	3.1.3.1.1 Cholesterol Transport and Steroidogenesis	35

26	3.1.3.1.2 Insl3 mRNA Expression	37

27	3.1.3.2 Fetal Testicular Testosterone	42

28	3.1.3.3 Anogenital Distance (AGD)	51

29	3.1.3.4 Nipple Retention	59

30	3.1.3.5 Hypospadias	64

31	3.1.3.6 Seminiferous Tubule Atrophy	69

32	3.1.3.7 Multinucleated Gonocyte (MNG) Formation	74

33	3.1.4 Phthalate Syndrome in Humans	77

34	3.1.4.1 Human Explant and Xenograft Studies	77

35	3.1.4.2 Epidemiologic Studies	79

36	3,1.5 Species Differences in Sensitivity	81

37	3.1.5.1 Species Difference in Metabolism and Toxicokinetics	83

38	3.1.6 Data Integration and Weight of Evidence Analysis	84

39	3.1.6.1 Temporal Concordance	85

40	3.1.6.2 Dose-Response Concordance	86

41	3.1.6.3 Strength, Consistency, and Specificity	88

42	3.1.6.4 Biological Plausibility and Coherence	90

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3.1.6.5 Uncertainties	90

3.1,7 Proposed Conclusions on Toxicologic Similarity	91

3.2	Evidence of Co-exposure over a Relevant Timeframe	92

3.3	Proposed Cumulative Chemical Group (Step 1 in Conceptual Model [Figure 2-1])	94

4	PROPOSED OPTIONS FOR ADDRESSING PHTHALATE SYNDROME	96

4.1	Addressing Phthalate Syndrome as a Whole Versus Focusing on the Most Sensitive Effect	96

4.1.1	Addressing Phthalate Syndrome as a Whole	96

4.1.2	Focusing on the Most Sensitive Effect	97

4.1.3	EPA's Proposed Approach for Addressing Phthalate Syndrome	98

4.2	Applicability of Dose Addition for Phthalates	98

4.3	Approaches Based on Dose Addition	100

4.3.1	Hazard Index Approach	100

4.3.2	Relative Potency Factor Approach	101

4.3.3	Proposed Risk Characterization Approach for Phthalates under TSCA	101

4.4	Proposed Options for Deriving Relative Potency Factors	102

4.4.1	Strengths and Uncertainties of Key Outcomes Datasets for RPF Derivation	102

4.4.1.1	Decreased Fetal Testicular Testosterone Production	102

4.4.1.2	Decreased Fetal Testicular Expression of Cholesterol Transport and Steroidogenesis
Genes	103

4.4.1.3	Decreased Anogenital Di stance	103

4.4.1.4	Nipple/Areolae Retention	104

4.4.1.5	Seminiferous Tubule Atrophy	104

4.4.1.6	Hypospadias	105

4.4.1.7	Incidence of MNGs	105

4.4.2	Proposed Options for Deriving RPFs	105

5	PROPOSED POPULATIONS CONSIDERED: STEP 2 IN CONCEPTUAL MODEL
(FIGURE 2-1)	108

6	PROPOSED EXPOSURE AND RISK APPROACH FOR ASSESSING PHTHALATES
FOR CUMULATIVE RISK UNDER TSCA: STEPS 3 TO 10 IN CONCEPTUAL MODEL
(FIGURE 2-1)	110

6.1	Overview	110

6.2	Summary of COUs and Pathways for Phthalates from Individual Scope Documents	110

6.2.1	Conditions of Use Listed in Final Scopes for Individual Phthalate Risk Evaluations (Step

3 in Conceptual Model [Figure 2-1])	110

6.2.2	Pathways and Routes of Exposure Considered in Risk Evaluation as Stated in Final
Phthalate Scopes	114

6.3	Scenario-Building for Pathways of Exposure (Steps 4 and 5 in Conceptual Model)	115

6.3.1	TSCA COUs (Step 4 in Conceptual Model [Figure 2-1])	115

6.3.2	Estimating Non-attributable and Non-TSCA Exposures (Step 5 in Conceptual Model
[Figure 2-1])	115

6.3.2.1	Scenario-Based Exposure Evaluation for Estimating Non-attributable and Non-TSCA
Exposures	117

6.3.2.2	Reverse dosimetry and Biomonitoring Approach for Estimating Non-attributable
Exposure	119

6.3.2.3	Comparison of Reverse Dosimetry and Scenario-Based Approaches	124

6.3.2.4	Uncertainties and Limitations of Approaches	125

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6.3.2.5 Proposed Approach for Estimating Exposure from Non-attributable and Non-TSCA

Sources	128

6.4 Combining Exposure and Estimating Cumulative Risk (Steps 6 to 10 in Conceptual Model

[Figure 2-1])	129

6.4.1	Consumer Exposures and Risk	129

6.4.1.1	Data Needs for Consumer Co-exposure Analysis	130

6.4.1.2	Co-exposure Resulting from TSCA Consumer COUs (Step 7 in Conceptual Model
[Figure 2-1])	131

6.4.1.2.1 Survey of Consumer Behavior for Determining Co-exposure	131

6.4.1.2.1	Purchase Data for Determining Co-exposure	131

6.4.1.2.2	Product Formulation Data for Determining Co-exposure	131

6.4.1.3	Combining Exposure to Consumers to Estimate Cumulative Exposure (Steps 8 and 9

in Conceptual Model [Figure 2-1])	134

6.4.1.4	Estimating Cumulative Risk for Consumers (Steps 10 in Conceptual Model [Figure
2-1])	135

6.4.2	Occupational Exposures and Risk	135

6.4.2.1	Data Needs for Releases and Cumulative Occupational Exposure Assessment	136

6.4.2.2	Co-exposure Resulting from TSCA Occupational COUs (Step 7 in Conceptual Model
[Figure 2-1])	138

6.4.2.2.1	EPA Program Data for Identifying Sites to Determine Co-exposure	138

6.4.2.2.2	NIOSH HHE, OSHA CEHD, and Other Literature Sources Data for Identifying
Sites to Determine Co-exposure	139

6.4.2.2.3	Product Information Data for Identifying Sites to Determine Co-exposure	139

6.4.2.2.4	Identifying Additional Unknown Sites with Release and Exposure Potential to
Determine Co-exposure	139

6.4.2.2.5	Workplace Monitoring Data for Determining Co-Exposure	141

6.4.2.3	Combining Exposure to Workers to Estimate Cumulative Risk (Steps 8 and 9 in
Conceptual Model [Figure 2-1])	141

6.4.2.4	Estimating Cumulative Risk for Workers (Step 10 in Conceptual Model [Figure 2-l])142

6.4.3	General Population (Fenceline Communities) Exposures and Risk	142

6.4.3.1	Data Needs for General Population/Fenceline Community Exposure Assessment	143

6.4.3.2	Co-exposure Resulting from TSCA COUs (Step 7 in Conceptual Model [Figure 2-l])144

6.4.3.2.1	Using Reported Release Data to Determine Co-exposure	144

6.4.3.2.2	Surrogate Release Data for Determining Co-exposure	145

6.4.3.3	Combining Exposure to General Population (Fenceline Communities) to Estimate
Cumulative Risk (Steps 8 and 9 in Conceptual Model [Figure 2-1])	145

7 SUMMARY OF PROPOSED APPROACH AND NEXT STEPS	149

REFERENCES	152

APPENDICES	170

Appendix A Phthalate Cumulative Risk Assessment Initiatives	170

A,1	United States Consumer Product Safety Commission	170

A.2	Health Canada	173

A.3	Danish EPA	174

A.4	Australia NICNAS	175

A,5	European Food Safety Authority	177

A.6	PODs Used in Previous Phthalate CRAs	178

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Appendix B Additional Toxicity Information	181

B.l	Dose-Response Data for Effects on Fetal Testicular Gene Expression and Testosterone

Production	181

B.2	DEHP Study Summaries	182

B,3	BBP Study Summaries	187

B.4	DBP Study Summaries	189

B,5	DIBP Study Summaries	195

B.6	DC HP Study Summaries	197

B,7	DINP Study Summaries	198

B.8	DIDP Study Summaries	200

Appendix C Methodology for Preliminary Dose-Response Modeling	202

C.l	General Approach	202

C,2 Anogenital Distance (AGD)	202

C.2.1 Calculation of Individual Phthalate Ester AGD Dose-Response Models	202

C.3 Nipple/Areolae Retention in 13 to 14 Day Old Infant Male Rats	203

C.3.1 Calculation of Individual Phthalate Ester Dose-Response Models	203

C.4 Testicular Pathology - Seminiferous Tubule Atrophy	203

C.4.1 Calculation of Individual Phthalate Ester Dose-Response Models	203

C.5 Hypospadias	203

Appendix D Occupational Exposure Assessment	205

Appendix E Glossary of Key Terms	206

LIST OF TABLES

Table 3-1. Summary of Critical Effects Selected for Use in Previous Phthalate CRAs	33

Table 3-2. Summary of Studies Supporting the Proposed Key Outcomes	34

Table 3-3. Studies Evaluating Fetal Testicular Steroidogenic Gene and Ins 3 mRNA Expression	38

Table 3-4. ED50 Values (mg/kg/day) for Reduced mRNA Expression of Steroidogenic Genes and Insl3

	41

Table 3-5. Studies Evaluating Fetal Testicular Testosterone	44

Table 3-6. ED50 Values for Reduced ex vivo Fetal Testicular Testosterone Production	49

Table 3-7. Summary of NASEM (2017) Systematic Review and Meta-Analysis Results for Effects on

Fetal Testosterone	49

Table 3-8. Studies Evaluating Anogenital Distance in Male Pups	53

Table 3-9. Summary of ED50 Values for Reduced (% Control) Male AGD	58

Table 3-10. Summary of NASEM (2017) Systematic Review and Meta-Analysis Results for Effects on

AGD	58

Table 3-11. Studies Evaluating Nipple Retention in Male Pups	61

Table 3-12. Summary of ED50 Values for Percent Males per Litter with Retained Nipples/Areolas	64

Table 3-13. Studies Evaluating Incidence of Hypospadias	66

Table 3-14. Summary of ED50 Values for Hypospadias	69

Table 3-15. Studies Reporting Seminiferous Tubule Atrophy	71

Table 3-16. Summary of ED50 Values for Incidence of Seminiferous Tubule Atrophy	74

Table 3-17. Studies Reporting on the Incidence of MNGs	75

Table 3-18. Summary of NASEM (2017) Systematic Review and Meta-Analysis for Epidemiologic

Studies of AGD	80

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Table 3-19. Summary of Epidemiologic Evidence of Male Reproductive Effects Associated with

Phthalates	81

Table 3-20. Comparison of Lipase Activity across Species	84

Table 3-21. Comparison of RatED50 Values (mg/kg/day) across Key Outcomes	87

Table 3-22. Summary of Phthalate Syndrome-Related Effects Observed in Rat Studies	89

Table 3-23. Summary of Phthalate Metabolite Detection Frequencies in NHANES	93

Table 3-24. Summary of Information Supporting EPA's Proposed Cumulative Chemical Group for CRA

under TSCA	95

Table 6-1. Categories of Conditions of Use for High-Priority Phthalates and a Manufacturer-Requested

Phthalate	Ill

Table 6-2. Urinary Phthalate Metabolites Included in NHANES	120

Table 6-3. Summary of Studies Providing Estimates of the Urinary Excretion Fractions (F ue )of

Phthalate Metabolites	123

Table 6-4. U.S. CPSC Estimated Median and 95th Percentile Phthalate Daily Intake Values for Women

of Reproductive Age	125

Table 6-5. Summary of Uncertainties and Limitations Associated with Use of Scenario-Based and

Reverse Dosimetry Approaches	126

Table 6-6. Sample of Consumer Products Containing Phthalates	132

Table 6-7. Available EPA Program and Common Source Data for Each Phthalate	137

Table 6-8: Conditions of Use for Each High-Priority and Manufacturer-Requested Phthalate	140

Table 6-9. Media of Release Covered by EPA Programs	145

LIST OF FIGURES

Figure 2-1. Cumulative Risk Assessment Conceptual Model	25

Figure 3-1. Decision Tree for Grouping Phthalates for CRA	26

Figure 3-2. Chemical Structures of Phthalates Being Evaluated under TSCA	29

Figure 3-3. Hypothesized Phthalate Syndrome Mode of Action Following Gestational Exposure	30

Figure 3-4. Testicular Steroidogenesis Pathway	35

Figure 3-5. CYP11A and StAR mRNA Expression in SD and Wistar Rats	36

Figure 3-6. Insl3 mRNA Expression in SD and Wistar Rats	37

Figure 4-1. Proposed Severity Classifications for Phthalate Syndrome-Related Outcomes (from

Blessinger et al. (2020))	 97

Figure 5-1. Diagram of Initial Proposed Populations Identified Based on Susceptibility to Phthalate

Syndrome	109

Figure 6-1. Scenario-Based and Reverse Dosimetry Approaches for Estimating Non-attributable and

Non-TSCA Exposure	116

Figure 6-2. Diagram of Hypothetical NHANES Population Distribution of Phthalates and Illustration of
Assumptions about Exposure Profiles	121

LIST OF EQUATIONS

Equation 4-1. Calculating the hazard index	100

Equation 4-2. Calculating RPFs	101

Equation 4-3. Calculating index chemical equivalents	101

Equation 6-1. Calculating a phthalate daily intake value from urinary biomonitoring data	121

Equation 6-2. Example estimation of cumulative phthalate exposure to consumers	135

Equation 6-3. Example estimation of cumulative exposure to occupational subpopulations	142

Equation 6-4. Example estimation of cumulative exposure from single facility releases to air	146

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227	Equation 6-5. Example estimation of phthalate cumulative exposure from single facility releases to

228	surface water	146

229	Equation 6-6. Example estimation of cumulative exposure from multiple facility releases to air	146

230	Equation 6-7. Example estimation of cumulative exposure from multiple facility releases to surface

231	water	146

232	Equation 6-8. Example estimation of cumulative exposure to fenceline populations who are not

233	consumers or workers	147

234	Equation 6-9. Example estimation of cumulative exposure to fenceline populations who are also

235	consumers and workers	147

236

237	LIST of appendix tables

238	Table_Apx A-l. Summary of Phthalates Included in Previous CRAs	170

239	Table_Apx A-2. Summary of Australia NICNAS Cumulative Phthalate Assessments	176

240	Table Apx A-3. Summary of PODs for High-Priority and Manufacturer-Requested Phthalates

241	Considered in Previous CRAs	178

242

243	LIST of appendix figures

244	FigureApx A-l. Estimated Phthalate Exposure by Individual Exposure Scenario for Women	172

245	Figure_Apx B-l. Dose-Response Data from Gray et al. (2021)	 181

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ACKNOWLEDGEMENTS

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).

Acknowledgements

The OPPT Assessment Team gratefully acknowledges participation or input from intra-agency
reviewers that included multiple offices within EPA and assistance from EPA contractors ERG
(Contract No. 68HERD20A0002) and ICF (Contract No. 68HERD22A0001). EPA also acknowledges
the contributions of technical experts from EPA's Office of Research and Development.

Docket

Supporting information can be found in public docket, Docket ID: EPA-HQ-OPPT-2022-0918

(https://www.reeiilations.eov/dociiment/EPA-HQ-OPPT-2022- '001)

Disclaimer

Reference herein to any specific commercial products, process, or service by trade name, trademark,
manufacturer or otherwise does not constitute or imply its endorsement, recommendation, or favoring by
the United States Government.

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ABBREVIATIONS AND ACRONYMS

AGD

Anogenital distance

ATSDR

Agency for Toxic Substances and Disease Registry

BBP

Butyl benzyl phthalate

BMD

Benchmark dose

CASRN

Chemical Abstracts Service Registry Number

CDC

U.S. Centers for Disease Control and Prevention

CDR

Chemical Data Reporting (Database)

CHAP

Chronic Hazard Advisory Panel

CI

Confidence interval

COU

Condition of use

CPHEA

Center for Public Health and Environmental Assessment

CPSC

U.S. Consumer Product Safety Commission

CRA

Cumulative risk assessment

DBP

Dibutyl phthalate

DCHP

Dicyclohexyl phthalate

DEHP

Di-ethylhexyl phthalate

DEP

Diethyl phthalate

DHT

Dihydrotestosterone

DI

Dietary intake

DIBP

Di-isobutyl phthalate

DIDP

Di-isodecyl phthalate

DINP

Di-isononyl phthalate

DMP

Dimethyl phthalate

DMR

Discharge Monitoring Report

DNEL

Derived no effect level

DPP

Dipentyl phthalate

ECHA

European Chemicals Agency

ECRAD

Existing Chemical Risk Assessment Division

ED50

Effective dose (causing a 50 percent response)

EFSA

European Food Safety Authority

EPA

U.S. Environmental Protection Agency

ESD

Emission Scenario Documents

EU

European Union

FDA

U.S. Food and Drug Administration

FHSA

Federal Hazardous Substances Act

GD

Gestational day

GS

Generic scenario

HI

Hazard index

HQ

Hazard quotient

IC

Index chemical

INSL3

Insulin-like Growth factor 3

IRIS

Integrated Risk Information System

LABC

Levator Ani/bulbocavernosus

LOAEL

Lowest-observed-adverse-effect-level

LOEL

Lowest-ob served-effect-level

MBP

Monobutyl phthalate

MEHP

Mono-2-ethylhexyl phthalate

MIE

Molecular initiating event

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MNG

Multinucleated gonocyte

MOA

Mode of action

MOE

Margin of exposure

NAS

National Academy of Sciences (now National Academies of Sciences, Engineering, and



Medicine [NASEM])

NHANES

National Health and Nutrition Evaluation Surveys

NEI

National Emissions Inventory

NICNAS

National Industrial Chemicals Notification and Assessment Scheme

NIOSHHHE

National Institute of Occupational Safety and Health: Health Hazard Evaluation

NPDES

National Pollutant Discharge Elimination System

NR

Nipple retention

NRC

National Research Council (now NASEM)

NTP

National Toxicology Program

OCSPP

EPA's Office of Chemical Safety and Pollution Prevention

OECD

Organisation for Economic Co-operation and Development

OHAT

NTP's Office of Health Assessment and Translation

OLEM

EPA's Office of Land and Emergency Management

OPP

EPA' Office of Pesticide Programs

OPPT

EPA's Office of Pollution Prevention and Toxics

ONU

Occupational non-user

ORD

EPA's Office of Research and Development

OSHA CEHD

Occupational Safety and Health Administration: Chemical Exposure Health Data

PESS

Potentially exposed or susceptible subpopulations

PND

Postnatal day

PNW

Postnatal week

POD

Point of departure

POTW

Publicly owned treatment work

PPS

Preputial separation

RCR

Risk characterization ratio

RCRA

Resource Conservation and Recovery Act

RPF

Relative potency factor

RfV

Reference Value

SACC

Science Advisory Committee on Chemicals

SAR

Structure-activity relationship

SD

Sprague-Dawley (rats)

SDS

Safety Data Sheet

StAR

Steroidogenic acute regulatory protein

SV

Seminal vesicle

TD

Tolerable daily intake

TP

Testicular pathology

TRI

Toxics Release Inventory

TSCA

Toxic Substances Control Act

WWTP

Wastewater treatment plant

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EXECUTIVE SUMMARY

The U.S. Environmental Protection Agency (EPA or the Agency) is currently conducting risk
evaluations for five phthalates designated as high-priority substances under the Toxic Substances
Control Act (TSCA)—di-ethylhexyl phthalate (DEHP), butyl benzyl phthalate (BBP), dibutyl phthalate
(DBP), di-isobutyl phthalate (DIBP), and dicyclohexyl phthalate (DCHP)—as well as two phthalates
subject to manufacturer-requested risk evaluation: di-isononyl phthalate (DINP) and di-isodecyl
phthalate (DIDP).

Phthalates are a group of ubiquitous environmental chemicals that are used in many industrial and
consumer products, including cosmetics, building and construction materials, and polyvinyl chloride
products, to make plastics more flexible and durable. Some phthalates are used in food contact materials
and have been measured in food. Studies investigating human exposure to phthalates have demonstrated
widespread exposure to some phthalates and that humans may become co-exposed to multiple phthalates
at the same time. Further, some phthalates have been shown to cause common adverse effects on the
developing male reproductive system, sometimes referred to as "phthalate syndrome." Because humans
are co-exposed to some phthalates and because some phthalates can cause common adverse effects on
the developing male reproductive system, EPA believes that the best approach to assess risk to human
health may be to look at the combined risk to health from exposure to multiple phthalates.

As one of the first steps in the risk evaluation process, EPA published the final scope documents for the
seven phthalates between 2020 and 2021. During the public comment periods for the draft scope
documents, EPA received comments from multiple stakeholders urging the Agency to assess phthalates
for cumulative risk to human health because humans are co-exposed to multiple phthalates and because
some phthalates can cause common adverse effects. The next step in the risk evaluation process is to
conduct individual risk evaluations for DEHP, BBP, DBP, DIBP, DCHP, DINP, and DIDP, which will
characterize risk from their conditions of use (COUs). EPA's Office of Pollution Prevention and Toxics
(OPPT) has not yet conducted a cumulative risk assessment (CRA) under TSCA, as it is still developing
the methods and approaches for conducting CRA under TSCA. Moreover, the results of the individual
phthalate risk evaluations are important inputs into the CRA and the development of individual risk
evaluations is still ongoing.

This draft document provides a description of a proposed approach to conduct a CRA on the phthalates.
but is not itself a CRA as no risk estimates are presented nor has any work on risk evaluation been
completed. This draft document, along with the Draft Proposed Principles of Cumulative Risk
Assessment under the Toxic Substances Control Act (hereafter referred to as Draft Proposed Principles
of CRA under TSCA), will be released for public comments and reviewed by the Science Advisory
Committee on Chemicals (SACC) in 2023. EPA will then use the peer review and public input to guide
the development of the CRA for phthalates. Although EPA is required to draft individual risk
determinations for each individual phthalate risk evaluation, the phthalate CRA will not contain a risk
determination. Instead, results from the CRA are anticipated to inform EPA's individual phthalate risk
determinations, pending completion of the CRA in parallel with individual phthalate risk evaluations.

TSCA does not expressly require EPA to conduct CRAs. However, TSCA does require that EPA, when
conducting TSCA risk evaluations in 3 to 3.5 years [15 U.S.C. § 2605(b)(4)(G)], consider the reasonably
available information, consistent with the best available science, and make decisions based on the
weight of scientific evidence [15 U.S.C. § 2625(h), (i), (k)]. EPA is also required to conduct the risk
evaluations in consideration of potentially exposed or susceptible subpopulations (PESS) [15 U.S.C. §
2605(b)(4)] and, among other requirements at 15 U.S.C. § 2605(b)(4)(F), "integrate and assess available
information on hazards and exposures for the conditions of use of the chemical substance, including

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information that is relevant to specific risks of injury..." EPA recognizes that for some chemical
substances undergoing risk evaluation, the best available science may indicate that the development of a
CRA is appropriate to ensure that any risks to human health are adequately characterized. To support
CRA of chemical substances under TSCA, and as noted above, EPA has developed the Draft Proposed
Principles of CRA under TSCA, which describes the proposed principles of CRA as potentially
conducted in support of TSCA risk evaluations and relies heavily on long-standing EPA practice and
guidance documents for mixtures risk assessment. The draft principles document lays the foundation for
EPA's proposed approach for CRA of chemical substances undergoing risk evaluation under TSCA
section 6(b).

EPA has conducted a preliminary review of
stakeholder comments received during the phthalate
scoping process, previous phthalate CRAs
conducted by other regulatory agencies (ECCC/HC.

2020; EFSA. 2019; NICNAS. 2015a. 2014a. b; U.S.

CPSC. 2014; NICNAS. 2013. 2012; EC HA, 201 1),
and recommendations of the National Research
Council (NRC) (2008). Based in part on this
information, EPA believes that the best available
science indicates that several phthalates undergoing
risk evaluation should be assessed for cumulative
risk to human health. This draft document describes
EPA's proposed approach for assessing these high-
priority and manufacturer-requested phthalates for
cumulative risk to human health under TSCA. Text
Box ES-1 provides a high-level summary of EPA's
proposed approach for CRA.

Individual phthalate risk evaluations are required to
consider exposures from the COUs of a single
phthalate and will include evaluation of all observed
hazards, consideration of all age groups and
lifestages, and assessment of aggregate exposures. In
contrast, the scope and purpose of CRAs are more
focused on the shared toxicological properties and
relevant lifestages. In addition, cumulative exposure
assessment is more complicated due to combining
exposures across multiple phthalates.

EPA has developed a conceptual model to outline its
proposed approach for estimating cumulative risk to
phthalates within the cumulative chemical group. EPA's draft conceptual model, which is shown in
Figure 2-1 and described in Section 2, outlines 10 proposed steps for conducting a phthalate CRA under
TSCA. A brief description and summary of the outcome of each step follows:

Step 1 in EPA's draft conceptual model is to determine which high-priority and manufacturer-
requested phthalates to include in the cumulative chemical group. As described in EPA's Draft
Proposed Principles of CRA under TSCA document (and in Section 3 of this document), chemicals
included in a cumulative chemical group should be toxicologically similar and there should be evidence

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Text Box ES-1. Summary of EPA's Proposed
Approach for CRA of High-Priority and
Manufacturer-Requested Phthalates

EPA proposes to:

-	Group DEHP, BBP, DBP, DIBP, DCHP and
DINP, but not DIDP, for CRA under TSCA.

-	Address phthalate syndrome by focusing on the
most sensitive effect (versus addressing the
syndrome as a whole).

-	Assess DEHP, BBP, DBP, DIBP, DCHP and
DINP for cumulative risk to human health
under an assumption of dose addition.

-	Use a relative potency factor approach for the
phthalate CRA conducted in support of TSCA.

-	Focus its CRA efforts on PESS susceptible to
phthalate syndrome (i.e., pregnant women,
women of reproductive age, male infants, male
toddlers, male children).

-	Consider exposures from TSCA COUs, as well
as non-attributable and non-TSCA exposures.

-	Use a scenario-building approach to estimate
cumulative exposure for susceptible
populations who may also be workers,
consumers, and members of the general
population (e.g., fenceline communities).

-	Use biomonitoring data when available to
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of co-exposure to the chemicals over a relevant timeframe (e.g., exposed to multiple phthalates during a
known sensitive lifestage).

To determine which high-priority and manufacturer-requested phthalates are toxicologically similar,
EPA reviewed data for seven key outcomes associated with phthalate syndrome; that is, decreased fetal
testicular gene expression and testosterone production, decreased male pup anogenital distance,
nipple/areolae retention in male pups, hypospadias, seminiferous tubule atrophy, and multinucleated
gonocyte formation (Sections 3.1.3.1 to 3.1.3.7). These key outcomes were selected based on EPA's
current understanding of phthalate syndrome and its underlying mode of action. Notably, many of the
key outcomes have also been selected as the critical effect (or co-critical effect) in previous phthalate
CRAs (Table 3-1). Based on the weight of evidence, EPA proposes that DEHP. BBP. DBP. DIBP.
DC HP. and DINP. but not DIDP. are toxicologically similar and induce effects on the developing male
reproductive system consistent with phthalate syndrome (Section 3.1.7). Of note, the TSCA Work Plan
includes one additional phthalate (i.e., di-n-octyl phthalate) that is not currently prioritized for risk
evaluation. However, Environment Canada/Health Canada (EC/HC. 2015 e) concluded that di-n-octyl
phthalate does not induce effects on the developing male reproductive system consistent with phthalate
syndrome (EC/HC. 2015e). Di-n-octyl phthalate is not discussed further in this document.

When considering phthalates for grouping, EPA also considered how to address phthalate syndrome,
which is currently identified as the common adverse effect, as part of a CRA. EPA is proposing to focus
on the most sensitive effect(s) (as opposed to assessing the syndrome as a whole) (Section 4.1). As
described in Section 4.2, empirical evidence from in vivo phthalate mixture studies indicate that
phthalates induce effects on the developing male reproductive system in a manner consistent with dose
addition. Therefore, EPA is proposing to assess DEHP. BBP. DBP. DIBP. DCHP. and DINP for
cumulative risk to human health under an assumption of dose addition, which is consistent with the
recommendations of the NRC (2008). EPA is considering the applicability of two component-based,
dose additive approaches, including the hazard index (HI) and relative potency factor (RPF) approaches.
EPA considers there to be sufficient data available to support RPF derivation for DEHP. BBP. DBP.
DIBP. DCHP. and DINP (Section 4.3.3) and is proposing to use an RPF approach to assess these
phthalates for cumulative risk. EPA has identified six potential options that are being considered for
deriving RPFs for phthalates, which are described in Section 4.4.2.

To determine if the U.S. population is co-exposed to multiple high-priority and manufacturer-requested
phthalates, EPA conducted a high-level review of National Health and Nutrition Evaluation Surveys
(NHANES) urinary biomonitoring data (Section 3.2). Available NHANES data demonstrate that the
U.S. population is co-exposed to multiple phthalates, including DEHP, BBP, DBP, DIBP, DINP, and
DIDP. Recent NHANES data are not available for DCHP. However, DCHP has been identified to be
used in various industrial, commercial, and consumer uses covered under TSCA. Based on exposure to
DCHP through identified TSCA uses, EPA anticipates there will be co-exposure to DCHP and other
high-priority and manufacturer-requested phthalates for certain populations and exposure scenarios
(Section 3.2). These data qualitatively demonstrate that humans are co-exposed to DEHP, BBP, DBP,
DIBP, DCHP, DINP, and DIDP. EPA's proposed approach for quantifying phthalate co-exposure is
outlined in Section 6.

Because the weight of evidence indicates that DEHP. BBP. DBP. DIBP. DCHP and DINP (but not
DIDP) are toxicologically similar and that the U.S. population is co-exposed to these phthalates over a
relevant timeframe. EPA is proposing to group these phthalates for CRA under TSCA.

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Step 2 in EPA's draft conceptual model (Figure 2-1) is to identify populations with potentially
increased susceptibility to phthalate syndrome. As part of the individual phthlate risk evaluations,
EPA will conduct consumer, occupational, and general population exposure assessments. Within these
populations, potentially exposed or susceptible subpopulations (PESS) with greater susceptibility to the
developmental and reproductive effects associated with phthalate syndrome, include pregnant women,
women of reproductive age, male infants, male toddlers, and male children. These PESS are proposed to
be the focus of EPA's approach for CRA of DEHP, BBP, DBP, DIBP, DCHP and DINP (Section 5).

Step 3 in EPA's draft conceptual model (Figure 2-1) is to identify TSCA COUs1 and other
potential sources of exposure. Sources of exposure including TSCA COUs, non-attributable, and non-
TSCA sources relevant to cumulative exposure and release will be identified using conceptual models in
individual phthalate scopes and literature reviews.

Step 4 in EPA's draft conceptual model (Figure 2-1) is exposure scenario-building for individual
phthalates for TSCA COUs. For identified TSCA COUs and populations, specific routes of exposure
and pathways for each exposure source are identified. Prior to the development of the phthalate CRA,
exposure scenarios for individual TSCA COUs and estimates of exposure will be completed in the
individual risk evaluations. Determination of co-exposure to multiple TSCA COUs or multiple
phthalates in a single TSCA COU will be completed in Step 7 of the conceptual model for consumers
(Section 6.4.1), workers (6.4.2), and the general population (Section 6.4.3).

Step 5 in EPA's draft conceptual model (Figure 2-1) is to build exposure scenarios of individual
phthalates for non-attributable and non-TSCA sources. EPA is proposing to include both non-
attributable and non-TSCA exposures as part of the phthalate CRA because certain non-TSCA (e.g.,
dietary) and non-attributable (e.g., household dust) exposure pathways are anticipated to be major
contributors to phthalate exposure leading to cumulative risk (discussed further in Section 6.2.2). The
Agency is considering two approaches for estimating non-attributable and non-TSCA phthalate
exposure, including a scenario-based approach (Section 6.3.2.1) and a reverse dosimetry-based approach
(Section 6.3.2.2). Because the reverse dosimetry approach, using biomonitoring data such as NHANES,
does not distinguish between routes or pathways of exposure and does not allow for source
apportionment, it provides an estimate of total non-attributable phthalate exposure. NHANES data may
reflect exposure from TSCA, non-attributable, and other non-TSCA sources, but exposures from TSCA
COUs cannot necessarily be source apportioned. As described in Section 6.3.2.5, EPA is proposing to
estimate non-attributable and non-TSCA exposures for DEHP. BBP. DBP. DIBP. DCHP. and DINP
from major exposure pathways using a scenario-based approach. The reverse dosimetry approach, which
does not allow for source apportionment, may be used to help characterize phthalate exposure and serve
as a comparator for scenario-based intake estimates (i.e., help contextualize whether scenario-based
estimates are an over- or underestimation of total exposure).

Steps 6 and 7 in EPA's draft conceptual model (Figure 2-1) are to identify major pathways of
exposure (Step 6) and determine the likelihood of phthalate co-exposure (Step 7). As shown in
EPA's draft conceptual model (Figure 2-1), EPA is proposing to assess PESS who are consumers
(Section 6.4.1), workers (Section 6.4.2), and fenceline communities as part of the general population

1 Condition of use (COU) (.1.5 U.S.C. § 2602(4)1: "means the circumstances, as determined by the Administrator, under which
a chemical substance is intended, known, or reasonably foreseen to be manufactured, processed, distributed in commerce,
used, or disposed of."

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(Section 6.4.3) for cumulative risk from exposure to DEHP, BBP, DBP, DIBP, DCHP, and DINP
through TSCA COUs. EPA proposes to identify major pathways of exposure and likelihood of co-
exposure to these phthalates through various pathways for combining to estimate cumulative exposure to
identified PESS (Steps 6 to 7 in conceptual model).

•	Major pathways of exposure for individual phthalates are combined to estimate aggregate
exposure and can be considered exposures attributable to TSCA COUs, non-attributable, or non-
TSCA.

•	To estimate cumulative exposure to consumers (Section 6.4.1), EPA proposes to combine the
non-attributable and non-TSCA exposures across phthalates with exposure from individual
consumer COUs, as reasonable. Determining reasonable cumulative exposure scenarios may
involve considering the likelihood of co-exposure, the possibility of double counting, and of
over- or under-estimating exposures.

•	To estimate cumulative exposure to workers (Section 6.4.2), EPA proposes to combine the non-
attributable and non-TSCA exposure with cumulative occupational exposure from TSCA COUs
in a work setting, as reasonable.

•	For cumulative exposure to the general populations, specifically fenceline communities (Section
6.4.3), EPA proposes estimating cumulative exposures from single or multiple facility releases to
ambient air and/or water and combining with non-attributable and non-TSCA exposure, as
reasonable.

•	EPA recognizes that some individuals may be part of multiple populations and may require
additional combinations of exposures. For example, combining occupational exposures with
consumer exposures and fenceline exposures for workers who use consumer products at home
and who live near the fenceline of a facility with phthalate releases.

Steps 8 to 10 in EPA's draft conceptual model (Figure 2-1) are to convert individual phthalate
exposure estimates to index chemical equivalents using RPFs (Step 8), and then to combine
exposures to estimate cumulative exposure (Step 9) and cumulative risk (Step 10). Because EPA is
proposing to use an RPF approach (Section 4.3.3), exposure from individual phthalates for each
exposure scenario will be scaled to the potency of an index chemical and expressed as index chemical
equivalents (Step 8 in conceptual model), which will then be summed to estimate cumulative exposure
for each exposure scenario (expressed as index chemical equivalents) (Step 9 in conceptual model).
Cumulative risk may then be estimated using a margin of exposure (MOE) approach (Section 4.3.3)
(Step 10 in conceptual model).

EPA is soliciting comments from the SACC on charge questions and comments from the public for the
SACC meeting scheduled on May 8-11, 2023.

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1 BACKGROUND

In December 2019, the U.S. Environmental Protection Agency (EPA or the Agency) designated butyl
benzyl phthalate (BBP, Chemical Abstracts Service Registry Number [CASRN] 85-68-7), dibutyl
phthalate (DBP, CASRN 84-74-2), dicyclohexyl phthalate (DCHP, CASRN 84-61-7), di-ethylhexyl
phthalate (DEHP, 117-81-7), and di-isobutyl phthalate (DIBP, CASRN 85-69-5) as high-priority
substances for risk evaluation under the Toxic Substances Control Act (TSCA) (	2019b. c, d,

e, f). Additionally, on May 24, 2019, EPA received requests from industry, pursuant to 40 CFR 702.37,
to conduct risk evaluations for di-isodecyl phthalate (DIDP, CASRNs 26761-40-0 and 68515-49-1)
(	PP. 2019a") and di-isononyl phthalate (DINP, CASRNs 28553-12-0 and 68515-48-0) ( :

H	). The Agency determined that the requests met the applicable regulatory criteria and

requirements, as prescribed under 40 CFR 702.37, and granted the manufacturer-requested risk
evaluations for DIDP and DINP on December 2, 2019. As one of the first steps in the risk evaluation
process, EPA published the final scope documents for BBP (• c. « i1 \ . 220a), DBP (• c. « i1 \
2020(f), DCHP (U.S. EPA. 2020e). DEHP (\ ^ \ 2020b), and DIBP (1 ^ \ :020c) in August
2020, fulfilling TSCA requirements under TSCA section 6(b)(4)(D) and as described in 40 CFR
702.41(c)(8). In August 2021, EPA published the final scope documents for DIDP (	2021b)

and DINP (U.S. EPA. 2021c).

During the public comment periods for the draft scope documents for the high-priority phthalates and
phthalates subject to manufacturer-requested risk evaluation, EPA received comments from multiple
stakeholders urging the Agency to assess phthalates for cumulative risk to human health.2'3 Recognizing
that human exposure to phthalates is widespread and that multiple phthalates can disrupt development of
the male reproductive system in laboratory animals at potentially human relevant doses, in 2007 EPA
asked the National Research Council (NRC) of the National Academy of Sciences (NAS; now National
Academies of Sciences, Engineering, and Medicine [NASEM]) to form a committee to review the health
effects of phthalates and determine whether a cumulative risk assessment (CRA) of phthalates is
appropriate. Additionally, EPA asked the NRC to provide recommendations on specific approaches that
could be used to assess phthalates for cumulative risk. NRC published their findings and
recommendations to EPA in a 2008 report Phthalates and Cumulative Risk Assessment: The Tasks
Ahead (NRC. 2008). Ultimately, the NRC concluded that "sufficient data are available to proceed with
the cumulative risk assessment of phthalates..." [p. 10 of (NRC. 2008)1.

In 2010, and in response to the NRC recommendations, EPA's Office of Research and Development's
Integrated Risk Information System (IRIS) Program convened a 2-day peer consultation workshop to
discuss and evaluate the NRC recommendations. As summarized in the final workshop report (U.S.
E	), there was broad support by both expert panelists and stakeholders to continue developing a

cumulative hazard assessment.

Other regulatory agencies have assessed phthalates for cumulative risk since NRC published their
recommendations (NRC. 2008)—including the Chronic Hazard Advisory Panel (CHAP) of the U.S.

2	For example, see comments submitted to the DEHP Docket (EPA-HQ-OPPT-2018-0433) received from the Environmental
Defense Fund (EP A-HO-OPPT-2018-043 3 4)03 31. Environmental Protection Network (EP A-HO-OPPT-2018-043 3 -00281.
Project TENDR (EPA-HQ-OPPT-20.1.8-0433-0045): and University of California. San Francisco Program on Reproductive
Health and the Environment (! r x IIO-OPPT-20.1.8-043 3 -0013).

3	For example, see comments submitted to the DINP Docket (EPA-HQ-OPPT-20.1.8-0436) received from University of
California. San Francisco Program on Reproductive Health and the Environment (EPA-HQ-OPPT-20.1.8-0436-0009);
Environmental Protection Network (EPA-HQ-OPPT-2018-0436-0026); Earthjustice (EPA-HQ-C	1.8-0436-0028.
EPA-HQ-OPPT-20.18-0436-0033); and Defend Our Health. Black Women for Wellness. Alaska Community Action on
Toxics and Breast Cancer Prevention Partners (EPA-HQ-OPPT-2018-0436-0042).

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Consumer Product Safety Commission 0 ; S CP SC. 2014); Environment and Climate Change Canada,
Health Canada (ECCC/HC. 2020); the National Industrial Chemicals Notification and Assessment
Scheme (N1CNAS) of Australia (NICNAS. 2015a. 2014a. b, 2013. 2012); the European Food Safety
Authority (EFSA. 2019). and the Danish EPA (	). Although the phthalate CRAs conducted

by these regulatory agencies vary in scope and regulatory purpose, they generally adhere to NRC
recommendations (NRC. 2008). For example, the CRAs primarily focus on assessing phthalates based
on their shared ability to disrupt development of the male reproductive system through a disruption of
androgen action {i.e., cause phthalate syndrome), and have all relied upon an assumption of dose
addition (see Appendices A.l to A.5 for summaries of phthalate CRAs conducted by these agencies).

1.1 What Is EPA Proposing in this Work?

As required under section 6(b)(4) of TSCA, EPA issued a final rule, Procedures for Chemical Risk
Evaluation Under the Amended Toxic Substances Control Act (	'26) (hereinafter "Risk

Evaluation Rule"), in July 2017, which provides the procedural requirements for EPA's risk evaluations,
including for chemicals designated as High-Priority Substances and chemical substances subject to a
Manufacturer-Requested Risk Evaluation. To date, EPA's Office of Pollution Prevention and Toxics
(OPPT) has focused risk evaluations on individual chemical substances, not the evaluation of multiple
chemical substances for cumulative risk to human health. TSCA does not define cumulative risk nor
explicitly require EPA to conduct CRAs. However, TSCA does require EPA, when conducting TSCA
risk evaluations, to (1) consider the reasonably available information, (2) use the best available science,
and (3) make decisions based on the weight of the scientific evidence [15 U.S.C. § 2625(h), (i), (k)].
EPA is also required to conduct the risk evaluations in consideration of potentially exposed or
susceptible subpopulations (PESS) [15 U.S.C. § 2605(b)(4)] and, among other requirements at 15 U.S.C.
§ 2605(b)(4)(F), "integrate and assess available information on hazards and exposures for the conditions
of use of the chemical substance, including information that is relevant to specific risks of injury..." EPA
recognizes that for some chemical substances undergoing risk evaluation, the best available science may
indicate that the development of a CRA is appropriate to ensure that risks of injury to human health and
the environment are adequately characterized. Although EPA is required to draft individual risk
determinations for each individual phthalate risk evaluation, the phthalate CRA will not contain a risk
determination. Instead, results from the CRA are anticipated to inform EPA's individual phthalate risk
determinations, pending completion of the CRA in parallel with individual phthalate risk evaluations. To
support CRA of chemical substances undergoing TSCA section 6(b) risk evaluations, EPA has
developed a document titled Draft Proposed Principles of Cumulative Risk Assessment under the Toxic
Substances Control Act (hereafter referred to as Draft Proposed Principles of CRA under TSCA). EPA's
Draft Proposed Principles of CRA under TSCA document describes the proposed principles of CRA,
which form the underpinning of EPA's draft approach for CRA of high-priority and manufacturer-
requested phthalates.

The Agency has reviewed the recom m endati on s of the NRC (2008). comments received from
stakeholders on the draft scope documents (see footnotes in Section 1), and CRAs conducted by other
regulatory agencies (see Appendices A.l to A.5). Based on this information, EPA believes the best
available science indicates that several high-priority and manufacturer-requested phthalates should be
assessed for cumulative risk to human health.

As part of conducting a risk evaluation under TSCA section 6(b), EPA must "determine whether a
chemical substance presents unreasonable risk of injury to health . . . including an unreasonable risk to a
potentially exposed or susceptible subpopulation [(PESS)] identified as relevant to the risk evaluation by
[EPA] . . ." [15 U.S.C. 2605(b)(4)(A)], EPA has identified phthalate syndrome as a specific risk from a
number of the phthalates undergoing risk evaluation. The Agency has also identified a number of PESS

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that have a greater susceptibility to phthalate toxicity—including pregnant women/women of
reproductive age, male infants, male toddlers, and male children (discussed in Section 5). Due to
toxicological similarity, shared ability to elicit key markers of phthalate syndrome, and co-exposures to
multiple phthalates to the aforementioned PESS (one of the factors laid out in Section 3.4 of the Draft
Proposed Principles of CRA under TSCA), EPA is proposing that a subset of the phthalates undergoing
risk evaluation represent a cumulative chemical group, and that a cumulative risk assessment is
necessary to ensure that individual risk evaluations on the phthalates in the cumulative chemical group
have considered the reasonably available information, are consistent with the best available science, and
based on the weight of the scientific evidence (15 U.S.C. 2625(h), (i), & (k)).

This draft document describes EPA's proposed approach for evaluating the phthalates in the cumulative
chemical group for cumulative risk to human health under TSCA. The phthalates included in OPPT's
proposed CRA are limited, at this time, to those undergoing risk evaluation under TSCA and are
inclusive of the phthalates that have been most commonly considered for CRA by other agencies (see
Appendix A).

This document describes EPA's draft proposed approach for assessing high-priority and manufacturer-
requested phthalates for cumulative risk to human health under TSCA based on the principles of CRA
described in the Draft Proposed Principles of CRA under TSCA. The proposed approach described in
this document follows many of the recommendations made by NRC (2008). Individual phthalate risk
evaluations will consider exposures from a single phthalate and will include evaluation of all observed
hazards, consideration of more age groups and lifestages, and assessment of aggregate exposures. In
contrast, the scope and purpose of CRAs are more focused on the shared toxicological properties and
relevant lifestages. In addition, cumulative exposure assessment is more complicated due to combining
exposures across multiple phthalates.

At the date of publication of this document, EPA has not yet completed all the expected systematic
review or data quality evaluation for the individual high-priority and manufacturer-requested phthalates.
Although this document is not reflective of complete systematic review, EPA has reviewed several key
documents prepared by various authoritative bodies and regulatory agencies along with numerous
studies and databases of toxicological and exposure information. As appropriate, EPA's proposed
approach may be revised based on any new information that is identified through the systematic review
process. Some key documents used to develop this proposed approach include

•	Phthalates and Cumulative Risk Assessment: The Tasks Ahead (NRC. 2008)

•	Application of Systematic Review Methods in an Overall Strategy for Evaluating Low-Dose
Toxicity from Endocrine Active Chemicals (NA.SEM. 2017)

•	Report to the U.S. Consumer Product Safety Commission by the Chronic Hazard Advisory
Panel on Phthalates and Phthalate Alternatives (	>C. 2014) and supporting toxicity
reviews of DEHP (\ < S CPSC. 2010c). BBP (\ « i PSC. 2010a). DBP (I > i PSC. 2010b).
DIBP (U.S. CPSC. 2011). DC HP Q ' i\ SC. 2010e). DINP (1 ^ i \™C. 2010fl. and DIDP
0 v <• 2010d)

•	Screening Assessment, Phthalate Substance Grouping (ECCC/HC. 2020) and supporting reports
(EC/HC. 2015a. b, c, e; Health Canada.: )

•	Existing Chemical Hazard Assessment Reports for DIBP (NICNAS. 2008b) and DEHP
(NICNAS. 2008a) and Priority Existing Chemical Assessment Reports for BBP (NICNAS.

2015a). DBP (NICNAS. 2013). DINP (NICNAS. 2012). DIDP (NICNAS. 2015b)

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717	• Update of the risk assessment of di-butylphthalate (DBP), butyl-benzyl-phthalate (BBP), bis(2-

718	ethylhexyl)phthalate (DEHP), di-isononylphthalate (DINP) and diisodecylphthalate (DIDP) for

719	use in food contact materials (EFSA. 2019)

720	This draft document, along with the Draft Proposed Principles of CRA under TSCA, will be reviewed

721	by the Science Advisory Committee on Chemicals (SACC) and receive public comments in 2023. EPA

722	will use the peer review and public input to guide the subsequent development of the CRA for

723	phthalates.

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2 KEY CONCEPTS AND PROPOSED CONCEPTUAL MODEL

Individual phthalate risk evaluations will consider exposures from a single phthalate and will include
evaluation of all observed hazards, consideration of all age groups and lifestages, and assessment of
aggregate exposures. In contrast, the scope and purpose of CRAs are more focused on the shared
toxicological properties and relevant lifestages. In addition, cumulative exposure assessment is more
complicated due to combining exposures across multiple phthalates. Therefore, EPA has provided some
definitions to key concepts relevant to CRAs in Section 2.1 and developed a draft conceptual model
described in Section 2.2 and shown in Figure 2-1 to outline its proposed approach for estimating
cumulative risk to several of high-priority and manufacturer-requested phthalates.

2.1 Key Concepts

•	Cumulative chemical group: A group of chemical substances included in a CRA. As discussed
in EPA's Draft Proposed Principles of CRA under TSCA, the cumulative chemical group is
developed based on evidence of toxicologic similarity and co-exposure over a relevant
timeframe.

•	Co-exposure: Characterizing co-exposure requires consideration of the source of chemical
exposure, populations impacted by exposure, and the possible varying routes and pathways of
exposure. Additionally, the magnitude, frequency, and duration of exposure to multiple chemical
substances influence the potential for co-exposure to occur within a given period of time (e.g., 24
hours, 1 year, a lifetime); where the magnitude of exposure is the level of exposure dictated by
the physical and chemical properties of the chemical substance and exposure scenario, frequency
is the number of exposure events over a given time, and duration is the length of exposure time
per event (()HCLL2 ' < \ v ii \ .0011

•	Relevant timeframe of exposure: Timeframes in which exposure duration or frequency is
relevant to effects of concern. This can include, but may not be limited to, exposure to multiple
chemicals at the same time, exposure to persistent chemicals at different times that may
bioaccumulate in the body or having persistent effects from exposure to multiple chemicals at
different times. Relevant timeframes of exposure can vary by factors including, but not limited
to, chemical properties, lifestages, or effect. Relevant timeframes of exposure for phthalates will
be determined through the risk evaluation process.

•	Relative potency factor: A numerical quantity used to scale the dose of one chemical to an
equitoxic dose of another chemical based on differences in potencies. The latter chemical is
typically termed the "index chemical" and is usually the chemical in the cumulative chemical
group with the most robust toxicological database and/or is considered to be the most
representative of the type of toxicity caused by other chemicals in the cumulative chemical group
(	2000).

•	Scenario-based evaluation:4 Estimates that use available information on concentrations of
chemicals in the exposure medium, and information about when, where, and how individuals
might contact the exposure medium—activities that can lead to transfer of the agent from the
exposure medium to the individual. Approach develops specific exposure scenarios and then
uses data, a series of exposure factors, and models to estimate exposure within the scenario (U.S.
EPA. 2019aY

4 Referred to as indirect estimation in EPA's Guidelines for Human Exposure Assessment (U.S. EPA. 2019a').

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•	Reverse dosimetry: Estimates chemical intake using empirical biomonitoring data and
information about chemical absorption, distribution, metabolism, and excretion rates (	L,
2019a).

•	TSCA COU exposure:5 Exposure that can be attributed to a specific TSCA COU (e.g.,
inhalation exposure during consumer use of an adhesive). Note that exposure scenarios for
TSCA COUs will be completed in individual phthalate risk evaluations and evaluated for
different populations such as consumers, workers, and general population.

•	Non-attributable exposure: Exposure from pathways that cannot be attributed to a specific
TSCA COU or another specific source. Household dust or human milk are a few examples in
which phthalate concentrations measured in those media may result from multiple sources of
phthalates that may nor may not be attributed to a TSCA COU or another specific source.

•	Non-TSCA exposure:6 Exposure that can be attributed to specific activities that are excluded
from the TSCA definition of "chemical substance," under TSCA section 3(2), such as a
pesticide, food, food additive, drug, cosmetic, or medical device.

2.2 Proposed Conceptual Model

EPA has developed a conceptual model to outline its proposed approach for estimating cumulative risk
to several of the high-priority and manufacturer-requested phthalates. EPA's draft conceptual model,
which is shown in Figure 2-1, outlines 10 proposed steps for conducting a phthalate CRA under TSCA
using a scenario-based approach. The conceptual model provides illustrative steps that may not be
inclusive of all details, such as all populations or all pathways of exposure, to be considered in an actual
cumulative assessment. The remainder of this document is structured around this draft conceptual
model. Some steps are described in greater detail in the document while others require risk evaluation
work to be conducted to be developed further.

The steps included in the conceptual model are provided below:

•	Step 1. Identifying the Cumulative Chemical Group: Identified based on a shared ability to
elicit key markers of phthalate syndrome and evidence of human co-exposure. EPA's proposed
cumulative chemical group includes DEHP, BBP, DBP, DIBP, DCHP, and DINP (Section 3.3).

•	Step 2. Populations: EPA will conduct consumer, occupational, and general population (e.g.,
fenceline) exposure assessments for each individual phthalate. The key human populations
considered in these exposure assessments include consumers, workers, and the general
population. Within these groups, there are PESS with increased susceptibility to the
developmental and reproductive effects associated with phthalate syndrome, including pregnant
women/women of reproductive age, male infants, male toddlers, and male children (described
further in Section 5).

5	Condition of use (COU) (40 CFR § 702.33): "means the circumstances, as determined by the Administrator, under which a
chemical substance is intended, known, or reasonably foreseen to be manufactured, processed, distributed in commerce, used,
or disposed of."

6	TSCA section 3(2) also excludes from the definition of "chemical substance" "any food, food additive, drug, cosmetic, or
device (as such terms are defined in section 201 of the Federal Food, Drug, and Cosmetic Act [21 U.S.C. 321]) when
manufactured, processed, or distributed in commerce for use as a food, food additive, drug, cosmetic, or device" as well as
"any pesticide (as defined in the Federal Insecticide, Fungicide, and Rodenticide Act [7 U.S.C. 136 et seq.]) when
manufactured, processed, or distributed in commerce for use as a pesticide." Section 2.2.2 of each final scope document for
BBP (U.S. EPA. 2020a). DBP (U.S. EPA. 2020d). DCHP (U.S. EPA. 2020e). DEHP (U.S. EPA. 2020b). DIBP (U.S. EPA.
2020c). and DINP (U.S. EPA. 2021c) outline the uses of each phthalate that EPA has determined to be non-TSCA uses.

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•	Step 3. Identify TSCA COUs and Other Sources of Exposure: After gathering the specific
COUs for each phthalate from their individual risk evaluation scope documents, the cross-
chemical comparisons are used to establish the COUs likely to result in co-exposure to multiple
phthalates under TSCA (Section 6.2.1). Other sources of exposure that are not considered TSCA
COUs may also be identified as major sources of exposure for the identified populations through
a review of the literature.

•	Step 4. Exposure Scenario-Building for Individual Phthalates for TSCA COUs: For TSCA
COUs and populations, specific routes of exposure and pathways for each exposure source are
identified. Exposure scenarios for individual TSCA COUs and estimates of exposure will be
completed in the individual risk evaluations. Determinations on the likelihood of co-exposure to
multiple phthalates in multiple TSCA COUs or multiple phthalates in a single TSCA COU will
be completed in Step 7 of the conceptual model for consumers (Section 6.4.1), workers (6.4.2),
and the general population, specifically fenceline communities (Section 6.4.3).

•	Step 5. Exposure Scenario-Building for Individual Phthalates for Non-Attributable and
Non-TSCA Sources: For identified sources of exposure (non-attributable or non-TSCA) and
populations, specific routes of exposure and pathways for each exposure source are considered.
Exposure scenarios are considered for major sources of exposure and exposure is estimated for
the various pathways of exposure. Scenario-building to estimate non-attributable and non-TSCA
exposures is discussed in Section 6.3.2.1.

•	Steps 6 to 9. Determining Cumulative Exposure Estimates: Cumulative exposure potentially
assessed under TSCA may be estimated by combining exposures from major exposure pathways
from TSCA COUs, non-attributable, and non-TSCA sources that may lead to co-exposure over a
relevant timeframe, which can mean exposure to multiple chemicals at the same time, exposure
to persistent chemicals at different times that may bioaccumulate in the body, or having
persistent effects from exposure to multiple chemicals at different times. This process involves:

o Step 6. Identifying Major Pathways of Exposure: Determining the major pathways of
exposure from TSCA COUs (completed in individual risk evaluations), non-attributable,
and non-TSCA sources for each phthalate. Different pathways of exposure may be
relevant for different populations and for different phthalates. For example, the human
milk and formula-fed pathways are most relevant for infant scenario-building, while
mouthing may be most relevant to infants and toddlers. Major pathways of exposure for
individual phthalates may be combined at this step to estimate aggregate exposure.

o Step 7. Determining Co-exposure: Determining likelihood of co-exposure across TSCA
COUs, non-attributable sources, and non-TSCA sources for the various phthalates.
Estimating the exposure associated with the consumer (Section 6.4.1), occupational
(Section 6.4.2), and general population (fenceline) (Section 6.4.3) TSCA COU exposures
and adding these exposures across COUs and across phthalates if reasonable.

Determining reasonable cumulative exposure scenarios may involve considering the
likelihood of co-exposure, the possibility of double counting, and of over- or under-
estimating exposures

o Step 8. Convert Exposures to Index Chemical Equivalents: Because EPA is proposing
to use an RPF approach (Section 4.3.3), phthalate exposure from each individual
phthalate will be scaled to the potency of an index chemical and expressed in units of
index chemical equivalents.

o Step 9. Estimating Cumulative Exposure: Combining the TSCA COU or release
cumulative exposure, the relevant non-attributable TSCA cumulative exposure, and the

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846	non-TSCA cumulative exposure to estimate cumulative exposure in a reasonable manner

847	for consumer (Section 6.4.1), occupational (Section 6.4.2), and general population

848	(Section 6.4.3).

849	• Step 10. Estimate Cumulative Risk: To estimate cumulative risk for each specific exposure

850	scenario, an MOE (ratio of index chemical point of departure [POD] to cumulative exposure

851	estimate expressed in index chemical equivalents [Step 9]) is calculated for comparison to the

852	benchmark MOE {i.e., the total uncertainty factor associated with the assessment) (Section

853	4.3.2). The lower the MOE (margin between the toxicity effect level and the exposure dose), the

854	more likely a chemical is to pose a risk.

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Step 1. Identify

Cumulative
Chemical Group

BBP

DBP

DCHP

DEHP

DIBP

DINP

Step 2. Identify Subpopulations

susceptible to phthalate syndrome

Consumer

Occupational

General Population

Consumers using
products and articles



Pregnant women

& women of
reproductive age



Male infants,
toddlers, and
children

Workers in
facility



Pregnant women

& women of
reproductive age

Population exposed to environmental
releases

Among fenceline
communities:

o

Fenceline
communities

v

Pregnant women

& women of
reproductive age

Male infants,
toddlers, and
children

Step 3. Identify TSCA COUs and
Other Sources of Exposure

relevant to cumulative exposure and
release(s) using conceptual models in
individual phthalate scopes and
literature reviews

Step 4. TSCA COU Scenario-Building for Individual Phthalates

by identifying exposure pathways and routes

BBP

DBP

DCHP

DEHP

DIBP

DINP

TSCA COU(s)

«.1.I,UII„IJI.I.HI14»	WIMfrM'IW

Pathways may include Pathways may include Pathways may include but
but are not limited to:	but are not limited to:	are not limited to:

Inhalation

&
Oral

ft

Dermal

Step 5. Scenario-Building for Non-Attributable Sources &
Non-TSCA Sources for Individual Phthalates

by identifying exposure pathways and routes

BBP

DBP

DCHP DEHP

DIBP

DINP

Non-Attributable Sources

Exposure not attributable to a
specific source. Pathways may
include but are not limited to:

Non-TSCA Sources

Exposure to phthalates from

non-TSCA sources may
include but are not limited to:

855

Figure Continued on Next Page

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856

1

Figure Continued from Previous Page

Step 6. Identify Major Pathways

for inclusion in cumulative estimate

Step 7. Determine Co-Exposure

across all phthalates

BBP DBP DCHP DEHP DIBP DINP

> Determine relative importance of scenarios and need for
inclusion for each individual phthalate

Non-Attributable Sources

Non-TSCA Sources

Identify pathways that may lead to co-
exposure

>	Considertimeframeof exposure

>	Consider probability of co-exposure



.IIInEWill

iJM.M.IJIJIIfeUMAIIIJJJA



BBP

Non-TSCA Sources

DBP

DCHP

DEHP

DIBP

DINP

Step 8. Convert Exposures to Index
Chemical Equivalents

Individual phthalate exposures are
converted to index chemical equivalents
using RPFs.

Index chemical equivalents =
Exp0SUrePhthaiate 1 X RPFphrlialate 1

Step 9. Estimate Cumulative Exposure Across
Phthalates in the Cumulative Chemical Group for
Various Receptors

Combine exposure across pathways and phthalates

| Exposure

Step 10. Estimate Cumulative Risk

Calculate the MOE for comparison to the
benchmark MOE {i.e., total uncertainty factor)


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3 CONSIDERATIONS FOR GROUPING PHTHALATES FOR CRA:
STEP 1 IN CONCEPTUAL MODEL (Figure 2-1)	

As described in EPA's Draft Proposed Principles of CRA under TSCA, there are two primary
considerations for grouping chemicals for inclusion in a CRA, including (1) toxicologic similarity, and
(2) evidence of co-exposure over a relevant timeframe. Figure 3-1 presents a decision tree for
determining which of the high-priority (DEHP, BBP, DBP, DIBP, DCHP) and manufacturer-requested
(DINP, DIDP) phthalates currently undergoing risk evaluation to group for CRA. The establishment of
cumulative chemical group(s) for purposes of CRA is developed using a weight of evidence narrative
that clearly characterizes the strengths and uncertainties of the evidence of toxicological similarity and
potential co-exposure for each chemical considered. Evidence supporting the toxicologic similarity of
the high-priority and manufacturer-requested phthalates is discussed in Section 3.1, evidence
demonstrating co-exposure of humans to the high-priority and manufacturer-requested phthalates is
discussed in Section 3.2, and EPA's proposed chemical substance grouping for CRA is summarized in
Section 3.3.

High-priority (DEHP, BBP, DBP, DIBP,
DCHP) and manufacturer-requested
phthalates (DINP, DIDP)

I

Hazard Filter

Does the phthalate cause effects
consistent with phthalate syndrome?

No.

^Yes

Exposure Filter

Evidence of
exposure in the
U.S. through
industrial,
commercial, or
consumer uses?

and/
or

Is there evidence
to support co-
occurrence or co-
exposure in the
U.S. population?

No

^ Yes

Exposure Timeframe Filter

Is there evidence that exposure is occurring

over a relevant timeframe in the US
	population?	

No

Figure 3-1. Decision Tree for Grouping Phthalates for CRA

Adapted from (EC/HC. 2015a).

3.1 Evidence of Toxicologic Similarity

As described in EPA's Draft Proposed Principles of CRA under TSCA, evidence for toxicological
similarity exists along a continuum and includes (from most to least informative/restrictive):

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•	identical toxicodynamics {i.e., same mode of action [MOA]) (same molecular initiating event
[MIE], downstream key events, and apical outcome));

•	similar toxicodynamics {e.g., similar MOA [different MIE, convergent toxicodynamic pathways
leading to a common downstream effect, and same apical outcome]);

•	shared syndrome;

•	shared apical outcome (MIE and other key events unknown);

•	effect on the same target organ;

•	structural similarity; and

•	similarly shaped dose-response curves in comparable toxicity studies.

In considering which chemicals to include in a CRA, the NRC (2008) concluded that "...the effects that
make up the androgen-insufficiency syndrome" should be included regardless of mechanism of action or
chemical structure. In part, NRC's recommendation was based on the availability of in vivo mixture
studies of phthalates and other antiandrogens with mixed MO As that provide empirical evidence
demonstrating the applicability of dose additive models (NRC. 2008). However, NRC also emphasized
that mechanism of action data is still desirable for defining critical pathways, determining human
relevance of observed effects, and reducing uncertainty in risk estimates.

Although NRC (2008) focused on the antiandrogenic effects of phthalates, the committee acknowledged
that other health effects of phthalates may also be important. For example, liver toxicity, female
reproductive toxicity, and neurodevelopmental outcomes have also been observed following exposure to
some phthalates (as discussed in (ATSDR. 2022; « t v \ , U.S. CPSC. 2014)). Further, stakeholders
have urged EPA to consider assessing phthalates for cumulative risk based on not just their
anti androgenic effects on the male reproductive system, but also on the growing epidemiologic evidence
of adverse neurodevelopmental outcomes (see Project TENDR and EarthJustice comments cited in
footnotes in Section 1). EPA will consider these and the other health effects of phthalates as part of the
individual phthalate risk evaluations. However, for these health effects, data appear more limited across
the high-priority and manufacturer-requested phthalates and effects tend to occur at higher doses than
observed for anti androgenic effects. For example, with the potential exceptions of DIDP and DINP,
recent phthalate risk assessments have concluded that the developing male reproductive system is more
sensitive than the liver for most phthalate diesters (EFSA. 2019). This is further supported by a recent
systematic review of DIBP animal toxicology studies conducted by EPA researchers in the Center for
Public Health and Environmental Assessment (CPHEA), who found only slight evidence for female
reproductive toxicity and liver toxicity but robust evidence for DIBP-induced male reproductive toxicity
{i.e., phthalate syndrome) (Yost et al.. 2019). Additionally, the Agency for Toxic Substances and
Disease Registry (ATSDR) recently identified neurodevelopmental effects in rodent models as a
sensitive outcome following acute developmental exposures to DEHP (ATSDR. 2022). However,
ATSDR also identified inconsistencies in the toxicological database and refrained from using this health
outcome as the basis of a minimal risk level due to uncertainty in the database (see Appendix A [p. A-9]
of (ATSDR. 2022) for further details).

Additionally, EPA CPHEA researchers recently conducted a systematic review and meta-analysis of
epidemiologic studies of five phthalates {i.e., DEHP, DINP, DBP, DIBP, BBP), which are also
undergoing TSCA risk evaluation, and concluded that there is limited evidence supporting an
association between prenatal phthalate exposure and neurodevelopmental outcomes such as cognition,
motor effects, behavior {e.g., attention-deficit/hyperactivity disorder [ADHD]), infant behavior, and
social behavior {e.g., autism spectrum disorder) (Radke et al.. 2020).

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Given the limitations and uncertainties discussed above, EPA believes that the most robust reasonably
available dataset to support conducting a human health CRA is based on phthalate syndrome. Other
health effects of the high-priority and manufacturer-requested phthalates will be evaluated as part of the
individual phthalate risk evaluations. Following completion of systematic review for the individual
phthalates, EPA may consider whether any new information would change this conclusion. Notably,
EPA's proposal to focus on the shared ability of phthalates to disrupt androgen action and cause a
common syndrome {i.e., phthalate syndrome) is consistent with the recommendations of the NRC (2008)
and with how other regulatory agencies {i.e., U.S. CPSC, Australia NICNAS, EFSA, Danish EPA, and
Health Canada) have evaluated phthalates for cumulative risk to human health (see Appendix A).
The remainder of Section 3.1 is organized as follows:

•	Section 3.1.1, Phthalate Syndrome Mode of Action (MO A), provides a summary of the current
state of the science regarding the proposed MOA for phthalate syndrome.

•	Section 3.1.2, Key Outcomes for Grouping High-Priority and Manufacturer-Requested
Phthalates for CRA, provides a description of the key outcomes assessed by EPA to support the
proposed cumulative chemical group for CRA.

•	Section 3.1.3, Key Outcomes Data, provides a summary of data available for each of the high-
priority and manufacturer-requested phthalates underlying the key outcomes that EPA is
evaluating to support the proposed cumulative chemical group for CRA.

•	Section 3.1.4, Phthalate Syndrome in Humans, provides a summary of mechanistic explant and
xenograft studies investigating phthalate syndrome in human fetal testis tissue and outlines
several recent systematic reviews of human epidemiologic studies examining effects on the male
reproductive system.

•	Section 3.1.5, Species Differences in Sensitivity, provides a summary of differences in species
sensitivity to phthalate-induced male reproductive toxicity.

•	Section 3.1.6, Data Integration and Weight of Evidence Analysis, provides EPA's weight of
evidence narrative to support development of a cumulative chemical group for CRA.

•	Section 3.1.7, Proposed Conclusions on Toxicologic Similarity, summarizes EPA's proposed
conclusions on the toxicological similarity of the high-priority and manufacturer-requested
phthalates.

3.1.1 Phthalate Syndrome Mode of Action (MOA)

As can be seen from Figure 3-2, DEHP, DBP, BBP, DIBP, DCHP, DINP, and DIDP are structurally-
related ortho phthalate diesters with varying length linear or branched alkyl or aryl ester chains.
Gestational and/or postnatal exposure to certain structurally-related phthalates can lead to a spectrum of
effects on the developing male reproductive system, known as phthalate syndrome. Phthalate syndrome
is characterized by both androgen-dependent and -independent effects on the male reproductive system.
The MOA for rat phthalate syndrome has been discussed by various organizations (NA.SEM. 2017;
NRC. 2008). regulatory agencies (Health Canada. 2015; U.S. CPSC. 2014). and other research groups
(Gray et al. 2021; Arzuaea et al. 2020; Howdeshell et al..: ). To date, the MOA underlying
phthalate syndrome has not been fully established; however, key cellular-, organ-, and organism-level
effects are generally understood (Figure 3-3). Nevertheless, the molecular events preceding cellular
changes remain unknown. Although androgen receptor antagonism and peroxisome proliferator-
activated receptor alpha activation have been hypothesized to play a role, studies have generally ruled
out the involvement of these receptors (Foster. 2005; Foster et al.. 2001; Parks et al.. 2000).

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ch3

oa

CHj

Di-ethylhexyl Phthalate

(DEHP, CASRN 117-81-7)

Dibutyl Phthalate

(DBP, CASRN 84-74-2)

Di-isobutyl Phthalate

(D1BP, CASRN 84-69-5)

Butyl Benzyl Phthalate

(BBP, CASRN 85-68-7)

)

Dicyclohexyl Phthalate

(DCHP, CASRN 84-61-7)

Figure 3-2. Chemical Structures of Phthalates Being Evaluated under TSCA

Representative structures are shown for DIDP and DINP, which are isomeric mixtures with branched ester carbon
backbones varying in length (discussed further in Section 3.1.2.1).

Studies have demonstrated that gestational exposure to certain phthalate diesters, and their subsequent
hydrolysis to monoester metabolites, which occur during a critical window of development (i.e., the
masculinization programming window) can lead to antiandrogenic effects on the developing male
reproductive system (NRC, 2008). In rats, the masculinization programming window in which androgen
action drives development of the male reproductive system occurs between days 15.5 to 18.5 of
gestation, while the mouse critical window corresponds to gestational days 14 to 16, and the human
masculinization programming window is between gestational weeks 8 to 14 (MacLeod et al.. 2010;
Welsh et al.. 2008; Carruthers and Foster. 2005).

In vivo pharmacokinetic studies with rats have demonstrated that the monoester metabolites of DEHP,
DBP, BBP, and DINP can cross the placenta and be delivered to the target tissue, the fetal testes
(Clewell et al.. 2013a; Clewell et al.. 2010). In utero phthalate exposure can affect both Ley dig and
Sertoli cell function in the fetal testes. Histologic effects observed following phthalate exposure include
Ley dig cell aggregation and/or altered tissue distribution, as well as reductions in Leydig cell numbers.
Functional effects on Leydig cells have also been reported. Leydig cells are responsible for producing
hormones required for proper development of the male reproductive system, including insulin4ike
growth factor 3 (INSL3), testosterone, and dihydrotestosterone (DHT) (Scott et al.. 2009). Phthalate
exposure during the critical window reduces mRNA and/or protein levels of INSL3, as well as genes
involved in steroidogenesis, sterol synthesis, and steroid and sterol transport (Figure 3-3) (Gray et al..
2021; Hannas et al.. 2012).

Gene array experiments have demonstrated that phthalates known to disrupt testicular testosterone
production alter a distinct cluster of genes (Gray et al.. 2021). Key genes in this cluster are depicted in
Figure 3-3 and include reductions in mRNA for proteins involved in steroid hormone and sterol
transport (Scarbl, SlAR)\ testis steroid hormone biosynthesis (CypllAl, Hsd3b, Cypl7Al, l)hcr7)\
testicular testosterone and peptide hormone INSL3 syntheses (InsI3)\ pituitary stimulation of Leydig cell
testosterone synthesis (Lhcgr)\ testis development (Inha)\ and mRNA for enzymes involved in adrenal
hormone synthesis (e.g., Cypllbl, Cypllbl). Decreased steroidogenic mRNA expression leads to
decreased fetal testicular testosterone production, as well as reductions in DHT levels, which is

Diisononyl Phthalate	Diisodecyl Phthalate

(DINP, CASRNs 28553-12-0 (DIDP, CASRNs 26761-40-0
& 68515-48-0)	& 68515-49-1)

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produced from testosterone by 5a-reductase in the peripheral tissues. Because DHT is required for
growth and differentiation of the perineum and for normal apoptosis of nipple anlage in male rats,
reduced DHT levels can lead to phenotypic changes {i.e., nipple/areolae retention [NR] and reduced
anogenital distance [AGD] in males) indicative of reduced Leydig cell function and androgen action.

Chemical Structure
and Properties

Phthalate
exposure during
critical window of
development



Metabolism to
monoester &
transport to fetal
testes



Unknown MIE

(not believed to be
AR or PPARa
mediated)

^ Key genes involved in the AOP
for phthalate syndrome

Scarbl	Chcr7	Mvd	Elo3b

StAR	Ebp	Nsdhl	Insl3

Cypllal	Fdps	RGD1564999 Lhcgr

Cypllbl	Hmgcr	Tm7sf2	In ha

Cypllb2	Hmgcsl	Cyp46al	NrObl

Cypl7al	Hsd3b	Ldlr	RhoxlO

CypSl	Fldil	Insigl

I

Wnt7a I

y

Fetal Male Tissue

4, AR dependent
mRNA/protein
synthesis







4- Testosterone
synthesis

, O -

4- Gene
expression

(INSL3, lipid
> metabolism,

cholesterol and
androgen synthesis
and transport)

3E

4, INSL3 synthesis

Fetal Leydig cell

Abnormal cell
apoptosis/
proliferation

(Nipple/areolae
retention, 4- AGD,

Disrupted testis
tubules, Leydig cell
clusters, MNGs,
agenesis of
reproductive tissues)

Suppressed
gubernacular cord
development

(inguinoscrotal phase)

>=>





Suppressed
gubernacular cord
development

(transabdominal
	phase)	

Adverse Organism
Outcomes

4< Androgen-
dependent tissue
weights, testicular

pathology (e.g.,
seminiferous tubule

atrophy),
malformations (e.g.,
hypospadias), 4'
sperm production

0



Impaired





fertility





Undescended
testes



Figure 3-3. Hypothesized Phthalate Syndrome Mode of Action Following Gestational Exposure

Figure adapted from (Conlev et al.. 2021; Gray et al.. 2021; Schwartz et al.. 2021; Howdeshell et al.. 2017).
AR = androgen receptor; INSL3 = insulin-like growth factor 3; MNG = multinucleated gonocyte; PPARa =
peroxisome proliferator-activated receptor alpha.

Gestational exposure to certain phthalate diesters can also affect Sertoli cell function, development, and
interactions with germ cells contributing to seminiferous tubule degeneration ( ioekelheide et al.. 2009).
Immature Sertoli cells secrete Anti-Miillerian hormone and play an essential role in gonadal
development (Lucas-Iierald and Mitchell, 2022). Reported Sertoli cell effects include decreased Sertoli
cell numbers, changes in mRNA and/or protein levels of genes involved in Sertoli cell function, and
altered cellular development and Sertoli-germ cell interactions. Because proper Sertoli cell function is
necessary for germ cell proliferation and development, altered Sertoli cell function can contribute to
increased germ cell death, decreased germ cell numbers, and increased formation of multinucleated
gonocytes (MNGs) (Arzuaga et al.. 2020).

At the organ level, a disruption of androgen action can lead to reduced testes and accessory sex gland
(e.g., epididymis, seminal vesicle [SV], prostate, etc.) weight; agenesis of accessory organs; delayed
preputial separation (PPS); testicular pathology (e.g., interstitial cell hyperplasia); and severe
reproductive tract malformations such as hypospadias. INSL3 is crucial for gubernacular cord
development and the initial transabdominal descent of the testes to the inguinal region (Adham et al ..
2000). while androgen action is required for the inguinoscrotal phase of testicular descent. Thus,
reduced INSL3 and testosterone levels following gestational phthalate exposure can prevent
gubernaculum development and testicular descent into the scrotum. Collectively, these effects can lead

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to reduced spermatogenesis, increased sperm abnormalities, and reduced fertility and reproductive
function (Gray et ai. 2021; Arzuaga et ai. 2020; Howdeshell et al.. 2017; NASEM. _' l , \1'R€. 2008).

Postnatal exposure of male pups to phthalate diesters that cause phthalate syndrome following exposure
during the critical window of development can also lead to a disruption of Ley dig and Sertoli cell
function when exposure occurs at the peripubertal lifestage. The MOA for postnatal effects on male
reproduction is similar to the MOA for gestational effects, with some distinctions. EPA researchers in
CPHEA recently reviewed the MOA for DBP-induced male reproductive effects following postnatal
exposure (Arzuaga et al.. 2020). Briefly, cellular effects observed following peripubertal phthalate
exposure include altered Leydig cell development and function and reduced steroidogenic enzyme
expression and/or activity in the testes leading to reductions in testicular and/or serum testosterone
levels. In the seminiferous cord, effects on Sertoli and germ cells have also been observed—including
altered Sertoli cell development and function, altered interactions between Sertoli and germ cells, and
disrupted germ cell development. The molecular events preceding these cellular changes have not been
established. At the organ level, effects include incomplete development and/or reduced testes and
accessory sex gland weight, as well as a disruption (e.g., decreased organ weight, altered hormone
levels) of the hypothalamic-pituitary-gonadal axis, which plays an important role in the development
and function of the male reproductive system. Collectively, these effects can lead to decreased
spermatogenesis and male fertility (Arzuaga et al.. 2020).

3,1.2 Key Outcomes for Grouping High-Priority and Manufacturer-Requested Phthalates
	for CRA	

To determine which high-priority and manufacturer-requested phthalates are toxicologically similar and
appropriate for grouping for inclusion in a CRA, EPA reviewed studies that addressed seven key
outcomes associated with phthalate syndrome.7 The selected outcomes are not comprehensive of all the
effects associated with phthalate syndrome, but instead were selected to inform EPA's cumulative
chemical grouping for CRA based on EPA's current understanding of phthalate syndrome and its
underlying MOA. Notably, many of the key outcomes have also been selected as the critical effect (or
co-critical effect) in previous phthalate CRAs (Table 3-1). Key outcomes examined to support phthalate
grouping based on toxicologic similarity include

1)	Effects on fetal testicular expression of genes involved in steroidogenesis and Insl3 (Section
3.1.3.1). Reduced mRNA expression of cholesterol transport and steroidogenesis genes is
believed to play an early role in the development of phthalate syndrome. Reduced expression of
steroidogenic genes in the fetal testes leads to reduced testosterone production. Insl3 expression
was also selected to inform EPA's approach as it represents an androgen-independent
mechanism that contributes to the development of phthalate syndrome. INSL3 is crucial for
gubernacular cord development and the initial transabdominal descent of the testes to the
inguinal region (Adham et al.. 2000).

2)	Effects on fetal testicular testosterone (Section 3.1.3.2). Testosterone is an androgen produced
by fetal Leydig cells that is required for the proper development of the male reproductive system.

3)	Effects on anogenital distance (AGD) (Section 3.1.3.3). Under the Organisation for Economic
Co-operation and Development (OECD) guidance, decreased male AGD is considered a
hallmark of antiandrogenic substances and should be considered an adverse effect relevant for

7 The TSCA Work Plan includes one additional phthalate (i.e., di-n-octvl phthalate) that is not currently prioritised for risk
evaluation. However, Environment Canada/Health Canada (EC/HC. 2015e") concluded that di-n-octvl phthalate does not
induce effects on the developing male reproductive system consistent with phthalate syndrome. Di-n-octyl phthalate is not
further discussed in this document.

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setting the NOAEL (OECD. 2013). DHT is an androgen derived from testosterone by the
enzyme 5a-reductase. DHT functions to lengthen the perineum in fetal males relative to females.
Reduced AGD in males at birth is indicative of a disruption of androgen action during
development.

4)	Nipple/areolae retention (NR) (Section 3.1.3.4). NR in male rats is a biomarker of disrupted
androgen action during fetal development. During development, DHT, derived from testosterone
produced by Ley dig cells, is required for the normal regression of nipple anlage in male rats.
Disrupted fetal testicular testosterone production is believed to contribute to NR in male pups by
reducing DHT levels (Schwartz et at.. 2021). Under OECD guidance NR in male pups is
considered an adverse effect of exposure and should be considered relevant for setting the
NOAEL (OECD. 2013).

5)	Hypospadias (Section 3.1.3.5). Hypospadias is a malformation of the external male genitalia
in which the urethra does not open on the tip of the penis. As discussed in NASEM ( ),
mechanistic studies conducted with rats provide evidence that the formation of hypospadias (and
other male reproductive tract malformations) is linked with reduced testosterone production by
fetal Leydig cells (Howdeshell et at.. 2015).

6)	Seminiferous tubule atrophy/degeneration (Section 3.1.3.6). Germ cells develop into
spermatozoa in close proximity to Sertoli cells in seminiferous tubules. Seminiferous tubule
atrophy/degeneration is a pathologic lesion associated with phthalate syndrome frequently
reported following in utero exposure to certain phthalates. Although there is uncertainty
underlying the MOA associated with phthalate-induced effects on the seminiferous cord,
seminiferous tubule atrophy was selected to serve as a key outcome because it is a sensitive
adverse effect frequently reported by board-certified pathologists.

7)	Multinucleated gonocytes (MNGs) (Section 3.1.3.7). Phthalates can affect Sertoli cell
function, development, and interactions with germ cells. Proper Sertoli cell function is necessary
for germ cell proliferation and development and altered Sertoli cell function contributes to
increased germ cell death, decreased germ cell numbers, and increased formation of MNGs.
Although there is uncertainty underlying the MOA associated with MNG formation, induction of
MNGs is a sensitive indicator of exposure to a number of phthalates, and may serve as an
indicator of altered Sertoli-germ cell interactions (Spade et at.. JO IN, Spade et at.. 2014).

EPA's decision to focus its review on seven key outcomes associated with phthalate syndrome for
purposes of grouping phthalates for CRA under TSCA is consistent with the approach used by Health
Canada (EC/HC. 2015a; Health Canada.! ). Health Canada developed a chemical category of
phthalates for effects on the developing male reproductive system based on a structure-activity
relationship (SAR) analysis. The SARs analysis focused on three key outcomes associated with the
phthalate syndrome MOA, including effects on (1) steroidogenic gene expression, (2) fetal testicular
testosterone production, and (3) AGD. The chemical category was then assessed for cumulative risk to
human health (ECCC/HC. 2020). EPA's current approach expands on Health Canada's approach by
assessing several additional key outcomes associated with phthalate syndrome, including testicular
INSL3 mRNA expression, NR, hypospadias, seminiferous tubule atrophy, and MNG formation.

Through EPA's systematic review of the individual phthalates additional key outcomes may be
identified and EPA will assess these additional outcomes for relevance for inclusion in the CRA.

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Table 3-1. Summary of Critical Effects Selected for Use in Previous Phthalate CRAsa&

Regulatory
Agency

DEHP

BBP

DBP

DIBP

DCHP

DINP

Danish EPA

I testes weight,

| AGD

I spermatocyte

| AGD, NR

—

—

(ECHA. 2011)

seminiferous
tubule atrophy



development





















Australia
NICNAS

I testes weight,
seminiferous

I testicular
testosterone

I testicular
testosterone

—

—

1 testicular
testosterone

(2015a. 2013.
2012)

tubule atrophy

























Health Canada

Small and/or

| AGD

I testicular

| AGD, NR,

| AGD,

MNGs,

(ECCC/HC.

aplastic



testosterone, fertility

I testicular

TP, t

Leydig cell

2020)

epididymis, TP
(including
tubular atrophy),
other rat
phthalate
syndrome effects



effects, I tubular &

interstitial

cell #, altered

seminiferous tubule

structure, effects on

spermatocyte

development

testosterone,
effects on
fertility

resorption

aggregation















EFSA (2019)

I testes weight,
seminiferous
tubule atrophy

| AGD

I spermatocyte
development





I testicular

testosterone,

MNGs















U.S. CPSC
(2014)c

I Spermatocytes
& spermatids,
reproductive tract
malformations,
delayed vaginal
opening

J.AGD, NR

J.AGD, NR

J.AGD



NR

" Effects highlighted in gray indicate overlap with key outcomes selected for review by EPA in this document.
b DIDP is not shown in this table because it has not been included in previous phthalate CRAs. Studies have demonstrated that
gestational exposure to DIDP does not disrupt development of the male reproductive system in a manner consistent with
phthalate syndrome.

c Case 3 point of departures identified by U.S. CPSC's de novo literature review are shown.

AGD = anogenital distance; CPSC = Consumer Product Safety Commission; EFSA = European Food Safety Authority; MNG
= multinucleated gonocytes; NR = nipple retention; TP = testicular pathology

3.1.2.1 Study Selection Strategy

As of the publication of this document, EPA has not completed its systematic review or data quality
evaluation for the high-priority and manufacturer-requested phthalates. Therefore, a systematic review
protocol was not employed by EPA to identify studies supporting the seven key outcomes assessed in
this document. Instead, EPA conducted targeted literature searches and reviewed several documents
prepared by various authoritative bodies and regulatory agencies (see list of documents in Section 1.1) to
identify studies that support each of the seven key outcomes.

EPA focused its review on studies that were conducted using in vivo models and included an exposure
that at a minimum covered the critical window of development (i.e., GDs 15.5 to 18.5 in rats; GDs 14 to
16 in mice). This included both guideline and non-guideline studies and may include prenatal exposure
studies, perinatal exposure studies, and single or multi-generation reproductive studies. Although, oral,
dermal, and inhalation exposure studies were considered, the only studies identified that covered the

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critical window were oral exposure studies. Lack of inhalation and dermal exposure studies is
considered a data gap and is discussed further in Section 3.1.6.5. The majority of studies identified were
conducted using rat models, however, studies conducted with other species (i.e., mouse, rabbit, and
primate) were also considered.

Finally, while DEHP, DIBP, DBP, BBP, and DCHP are discrete chemical substances, DIDP and DINP
are isomeric mixtures with multiple CASRNs. DIDP (CASRNs 26761-40-0 and 68515-49-1) is an
isomeric mixture with branched ester carbon backbones composed of 7 (approximately 0 to 10 percent)
or >8 (approximately 70 to 90 percent) carbons (ECHA 2013). Two different isomeric mixtures of
DINP are commercially available, including DINP-1 (CASRN 68515-48-0) and DINP-2 (CASRN
28553-12-0), which contain linear and branched ester carbon backbones composed of 6 (5 to 10 percent
for DINP-1 and -2), 7 (45 to 55 percent for DINP-1; 40 to 45 percent for DINP-2), or >8 (20 to 45
percent for DINP-1; 35 to 50 percent for DINP-2) carbons (ECHA 2013). In the final scope documents
for DINP (U.S. EPA 2021c) and DIDP (U.S. EPA 2021b). EPA determined that the two CASRNs for
DINP and DIDP should be treated as categories of chemical substances as defined in 15 U.S.C §
2625(c). Therefore, EPA considered studies of both CASRNs for DINP and DIDP relevant for
informing toxicological similarity.

3.1.2.2 Availability of Studies to Inform Key Outcomes

EPA reviewed the toxicology studies available for the high-priority and manufacturer-requested
phthalates. While the amount of available data varies for each phthalate, data for all the proposed key
outcomes were available for DEHP, BBP, DBP, DIBP, DCHP, DINP and DIDP, except data for MNGs
for DIDP (Table 3-2). Additionally, although EPA's review focused on studies that assessed seven key
outcomes, EPA extracted data for all phthalate syndrome-related effects reported in each reviewed
study. Tables summarizing all observed phthalate syndrome-related effects for each study and phthalate
can be found in Appendices B.2 (DEHP), B.3 (BBP), B.4 (DBP), B.5 (DIBP), B.6 (DCHP), B.7 (DINP),
and B.8 (DIDP).

Table 3-2. Summary of Studies Supporting the Proposed Key Outcomes

Key Outcome

DEHP

BBP

DBP

DIBP

DCHP

DINP

DIDP

[ Steroidogenic gene and Ins 13 expression
(Section 3.1.3.1)

S

s

s

s

s

s

X

[ Fetal testicular testosterone (Section 3.1.3.2)

S

s

s

s

s

s

X

[ Anogenital distance (Section 3.1.3.3)

s

s

s

s

s

s

X

Nipple retention (Section 3.1.3.4

s

s

s

s

s

s

X

t Hypospadias (Section 3.1.3.5)

s

s

s

s

s

X

X

Seminiferous tubule atrophy (Section 3.1.3.6)

s

s

s

s

s

s

X

t Multinucleated gonocytes (Section 3.1.3.7)

s

s

s

s

s

s

-

S Studies available, effects observed
x Studies available, no effects were observed

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3,1,3 Key Outcomes Data

3.1.3.1 Fetal Testicular Gene Expression

3.1.3.1.1 Cholesterol Transport and Steroidogenesis

An early step in the hypothesized MO A for phthalate syndrome is a disruption of expression of
cholesterol transport and steroidogenesis genes in the fetal testes. The molecular events preceding these
cellular changes are unknown. The testicular steroidogenesis pathway is depicted in Figure 3-4. The
scavenger receptor class B member 1 gene (scarbl) encodes the SR-B1 protein, which transports
cholesterol into Leydig cells. The steroidogenic acute regulatory protein (encoded by StAR gene)
transports cholesterol across the mitochondrial membrane, which is the rate-limiting step in testicular
steroidogenesis (Petrescu et ai. 2001). Cytochrome P450 family 11 subfamily A member (CYP11 Al,
also referred to as P450 side-chain cleavage enzyme [P450scc]) catalyzes the conversion of cholesterol
to pregnenolone, which is next converted to progesterone by 3-beta-hydroxysteroid dehydrogenase (3P-
HSD). Progesterone is then converted to androstenedione by CYP17A1, and then to testosterone by 17P-
HSD.

7-dehydrocholesterol ]

Cholesterol

(star)! (gmiTj)
*

Pregnenolone

Progesterone

17a-Hydroxy-progesterone

(fyi7flj)
Androstenedione
17P-BSD
Testosterone

Figure 3-4. Testicular Steroidogenesis Pathway

Adapted from Hannas et al. (2012).

Red circles indicate genes assessed as part of the fetal

testicular gene expression key outcome.

EPA identified 20 in vivo experimental studies published by multiple research groups that evaluated
fetal testicular expression of key cholesterol transport {i.e., Scarbl, SlAR) and steroidogenesis {i.e.,
Cypllal, 3bHSD, Cypl7Al) genes following exposure during the critical window of development
(Table 3-3). Identified studies were primarily conducted using rat models (18 rat studies and 2 mouse
studies). Studies evaluating steroidogenesis were available for DEHP (six rat studies), BBP (one rat
study), DBP (seven rat studies and one mouse study), DIBP (five rat studies and one mouse study),
DCHP (two rat studies), DINP (five rat studies), and DIDP (two rat studies).

Across available rat studies (conducted with multiple strains, including Sprague-Dawley [SD], Wistar,
and Long-Evans) of DEHP, BBP, DBP, DIBP, and DCHP, consistent dose-dependent decreases in
mRNA expression of cholesterol transport and steroidogenesis genes in fetal testes were observed
(Table 3-3; Figure Apx B-l). In a study by Hannas et al. (2011). SD and Wistar rats were orally

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exposed to DEHP during the critical window and then StAR and Cypllal mRNA was measured in the
fetal testis. Similar dose-dependent decreases in mRNA expression of both genes were observed for both
rat strains (Figure 3-5).

DEHP (mg/kg/d)	DEHP (mg/kg/d)

Figure 3-5. CYP11A and StAR mRNA Expression in SD and Wistar Rats

Adapted from (Harmas et al.. 2011).

<
z
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that caused a 50 percent response) values for cholesterol transport and steroidogenesis genes for each
phthalate (except DIDP) (Table 3-4). As can be seen in Table 3-4, estimated 95 percent confidence
intervals overlap for most genes across phthalates, which limits the comparisons that can be made.
However, several trends in the dataset are apparent. First, DCHP, DEHP, BBP, and DBP appear to be
consistently more potent than DIBP at reducing fetal testicular mRNA expression, while DINP is
consistently the least potent phthalate.

3.1.3.1.2 Insl3 mRNA Expression

INSL3 is crucial for gubernacular cord development and the initial transabdominal descent of the testes
to the inguinal region (Adham et al. 2000). while androgen action is required for the inguinoscrotal
phase of testicular descent. Reduced INSL3 and testosterone levels following gestational phthalate
exposure can prevent gubernaculum development and testicular descent into the scrotum. EPA identified
12 in vivo experimental studies published by multiple research groups that evaluated fetal testicular
expression of Insl3 mRNA following exposure to the high-priority and manufacturer-requested
phthalates (Table 3-3). All identified studies were conducted using rat models. Studies evaluating Insl3
mRNA were available for DEHP (6 rat studies), BBP (2 rat study), DBP (3 rat studies), DIBP (3 rat
studies), DCHP (2 rat studies), DINP (4 rat studies), and DIDP (2 rat studies).

Consistent, dose-dependent reductions in Insl3 mRNA were observed for DEHP, BBP, DBP, DIBP and
DCHP across the available rat studies, regardless of strain tested (Table 3-3 and FigureApx B-l). In a
study by Hannas et al. ( ), SD and Wistar rats were orally exposed to DEHP during the critical
window and then Insl3 mRNA was measured. A similar dose-response was observed for both strains,
indicating no strain-specific differences (Figure 3-6).

Q£

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1274 Table 3-3. Studies Evaluating Fetal Testicular Steroidogenic Gene and Ins3 mRNA Expression

Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

Scarbl

(NOEL/

LOEL,

mg/kg/d)c

StAR
(NOEL/
LOEL,

mg/kg/d)c

Cypllal
(NOEL/
LOEL,

mg/kg/d)c

3bHSD
(NOEL/
LOEL,

mg/kg/d)c

Cypl 7al
(NOEL/
LOEL,

mg/kg/d)c

Insl3
(NOEL/
LOEL,

mg/kg/d)c



(Gray et al..

Rat (HSD)

GD 14-18 (GD 18)

0, 100, 300,
600, 900

100/300

100/300

100/300

100/300

100/300

100/300



2021)

Rat (CRSD)

GD 14-18 (GD 18)

0, 100, 300,
600, 900

300/600

100/300

300/600

300/600

300/600

300/600



(Hannas et

Rat (SD)

GD 14-18 (GD 18)

0, 100, 300,
500, 625,
750, 875

~

300/500

300/500

~

~

500/625

DEHP

al.. 2011)

Rat (W)

GD 14-18 (GD 18)

0, 100, 300,
500, 625,
750, 875

—

300/500

300/500

—

—

300/500

(Wilson et
al.. 2004)

Rat (SD)

GD 14-18 (GD 18)

0, 750

-

-

-

-

-

None/750



(Saillenfait

et al.. 2013)

Rat (SD)

GD 12-19 (GD 19)

0, 50, 625

None/50

None/50

50/625

50/625

50/625

-



(Borch et
al.. 2006b)

Rat (W)

GD 7-21 (GD 21)

0, 10, 30,
100, 300

100/300

30/100

100/300

-

NE °

100/300



(Cultv et
al.. 2008)

Rat (SD)

GD 14-19 (GD 19)

0, 234, 469,
938

-

NE

234/469

-

234/469

234/469



(Lin et al..
2008)

Rat (LE)

GD 2-20 (GD 21)

0, 10, 100,
750

100/750

100/750

10/100

-

-

100/750



(Grav et al..
2021)

Rat (HSD)

GD 14-18 (GD 18)

0, 11,33,
100, 300,
600, 900

11/33

33/100

11/33

33/100

100/300

11/33

BBP

Rat (CRSD)

GD 14-18 (GD 18)

0, 100, 300,
600, 900

300/600

300/600

300/600

600/900

300/600

300/600



(Wilson et
al.. 2004)

Rat (SD)

GD 14-18 (GD 18)

0, 750

-

-

-

-

-

None/750

DBP

(Grav et al..
2021)

Rat (HSD)

GD 14-18 (GD 18)

0, 1, 10, 33,
50, 100,
300, 750

50/100

50/100

100/300

50/100

50/100

50/100

Page 38 of 209


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Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

Scarbl

(NOEL/

LOEL,

mg/kg/d)c

StAR
(NOEL/
LOEL,

mg/kg/d)c

Cypllal
(NOEL/
LOEL,

mg/kg/d)c

3bHSD
(NOEL/
LOEL,

mg/kg/d)c

Cypl 7al
(NOEL/
LOEL,

mg/kg/d)c

Insl3
(NOEL/
LOEL,

mg/kg/d)c

DBP

(Wilson et
al.. 2004)

Rat (SD)

GD 14-18 (GD 18)

0, 750

-

-

-

-

-

None/750

(Lehmann
et al.. 2004)

Rat (SD)

GD 12-19 (GD 19)

0,0.1, 1, 10,
50, 100, 500

0.1/1

10/50

10/50

None/0.1

100/500

100/500

(Strove et

Rat (SD)

GD 12-19 (GD 19)

0, 112, 582

None/112

None/112

None/112

-

None/112

-

al.. 2009)

GD 12-19 (GD 20)

0, 112, 582

112/582

NE

112/582

-

112/582

-

(Kuhl et al..
2007)

Rat (SD)

GD 18 (GD 19)

0, 100, 500

None/100

None/100

None/100

-

None/100

-

(Drake et
al.. 2009)

Rat (W)

el3.5-16.5 (el7.5)

0, 500

-

None/500

None/500

-

-

-

(Johnson et
al.. 2012)

Rat (SD)

GD 19 (lhpost
dose)

0, 500

-

NE

NE

-

NE

-

GD 19 (3 h)

0, 500

-

NE

NE

-

None/500

-

GD 19 (6 and 18 h)

0, 500

-

None/500

None/500

-

None/500

-

(Thompson
et al.. 2005)

Rat (SD)

GD 19(0.5, l.and 2
h post dose)

0, 500

NE

NE

NE

-

NE

-

GD 19 (3 h)

0, 500

NE

None/500

NE

-

NE

-

GD 19 (6, 12, 18,
and 24 h)

0, 500

None/500

None/500

None/500

-

None/500

-

(Gaido et

Mouse
(CD-I)

GD 18 (2, 4, 8 hours
after final dose)

0, 500

NE

NE

NE

-

NE

-

al.. 2007)

GD 14-17 (2 hours
after final dose)

0, 250

NE

NE

NE

-

NE

-

DIBP

(Grav et al..
2021)

Rat (HSD)

GD 14-18 (GD 18)

0, 100, 200,
300, 500,
600, 750,
900

100/200

100/200

200/300

100/200

100/200

200/300



Rat (CRSD)

GD 14-18 (GD 18)

0, 100, 300,
600, 900

-

-

-

-

-

-

Page 39 of 209


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Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

Scarbl

(NOEL/

LOEL,

mg/kg/d)c

StAR
(NOEL/
LOEL,

mg/kg/d)c

Cypllal
(NOEL/
LOEL,

mg/kg/d)c

3bHSD
(NOEL/
LOEL,

mg/kg/d)c

Cypl 7al
(NOEL/
LOEL,

mg/kg/d)c

Insl3
(NOEL/
LOEL,

mg/kg/d)c



(Hannas et
al.. 2011)

Rat (SD)

GD 14-18 (GD 18)

0, 100, 300,
600, 900

-

100/300

None/100

-

-

-

DIBP

(Hannas et
al.. 2012)

Rat (SD)

GD 14-18 (GD 18)

0, 100, 300,
600, 900

100/300

100/300

100/300

100/300

100/300

100/300



(Saillenfait
et al.. 2017)

Rat (SD)

GD 13-19 (GD 19)

0, 250

None/250

None/250

NE

NE

None/250

-



(Bobere et
al.. 2008)

Rat (W)

GD 17-21 (GD 19
or21)

0, 600

None/600

None/600

None/600

-

None/600

None/600



(Wans et
al.. 2017)

Mouse
(ICR)

GD 0-21 (PND21)

0, 450

-

NE

None/450

NE

NE

-



GD 0-PND 21
(PND21)

0, 450

-

NE

None/450

None/450

None/450

-



(Grav et al..
2021)

Rat (HSD)

GD 14-18 (GDI8)

0, 33, 100,
300, 600,
900

33/100

33/100

33/100

33/100

33/100

33/100

DCHP

Rat (CRSD)

GD 14-18 (GDI8)

0, 100, 300,
600, 900

-

-

-

-

-

-



(Li et al..
2016)

Rat (SD)

GD 12-21 (GD
21.5)

0, 10, 100,
500, 1,000

100/500

None/10

NE °

None/10

NE °

10/100



(Grav et al..
2021)

Rat (HSD)

GD 14-18 (GD 18)

0, 500, 750,
1,000, 1,500

None/500

None/500

None/500

None/500

None/500

None/500



(Hannas et
al.. 2011)

Rat (SD)

GD 14-18 (GD 18)

0, 500, 750,
1,000, 1,500

-

750/1,000

750/1,000

-

-

-

DINP

(Hannas et
al.. 2012)

Rat (SD)

GD 14-18 (GD 18)

0, 500, 750,
1,000, 1,500

None/500

None/500

None/500

None/500

None/500

None/500



(Li et al..
2015a)

Rat (SD)

GD 12-21 (GD
21.5)

0, 10, 100,
500, 1000

NE

100/500

10/100

10/100

10/100

None/10



(Adamsson
et al.. 2009)

Rat (SD)

el3.5-17.5 (el9.5)

0, 250, 750

-

NE

NE

NE

-

250/750 6

DIDP

(Grav et al..
2021)

Rat (CRSD)

GD 14-18 (GD 18)

0, 300, 750,
1,000, 1,500

NE

NE

NE

NE

NE

NE

Page 40 of 209


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Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

Scarbl

(NOEL/

LOEL,

mg/kg/d)c

StAR
(NOEL/
LOEL,

mg/kg/d)c

Cypllal
(NOEL/
LOEL,

mg/kg/d)c

3bHSD
(NOEL/
LOEL,

mg/kg/d)c

Cypl 7al
(NOEL/
LOEL,

mg/kg/d)c

Insl3
(NOEL/
LOEL,

mg/kg/d)c



(Hannas et
al.. 2012)

Rat (SD)

GD 14-18 (GD 18)

0, 500, 750,
1,000, 1,500

NE

NE

NE

NE

NE

NE

" Apparent dose-related decrease in gene expression however, statistical significance was not achieved.

h Adamsson et al. (2009) reoort a slieht. but statistically significant, increase in inRNA expression of Insl3 at the highest dose tested.
c NOEL/LOEL values reflect study authors statistical analysis (i.e., the LOEL is the lowest value where a statistically significant effect was observed).

= gene was not measured in the study; CRSD = Charles River Sprague-Dawley; e = embryonic day; GD = gestational day; HSD = Harlan Sprague-Dawley; LE =
Long Evans; NE = no effect; LOEL = lowest observed effect level; NOEL = no observed effect level; PND = postnatal day; SD = Sprague-Dawley; W = Wistar

1275

1276

Table 3-4. EI

>50 Values (mg/kg/day) for Reduced ml

tNA Expression of Steroidogenic Genes and Insl3

Phthalate

Star
ED50
(95% CI)

Scarbl
ED50
(95% CI)

Cypllal

ED50
(95% CI)

Cvpl7al

ED50
(95% CI)

3bHSD
ED50
(95% CI)

Insl3
ED50
(95% CI)

DCHP

99
(48, 202)

62
(40, 96)

129
(49, 338)

53
(30, 92)

95
(37, 244)

162
(97, 270)

DEHP

109
(33, 196)

120
(62, 178)

173
(102, 249)

134
(101, 168)

242
(80, 503)

158
(104,215)

BBP

77
(46, 129)

50
(20, 121)

126
(59, 266)

180
(129, 251)

164

(72, 372)

167
(65, 434)

DBP

247
(74, 824)

295
(111, 779)

367
(170, 793)

285
(186, 437)

530
(288, 974)

237
(149, 376)

DIBP

324
(201,523)

287
(159, 519)

407
(253, 654)

371
(219, 626)

595
(325, 1,089)

414

(261, 656)

DINP

592
(493, 709)

594
(440, 802)

1148

(862, 1,530)

802
(698, 921)

1,016
(750, 1,376)

1,537
(730, 3,236)

ED50 values and 95% confidence intervals (95% CI) were estimated usins data from dosc-rcsDonsc experiments conducted bv Grav et al. (2021) with Harlan SD rats
(Fieure Adx B-l). ED50s were calculated usins methods described in Grav et al. (2021).

1278

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3.1.3.2 Fetal Testicular Testosterone

Testosterone is necessary for the proper development of the male reproductive system and a disruption
of testosterone levels during the masculinization programming window {i.e., GDs 15.5 to 18.5 in rats
(Welsh et ai. 2008)) contributes to the spectrum of effects that make up phthalate syndrome. EPA
identified a large amount of in vivo experimental data (38 studies from multiple laboratories) that
support this key outcome (Table 3-5). Available studies have primarily been conducted using rat models
(34 rat and 4 mouse studies identified). DEHP (13 rat studies and 2 mouse studies), DBP (16 rat studies
and 1 mouse study) and DINP (9 rat studies) have the largest amount of data available, while fewer
studies are available for DIBP (5 rat studies and 1 mouse study), BBP (5 rat studies), DCHP (3 rat
studies) and DIDP (3 rat studies).

As can be seen in Table 3-5, available rat studies (conducted with Wistar, SD, and Long-Evans strains)
of DEHP, BBP, DBP, DIBP, and DCHP provide consistent evidence that gestational exposure during
the critical window of development leads to reduced fetal testicular testosterone and/or ex vivo fetal
testicular testosterone production. Notably, the effect on fetal testicular testosterone consistently
occurred in a dose-dependent manner (see FigureApx B-l, which presents dose-response data for the
five high-priority and two manufacturer-requested phthalates (Gray et ai. 2021)) and was large in
magnitude at the lowest LOEL identified for DEHP (28 percent decrease at 50 mg/kg/d (Saillenfait et
at.. 2013)). BBP (53 percent decrease at 100 mg/kg/d (Furr et at.. 2014)). DBP (40 percent decrease at
50 mg/kg/d (Lehmann et at.. 2004)). DIBP (55 percent decrease at 250 mg/kg/d (Saillenfait et at.. 2017))
and DCHP (25 percent decrease at 33 mg/kg/d (Gray et at.. 2021)). Time course experiments conducted
with SD rats have demonstrated rapid reductions in fetal testicular testosterone following phthalate
exposure (Johnson et at.. 2012; Thompson et at.. 2005). Thompson et al. reported a 50 percent reduction
in fetal testicular testosterone as early as 1 hour after a single gavage dose to 500 mg/kg DBP, while
Johnson et al. reported an approximate 60 percent reduction in testosterone starting 18 hours after a
single gavage dose of 500 mg/kg DBP.

The four available mouse studies (one each of DBP and DIBP and two of DEHP) provide somewhat
contrasting results. Gestational exposure during the critical window to 450 mg/kg/day DIBP reduced
postnatal testicular testosterone and ex vivo testicular testosterone production in ICR mice on PND 21
(Wane et al.. 2017); however, effects on testicular testosterone were not evaluated during the fetal
lifestage in this study. In contrast, exposure to 1,000 to 1,500 mg/kg/day MBP or DBP during the critical
window did not affect fetal testicular testosterone in C57B1/6J mice (Gaido et al.. 2007). Similarly,
gestational exposure of CD-I mice to doses of up to 500 mg/kg/day DEHP (Do et al.. 2012) or
C57B 1/6J mice to doses of 500 to 1,000 mg/kg/day MEHP did not affect testicular testosterone (Gaido
et at.. 2007).

For DINP, effects on testosterone from the nine available rat studies were slightly less consistent. In 7
out of 9 studies, gestational exposure to DINP throughout the critical window of development dose-
dependently reduced fetal testicular testosterone levels and/or ex vivo testosterone production (Table 3-5
and Figure Apx B-1). Two studies (Ctewett et al.. 2013b; Adamsson et al.. 2009) did not report an
effect on fetal testicular testosterone following gestational exposure at doses that caused an effect in
other studies {i.e., 720 to 750 mg/kg/day DINP). Inconsistencies may be due to differences in phthalate
potency {i.e., DINP is less potent than other phthalates at disrupting steroidogenic gene expression
(Table 3-4) and testosterone (Table 3-6)), as well as timing of when testosterone was measured. For
example, Clewell et al. (2013a) gavaged rats with up to 750 mg/kg/day DINP on GDs 12 to 19 and
measured testicular testosterone levels at 2 and 24 hours after the final dose. Testicular testosterone

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levels were reduced 50 to 65 percent in the two highest treatment groups two hours, but not 24 hours,
after the final dose indicating a transient effect on testosterone.

For DIDP, three studies (all conducted with SD rats) were identified that investigated effects on fetal
testicular testosterone production. All three studies consistently found that exposure to DIDP throughout
the critical window had no effect on ex vivo fetal testicular testosterone production at doses as high as
1,500 mg/kg/day (Table 3-5). This is consistent with studies showing DIDP not affecting mRNA
expression of steroidogenic genes (Section 3.1.3.1.1).

Differences in the potency of the high-priority and manufacturer-requested phthalates to reduce fetal
testicular testosterone in rats are apparent. ED50 values for reduced ex vivo fetal testicular testosterone
production for these phthalates are reported by Furr et al. (2014) and shown in Table 3-6. Similarly,

Gray et al. (2021) report dose-response studies evaluating ex vivo fetal testicular testosterone production
in rats for the high-priority and manufacturer-requested phthalates. EPA used this dose-response data
(Figure Apx B-l) to calculate ED50 values for reduced ex vivo fetal testicular testosterone production.
As can be seen from Table 3-6, estimated 95 percent confidence intervals overlap for some phthalates.
However, data from both Furr et al. and Gray et al. indicate that DCHP, DEHP, and DBP are slightly
more potent than BBP and DIBP at reducing fetal testicular testosterone production, while DINP is
clearly the least potent.

EPA's findings are consistent with a recent systematic review and meta-analysis conducted by NASEM
(2017). NASEM assessed experimental animal evidence for effects on fetal testicular testosterone
following in utero exposure to DEHP, BBP, DBP, DIBP, and DINP (DCHP not included in analysis)
using the systematic review methodology developed by the National Toxicology Program's (NTP)
Office of Health Assessment and Translation (OHAT). NASEM found high confidence in the body of
evidence and a high level of evidence that fetal exposure to DEHP, BBP, DBP, DIBP, and DINP is
associated with a reduction in fetal testosterone in rats. Furthermore, NASEM found a statistically
significant overall effect and linear trends in logio(dose) and dose, with an overall large magnitude of
effect (>50 percent), for DEHP, BBP, DBP, DIBP, and DINP in their respective meta-analyses. For
DEHP, NASEM found that data were amenable to conducting separate subgroup analyses of SD and
Wistar rat strains. Meta-analysis found that SD rats were slightly more sensitive to DEHP than Wistar
rats (Table 3-7). Benchmark dose (BMD) values based on benchmark response (BMR) values of 5 and
40 percent were calculated by NASEM and are shown in Table 3-7. A comparison of BMD values
indicates similar trends in potency as was observed based on ED50 values calculated by EPA.

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Table 3-5.

Studies Evaluating Feta

Testicular Testosterone

Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

Dose-
Response
Observed?

NOEL

(mg/kg/d)/

LOEL

(mg/kg/d)/

% Decrease
from Control
at LOEL



(Saillenfait ct al.. 2013) h

Rat (SD)

GD 12-19 (GD 19)

0, 50, 625

Yes

None

50

28%



(Furr et al.. 2014) "

Rat (SD)

GD 14-18 (GD 18)

0, 100, 300, 600,
900

Yes

None

100

21-63%c



(Grav et al.. 2021) °

Rat (HSD)

GD 14-18 (GD 18)

0, 100, 300, 600,
900

Yes

None

100

38%



Rat

(CRSD)

GD 14-18 (GD 18)

0, 100, 300, 600,
900

Yes

100

300

29%



(Cultv et al.. 2008) °

Rat (SD)

GD 14-20 (GD 20)

0, 117, 234, 469,
938

Yes

None

117

60% d



(Hannas et al.. 2011) °

Rat (SD)

GD 14-18 (GD 18)

0, 100, 300, 500,
625, 750, 875

Yes

100

300

39%



Rat (W)

GD 14-18 (GD 18)

0, 100, 300, 500,
625, 750, 875

Yes

100

300

50%

DEHP

(Howdeshell et al.. 2008) "

Rat (SD)

GD 8-18 (GD 18)

0, 100, 300, 600,
900

Yes

100

300

42%



(Borch et al.. 2006b)ab

Rat (W)

GD 7-21 (GD 21)

0, 10, 30, 100, 300

Yes

100

300

60-80% abd



(Borch et al.. 2004)ah

Rat (W)

GD 7-21 (GD 21)

0, 300, 750

Yes

None

300

10%abd



(Linet al.. 2008)b

Rat (LE)

GD 2-20 (GD 21)

0, 10, 100, 750

Yes

100

750

67%





Rat (SD)

GD 14-17 (GDI7)

0, 750

-

None

750

54%







GD 14-18 (GD 18)

0, 750

-

None

750

59%



(Parks et al.. 2000)"



GD 14-20 (GD 20)

0, 750

-

None

750

57%







GD 14-PND 2
(PND 2)

0, 750

-

None

750

42%



(Soade et al.. 2018) °

Rat (SD)

GD 17-21 (GD 21)

0, 750

-

None

750

62%



(Wilson et al.. 2004) °

Rat (SD)

GD 14-18 (GD 18)

0, 750

-

None

750

50% d



(Martino-Andrade et al..
2008) b

Rat (W)

GD 13-21 (GD 21)

0, 150

-

150

None

-

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Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

Dose-
Response
Observed?

NOEL

(mg/kg/d) ^

LOEL

(mg/kg/d) ^

% Decrease
from Control
at LOEL



(Do et al.. 2012)b

Mouse
(CD-I)

GD 9-18 (GD 18)

0, 0.0005, 0.001,
0.005, 0.5, 50, 500

No

500

None

-



(Gaido et al.. 2007)b

Mouse

GD 14-16 (GD 17)

0, 500 (MEHP)

No

500

None

-





(C57B1/6J)

GD 15-17 (GD 17)

0, 1000 (MEHP)

No

1000

None

-



(Furr et al.. 2014) "

Rat (SD)

GD 14-18 (GD 18)

0, 100, 300, 600,
900

Yes

None

100

53%









0, 11, 33, 100

No

100

None

-



(Howdeshell et al.. 2008)"

Rat (SD)

GD 8-18 (GD 18)

0, 100, 300, 600,
900

Yes

100

300

22%

BBP

(Grav et al.. 2021)°

Rat (HSD)

GD 14-18 (GD 18)

0, 11, 33, 100, 300,
600, 900

Yes

33

100

27%



Rat

(CRSD)

GD 14-18 (GD 18)

0, 100, 300, 600,
900

Yes

100

300

38%



(Soade et al.. 2018) °

Rat (SD)

GD 17-21 (GD 21)

0, 750

-

None

750

69%



(Wilson et al.. 2004) °

Rat (SD)

GD 14-18 (GD 18)

0, 750

-

None

750

80% d



(Lehmann et al.. 2004)h

Rat (SD)

GD 12-19 (GD 19)

0,0.1, 1, 10, 50,
100, 500

Yes

10

50

40% d



(Furr et al.. 2014) "

Rat (SD)

GD 14-18 (GD 18)

0, 33, 50, 100, 300

Yes

50

100

35%



(Grav et al.. 2021) °

Rat (HSD)

GD 14-18 (GD 18)

0, 1, 10,33,50,
100, 300, 750

Yes

50

100

32%

DBP

(Mahood et al.. 2007)h

Rat (W)

GD 13.5-20.5
(GD 21.5)

0, 4, 20, 100, 500

Yes

20

100

14% d



(Strove et al.. 2009)h

Rat (SD)

GD 12-19 (GD 20)

0, 112, 582

Yes

None

112

70%



(Howdeshell et al.. 2008)"

Rat (SD)

GD 8-18 (GD 18)

0, 33, 50, 100, 300,
600

Yes

100

300

34%







el2.5-20.5 (el7.5)

0, 100, 300, 900

Yes

100

300

75% d



(Li et al.. 2015b)b

Rat (W)

el2.5-20.5 (el9.5)

0, 100, 300, 900

Yes

300

900

50% d







el2.5-20.5 (e21.5)

0, 100, 300, 900

Yes

300

900

40% d

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Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

Dose-
Response
Observed?

NOEL

(mg/kg/d) ^

LOEL

(mg/kg/d) ^

% Decrease
from Control
at LOEL



(Martino-Andrade et al..
2008) b

Rat (W)

GD 13-21 (GD 21)

0, 100, 500

Yes

100

500

63%



(Kuhl et al.. 2007)b

Rat (SD)

GD 18 (GD 19)

0, 100, 500

Yes

100

500

85%



(Mvlchreest et al.. 2002)b

Rat (SD)

GD 12-21 (GD 18)

0, 500

-

None

500

66%



GD 12-21 (GD 21)

0, 500

-

None

500

74%



(MacLeod et al.. 2010)h

Rat (W)

el3.5-21.5 (e20.5)

0, 500

-

None

500

-



(Drake et al.. 2009)h

Rat (W)

el3.5-16.5 (el7.5)

0, 500

-

None

500

40% d



(Johnson et al.. 2012)h

Rat (SD)

GD 19(1, 3, and 6
hours post-dose)

0, 500

-

500

None

-





GD 19 (18 hours)

0, 500

-

None

500

60% d

DBP





GD 19 (0.5 hours
post-dose)

0, 500

-

500

None

-



(Thompson et al.. 2005)h

Rat (SD)

GD 19 (1, 2, 3, 6
hours)

0, 500

-

None

500

50% d







GD 19 (12, 18, 24
hours)

0, 500

-

None

500

75% d



(van den Driesche et al..

Rat (W)

el3.5-20.5 (e21.5)

0, 500, 750

Yes

None

500

70% d



2012) b

el3.5-20.5 (e21.5)

0, 750

-

None

750

35% d



(Soadc et al.. 2018)"

Rat (SD)

GD 17-21 (GD 21)

0, 750

-

None

750

75%



(Wilson et al.. 2004) °

Rat (SD)

GD 14-18 (GD 18)

0, 1000

-

None

1000

85% d







GD 14-16 (GD 17)

0, 1000 (MBP)

-

1000

None

-



(Gaido et al.. 2007)b

Mouse

GD 14-16 (GD 17)

0, 1500 (DBP)

-

1500

None

-



(C57B1/6J)

GD 15-17 (8 hours
post-final dose)

0, 1000 (MBP)

-

1000

None

-



(Saillenfait et al.. 2017)"

Rat (SD)

GD 13-19 (GD 19)

0, 250

-

None

250

55%

DIBP

(Grav et al.. 2021)°

Rat (HSD)

GD 14-18 (GD 18)

0, 100, 200, 300,
500, 600, 750, 900

Yes

200

300

66%

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Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

Dose-
Response
Observed?

NOEL

(mg/kg/d) ^

LOEL

(mg/kg/d) ^

% Decrease
from Control
at LOEL





Rat

(CRSD)

GD 14-18 (GD 18)

0, 100, 300, 600,
900

Yes

100

300

34%

(Hannas et al.. 2011) °

Rat (SD)

GD 14-18 (GD 18)

0, 100, 300, 600,
900

Yes

100

300

66%

(Howdeshell et al.. 2008) °

Rat (SD)

GD 8-18 (GD 18)

0, 100, 300, 600,
900

Yes

100

300

40%

(Borch et al.. 2006a)"h

Rat (W)

GD 7-20/21 (GD
20/21)

0, 600

-

None

600

90%a bd

(Wane etal.. 2017)*

Mouse
(ICR)

GD 0-21 (PND 21)

0, 450

-

None

450

50%"

GD 0-PND 21
(PND 21)

0, 450

-

None

450

50% bd

DCHP

(Grav et al.. 2021) °

Rat (HSD)

GD 14-18 (GD 18)

0, 33, 100, 300,
600, 900

Yes

None

33

25%

Rat

(CRSD)

GD 14-18 (GD 18)

0, 100, 300, 600,
900

Yes

None

100

41%

(Furr et al.. 2014)"

Rat (SD)

GD 14-18 (GD 18)

0, 100, 300, 600,
900

Yes

33

100

69%



0, 33, 100, 300

Yes

33

100

55%

(Li et al.. 2016)6

Rat (SD)

GD 12-21 (GD
21.5)

0, 10, 100, 500

Yes

10

100

38%

DINP

(Clewell et al.. 2013a)b

Rat (SD)

GD 12-19 (2 hours
post final dose)

0, 50, 250, 750

Yes

50

250

50%

GD 12-19 (24
hours post final
dose)

0, 50, 250, 750

No

750

None



(Bobere et al.. 2011)ah

Rat (W)

GD 7-PND 17 (GD
21)

0, 300, 600, 750,
900

	e

300

600

50% bd

(Grav et al.. 2021) °

Rat (SD)

GD 14-18 (GD 18)

0, 500, 750, 1,000,
1,500

Yes

None

500

29%

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Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

Dose-
Response
Observed?

NOEL

(mg/kg/d) ^

LOEL

(mg/kg/d) ^

% Decrease
from Control
at LOEL



(Hannas et al.. 2011)"

Rat (SD)

GD 14-18 (GD 18)

0, 500, 750, 1,000,
1,500

Yes

None

500

30%



(Furr et al.. 2014)"

Rat (SD)

GD 14-18 (GD 18)

0, 750

-

None

750

24-50%c



(Borch et al.. 2004)ah

Rat (W)

GD 7-21 (GD 21)

0, 750

-

None

750

65% ad
15%bd



(Li et al.. 2015a)b

Rat (SD)

GD 12-21 (GD
21.5)

0, 10, 100, 500,
1,000

Yes

500

1000

57%



(Adamsson et al.. 2009)b

Rat (SD)

el3.5-17.5 (el9.5)

0, 250, 750

No

750

None

-



(Clewell et al.. 2013b)b

Rat (SD)

GD 12-PND 14
(PND 49-50)

0, 56, 288, 720

No

720

None

-



(Hannas et al.. 2012) °

Rat (SD)

GD 14-18 (GD 18)

0, 500, 750, 1,000,
1,500

No

1500

None

-

DIDP

(Furr et al.. 2014)"

Rat (SD)

GD 14-18 (GD 18)

0, 500, 750, 1,000,
1,500

No

1500

None

-



(Grav et al.. 2021) °

Rat

(CRSD)

GD 14-18 (GD 18)

0, 300, 750, 1,000,
1,500

No

1500

None

-

"Ex vivo fetal testicular testosterone production measured.
h Testes testosterone level measured.

c Range reflects results from multiple studies conducted using the same doses and methods reported within the publication (Blocks 31-32 for DEHP; Blocks 1, 5, and
7 for DINP).

J Value estimated based on graphical presentation of data.

e Fetal testicular testosterone was significantly reduced at 600 mg/kg/d DINP and appear reduced at higher doses, however, the effect at higher doses was not
statistically significant. Testicular testosterone production appeared reduced by >50% at doses >300 mg/kg/d DINP, however, the effect was not statistically
significant due to variability in the control samples (Bobere et al.. 2011).

' NOEL/LOEL values reflect study authors statistical analysis (i.e., the LOEL is the lowest value where a statistically significant effect was observed).

CRSD = Charles River Sprague-Dawley; e = embryonic day; GD = gestational day; HSD = Harlan Sprague-Dawley; LE = Long Evans; LOEL = lowest-observed-

effect-level; NOEL = no-observed-effect-level; PND = postnatal day; SD = Sprague-Dawley; W = Wistar

1362

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1363

Table 3-6. EE

•50 Values for Reduced ex vivo Fetal Testicular Testosterone Production

Phthalate

ED50 (95% CI)

(mg/kg/day)
(Furr et al.. 2014)

ED50 (95% CI)
(mg/kg/day)
(Grav et al.. 2021Y1

DCHP

62 (40, 96)

91 (46, 180)

DEHP

121 (92,160)

143 (132, 156)

DBP

158 (101,248)

154 (88,268)

BBP

172(116, 257)

228 (150, 347)

DIBP

288 (248, 335)

275 (226, 334)

DINP

738 (617, 884)

918 (780, 1,081)

11ED50 values and 95% confidence intervals (95% CI) were estimated using data from dose-response
experiments conducted bv Grav et al. (2021) with Harlan Sprague-Dawlev rats (Figure Apx B-l).
ED50s calculated using methods described in Grav et al. (2021).

1364

1365

1366

Table 3-7. Summary of NASEM (2017) Systematic Review and Meta-Analysis Results for Effects on Fetal Testosterone

Phthalate

Database
Supporting
Outcome

Confidence
in Evidence

Evidence of
Outcome

Heterogeneity

Model with
Lowest AIC

BMDS
mg/kg/day
(95% CI)

BMD40C
mg/kg/day
(95% CI)

DEHP17

11 rat studies
& 1 mouse
study

High

High

I2 > 90%
(combined)

Linear
quadratic

15 (11,24)

160 (120, 240)

I2 > 95% (SD)

Linear
quadratic

13 (9, 23)

140 (100, 230)

I2 = 21% (W)

Linear
quadratic

23 (21,24)

230 (210, 240)

BBP

2 rat studies

High

High

I2 > 85%

Linear
quadratic

23 (13,74)

230 (140, 390)

DBP

12 rat studies

High

High

I2 > 80%

Linear
quadratic

12 (8, 22)

130 (85,210)

DIBP

2 rat studies

High

High

I2 > 60%

Linear

NDfe

270 (225, 340)

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Phthalate

Database
Supporting
Outcome

Confidence
in Evidence

Evidence of
Outcome

Heterogeneity

Model with
Lowest AIC

BMDS

mg/kg/day
(95% CI)

BMD40c
mg/kg/day
(95% CI)

DINP

4 rat studies

High

High

I2 > 20%

Linear
quadratic

76 (49,145)

701 (552, 847)

11 Meta-analyses were conducted for combined strain data, as well as individual Wistar (W) and Sprague-Dawley (SD) data.
b The 5% change was well below the ranee of the data (NASEM. 2017).

c NASEM (2017) calculated BMD40s for this endooint because "previous studies have shown that reproductive-tract malformations were seen in male rats
when fetal testosterone production was reduced by about 40%."

1367

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3.1.3.3 Anogenital Distance (AGD)

DHT is an androgen derived from testosterone by the enzyme 5a-reductase. In rodents, DHT functions
to lengthen the perineum in males relative to females. Reduced AGD in males at birth is indicative of a
disruption of testosterone production by Leydig cells and is considered a biomarker of disrupted
androgen action during development. Compared to rodent models, the role of androgen action on AGD
is less well established in humans. However, observational human data are consistent with androgen
action during gestation playing a role in lengthening the perineum in humans (reviewed in (Thankamony
et ai. 2016)). This is consistent with the conclusions of NASEM Q ). After reviewing available
mechanistic information, NASEM concluded that "androgen-dependent development of the male
reproductive tract and androgen-dependent AGD appear to be well conserved across mammalian species
(including humans)."

EPA identified a large number of in vivo experimental studies (55 studies total from multiple research
groups) that evaluated AGD following gestational exposure to the high-priority and manufacturer-
requested phthalates (Table 3-8). Available studies were of varying design {i.e., gestational, perinatal,
and multi-generation exposure studies) and were primarily conducted using rat models (51 rat and 4
mouse studies). DEHP (16 rat and 3 mouse studies) and DBP (18 rat studies) have the largest amount of
data available. Fewer studies investigating AGD are available for BBP (five rat studies), DIBP (three rat
studies and one mouse study), DCHP (five rat studies), DINP (six rat studies), and DIDP (one rat study).

Available experimental rat studies (conducted with Wistar, SD, and Long-Evans strains) of DEHP, BBP,
DBP, DIBP, and DCHP provide consistent evidence that gestational exposure during the critical window
leads to a dose-dependent reduction in male pup AGD (Table 3-8). Importantly, statistically significant
reductions in AGD were consistently observed for both absolute AGD {i.e., measured in mm) and body
weight normalized AGD {i.e., mm/body weight or mm/cube root of body weight) for DEHP, BBP, DBP,
DIBP, and DCHP—indicating that the effect on AGD was not due to differences in pup size or body
weight. One out of 14 rat studies of DEHP reported no effect on AGD (Martin rade et ai. 2008).
however, this study included only a single dose group {i.e., 150 mg/kg/day) at a level that inconsistently
reduced AGD across the other available studies of DEHP (Table 3-8).

Effects on AGD are less consistent across the four available mouse studies of DEHP and DIBP. One
study in which C57BL/6 mice were gavaged with 100 to 500 mg/kg/day DEHP on embryonic days 12 to
17 reported a dose-dependent decrease in absolute fetal male AGD starting at the lowest dose (Liu et ai.
2008). In contrast, AGD was not reduced in CD-1 mice gestationally exposed to up to 500 mg/kg/day
DEHP via gavage (Do et ai. 2012) or 5 mg/kg/day DEHP via diet (Pocar et ai. 2012) or to ICR mice
exposed to 450 mg/kg/day DIBP via diet (Wane et ai. 2017). However, in the study of DIBP, AGD was
evaluated on PND 21, which is considered a less sensitive timepoint for AGD evaluation because AGD
can be affected by growth and changes in body weight (OECD guidance recommends AGD be
measured between PND 0 to PND 4 (	13)). Studies evaluating AGD in mice for other

phthalates were not identified, and it is unclear whether inconsistencies across mouse studies are due to
strain differences in sensitivity or some other factor.

For DINP, there is inconsistent evidence of an effect on male pup AGD following exposure during the
critical window (Table 3-8). Two out of six rat studies reported reduced AGD following exposure to
DINP. Bob erg et ai (2011) dosed Wistar rats with 300 to 900 mg/kg/day DINP on GD 7 through PND
17 and reported a dose-dependent reduction in both absolute and bodyweight normalized AGD on PND
13 in the highest treatment group; however, the effect was no longer apparent at PND 90. In a second
study conducted by Clewell et ai (2013b). SD rats were dosed with 56 to 720 mg/kg/day DINP from

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GD 12 through PND 14. Absolute and bodyweight normalized AGD was reduced in a dose-dependent
manner at PND 14, but not at PNDs 2 or 49. Four additional studies conducted with SD rats found no
effect on AGD following exposure during the critical window to doses of DINP ranging from 750 to
1,165 mg/kg/day (Li et al.. 201 Clewell et al. 2013a; Masutomi et ai. 2003; Gray et al. 2000).
Inconsistent effects on AGD are consistent with DINP being a less potent antiandrogen, as demonstrated
by potency comparisons for effects on gene expression (Table 3-4) and testosterone (Table 3-6).

For DIDP, AGD has only been evaluated in one study, a two-generation reproduction study of SD rats
(Hushka et al.. 2001). Absolute AGD was unaffected in male pups of both the F1 and F2 generations
when exposed to 300 to 400 mg/kg/day DIDP (highest dose tested). This is consistent with DIDP having
no effect on fetal testicular expression of steroidogenic genes (Section 3.1.3.1) or fetal testosterone
(Section 3.1.3.2).

To support relative potency comparisons, EPA conducted preliminary dose-response modeling of data
from studies that reported reduced male pup AGD following gestational exposure to each of the high-
priority and manufacturer-requested phthalates. For this preliminary analysis, data for DEHP, DBP,
BBP, DIBP, and DCHP were modeled to estimate an ED50 value for each phthalate. DINP was not
included in the initial dose-response analysis because effects on AGD were generally not large enough
in magnitude to support an accurate ED50 prediction. As can be seen from Table 3-9, 95 percent
confidence intervals for ED50 estimates generally overlapped, which prohibits direct potency
comparisons. A comparison of ED50 values for reduced AGD with those for changes in testosterone and
gene expression indicate AGD is a less sensitive outcome.

EPA's findings are consistent with a recent systematic review and meta-analysis conducted by NASEM
(2017) (summarized in Table 3-10). NASEM evaluated experimental animal evidence for effects on
AGD following in utero exposure to DEHP, BBP, DBP, and DINP (DIBP, DCHP, and DIDP were not
included) using the systematic review methodology developed by NTP's OHAT. NASEM found high
confidence in the body of evidence and a high level of evidence that fetal exposure to DEHP, BBP, and
that DBP is associated with reduced AGD in male rats. For DINP, NASEM had very low confidence in
the body of evidence and determined that there was inadequate evidence to support an association.
Meta-analyses found statistically significant overall effects and linear trends in logio(dose) and dose for
DEHP, BBP, and DBP. Additional meta-analyses of mouse as well as Wistar and SD rat data were
conducted for DEHP. Wistar rats were found to be more sensitive than SD rats, which is in contrast to
what was observed for effects on testosterone (i.e., SD rats were slightly more sensitive than Wistar
rats). For mice, the overall effect was not statistically significant; however, significant linear trends in
logio(dose) and dose were reported and mice were found to be similarly sensitive to DEHP-induced
effects on AGD as SD rats.

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1453 Table 3-8. Studies Evaluating Anogenital Distance in Male Pupsa&		

Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

NOAEL

(mg/kg/d)h

LOAEL

(mg/kg/d)h

Dose-Response Data



(Christiansen et al..
2010)

Rat (W)

GD 7-PND 16 (PND 1)

0, 10, 30, 100, 300,
600, 900

None

10

3.7±0.1, 3.4±0.1, 3.4±0.1, 3.4±0.1,
3.4±0.1, 3.3±0.1, 3.2±0.1 (mm.
x±SE)



Rat (W)

GD 7-PND 16 (PND 1)

0, 3, 10, 30, 100

30

100

3.40±0.1, 3.4±0.1, 3.4±0.1,
3.4±0.1, 3.25±0.1 (mm, x±SE)



(Vo et al.. 2009)

Rat (SD)

GD 11-21 (PND 63)

0, 10, 100, 500

10

o

O

o

38±1.3, 37±0.9, 31±1.2, 36±2.5
(mm, x±SE)



(Liu et al.. 2008)

Mouse
(C57BL/6)

el2-17 (el9)

0, 100, 200, 500

None

O
O

o

0.208±0.01, 0.198±0.01,
0.193±0.01, 0.181±0.12 (mm,
x±SD)



(Grav et al.. 2009)

Rat (SD)

GD 8-PND 17 (PND 2)

0, 11,33, 100, 300

100

300 c

3.3±0.11, 3.2±0.05, 3.2±0.09,
3.2±0.05, 2.7±0.08 (mm, x±SE)

DEHP







0.12,0.78,2.4, 7.9,
23, 77, 592, 775
(Fl)

77 (Fl)

592 (Fl)crf

	a

(Therlmmune
Research Co monition.
2004) f

Rat (SD)

GD 0-21 (PND 1)

0.09,0.48, 1.4,4.9,
14, 48,391,543
(F2)

48 (F2)

391 (¥2)cd

	a









0.1,0.47, 1.4,4.8,
14, 46, 359 (F3)

46 (F3)

359(F3)cd

	a



(Jarfelt et al.. 2005)

Rat (W)

GD 7-PND 17 (PND 3)

0, 300, 750

None

300 c

4.6±0.1, 4.0±0.6, 3.8±0.4 (mm,
x±SE)



(Andrade et al..
2006b)

Rat (W)

GD 6-21 (PND 22)

0.015,0.045,0.135,
0.405, 1.215,5, 15,
45, 135, 405

135

405 c

	a



(Li et al.. 2013)

Rat (SD)

GD 12-19 (PND 1)

0, 500, 750, 1,000

500

750 e

0.63±0.12, 0.61±0.12, 0.55±0.06,
0.56±0.03 (mm/bw, x±SE)



(Saillenfait et al..
2009a)

Rat (SD)

GD 12-21 (PND 1)

0, 500

None

500 ce

1.32±0.08, 1.08±0.05 (mm/\W
x±SD)



(Howdeshell et al..
2007)

Rat (SD)

GD 14-18 (PND 3)

0, 500

None

500 c

	a

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Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

NOAEL

(mg/kg/d)h

LOAEL

(mg/kg/d)h

Dose-Response Data

DEHP

(Moore et al.. 2001)

Rat (SD)

GD 9-PND 21 (PND 1)

0, 375, 750, 1,500

375

750 ce

	a

(Lin ct al.. 2008)

Rat (LE)

GD 2-20 (GD 21)

0, 10, 100, 750

100

750 c

4.5:0.1. 4.3:0.1. 4.8:0.1. 4.1 : 0.1
(mm, x±SE)

(Parks et al.. 2000)

Rat (SD)

GD 14-PND 2 (PND 2)

0, 750

None

750 c

	a

(Borch et al.. 2004)

Rat (W)

GD 7-21 (PND 3)

0, 750

None

750 c

	a

(Grav et al.. 2000)

Rat (SD)

GD 14-PND 3 (PND 2)

0, 750

None

750 c

3.42±0.08, 2.41±0.08 (mm. x±SE)

(Cultv et al.. 2008)

Rat (SD)

GD 14-PND 0 (PND 60)

0, 234, 469, 700,
750, 938, 1,250

938

1,250 c

	a

(Martino-Andrade et
al.. 2008)

Rat (W)

GD 3-21 (GD 21)

0, 150

150

None c e

-

(Do et al.. 2012)

Mouse
(CD-I)

GD 9-18 (GD 18)

0, 0.0005, 0.001,
0.005, 0.5, 50, 500

500

None c

-

(Pocaret al.. 2012)

Mouse
(CD-I)

GD 0.5-PND 21 (PND
42)

0, 0.05, 5

5

None e

-

BBP

(Asoetal.,2005)/

Rat (SD)

GD 0-21 (PND 4)

0, 100, 200, 400
(Fl, F2)

400 (Fl)

None (Fl)ce

-

None (F2)

100 (F2)ce

2.12±0.16, 1.96±0.11, 1.94±0.16,
1.87±0.21 (lnm/fbw, x±SE)

(Ema et al.. 2003)

Rat (W)

GD 15-17 (GD 21)

0 , 167, 250, 375

167

250 ce

	a

(Tvl et al.. 2004)^

Rat (SD)

GD 0-21 (PND 0)

0, 50, 250, 750 (Fl)

50

250 c

2.06±0.03, 2.01±0.04, 1.89±0.02,
1.71±0.03 (mm, x±SE)

0, 50, 250, 750 (F2)

50

250 c

2.05±0.01, 2.05±0.02, 1.99±0.01,
1.77±0.03 (mm, x±SE)

(Nasao et al.. 2000)f

Rat (SD)

GD 0-21 (PND 0)

0, 20, 100, 500 (Fl)

100 (Fl)

500 (Fl)c

2.6±0.2, 2.6±0.2, 2.5±0.1, 2.4±0.3
(mm, x±SD)

(Grav et al.. 2000)

Rat (SD)

GD 14-PND 3 (PND 2)

0, 750

None

750 c

3.42±0.08, 2.53±0.09 (mm, x±SE)

DBP

(Mvlchreest et al..
1999)

Rat (SD)

GD 3-21 (PND 1)

0, 100, 250, 500,
750

250

500 c

	a

GD 12-21 (PND 1)

0, 100, 250, 500

100

250 c

	a

(Zhane et al.. 2004)

Rat (SD)

GD 1-PND 21 (PND 4)

0, 50, 250, 500

50

250 cd

	a

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Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

NOAEL

(mg/kg/d)h

LOAEL

(mg/kg/d)h

Dose-Response Data



(Li et al.. 2009)

Rat (W)

GD 6-PND 1 (PND 1)

0,31,94, 291,797

94

291 c

3.80±0.15, 3.67±0.13, 3.72±0.20,
3.59±0.22, 2.78±0.16 (mm. x±SD)



(Li etal.. 2015b)

Rat (W)

el2.5-20.5 (PND 2)

0, 100, 300, 900

100

300 c

3.0±0.3, 2.9±0.2, 2.5±0.3, 2.2±0.2
(mm, x±not specified)



(Mvlchreest et al..
1998)

Rat (SD)

GD 3-PND 20 (PND 1)

0, 250, 500, 750

250

500 c

	a



(Jians et al.. 2007)

Rat (SD)

GD 14-18 (PND 1)

0, 250, 500, 750

250

500 d

0.65±0.08, 0.64±0.08, 0.61±0.05,
0.59±0.03 (mm/bw, x±SD)



(Kim etal.. 2010)

Rat (SD)

GD 10-19 (PND 11)

0, 250, 500, 700

250

500 d

	a



(Drake et al.. 2009)

Rat (W)

el3.5-21.5 (>12 weeks)

0, 100, 500

100

500 c

	a



(Mvlchreest et al..
2000)

Rat (SD)

GD 12-21 (PND 1)

0, 0.5, 5, 50, 100,
500

100

500 c

	a



(Strove et al.. 2009)

Rat (SD)

GD 12-19 (GD 20)

0, 100, 500

100

500 c

1.95±0.28, 1.90±0.2, 1.67±0.18
(mm, x±SE)

DBP

(Martino-Andrade et
al.. 2008)

Rat (W)

GD 12-21 (GD 21)

0, 100, 500

None

100 e

2.04±0.03, 1.88±0.04, 1.79±0.04
(mm/^bw, x±SE)



(Barlow et al.. 2004)

Rat (SD)

GD 12-21 (PND 1)

0, 100, 500

100

500 c

	a



GD 12-21 (PND 180)

0, 100, 500

100

500 c

	a



(MacLeod et al.. 2010)

Rat (W)

el3.5-20.5 (e21.5)

0, 500

None

500 c

	a



el3.5-21.5 (PND 25)

0, 100, 500

100

500 c

	a



(Howdeshell et al..
2007)

Rat (SD)

GD 14-18 (PND 3)

0, 500

None

500 c

	a



(van den Driesche et

Rat (W)

el3.5-20.5 (e21.5)

0, 500, 750

None

500 c

	a



al.. 2012)°

Rat (W)

el9.5-20.5 (e21.5)

0, 500, 750

750

None c

-



(Ema et al.. 1998)

Rat (W)

GD 11-21 (GD 21)

0, 331, 555,661

331

555 cd

	a



(Clewell et al.. 2013b)

Rat (SD)

GD 12-PND 14 (PND 2)

0, 642

None

642 ce

2.27±0.04, 2.04±0.04 (inin/^bw,
x±SE)



GD 12-PND 14 (PND 14)

0, 642

None

642 ce

3.40±0.04, 3.11±0.04 (inin/^bw,
x±SE)

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Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

NOAEL

(mg/kg/d)h

LOAEL

(mg/kg/d)h

Dose-Response Data







GD 12-PND 14 (PND 50)

0, 642

642

None

-



(Lee et al.. 2004)

Rat (SD)

GD 15-PND 21 (PND 2)

0, 2, 14, 148, 712

148

712 c

3.7±0.2, 3.9±0.2, 3.8±0.3, 3.8±0.2,
3.0±0.1 (mm, x±SD)



(Saillenfait et al..
2008)

Rat (SD)

GD 12-21 (PND 1)

0, 125, 250, 500,
625

125

250 c

2.55±0.17, 2.44±0.15, 2.28±0.30,
2.02±0.13, 1.98±0.16 (mm, x±SD)

DIBP

(Saillenfait et al..
2017)

Rat (SD)

GD 13-19 (GD 19)

0, 250

None

250 e

1.77±0.07, 1.68±0.07 (mm/^bw,
x+SD)

(Borch et al.. 2006a)

Rat (W)

GD 7-20/21 (GD 20/21)

0, 600

None

600 cd

	a



(Wane etal.. 2017)

Mouse

GD 0-21 (PND 21)

0, 450

450

None

-



(ICR)

GD 0-PND 21 (PND 21)

0, 450

450

None

-



(Ahbab and Barlas.
2015)

Rat (W)

GD 6-19 (GD 20)

0, 20, 100, 500

None

20 cde

	a



(Li et al.. 2016)

Rat (SD)

GD 12-21 (GD 21.5)

0, 10, 100, 500

10

100 c

3.3±0.3, 3.0±0.5, 2.7±0.2, 2.6±0.2
(mm, x±SE)

DCHP

(Saillenfait et al..
2009b)

Rat (SD)

GD 6-20 (GD 21)

0, 250, 500, 750

None

250 e

1.66±0.07, 1.52±0.09, 1.47±0.09,
1.43±0.08 (mm/^bw, x±SD)

(Yamasaki et al..
2009)

Rat (SD)

GD 6-PND 20 (PND 4)

0, 20, 100, 500

100

500 e

1.90±0.15,-, 1.66±0.11
(mm/^bw, x±SD)h



(Hoshino et al.. 2005) '

Rat (SD)

GD 0-21 (PND 4)

0,21, 104,511 (Fl)

104 (Fl)

511 (Fl)e

2.2±0.22,2.2±0.21,2.1±0.15,
2.0±0.15 (mm/^bw, x±SD)



0, 21, 107, 534 (F2)

21 (F2)

107 (F2)e

2.1±0.15,2.0±0.13, 1.9±0.16,
1.9±0.13 (mm/^bw, x±SD)







GD 12-PND 14 (PND 2)

0, 56, 288, 720

720

None c e

-



(Clewell et al.. 2013b)

Rat (SD)

GD 12-PND 14 (PND 14)

0, 56, 288, 720

288

720 ce

	a

DINP





GD 12-PND 14 (PND 49)

0, 56, 288, 720

720

None c e

-



(Bobers et al.. 2011)

Rat (W)

GD 7-PND 17 (PND 13)

0, 300, 600, 750,
900

750

900 e

11.6±1.0, 11.4±0.8, 11.3±0.2,
11.3±0.8, 11.0±0.9 (mm/^bw,
x±SD)

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Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

NOAEL

(mg/kg/d) '¦

LOAEL

(mg/kg/d) '¦

Dose-Response Data







GD 7-PND 17 (PND 90)

0, 300, 600, 750,
900

900

None e

-

(Masutomi et al..
2003)

Rat (SD)

GD 15-PND 2 (PND 2)

0, 30,307, 1,165

1,165

None c

-

(Li et al.. 2015a)

Rat (SD)

GD 12-21 (GD 21.5)

0, 10, 100, 500,
1,000

1,000

None c e

-

(Gray et al.. 2000)

Rat (SD)

GD 14-PND 3 (PND 2)

0, 750

750

None c d

-

(Clewell et al.. 2013a)

Rat (SD)

GD 12-PND 14 (GD 20)

0, 50, 250, 750

750

None c e

-

DIDP

(Hushka et al.. 2001)

Rat (SD)

GD 0-21 (PND 0)

0, 15, 50, 165, 300-
400

300-100
(Fl, F2)

None c

-

" Dose-response observed, but data not extracted because data was only presented graphically or, in some cases, data was reported at the pup level and was not extracted for

the Diirooscs of this document (e.s., see (Therlmmune Research Corporation. 2004)).
b AGD not reported for all dose groups.
c AGD reporting metric: mm
d AGD reporting metric: mm/body weight
' AGD reporting metric: mm/^bodyweight

^Multi-generation reproduction study. F1 and F2 indicates pups produced by F0 and F1 parental generations, respectively.
g Statistical analysis of combined data from both studies indicates a significant effect at 10 me/me/dav (Christiansen et al.. 2010).

h NOAEL/LOAEL values reflect study authors statistical analysis (i.e., the LOAEL is the lowest value where a statistically significant effect was observed),
e = embryonic day; GD = gestational day; LOAEL = lowest-observed-adverse-effect-level; NOAEL = no-observed-adverse-effect-level; PND = postnatal day; SD =
Sprague-Dawley; W = Wistar

1454

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Table 3-9. Summary of EE

~50 Values for Reduced (% Control) Male AGD

Phthalate

ED50 (95% CI)
(mg/kg/day)

DCHP

1,128 (825,2,042)

DEHP

1,314 (1068, 1,846)

DBP

920 (775, 1,149)

BBP

813 (685, 1,002)

DIBP

777 (594, 1,177)

ED50 value indicates the dose at which male pup AGD was reduced to 50% of the
control value. A description of the methodology used to estimate the ED50 values is
provided in Appendix C.

1456

1457

1458	Table 3-10. Summary of NASEM (2017) Systematic Review and Meta-Analysis Results for Effects

1459	on AGD

Phthalate

Database

Confidence
in Evidence

Evidence of
Outcome

Heterogeneity

Model with
Lowest AIC

BMDS

(mg/kg/day) (95%
CI)

DEHP17

16 rat studies
& 3 mouse
study

High

High

I2 > 20%

Linear
quadratic

270 (180, 420)
(combined)

290 (170, >1,000)
(SD)

150 (100, 280) (W)

BBP

6 rat studies

High

High

I2 > 75%

Linear
quadratic

250 (160,380)

DBP

22 rat studies

High

High

I2 > 75%

Linear
quadratic

150 (120, 220)

DINPfe

4 rat studies

Very Low

Inadequate

-

-

-

"Meta-analyses were conducted for combined strain data, as well as individual Wistar (W) and Sprague-Dawley data.
b NASEM did not conduct a meta-analysis for DINP due to their conclusion of inadequate evidence for this outcome.

1460

Page 58 of 209


-------
1461

1462

1463

1464

1465

1466

1467

1468

1469

1470

1471

1472

1473

1474

1475

1476

1477

1478

1479

1480

1481

1482

1483

1484

1485

1486

1487

1488

1489

1490

1491

1492

1493

1494

1495

1496

1497

1498

1499

1500

1501

1502

1503

1504

1505

1506

1507

1508

PUBLIC COMMENT DRAFT - DO NOT CITE OR QUOTE

3.1.3.4 Nipple Retention

DHT is an androgen derived from testosterone by the enzyme 5a-reductase. DHT is necessary for proper
apoptosis and regression of nipple anlagen in male rats. Because phthalate exposure reduces fetal
testicular testosterone production, DHT levels in peripheral tissues are also reduced leading to retained
nipples/areolas (NR). EPA identified 26 in vivo experimental animal studies from multiple research
groups that evaluated NR in male pups following phthalate exposure during the critical window (Table
3-11). Available studies were of varying design {i.e., gestational, perinatal, multi-generation exposure
studies), but were all conducted using either SD or Wistar strains. DEHP (12 rat studies) and DBP (eight
rat studies) have the largest amount of data available. Fewer studies investigating NR are available for
BBP (two rat studies), DIBP (one rat study), DCHP (two rat studies), DINP (three rat studies), and
DIDP (one rat study).

As noted in Table 3-11, there is variability in how publications report NR (e.g., NR is reported as mean
number of nipples/areolas per male, incidence of males with NR, or mean percent of litters with males
with NR, etc.). Furthermore, publications may or may not distinguish between retained areolas versus
retained nipples. These discrepancies in data reporting can make comparisons between studies difficult.
However, across available studies a consistent dose-dependent increase in NR was observed for male
pups gestationally exposed to DEHP, BBP, DBP, DIBP, or DCHP when evaluated between PNDs 11 to
14, which is consistent with OECD recommendations for timing of when evaluation of this outcome
should occur (OECD. 2013). For one study of DEHP (Martino-Andrade et al. 20081 retained nipples in
male pups was not observed, however, the study tested a single dose level (i.e., 150 mg/kg/day), which
produced inconsistent effects on NR across the other available studies of DEHP. For DINP, rat studies
are somewhat inconsistent. Two studies conducted with Wistar and SD rats demonstrate a dose-related
increase in male NR at doses ranging from 750 to 900 mg/kg/day, while a third study found no increase
in NR at a high dose of 720 mg/kg/day.

Several studies have examined whether or not NR is a permanent malformation in adult male rats that
were gestationally or perinatally exposed to phthalates. Available studies consistently report permanent
nipples in adult male rats exposed to DEHP (Gray et al.. 2009; Saillenfait et al.. 2009a; Howdeshell et
al.. 20^ , ray et al.. 2000). BBP (Gray et al.. 2000). DBP (Clewell et d k nowdeshell et al..
20*.1 , Tn ..flow et al.. 2004) and DIBP (Saillenfait et al.. 2008). No studies were identified that evaluated
permanent nipples in adult male rats exposed to DCHP. For DINP, there is inconsistent evidence of
permanent nipples. Boberg et al. (. ) found that Wistar rats exposed to doses of DINP >750
mg/kg/day had increased NR at PND 13, but permanent nipples were not observed at PND 90, while
Gray et al. (2000) reported permanent nipples in two out of 52 adult (3 to 7 months of age) males
perinatally exposed to 750 mg/kg/day DINP.

For DIDP, NR has only been evaluated in one study—a two-generation reproduction study of SD rats
(Hushka et al.. 2001). No increase in NR was reported in F1 or F2 male pups following exposure to up
to 300 to 400 mg/kg/day DIDP (highest dose tested). This is consistent with DIDP having no effect on
fetal testicular expression of steroidogenic genes (Section 3.1.3.1), fetal testicular testosterone (Section
3.1.3.2), and AGD (Section 3.1.3.3).

To support relative potency comparisons, EPA conducted preliminary dose-response modeling of data
from studies that reported NR for males as the percent of males per litter showing any retained
nipples/areolas. For this preliminary analysis, data for DEHP, DBP, BBP, DIBP, and DCHP were
modeled to estimate the ED50 for each phthalate. DINP was not included in this preliminary analysis
because the two available studies either did not report data as percent of males per litter showing any

Page 59 of 209


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1509	retained nipples/areolas {i.e., Boberg et al. ( ) reported data as the number of nipples per male) or

1510	only tested one dose level (Gray et al. 2000) and do not support ED50 predictions. As can be seen from

1511	Table 3-11, 95 percent confidence intervals overlapped for some ED50 estimates; however, based on

1512	this initial analysis DEHP and DBP appeared to be more potent than DCHP, DIBP, and BBP at inducing

1513	male pup NR.

1514

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1515 Table 3-11. Studies Evaluating Nipple Retention in Male Pups

Phthalate

Refence

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

NOAEL

(mg/kg/d) k

LOAEL

(mg/kg/d*k

Dose-Response Data"



(Christiansen et al..
2010)

Rat (W)

GD 7-PND 16 (PND 12)

0, 10, 30, 100, 300,
600, 900

3

10 '

0.22±0.08, 3.14±0.94, 1.81±0.82,
1.23±0.68, 5.21±1.25, 4.63±1.72,
5.01±1.36 b







0, 3, 10, 30, 100

100

None'

0.38±0.92, 0.59±0.99, 1.13±1.26,
0.31±0.40, 0.86±1.23 b



(Grav et al.. 2009)

Rat (SD)

GD 8-PND 17 (PND 13)

0, 11, 33, 100, 300

100

300

0.7±0.4, 0.8±0.3, 0.3±0.1, 0.7±0.3,
2.9±0.6 b

11±5.5, 21±8.9, 10±4.7, 16±6.7,
55%±10.1c







GD 8-PND 17 (PNM 7)

0, 11, 33, 100, 300

100

300

0±0, 0.08±0.08, 0±0, 0.15±0.12,
1.22±0.41 b



(Jarfelt et al.. 2005)

Rat (W)

GD 7-PND 17 (PND 13)

0, 300, 750

None

300

0.1±0.2, 3.9±2.7, 5.2±1.7 b

DEHP

(Andrade et al..
2006b)

Rat (W)

GD 6-PND 21 (PND 13)

0, 0.015, 0.045,
0.135,0.405,
1.215,5, 15,45,
135,405

135

405

0/60, 0/45, 0/46, 0/54, 0/58, 0/63,
0/42, 0/50, 0/41,0/56, 13/41rf



(Vo et al.. 2009)

Rat (SD)

GD 11-21 (PND 13)

0, 10, 100, 500

100

500

0, 0, 0, 9.06±1.83 i(±SDl



(Saillenfait et al..

Rat (SD)

GD 12-21 (PND 12-14)

0, 500

None

500

0/43, 42/56 d



2009a)

GD 12-21 (PND 70-120)

0, 500

None

500

0/42, 25/54 d



(Howdeshell et al..

Rat (SD)

GD 14-18 (PND 14)

0, 500

None

500

6.3±6.3, 55.8±16.4%c



2007)



GD 14-18 (PNM 7-11)

0, 500

None

500

0, 41.3±16.7%c



(Moore et al.. 2001)

Rat (SD)

GD 9-PND 21 (PND 14)

0, 375, 750, 1500

None

375

Jg



(Borch et al.. 2004)

Rat (W)

GD 7-21 (PND 13)

0, 750

None

750

_g



(Grav et al.. 2000)

Rat (SD)

GD 14-PND 3 (PND 13)

0, 750

None

750

0,6.3=^.1*; 0, 86.9±5%c



GD 14-PND 3 (PNM 3-7 )

0, 750

None

750





(Martino-Andrade et
al.. 2008)

Rat (W)

GD 13-21 (PND 13)

0, 150

150

None

2/18, 5/26 d

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Phthalate

Refence

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

NOAEL

(mg/kg/d) k

LOAEL

(mg/kg/d*k

Dose-Response Data"



(Therlmmune
Research

Corporation. 2004)

Rat (SD)

GD 0-PND 13 (PND 13)

0.1,0.47, 1.4,4.8,
14, 46, 359 (F2)

46 (F3)

359 (F3)

0±1, 0±0, 0±0, Oil, o±o, o±o,
11±7%e-'



(Gray et al.. 2000)

Rat (SD)

GD 14-PND 3 (PND 13)

0, 750

None

750

0, 5.1=^.9*; 0, 70±11%c







GD 14-PND 3 (PNM 3-7)

0, 750

None

750



BBP

(Tvl et al.. 2004)

Rat (SD)

GD 0-PND 13 (PND 11-

13)

0, 50, 250, 750

250

(Fl, F2)

750

(Fl, F2)

0.07±0.04, 0.00±0.00, 0.02±0.00,
1.29±0.33 (Fl) *¦'















0.05±0.03, 0.12±0.04, 0.19±0.08,
3.14±0.50 (F2) *¦'



(Mvlchreest et al..
2000)

Rat (SD)

GD 12-21 (PND 14)

0, 0.5, 5, 50, 100,
500

50

100

9/134, 8/119, 13/103, 12/120,
44/141, 52/58 d



(Barlow et al.. 2004)

Rat (SD)

GD 12-21 (PND 13)

0, 100, 500

None

100

_g



GD 12-21 (PND 180)

0, 100, 500

100

500

_g



(Mvlchreest et al..
1999)

Rat (SD)

GD 12-21 (PND 14)

0, 100, 250, 500

100

250

0/57, 0/58, 35/62, 47/54 d



(Kim etal.. 2010)

Rat (SD)

GD 10-19 (PND 11)

0, 250, 500, 700

250

500

0/201, 0/53, 3/36, 31/55 d

DBP

















(Martino-Andrade et
al.. 2008)

Rat (W)

GD 13-21 (PND 13)

0, 100, 500

100

500

2/18, 5/31, 7/8 d



(Howdeshell et al..

Rat (SD)

GD 14-18 (PND 14)

0, 500

None

500

6.3±6.3, 41.3±18.7%c



2007)

GD 14-18 (PNM 7-11)

0, 500

500

None

0±0, 21.8±13.4%c



(Clewell et al..

Rat (SD)

GD 12-PND 14 (PND 14)

0, 642

None

642

1.8±0.4, 5.8±0.8 b



2013b)

GD 12-PND 14 (PND 50)

0, 642

None

642

0.6±0.4, 2.5±0.5 6



(Lee et al.. 2004)

Rat (SD)

GD 15-PND 21 (PND 14)

0, 2, 14, 148, 712

148

712

0, 4, 13, 15, 100% c

DIBP

(Saillenfait et al..

Rat

GD 12-21 (PND 12-14)

0, 125, 250, 500,
625

125

250

0/76, 0/78, 8/96, 47/79, 56/76 d

2008)

(SD)

GD 12-21 (PNW 11-21 or
16-17)

0, 125, 250, 500,
625

125

250

0/46, 0/40, 4/55, 24/44, 29/38 d

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Phthalate

Refence

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

NOAEL

(mg/kg/d) k

LOAEL

(mg/kg/d*k

Dose-Response Data"

DCHP

(Yamasaki et al..
2009)

Rat
(SD)

GD 6-PND 20 (PND 13)

0, 20, 100, 500

100

500

_g

(Hoshino et al.. 2005)

Rat (SD)

GD 0-PND 14 (PND 14)

0, 14, 70, 349 (Fl)

70

349

0, 0, 0, 16% (Fl)/-'



GD 0-PND 12 (PND 12)

0, 14, 72, 351 (F2)

72

351

0, 0, 18, 63% (F2) ^'



(Bobere et al.. 2011)

Rat (W)

GD 7-PND 17 (PND 13)

0, 300, 600, 750,
900

600

750

1.98±0.83, 2.00±0.64, 2.91±0.69,
3.14±1.21, 3.17±0.92 i(±SDl



GD 7-PND 17 (PND 90)

0, 300, 600, 750,
900

900

None

-

DINP

(Gray et al.. 2000)

Rat (SD)

GD 14-PND 3 (PND 13)

0, 750

None

750

0,0.11±0.09*; 0, 22.4±8.9%c



GD 14-PND 3 (PNM 3-7)

0, 750

None

750

0, 2/52 d



(Clewell et al..

Rat (SD)

GD 12-PND 14 (PND 14)

0, 56, 288, 720

720

none

_b



2013b)

GD 12-PND 14 (PND 49)

0, 56, 288, 720

720

none

_ b

DIDP

(Hushka et al.. 2001)

Rat (SD)

GD 0-PND 14 (PND 12-
14)

0, 15, 50, 165,
300-100 (Fl, F2)

300-100
(Fl, F2)

None (Fl,
F2)

o±o, o±o, o±o, o±o, o±o H±SD)

" Response data is provided for each respective treatment group included in the study, starting with the control response.
h Mean number of nipples/areolas per male. Unless otherwise indicated, variation is reported as ± SEM.
c Mean (± SEM) percent of males with nipples/areolas.

J Incidence of males with nipples/areolas to total number of examined animals.
e Mean (± SEM) percent of male pups per litter with retained nipples/areolas.

' Mean percent of litters with males with retained areolas.

g Dose-response observed, but data not extracted because data was only presented graphically.

h Studv authors reported that most DEHP and BBP cxooscd adult males had Dcrmancntlv retained ninnies, however, the effect is not auantified (Gray et al.. 2000).
' Statistical analysis of combined data from both studies indicates a significant effect at 10 me/me/dav (Christiansen et al.. 2010).

' Multi-generation reproduction study. F1 and F2 indicate pups produced by F0 and F1 parental generations, respectively.

k NOAEL/LOAEL values reflect study authors statistical analysis (i.e., the LOAEL is the lowest value where a statistically significant effect was observed).

GD = gestational day; LOAEL = lowest-observed-adverse-effect-level; NOAEL = no-observed-adverse-effect-level; PND = postnatal day; PNM = postnatal month;

PNW = postnatal week; SD = Sprague-Dawley; W = Wistar

1516

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Table 3-12. Summary of ED50 Values for Percent Males
per Litter with Retained Nipples/Areolas	

Phthalate

ED50 (95% CI)
(mg/kg/day)

DCHP

588 (324, 1,067)

DEHP

368 (275,491)

DBP

331 (240, 463)

BBP

749 (551,2,020)

DIBP

479 (366, 628)

ED50 values indicate the dose at which 50% of males per litter
had retained nipples/areolas. A description of the methodology
used to estimate the ED50 values is provided in Appendix C.

3.1.3.5 Hypospadias

As discussed by NASEM ( ), mechanistic studies conducted with rats provide evidence that link the
formation of hypospadias (and other male reproductive tract malformations) with reduced fetal
testosterone production by fetal Leydig cells (Howdeshell et al. 2015). EPA identified 27 in vivo
experimental studies conducted by multiple research groups that evaluated hypospadias in experimental
models. Available studies have primarily been conducted using rats (24 rat studies and 1 study
conducted each with mice, rabbits, and marmosets). DEHP (9 rat studies and 1 mouse study) and DBP
(9 rat studies, 1 rabbit study, and 1 marmoset study) have the most available data, while fewer studies
are available for BBP (3 rat studies), DIBP (1 rat study), DCHP (1 rat study), DINP (3 rat studies), and
DIDP (1 rat study).

For DEHP, available data are suggestive of a strain-specific difference in sensitivity. Across the six
available studies conducted with SD rats, consistent dose-related increases in hypospadias were
observed starting at doses as low as 100 mg/kg/day DEHP (Table 3-13). In contrast, no hypospadias
were observed in two studies in which Wistar rats were exposed to up to 405 mg/kg/day (Andrade et al..
2006b) or 900 mg/kg/day DEHP (Christiansen et al.. 2010). In a third study, a slight (3 percent) increase
in hypospadias was observed in Wistar rats administered 300, but not 750 mg/kg/day DEHP (Jarfelt et
al.. 2005). For DBP, consistent dose-related increases in hypospadias were observed across all available
studies of SD (6 studies) and Wistar (2 studies) rats. Furthermore, hypospadias were observed starting at
comparable levels of exposure to DBP across strains {i.e., the lowest LOAELs were 250 and 300
mg/kg/day for SD and Wistar rats, respectively) (Table 3-13). Presently, it is unclear why strain-specific
differences in sensitivity exist for DEHP, but not DBP, for hypospadias.

Sufficient studies are not available to assess whether or not strain differences in sensitivity exist for BBP
(three SD rat studies), DIBP (one SD rat study), or DCHP (one SD rat study). Regardless, the available
studies of BBP, DIBP and DCHP report consistent dose-related increases in hypospadias following
gestational exposure throughout the critical window (Table 3-13). In one two-generation study of BBP
(Naeao et al.. 2000). hypospadias were not observed at the highest dose tested {i.e., 500 mg/kg/day),
however, this dose is lower than that shown to induce hypospadias in other studies of BBP {i.e., 750
mg/kg/day), including a two-generation study (Tyl et al.. 2004).

In the one available mouse study, doses of DEHP ranging from 100 to 500 mg/kg/day caused a dose-
dependent increase in the hypospadias (Liu et al.. 2008). Similarly, in the one available rabbit study of
DBP, hypospadias were observed in 1 out of 17 male pups (representing eight litters) exposed to 400

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mg/kg/day (only dose level tested) (Higuchi et ai. 2003). In contrast, no hypospadias were reported in
11 male offspring originating from 9 pregnant marmosets treated with 500 mg/kg/day MBP from weeks
7 through 15 of gestation (McKinnell et ai. 2009).

No significant increases in hypospadias were observed in any of the three rat studies (one with Wistar
and two with SD rats) of DINP at doses as high as 720 to 900 mg/kg/day. This is consistent with
findings for steroidogenic gene expression, fetal testicular testosterone, AGD, and NR results, all of
which indicate DINP is a less potent antiandrogen than other phthalates.

For DIDP, no hypospadias were reported in the two available two-generation studies of SD rats at doses
as high as 600 mg/kg/day (Hushka et ai. 2001). which is consistent with DIDP not disrupting fetal
testicular steroidogenesis or causing reduced AGD and NR.

EPA's findings are consistent with a recent systematic review conducted by NASEM (2017). NASEM
evaluated experimental animal evidence of hypospadias following in utero exposure to DEHP, BBP, and
DBP (DINP, DIBP and DCHP were not included) using the systematic review methodology developed
by NTP's OHAT. For both DEHP (8 rat studies and 1 mouse study) and BBP (two rat studies), NASEM
concluded that there is moderate confidence in the body of evidence and a moderate level of evidence
that gestational exposure to DEHP and BBP are associated with an increase in hypospadias in male rats.
In part, NRC downgraded confidence in the body of evidence for DEHP due to unexplained
inconsistency in response across rat strains {i.e., SD rats were more sensitive than Wistar rats). For DBP
(eight studies in rats), NRC concluded that there is high confidence in the body of evidence and a high
level of evidence that gestational exposure to DBP is associated with hypospadias in male rats. NASEM
did not conduct a meta-analysis of incidence data for hypospadias.

To support relative potency comparisons, EPA conducted preliminary dose-response modeling of data
from studies reporting increased incidence of hypospadias in adult F1 males following gestational or
perinatal exposure. For this preliminary analysis, data for DEHP, DBP, BBP, DIBP, and DCHP was
modeled to estimate the ED50 for each phthalate. DINP was not included in this preliminary analysis
because statistically significant increases in hypospadias have not been reported in male rats following
gestational exposure to DINP. As can be seen from Table 3-14, 95 percent confidence intervals
overlapped for some ED50 estimates; however, based on this initial analysis, DIBP and DCHP appeared
more potent than DEHP, BBP, and DBP.

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Table 3-13

. Studies Evaluating

ncidence of Hypospadias

Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome Evaluation)

Doses
(mg/kg/d)

NOAEL

(mg/kg/d)/

LOAEL

(mg/kg/d)

/

Response
(% Males Affected)"

DEHP

(Grav et al.. 2009)

Rat (SD)

GD 8-PND 17 or 63 (PND 63-
65)

0, 11,33, 100, 300

33

100

0, 0, 0, 1.1, 1.4% h

(Liu et al.. 2008)

Mouse
(C57BL/6)

el2-17 (el9)

0, 100, 200, 500

None

100

0, 7.1, 14, 76%

(Howdeshell et al..
2007)

Rat (SD)

GD 14-18 (PNM 7)

0, 500

None

500

0, 1.9%

(Li et al.. 2013)

Rat (SD)

GD 12-19 (PND 1)

0, 500, 750, 1000

None

500

0, 11,31,37%

(Grav et al.. 2000)

Rat (SD)

GD 14-PND 3 (PNM 3-7)

0, 750

None

750

0, 42%

(Vo et al.. 2009)

Rat (SD)

GD 11-21 (PND 63)

0, 10, 100, 500

100

500

0, 0, 0, 100%

(Saillenfait et al..

Rat (SD)

GD 12-21 (PND 70-120)

0, 500

None

500

0, 14.8%

2009a)

GD 12-21 (PND 70-84)

0, 625

None

625

0, 37%

(Jarfelt et al.. 2005)

Rat (W)

GD 7-PND 17 (PND 22)

0, 300, 750

None

300

0, 3, 0%

(Andrade et al.. 2006b)

Rat (W)

GD 6-PND 21 (PND 33)

0.015,0.045,0.135,
0.405, 1.215, 5, 15,
45, 135, 405

405

None

-

(Christiansen et al..
2010)

Rat (W)

GD 7-PND 16 (PND 16)

0, 10, 30, 100, 300,
600, 900

900 g

None

-

0, 10, 30, 100

100 «

None

-

BBP

(Grav et al.. 2000)

Rat (SD)

GD 14-PND 3 (PNM 3-7)

0, 750

None

750

0, 29%

(Tvl et al.. 2004)c

Rat (SD)

GD 0-21 (PND 4)

0, 50, 250, 750

250 (F2)

750 (F2)

0, 0, 0, 7%

(Nasao et al.. 2000)c

Rat (SD)

GD 0-21 (PND 21-22)

0, 20, 100, 500

500 (Fl, F2)

None

-

DBP

(Mvlchreest et al..
1998)

Rat (SD)

GD 3-PND 20 (PND 100)

0, 250, 500, 750

None

250

0,3,21,43%

(Li et al.. 2015b)

Rat (W)

GD 12.5-20.5 (PND 63)

0, 100, 300, 900

100

300

0, 0, 23, 44%

(Mvlchreest et al..
1999)

Rat (SD)

GD 12-21 (PND 100-105)

0, 100, 250, 500

250

500

0, 0, 0, 40%

(Jians et al.. 2007)

Rat (SD)

GD 14-18 (PND 1)

0, 250, 500, 750

250

500

0, 0, 7, 41%

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Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome Evaluation)

Doses
(mg/kg/d)

NOAEL

(mg/kg/d) f

LOAEL

(mg/kg/d)

/

Response
(% Males Affected)"



(Mvlchreest et al..
2000)

Rat (SD)

GD 12-21 (PND 110-120)

0, 0.5, 5, 50, 100,
500

100

500

0, 0, 0, 0, 0, 9%







GD 12-21 (PND 180)

0, 100, 500

100

500

0, 0, 16%



(Barlow et al.. 2004)

Rat (SD)

GD 12-21 (PND 370)

0, 100, 500

100

500

0, 0, 22%







GD 12-21 (PND 540)

0, 100, 500

100

500

0, 0, 26%



(Drake et al.. 2009)

Rat (W)

el3.5-21.5 (>12 weeks)

0, 100, 500

100

500

0, 0,31%

DBP

(Clewell et al.. 2013b)

Rat (SD)

GD 12-PND 14 (PND 49-50)

0, 642

None

642

0.1, 11%



(Kim etal.. 2010)

Rat (SD)

GD 10-19 (PND 11)

0, 250, 500, 700

500

700

0, 0, 0, 47%



(Hieuchi et al.. 2003)

Rabbit

(Dutch-

Belted)

GD 15-29 (PNW 12)

0, 400

None

400

0, 5.9% d



(McKinnell et al.. 2009)

Marmoset

GW 7-15 (PND 1-5)

0, 500 (MBP)

500

None

-



GW 7-15 (PNM 18-21)

0, 500 (MBP)

500

None

-

DIBP

(Saillenfait et al.. 2008)

Rat (SD)

GD 12-21 (PND 76-122)

0, 125, 250, 500,
625

250

500

0, 0, 0, 11,56%

DCHP

(Yamasaki et al.. 2009)

Rat (SD)

GD 6-PND 20 (PND 70)

0, 20, 100, 500

100

500

0, 0, 0, 12.5%



(Bobers et al.. 2011)

Rat (W)

GD 7-PND 17 (PND 90)

0, 300, 600, 750,
900

900

None

-

DINP

(Grav et al.. 2000)

Rat (SD)

GD 14-PND 3 (PNM 3-7)

0, 750

750

None

-



(Clewell et al.. 2013b)

Rat (SD)

GD 12-PND 14 (PND 49)

0, 56, 288, 720

720

None

0.9, 0, 0, 2.4%e

DIDP

(Hushka et al.. 2001)c

Rat (SD)

GD 0-21 (PND 0)

0, 165, 300-100,
600

600

(Fl, F2)c

None

-

0, 15, 50, 165, 300-
400

300—4-00
(Fl, F2)c

None

-

" Response data is provided for each respective treatment group included in the study, starting with the control response.
b Combined data from PUB (cxooscd from GD 8-PND 65) and IUL (cxooscd from GD 8-PND 17) cohorts (Table 6 of (Gray et al.. 2009)).
c Multi-generation reproduction study. F1 indicates male pups produced by F0 mating pairs, while F2 indicates male pups produced by F1 mating pairs.
d One out of 17 male pups (representing 8 litters) manifested hypospadias.

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Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome Evaluation)

Doses
(mg/kg/d)

NOAEL

(mg/kg/d) f

LOAEL

(mg/kg/d)

/

Response
(% Males Affected)"

e One out of 111 control (0.9%) and two out of 84 high-dose (2.4%) male pups manifested hypospadias described as mild/slight. The effect was not significant.
' NOAEL/LOAEL values reflect study authors statistical analysis (i.e., the LOAEL is the lowest value where a statistically significant effect was observed). In some
cases statistical analyses were not reported. In these cases, the NOAEL reflects the lowest dose where no hypospadias were observed.

g Study authors report mild dysgenesis of the external genitalia in all dose groups. Moderate to severe dysgenesis of the external genitalia, which includes hypospadias,
was not reported. EPA interpreted this to indicate that no hypospadias were observed at any dose in the study (Christiansen et al„ 2010).

e = embryonic day; GD = gestational day; NOAEL = no-observed-adverse-effect-level; PND = postnatal day; PNM = postnatal month; SD = Sprague-Dawley; W =
Wistar

1588

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1591

1592

1593

1594

1595

1596

1597

1598

1599

1600

1601

1602

1603

1604

1605

1606

1607

1608

1609

1610

1611

1612

1613

1614

1615

1616

1617

1618

1619

1620

1621

1622

1623

1624

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Table 3-14. Summary of ED50 Values for Hypospadias

I'hlluihitc

i: 1)50 (<)?%( i)

DCHP

699 (631, 825)

DEHP

846 (804, 904)

DBP

958 (919, 999)

BBP

878 (829, 948)

DIBP

626 (603,653)

ED50 values indicate the dose at which 50% of males exhibited
hypospadias. A description of the methodology used to estimate the
ED50 values is provided in Appendix C.

3.1.3.6 Seminiferous Tubule Atrophy

Seminiferous tubule atrophy/degeneration is a pathologic lesion frequently reported in adult animals
following in utero exposure to certain phthalates. Although there is uncertainty underlying the
mechanisms associated with phthalate-induced effects on the seminiferous cord, seminiferous tubule
atrophy was selected to serve as a key outcome because it is a sensitive, adverse effect frequently
reported by board certified pathologists. EPA identified 22 in vivo experimental studies that evaluated
testicular pathology and reported seminiferous tubule atrophy following gestational exposure to the
high-priority and manufacturer-requested phthalates. All studies were conducted using rat models. Data
were available for DEHP (three studies), BBP (three studies), DBP (eight studies), DIBP (one study),
DCHP (two studies), DINP (five studies), and DIDP (one study).

As can be seen from Table 3-15, available studies consistently demonstrate that exposure to DEHP,
BBP, DBP, DIBP, and DCHP lead to a dose-dependent increase in incidence of seminiferous tubule
atrophy. Studies reporting seminiferous tubule atrophy are of varying design, and increased incidence of
seminiferous tubule atrophy has been reported consistently across studies utilizing different exposure
paradigms {i.e., gestational, perinatal, and one or two-generation continuous exposure studies). Notably,
studies have demonstrated that gestational exposure to DEHP on GDs 14 to 18 (Howdeshell et al.
2007): DBP on GDs 14 to 18 (Hotchkiss et al. 2010; Howdeshell et al. 2007) or GDs 12 to 21 ( xlow
et al.. 2004; Mylchreest et al.. 2000; Mvlchreest et al.. 1999); and DIBP on GDs 12 to 21 (Saillenfait et
al.. 2008) is sufficient to cause increased seminiferous tubule atrophy in adults. This demonstrates that
exposure during gestation is sufficient to cause tubular atrophy later in life, well after cessation of
exposure.

For DINP, effects on seminiferous tubule atrophy are less consistent. Three studies reported no
significant increase in seminiferous tubule atrophy at doses ranging from 577 to 1,165 mg/kg/day DINP
(Clewell et al.. 201 Uv Masutomi et al.. 2003; Waterman et al.. 2000). Bob erg et al. (JO I I) reported that
a "few animals had small areas of tubular degeneration in areas of focal Leydig cell hyperplasia."
However, the dose levels at which tubular degeneration was observed were not reported. A fifth study
reported low incidence of tubular atrophy in adult rats exposed to 750 mg/kg/day DINP on GD 14
through PND 3 (Gray et al.. 2000). For DIDP, no seminiferous tubule atrophy was reported in available
two-generation reproductive studies at doses as high as 600 mg/kg/day.

To support relative potency comparisons, EPA conducted preliminary dose-response modeling of data
for incidence of seminiferous tubule atrophy in adult F1 male rats following gestational or perinatal
phthalate exposure. For this preliminary analysis, data for DEHP, DBP, BBP, DIBP, and DCHP was

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1625	modeled to estimate the ED50 for each phthalate. DINP was not included in this preliminary analysis, as

1626	tubular atrophy was reported qualitatively and at a low incidence in the one available study in which this

1627	pathologic lesion was observed. As can be seen from Table 3-16, 95 percent confidence intervals

1628	overlapped for some ED50 estimates. However, based on this initial analysis, DIBP and DCHP appear

1629	to be slightly more potent than DEHP and BBP, while DBP appears to be the least potent.

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1630 Table 3-15. Studies Reporting Seminiferous Tubule Atrophy 		

Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

NO A EL4'
(mg/kg/d)

LOAEI/

(mg/kg/d)

Seminiferous Tubule Atrophy
(# Affected Males/Total Males
Examined or % Males Affected)
(Severity, if Reported)



(Therlmmune Research
Corooration. 2004)

Rat
(SD)

-°(F1 Adults)

0.09,0.48, 1.4, 4.9,
14, 48,391,543

48

391

0/10, 0/10, 0/10, 1/10, 1/10, 0/10, 10/10,
10/10

DEHP

(Grav et al.. 2009)

Rat
(SD)

GD 8-PND 17 (PNM 7)

0, 11,33, 100, 300

	C

	C

_b



(Howdeshell et al..
2007)

Rat
(SD)

GD 14-18 (7-11 months)

0, 500

None

500

0, 33%



(Aso et al.. 2005)

Rat
(SD)

-°(F1 Adults)

0, 100, 200, 400

200

400

1/24, 1/24, 3/24, 9/24

BBP

(Naeao et al.. 2000)

Rat
(SD)

-°(F1 Adults)

0, 20, 100, 500

100

500

0/10, 0/10, 0/10, 6/10



(Tvl et al.. 2004)

Rat
(SD)

-°(F1 Adults)

0, 50, 250, 750

250

750

3/30, 0/29, 4/28, 23/28



(Mvlchreest et al..
1999)

Rat
(SD)

GD 12-21 (PND 100-
105)

0, 100, 250, 500

100

250

5/51, 0/51, 3/55, 11/45 (Grade 1)h
1/51, 1/51, 0/55, 1/45 (Grade 2)
0/51, 0/51, 1/55, 2/45 (Grade 3)
2/51, 0/51, 5/55, 19/45 (Grade 4)



(Mvlchreest et al..
1998)

Rat
(SD)

GD 3-PND 20 (PND
100)

0, 250, 500, 750

None

250

_b



(Wine et al.. 1997)

Rat
(SD)

-° (F1 Adults)

0, 256-385, 509-794

None

256-385

1/10, 3/10, 8/10

DBP

(Hotchkiss et al.. 2010)

Rat
(SD)

GD 14-18 (PND 123-
135)

0, 250, 500, 750,
1,000

250

500

1/15, 0/4, 2/7, 5/5, 6/6



(Mvlchreest et al..
2000)

Rat
(SD)

GD 12-21 (PND 110-
120)

0, 0.5, 5, 50, 100, 500

100

500

5/134, 6/118, 3/103, 3/120, 5/140, 4/58
(Grade 1)h

0/134, 1/118, 0/103, 0/120, 1/140, 2/58
(Grade 2)

0/134, 0/118, 0/103, 0/120, 0/140, 0/58
(Grade 3)

0/134, 0/118, 0/103, 0/120, 1/140, 25/58
(Grade 4)



(Barlow et al.. 2004)'



GD 12-21 (PND 180)

0, 100, 500

100

500

0/60, 0/65, 22/45

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Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Doses
(mg/kg/d)

NOAEI/

(mg/kg/d)

LOAEI/

(mg/kg/d)

Seminiferous Tubule Atrophy
(# Affected Males/Total Males
Examined or % Males Affected)
(Severity, if Reported)





Rat
(SD)

GD 12-21 (PND 370)

0, 100, 500

100

500

2/61,0/61,20/74

GD 12-21 (PND 540)

0, 100, 500

100

500

0/45, 0/49, 20/35

(Howdeshell et al..
2007)

Rat
(SD)

GD 14-18 (PNM 7-11)

0, 500

None

500

0, 14%

(Clewell et al.. 2013b)

Rat
(SD)

GD 12-PND 14 (PND
49-50)

0, 642

None

642

2/24, 6/26

DIBP

(Saillenfait et al.. 2008)

Rat
(SD)

GDs 12-21
(PNW 11-12)

0, 125, 250, 500, 625

125

250

2/24, 0/20, 1/28, 3/22, 1/20 (Grade 1)'
0/24, 1/20, 1/28, 1/22, 0/20 (Grade 2)
0/24, 0/20, 2/28, 0/22, 2/20 (Grade 3)
0/24, 0/20, 1/28, 4/22, 0/20 (Grade 4)
0/24, 1/20, 2/28, 8/22, 17/20 (Grade 5)

DCHP

(Ahbab and Barlas.
2015)

Rat
(SD)

GD 6-19 (GD 20)

0, 20, 100, 500

None

20

0/10, 8/10, 10/10, 10/10

(Hoshino et al.. 2005)

Rat
(SD)

- (F1 Adults)

0, 18, 90, 457

90

457

1/20, 0/23, 2/20, 6/22 (slight)
0/20, 0/23, 0/20, 3/22 (severe)

DINP

(Grav et al.. 2000)

Rat (SD

GD 14-PND 3 (PNM 3-
7)

0, 750

None

750

_b

(Masutomi et al.. 2003)

Rat
(SD)

GD 15-PND 10 (PNW
11)

0, 30, 307, 1,165

1165

None

NAe

(Clewell et al.. 2013b)

Rat
(SD)

GD 12-PND 14 (PND
49-50)

0, 56, 288, 720

720

None

2/24, 1/20, 0/20, 1/20 d

(Bobers et al.. 2011)

Rat (W)

GD 7-PND 17 (PND 90)

0, 300, 600, 750, 900

900

None'

NA

(Waterman et al.. 2000)

Rat
(SD)

-°(F1 Adults)

0, 133-153, 271-307,
543-577

577

None

NA d

DIDP

(Hushka et al.. 2001)

Rat
(SD)

-°(F1 Adults)

0, 15, 50, 165, 300-
400

300-400

None

NA d



0, 165, 300-400, 600

600

None

NA d

h Incidence of lesion reported qualitatively.

c Grav et al. (2009) reoort mild to moderate testicular seminiferous tubular degeneration. however, doses at which lesions were observed are not stated.
d Histologic examination of testes revealed no significant increase in seminiferous tubule atrophy or any other testicular pathologies.

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Phthalate

Reference

Species

Exposure Window
(Time of Outcome
Evaluation)

Doses

NOAEI/

LOAEI/

Seminiferous Tubule Atrophy
(# Affected Males/Total Males

(Strain)

(mg/kg/d)

(mg/kg/d)

(mg/kg/d)

Examined or % Males Affected)
(Severity, if Reported)

e Tubule atrophy not reported. Testicular pathology limited to degeneration of meiotic spermatocytes and vacuolar degeneration of Sertoli cells in high dose males.

' Boberg et al. (2011) report "Testicular histology at PND 90 appeared unaffected although a few animals had small areas of tubular degeneration in areas of focal Leydig cell
hyperplasia." However, the doses at which this effect was observed were not reported.

NOAEL/LOAEL values as reported by study authors.
h Severity grades reflect the percentage of degenerated tubules (Grade 1 = less than 5%; Grade 2 = 6-20%; Grade 3 = 21-50%; Grade 4 = greater than 50%).

' Severity grades reflect the percentage of degenerated tubules (Grade 1 = less than 5%; Grade 2 = 5-25%; Grade 3 = 26-45%; Grade 4 = 46-85%; Grade 5 = 86-100%).
¦' Reported as unilateral testicular dysgenesis, which is defined as "areas of aberrant or immature seminiferous tubules associated with proliferative Leydig cells."
GD = gestational day; LOAEL = lowest observed adverse effect level; NOAEL = no observed adverse effect level; PND = postnatal day; PNM = postnatal month; PNW =
postnatal week; SD = Sprague-Dawley; W = Wistar

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Table 3-16. Summary of ED50 Values for Incidence of Seminiferous Tubule Atrophy

Phthalate

ED50 (95% CI)
(mg/kg/day)

DCHP

380 (350,412)

DEHP

472 (438, 508)

DBP

628 (576,683)

BBP

417 (392, 444)

DIBP

344 (313,377)

11ED50 values indicate the dose at which a 50% incidence of seminiferous tubule atrophy was
observed. A description of the methodology used to estimate ED50 values is provided in Appendix C.

3.1.3.7 Multinucleated Gonocyte (MNG) Formation

Phthalates can affect Sertoli cell function, development, and interactions with germ cells. Proper Sertoli
cell function is necessary for germ cell proliferation and development and altered Sertoli cell function
contributes to increased germ cell death, decreased germ cell numbers, and increased formation of
MNGs. There is uncertainty underlying the mechanisms associated with MNG formation; however, it
may serve as a biomarker of altered Sertoli-germ cell interactions (Spade et al.. 2018; Spade et al..
2014). EPA identified 24 in vivo experimental studies that evaluated MNG formation following
gestational exposure to the high-priority and manufacturer-requested phthalates. The majority of
available studies were conducted using rat models (22 rat studies, 1 mouse studies, and 1 marmoset
study). The most data was available for DEHP (seven rat studies) and DBP (nine rat studies, one mouse
study, and one marmoset study), while less data was available for BBP (one rat study), DIBP (one rat
study), DCHP (two rat studies) and DINP (four rat studies). No studies were available for DIDP.

As can be seen from Table 3-17, there is variability in how publications report MNGs, which makes
comparisons across studies difficult (e.g., this outcome may be reported as MNGs per testis or
seminiferous cross-section, incidence of animals with MNGs in testes, percentage of total germ cells
multinucleated, average number of nuclei per germ cell, etc.). However, the available rat studies
(conducted with both SD and Wistar rats) consistently demonstrate that gestational exposure to DEHP,
DBP, DCHP, and DINP can increase MNG formation in a dose-dependent manner (Table 3-17). One rat
study investigating MNGs is available each for BBP (Spade et al.. 2018) and DIBP (Borch et al.. 2006a).
and these studies only tested a single dose level (i.e., 600 mg/kg/day DIBP; 750 mg/kg/day BBP).
However, both studies reported marked increases in MNGs following gestational exposure to BBP or
DIBP. In one mouse study of DBP, a dose-dependent increase in the number of MNGs per seminiferous
cord cross-section was reported at all dose-levels tested. In contrast to the results observed in rats and
mice, MNGs were not observed in marmosets gestationally exposed to 500 mg/kg/day MBP (a dose that
causes MNGs in mice and rats), however, unusual clusters of undifferentiated germ cells were found in
two of six MBP-exposed animals (McKinnell et al.. 2009).

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1660 Table 3-17. Studies Reporting on the Incidence of MNGs			

Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Dose
(mg/kg/d)

NOEL

(mg/kg/d)!

LOEL

(mg/kg/d)!

Dose-Response
Data"

DEHP

(Borch et al.. 2006b)

Rat (W)

GD 7-21 (GD 21)

0, 10, 30, 100, 300

30

100

_b

(Nardelli et al.. 2017)

Rat (SD)

GD 8-PND 21 (PND 3)

0, 30, 300

30

300

-cf

(Andrade et al.. 2006a)

Rat (W)

GD 6-PND 21 (PND 144)

0, 5, 15, 45, 135,
405

135

405

_b

(Parks et al.. 2000)

Rat (SD)

GD 14-PND 2 (PND 2)

0, 750

None

750

_b

(Grav et al.. 2000)

Rat (SD)

GD 14-PND 3 (PND 3)

0, 750

None

750

_b

(Srade et al.. 2018)

Rat (SD)

GD 17-21 (GD 21)

0, 750

None

750

2, 76 d

(Marti no-Andrade et al..
2008)

Rat (W)

GD 13-21 (GD 21)

0, 150

150

None

None

BBP

(Soade et al.. 2018)

Rat (SD)

GD 17-21 (GD 21)

0, 750

None

750

2, 64 d

DBP

(Boekelheide et al.. 2009)

Rat (SD)

GD 12-20 (GD 21)

0.1, 1, 10, 30, 50,
100, 500

50

100'

	C

(Mahood et al.. 2007)

Rat (W)

GD 13.5-20.5 (GD 21.5)

0, 4, 20, 100, 500

20

100/

	c

(Strove et al.. 2009)

Rat (SD)

GD 12-19 (GD 19)

0, 112,582

None

112

0/8, 1/6, 1/6 e

GD 12-19 (GD 20)

0, 112,582

None

112

0/9, 2/7, 6/7 e

(Gaido et al.. 2007)

Mouse
(C57B16)

GD 16-18 (GD 19)

0, 250, 500

None

250

	c h

(Mvlchreest et al.. 2002)

Rat (SD)

GD 12-21 (GD 21)

0, 500

None

500

_b

(van Den Driesche et al..
2015)

Rat (W)

el3.5-21.5 (e21.5)

0, 500

None

500

0, 3.9%g

(Martino-Andrade et al..
2008)

Rat (W)

GD 13-21 (GD 21)

0, 100, 500

100

500

-Cf

(Clewell et al.. 2013b)

Rat (SD)

GD 12-PND 14 (PND 2)

0, 642

None

642

1/24, 21/21e

(Ferrara et al.. 2006)

Rat (W)

el3.5-17.5 (el7.5)

0, 500

500

None

None

el3.5-19.5 (el9.5)

0, 500

None

500

sf

el3.5-20.5 (e21.5)

0, 500

None

500

_pf

el3.5-21.5 (e21.5)

0, 500

None

500

-Cf

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Phthalate

Reference

Species
(Strain)

Exposure Window
(Time of Outcome
Evaluation)

Dose
(mg/kg/d)

NOEL

(mg/kg/d)!

LOEL

(mg/kg/d)!

Dose-Response
Data"







el3.5-21.5 (PND 4)

0, 500

None

500

-Cf

el9.5-20.5 (e21.5)

0, 500

None

500

-Cf

(Srade et al.. 2018)

Rat (SD)

GD 17-21 (GD 21)

0, 750

None

750

2, 60 d

(McKinnell et al.. 2009)

Marmoset

GW 7-15 (PND 1-5 or 18-
21 months)

0, 500

500

None

None

DIBP

(Borch et al.. 2006a)

Rat (W)

GD 7-20/21 (GD 21)

0, 600

None

600

1/10, 10/16 e

DCHP

(Ahbab andBarlas. 2015)

Rat (W)

GD 6-19 (GD 20)

0, 20, 100, 500

20

100

0/10, 2/10, 5/10, 9/10 e

(Li etal.. 2016)

Rat (SD)

GD 12-21 (GD 21.5)

0, 10, 100, 500

10

100

0.4, 2, 16, 21%-f

DINP

(Lietal.. 2015a)

Rat (SD)

GD 12-21 (GD 21.5)

0, 10, 100, 500,
1,000

10

100

-Cf

(Clewell et al.. 2013a)

Rat (SD)

GD 12-19 (GD 20)

0, 50, 250, 750

50

250

0, 0 ,0.75, 1.25 d

(Clewell et al.. 2013b)

Rat (SD)

GD 12-PND 14 (PND 2)

0, 56, 288, 720

56

288

1/24, 2/20, 7/20, 18/19 e

(Bobere et al.. 2011)

Rat (W)

GD 7-21 (GD 21)

0, 300, 600, 750,
900

300

600

0/7, 2/8, 3/5, 6/7, 6/6 e

" Response data is provided for each respective treatment group included in the study, starting with the control response.
b MNGs reported qualitatively in text.

c Dose-response observed, but data not extracted because data was only presented graphically.

''MNGs per testis cross-section.
e Incidence of animals with MNGs in testes.

^Percent seminiferous cords with MNGs.
g Expressed as a percentage of all germ cells.
h MNGs per seminiferous cord cross-section.

' NOEL/LOEL values reflect study authors statistical analysis (i.e., the LOEL is the lowest value where a statistically significant effect was observed),
e = embryonic day; GD = gestational day; LOEL = lowest-observed-effect-level; NOEL = no-observed-effect-level; PND = postnatal day; SD = Sprague-Dawley;
W = Wistar

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3,1,4 Phthalate Syndrome in Humans

As discussed by NRC (2008) and NASEM (2017). rat phthalate syndrome shows similarities with the
hypothesized human testicular dysgenesis syndrome fWohlfahrt-Veie et al. 2009). However, the
etiology of the human syndrome is unknown, and it is unclear if endocrine disrupting chemicals such as
phthalates play a role.

To help inform EPA's understanding of the human relevance of phthalate syndrome, EPA reviewed two
types of studies, (1) mechanistic explant and xenograft studies of human fetal testis tissue (discussed in
Section 3.1.4.1), and (2) human epidemiologic studies evaluating associations between phthalate
exposure and effects on the male reproductive system (discussed in Section 3.1.4.2). Several recent
systematic reviews of human epidemiology studies have been conducted by NASEM (2017) and EPA
CPHEA scientists (Radke et al.. 2018). These reviews include five of the high-priority and
manufacturer-requested phthalates, including DEHP, BBP, DBP, DIBP, and DINP; results from these
systematic reviews are the focus of discussion in Section 3.1.4.2. Neither review included DCHP, so
EPA further reviewed several risk assessments conducted by other regulatory agencies to identify
epidemiologic studies of DCHP (ECCC/HC. 2020; ECB \ S, I * i PSC. 2014. 2010e); however,
no epidemiologic studies of DCHP and male reproductive outcomes were identified.

3.1.4.1 Human Explant and Xenograft Studies

Several explant (Lambrot et al.. 2009; Hallmark et al.. 2007) and xenograft studies (van Den Driesche et
al.. 2015; Spade et a I JO I I; s. [eger et al.. 2012; Mitchell et al.. 2012) using human donor fetal testis
tissue have been conducted to investigate the antiandrogenicity of mono-2-ethylhexyl phthalate (MEHP;
a monoester metabolite of DEHP), DBP, and monobutyl phthalate (MBP; a monoester metabolite of
DBP) in a human model. Hallmark et al. (2007) dosed human fetal testis explants (obtained from four
donors during gestational weeks 15 to 20) with DBP or MBP for 24 to 48 hours and observed no effect
on basal, human chorionic gonadotropin (hCG) stimulated, or 22R-hydroxy-cholesterol (22-R-CHO)
stimulated testosterone production. In contrast, MBP reduced hCG (but not 22-R-CHO) stimulated
testosterone production and caused a slight but significant increase in Ley dig cell aggregate size in fetal
testes explants obtained from Wistar rats at GD 19.5. Similarly, Lambrot et al. (2009) observed no effect
on basal or luteinizing hormone stimulated testosterone production or expression of Insl3 and
steroidogenic genes (P450cl7, P450scc, SlAR) in human fetal testes explants (obtained from donors
between gestational weeks 7 to 12) exposed to MEHP for 3 days. However, the researchers did observe
decreased germ cell numbers and an increase in the number of apoptotic germ cells.

Two separate research groups have developed xenograft protocols to evaluate the effects of phthalates
on human fetal testis tissue. Mitchell et al. (2012) grafted human fetal testis tissue (obtained from 12
donors at 14 to 20 weeks of gestation) into castrate male CD-I nude mice, which were then gavaged
with 500 mg/kg/day DBP or MBP for up to 21 days. Treatment with DBP had no effect on host serum
testosterone or seminal vesicle weight after 21 days of exposure, while MBP had no effect on host
seminal vesicle weight. Treatment with MBP appeared to reduce host serum testosterone by around 50
percent; however, this effect was highly variable and was not statistically significant. Concurrent studies
in which Wistar rat fetal testis tissue (obtained on GD 17.5) was grafted into castrate male mice were
also conducted to help validate human results. After 4 days of oral exposure to 500 mg/kg/day DBP, a
trend in reduced host serum testosterone level, reduced host seminal vesicle weight, and reduced mRNA
expression of StAR and Cypllal in retrieved rat xenografts was observed. These affects are consistent
with a disruption of androgen action. In a subsequent study by the same research group, the effects of
DBP on Sertoli and germ cells in human xenografts was investigated (van Den Driesche et al.. ^ ).
Briefly, human fetal testis tissue (obtained from seven donors between gestational weeks 14 to 20) was

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grafted into male CD-I nude mice, which were then gavaged with 500 mg/kg/day DBP for 21 days. In
retrieved xenografts, DBP was found to reduce germ cell numbers, increase the incidence of MNGs,
while immunostaining demonstrated a disruption of Sertoli cell cytoplasm distribution in xenografts
with germ cell aggregation.

Concomitantly with the study by Mitchell et al. ( ), Heger et al. (2012) grafted human (obtained at
10 to 24 weeks of gestation), Fischer rat (obtained on GD 16), and CD-I mouse (obtained on GD 15)
fetal testis tissue into male Crl:NIH-Foxnlmunude rats. Hosts were then gavaged with 100, 250, or 500
mg/kg/day DBP for 1 to 3 days and effects on steroidogenesis were measured in retrieved xenografts.
For human xenografts, testosterone production could not be accurately measured and was only reported
to be "highly variable" by study authors, while no effect of DBP was observed on steroidogenic gene
expression. Similarly, DBP had no effect on testosterone production or steroidogenic gene expression in
mouse grafts; however, an increased incidence of MNGs was observed in both human and mouse grafts
following exposure to DBP. In contrast, a reduction in testosterone production and steroidogenic gene
expression as well as an increase in MNGs were observed in rat xenografts following exposure to DBP.
In a second study by the same research group, Spade et al. (2014) grafted human fetal testis tissue
(obtained from six donors at 16 to 22 weeks of gestation) into adult castrated male nude mice. Hosts
were then gavaged with 500 mg/kg/day DBP or 75 mg/kg/day abiraterone acetate (CYP17A1 inhibitor)
for 14 days. Treatment with DBP had no effect on host serum testosterone or progesterone levels, host
seminal vesicle, prostate or LABC weight, and microarray analysis indicated no widespread impact on
steroidogenic gene expression. In contrast, abiraterone reduced host serum testosterone levels as well as
SV, LABC, and prostate weight.

Collectively, human explant and xenograft studies suggest that human fetal testis tissue is not sensitive
to the antiandrogenic effects of MEHP, DBP, or MBP. These results call into question the relevance of
the rat model for use in human health risk assessment. However, there are limitations associated with
these studies, which have been discussed extensively (Arzuaga et al.. 2020; ECHA. 2017; EC/HC.
2015c; Albert and Jeeou. 2014; Habert et jI JO I I; 1 c. t _ JO I I). First, the majority of human fetal
testis tissue used in xenograft and explant studies was obtained from fetuses older than 14 weeks of
gestational age. Male programming of the testes occurs during gestational weeks 8 to 14 in humans
(MacLeod et al.. 2010); therefore, it is possible that effects on testosterone and steroidogenic gene
expression were not observed due to the age of the fetal material. However, in the only study that
utilized human fetal testis tissue obtained from donors between gestational weeks 7 to 12 (Lambrot et
al.. 2009). no effect on testosterone production or steroidogenic gene expression was observed in
explants following exposure to MEHP, which would seem to argue against this possibility. Additionally,
Hallmark et al. (2007) and Spade et al. (2014) exposed human fetal testis explants and xenografts,
respectively, to CYP17 inhibitors {i.e., ketoconazole and abiraterone) known to disrupt testicular
steroidogenesis and observed reductions in testosterone. These findings indicate that steroidogenesis can
be disrupted in human explants and xenografts, regardless of fetal age, at least through certain
mechanisms.

Secondly, compared to rat explant and xenograft studies in which fetal testis tissue was obtained using a
standard protocol and at a consistent gestational age, human fetal testis tissue was obtained from donors
of variable age and by more variable methods, which likely contributed to the observed variability.

Other potential issues raised with the human fetal testis explant studies (Lambrot et al.. 2009; Hallmark
et al.. 2007) include the short phthalate exposure durations {i.e., 1 to 3 days) that were necessary because
explants were only viable in vitro for a few days. This raises the possibility that longer exposure
durations that are more reflective of in utero phthalate exposure in humans might have resulted in an
effect on steroidogenesis ( ;rt and Jeeou. 2014). Further, other hormonal effects {e.g., the

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hypothalamic-pituitary-gonadal axis) that are known to play a role in testis development cannot be
accounted for using in vitro explant models (ECUA. 2017; EC/HC. 2015c; Habert et ai. 2014).

Another potential issue that has been raised with human xenograft studies is variability in the
testosterone assays. For example, in the study by Mitchell et al. (2012) there appeared to be a 50 percent
reduction in host serum testosterone after 21 days of exposure to MBP; however, this result did not
reach statistical significance due to variability and small sample size (N = 3-4). Similarly, results from
the testosterone assay by Heger et al. (2012) was also reported to be highly variable and study authors
did not quantitatively report results. Assay variability is likely in part due to small sample sizes and
inherent biological variability in human tissue, as well as due to use of pooled results from human fetal
testis tissue of varying ages.

NASEM (2017) attempted to address the human tissue sample size issue by conducting a meta-analysis
of human xenograft studies of DBP and MBP and serum testosterone (Spade et al.. 2014; Mitchell et al..
2012). Overall, NASEM observed a trend toward decreased serum testosterone (-14.5 percent [95
percent CI: -40.4, 22.6]); however, this effect was not statistically significant due to the low precision of
the estimate (see Figure 3-17 in NASEM (2017)).

Generally, results from human explant and xenograft studies suggest that human fetal testes are not
sensitive to the antiandrogenic effects of phthalates, which has led some to conclude that rats are not an
appropriate model for use in human health risk assessment. However, as discussed above, human
explant and xenograft studies have limitations, and therefore the human relevancy of anti androgenic
effects of phthalates should not be ruled out. Notably, other authoritative agencies have drawn similar
conclusions regarding the human explant and xenograft studies, and concluded that the rat is an
appropriate model for use in human health risk assessment (ECHA. 2017; NASEN I < s , < v B'.€.
2015c; Is S (TSC.2014).

3.1.4.2 Epidemiologic Studies

Two recent systematic reviews investigating associations between phthalate exposure and effects on the
male reproductive system were evaluated (Radke et al.. 2018; NASEM. 2017). NASEM employed
NTP's OHAT systematic review methodology to evaluate the relationship between gestational exposure
to metabolites of DEHP, BBP, DBP, DIBP, and DINP (DCHP and DIDP not included) and several
outcomes, including decreased AGD, hypospadias, and testosterone concentrations during gestation or at
birth. For hypospadias and testosterone, NASEM identified a limited number of epidemiologic studies
and identified a number of confounding factors within the available studies, which led NASEM to
conclude that there was inadequate evidence to determine if fetal exposure to DEHP, DBP, DIBP, BBP,
or DINP is associated with hypospadias or a reduction in fetal testosterone. In contrast, NASEM
identified a number of prospective cohort studies investigating the effects of gestational phthalate
exposure on AGD at birth. NASEM found moderate confidence in the body of evidence for DEHP,
BBP, DBP, DIBP, and DINP and conducted further meta-analyses of each phthalate. Although the meta-
analyses found no statistically significant overall effect to support gestational exposure to BBP, DIBP,
and DINP being associated with reduced AGD, they found statistically significant estimates of 4 and 3
percent decreases in AGD for DEHP and DBP, respectively, per logio increase in exposure. These
findings led NASEM to conclude that there is a moderate level of evidence to support an association
between gestational exposure to DEHP and DBP and reduced AGD (Table 3-18).

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Table 3-18. Summary of NASEM (2017) Systematic Review and Meta-Analysis for Epidemiologic
Studies of AGD

Phthalate

# of
Studies

Confidence
in Evidence

Heterogeneity
(I2)

Summary Estimate
(% Change)
(95% CI)

Meta-Analysis
Conclusion

Evidence of
Outcome

DEHP

6 prospective

Moderate

0%

-4.07 (-6.49, -1.66)
(p = 0.001)

Evidence of
effect

Moderate

BBP

4 prospective

Moderate

0%

-1.43 (-3.47, 0.61)
(p = 0.17)

No evidence of
effect

Inadequate

DBP

4 prospective

Moderate

0%

-3.13 (-5.63,-0.64)
(p = 0.014)

Evidence of
effect

Moderate

DIBP

3 prospective

Moderate

0%

-2.23 (-5.15,0.70)
(p = 0.13)

No evidence of
effect

Inadequate

DINP

3 prospective

Moderate

58%

-0.96 (-4.17, 2.25)
(p = 0.56)

No evidence of
effect

Inadequate

In a second systematic review of human epidemiologic studies, Radke et al. (2018) evaluated the
strength of evidence supporting an association between phthalate exposure and male reproductive
effects. The review included DEHP, BBP, DBP, DIBP, and DINP (but not DCHP or DIDP) and focused
on outcomes such as AGD, semen parameters (i.e., concentration, motility, and morphology), time to
pregnancy (male exposure), testosterone, timing of pubertal development, hypospadias, and
cryptorchidism. Notably, both DEHP and DBP showed moderate evidence of an association between
gestational exposure and reduced AGD, while evidence of an association was slight for other evaluated
phthalates, which is consistent with the results of NASEM (2017). For other outcomes (i.e.,
hypospadias/cryptorchidism, testosterone in infants, and timing of pubertal development) associated
with gestational and/or childhood phthalate exposure, the study authors identified a limited number of
studies and concluded that there was slight or inadequate evidence to support an association (Table
3-19).

For outcomes associated with adult phthalate exposure, Radke et al. found (1) moderate evidence of
postnatal exposure to DBP and BBP and time to pregnancy; (2) moderate evidence of postnatal exposure
to DEHP, DINP, and DIBP and reduced testosterone levels in adults; and (3) moderate to robust
evidence of an association between DEHP, DINP, DBP, and BBP with effects on semen parameters
such as concentration, motility, and morphology (Table 3-19). As noted by the study authors, because
DEHP and DBP tended to have the most available studies and higher exposure levels compared to some
of the other evaluated phthalates, it may explain the difference in confidence in the strength of
associations for certain outcomes. Regardless, Radke et al. generally concluded that there is robust
evidence of an association between exposure to DEHP and DBP with male reproductive effects and
moderate evidence of an association for DINP, DIBP, and BBP.

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Table 3-19. Summary of Epidemiologic Evidence of Male Reproductive Effects Associated with
Phthalates"

Timing of
Exposure

Outcome

DEHP

DINP

DBP

DIBP

BBP

In utero

| AGD

Mb

(0/3/3)c

S

(0/3/0

M

(0/3/2)

S

(0/2/1)

S

(0/3/2)

Hypospadias/
cryptorchidism

I

(0/2/2)

S

(0/2/1)

S

(0/2/1)

S

(0/2/1)

S

(0/2/1)

In utero or
childhood

Testosterone in infants

I

(0/0/1)

I

(0/0/1)

I

(0/0/1)

I

(0/0/1)

I

(0/0/1)

Timing of pubertal development

I

(0/1/2)

I

(0/0/1)

I

(0/1/1)

I

(0/0/1)

I

(0/1/2)

Adult

Semen parameters

M

(0/12/2)

M

(0/3/0)

R

(0/11/1)

s

(0/4/0)

M

(0/9/1)

Time to pregnancy

S

(1/0/0)

I

(1/0/0)

M

(1/0/0)

S

(1/0/0)

M

(1/0/0)

Testosterone in adults

M

(0/9/4)

M

(0/5/0)

s

(0/7/3)

M

(0/2/2)

I

(0/6/2)

Overall Evic

ence

R

M

R

M

M

"Table adapted from Radke et al. (2018).

h Strength of evidence descriptors: R = robust (bolded and cell shaded dark gray); M = moderate (bolded and light
gray); S = slight; I = indeterminant. Robust and moderate descriptors indicate evidence supports a hazard based on
the quantity and quality of available information, which rules out alternative explanations for the results. Slight and
indeterminant descriptors indicate that evidence could support the presence or absence of a hazard and is typically
limited by quantity or confidence level in available studies.

c Numbers in parentheses indicate the number of high, medium, and low confidence studies, respectively, used as
part of the overall strength of evidence evaluation.

3.1.5 Species Differences in Sensitivity

Differences in species sensitivity to testicular toxicity of phthalate diesters has been recognized for
decades (Gray et al.. 1982) and has been discussed extensively by various authoritative agencies (e.g.,
see (NASEM. 2017)1 regulatory bodies (e.g., see (ECHA. 2017; U.S. CPSC. 2014)). and research
groups (e.g., see (Arzuaga et al.. 2020; Johnson et al.. 2012)). As discussed in Sections 3.1.3.1 to 3.1.3.6,
the majority of in vivo studies investigating the effects of gestational phthalate exposure have been
conducted using rat models. Available rat data provide consistent evidence that gestational exposure to
certain phthalates during the critical window of development can lead to a spectrum of effects on the
developing male reproductive system consistent with phthalate syndrome. Studies that investigated
phthalate syndrome following gestational exposure during the critical window are available for mice
(Wang et al.. 2017; Do et al.. 2012; Pocar et al.. 2012; Liu et al.. 2008; Gaido et al.. 2007). rabbits,
(Higuchi et al.. 2003). and marmosets (McKinnell et al.. 2009). Results from mouse and marmoset
studies are inconsistent with findings from rat studies and indicate species differences in sensitivity.
Mouse, rabbit, and marmoset data are discussed below, while available rat data are discussed further as
part of the data integration and weight of evidence analysis in Section 3.1.6.

Consistent with findings from rat studies, gestational exposure of Dutch-Belted rabbits to 400 mg/kg
DBP on GDs 15 to 29 caused numerous effects consistent with phthalate syndrome (Higuchi et al..
2003). Observed effects included (1) reduced absolute paired testis weight at postnatal week (PNW) 12
(but not at PNW 25) and accessory sex gland weight at PNWs 12 and 25; (2) sperm effects (i.e., reduced
ejaculate volume, sperm concentration and total sperm per ejaculate; morphologically abnormal sperm
with acrosomal and nuclear defects); (3) pathological changes of the seminiferous epithelium; and (4)

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reduced serum testosterone at PNW 6 (but not at PNWs 12 and 25). Additionally, 1 out of 17 male pups
exposed to DBP in utero exhibited gross malformations of the reproductive tract, including undescended
testes, malformed prepuce, hypospadias, hypoplastic seminal vesicle and prostate, and agenesis of the
bulbourethral gland. This study indicates that like rats, rabbits are also sensitive to phthalate-induced
effects on the developing male reproductive system.

In contrast to rats and rabbits, no effects on testicular morphology or development {i.e., no hypospadias,
cryptorchidism, small testes, impaired spermatogenesis, or testicular dysgenesis), serum testosterone,
germ cell number or proliferation, or Sertoli cell number were noted at birth in marmosets dosed with
500 mg/kg/day MBP between gestational weeks 7 through 15 (McKinnell et ai. 2009). Large clumps of
undifferentiated germ cells were found in two of six marmosets exposed to MBP; however, the
significance of this finding is unclear.

In vivo mouse studies investigating phthalate syndrome provide inconsistent results. Gaido et al. (2007)
observed (1) no effect on fetal testicular testosterone in mice exposed to DBP, MBP, or MEHP at doses
ranging from 500 to 1,500 mg/kg/day; and (2) no effect on expression of steroidogenic genes in mice
exposed to a single dose of 500 mg/kg DBP on GD 18 or multiple doses of 250 mg/kg DBP on GDs 14
to 17. Similarly, Do et al. ( ) found no reduction in fetal testicular testosterone in mice exposed up to
500 mg/kg/day DEHP on GDs 9 to 18. Furthermore, as discussed in Section 3.1.4.1, xenograft studies of
mouse fetal testis tissue found no effect of DBP on testosterone production or steroidogenic gene
expression in grafts retrieved from exposed hosts (Heger et al.. 2012). For DBP, MBP, DEHP, and
MEHP, results consistently indicate that gestational phthalate exposure does not disrupt steroidogenesis
during the critical window in in vivo mouse models, which is inconsistent with rat models.

In contrast to results for DBP and DEHP, Wang et al. (2017) observed a disruption of testicular
steroidogenesis in ICR mice exposed to 450 mg/kg/day DIBP from GD 0 to 21 or GD 0 through PND
21. Observed effects include reduced testicular mRNA and protein expression of cholesterol transport
and steroidogenic genes in offspring at PND 21 and PND 80 and reduced serum and testis testosterone
levels in offspring at PND 21 and PND 80. These results indicate a persistent disruption of testicular
steroidogenesis following gestational and/or perinatal exposure to DIBP. However, the study authors did
not measure testosterone or steroidogenic gene expression in fetal testis, and it is unclear if
steroidogenesis was disrupted during the critical window following exposure to DIBP (Wane et al..
2017).

Although studies indicate that steroidogenesis is not disrupted during the critical window in mouse
models following gestational exposure to DEHP or DBP, other effects consistent with phthalate
syndrome have been observed in mice. One study in which mice were gavaged with 100 to 500 mg/kg
DEHP on embryonic days 12 to 17 reported a dose-dependent reduction in AGD (Liu et al.. 2008);
however, three other studies in which mice were exposed to up to 500 mg/kg/day DEHP (Do et al..
2012). 5 mg/kg/day DEHP (Pocar et al.. 2012) or 450 mg/kg/day DIBP (Wane et al.. 2017) throughout
the critical window found no effect on AGD. Effects on male reproductive organ and accessory gland
weight have been observed following gestational exposure to DEHP and DIBP in mouse models. For
example, Do et al. reported a dose-dependent reduction in absolute testis weight in mice gestationally
exposed to 50 mg/kg/day or more of DEHP; Pocar et al. (2012) reported a dose related decrease in
absolute seminal vesicle, but not testis, weight at low doses of DEHP (>0.05 mg/kg/day); and Wang et
al. (2017) reported decreased absolute testis, but not epididymis, weight in mice exposed to 450
mg/kg/day DIBP. Other notable effects consistent with the development of phthalate syndrome after
gestational and/or perinatal exposure to phthalates have been reported, including: (1) hypospadias and
reduced anterior urethra length at doses of 100 mg/kg/day or greater of DEHP (Liu et al.. 2008); (2)

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decreased sperm concentration and viability after low dose {i.e., 0.05 and 5 mg/kg/day) exposure to
DEHP (Pocar et al. 2012); (3) decreased sperm concentration and motility after perinatal exposure to
450 mg/kg DIBP (Wang et al. 2017); and (4) dose-related increases in seminiferous cord diameter and
MNG formation at dose of 250 to 500 mg/kg/day DBP (Gaido et al.. 2007). Presumably, these effects on
the male reproductive system are occurring in the absence of a disruption of fetal testicular
steroidogenesis during the critical window in mouse models, which is inconsistent with rat models.

Finally, and as discussed in Section 3.1.4.1, mechanistic explant and xenograft studies conducted with
human fetal testis tissue generally indicate that the human fetal testis may be less sensitive to phthalate-
induced disruptions of steroidogenesis. However, these studies have limitations and their results should
be interpreted cautiously. For example, testosterone results were highly variable in the xenograft studies
conducted by both Mitchell et al. ( ) and Heger et al. (2012). Variability may be attributable to small
sample size, variable methods by which the human tissue was obtained, variable age of fetal material,
and/or the fact that most studies were conducted using testis obtained from fetuses outside of the male
programming window.

Notably, scientists from EPA's CPHEA recently conducted a species concordance analysis for DBP that
incorporated additional study types {i.e., mechanistic studies conducted using in vitro cell culture models
and ex vivo tissue culture models) and exposure periods {i.e., postnatal/peripubertal exposures) (Arzuaga
et al.. 2020). The study authors draw several notable conclusions based on the totality of data for DBP.
First, fetal rats appear to be more sensitive than other mammalian species to the antiandrogenic effects
of DBP. Second, effects on the seminiferous cord and germ cells {e.g., decreased Sertoli cell numbers,
altered interactions between germ and Sertoli cells, impaired germ cell development, increased germ cell
apoptosis) appear to be conserved across most mammalian species, including human xenografts. Third,
that anti androgenic effects, as well as effects on Sertoli cells and germ cells, appear to be conserved
across most mammalian species, including human xenografts, following postnatal exposure to DBP.

3.1.5.1 Species Difference in Metabolism and Toxicokinetics

Species differences in phthalate metabolism and toxicokinetics have been reported, and discussed
extensively by various agencies {e.g., see (ATSDR. 2022; NASEM. 2017)) and regulatory bodies {e.g.,
see (ECB \ n / * ^ 'SC. 2014)). Most recently, ATSDR (2022) summarized available
toxicokinetic data for DEHP and reached several notable conclusions based on the totality of available
information across species and exposure routes. First, ATSDR concluded that DEHP can be absorbed
via the (1) oral (>70 percent for humans; >30 percent in monkeys, rats, mice, and hamsters [because
fecal excretion is generally not accounted for, absorption values based on urinary excretion are
considered underestimates]), (2) dermal (2 percent for humans; 6 percent for rats), and (3) inhalation
routes (98 percent absorption for rats; demonstrated qualitatively in humans). Second, animal studies
indicate that for all routes of exposure, DEHP is systemically distributed, including to the testes and
fetus; however, distribution has not been reliably evaluated in humans. Third, across species, DEHP is
metabolized to MEHP by lipase, for which significant species differences in enzyme activity exist (see
additional discussion below). Fourth, DEHP metabolites are excreted primarily in the urine and feces
(urinary :biliary excretion ratios vary widely across studies), with blood, serum, or plasma elimination
half-lives for MEHP ranging from 2 to 4 hours in humans and marmosets and 1.1 to 9.4 hours in rats.
Finally, ATSDR concluded that metabolite excretion profiles observed in humans are similar to those
observed in other mammalian species {i.e., monkeys, rats, mice, hamsters, and guinea pigs), although
differences in abundance of certain metabolites and glucuronide conjugates have been reported between
species.

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Several quantitative pharmacokinetic studies have noted species specific differences. For example,
Kessler et al. (2004) administered repeated doses 30 or 500 mg/kg/day DEHP to pregnant SD rats and
marmosets and reported peak blood concentrations (Cmax) to be 1.6 to 4.3 times higher in the rat
compared to the marmoset, while AUC values were 2.6 to 15.6 times higher in rats compared to
marmosets. These results indicate differences in dosimetry that may help explain observed differences in
DEHP response between rats and marmosets. Kinetic experiments conducted on human volunteers are
available (Kessler et al.. 2012; Koch et al.. 2012; Koch et al.. 2004). Notably, Kessler et al. (2012) found
that Cmax and AUC values for MEHP and DEHP in human serum were much higher than reported for
rats and marmosets at similar doses, which led the study authors to conclude that the "MEHP blood
burden at a given DEHP dose per kg body weight will be higher in humans than in the animals";
however, this study only included four human volunteers.

Large species differences in tissue lipase activity have been reported. As discussed in Section 3.1.1, a
critical first step in phthalate toxicity is the metabolism of the diester parent phthalate to its monoester
metabolite by lipases in the intestine or liver. Monoester metabolites are implicated as being the toxic
moiety associated with the reproductive toxicity of phthalates. Ito et al. (2005) found lipase activity,
measured as the rate of conversion of DEHP to MEHP, to be 2.3 to 29.5 times higher in mice compared
to rats in liver and small intestine microsomes and 26.7 to 357 times higher in mice compared to
marmosets (Table 3-20). In a follow-up study, Ito et al. (2014) evaluated liver lipase activity in mice and
humans, and found mouse lipase activity to be 5.1 times higher than human lipase activity, however, it is
worth noting that human lipase activity varied by approximately 10-fold across 28 individuals. Intrinsic
clearance {i.e., the ratio of Vmax [maximum velocity]-to-Km [Michaelis constant]) varied significantly
across species, indicating species difference in enzyme affinity for DEHP exist. Notably, human liver
lipase activity was considerably higher than marmosets (~6.5-fold), and only modestly lower than rats
{i.e., less than a factor of 2), indicating liver lipase activity may not vary dramatically between rats and
humans. Ito et al. (2014; 2005) also measured the activity of several other enzymes involved in phthalate
metabolism {i.e., UDP-glucuronocyltransferase, alcohol dehydrogenase, and aldehyde dehydrogenase);
however, the extent of species differences in activity were not as great as for lipase for these enzymes.

Table 3-20. Comparison of Lipase Activity across Species

Species
(Strain)

Small
Intestine
Lipase
Activity
(pmol/mg)

Liver
Lipase
Activity
(pmol/mg)

Liver
Km
(mmol L_1)

Liver
Vmax
(nmol
mg_1min_1)

Liver Vmax/Km

Mouse (CD-l)fl

11,790

4,964

0.012

3.91

333

Mouse
(129/Sv)fe

-

6,220

0.0076

5.45

714

Rat (SD)fl

400

2,129

0.006

1.32

227

Marmoset0

33

186

1.357

0.49

1.38

Human6

-

1,210

0.0144

1.52

106

a Source: dto et al.. 2005)
b Source: dto et al.. 2014)

3.1.6 Data Integration and Weight of Evidence Analysis

Sections 3.1.3.1 to 3.1.3.7 of this document review the available data for the five high-priority and two
manufacturer-requested phthalates for several key outcomes associated with phthalate syndrome. As

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described in Section 3.1.2, these key outcomes were selected to help inform EPA's development of a
cumulative chemical group for CRA based on EPA's current understanding of phthalate syndrome and
its underlying MOA.

To support data integration and the weight of evidence analysis, EPA applied modified Bradford Hill
criteria, which are typically applied in the context of evaluating the relevance of a non-cancer or a
cancer MOA for humans (WHQ/IPCS. 200 ; 1 c. « i1 \ J005). Although the purpose of this document
is not to establish a MOA for phthalate syndrome, modified Bradford Hill criteria {i.e., temporal and
dose-response concordance; strength, consistency and specificity; biological plausibility and coherence)
provide a useful structure for discussing the weight of evidence supporting EPA's proposed cumulative
chemical group. As discussed in Sections 3.1.3.1 to 3.1.3.7, rat models provide the most available in
vivo data supporting key outcomes associated with phthalate syndrome, and the discussion of modified
Bradford Hill criteria is primarily focused on data from available rat studies. However, inconsistencies,
as well as consistencies, observed across species are emphasized throughout the sections that follow.

3.1.6.1 Temporal Concordance

The temporal relationship between phthalate exposure and certain key outcomes associated with
phthalate syndrome is generally well recognized. As discussed by NRC (2008) and NASEM (2017). the
male programming window in which androgen action drives development of the male reproductive
system is from gestation days 15.5 to 18.5 in rats, which corresponds to gestation weeks 8 to 14 in
humans (Welsh et ai. 2008; Carruthers ami loner. 2005). As discussed in Sections 3.1.3.1 and 3.1.3.2,
rat data demonstrate that exposure to DEHP, BBP, DBP, DIBP, DCHP, and DINP (but not DIDP)
during the male programming window result in reduced expression of cholesterol transport and
steroidogenic genes, as well as reduce fetal testicular testosterone content and/or testosterone
production. Time course studies investigating testicular gene expression and testosterone provide
somewhat conflicting results regarding temporality. Johnson et al. (2012) gavaged rats with a single
dose of 500 mg/kg DBP on GD 19 and observed reductions in gene expression 3 (Cypl7al) to 6
{Cypllal, StAR) hours post-exposure, while fetal testicular testosterone was not reduced until 18 hours
post-exposure, supporting a temporal relationship. In contrast, Thompson et al. (2005) reported a 50
percent reduction in fetal testicular testosterone 1 hour after a single dose of 500 mg/kg DBP on GD 19,
while changes in gene expression occurred 3 (StAR) to 6 {Cypllal, Cypl7al, Scarbl) hours post-
exposure and protein levels of these genes were reduced 6 to 12 hours post-exposure. Of note,
testosterone levels were reduced by approximately 50 percent until the 6-hour time point and then
further declined to approximately 20 percent of control values from the 12-hour time point onwards.
This further decline in testosterone correlated with the reduction in mRNA and protein levels of
cholesterol transport and steroidogenic genes.

As discussed in Section 3.1.3.3 to 3.1.3.7, rat data indicate that reductions in fetal testicular testosterone
during the critical window are associated with development of phthalate syndrome-related effects later
in life—including reduced AGD, increased incidence of NR, seminiferous tubule atrophy, hypospadias
and other reproductive malformations. In support of these findings, Howdeshell et al. (2015)
demonstrate an inverse relationship between reduced fetal testicular testosterone production on
gestational day 18 and the frequency and severity of phthalate syndrome-related effects {i.e., decreased
AGD, NR, reproductive tract malformations) observed in prepubertal and adult rats well after cessation
of exposure. Studies by Carruthers et al. (2005) further demonstrate that exposure to as few as two oral
doses of 500 mg/kg DBP on successive days between GDs 15 to 20 can reduce male pup AGD, cause
permanent NR, and increase the frequency of reproductive tract malformations and testicular pathology
in adult rats. These effects were absent when exposure occurred prior to the male programming window

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Collectively, these studies demonstrate the temporal relationship between gestational exposure during
the critical window and the occurrence of adverse effects on the male reproductive system later in life
well after cessation of exposure.

3.1.6.2 Dose-Response Concordance

As discussed in Sections 3.1.3.1 to 3.1.3.7, data from rat studies supporting the key outcomes generally
exhibit strong dose-response concordance (inconsistencies observed across species discussed in Section
3.1.5). For DEHP, BBP, DBP, DIBP, DCHP and DINP, rat studies consistently demonstrate that
gestational exposure during the critical window leads to dose-dependent decreases in fetal testicular
mRNA expression of cholesterol transport genes {i.e., Scarbl, Star), steroidogenic genes {i.e., Cypllal,
Cypl7al, 3bHSD), and Ins/3 (Section 3.1.3.1). Additionally, consistent dose-dependent reductions in
fetal testicular testosterone were observed for DEHP, BBP, DBP, DIBP, DCHP, and DINP (Section

3.1.3.2).

Consistent with a disruption of steroidogenesis, dose-dependent decreases in male pup AGD (Section

3.1.3.3),	increases in nipple/areolae retention in male pups (Section 3.1.3.4) and hypospadias (Section
3.1.3.5) are observed following gestational exposure to DEHP, BBP, DBP, DIBP, and DCHP across
available rat studies. Increased incidence of seminiferous tubule atrophy is also observed following
gestational and/or perinatal exposure to these phthalates across available rat studies (Section 3.1.3.6).
These effects generally occur regardless of rat strain tested; however, NASEM's meta-analysis and
BMD analysis of Wistar and SD rat data for DEHP did note several unexplained inconsistencies in strain
sensitivity (NASEM. 2017). First, SD rats appeared slightly more sensitive to DEHP than Wistar rats for
effects on fetal testicular testosterone (see Table 3-7), while Wistar rats were more sensitive than SD rats
for effects on AGD (see Table 3-9). Also, as noted by both EPA and NASEM, dose-dependent increases
in hypospadias are observed in SD, but not Wistar, rats administered similar doses of DEHP. In contrast,
DBP consistently increased hypospadias in both SD and Wistar rats in a dose dependent manner.
Currently, the biological significance of the strain differences in sensitivity to DEHP are unclear.

For DINP, data indicate less consistent dose-related effects on AGD, nipple/areolae retention, and
seminiferous tubule atrophy following gestational exposure during the critical window. Two out of six
rat studies found that DINP reduced male AGD (Section 3.1.3.3), while two out of three rat studies
report a dose-related increase in male nipple/areolae retention (Section 3.1.3.4). For tubular atrophy, one
study reported a low incidence of this lesion at 750 mg/kg/day (Gray et al. 20001 while a second study
reported that a few rats gestationally exposed to DINP had "areas of tubular degeneration in areas of
focal Ley dig cell hyperplasia." However, doses at which this effect was observed were not consistently
reported (Sobers et al. 2011). and three other studies found no significant incidence of tubule atrophy at
similar or higher doses (Section 3.1.3.6). In a study conducted by Clewell et al. Q ), mild
hypospadias were reported in 1 out of 111 control and 2 out of 84 high-dose {i.e., dosed with 720
mg/kg/day DINP) pups; however, the effect was not statistically significant. Hypospadias have not been
observed in other studies following gestational exposure to DINP at doses as high as 900 mg/kg/day
(Section 3.1.3.5).

Finally, as discussed in Section 3.1.3.7, data are available for DEHP, DBP, DCHP, and DINP that
consistently demonstrate dose-response concordance for formation of MNGs. For BBP and DIBP, only
a single study evaluating MNGs formation was identified for each phthalate and the available studies
only tested a single, relatively high dose {i.e., 600 mg/kg/day DIBP; 750 mg/kg/day BBP) (Spade et al..
2018; Borch et al.. 2006a); however, both studies reported effects on MNG formation that were large in
magnitude.

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To better understand dose-response relationships across phthalates and across evaluated outcomes, EPA
conducted preliminary dose-response analyses on gene expression, testosterone, AGD, NR,
hypospadias, and seminiferous tubule atrophy data. For this preliminary analysis, ED50 values were
calculated (Table 3-21). As can be seen from Table 3-21, 95 percent confidence intervals overlapped
across ED50 values for some phthalates and outcomes, which limits this comparative analysis. However,
certain trends in potency are apparent. First, for effects on fetal testicular gene expression DEHP,

DCHP, and BBP appear to be the most potent, followed by DBP and then DIBP, while DINP is clearly
and consistently the least potent. For effects on fetal testicular testosterone production, DCHP, DEHP,
and DBP appear to be the most potent, followed by BBP and DIBP, while DINP is the least potent. For
effects on male pup nipple/areolae retention, DEHP and DBP appeared to be more potent than DCHP,
DIBP and BBP, while for hypospadias, DIBP and DCHP appear to be more potent than DEHP, BBP,
and DBP. Finally, for seminiferous tubule atrophy, DCHP and DIBP appear to be more potent than BBP
and DEHP, while DBP appears to be the least potent. These preliminary results indicate that although
phthalate potency may vary by outcome, further comparative studies using lower effect levels are
needed. These results also consistently indicate that DINP is less potent than other phthalates, such as
DEHP, which is consistent with how other authoritative agencies have characterized DINP {i.e., as a
weak antiandrogen) (EC/HC. 2015b; NICNAS. 20 I J; 1 c. t ^SC. JO I Of). Finally, it is worth noting that
comparative pharmacokinetic studies indicate that differences in potency are not due to differences in
dosimetry at the target tissue {i.e., fetal testis) (Clewell et at.. 2010).

Collectively, these studies demonstrate dose-response concordance between gestational exposure during
the critical window and the occurrence of adverse effects on the male reproductive system.

Table 3-21. Comparison of Rat ED50 Values (mg/kg/day) across Key Outcomes

Key Outcome"

DEHP
ED50

(95% CI)

DCHP
ED50

(95% CI)

DBP
ED50

(95% CI)

BBP
ED50
(95% CI)

DIBP
ED50
(95% CI)

DINP
ED50
(95% CI)

Star mRNA

109

(33, 196)

99

(48, 202)

247

(74, 824)

77

(46, 129)

324

(201, 523)

592

(493, 709)

Scarbl mRNA

120

(62, 178)

62

(40, 96)

295

(111,779)

50

(20, 121)

287

(159,519)

594

(440, 802)

Cypllal mRNA

173

(102, 249)

129

(49, 338)

367

(170, 793)

126

(59, 266)

407

(253, 654)

1148

(862, 1,530)

CypHal mRNA

134

(101, 168)

53

(30, 92)

285

(186, 437)

180

(129, 251)

371

(219, 626)

802

(698, 921)

3bHSD mRNA

242

(80, 503)

95

(37, 244)

530

(288, 974)

164

(72, 372)

595

(325, 1,089)

1016

(750, 1,376)

Insl3 mRNA

158

(104,215)

162

(97, 270)

237

(149, 376)

167

(65, 434)

414

(261, 656)

1537

(730, 3,236)

Testicular
Testosterone

143

(132, 156)

91

(46, 180)

154

(88, 268)

228

(150,347)

275

(226, 334)

918

(780, 1,081)

i Anogenital
Distance

1314

(1068,1846)

1128

(825, 2042)

920

(775, 1149)

813

(685, 1002)

777

(594, 1,177)

_b

Nipple/areolae
Retention

368

(275, 491)

588

(324, 1067)

331

(240, 463)

749

(551,2,020)

479

(366, 628)

_b

Hypospadias

846

(804, 904)

699

(631, 825)

958

(919, 999)

878

(829, 948)

626

(603, 653)

_b

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Key Outcome"

DEHP
ED50

(95% CI)

DCHP
ED50

(95% CI)

DBP
ED50

(95% CI)

BBP
ED50

(95% CI)

DIBP

ED50

(95% CI)

DINP
ED50
(95% CI)

Seminiferous
Tubule Atrophy

472

(438, 508)

380

(350,412)

628

(576, 683)

417

(392, 444)

344

(313,377)

_b

a Rat ED50 values and 95% confidence intervals (95% CI) as reported in Sections 3.1.3.1 to 3.1.3.6.
b Rat ED50 values were not estimated for DINP for reduced AGD, NR, hypospadias, or seminiferous tubule atrophy
because sufficient dose-response data was not available to support accurate ED50 predictions (see Sections 3.1.3.3 to
3.1.3.6 for more details).

3.1.6.3 Strength, Consistency, and Specificity

As discussed in Sections 3.1.3.1 to 3.1.3.7, rat models provide the most available in vivo data supporting
key outcomes associated with phthalate syndrome. Available rat studies have been conducted by
multiple research groups and are of varying design {i.e., gestational, perinatal, and multigeneration
studies). Available rat studies provide remarkably consistent evidence demonstrating that gestational
exposure to DEHP, BBP, DBP, DIBP, and DCHP during the critical window of development effects all
key outcomes associated with phthalate syndrome (Table 3-22). Although EPA's review focused on
studies that evaluated seven key outcomes, EPA extracted data for all phthalate syndrome-related effects
reported in each reviewed study (see Appendices B.2 to B.8). As can be seen from Table 3-22, other
phthalate syndrome-related effects have been observed following gestational exposure to DEHP, BBP,
DBP, DIBP, and DCHP—including decreased absolute reproductive organ and accessory sex gland
weight, testicular pathology, epididymal and/or gubernaculum agenesis, undescended testes, sperm
effects, and impairment of male fertility and reproductive function. These observations further add to the
weight of evidence demonstrating that gestational exposure to DEHP, BBP, DBP, DIBP, and DCHP
disrupt development of the male reproductive system in rat models.

For DINP, gestational exposure during the critical window results in consistent reductions in fetal
testicular mRNA expression of Insl3, cholesterol transport, and steroidogenesis genes (discussed in
Section 3.1.3.1). Consistent with a disruption of steroidogenesis at the mRNA level, gestational
exposure to DINP also results in consistent reductions in fetal testicular testosterone production (Section
3.1.3.2). However, effects on AGD, NR, and seminiferous tubule atrophy are less consistently observed
across available rat studies (Sections 3.1.3.3, 3.1.3.4, 3.1.3.6). In contrast to other high-priority
phthalates, gestational exposure to DINP does not appear to cause hypospadias or other severe
reproductive tract malformations, such as cryptorchidism (Table 3-22), and did not alter male fertility or
reproductive function in available one- and two-generation reproduction studies (Waterman et ai. 2000).
However, as can be seen from Table 3-22, other effects consistent with phthalate syndrome have been
observed following gestational exposure to DINP, including (1) decreased sperm motility (Sobers et ai.
2011). (2) epididymal agenesis (Gray et ai. 2000). and (3) other testicular pathologies (e.g., Leydig cell
aggregation, enlarged diameter seminiferous chords, many gonocytes centrally located in chords) (Li et
ai. 2015a; Clewell et ai. 2013a; Clewell et a	). These effects provide further evidence that

gestational exposure to DINP can have adverse effects on the developing male reproductive system. As
was discussed in Section 3.1.6.2, comparative dose-response studies demonstrate that DINP is less
potent at disrupting fetal testicular steroidogenesis compared to other high-priority phthalates, and
therefore less consistent effects on apical outcomes are not unexpected.

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2136 Table 3-22. Summary of Phthalate Syndrome-Related Effects Observed in Rat Studies"

Phthalate Syndrome-Related Effect

DEHP

BBP

DBP

DIBP

DCHP

DINP

DIDP

I Steroidogenic gene and fns/3
expression

S

S

S

S

S

S

X

I Fetal testosterone

S

S

S

S

S

S

X

I Anogenital distance

s

s

s

s

s

i

X

Nipple retention

s

s

s

s

s

i

X

Hypospadias

s

s

s

s

s

X

X

Seminiferous tubule atrophy

s

s

s

s

s

i

X

MNGs

s

s

s

s

s

S

-

I Reproductive organ weight6

s

s

s

s

s

i

X

Testicular pathology6

s

s

s

s

s

S

X

Epididymal agenesis

s

s

s

s

-

S

X

Gubemaculum agenesis

s

-

s

-

-

-

X

Undescended testes

s

s

s



X

X

X

Sperm effects1'

s

s

s

-

s

S

X

I Male fertility6

s

s

s

-

X

X

X

Appendix

B.2

B.3

B.4

B.5

B.6

B.7

B.8

^ = Studies available, effects observed,
x = Studies available, no effects observed.
i = Studies available, inconsistent effects observed.

- = No study available.

" See reference list (below) for examples of studies demonstrating each observed effect. Rows shaded white indicate key
outcomes selected by EPA for in depth review. Rows shaded gray are additional phthalate syndrome-related effects
observed during study review of key outcome data. See cited Appendices for study summaries.
h May include decreased absolute testis, epididymis, seminal vesicle, and/or prostate weight.
c May include, but is not limited to, Leydig cell aggregation, interstitial cell hyperplasia or adenoma, Sertoli cell only
tubules, and/or epididymal oligospermia or azoospermia.

''May include, but is not limited to, decreased sperm motility and/or concentration.
e May include, but is not limited to decreased mating, pregnancy, and/or fertility indices.

References

DEHP: Orsan weisht (Gray et al.. 2009; Lin et al.. 2008); testicular oatholosv (Saillenfait et al.. 2009a; Borch et al..
2006b); emdidvmal & subernaculum asenesis (Howdeshell et al.. 2007; Grav et al.. 2000); undescended testes
(Saillenfait et al.. 2009a; Vo et al.. 2009; Cultv et al.. 2008); SDcrm & fertility effects (Grav et al.. 2009; Vo et al.. 2009;
Therlmmune Research Corporation. 2004)

BBP: Orsan weieht (Ahmad et al.. 2014; Aso et al.. 2005; Tvl et al.. 2004); testicular oatholosv (Aso et al.. 2005; Tvl et
al.. 2004); emdidvmal asenesis (Grav et al.. 2000); undescended testes (Tvl et al.. 2004; Ema et al.. 2003; Grav et al..
2000); sperm and fertility effects (Alunad et al.. 2014; Tvl et al.. 2004)

DBP: Orsan weisht (Clewell et al.. 2013b; Mvlchreest et al.. 2000); testicular oatholosv (Clewell et al.. 2013b; Barlow et
al.. 2004); epididvmal & subernaculum asenesis (Howdeshell et al.. 2007; Mvlchreest et al.. 1999); undescended testes
(Li et al.. 2015b; Drake et al.. 2009); SDcrm & fertility effects (Mahood et al.. 2007; NTP. 1995)

DIBP: Orsan weisht (Saillenfait et al.. 2008); testicular oatholosv (Saillenfait et al.. 2008; Borch et al.. 2006a); emdidvmal

asenesis (Saillenfait et al.. 2008); undescended testes (Saillenfait et al.. 2008; Saillenfait et al.. 2006)

DCHP: Orsan weisht (Yamasaki et al.. 2009; Hoshino et al.. 2005); testicular oatholosv (Li et al.. 2016; Ahbab and

Barlas. 2015); undescended testes (Saillenfait et al.. 2009b); fertility & SDcrm effects (Hoshino et al.. 2005)

DINP: Testicular oatholosv (Li et al.. 2015a; Clewell et al.. 2013a; Bobers et al.. 2011); epididvmal asenesis (Grav et al..

2000); sperm & fertility effects (Bobers et al.. 2011; Waterman et al.. 2000)

DIDP: See (Hushka et al.. 2001)

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For DIDP, there is no evidence of effect on the male reproductive system consistent with phthalate
syndrome. Three studies have demonstrated no effect on fetal testicular testosterone production and/or
steroidogenic gene and Insl3 mRNA expression in rats gestationally exposed to up to 1,500 mg/kg/day
DIDP (Sections 3.1.3.1 to 3.1.3.2). In the available two-generation reproduction studies of DIDP,
continuous exposure of up to 400 to 600 mg/kg/day DIDP had no effect on AGD, NR, or hypospadias in
male pups of either generation. Additionally, DIDP did not affect any other phthalate syndrome-related
outcomes in the available two-generation studies, including reproductive indices (e.g., mating, fertility,
gestation and birth index), weight of androgen-sensitive organs (e.g., prostate, testes, epididymis, and
SV), sperm parameters (i.e., sperm count, motility, and morphology) or preputial separation (Hushka et
ai. 2001). Notably, the European Commission (ECJRC. 2003). ECHA (2013). EFSA (2019). Australia
NICNAS (2015b). Health Canada (EC/HC. 2015 e). and the U.S. CPSC (2010d) have also concluded
that DIDP does not induce antiandrogenic effects on the developing male reproductive system.

3.1.6.4	Biological Plausibility and Coherence

As discussed by NRC (2008) and NASEM (2017). androgen action has a conserved role in the
development of the male reproductive system across mammalian species, including humans. In rats,
exposure to certain phthalates during the critical window can disrupt fetal testicular steroidogenesis
leading to reduced testosterone production and a cause spectrum of effects on the developing male
reproductive system. In humans, rat phthalate syndrome shows similarities with the hypothesized
testicular dysgenesis syndrome, which includes adverse effects such as infertility, decreased sperm
count, cryptorchidism, hypospadias, testicular tumors, and reproductive tract malformations (reviewed
in (NRC. 2008)). Further, androgen insufficiency is well described in humans. For example, mutations
in the gene encoding 5a-reductase can result in male pseudohermaphroditism and delay development of
male physical characteristics, resulting in effects ranging from external feminization to male infertility.
These effects demonstrate a conserved role for androgen action in humans.

Given the conserved role that androgens play in development of the male reproductive system across
mammalian species, it is biologically plausible that in utero exposure to phthalates may lead to a
disruption of androgen action and cause adverse effects on the developing male reproductive system in
humans. Biological plausibility is further strengthened by systematic reviews and meta-analyses of
epidemiologic studies conducted by EPA (Radke et ai. 2018) and NASEM (2017). both of which found
moderate evidence of an association between in utero exposure to DEHP and DBP and reduced AGD in
male infants (discussed in Section 3.1.4.2). Notably, NRC (2008). NASEM Q ), and other
authoritative regulatory agencies have drawn similar conclusions regarding biological plausibility of rat
phthalate syndrome in humans and have determined rat models are appropriate for characterizing risk to
human health (ECCC/HC. 2020; EFSA. 2019; ECHA.. 2017; NICNAS. 2015a;	SC. 2014).

3.1.6.5	Uncertainties

Several areas of uncertainty are associated with EPA's current analysis. First, there are differences in
species sensitivity to phthalate-induced reproductive toxicity (discussed in Section 3.1.5). Rats and
rabbits appear to be sensitive species based on numerous studies in rats and the one gestational exposure
study in rabbits, while no effects consistent with phthalate syndrome were observed in one study of
marmosets exposed during the critical window. For mice, no effects of fetal testicular steroidogenesis
are observed following exposure to DBP or DEHP during the critical window. However, some effects
consistent with phthalate syndrome have been observed, albeit inconsistently, including reduced AGD,
nipple/areolae retention, decreased testes and accessory sex gland weights, hypospadias, and sperm
effects. These effects are presumably occurring in the absence of a disruption of fetal testicular
steroidogenesis. Human xenograft and explant studies suggest that the human fetal testis is insensitive to
phthalate-induced perturbations of steroidogenesis. As discussed in Section 3.1.4.1, these studies have

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limitations and their results must be interpreted with caution. Species differences in metabolism and
toxicokinetics have been implicated in playing a role in species differences in sensitivity. For example,
monoester metabolites, formed through the enzymatic action of lipase, are thought to be one of the toxic
moieties associated with phthalate reproductive toxicity.

As discussed in Section 3.1.5.1, studies have shown that mice and rats have significantly higher lipase
activity than marmosets (241- [mice] to 164- [rat] fold higher; Table 3-20). Additionally, comparative
pharmacokinetic studies have found peak blood concentrations and AUC values to be 1.6 to 4.3 and 2.6
to 15.6 times higher, respectively, in rats compared to marmosets when administered equivalent doses of
DEHP. These difference in metabolism and toxicokinetics may explain observed differences in
sensitivity between rats and marmosets. However, lipase activity appears to be significantly higher in
humans compared to marmosets, and is within a factor of two of rat lipase activity (Table 3-20).
Additionally, kinetic experiments with a small number of human volunteers indicate that MEHP blood
levels may be higher in humans compared to rats at comparable doses. These findings raise uncertainty
and seem to indicate that observed differences in species sensitivity cannot be fully explained by
differences in metabolism and toxicokinetics.

Another source of uncertainty is that the molecular initiating event(s) associated with the phthalate
syndrome MOA have not been established (discussed in Section 3.1.1). Establishing the molecular
initiating event(s) associated with phthalate syndrome may help to explain the observed differences in
species sensitivity.

Another source of uncertainty is lack of inhalation and dermal studies that include an exposure that
covers the critical window of development. As discussed in Section 6, EPA is evaluating the oral,
dermal, and inhalation exposure routes for the five high-priority and two manufacturer-requested
phthalates. Lack of inhalation and dermal studies that include exposure throughout the critical window is
a data gap. To address this data gap, EPA may employ route-to-route extrapolation, which can introduce
uncertainty into assessments, as it generally does not account for route-specific differences in
toxicokinetics (IGHRC. 2006).

3.1.7 Proposed Conclusions on Toxicologic Similarity

The totality of rat data indicates that gestational exposure to DEHP, BBP, DBP, DIBP, and DC HP
during the critical window of development leads to a disruption of fetal testicular steroidogenesis, which
results in reduced fetal testicular testosterone production, reduced AGD, nipple/areolae retention, and
hypospadias. Seminiferous tubule atrophy is also consistently observed following exposure to these five
phthalates. Available rat data are remarkably consistent and support temporal and dose-response
concordance. For DINP, available rat data also provide consistent evidence that gestational exposure to
DINP disrupts steroidogenesis in the fetal testes in a dose-related manner. Comparative dose-response
studies indicate that DINP is less potent than other phthalates such as DEHP. Dose-related effects on
AGD and NR were also observed for DINP, albeit less consistently than for other phthalates, while
severe reproductive tract malformations such as hypospadias have not been reported following
gestational exposure. Finally, available data indicate consistent, dose-related increases in incidence of
MNGs following gestational exposure to DEHP, BBP, DBP, DIBP, DCHP, and DINP. In contrast, for
DIDP, the totality of evidence indicates that gestational exposure to very high doses (e.g., 1,500
mg/kg/day) of DIDP does not disrupt fetal testicular steroidogenesis or cause any other effects consistent
with phthalate syndrome in rat models. Based on the totality of data from rat studies. EPA has reached a
preliminary conclusion that DEHP. DBP. BBP. DIBP. DCHP. and DINP. but not DIDP. are
toxicologically similar.

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As discussed above in Section 3.1.6.5, there are several sources of uncertainty that reduce EPA's
confidence in this preliminary conclusion, including differences in species sensitivity observed across
certain mammalian species that do not appear to be fully explained by differences in toxicokinetics. As
discussed in Section 3.1.1, the molecular events associated with the development of phthalate syndrome
are unknown. Establishing the molecular events preceding cellular, organ, and organism-level changes
may help to further explain species differences in sensitivity. Given the conserved role that androgens
play in the development of the male reproductive system across mammalian species, it is biologically
plausible that in utero exposure to phthalates may adversely affect development of the male reproductive
system in humans. Recent systematic reviews and meta-analyses of human epidemiologic data have
linked in utero exposure to DEHP and DBP to reduced AGD at birth, which further strengthens EPA's
conclusion on biological plausibility.

Further, compared to other phthalates such as DEHP, DINP is less potent at disrupting fetal testicular
steroidogenesis and subsequent apical outcomes associated with phthalate syndrome are either
inconsistently reported (e.g., decreased AGD, NR, seminiferous tubule atrophy) or not reported at all
(e.g., hypospadias). These inconsistencies in response are another source of uncertainty and reduce
EPA's confidence in the conclusion that DINP is toxicologically similar to DEHP, DBP, DIBP, BBP,
and DCHP.

3.2 Evidence of Co-exposure over a Relevant Timeframe

In addition to considerations of toxicological similarity, inclusion and grouping phthalates into a CRA
requires consideration of whether co-exposure is occuring over a relevant timeframe for the populations
of concern. Relevant timeframe of exposure could mean exposure to multiple chemical in the same
timeframe or overlapping of persistent effects from exposure to multiple chemicals. Characterizing co-
exposure requires consideration of the source of chemical exposure, populations impacted by exposure,
and the possible varying routes and pathways of exposure. Sources of data or information that can help
determine whether the general population or subpopulations considered under TSCA are potentially co-
exposed to the seven phthalates of interest include

•	biomonitoring data showing the presence of multiple phthalates in a human population;

•	monitoring data of environmental media including ambient air, drinking water, surface water,
and soil showing the co-occurrence of multiple phthalates;

•	product formulation information showing multiple phthalates in a single product; or

•	workplace monitoring information showing that workers may encounter multiple phthalates in an
occupational setting.

The U.S. CPSC and Health Canada applied similar exposure filters for inclusion of individual phthalates
into a CRA and concluded that there was evidence of co-exposure to multiple phthalates to the general
population, pregnant women, women of reproductive age, and infants, based on biomonitoring and
environmental monitoring data (ECCC/HC. 2020;	SC. 2014V

Specifically for the U.S. population, U.S. CPSC (2014) utilized U.S. Centers for Disease Control and
Prevention (CDC) National Health and Nutrition Evaluation Surveys (NHANES) biomonitoring data
from the 2005 to 2006 cycle, which reports urinary concentrations for 15 phthalate metabolites specific
to individual phthalate diesters. U.S. CPSC utilized 12 of the reported metabolites to determine exposure
of pregnant women in the population to nine phthalate diesters, which included the following
toxicologically similar phthalates: BBP, DBP, DEHP, DIBP, and DINP that are being considered for
CRA under TSCA. U.S. CPSC also analyzed urinary biomonitoring data collected through the Study for
Future Families and found that infants (0 to 37 months of age), as well as their mothers, had measurable
levels of BBP, DBP, DEHP, DIBP, and DINP metabolites in their urine (Sathyanaravana et ai. 2008b;

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Sathyanaravana et ai. 2008a). Notably, U.S. CPSC's analysis demonstrated that the general population
and relevant subpopulations of concern had similar exposure levels based on measured urinary phthalate
metabolites, and metabolites were measured above the analytical limit of detection in close to 100
percent of samples.

Analyses of more recent NHANES urinary biomonitoring data by EPA demonstrate continued co-
exposure to the high-priority and manufacturer-requested phthalates. For example, researchers from
EPA's Office of Research and Development report that the frequency of detection of most phthalate
metabolites associated with exposure to BBP, DBP, DEHP, DIBP, DINP, and DIDP was greater than 97
percent across all NHANES participants from the 2013 to 2014 cycle (Reyes and Price. 2018) (Table
3-23). Similarly, as part of America's Children and Environment program, EPA analyzed a subset of
2015 to 2016 NHANES urinary phthalate metabolite data for BBP and DEHP in women ages 16 to 49
years and children ages 6 to 17 years and demonstrate a greater or equal to 97 percent frequency of
detection in these populations for most metabolites. The high frequency of detection of phthalate
metabolites in NHANES urinary biomonitoring data provides strong evidence of co-exposure to the
high-priority and manufacturer-requested phthalates for the U.S. population. As discussed in Section
3.1.5.1, phthalates have elimination half-lives on the order of several hours and are quickly excreted
from the body in urine. Therefore, the presence of phthalate metabolites in NHANES urinary
biomonitoring data indicates recent phthalate exposure.

Table 3-23

. Summary of Phthalate Met

tabolite Detection Frequencies in NHANES

Parent
Phthalate

Urinary Metabolite

Percentage Below the Limit of Detection

2013-2014 NHANES
(All Participants;

N=2663)"

2015-2016 NHANES
(Women Aged 16—49;

N=585) b

2015-2016 NHANES
(Children Ajjcd 6-17;

N=789)b

BBP

Mono-benzyl phthalate (MBzP)

2.4%

3%

2%

DBP

Mono-n-butyl phthalate (MnBP)

1.6%

-

-

DEHP

Mono-2-ethylhexyl phthalate
(MEHP)

37.66%

35%

35%

Mono-(2-ethyl-5-hydroxyhexyl)
phthalate (MEHHP)

0.3%

0%

1%

Mono-(2-ethyl-5-oxohexyl)
phthalate (MEOHP)

0.5%

0.4%

1%

Mono-(2-ethyl-5-carboxypentyl)
phthalate (MECPP)

0.2%

-

-

DIBP

Mono-isobutyl phthalate (MiBP)

2.7%

-

-

DINP

Mono-isononyl phthalate
(MiNP)

59.56%

-

-

Mono-(carboxyoctyl) phthalate
(MCOP)

0.1%

-

-

DIDP

Mono(carboxynonyl) phthalate

1.2%

-

-

" As reported in Reyes et al. (2018)

b As reported in EPA's Detailed Methods for Indicators B9 and B10. prepared is support of America's Children and the
Environment program.

- Indicates that the metabolite was not included as part of the analysis.

Of note, although the DCHP metabolite, monocyclohexyl phthalate, was included in NHANES from
1999 to 2010, it has since been excluded from the NHANES survey due to low detection levels and a
low frequency of detection in human urine (CDC. 2013a). U.S. CPSC (2014) did not report any
exposure to DCHP and stated that current exposure to DCHP individually does not indicate a high level

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of concern. Biomonitoring data used by Health Canada (2020) in their cumulative assessment also did
not include monitoring for DCHP but based on the in-commerce status and measured presence in dust
samples from Canadian homes, DCHP was included in their CRA. Recent human urinary biomonitoring
data is not available to support the conclusion that there is co-exposure to DCHP and other high-priority
and manufacturer-requested phthalates in the U.S. population. However, there is evidence that exposure
to DCHP can occur through various industrial, commercial and consumer uses under TSCA jurisdiction
(see COU Table 2-2 in DCHP Final Scope Document (	2020e)). Based on exposure to DCHP

through the above uses, EPA anticipates there will be co-exposure to DCHP and other phthalates
currently undergoing risk evaluation, for certain subpopulations and exposure scenarios. For example,
an individual might be exposed to DCHP through an occupational exposure or consumer use, and this
exposure may co-occur with other phthalates due to concurrent exposure to DEHP, BBP, DBP, DIBP,
and DINP, as demonstrated by NHANES biomonitoring data.

Based on manufacturers reporting to the Chemical Data Reporting (CDR) database, indicating that they
produce domestically or import into the U.S. generally above 25,000 lb per site per year, all phthalates
currently undergoing risk evaluation under TSCA section 6 are expected to be in commerce. Further
details on the sources of phthalate exposure regulated under TSCA is presented in Section 6.
Additionally, as described in the final scope documents for BBP (	2020a). DBP (U.S. EPA.

2020(f), DCHP (U.S. EPA. 2020e). DEHP (\ ^ \ 2020b). DIBP (1 ! V \ :020c), and DINP
(	2021c). COUs were identified with expected use by consumer, commercial, and industrial

users, further indicating potential for co-exposure.

Based on biomonitoring data and use in commerce, EPA anticipates that there may be co-exposure to
BBP, DBP, DCHP, DEHP, DIBP, DINP, and DIDP for certain populations.

3.3 Proposed Cumulative Chemical Group (Step 1 in Conceptual Model
[Figure 2-1])

As described in EPA's Draft Proposed Principles of CRA under TSCA, there are two primary
considerations for grouping chemicals for inclusion in a CRA, (1) toxicologic similarity, and (2)
evidence of co-exposure over a relevant timeframe. The establishment of a cumulative chemical group
for purposes of CRA is developed using a weight of evidence narrative that clearly characterizes the
strengths and uncertainties of the evidence of toxicological similarity and potential co-exposure for each
chemical considered.

As described in Section 3.2, human urinary biomonitoring data indicate that the U.S. population is
concurrently exposed to DEHP, DBP, BBP, DIBP, DIDP, and DINP. For DCHP, recent human urinary
biomonitoring data are not available; however, DCHP has been detected in house dust samples and has
various industrial, commercial and consumer uses that fall under TSCA jurisdiction, which indicates
there is potential for humans to be co-exposed to DCHP and the other six phthalates.

As described in Section 3.1.7, the weight of evidence indicates that DEHP, DBP, BBP, DCHP, DIBP,
and DINP are toxicologically similar. Data indicate that gestational exposure to these toxicologically
similar phthalates during the critical window of development leads to a spectrum of effects on the
developing male reproductive system consistent with phthalate syndrome. However, DINP is less potent
than other toxicologically similar phthalates. As described in EPA's Guidance on Cumulative Risk
Assessment of Pesticide Chemicals That Have a Common Mechanisms of Toxicity (	'002). not

all chemicals identified as part of common mechanism group need to be carried forward for quantitative
CRA. For example, a chemical with low hazard potential may be excluded.

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As described in EPA's supplemental mixtures guidance (	MX)), quantitative CRAs should

focus on chemicals and exposure scenarios that are likely to be the largest contributors to risk. Further,
uncertainties and biases can be substantial, even for CRAs focusing on a small number of chemicals.
EPA considered whether DINP should be excluded from the phthalate cumulative chemical group on the
basis of its lower potency. In the phthalate CRA conducted by Health Canada (ECCC/HC. 2020). HQs
for DINP were found to be one of the largest contributors to the calculated HI for pregnant
women/women of childbearing age and infants due to relatively higher exposure (see Appendix A.2).
Thus, although DINP is less potent compared to other high-priority phthalates, it may still significantly
contribute to risk for human populations being considered under TSCA due to relatively higher exposure
and therefore EPA does not believe available data support the exclusion of DINP from the phthalate
cumulative chemical group for CRA.

Based on currently available hazard and exposure data (summarized in Table 3-24). EPA proposes a
cumulative chemical group of DEHP. BBP. DBP. DIBP. DCHP. and DINP for human health CRA
under TSCA.

Although NHANES urinary biomonitoring data indicates that there is potential for co-exposure to DIDP
and other phthalates being evaluated under TSCA (DEHP, BBP, DBP, DIBP, DINP) (Section 3.2), the
weight of evidence indicates that DIDP is not toxicologically similar to these phthalates (Section 3.1.7).
Available data indicate that DIDP does not cause effects on the developing male reproductive system
consistent with phthalate syndrome (Section 3.1.7). As shown in Figure 3-1, chemicals included in a
cumulative chemical group should be toxicologically similar and there should be evidence to support co-
exposure over a relevant timeframe. Because DIDP does not satisfy both criteria, EPA proposes to
exclude DIDP from the phthalate cumulative chemical group.

Table 3-24. Summary of Information Supporting EPA's Proposed Cumulative Chemical Group

for CRA under TSCA

Phthalate

High-Priority
or

Manufacturer-
Requested?

Toxicologically
Similar?

Evidence of
Co-exposure
(Biomonitoring)?

Evidence of Exposure
through Manufacturing
and/or Use (Industrial
Commercial,
Consumer)?

Include in
Cumulative
Chemical
Group?

DEHP

High-Priority

Yes

Yes

Yes

Yes

BBP

High-Priority

Yes

Yes

Yes

Yes

DBP

High-Priority

Yes

Yes

Yes

Yes

DIBP

High-Priority

Yes

Yes

Yes

Yes

DCHP

High-Priority

Yes

Limited data"

Yes

Yes

DINP

Manufacturer-
Requested

Yes

Yes

Yes

Yes

DIDP

Manufacturer-
Requested

No

Yes

Yes

No

" The DCHP metabolite, monocyclohexyl phthalate, was included in NHANES from 1999-2010; however, it has since
been excluded from the NHANES survey due to low detection levels and a low frequency of detection in human urine
rCDC. 2013aV

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4 PROPOSED OPTIONS FOR ADDRESSING PHTHALATE
SYNDROME

4.1 Addressing Phthalate Syndrome as a Whole Versus Focusing on the
Most Sensitive Effect

NRC laid out two options for addressing phthalate syndrome, including (1) assessing the syndrome as a
whole, and (2) focusing on the most sensitive effect associated with the syndrome (NRC. 2008). As
discussed further below, EPA considered the applicability of both of these approaches for use in a
phthalate CRA under TSCA.

4,1,1 Addressing Phthalate Syndrome as a Whole	

As discussed by NRC (2008), when addressing phthalate syndrome as a whole there are two potential
approaches that can be used. First, effects associated with phthalate syndrome can be combined by
evaluating individual pup level data for the presence or absence of phthalate syndrome-related effects.
Individual pups can then be classified as exhibiting phthalate syndrome or not. Under this approach,
each dichotomized endpoint is assumed to have an equal level of toxicity, and each pup is simply
classified as having the syndrome or not. A second option for addressing phthalate syndrome as a whole
is to develop and incorporate a scoring method that adjusts individual pup level data for the severity of
each observed effect. For this approach, (1) data for individuals is evaluated, (2) each observed phthalate
syndrome-related effect is scored for severity, and (3) a composite toxicity score is developed for each
exposed individual.

Recently, researchers in EPA's Office of Research and Development (ORD) developed an ordinal dose-
response modeling approach for addressing phthalate syndrome as a whole (Blessinger et ai. 2020).
Under this approach, data for several phthalate syndrome-related outcomes are evaluated for each
individual pup, and individual pups are categorized into ordinal levels based on the expected effect on
male fertility. Ordinal levels include, level 0 (no phthalate syndrome-related effects observed), level 1
(>1 phthalate syndrome-related effect observed with no to moderate impacts on fertility), or level 2 (>1
phthalate syndrome-related effect observed with severe impacts on fertility). Figure 4-1 shows the
phthalate syndrome-related outcomes and associated levels developed by Blessinger et al. Level binning
decisions were determined by study authors in partnership with EPA's National Center for Risk
Assessment's Reproductive, Developmental, and Neurological Toxicology Workgroup and were
consistent with recommendations of toxicologic pathologists (Lamming et al.. 2002). Once individual
pups are categorized into ordinal levels, benchmark dose (BMD) modeling is conducted to estimate
BMD values for ordinal level 1 and 2 data using a benchmark response of 5 percent and 1 percent extra
risk, respectively.

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Level 1	Level 2

Male developmental reproductive effects

•	Areola/nipple Retention	• Cleft prepuce

•	Hypospadias (mild, moderate, or severe)

•	Exposed os penis

•	Undescended testes (left, right, or both)

•	Small penis

•	Vaginal pouch

•	Prostate absent

•	Seminal vesicles abnormal

Epididymis histopathologyl

•	Interstitial mononuclear cells: grade 2-5	• Bilateral grade 5 oligospermia or azoospermia

•	Oligospermia: grade 2-4

•	Grade 5 oligospermia, unilateral (i.e., in only one epid), w/o azoospermia in other epididymis; or
azoospermia, unilateral, w/o grade 5 oligospermia in other epididymis

•	Sloughed cells (caput, cauda, or corpus): grade 2-5

•	Granulomatous inflammation: grade 2-4b

•	Tubular necrosis: grade 2-5

Testis histopatholog/1

•	Interstitial cell hyperplasia: grade 2-4J

•	Tubular necrosis/mineralization: grade 2-5

•	Tubular vacuolation/loss of germ cells: grade 2-4b

•	Seminiferous tubule degeneration-atrophy/hypoplasia: grade 2-4

•	Seminiferous tubule degeneration-atrophy/hypoplasia: grade 5 in one testis, grade 1-4 or not present in
other testis

•	Loss of seminiferous tubules: grade 2-4

a Unless otherwise indicated, an animal was designated as having the level 1 endpoint if either side (left or right) had the endpoint.

b No animals had grade 5 granulomatous inflammation, grade 2-5 interstitial fibrosis, or grade 5 interstitial cell hyperplasia, tubular vacuolation, or loss of
seminiferous tubules.

Figure 4-1. Proposed Severity Classifications for Phthalate Syndrome-Related Outcomes (from
Blessinger et al. (2020))

Although the ordinal dose-response modeling approach presented by Blessinger et al. provides a
relatively straightforward approach for addressing phthalate syndrome as a whole, limitations are
apparent. First, the approach requires individual-level pup data that is infrequently available. This would
limit the number of studies available to EPA for BMD modeling. Second, the current approach only
incorporates a limited number of phthalate syndrome-related endpoints (e.g., a number of
malformations, such as testis and epididymal agenesis are not included) and outcomes measured as
continuous variables are not included in the scoring system (e.g., decreased AGD, delayed PPS,
decreased reproductive and accessory organ weight). In some cases, exclusion of these outcomes may
inappropriately lead to a pup being binned into level of 0, when level 1 or 2 is more appropriate.

Another consideration is that this approach may not always provide the most sensitive point of
departure. For example, to demonstrate the applicability of the approach, Blessinger et al. used pup data
from a gestational exposure study in which SD rats were orally dosed with 125 to 625 mg/kg/day DIBP
from GD 12 to 21 (Saillenfait et al.. 2008). BMDs for phthalate syndrome ordinal levels 1 and 2 were
215 mg/kg (BMDL = 98 mg/kg) and 234 mg/kg (BMDL =101 mg/kg), respectively, while modeling of
azoospermia and sloughed cells gave more conservative BMDs values of 117 mg/kg (BMDL = 60
mg/kg) and 112 mg/kg (BMDL = 67 mg/kg).

_4.1.2 Focusing on the Most Sensitive Effect	

A second option for addressing phthalate syndrome is to focus on the most sensitive effect associated
with the syndrome. For this approach, comparative dose-response studies are conducted to identify the
most sensitive common phthalate syndrome-related effect across the six toxicologically similar
phthalates under consideration. One potential challenge associated with this approach is that no single
outcome may be identified as the most sensitive across the six toxicologically similar phthalates.
However, failure to identify a single outcome as the most sensitive would likely be more a reflection of
the available literature for each phthalate, than biology. For example, across available gestational and
perinatal studies there is great deal of variation related to dose selection, exposure timing and duration,

• Seminiferous tubular degeneration-atrophy/hypoplasia:
grade 5 in both testes

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species/strain tested, and measured phthalate syndrome-related outcomes (e.g., fetal testicular
testosterone synthesis is sometimes, but not always measured). This variability has generally led to
regulatory agencies to select PODs based on different critical effects for use in phthalate CRAs (e.g., see
Appendix A). However, previous phthalate CRAs conducted by regulatory agencies have generally
utilized the NOAEL/LOAEL approach for determining the critical effect, not more robust dose-response
analyses.

4,1.3 EPA's Proposed Approach for Addressing Phthalate Syndrome	

As discussed in Section 4.1.1, there are numerous challenges and limitations associated with addressing
phthalate syndrome as a whole. Most notably, this approach requires individual pup level data, which is
infrequently available and would limit the number of studies available to EPA for BMD modeling. Due
to this limitation, EPA is proposing to address phthalate syndrome under TSCA by focusing on the most
sensitive effect. As discussed above (Section 4.1.2), one potential challenge with this approach is that no
single outcome may be identified as the most sensitive across the six toxicologically similar phthalates.
Potential options for addressing this challenge, if encountered, are discussed further in Section 4.4.
EPA's proposal to address phthalate syndrome by focusing on the most sensitive effect is consistent
with how U.S. CPSC (20141 Health Canada (ECCC/HC. 20201 Australia NICNAS (2015a. 2014a. b,
2013. 2012). Danish EPA (ECHA. 2011). and EFSA (2J> ) addressed phthalate syndrome (see
summary of CRA approaches in Appendices A. 1 to A. 5).

4.2 Applicability of Dose Addition for Phthalates

As described in EPA's Draft Proposed Principles of CRA under TSCA, several additivity approaches
can be used to evaluate multiple chemical substances for cumulative risk to human health, including
dose addition, response addition, and integrated addition, as well as approaches that account for
toxicologic interactions (	300. 1986). EPA is proposing to rely upon a default assumption of

dose addition when conducting CRAs for toxicologically similar chemical substances under TSCA. As
described in Section 3.1.7, EPA considers there to be sufficient evidence to conclude that DEHP, BBP,
DBP, DIBP, DCHP, and DINP are toxicologically similar and induce effects on the developing male
reproductive system consistent with phthalate syndrome. Therefore. EPA is proposing to evaluate
DEHP. BBP. DBP. DIBP. DCHP. and DINP for cumulative risk to human health under an assumption
of dose addition.

Consistent with EPA's proposal to evaluate phthalates under an assumption of dose addition, other
regulatory agencies that have evaluated phthalates for cumulative risk to human health have also done so
under an assumption of dose addition (ECCC/HC. 2020; EFSA. 2019; NICNAS. 2015a. 2014a. b; U.S.
CPSC. 2014; 'NICNAS. 2013. „ < < l „, H U \ i) In further support of EPA" s proposal to use dose
addition, NRC concluded that there is strong evidence to support the use of dose addition for assessing
anti androgenic phthalates, as well as phthalates and other antiandrogens (despite mixed MO As), for
cumulative risk to human health (NRC. 2008). Notably, NRC's conclusion was based upon empirical
evidence from multiple in vivo phthalate studies (Howdeshell et al.. 2008; Howdeshell et al.. 2007). in
vivo studies of anti androgenic pesticides and pharmaceuticals (Hass et al.. 2007; MetzdorfF et al.. 2007;
Birkhgi et al.. 2004; Nellemann et al.. 2003). and in vivo studies of phthalates and anti androgenic
pesticides and pharmaceuticals with mixed MOAs (Rider et al.. 2008; Hotchkiss et al.. 2004). Although
NRC noted that in many cases both dose addition and response addition can accurately predict observed
effects, in several cases response addition underestimated the observed effects, while dose addition
provided equal or better predictions of observed effects for phthalates, other antiandrogens, and
phthalates in combination with other antiandrogens, despite mixed MOAs.

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Since NRC published their 2008 report, additional in vivo phthalate mixtures studies (Howdeshell et al.
2015; Hannas et al.. 2011) and studies of phthalates and other antiandrogens (Conlev et al.. 2021;

Conlev et al.. 2018; Beverly et al.. 2014; Hotchkiss et al.. 2010; Rider et al.. 2010; Christiansen et al..
2009; Rider et al.. 2009) have been published, and results from these studies further support the
conclusions of NRC (2008) and EPA's proposal to use dose addition for phthalates. For example,

Hannas et al. (2011) report the results of a nine phthalate {i.e., DEHP, D1HP, DIBP, DBP, BBP, DC HP,
DPP, di(n)heptyl phthalate, di-n-hexyl phthalate) fixed ratio mixture study. In this study, SD rats were
gavaged with dilutions of mixture containing 54 to 650 mg/kg total phthalates on GDs 14 to 18, and then
ex vivo fetal testicular testosterone production was evaluated. When observed phthalate mixture effects
were compared to dose addition and response addition model predictions, the study authors found that
dose addition provided the best prediction of the observed mixture effect. In a subsequent fixed ratio
mixture study, Howdeshell et al. (. ) gavaged pregnant SD rats with dilutions of a mixture of five
phthalates (i.e., BBP, DBP, DEHP, DIBP, DPP) from GD 8 to PND 3. Administered mixture dilutions
contained 0, 65, 130, 260, 520 and 780 mg/kg/day total phthalates. Male pups and adult offspring (aged
40 to 46 weeks) were evaluated for 14 phthalate syndrome-related effects, including neonatal mortality,
AGD (PND2), nipple retention (PND 13 and adults), hypospadias, epididymal and testicular
malformations, SV and ventral prostate agenesis, and absolute testes, epididymal, SV, and ventral
prostate weight. Overall, the study authors found that dose addition models accurately predicted 11 out
of 14 outcomes and better predicted observed mixture effects compared to response addition models.

Previously, stakeholders have raised concerns over the applicability of dose addition at very low doses
(i.e., at doses below the individual chemical LOAELs) (	). However, two recent

publications have addressed this uncertainty (Conlev et al.. 2021; Conlev et al.. 2018). Con ley et al.
(2018) administered an 18 chemical mixture that contained 9 phthalates (DEHP, DPP, DBP, DCHP,
BBP, DIBP, diisoheptyl phthalate, dihexyl phthalate, diheptyl phthalate) and 9 antiandrogenic pesticides
and pharmaceuticals (p, p'-DDE, linuron, prochloraz, procymidone, pyrifluquinazon, vinclozolin,
finasteride, flutamide, simvastatin) with mixed MO As to rats on GDs 14 to 18. Dosing solutions were
prepared as a fixed ratio dilution series based on the LOAEL for antiandrogenic effects for each
individual chemical such that the highest dose tested contained each chemical at its LOAEL divided by
5, followed by each chemical at its LOAEL divided by 10, 20, 40, and 80. Antiandrogenic effects (e.g.,
reduced paired testis, epididymal, LABC weight) were noted at the lowest dose tested (i.e., LOAEL/80).
Although, the primary goal of the study was not to evaluate how well dose addition and response
addition models predict observed mixture effects, study authors did compare observed mixture effects
on AGD in male pups at PND 2 with model predictions by comparing observed and model predicted
ED90 and ED60 values. For this outcome, the study authors found that response addition models better
predicted the observed mixture ED90, while dose addition models better predicted the observed mixture
ED60.

In a subsequent study. Con ley et al. (2021) administered a 15 chemical mixture containing 9 phthalates
(BBP, DBP, DCHP, DEHP, DIBP, DPP, diheptyl phthalate, dihexyl phthalate, diisoheptyl phthalate)
and 6 antiandrogenic pesticides and pharmaceuticals (linuron, p,p'-DDE, prochloraz, procymidone,
pyrifluquinazon, vinclozolin) to rats on GDs 14 to 18. Dosing solutions were prepared as a fixed ratio
dilution series based on the NOAEL for antiandrogenic effects for each individual chemical such that
the highest dose contained each chemical at two-fold its NOAEL, followed by a dilution series of each
chemical at its NOAEL and NOAEL divided by 2, 4, 8, 15, 100, and 1,000. Male fetuses (GD 18), pups
(PND 2, 9, 13), and adults (PND 120) were then examined for a suite of effects on the male reproductive
system associated a disruption of androgen action, including decreased AGD, reduced seminal vesicle
weight, and formation of hypospadias. The most sensitive effect was reduced testicular expression of
steroidogenic genes atNOAEL/15. For AGD, seminal vesical weight, and hypospadias, ED50 values

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were calculated based on the observed mixture effects and predicted using dose addition, response
addition, and integration addition models. For all three outcomes, dose addition models provided the
most accurate predictions of observed mixture effects.

Mixture studies by Conley et al. demonstrate several key points. First, they provide evidence to support
the concept of "something from nothing" since effects were observed at exposure levels below the
individual chemical LOAELs {i.e., LOAEL/80 in (Conley et al.. 2018)) and NOAELs {i.e., NOAEL/15
in (Conley et al.. 2021)). Secondly, these studies provide evidence to support the applicability of dose
addition at low doses for mixtures of phthalates and other antiandrogens. Finally, these studies further
demonstrate the applicability of dose addition for mixtures of antiandrogens with mixed MO As. For
example, although the tested chemicals disrupt androgen action through multiple molecular initiating
events {e.g., finasteride is a 5a-reductase inhibitor, flutamide and vinclozolin are androgen receptor
antagonists, linuron inhibits steroidogenic CYPs and is an androgen receptor antagonist, while the
molecular initiating event for phthalates is unknown), these chemicals cause common key cellular events
and lead to common adverse effects on development of the male reproductive tract in a manner
consistent with dose addition.

4.3 Approaches Based on Dose Addition

The final rule for Procedures for Chemical Risk Evaluation Under the Amended Toxic Substances
Control Act (82 FR 33726, July 20, 2017) provides EPA flexibility to select the most appropriate risk
characterization method based on the best available science (TSCA sections 26(h)). As described in
EPA's mixture guidances (2000. 1986). several component-based approaches can be used to evaluate
two or more chemical substances based on dose additivity. The HI approach and RPFs are two
component-based approaches frequently used by EPA. EPA's Office of Land and Emergency
Management (OLEM) frequently uses the HI approach for Superfund site risk assessment (U.S. EPA.
1989). while EPA's Office of Pesticide Programs (OPP) often uses the RPF and MOE approaches to
evaluate multiple pesticides when implementing the Food Quality Protection Act (	)02). The

HI and RPF approaches are described briefly below in Sections 4.3.1 and 4.3.2, respectively, and in
more detail in EPA's mixture guidances (2000. 1986). EPA is considering the applicability of both the
HI and RPF approaches for a phthalates CRA under TSCA.

4,3,1 Hazard Index Approach

The HI approach integrates estimated exposures with toxicity information to characterize the potential
for adverse effects. In the HI approach, hazard quotients (HQs) are calculated by dividing an estimate of
exposure by a reference value (RfV) for each component chemical in the mixture. These HQs are
summed to yield the HI for the mixture (Equation 4-1). For oral and inhalation exposures, EPA's
preferred RfVs are the oral reference dose and inhalation reference concentration, respectively, in health
risk assessments. Because the HI is dimensionless, exposure estimates and the RfV must have the same
units. The HI does not estimate risk,per se; it is not expressed as a probability and does not estimate a
toxicity measure. Instead, the HI is an indicator of potential hazard. In general, an HI that is greater than
or equal to 1 indicates potential concern.

Equation 4-1. Calculating the hazard index

Hi=iUHQi=iu^rt

where:

•	HI= hazard index (unitless)

•	HQi = hazard quotient for the ith chemical (unitless)

•	Ei = estimated exposure for the ith chemical (mg/kg/day or mg/m3)

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•	RjVi = reference for the ith chemical (e.g., RfD in mg/kg/day or mg/m3)

4.3.2	Relative Potency Factor Approach

For the RPF approach, chemicals being evaluated require data that support toxicologic similarity (e.g.,
components of a mixture share a known or suspected common MOA or share a common apical
endpoint/effect) and have dose-response data for the effect of concern over similar exposure ranges
(I	3). RPF values account for potency differences among chemicals in a mixture and scale

the dose of one chemical to an equitoxic dose of another chemical (typically called the index chemical
[IC]). The chemical selected as the IC is often among best characterized toxicologically and considered
to be representative of the type of toxicity elicited by other components of the mixture. Implementing an
RPF approach requires a quantitative dose response assessment for the IC and pertinent data that allow
the potency of the mixture components to be meaningfully compared to that of the IC. In the RPF
approach, RPFs are calculated as the ratio of the potency of the individual component to that of the
index chemical using either (1) the response at a fixed dose; or (2) the dose at a fixed response (Equation
4-2).

Equation 4-2. Calculating RPFs

=

1 BMDR_i

where:

•	BMD = benchmark dose (mg/kg/day or mg/m3)

•	R = magnitude of response (i.e., benchmark response)

•	i = ith chemical

•	IC = index chemical

After scaling the chemical component doses to the potency of the IC, the scaled doses are summed and
expressed as index chemical equivalents for the mixture (Equation 4-3).

Equation 4-3. Calculating index chemical equivalents

Index Chemical EquivalentsMIX = £f=i dj x RPFt
where:

•	Index chemical equivalents = dose of the mixture (mg/kg/day or mg/m3)

•	di = dose of the ith chemical in the mixture (mg/kg/day or mg/m3)

•	RPFi = relative potency factor of the ith chemical in the mixture (unitless)

Noncancer risk associated with exposure to the mixture can then be assessed by calculating an MOE,
which in this case is the ratio of the index chemical's non-cancer hazard value (e.g., the BMDL) to an
estimate of mixture exposure expressed in terms of index chemical equivalents. The MOE is then
compared to the benchmark MOE (i.e., the total uncertainty factor associated with the assessment) to
characterize risk. The lower the MOE (margin between the toxicity effect level and the exposure dose),
the more likely a chemical is to pose a risk.

4.3.3	Proposed Risk Characterization Approach for Phthalates under TSCA

Both the HI and RPF approaches have been used as part of previous phthalate human health CRAs. For
example, Health Canada and Danish EPA employed the HI and risk-characterization-ratio approaches
(analogous to HI approach), respectively (ECCC/HC. 2020:	), while EFSA employed an

RPF approach (EFSA.! ), and U.S. CPSC (2014) employed a hybrid approach that utilized both the
HI approach and relative potency assumptions (see Appendices A.1-A.5 for a summary of previous
phthalate CRA approaches). However, there are challenges associated with the RPF approach. In 2008,
NRC considered the applicability of RPFs for phthalates (NRC. 2008). NRC concluded that RPFs

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cannot be recommended for phthalates because phthalates have dose-response curves that have differing
slopes and shapes depending on the outcome being evaluated, which would result in differing potency
factors depending on the response level at which they are computed.

However, the science has evolved since the NRC made their recommendation against the use of RPFs.
RPFs can be applied for chemicals with dissimilar dose-response curves, as the establishment of a
known or suspected common MOA shared by members of the class of compounds is considered more
fundamental. It is common practice to estimate RPFs closer to the low-dose range of the dose-response
function {i.e., at the 5 or 10 percent effect level versus 50 percent) (	, 2007b. 2000). This

practice is intended to reduce possible high-dose influences on estimated RPFs that may arise due to
saturation of certain kinetic processes (e.g., receptor binding, metabolic elimination). However, this
approach also carries an implicit assumption that dose-response curve shapes will be the same below the
selected response level. In this case, special consideration should be given to the choice of IC, as the IC
should not have an extreme difference in shape compared to other chemicals under consideration.

As discussed above in Section 3.1.7, available data indicate thatDEHP, BBP, DBP, DIBP, DCHP, and
DINP are toxicologically similar. Gestational exposure to these phthalates leads to a common syndrome
(i.e., phthalate syndrome), and there is evidence that suggests a common MOA (i.e., a disruption of fetal
testicular steroidogenesis) for certain, androgen-dependent, aspects of the syndrome. Additionally,
robust dose-response data are available across the toxicologically similar phthalates for multiple key
outcomes associated with phthalate syndrome. Given the available data, EPA believes there is sufficient
information available to support the development of RPFs for phthalates. Therefore. EPA is proposing to
use an RPF approach for the phthalate CRA conducted in support of TSCA section 6 risk evaluations.

4.4 Proposed Options for Deriving Relative Potency Factors

As described in OPP's Guidance on Cumulative Risk Assessment of Pesticide Chemicals that have a
Common Mechanism of Toxicity (	2002). RPFs should be developed based on a uniform point

of comparison. For chemical substances grouped for CRA, this includes, whenever possible, using the
same common effect, same measure of potency, same species/strain and studies that were conducted
using relatively comparable methodology. Additionally, consideration should be given to the human
relevance of the effect.

To support RPF derivation, EPA considered the strengths and uncertainties associated with the dataset
for each evaluated key outcome (Section 4.4.1). Based on this, EPA identified several potential options
for deriving RPFs to address phthalate syndrome, which are discussed in Section 4.4.2.

4.4.1 Strengths and Uncertainties of Key Outcomes Datasets for RPF Derivation

4.4.1.1 Decreased Fetal Testicular Testosterone Production

As discussed in Section 3.1.3.2, testosterone is necessary for the proper development of the male
reproductive system, and a disruption of testicular testosterone production during the masculinization
programming window contributes to the spectrum of effects that make up phthalate syndrome. Further,
reduced testosterone production in the fetal testis plays an early role in the phthalate syndrome MOA.
Available data clearly and consistently demonstrate that gestational exposure to DEHP, BBP, DBP,
DIBP, DCHP, and DINP during the critical window of development leads to a dose-dependent reduction
in fetal testicular testosterone production (Table 3-6). Across these six phthalates, there are robust dose-
response data available from multiple studies that are similar in design (i.e., utilize the same
species/strain of rat, same route/method of exposure, similar exposure durations, similar timing of
measure, and similar method of measuring ex vivo testosterone production via radioimmunoassay).

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Further, several comparative dose-response studies are available that have evaluated fetal testicular
testosterone production following exposure to each of the six toxicologically similar phthalates under
consideration (Gray et al. 2021; Furr et al. 2014).

There are sufficient dose-response data from multiple studies of similar design to support deriving RPFs
for reduced fetal testicular testosterone production. Use of this outcome for deriving RPFs is
strengthened by the fact that androgen action has a conserved role in the development of the male
reproductive system across mammalian species, including humans. Further, reduced fetal testicular
testosterone production has been selected as the critical effect for use in risk characterization in previous
phthalate CRAs conducted by several regulatory agencies, including Australia NICNAS, Health Canada,
and U.S. CPSC (see Table 3-1 and Appendix A).

4.4.1.2	Decreased Fetal Testicular Expression of Cholesterol Transport and
Steroidogenesis Genes

As discussed in Section 3.1.3.1, reduced expression of cholesterol transport and steroidogenesis genes in
the fetal testis plays an early role in the phthalate syndrome MOA. It is biologically plausible that
reduced steroidogenic gene expression will lead to reduced fetal testicular testosterone production, and
some data are available to support the temporal relationship between these outcomes (Section 3.1.6.1,
Figure 3-3). Available data provided consistent evidence to support dose-response concordance for
DEHP, BBP, DBP, DIBP, DCHP, and DINP for this key outcome (Section 3 .1.6.2). For these six
phthalates, adequate data are available to support dose-response modeling for changes in expression of
Scarbl, StAR, Cypllal, 3bHSD, and Cypl7al. Available gene expression studies have been conducted
using similar methodologies (i.e., similar exposure route/method, timing/duration of exposure, timing of
outcome assessment) and have most frequently been conducted with SD rats, although some data are
available for other strains. Further, several comparative dose-response studies investigating gene
expression have been conducted that evaluate all or a subset of the high-priority and manufacturer-
requested phthalates (Gray et al.. 2021; Hannas et al.. 2012; Hannas et al.. 2011).

Here, gene expression data are being considered as a measure of the potency of one chemical relative to
that of another. One challenge with developing RPFs based on gene expression data is that this type of
data is typically not used to derive PODs for use in regulatory risk assessment. Generally, gene
expression data are used by EPA as part of the weight of evidence analysis to support the human
relevance of an effect or support a hypothesized MOA. However, transcriptomic dose-response
modeling approaches that enable transcriptomic PODs to be calculated for use in risk assessment have
been proposed, and this is an active area of research at EPA and NTP. For example, NTP has proposed
deriving a POD based on transcriptomics dose-response data for "active" gene sets (i.e., at least three
genes in the set are altered). The median BMD of affected genes in the active gene set is then derived to
get a central-tendency measure of potency (NTP. 2018). Thus, approaches are available that could
enable EPA to derive RPFs based on reduced testicular steroidogenic gene expression.

4.4.1.3	Decreased Anogenital Distance

As described in Section 3.1.3.3, decreased male AGD is mechanistically linked to reduced fetal
testicular testosterone production and is considered a biomarker of disrupted androgen action. As
described in OECD guidance (OB	), a decrease in male pup AGD that cannot be explained by

differences in animal size indicates an adverse effect that is relevant for setting the NOAEL.
Consistently, reduced male pup AGD has been selected as the critical (or co-critical) effect for
characterizing risk in previous phthalate CRAs (Table 3-1). As can be seen from Table 3-8, there are
sufficient data to support dose-response modeling for DEHP, BBP, DBP, DIBP, DCHP, and DINP.
Generally, there are available studies of similar design across these six phthalates that would facilitate a

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relatively uniform point of comparison {i.e., utilize the same species/strain of rat, same route/method of
exposure, similar exposure durations, and similar timing of measure). However, there are several
challenges with developing RPFs based on decreased AGD. First, OECD guidance recommends that
AGD should be normalized to body weight (preferrable the cubic root of body weight), since animal size
can influence AGD. Many of the available studies only report absolute pup AGD. For example, in the
case of DIBP only one dose-response study is available, and this study only reports absolute AGD.
However, in this case, effects on male pup body weight were only observed at the highest dose tested
(625 mg/mg), while effects on AGD were observed starting at lower doses (>250 mg/kg/day). Thus,
absolute AGD may be used for dose-response modeling, but care must be taken to ensure that
potentially confounding body weight changes are not occurring at lower doses. Another source of
uncertainty stems from the DINP dataset. In contrast to DEHP, BBP, DBP, DCHP, and DIBP where
consistent effects on AGD are reported, statistically significant effects on AGD are less consistently
reported for DINP across studies that test comparable doses {i.e., DINP reduced AGD in two of six
studies). Inconsistency in the DINP dataset reduces EPA's confidence in deriving RPFs based on this
outcome.

4.4.1.4	Nipple/Areolae Retention

As discussed in Section 3.1.3.4, male pup nipple/areolae retention is mechanistically linked to a
reduction in fetal testicular testosterone during gestation. As described in OECD guidance,
nipple/areolae retention is considered a biomarker of a disruption of androgen action and should be
considered in setting the NOAEL (OECD. 2013). Consistently, male pup nipple/areolae retention has
been selected as the critical (or co-critical) effect for characterizing risk in previous phthalate CRAs
(Table 3-1). As can be observed in Table 3-11, there are sufficient data to support dose-response
modeling for DEHP, BBP, DBP, DIBP, DCHP, and DINP. Generally, available studies are of similar
design (utilize the same species/strain of rat, same route/method of exposure, similar exposure durations,
similar timing of measure). However, there are several challenges associated with deriving RPFs for this
outcome. First, as discussed in Section 3.1.3.4, there is variability in how publications report
nipple/areolae retention {e.g., reported as mean number of nipples/areolas per male, incidence of males
with NR, or mean percent of litters with males with NR). Variability in data reporting makes
comparisons across studies difficult. However, sufficient studies reported NR as percent of males per
litter showing retained nipples/areolas to support EPA's preliminary dose-response modeling.
Additionally, although male pup nipple/areolae retention is a biomarker of disrupted androgen action in
rodents, it is not directly a human relevant effect. This uncertainty reduces EPA's confidence in deriving
RPFs based on nipple/areolae retention in male pups.

4.4.1.5	Seminiferous Tubule Atrophy

Seminiferous tubule atrophy is a pathologic lesion frequently reported in adult animals following
gestational and/or perinatal phthalate exposure. As discussed in Section 3.1.3.6, there is some
uncertainty underlying the mechanisms associated with phthalate-induced effects on seminiferous
tubules; however, available studies consistently demonstrate that exposure to DEHP, BBP, DBP, DIBP,
and DCHP lead to a dose-dependent increase in incidence of seminiferous tubule atrophy. Further, this
outcome has been selected as the critical (or co-critical) effect for use in risk characterization in previous
phthalate CRAs conducted by several regulatory agencies (Table 3-1). There appears to be relatively
robust dose-response data available for DEHP, BBP, DBP, DIBP, and DCHP to support dose-response
modeling (Table 3-15). However, there are several challenges associated with using this outcome for
deriving RPFs. First, available studies have utilized differing exposure durations. For example, the three
available studies of BBP and one available study of DCHP are two-generation reproduction studies in
which seminiferous tubule atrophy is reported in adult F1 males following continuous exposure to BBP
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contrast, the one available study of DIBP reports seminiferous tubule atrophy in adult F1 males after
dams were exposed throughout gestation only {i.e., on GDs 12 to 21). For DEHP and DBP, gestational,
perinatal, and continuous exposure studies are available. In contrast to DEHP, BBP, DBP, DCHP, and
DIBP where seminiferous tubule atrophy is consistently observed, effects on tubular atrophy are
inconsistently reported in studies of DINP that test comparable doses. Differences in exposure duration
across available studies and inconsistency in the DINP dataset reduces EPA's confidence in deriving
RPFs based on this outcome.

4.4.1.6	Hypospadias

Hypospadias are a severe malformation of the reproductive tract in which the urethra does not open on
the tip of the penis. As described in Section 3.1.3.5, mechanistic studies provide evidence that link the
formation of hypospadias with reduced fetal testosterone production. Data are available to support dose-
response modeling for the high-priority phthalates, including DEHP, DBP, BBP, DIBP, and DCHP.
However, there are several potential challenges associated with deriving RPFs for this outcome. First,
significant increases in incidence of hypospadias have not been observed following gestational exposure
to DINP. Additionally, as can be seen from Table 3-13, there are several studies available for DEHP,
DBP and BBP that could potentially be used for dose-response modeling; however, data for DIBP and
DCHP are limited to a single study for each phthalate. For BBP, the available study is significantly
different in design {i.e., a two-generation reproduction study in which hypospadias were reported in the
F2 generation) compared to the studies available for other phthalates {i.e., gestational and/or perinatal
exposure studies in which hypospadias are observed in adult F1 animals). Limitations in the hypospadias
dataset reduce EPA's confidence in deriving RPFs based on this outcome.

4.4.1.7	Incidence of MNGs

As discussed in Section 3.1.3.7, MNG formation may serve as a biomarker of altered Sertoli-germ cell
interaction. However, there is uncertainty underlying the MOA associated with MNG formation and the
biological significance of MNGs remains unclear. As can be seen from Table 3-17, although increased
incidence of MNGs has been observed following gestational exposure to DEHP, BBP, DBP, DIBP,
DCHP, and DINP, there is variability in how publications report MNGs. For example, MNGs may be
reported as MNGs per testis or seminiferous cross-section, incidence of animals with MNGs in testes,
percent seminiferous cords with MNGs, or percentage of germ cells multinucleated. These discrepancies
in data reporting make comparisons across studies difficult. Additionally, EPA only identified single
studies evaluating MNGs for BBP and DIBP, and these studies both tested a single high-dose of each
phthalate, which prohibits further dose-response analysis for BBP and DIBP. Given uncertainties related
to biological significance and the MOA underlying MNG formation, as well as data reporting limitations
and lack of adequate dose-response data for BBP and DIBP, EPA is not considering MNGs further for
RPF derivation.

4.4.2 Proposed Options for Deriving RPFs

EPA is proposing to address phthalate syndrome under TSCA by focusing on the most sensitive effect.
One potential challenge with this approach is that no single outcome may be identified as the most
sensitive across the six toxicologically similar phthalates. Potential options are under consideration for
addressing this challenge, if encountered.

Considering the strengths and uncertainties associated with the datasets for each key outcome, datasets
for reduced fetal testicular testosterone production and steroidogenic gene expression appear to have the
most robust datasets to support RPF derivation. Confidence in deriving RPFs for these outcomes is
further strengthened due to the conserved role that androgen action plays in the development of the male
reproductive system across mammalian species, including humans. Furthermore, given what is known

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about the phthalate syndrome MOA (discussed in Section 3.1.1), disrupted steroidogenesis during the
critical window of development is an upstream effect that appears to be necessary for antiandrogenic
effects on the male reproductive system {i.e., phthalate syndrome) to occur. Therefore, reduced fetal
testicular testosterone production and steroidogenic gene expression are appropriate measures of
toxicological potency because they reflect downstream apical outcomes associated with phthalate
syndrome. Datasets for reduced male AGD, NR, seminiferous tubule atrophy, and hypospadias are also
relatively robust, however, as discussed above, these datasets have additional uncertainties and
challenges that reduce EPA's confidence in using them for RPF derivation. The dataset for MNGs {i.e.,
no dose-response data for BBP or DIBP) does not appear sufficient to support RPF derivation.

Given the strengths and uncertainties associated with the datasets for each key outcome, EPA is
considering several options to derive RPFs based on gestational {i.e., reduced fetal testicular testosterone
content and reduced testicular steroidogenic gene expression, Options 1-4) and postnatal effects {i.e.,
AGD, NR, seminiferous tubule atrophy, hypospadias, Options 5 and 6).

Options under consideration by EPA include the following:

•	Option 1. For this option, EPA is proposing to derive RPFs based on reduced fetal testicular
testosterone production and reduced fetal testicular steroidogenic gene expression. Individual
studies of similar design and with sufficient dose-response data would be modelled to estimate
BMDs that would be used to derive RPFs. A range of BMRs in the low-end range of the dose
response curve would be modeled, including, but not limited to BMRs of 5, 10, 20 percent, as
well as the Agency's default of one control standard deviation (	). Modeling
multiple BMRs would allow EPA to consider the consistency of RPFs across a range effect
levels.

•	Option 2. This option is similar to Option 1. RPFs would be derived for testicular testosterone
and steroidogenic gene expression and multiple BMRs would be modelled. However, for this
option dose-response data from studies of similar design would be combined prior to modeling.
One approach to combing data is to conduct a meta-regression, which characterizes dose-
response across a group of studies and considers heterogeneity within and across studies through
random effects. This approach would help eliminate the estimated random effects of inter- and
intra-study variation. The feasibility of this approach for phthalates has been demonstrated by
NASEM (2017). who used meta-regression results to estimate BMDs for reduced fetal testicular
testosterone production for DEHP, DBP, BBP, DIBP, and DINP.

•	Option 3. This option is related to Option 1. RPFs would be derived for testicular testosterone
and steroidogenic gene expression using data from individual studies and multiple BMRs would
be modelled. RPFs derived from gene expression and testosterone data would then be combined
to get a composite RPF for each individual phthalate. This option may be preferable over Option
1 if variability in RPF values is observed between the two modeled outcomes.

•	Option 4. This option is related to Options 2. RPFs would be derived for testicular testosterone
and steroidogenic gene expression using combined dose-response data from multiple studies.
Multiple BMRs would also be modelled. RPFs derived from gene expression and testosterone
data would then be combined to get a composite RPF for each individual phthalate. This option
may be preferable over Option 2 if variability in RPF values is observed between the two
modeled outcomes.

•	Option 5. For this option, RPFs would be derived for postnatal effects {i.e., decreased AGD,
NR, seminiferous tubule atrophy, and hypospadias) using data from individual studies and
multiple BMRs. RPFs would then be combined across postnatal effects to get a composite RPF

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for each individual phthalate. For postnatal effects, EPA is proposing to use a composite RPF
approach for several reasons. First, as discussed above, there are a number of uncertainties and
limitations associated with several of the evaluated postnatal effects and it may not be possible
to derive RPFs for the six toxicologically similar phthalates for all four outcomes (e.g.,
hypospadias are not reported following exposure to DINP). Second, as discussed in Section
3.1.6.2, preliminary dose-response modeling results indicate phthalate potency may vary by
outcome. Developing a composite set of RPFs for postnatal effects would help to circumvent
these challenges.

• Option 6. This option is related to option 5. RPFs would be derived for postnatal effects (i.e.,
decreased AGD, NR, seminiferous tubule atrophy and hypospadias) using multiple BMRs,
however, for this option dose-response data from similarly designed studies would be combined
prior to modeling. RPFs would then be combined for each postnatal effect to get a composite
RPF for each individual phthalate.

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5 PROPOSED POPULATIONS CONSIDERED: STEP 2 IN
CONCEPTUAL MODEL (Figure 2-1)

As described in the final scope documents for BBP (V S i ]' \ JO20a), DBP (U.S. EPA. 2020d). DCHP
(I v	M DEHP (• ^ \ :020b\ DIBP (U.S. EPA. 2020c). and DINP 0 ! V \ :0211c),

EPA will conduct consumer, occupational, and general population exposure assessments for each
individual phthalate. Within these assessments, PESS will be considered, which are "a group of
individuals within the general population identified by [EPA] who, due to either greater susceptibility or
greater exposure, may be at greater risk than the general population of adverse health effects from
exposure to a chemical substance or mixture, such as infants, children, pregnant women, workers, or the
elderly" [15 U.S.C. § 2602(12)]. TSCA does not statutorily define what constitutes "greater
susceptibility" or "greater exposure," thereby providing EPA with flexibility in how PESS groups are
identified.

As discussed throughout Section 3, in utero exposure to DEHP, BBP, DBP, DIBP, DCHP and DINP can
disrupt testicular steroidogenesis and cause adverse effects on the developing male reproductive system.
Postnatal phthalate exposure can also cause male reproductive toxicity; however, the perinatal and
peri pubertal lifestages are believed to be the most sensitive to phthalate exposure (NRC. 2008). Based
on EPA's current understanding of the developmental and reproductive toxicity of phthalates and
susceptible populations identified in previous phthalate CRAs (ECCC/HC. 2020;	3. 20141

EPA initially proposes to focus its CRA for phthalates on two groups that may be more susceptible to
phthalate syndrome due to lifestages:

•	pregnant women/women of reproductive age, and

•	male infants, male toddlers, and male children.

It is important to note that although EPA is proposing to focus its CRA efforts on subpopulations
susceptible to phthalate syndrome based on lifestages. individual phthalate risk evaluations will consider
all relevant lifestages. populations, and PESS.

In addition to potentially being more susceptible to phthalate exposure, these subpopulations and
lifestages identified may also have higher exposure to phthalates from factors that may include, but are
not limited to, diet, mouthing, and exposure relative to bodyweight. These populations may also be
members of the general population living in communities near facilities that emit or release phthalates to
water or ambient air resulting in higher phthalate exposure (i.e., fenceline communities). Overburdened
communities in which the identified susceptible subpopulations are exposed to higher levels of
phthalates will be identified by EPA throughout the risk evaluation process, as appropriate. Additional
PESS based on factors that may include but are not limited to race, ethnicity, or socioeconomic status
who have higher exposures to phthalates may also be identified throughout the risk evaluation process
and incorporated into a CRA as appropriate.

The PESS, initially identified by susceptibility to phthalate syndrome based on lifestage, may be part of
the consumers, workers, and general population that are part of the phthalate exposure assessment.
EPA's proposed approach focuses on the assessment of cumulative risk to consumers, workers, and
general population—specifically fenceline communities that include, but may not be limited to, pregnant
women/women of reproductive age, and male infants, toddlers, and children (Figure 5-1).

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Consumer
Exposure
Assessment

Consumers using
products and articles

iiidt

O

V

Pregnant women

& women of
reproductive age



Male infants,
toddlers, and children

2920

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Occupational

Exposure
Assessment

Workers in
facility

i

%

%

Pregnant women

& women of
reproductive age

General
Population
Exposure
Assessment

Population exposed to

environmental
releases of phthalates



#

Fenceline
communities located
closest to facility

Among fenceline
communities:

Pregnant women

& women of
reproductive age



Male infants,
toddlers, and children

Figure 5-1. Diagram of Initial Proposed Populations Identified Based on
Susceptibility to Phthalate Syndrome

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6 PROPOSED EXPOSURE AND RISK APPROACH FOR

ASSESSING PHTHALATES FOR CUMULATIVE RISK UNDER
TSCA: STEPS 3 TO 10 IN CONCEPTUAL MODEL (Figure 2-1)

6.1	Overview

TSCA Section 6(b)(4)(D) requires EPA to identify the hazards, exposures, COUs, and the PESS the
Administrator expects to consider in a risk evaluation, which EPA did in the final scope documents for
each of the high-priority and manufacturer-requested phthalates. In this section, EPA is providing
similar information as it relates to the CRA for phthalates. As discussed in Section 3.3, EPA is
proposing to assess DEHP, BBP, DBP, DIBP, DCHP, and DINP (but not DIDP) for cumulative risk to
human health under TSCA. This section describes EPA's proposed approach for how exposure from
TSCA COUs may be combined with other exposures from other sources (non-attributable and non-
TSCA) of phthalate exposure to estimate cumulative exposure needed to determine cumulative risks
associated with these phthalates.

This section begins with a summary of information from the final scope documents for each individual
phthalate, including COUs (Section 6.2.1) and exposure pathways (Section 6.2.2). Other sources of
phthalate exposure are also introduced and discussed in those sections. Section 6.3 describes EPA's
proposed scenario-based approach to estimating cumulative phthalate exposure, and a proposed reverse
dosimetry approach to support exposure characterization. The proposed scenario-based approach
includes estimating TSCA, non-attributable, and non-TSCA exposures for reasonable combinations to
determine cumulative risk. A scenario-based method allows for source apportionment of TSCA COU
contributions to the total risk. The reverse dosimetry approach considers CDC's NHANES urinary
biomonitoring dataset and a single compartment toxicokinetic model to estimate total phthalate
exposure. As described in Section 4.3.3, EPA is proposing to use an RPF approach. Therefore, exposure
from each individual phthalate will be scaled to the potency of an IC and expressed in terms of IC
equivalents. This approach is proposed for consumer (Section 6.4.1), occupational (Section 6.4.2), and
general population/fenceline community (Section 6.4.3) exposures. An MOE approach is proposed for
use in characterizing cumulative risk.

6.2	Summary of COUs and Pathways for Phthalates from Individual Scope
Documents

6.2,1 Conditions of Use Listed in Final Scopes for Individual Phthalate Risk Evaluations
(Step 3 in Conceptual Model [Figure 2-1])

As discussed in Section 3.3, EPA's proposed phthalate cumulative chemical group includes BBP, DBP,
DCHP, DEHP, DBP, and DINP, but not DIDP. EPA plans to analyze human exposures and releases to
the environment resulting from the COUs within the scope of the risk evaluation for each of these
phthalates separately, as stated in the final scope documents for BBP (I. c. « i1 \ 2020a). DBP (I. c.
«^ :020d), DCHP (U.S. EPA. 2020el DEHP (U.S. EPA. 2020b\ DIBP (U.S. EPA. 2020c). and
DINP (1 c. 1 ^ \ -XV I < ). In each scope document for the individual chemical substance, EPA identified
and described the categories and subcategories of COUs which include information related to
manufacture, processing, distribution in commerce, use, and disposal that the EPA plans to consider in
the risk evaluation. EPA has gathered those COUs and compiled a list across the different phthalates; the
COUs associated with industrial, commercial, and consumer uses are summarized in Table 6-1. In
addition to these COUs, sites associated with manufacture, processing, distribution, use, and disposal of

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each phthalate that emit releases to the surrounding area are considered sources of possible phthalate
exposure assessed under TSCA.

Prior to the development of the phthalate CRA. exposure scenarios for TSCA CPUs will be completed
in individual phthalate risk evaluations.

Table 6-1. Categories of Conditions of Use for High-Priority Phthalates and a Manufacturer-

Use

Conditions of Use

DBP

BBP

DEHP

DCHP

DIBP

DINP



Adhesive and sealants



X



X

X

X



Automotive care products



X







X



Building/construction materials not
covered elsewhere



X





X

X



Castings



X











Chemical intermediate



X











Fabric, textile, and leather products
not covered elsewhere



X





X





Finishing agent







X







Floor coverings



X





X





Fuels and related products









X



Industrial

Hydraulic fluid



X











Hydraulic fracturing





X









Ink, toner, and colorant products



X



X

X





Laboratory chemicals



X

X









Paints and coatings



X

X



X





Plastic and rubber products not
covered elsewhere



X



X

X





Plasticizer











X



Solvent

X













Transportation equipment
manufacturing





X









Adhesives and sealants

X

X

X

X

X

X



Air care products









X

X



Arts, crafts and hobby materials





X





X



Automotive care products



X

X





X

Commercial

Batteries





X







Building/construction materials not
covered elsewhere



X

X

X



X



Castings



X











Chemical intermediate



X









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Use

Conditions of Use

DBP

BBP

DEHP

DCHP

DIBP

DINP



Chemiluminescent light stick

X













Cleaning and furnishing care
products

X









X



Dyes and pigments





X









Electrical and electronic products





X





X



Explosive materials

X













Fabric, textile, and leather products
not covered elsewhere



X

X





X



Floor coverings

X

X





X

X



Foam seating and bedding products











X



Furniture and furnishings not
covered elsewhere

X



X





X



Hydraulic fluid











X

Commercial

Ink, toner, and colorant products

X

X



X

X



Inspection penetrant kit

X













Laboratory chemical

X

X



X

X

X



Lawn and garden care products





X









Lubricants

X













Paints and coatings

X

X

X

X

X

X



Personal care products

X













Pigment











X



Plastic and rubber products











X



Plastic and rubber products not
covered elsewhere

X

X

X

X

X

X



Solvent











X



Toys, playground, and sporting
equipment





X





X

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Use

Conditions of Use

DBP

BBP

DEHP

DCHP

DIBP

DINP



Adhesives and sealants

X

X

X

X

X

X



Air care products









X

X



Arts, crafts and hobby materials

X

X

X

X



X



Automotive Care products



X

X





X



Batteries





X









Building/construction materials not
covered elsewhere



X

X





X



Chemilumine scent light stick

X













Cleaning and furnishing care
products

X

X







X



Dyes and pigments





X









Electrical and electronic products





X





X

Consumer

Fabric, textile, and leather products
not covered elsewhere

X

X

X



X

X

Floor coverings

X

X





X

X



Foam seating and bedding products











X



Furniture and furnishings not
covered elsewhere

X



X





X



Ink, toner, and colorant products



X



X

X

X



Lawn and garden care products





X









Paints and coatings

X

X

X

X

X

X



Paper products











X



Plastic and rubber products











X



Plastic and rubber products not
covered elsewhere

X

X

X

X

X

X



Reference material and/or
laboratory reagent





X









Toys, playground, and sporting
equipment

X

X

X



X

X

TSCA section 6(b)(4)(D) requires EPA to identify the hazards, exposures, conditions of use, and the PESS
the Administrator expects to consider in a risk evaluation. TSCA section 3(2) excludes from the definition
of "chemical substance" "any food, food additive, drug, cosmetic, or device (as such terms are defined
in section 201 of the Federal Food, Drug, and Cosmetic Act [21 U.S.C. 321]) when manufactured,
processed, or distributed in commerce for use as a food, food additive, drug, cosmetic, or device" as well
as "any pesticide (as defined in the Federal Insecticide, Fungicide, and Rodenticide Act [7 U.S.C. 136 et
seq.]) when manufactured, processed, or distributed in commerce for use as a pesticide." As a result,
EPA identified several non-TSCA uses in the final scope documents for BBP (	)20a). DBP

(U.S. EPA. 2020d\ DC HP (U.S. EPA. 2020el DEHP (\ ^ \ 2020b), D1BP (1 ! V \ :020c). and
DINP (1 c. 1 V \ :02jc) (e.g., use in food packaging materials, dental sealants and nail polish,
fragrances, medical devices, and pharmaceuticals; see Section 2.2.2 of final scope documents for

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additional discussion of non-TSCA uses). These non-TSCA uses are excluded from the definition of
"chemical substance" in TSCA § 3(2)(B)(vi) and are not included in Table 6-1.

EPA may not in a risk management rule under section 6(a) directly regulate non-TSCA uses; however,
incidental effects of 6(a) regulation on non-TSCA uses are not prohibited by TSCA's chemical
substance definition. Additionally, as described in EPA's Risk Evaluation Rule (see Procedures for
Chemical Risk Evaluation Under the Amended TSCA, 33726 Fed. Reg. 33735 (July 20, 2017), "[t]he
potential risks of non-TSCA uses may help inform the Agency's risk determination for the exposures
from uses that are covered under TSCA (e.g., as background exposures that would be accounted for,
should EPA decide to evaluate aggregate exposures)" (82 FR at 33735). Certain non-TSCA sources may
be major pathways of human exposure, and their exclusion from a CRA may lead to an underestimation
of risk. For example, previous phthalate CRAs conducted by U.S. CPSC (2014) and Health Canada
(ECCC/HC. 2020) found dietary sources to be a major pathway of exposure (see Appendix A. 1 to A.2).
Therefore, EPA would consider major non-TSCA sources of phthalate exposure as identified during its
process as part of a CRA.

6.2,2 Pathways and Routes of Exposure Considered in Risk Evaluation as Stated in Final

	Phthalate Scopes	

As stated in the final scope documents for BBP (U.S. EPA. 2020a). DBP (U.S. EPA. 2020d). DCHP
(I c- « ^ 2020e). DEHP (U.S. EPA. 2020b). DIBP (U.S. EPA. 2020c). and DINP (U.S. EPA. 2021c).
EPA plans to analyze exposure levels for indoor air, ambient air, surface water, groundwater, sediment,
human milk, and aquatic biota (e.g., fish) associated with exposure for each of the six phthalates being
considered for the CRA. The scope documents for the individual phthalate risk evaluations present an
exposure analysis plan based on the exposure from TSCA COUs for the individual risk evaluations. The
cumulative assessment, however, will consider exposure from each pathway combined across the
phthalates and has unique consideration for building scenarios that are not completed in the individual
risk evaluation.

Under TSCA section 6(b)(4)(F), EPA is required to "describe whether aggregate or sentinel exposures to
a chemical substance within the conditions of use were considered, and the basis for that determination."
In this definition and within the CRA

•	aggregate exposure is the combined exposures to an individual from a single chemical substance

across multiple routes and across multiple pathways (	), and

•	cumulative exposure is the aggregate exposure to multiple agents or stressors (	303).

Because the cumulative exposure assessment will focus on susceptible subpopulations (described in
Section 5) the increased exposure and/or susceptibility in these populations may support the need for
evaluating aggregate exposures across pathways which can be combined across phthalates to properly
characterize cumulative risk. In their aggregate exposure assessments, both U.S. CPSC (2014) and
Health Canada (ECCC/HC. 2020) found dietary sources to be major contributors to aggregate exposures
in pregnant women and infants (see Appendix A. 1 and A.2). Thus, excluding dietary exposure from
estimation of aggregate and cumulative exposure may lead to an underestimate of risk. In addition,
levels of phthalates are generally detectable in the indoor air, indoor dust, and soil media as
demonstrated by (t; S CPSC. 2014) (see Figure Apx A-1). Concentration of phthalates in these media
may be apportioned to one or more TSCA conditions of use; however, the media concentrations may not
be attributable to specific releases. Even where the phthalate exposures may not be attributed to a
specific condition of use, the exposures/concentrations found in the media may still pertain to the
"chemical substance."

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To account for exposures from different sources expected to impact cumulative risk, the CRA may
include estimations of the following and appropriate combinations of the following exposures:

•	TSCA COU exposure: Exposure that can be attributed to a specific TSCA COU (e.g.,
inhalation exposure during consumer use of an adhesive). Note that exposure scenarios for
TSCA COUs will be completed in individual phthalate risk evaluations and evaluated for
different populations such as consumers, workers, and general population.

•	Non-attributable exposure: Exposure from pathways that cannot be attributed to a specific
TSCA COU or another specific source. Household dust or human milk are a few examples in
which phthalate concentrations measured in those media may result from multiple sources of
phthalates that may nor may not be attributed to a TSCA COU or another specific source.

•	Non-TSCA exposure: Exposure that can be attributed to specific activities that are excluded
from the TSCA definition of "chemical substance," under TSCA Section 3(2), such as a
pesticide, food, food additive, drug, cosmetic, or medical device.

6.3 Scenario-Building for Pathways of Exposure (Steps 4 and 5 in
Conceptual Model)

EPA proposes to combine non-attributable and non-TSCA exposures with exposures from TSCA COUs
when appropriate to determine cumulative exposure. The scenario-building needed to estimate the
various exposures is described below.

6.3.1	TSCA COUs (Step 4 in Conceptual Model [Figure 2-1])	

EPA plans to analyze human exposures and releases to the environment resulting from the COUs stated
in the final scope documents for BBP (I v \ .020a), DBP(l v \ -Q20d), DCHP (U.S. EPA.
2020e), DEHP (U.S. EPA. 2020b\ DIBP (US^EP^JOZOc), and DINP (1 IV \ 2021c). COUs for
each of these phthalates are shown Table 6-1. Prior to the development of the phthalate CRA. scenario-
building and estimations of exposure from TSCA COUs will be completed in individual phthalate risk
evaluations.

6.3.2	Estimating Non-attributable and Non-TSCA Exposures (Step 5 in Conceptual
	Model [Figure 2-1])	

EPA outlines the process for estimating the exposure from sources that are not directly attributable to
TSCA COUs (non-attributable sources) and attributable to non-TSCA COUs (i.e., excluded from the
definition of chemical substance) that will be combined with exposures from TSCA COUs to determine
cumulative exposure.

EPA is considering the applicability of two approaches for estimating non-attributable and non-TSCA
exposures to DEHP, BBP, DBP, DIBP, DCHP, and DINP—including scenario-based and reverse
dosimetry approaches (Figure 6-1).

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Scenario-Based Approach

Non-Attributable Sources

Non-TSCA Sources

BBP DBP DCMP DEHP DIBP DINP l:|-]jil5lS3"IM=l5lla3Ba"""MBIm3

- Combine phthalate exposures from relevant > Combine phthalate exposures from relevant
non-attributable sources	non-TSCA sources

halation













Oral

	Biomonitoring (urinary) data representative
of the U.S. population by age

>	Estimate aggregate exposure for each
individual phthalate

>	Aggregate exposure estimates are not
source apportioned

Urinary concentration
of phthalate metabolite

Dermal

Dermal contact

Non-Attributable Daily Intake by Phthalate

Scale individual phthalate daily intakes
by relative potency and combine

Scale individual phthalate daily intakes
by relative potency and combine

Non-TSCA Cumulative Daily Intake

Reverse dosimetry
model

Daily intake of
parent phthalate

Non-Attributable Cumulative Daily Intake

Figure 6-1. Scenario-Based and Reverse Dosimetry Approaches for Estimating Non-attributable
and Non-TSCA Exposure

•	Scenario-Based Approach. This approach involves estimating exposure for specific
populations based on distinct behaviors, exposure factors, assumptions, and inferences about
how exposure takes place under a specific set of conditions and often relies on monitoring data
for determining the concentrations of chemicals in the various exposure media (described
further in Section 6.3.2.1). Scenarios can be built for individual pathways of exposure and can
be combined to determine cumulative exposure. As shown in Figure 6-1, phthalate exposure
estimates (expressed as a daily intake value) from multiple pathways (e.g., ambient air,
mouthing, etc.) and exposure routes (e.g., inhalation, oral, dermal) for non-attributable and non-
TSCA sources for DEHP, BBP, DBP, DIBP, DCHP, and DINP can be combined with exposure
from TSCA COUs to determine cumulative exposure. The availability of current and reliable
monitoring data, which the approach relies on, is one limitation of this approach. Further
limitations and uncertainties are discussed in Section 6.3.2.4.

•	Reverse Dosimetry Approach. As shown in Figure 6-1, this approach involves estimating
aggregate exposure (expressed as a daily intake value) for each individual phthalate from human
urinary biomonitoring data for metabolites unique to each individual parent phthalate to be
combined for an estimate of cumulative exposure. As described further in Section 6.3.2.2,
reverse dosimetry modeling for phthalates involves use of a single compartment toxicokinetic
model and does not distinguish between routes or pathways of exposure, and does not allow for
source apportionment (i.e., exposure from TSCA COUs cannot be isolated), which are a
limitations of this approach for use under TSCA. Further limitations and uncertainties are
discussed in Section 6.3.2.4.

These approaches are based on the needs of the TSCA exposure assessment as well as review of
available data and approaches utilized in previous phthalates CRAs (ECCC/HC. 2020; U.S. CPSC.
2014). The scenario-based and exposure reverse dosimetry approaches are described further in Sections
6.3.2.1 and 6.3.2.2, respectively. Based on the limitations and uncertainties associated with each

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approach which are described in Section 6.3.2.4. EPA proposes to primarily use a scenario-based
approach to estimate non-attributable and non-TSCA exposures that may be combined with exposure
from TSCA COUs to determine cumulative risk. EPA proposes to use reverse dosimetry as a
comparator for scenario-based exposure estimates as described in Section 6.3.2.5.

6.3.2.1 Scenario-Based Exposure Evaluation for Estimating Non-attributable and
Non-TSCA Exposures

The first approach EPA is considering for estimating non-
attributable and non-TSCA exposures to phthalates is the
scenario-based approach. As shown in EPA's conceptual
model for estimating cumulative exposure (Figure 2-1), the
approach for determining cumulative exposure would be to
combine exposures from TSCA COUs as estimated from
scenario-based approaches to exposures from non-
attributable and non-TSCA exposures estimated using a
scenario-based approach. As described in EPA's Guidelines
for Raman Exposure Assessment (U.S. EPA. 2019a).
scenario-based approaches can be used to define exposure for
specific populations based on distinct behaviors, exposure
factors, assumptions, and inferences about how exposure
takes place under a specific set of conditions and often relies
on monitoring data for determining the concentrations of
chemicals in the various exposure media. Scenario-based
assessments estimate exposure based on intensity, duration,
and frequency of exposure.

Both the U.S. CPSC (2014) and Health Canada (ECCC/HC.

2020) phthalate CRAs estimated aggregate exposure for
multiple phthalates using a scenario-based approach and
calculated a total daily intake value for each individual
phthalate (Text Box 6-1). U.S. CPSC (2014) states their scenario-based approach to be a step-by-step
approach with four steps including compiling concentrations, compiling human exposure factors,
estimating route-specific exposures, and estimating aggregate exposures.

EPA utilizes primarily a scenario-based approach to estimate exposure to TSCA COUs in individual
chemical risk evaluations. Major pathways of exposure may vary based on age group; therefore, non-
attributable and non-TSCA exposures would be assessed separately for relevant populations. For EPA to
utilize a scenario-based exposure assessment to determine non-attributable and non-TSCA exposure
levels to all phthalates, EPA could reconstruct an aggregated daily exposure profile for individuals
varied by lifestages (women of reproductive age, male infants, toddlers, and children) using similar
methods to Health Canada (ECCC/HC. 2020) and U.S. CPSC (2014). In a scenario-based assessment,
unique exposure factors including but not limited to ingestion and inhalation rate, body weight, body
surface area, and dietary intake differences are applied to determine the non-attributable and non-TSCA
exposures to each subpopulation of interest (U.S. EPA. 2021a). For example, given childrens' crawling
and hand-to-mouth behaviors, relevant routes of exposure may include oral in addition to inhalation.
Because exposures can be estimated for various pathways and populations through unique built
scenarios, exposure estimates for non-attributable or non-TSCA pathways can be varied for different
populations and combined differently for an aggregated daily exposure profile for specific populations
to limit the possibility of "double counting."

Text Box 6-1. Sources of Exposure
Identified in CRAs Conducted by U.S.
CPSC and Health Canada

For their phthalate CRAs, both U.S. CPSC
and Health Canada used a scenario-based
approach employing indirect exposure
estimates. U.S. CPSC found the majority
of women's exposure to DEHP, DINP,
and DIBP was from diet (DCHP was not
included in their analysis). Their estimates
were in general agreement (within an
order of magnitude) with two other studies
estimating phthalate exposure using
scenario-based exposure assessment
methods with differences attributable to
differing approaches for dietary exposure
estimation (Clark et al.. 2011; Wonnuth et
al.. 2006).

Health Canada concluded that the main
sources of exposure to the general
Canadian population for medium-chain
phthalates were food, indoor air, dust, and
breast milk (ECCC/HC. 2020).

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Examples of different combinations of exposures from non-attributable and non-TSCA sources are
mentioned in Section 6.4.1 for consumers, Section 6.4.2 for occupational, and Section 6.4.3 for a
fenceline community.

Concentrations of phthalates measured in various environmental media from relevant and reliable
monitoring studies or databases would be considered alongside human exposure factors to determine an
estimated aggregate exposure for each phthalate. Exposure via relevant environmental exposure
pathways that may not be source-attributable may include, but are not limited to, drinking water, surface
water, groundwater, ambient air, indoor air, and soil. Because the scenario-based approach would
consider exposures from various phthalates, sources, and pathways separately, EPA could conduct
sensitivity analyses to determine relative contributions of
the various phthalates and sources of exposure to
cumulative risk to inform risk determinations for individual
phthalates. Determination of relative contributions to
exposure can also help determine the major pathways of
exposure for inclusion in the cumulative estimate as shown
in Step 6 of the conceptual model (Figure 2-1). Once major
pathways of exposure are identified for each individual
phthalate, consideration of the magnitude, frequency, and
duration of exposure must be considered for a relevant
exposure timeframe to determine if co-exposure to multiple
phthalates is occuring from the major relevant pathways of
exposure.

TSCA section 3(2) excludes from the definition of
"chemical substance" "any food, food additive, drug,
cosmetic, or device (as such terms are defined in Section
201 of the Federal Food, Drug, and Cosmetic Act [21
U.S.C. 321]) when manufactured, processed, or distributed
in commerce for use as a food, food additive, drug,
cosmetic, or device." However, as discussed by U.S. CPSC
(2014) and Health Canada (ECCC/HC. 2020). dietary
intake from food and beverages comprises the majority of
daily intake of phthalates and are important sources of
exposure to consider in a cumulative assessment.

There are several approaches EPA may consider for
estimating dietary intake as part of a scenario-based
approach for determining non-TSCA exposure.

First, EPA may identify and evaluate key data sources through a review of the literature for
concentrations of phthalates in food products consumed by the U.S. population and food consumption
patterns, similar to the approach employed by U.S. CPSC (Text Box 6-2). A second option is to use total
diet study data, the method selected by Health Canada (Text Box 6-2). However, currently, there is no
national total diet study measuring phthalate residue in U.S. food products. The Food and Drug
Administration (FDA) conducts an ongoing Total Diet Study to monitor levels of contaminants in foods
eaten by the US population, but phthalates are not measured as part of this study (FDA 2022).

Therefore, EPA may consider if total diet studies of other countries are reflective of the U.S. diet and
use that data if appropriate. U.S. food consumption patterns such as the Food Commodity Intake
Database and understanding of other nations' phthalate regulations will provide insight into which

Text Box 6-2. Approaches Used by U.S.
CPSC and Health Canada to Estimate
Phthalate Dietary Intake

U.S. CPSC estimated dietary exposure using
two datasets of phthalate residues in food
items (Bradley et al.. 2013; Page and
Lacroix. 1995). Additional studies were
used for food categorization and
consumption estimates, including the U.S.
EPA National Center for Environmental
Assessment's analysis of food intake and
diet composition (Clark et al.. 2011; U.S.
EPA. 2007a; Wormuth et al.. 2006).

Health Canada estimated dietary intake of
DIBP, BBP, DBP, and DEHP using the
2013 Canadian Total Diet Study
(ECCC/HC. 2020). For other phthalates, the
2013-2014 and 2014-2015 Food Safety
Action Plan (Canadian Food Inspection
Agency) and/or a dietary exposure study
from the United States (Schecter et al..
2013) were used. A United Kingdom total
diet study (Bradley et al.. 2013) was used to
fill in data gaps. The phthalate
concentrations were matched to 2004
Canadian Community Health Survey on
nutrition (Statistics Canada. 2004)
consumption values for each individual
food.

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nations' total diet studies may be best suited for estimating U.S. intake. EPA's dietary intake assessment
may be deterministic or probabilistic based on available data.

EPA's tiered approach to exposure assessment uses a step-by-step, iterative process in which risk
assessment advances from relatively simple to increasingly more complex analyses as required by the
specific scenario (	i). Each tier corresponds to increased complexity of exposure, risk,

and uncertainty characterization, progressing from screening-level deterministic modeling to advanced
deterministic/mechanistic modeling, and ultimately probability modeling (uncertainty and variability
assessment) following a similar process found in the WHO/IPCS framework for risk assessment of
combined exposure to multiple chemicals (Meek et al..: ). For example, in the WHO/IPCS
framework, the tier of exposure assessment can vary from lower tiers employing simple semi-
quantitative estimates of exposure to higher tiers employing probabilistic exposure estimates. After an
exposure assessment scenario has been conducted, the results can help inform whether additional
refinement of the assessment is needed, either by improving data specificity or by utilizing higher
precision analysis techniques.

Data availability will dictate the tier of exposure assessment employed and may vary based on exposure
scenario. The limited data available for the U.S. diet, for example, may lead to uncertainties in estimates
of total phthalate intake as food and beverages are generally responsible for the majority of total intake
in comparison to other sources (Figure Apx A-l). The recency of food residue data may also introduce
uncertainty in exposure estimates that should be reflective of current populations. In general, varying
levels of data, both in terms of availability and quality, across phthalates and for the various
environmental media concentrations, adds uncertainty to the aggregation of exposure across pathways
and across phthalates that may be quantified using differing tiers of assessment.

6.3.2.2 Reverse dosimetry and Biomonitoring Approach for Estimating Non-
attributable Exposure

A second approach EPA is considering for estimating non-attributable exposures that may include
TSCA and non-TSCA exposure to phthalates is reverse dosimetry. Reverse dosimetry is the process of
estimating an external exposure or intake dose to a chemical using biomonitoring data (

2019a). Reverse dosimetry modeling does not distinguish between routes or pathways of exposure,
instead reverse dosimetry provides an estimate of the total dose (or aggregate exposure) responsible for
the measured biomarker.

Urinary biomonitoring data are available to support estimating exposures for most of the high-priority
and manufacturer-requested phthalates for various lifestages. CDC's NHANES dataset is a national,
statistical representation of the general, non-institutionalized, civilian U.S. population. As can be seen
from Table 6-2, monoester metabolites of BBP, DBP, DEHP, DIBP. and DINP in human urine are
regularly measured as part of the NHANES biomonitoring program, including during the most recent
NHANES survey period for which biomonitoring data is available {i.e., 2017 to 2018). However, DCHP
is an exception. The DCHP metabolite, monocyclohexyl phthalate, was included in NHANES from
1999 to 2010; however, it has since been excluded from the NHANES survey due to low detection
levels and a low frequency of detection in human urine (CDC. ). NHANES urinary biomonitoring
data is also available to support estimating non-attributable exposures for some of the susceptible
subpopulations EPA identified in Section 5, including women of reproductive age (all survey years),
children aged 6 years or older (all survey years), and children aged 3 to 5 years (only included in two
most recent surveys, 2015 to 2016 and 2017 to 2018). However, a limitation of the NHANES dataset is
that it does not include biomonitoring data for infants and generally too few pregnant women are
sampled to support statistical analysis in survey years after 2005 to 2006 (CDC. 201 'h; NCHS. 2012).

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Data from other recent studies that include urinary biomonitoring data for DCHP metabolites, infants,
and pregnant women may be used to help overcome limitations of NHANES, if identified during
systematic review.

Table 6-2. Urinary Phthalate Metabolites Included in NHANES

High-Priority and
Manufacturer-Requested
Phthalates

NHANES Urinary Metabolite"

Associated

Parent
Compound

NHANES
Reporting
Years6

Butyl benzyl phthalate
(BBP)

Mono-benzyl phthalate (MBzP)

BBP

1999-2018

Dibutyl benzyl phthalate
(DBP)

Mono-3-hydroxybutyl phthalate (MHBP)

DBP

2013-2018

Mono-n-butyl phthalate (MnBP)

DBP, BBP

1999-2018

Di-ethylhexyl phthalate
(DEHP)

Mono-2-ethylhexyl phthalate (MEHP)

DEHP

1999-2018

Mono-(2-ethyl-5 -hydroxyhexyl) phthalate
(MEHHP)

DEHP

2001-2018

Mono-(2-ethyl-5 -oxohexyl) phthalate
(MEOHP)

DEHP

2001-2018

Mono-(2-ethyl-5 -carboxypentyl) phthalate
(MECPP)

DEHP

2003-2018

Diisobutyl phthalate
(DIBP)

Mono-isobutyl phthalate (MBP)

DIBP

2001-2018

Mono-2-methyl-2-hydroxypropyl Phthalate
(MHiBP)

DIBP

2013-2018

Dicyclohexyl phthalate
(DCHP)

Mono-cyclohexyl phthalate (MCHP)

DCHP

1999-2010

Di-isononyl phthalate (DINP)

Mono-isononyl phthalate (MiNP)

DINP

1999-2018

Mono-oxoisononyl phthalate (MONP)

DINP

2015-2018

Mono-(carboxyoctyl) phthalate (MCOP)

DINP

2005-2018

"NHANES reports uncorrected and creatinine corrected urine concentrations for each metabolite.
b 2017-2018 is the most recently available NHANES dataset.

NHANES provides data across the population that can be used to create cumulative distribution
functions (percentiles). A major challenge in using the NHANES is selection of the specific metric (e.g.,
median, arithmetic mean, geometric mean, lower or upper percentiles) that represents the non-
attributable exposure. One approach could be to assume that the median exposure represents typical
exposure to the U.S. population and may not include those exposed to specific TSCA COUs. Figure 6-2
shows an example cumulative distribution of NHANES where the central tendency might be
representative of individual exposures that do not include exposure to TSCA COUs and the upper
percentile represents highly exposed individual which may include those exposed to TSCA COUs.

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Upper Percentile

CO

d

Median (50th Percentile)

Central Tendency
Estimate of Aggregate
Exposure: May not
include highly exposed
TSCA source individuals

Upper Percentile Estimate of
Aggregate Exposure:

Includes individuals with higher
exposure due to either TSCA or
non-TSCA sources

Aggregate Phthalate Exposure

Figure 6-2. Diagram of Hypothetical NHANES Population Distribution
of Phthalates and Illustration of Assumptions about Exposure Profiles

Based on the assumption that the median exposure represented by NHANES data may not include
individuals exposed to TSCA sources, EPA could combine the median exposures with exposure from
TSCA COUs to estimate a cumulative exposure where a portion of the exposure was attributable to
TSCA and could be used to inform individual phthalate risk determinations. This approach may be
supported by analyses conducted by both U.S. CPSC (2014) and Health Canada (ECCC/HC. 2020)
indicating that using reverse dosimetry and scenario-based approaches provide similar (within an order
of magnitude) estimates of daily intake for phthalates. Dietary intake, primarily a non-TSCA exposure,
comprised the majority of total cumulative daily intake (Figure Apx A-l), while non-attributable
sources such as dust, indoor air, ambient air, and drinking water were smaller contributors to total
exposure. This assumption necessarily introduces additional uncertainty in the non-attributable exposure
estimates with the potential for "double counting" if estimates from NHANES data already include
exposures from TSCA COUs. However, this approach may prevent underestimations of exposures
attributable to TSCA COUs that may be unique and not captured in a nationally representative dataset,
which has its own limitations discussed below.

A further assumption can then be that the upper percentiles include individuals with TSCA exposures as
well as highly exposed individuals and individuals with differences in kinetics that make them more
susceptible to phthalate exposure. Using the upper percentile, with that assumption, however, would not
allow source apportionment of the TSCA source of exposure to cumulative exposure which is necessary
to inform individual phthalate risk determinations under TSCA. Additionally, there may not be data to
support NHANES being representative of occupational or fenceline populations.

Reverse dosimetry approaches that incorporate basic pharmacokinetic information are available for
phthalates (Koch et al.. 2007; Koch et al.. 2003; David. 2000) and have been used in previous human
health CRAs conducted by U.S. CPSC (2014) and Health Canada (2020). For phthalates, reverse
dosimetry can be used to estimate a daily intake (DI) value for a parent phthalate diester based on
phthalate monoester metabolites measured in human urine using Equation 6-1 (Koch et al.. 2007).

Equation 6-1. Calculating a phthalate daily intake value from urinary biomonitoring data.

(UE5uTn x CE)

Phthalate DI = -—^		 x MWParent

Where:



Phthalate DI = The daily intake (|ig/kgbw/day) value for the parent phthalate diester.

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•	UEsum = The sum molar concentration of urinary metabolites associated with the parent phthalate
diester (in units of |imole per gram creatinine).

•	CE = The creatinine excretion rate normalized by body weight (in units of mg creatinine per kg
bodyweight per day). CE can be estimated from the urinary creatinine values reported in
biomonitoring studies {i.e., NHANES) using the equations of Mage et al. (2008) based on age,
gender, height, and race, as was done by Health Canada (ECCC/HC. 2020) and U.S. CPSC
(2014).

•	FueSUm = The summed molar fraction of metabolites. The molar fraction describes the molar ratio
between the amount of metabolite excreted in urine and the amount of parent compound taken
up.

•	MWparent = The molecular weight of the parent phthalate diester (in units of g/mole).

Using this approach, DI values can be calculated for each of the high-priority and manufacturer-
requested phthalates, scaled to the relative potency of an index chemical, and then scaled daily intake
values can be summed to yield an estimate of non-attributable exposure expressed as index chemical
equivalents.

Controlled human exposure studies have been conducted and provide estimates of the urinary molar
excretion factor {i.e., the Fue) to support use of a reverse dosimetry approach (Table 6-3). These studies
most frequently involve oral administration of an isotope-labelled {e.g., deuterium or carbon-13)
phthalate diester to a healthy human volunteer and then urinary excretion of monoester metabolites is
monitored over 24 to 48 hours. Fue values estimated from these studies have been used by both U.S.
CPSC (2014) and Health Canada (2020) to estimate phthalate DI values using urinary biomonitoring
data. As can be seen from Table 6-3, human Fue values have been estimated for DEHP, BBP, DBP,
DIBP and DINP. However, an Fue value is not available for DCHP and the Fue value for DIBP is
estimated from a single volunteer (Koch et al.. 2012). It may be possible to use analogue data to address
these data gaps. For example, U.S. CPSC (2014) used the DBP Fue value to estimate a daily intake
value for DIBP using reverse dosimetry. Another uncertainty associated with estimated Fue values is
whether or not they are reflective of human variability in phthalate metabolism and excretion. As can be
seen from Table 6-3, Fue values were estimated from a relatively small number (N = 1-20) of adult
human volunteers, and in some cases the age and gender of volunteers is unknown. It is unclear if these
Fue values are reflective of the larger population or susceptible subpopulations based on lifestages
identified in Section 5.

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3327 Table 6-3. Summary of Studies Providing Estimates of the Urinary Excretion Fractions (Fue) of

Phthalate

Metabolites

Parent
Phthalate

Study Population

Metabolite(s)

Fue'1

Fue
Sum6

Reference

DEHP

N = 10 men (20-42 years of
age) and 10 women (18-77
years of age)

MEHP

0.062

0.453

(Anderson et al.. 2011)

MEHHP

0.149

MEOHP

0.109

MECPP

0.132

N = 1 man (61 years of age)

MEHP

0.073

0.469

(Koch et al.. 2004)

MEHHP

0.247

MEOHP

0.149

N - 1 man (61 years of age)

MEHP

0.059

0.627

(Koch et al.. 2005)

MEHHP

0.233

MEOHP

0.150

MECPP

0.185

N = 4 men (28-61 years of age)

MEHP

0.025

0.291c

(Kessler et al.. 2012)

MEHHP

0.125

MEOHP

0.141

BBP

N = 14 volunteers (gender and
age not provided)

MBP

0.06

0.79

(Anderson et al.. 2001)

MBzP

0.73

DBP

N = 13 volunteers (gender and
age not provided)

MBP

0.69

0.69

(Anderson et al.. 2001)

DIBP

N = 1 man (36 years of age)

MiBP

0.703

0.903

(Koch etal.. 2012)

MHiBP

0.1928

30H-MiBP

0.0069

MCiPP

Not

detected

DINP

N = 10 men (20-42 years of
age) and 10 women (18-77
years of age)

MINP

0.030

0.305

(Anderson et al.. 2011)

MONP

0.063

70H-MMeOP

0.114

MCOP

0.099

N - 1 man (63 years of age)

MINP

0.0212

0.396

(Koch and Anserer. 2007)

MONP

0.0997

70H-MMeOP

0.184

MCOP

0.0907

" Fue values are presented on a molar basis and were estimated by study authors based on metabolite excretion over a 24-
hour period.

b Fue sum indicates the sum of Fue values for the measured metabolites.

c Fue calculated based on urinary excretion of metabolites over a 22-hour oeriod (Kcsslcr et al.. 2012).

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Use of reverse dosimetry and urinary biomonitoring data to estimate non-attributable exposures to
phthalates is consistent with approaches employed by both U.S. CPSC (2014) and Health Canada
(2020). However, there are challenges and sources of uncertainty associated with the use of reverse
dosimetry approaches. U.S. CPSC considered several sources of uncertainty associated with use of
human urinary biomonitoring data to estimate daily intake values and conducted a semi-quantitative
evaluation of uncertainties to determine the overall effect on daily intake estimates (see Section 4.1.3 of
(U.S. CPSC. 2014)). Identified sources of uncertainty include: (1) analytical variability in urinary
metabolite measurements; (2) human variability in phthalate metabolism and its effect on metabolite
conversion factors {i.e., the Fue); (3) temporal variability in urinary phthalate metabolite levels; (4)
variability in urinary phthalate metabolite levels due to fasting prior to sample collection; (5) variability
due to fast elimination kinetics and spot samples; and (6) creatinine correction models for estimating
daily intake values.

In addition to some of the limitations and uncertainties discussed above and outlined by U.S. CPSC, the
short half-lives of phthalates can be a challenge when using a reverse dosimetry approach. As discussed
in Section 3.1.5.1 and elsewhere (ATSDR. 2022; EC/HC. 2015c). phthalates have elimination half-lives
on the order of several hours and are quickly excreted from the body in urine and to some extent feces.
Therefore, spot urine samples, as collected through NHANES and many other biomonitoring studies, are
representative of relatively recent exposures. Spot urine samples were used by Health Canada
(ECCC/HC. 2020) and U.S. CPSC (2014) to estimate a daily intake values. The short half-lives of
phthalates, however, lead to single spot sample that may not be representative of average urinary
concentrations that are collected over a longer term or calculated using pooled samples (Shin et al.
2019; Aylward et al.. 2016). Multiple spot samples provide a better characterization of exposure, with
multiple 24-hour samples potentially leading to better characterization but are less feasible to collect for
large studies (Shin et al.. 2019). Due to rapid elimination kinetics, U.S. CPSC concluded that spot urine
samples collected at a short time (2 to 4 hours) since last exposure may overestimate human exposure,
while samples collected at a longer time (>14 hours) since last exposure may underestimate exposure
(see Section 4.1.3 of (IJ..S CPSC. 2014) for further discussion).

Overall, U.S. CPSC (2014) concluded that factors that might lead to an overestimation of daily intake
seem to be well balanced by factors that might lead to an underestimation of daily intake, and therefore
reverse dosimetry approaches "provide a reliable and robust measure of estimating the overall phthalate
exposure."

6.3.2.3 Comparison of Reverse Dosimetry and Scenario-Based Approaches

As discussed in Sections 6.3.2.1 and 6.3.2.2, Health Canada (ECCC/HC. 2020) and U.S. CPSC (2014)
both estimated phthalate daily intake values using reverse dosimetry with human urinary biomonitoring
data and scenario-based exposure assessment approaches. Health Canada and U.S. CPSC found that
both approaches resulted in daily intake values that were generally similar in magnitude. However, this
depended on the recency and quality of data available for use, particularly for data on major exposure
pathways like diet.

U.S. CPSC (2014) indicated that a comparison of the intake values estimated from both methods would
either reveal the presence of pathways of exposure not captured in their scenario-based approach if
estimates from biomonitoring were higher or reveal that worst-case scenarios may not be present in the
biomonitoring approach if their estimates from their scenario-based approach was higher. There were no
assumptions made that the two methods would yield identical results. U.S. CPSC found their estimates
of scenario-based modeled daily intake values to be higher than those estimated using reverse dosimetry
and 2005/2006 NHANES biomonitoring data for several phthalates {i.e., BBP and DINP) (Table 6-4),

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indicating that their scenario-based assessment included potentially worst-case scenarios. Yet, U.S.
CPSC concluded that their results were within an order of magnitude of those from biomonitoring data
and were useful in determining contributions of certain products or phthalates within the combined risk.
In comparing modeled daily intake values for BBP, DEHP, DINP, DBP, and DIBP estimated using a
scenario-based approach to those estimated using NHANES urinary biomonitoring data and reverse
dosimetry, U.S. CPSC demonstrated that while results for both approaches were similar in magnitude,
intake values estimated from biomonitoring data did vary based on the NHANES cycle used for
analysis. This indicates that potential exposure to each of the phthalates may vary over time. For
example, trends in decreasing DEHP exposure, and increasing DINP exposure were observed in
NHANES data from 2005/2006 to 2012/2013 (Table 6-4).

Table 6-4. U.S. CPSC Estimated Median and 95th Percentile Phthalate Daily Intake Values for
Women of Reproductive Age					

Scenario

BBP

DEHP

DINP

DBP

DIBP

Median daily intake (

ug/kg-day)

Scenario-based Estimates'1

1.1

1.6

5.1

0.3

0.1

NHANES 2005/20066

0.26

3.8

1.0

0.69

0.19

NHANES 2007/20086

0.29

4.1

1.5

0.79

0.29

NHANES 2009/2010fe

0.23

2.0

3.0

0.58

0.32

NHANES 2011/2012fe

0.19

1.7

5.0

0.33

0.26

NHANES 2012/2013c

0.15

1.3

5.0

0.33

0.29

95th Percentile daily inta


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Table 6-5. Summary of Uncertainties and Limitations Associated with Use of Scenario-Based and
Reverse Dosimetry Approaches		

Scenario-Based

Reverse Dosimetry

•	Monitoring data sources may not be
reflective of current exposure

•	Lack of data for all phthalates concentrations
in all environmental media

•	Models may utilize conservative assumptions
leading to higher exposure estimates

•	Data availability and quality to determine
exposure from different pathways of
exposure may vary leading to deterministic
estimates for some pathways and
probabilistic estimates for other pathways

•	Uncertainties may compound as individual
intake estimates are aggregated across routes
and pathways and then combined across
phthalates

•	Cannot be source apportioned

•	Relies on use of spot urine samples (may not be
representative of average daily exposure due to fast
elimination kinetics)

•	Urinary excretion factors estimated from controlled
human exposure studies conducted with a limited
number of adult volunteers (may not be reflective of
intraspecies variation in toxicokinetics)

•	No urinary excretion factor is available for DCHP

•	Lack of current biomonitoring data for DCHP
(excluded from NHANES after 2009-10) (systematic
review may identify newer data to address this)

•	Lack of recent infant urinary biomonitoring data
(youngest age group in NHANES is children aged 3 to
5 years) (systematic review may identify newer data to
address this)

•	May introduce additional uncertainties when combined
with scenario-based exposure estimates for specific
TSCA COUs for consumers and workers

As discussed in Section 6.3.2.2, reverse dosimetry is the process of estimating an intake dose for a
chemical based on biomonitoring data (	a). One limitation associated with reverse

dosimetry is that this approach cannot be used to distinguish between routes or pathways of exposure
and cannot be used to determine source apportionment of TSCA and non-TSCA sources. The inability to
source apportion exposure using a biomonitoring approach represents a challenge under TSCA because
aggregate exposure estimates may include exposure from non-TSCA and TSCA COUs leading to an
overestimate of risk due to "double-counting" if non-attributable and non-TSCA exposure is combined
with exposure from another TSCA COU. Additionally, because risk from individual TSCA COUs are
estimated using scenario-based approaches in individual chemical risk evaluations, there may be
uncertainties introduced when combining those exposure estimates with the aggregate exposure
estimated using reverse dosimetry.

Use of a reverse dosimetry approach requires availability of biomonitoring data. In the case of
phthalates, CDC's NHANES dataset provides a relatively recent (data available through 2017 to 2018)
and robust source of urinary biomonitoring data that is considered a national, statistically representative
sample of the non-institutionalized, U.S. civilian population. Further, the NHANES dataset has been
used in previous phthalate CRAs conducted by U.S. CPSC (2014) and Health Canada (ECCC/HC.
2020). However, there are several limitations associated with use of the NHANES urinary
biomonitoring data. First, NHANES does not include infants, one of the susceptible subpopulations
based on lifestages identified by EPA in Section 5 (the youngest age group currently included in
NHANES is children aged 3 to 5 years), nor does NHANES currently measure any urinary metabolites
for DCHP. The DCHP metabolite, monocyclohexyl phthalate, was included in NHANES from 1999 to
2010; however, it has since been excluded from the NHANES survey due to low detection levels and a
low frequency of detection in human urine (CDC. 2013a). These limitations with the NHANES dataset
present a challenge for the use of a reverse dosimetry approach for estimating non-attributable phthlate

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exposure. However, these limitations may be addressed using biomonitoring data from other recent
studies, if identified by EPA during systematic review.

Another source of uncertainty associated with the reverse dosimetry approach is use of spot urine
samples, which are collected as part of NHANES and many other human biomonitoring studies. As
discussed in Section 3.1.5.1 and elsewhere ( .DR. 2022; EC/HC. 2015c). phthalates have elimination
half-lives on the order of several hours and are quickly excreted from the body in urine and to some
extent feces. Therefore, spot urine samples, as collected through NHANES and many other
biomonitoring studies, are representative of relatively recent exposures. The short half-lives of
phthalates, however, lead to single spot sample that may not be representative of average urinary
concentrations that are collected over a longer term or calculated using pooled samples (Shin et at..
2019; Aylward et at.. 2016). Multiple spot samples provide a better characterization of exposure, with
multiple 24-hour samples potentially leading to better characterization but are less feasible to collect for
large studies (Shin et at.. 2019). As discussed by U.S. CPSC, spot urine samples collected at a short time
(2 to 4 hours) since last exposure may overestimate human exposure, while samples collected at a longer
time (>14 hours) since last exposure may underestimate exposure (see section 4.1.3 of (!; S CPSC.
2014) for further discussion).

Human variability in phthalate metabolism and excretion is another potential source of uncertainty
associated with the reverse dosimetry approach. As discussed in Section 6.3.2.2, reverse dosimetry relies
upon molar urinary excretion factors {i.e., the Fue) estimated from controlled human exposure studies to
estimate daily intake values from urinary phthalate metabolites. Fue values are available for DEHP,
DBP, BBP, DIBP and DINP, however, no Fue value is available for DCHP (Table 6-3). Additionally,
the Fue value for DIBP was estimated from a single male volunteer, and may not be reflective of the
larger population or certain PESS {e.g., infants, children, women of reproductive age or pregnant
women). To overcome this limitation, U.S. CPSC (2014) used the Fue value for DBP to estimate a
dietary intake value for DIBP. Finally, as can be seen from Table 6-3, Fue values for DEHP, DBP, BBP
and DINP were calculated based on relatively small sample sizes of 10 to 20 adult volunteers and there
is some uncertainty related to how reflective these Fue values are of variation in phthalate metabolism
and excretion for the broader population, including PESS {e.g., infants, children, women of reproductive
age, pregnant women).

As can be seen from U.S. CPSC's analysis of NHANES urinary biomonitoring data in Table 6-4,
phthalate daily intake estimates vary by year. The most notable trends appear to be that exposure to
DEHP is decreasing, while exposure to DINP is increasing. These trends in exposure are notable, and
may lead to differing conclusions regarding risk, depending upon which NHANES survey year is used.
Similarly, availability of the most recent data to support scenario-based exposure assessments may have
an impact on risk estimates. Dietary intake, for example, comprises a large portion of total estimated
intake for all subpopulations and may vary over time; yet, EPA may not have the data to assess the
dietary intake reflective of the current U.S. population because of the lack of availability of an ongoing
total diet study and may need to rely on older dietary intake data or data from other nations as discussed
in Section 6.3.2.1 leading to uncertainties in dietary intake estimates.

Data availability and data quality may affect estimations of exposure for other relevant pathways as
well, which is a source of uncertainty for estimating exposure using a scenario-based method. Many of
the inputs needed for either deterministic or probabilistic estimates, such as product use, body weight,
breathing rate, environmental media concentration, etc. each have associated variabilities and
uncertainties. Factors including but not limited to sampling methodology, study age, and location can all

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impact uncertainties associated with model inputs. Some of the phthalates may also have more data
available than others.

Data availability will dictate whether deterministic or probabilistic methods are appropriate for each
pathway. Deterministic models use point estimates as inputs and are most often screening level.
Conservative input variables may lead to overestimations of exposure. The type of models used to
estimate intake can also vary for different routes or pathways, some of which are better characterized,
and each have associated uncertainties.

Combining estimates of daily intake from various routes and pathways and across multiple phthalates
also introduces uncertainties. Uncertainties associated with each individual intake estimate may
compound when aggregating to estimate a total intake. Furthermore, aggregating exposure estimates
quantified using differing tiers of assessment as discussed in Section 6.3.2.1. introduces additional
uncertainty. Scenario-based assessments utilize many assumptions of human behavior and can include
conservative assumptions to evaluate risk to be protective of populations assessed. As seen in Table 6-4
U.S. CPSC estimated higher intake values using a scenario-based approach for many, although not all,
of the phthalates and noted it was potentially due to worst-case assumptions that were carried out for
their study (Is S ('PSC.2014Y

6.3.2.5 Proposed Approach for Estimating Exposure from Non-attributable and Non-
TSCA Sources

Given the strengths, limitations and uncertainties of scenario-based and reverse dosimetry approaches
described in Sections 6.3.2.1-6.3.2.4, EPA believes that the scenario-based approach for estimating non-
attributable and non-TSCA phthalate exposure is better suited to support conduct of a phthalate CRA
under TSCA. This is in part because the scenario-based approach provides EPA with more flexibility to
include and/or exclude major pathways of non-attributable and/or non-TSCA exposure when building
cumulative exposure scenarios for consumers (discussed further in Section 6.4.1), workers (Section
6.4.2), and fenceline communities (Section 6.4.3). Furthermore, the scenario-based approach allows for
source apportionment of non-attributable and non-TSCA exposures and estimates of exposure from non-
attributable or non-TSCA sources can be varied for specific subpopulations and exposure scenarios. In
contrast, the reverse dosimetry and biomonitoring approach provides an aggregate exposure estimate for
each individual phthalate, which cannot be source apportioned, and may include exposures from TSCA
sources, which may lead to double-counting if combined with exposure from specific TSCA COUs.

Therefore, EPA is proposing to use environmental monitoring data and modeling to build scenarios for
estimating non-attributable and non-TSCA human exposure to phthalates through relevant pathways of
exposure using a scenario-based approach. Under this approach, non-attributable and non-TSCA
phthalate exposure will be estimated for the susceptible subpopulations identified in Section 5 by
applying exposure factors specific to each lifestage.

Although there are limitations and uncertainties associated with the reverse dosimetry and biomonitoring
approach, EPA recognizes the potential utility of this approach to help characterize phthalate exposure,
and this information may be utilized by EPA in several ways, including:

•	as a comparator for scenario-based daily intake estimates, and

•	temporal trends analysis to better understand changes in phthalate exposure over time.

Recent NHANES urinary biomonitoring data is available for most of the high-priority and
manufacturer-requested phthalates (with the exception of DCHP). Daily intake values estimated using
urinary biomonitoring data and reverse dosimetry can be compared to scenario-based daily intake

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estimates to help EPA determine if reasonable scenarios are being considered in their scenario-based
assessment. EPA does not anticipate that reverse dosimetry and scenario-based approaches will yield
identical results, because both methods have their own sets of uncertainties and limitations (see Section
6.3.2.4); however, as was reported by Health Canada (ECCC/HC. 2020) and U.S. CPSC (20141 both
methods are anticipated to provide similar results (U.S. CPSC found that results were within an order of
magnitude; see Table 6-4).

Additionally, EPA intends to conduct its own updated analysis of the NHANES dataset starting with the
oldest NHANES cycle {i.e., 1999 to 2000) up to the most currently available cycle for each phthalate as
statistics and sampling methodology allows. By analyzing each cycle, EPA can examine the temporal
trend of phthalate exposure over time in women of reproductive age and other susceptible
subpopulations to understand changes in phthalate exposure in the U.S. population. Understanding
current phthalate exposure levels in the U.S. population for each phthalate may help inform the
Agency's risk determination including identifying which phthalates may be contributing to greater
proportions of exposure in women of reproductive age and other susceptible subpopulations over time.

Because NHANES did not include surveillance of children under 6 years of age at the time of their
analysis, U.S. CPSC (2014) used data from the Study for Future Families (Sathyanaravana et ai. 2008b;
Sathyanaravana et al. 2008a) to estimate exposure to children aged 2 to 36 months and to estimate
prenatal and postnatal measurements in women (II S. CPSC. 2014). EPA does not intend to update their
analysis to estimate infant daily intake values unless systematic review identifies new or updated sources
of biomonitoring data for infants. The lack of recent infant urinary phthalate biomonitoring data is a data
gap, which can be overcome by EPA's proposal to primarily rely upon a scenario-based approach to
estimate daily intake values for identified susceptible subpopulations based on lifestages.

6.4 Combining Exposure and Estimating Cumulative Risk (Steps 6 to 10 in
Conceptual Model [Figure 2-1])

6.4.1	Consumer Exposures and Risk	

This section describes EPA's proposed approach for building cumulative exposure scenarios generally
for consumers and estimating cumulative risk for consumers. As stated previously in Section 5, EPA
proposes to focus its CRA for phthalates on subpopulations that may be more susceptible to phthalate
syndrome, which include pregnant women/women of reproductive age, and male infants, male toddlers,
and male children who may be impacted by exposure from TSCA consumer COUs (but the proposed
approach will be presented as applicable to all consumers). This involves the following steps as outlined
in EPA's conceptual model (Figure 2-1):

•	Step 6. Identifying major pathways of exposure. Determining the major pathways of exposure
from TSCA consumer COUs (see purple box in Step 4 of conceptual model in Figure 2-1;
completed in individual risk evaluations), non-attributable, and non-TSCA sources. This step
would be completed after exposures are estimated for the various pathways of exposure and is
dependent on the magnitude of those estimates. Major pathways may vary by relevant population
and may also vary by phthalate. Identification of major pathways of exposure to relevant
populations may require sensitivity analysis for determining inclusion of a pathway into a
cumulative estimate. Description of this process is not detailed in this document as it will be
dependent on the identified pathways.

•	Step 7. Determining co-exposure. Determining likelihood of co-exposure across TSCA
consumer COUs, non-attributable sources, and non-TSCA sources (Section 6.4.1.2).

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•	Step 8. Convert exposures to index chemical equivalents. Phthalate exposure from each
individual phthalate is scaled to the potency of an index chemical using RPFs and expressed in
units of IC equivalents (Section 6.4.1.3).

•	Step 9. Estimating cumulative exposure. Combining TSCA consumer COU cumulative
exposure, the relevant non-attributable cumulative exposure, and the non-TSCA cumulative
exposure to estimate cumulative exposure in a reasonable manner (Section 6.4.1.3).

•	Step 10. Estimating cumulative risk. A cumulative MOE is calculated for comparison to the
benchmark MOE (total uncertainty factor associated with the assessment) (Section 6.4.1.4).

As shown in EPA's conceptual model (Figure 2-1), consumers may be exposed to multiple phthalates
through use of consumer products associated with TSCA COUs, as well through additional non-
attributable and non-TSCA sources (described in Section 6.3.2). Therefore, estimating cumulative risk to
consumers will involve combining major sources of phthalate exposure resulting from TSCA consumer
COU(s), as well as additional non-attributable and non-TSCA sources that can be reasonably expected
to co-occur over a relevant timeframe. Considerations for determining phthalate co-exposure from
TSCA consumer COUs are provided in Section 6.4.1.2, and EPA's proposed approach to estimating
cumulative risk to consumers is provided in Sections 6.4.1.3 and 6.4.1.4.

6.4.1.1 Data Needs for Consumer Co-exposure Analysis

A consumer exposure assessment in individual phthalate risk assessments will estimate magnitude of
exposure to a single phthalate during use of a consumer product, which will depend on the concentration
of the phthalate in the product, use patterns (including frequency, duration, amount of product used,
room of use) and/or application methods. Common data sources used to complete individual chemical
consumer exposure assessments include but are not limited to product formulation data, product use
data, EPA's Exposure Factors Handbook (	32la), and literature sources reporting on indoor

air concentrations or consumer products.

Data sources needed to determine the likelihood of co-exposure to multiple phthalates from a single
consumer product or co-exposure to multiple phthalates from the use of multiple products containing a
single or multiple phthalates may include, but may not be limited to the following:

•	Product Formulation Data. Consumers may encounter co-exposure to multiple phthalates
through exposures from the presence of multiple phthalates in a single product (e.g., plastic
products containing BBP and DBP). The presence of multiple phthalates in a single product may
be determined through process information or production formulation data provided by the
manufacturer of a product or through publicly available product MSDS (Material Safety Data
Sheet) or SDS (Safety Data Sheet) documents.

•	Survey of Consumer Behavior. Co-exposures to two or more chemical substances from
multiple COUs result from what is commonly referred to as the co-occurrence of use (or co-use)
and/or co-location of exposure sources. In other words, a determination of co-exposures is
dependent on evidence of co-use and/or co-location. In the context of TSCA, co-uses typically
refer to scenarios from which an individual (e.g., consumer) may be exposed to two or more
COUs such as when a spray and powdered cleaner are used concurrently to clean a bathtub. For
consumer co-exposures, which are primarily dependent on co-use data that are rare in the
literature, studies which report continuous emissions of chemicals even when products are not in
use can be used to determine which products consumers and bystanders may be co-exposed to
via specific rooms or space of use and periods of time. Usage surveys may also be used to
determine the length of time a product is used to determine if timeframes of exposure to multiple
products may overlap.

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•	Purchase/Market Data. If there is limited data on consumer behavior using products
concurrently, purchase data may provide insight into whether products are ending up in a single
household and potentially leading to co-exposure to multiple phthalates.

6.4.1.2 Co-exposure Resulting from TSCA Consumer COUs (Step 7 in Conceptual
Model [Figure 2-1])

Risks from individual phthalates across various exposure routes, presented in the individual risk
evaluations, will be combined based on the available evidence of co-exposure to determine a cumulative
risk across relevant phthalates. There are several considerations for estimating cumulative risk for
consumers, including consumers' use of

•	multiple products each containing a different high-priority or manufacturer-requested phthalate;

•	multiple products each containing more than one high-priority or manufacturer-requested
phthalate; and

•	a single product containing multiple high-priority and manufacturer-requested phthalates.

6.4.1.2.1 Survey of Consumer Behavior for Determining Co-exposure

As discussed in EPA's Draft Proposed Principles of CRA under TSCA and stated above in the data
needs section (Section 6.4.1.1), in general, there is limited information on the co-use and/or co-location
of consumer products to serve as evidence for co-exposure to different chemicals present in multiple
consumer products. Some studies have investigated co-use patterns for personal care products, which are
regulated by FDA (Safford et al. 201 cerbos et ai. 2013). Thus far, only one co-use study by Han
et al. has been identified, which considered multiple TSCA-relevant consumer products in its analysis,
including laundry detergents, fabric softeners, air fresheners, dishwashing detergents, and all-purpose
cleaners. However, the authors found no strong correlation of co-use between any pair of household and
personal care products (Han et al.. 2020).

6.4.1.2.1	Purchase Data for Determining Co-exposure

Another approach to determine co-use of products has been to use purchase data or presence of certain
consumer products in the home to extrapolate combined exposure and risk (Stanfield et al.. 2021;
Tornero-Velez et al.. 2021). Unfortunately, the presence of consumer products in the home is
insufficient to paint the realistic picture of daily exposure for consumers. This further emphasizes the
importance of co-use data that help to describe consumer use patterns (e.g., which combinations of
products are used, how often, how much, etc.) for products currently on the market.

Currently, available co-use studies indicate that there is lack of evidence of co-use specifically for the
TSCA COUs shown in Table 6-1. This may in part be because many of the TSCA COUs associated with
the phthalates are not necessarily common household products regularly studied for concurrent use.

6.4.1.2.2	Product Formulation Data for Determining Co-exposure

To better understand whether consumers may be exposed to multiple phthalates through the use of a
single product containing more than one phthalate, EPA reviewed products associated with TSCA COUs
listed in the use report for each phthalate (	, »20f. g, h, |,}). This analysis involved

review of product formulation data either from manufacturer websites or Safety Data Sheets (SDS) to
determine the presence of a phthalate in a product and the weight or volume percent of that phthalate in
the product. The same information was also used to determine if multiple phthalates were listed as being
part of a single product. Table 6-6 lists the products and manufacturers identified in each phthalate use
report as being associated with a COU that EPA reviewed. This preliminary analysis revealed little
evidence to suggest that many consumer products associated with TSCA COUs contain more than one
phthalate. Of the products listed in Table 6-6, EPA identified one that contains multiple high-priority
and manufacturer-requested phthalates, which is PSI PolyClav Canes and PSI PolvClav Bricks

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(containing <2.5 percent, but unspecified by weight or volume) each of DEHP, BBP, DBP, and DINP.
Although not always interchangeable, many phthalates serve a similar role as a plasticizer in many
products (Graham. 1973) and, in the identified products with TSCA uses, the phthalates were often
found independently even when identified as being in the same category of products (e.g., paints and
coating). One limitation of this preliminary analysis is that it did not include a review of consumer
product information for DINP. EPA may update this initial analysis as more up-to-date information for
TSCA consumer products is identified.

Table 6-6. Sample of Consumer Products Containing Phthalates

Phthalate

Productabc

Manufacturer d



Sakrete Blacktop Repair Tube

Sakrete of North America



Concrete Patching Compound

Quikrete Companies



Mortar Repair Sealant

Quikrete Companies



DAP Roof & Flashing Sealant, Polyurethane

DAP Products, Inc.



Pre-Mixed Stucco Patch

Quikrete Companies



Hercules Plumber's Caulk - White/Linen

HCC Holdings Inc.



Wilsonart Color Matched Caulk

Wilsonart LLC



Acrylic Caulk

Momentive Performance Materials - Daytona



Silicone Fortified Window & Door Sealant

Henry Company



Air Bloc 33

Henry Company



PSI PolyClay Canes and PSI PolyClay Bricks e

Penn State Industries



Double Bubble Urethane High Peel Strength D50 Part
A (04022)

Royal Adhesives & Sealants



Dymonic FC Anodized Aluminum

Tremco Canadian Sealants [Canada]



GE7000

Momentive Performance Materials



Hydrogel SX

Prime Resins Inc.



Pennatite Acrylic Sealant

Pennatite / Division of DSI



Protecto Sealant 25XL

Protecto Wrap Company

BBP

Spectrem 3 Aluminum Stone - 30 CTG

Tremco Canadian Sealants [Canada]



Spectrem 4

Tremco U.S Sealants



STP 17925 Power Steering Fluid & Stop Leak

Armored AutoGroup Inc.



126VR Disc Brake Quiet 0.25 Fl. Oz Pouch

ITW Pennatex



Steri-Crete SL Component A

Dudick, Inc.



Stonclad UT Resin Polyol

Stonhard, Division of StonCor Group, Inc.



ENSURE Sterilization Emulator

SciCanLtd. [Canada]



Phthalates in Polyvinyl chloride)

SPEX CertiPrep, LLC



Elmer's Model + Hobby Cement

Elmers Products, Inc.



Accent MBRU 6pk Silver Metallic 2oz

Rust-Oleum Corporation



Champion Spray on Acrylic Matte Finish

Chase Products Co.



6840 Ultra Black

BJB Enterprises, Inc.



Handstamp - Blue

Identity Group



Repair and Refinishing Spray

Multi-Tech Products Corp.



Annacell WB Finish

Mon-Eco Industries, Inc.



Black Tire Paint Concentrate

Akron Paint and Varnish (dba APV
Engineered Coatings)



IC 1-gl 2pk Gray Shop Coat Primer

Rust-Oleum Corporation

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Phthalate

Productabc

Manufacturer d

BBP

Klean-Strip Mask & Peel Paint Booth Coating

W. M. Ban-

Lacquer Touch-up Paint - Clear Topcoat

Ford Motor Company



SK Clear-Seal Satin Sealer 5 Gal

Rust-Oleum Corporation



3M Bondo Glazing & Spot Putty

3M Company



SureFlex Multi-Purpose Adhesive, SH-360

Barristo Enterprises, Inc. dba SureHold



Lanco Seal

Lanco Mfg. Corp.



PSI PolyClay Canes and PSI PolyClay Bricks e

Penn State Industries



Hydrostop Preiniumcoat Finish Coat

GAF



Hydrostop Preiniumcoat Foundation Coat

GAF



Hydrostop Trafficcoat Deck Coating

GAF

DBP

Pro 1-GL 2PK Flat Aluminum Primer

Rust-Oleum Corporation

DURALAQ-WB WATERBORNE WHITE
ACRYLIC FINISH DULL RUBBED

Benjamin Moore & Co.



Hydrostop Preiniumcoat Foundation Coat Summer

GAF



Bondo Gray Filler Primer

3M Company



Pettit XL Vivid 1861 Black

Kop-Coat, Inc. / Pettit Marine Paint



Accurate Solo 1000, Accurate LT-30, Accurate LT-





32, Accurate 2015, Accurate 2495, Accurate 4064,

Western Powders, Inc.



Accurate 4350





Cartridge 9 mm FX Marking, Toxfree primer

General Dynamics - Ordnance and Tactical
Systems - Canada Inc. [Canada]



Rimfire Blank Round - Circuit Breaker

Olin Corporation - Winchester Division, Inc.



Wizard 31 Epoxy Ball Plug Hardener

Brunswick Bowling Products, LLC



765-1553 BALKAMP VINYL REPAIR KIT

Pennatex, Inc.



Chocolate

Wellington Fragrance



PSI PolyClay Canes and PSI PolyClay Bricks8

Penn State Industries



DUPLI-COLOR BED ARMOR

Dupli-Color Products Company



DUPLI-COLOR High Performance Textured Metallic
Coating Charcoal

Dupli-Color Products Company



264 BLACK TRUCK BED LNR 6UC

The Valspar Corporation



RED GLAZING PUTTY 1# TUBE

The Valspar Corporation



Prime WPC/Prime Essentials/Prime SPC

Carlton Hardwood Flooring

DEHP

Lenox MetalMax

Lenox Tools



6.17 OZ 100040 FH FRESH SCENT PET TW 12PK

Fresh House



KRYLON Fusion All-in-One Textured Galaxy

Krylon Products Group



Self-cath pediatric 30 pack

Coloplast Corp.



3M™ Economy Vinyl Electrical Tape 1400, 1400C

3M



Pronto Putty

The Valspar Corporation



Red Glazing Putty 1# Tube

Quest Automotive Products



BD Loop Goop

Royal Adhesives and Sealants Canada Ltd.



SCOFIELD® CureSeal 350

Sika Corporation

DCHP

Duco Cement (bottle and tube)l

ITW Consumer - Devcon/Versachem

Fusor 108B, 109B Metal Bonding ADH PT B

LORD Corporation

DIBP

Blue Label Washable PVA Adhesive

Colorlord Ltd.

BETAKRIL TEXTURE

Betek Boya ve Kimya Sanayi A.S [Turkey]



Centerfire Pistol & Revolver and Rifle Cartridges

Companhia Brasileira de Cartuchos (CBC)

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Phthalate

Product

Manufacturer''

DIBP

Art Board

Ningbo Zhonghua Paper Co. Ltd.

Glitter Boards

DJECO

Painting - Oh, It's Magic

DJECO

" This table includes a sample of products listed in the Use Reports for each DBP, BBP, DIBP, DEHP, DCHP (U.S.
EPA. 202Id. 2020f. e. h. i. D.

h This table may represent updated information with products listed that are not identified in the published Use Reports.
c This is not a comprehensive list of products containing each phthalate nor does the presence of a product on this list
indicate its availability in the United States for consumer purchase

d Some manufacturers may appear over-represented in this table. This may mean that they are more likely to disclose
product ingredients online than other manufacturers, but this does not imply anything about use of the chemical
compared to other manufacturers in this sector.

e The SDS for PSI PolyClay Canes and PSI PolyClay Bricks, which lists the product as containing multiple phthalates is
available here: httDs://www.Dennstateind.com/MSDS/POLYCLAY MSDS.Ddf.

6.4.1.3 Combining Exposure to Consumers to Estimate Cumulative Exposure (Steps 8
and 9 in Conceptual Model [Figure 2-1])

As described in the final scope documents for BBP ('	). DBP (	). DCI IP

(I v	M DEHP (• ^ \ - >20b). DIBP ('. ! V \ .'.020c). and DINP ('. ! V \ :021c).

EPA will assess exposures to consumers for each COU outlined in the scope documents. Consumer
exposure to phthalate containing products will be assessed for individual COUs and will primarily focus
on inhalation and dermal exposures, however, oral exposures may be considered based on the use of the
product or the possibility for hand-mouth behavior to occur.

As described above in Section 6.4.1.2, there is currently a lack of evidence that multiple consumer
products are used concurrently by consumers. Therefore, EPA is not proposing to combine risk for co-
use of multiple consumer products for consumers, unless new information is identified to support doing
so. Similarly, due to the initial identification of a single product containing less than or equal to 2.5
percent (unspecified weight or volume) of DEHP, BBP, and DBP, EPA does not anticipate assessing the
cumulative exposure to multiple phthalates through the use of a single consumer product, unless new
product information is identified to support doing so. EPA believes that assessing risk for individual
COUs associated with a single phthalate is adequate because of the numerous product examples
containing greater than or equal to 10 percent (weight and volume) of single phthalates, which likely
exceeds the cumulative exposure associated with the use of the single identified product. Therefore,
EPA is unlikely to combine risk across multiple phthalates for a single consumer COU.

Consumers may have exposures to multiple phthalates through sources other than uses of consumer
products. Additional sources of exposure to multiple phthalates are captured in EPA's proposed non-
attributable and non-TSCA exposure estimate (Section 6.3.2). Therefore, to estimate cumulative
exposure to consumers, EPA proposes to combine non-attributable and non-TSCA exposure for the
high-priority and manufacturer-requested phthalates with exposure from use of a single consumer
product containing a single phthalate using Equation 6-2. Determining reasonable cumulative exposure
scenarios may involve considering the likelihood of co-exposure, the possibility of double counting, and
of over- or under-estimating exposures. The estimates for the non-attributable and non-TSCA portion of
Equation 6-2 may vary based on the consumer scenario and population being assessed and exposures
that comprise each category (i.e., TSCA, non-attributable, non-TSCA) can be adjusted to limit the
possibility of "double counting." For example, if the TSCA consumer exposures are known to highly
impact indoor air concentrations then the non-attributable indoor air concentrations may need to be
adjusted accordingly. Additionally, adults may have different exposure from infants or toddlers based on
exposure factors and interaction with different sources of exposure leading to potentially different

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estimates in all exposure categories with some phthalates being more or less impactful for different
lifestages.

Equation 6-2. Example estimation of cumulative phthalate exposure to consumers8

Consumer cumulative exposure (expressed as index chemical equivalents) = non-attributable exposure
+ non-TSCA exposure + individual consumer COU exposure

Because EPA is proposing to use an RPF approach (Section 4.3.3), phthalate exposure from each
individual COU will be scaled to the potency of an IC and expressed as IC equivalents, which will then
be summed with non-attributable and non-TSCA exposure (described in Section 6.3.2 and also
expressed as index chemical equivalents) to estimate consumer cumulative exposure (expressed as IC
chemical equivalents).

6.4.1.4 Estimating Cumulative Risk for Consumers (Steps 10 in Conceptual Model
[Figure 2-1])

To estimate cumulative risk for each specific consumer exposure scenario, an MOE (ratio of index
chemical POD to consumer cumulative exposure estimate (expressed as index chemical equivalents)
calculated using Equation 6-2) would be calculated for comparison to the benchmark MOE {i.e., the
total uncertainty factor associated with the assessment) (described in Section 4.3.2). The lower the MOE
(margin between the toxicity effect level and the exposure dose), the more likely a chemical is to pose a
risk.

6.4.2 Occupational Exposures and Risk

This section describes EPA's proposed approach for building cumulative exposure scenarios for workers
and estimating cumulative risk for workers. As stated previously in Section 5, EPA proposes to focus its
CRA for phthalates on subpopulations that may be more susceptible to phthalate syndrome which
include pregnant women/women of reproductive age who may be impacted by exposure from TSCA
occupational COUs, but the proposed approach will be presented as applicable to all workers. This
involves the following steps as outlined in EPA's conceptual model (Figure 2-1):

•	Step 6. Identifying major pathways of exposure. Determining the major pathways of exposure
from TSCA occupational COUs (see yellow box in Step 4 of conceptual model (Figure 2-1);
completed in individual risk evaluations), non-attributable, and non-TSCA sources. This step
would be completed after exposures are estimated for the various pathways of exposure and is
dependent on the magnitude of those estimates. Major pathways may vary by relevant
populations and may also vary by phthalate. Identification of major pathways of exposure to
relevant populations may require sensitivity analysis for determining inclusion of a pathway into
a cumulative estimate. Description of this process is not detailed in this document as it will be
dependent on the identified pathways.

•	Step 7. Determining co-exposure. Determining likelihood of co-exposure across TSCA
occupational COUs, non-attributable sources, and non-TSCA sources (Section 6.4.2.2).

8 EPA may consider an alternative to this proposed approach for COU categories such as floor coverings, fabric and textiles,
and building materials, which may contribute to concentrations of phthalates in indoor air or dust outside the period of direct
product use through ongoing releases. In this case, EPA may consider combining exposure from non-attributable sources,
non-TSCA sources, exposure from an individual consumer COU during use and exposure from a consumer COU due to
ongoing releases to estimate cumulative exposure. This approach will be considered by EPA as supported by available data.

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•	Step 8. Convert exposures to index chemical equivalents. Phthalate exposure from each
individual phthalate is scaled to the potency of an index chemical using RPFs and expressed in
units of index chemical equivalents (Section 6.4.2.3).

•	Step 9. Estimating cumulative exposure. Combining TSCA occupational COU cumulative
exposure, the relevant non-attributable cumulative exposure, and the non-TSCA cumulative
exposure to estimate cumulative exposure in a reasonable manner (Section 6.4.2.3).

•	Step 10. Estimating cumulative risk. A cumulative MOE is calculated for comparison to the
benchmark MOE (total uncertainty factor associated with the assessment) (Section 6.4.2.4).

As shown in EPA's conceptual model (Figure 2-1), workers may be exposed to multiple phthalates
through workplace exposures associated with TSCA COUs, as well through additional non-attributable
and non-TSCA sources (described in Section 6.3.2). Therefore, estimating cumulative risk to workers
will involve combining major sources of phthalate exposure resulting from TSCA occupational COU(s),
as well as additional non-attributable and non-TSCA sources that can be reasonably expected to co-
occur over a relevant timeframe.

In order to assess releases and cumulative risks in the workplace, EPA will rely on EPA program data
and data discovered through the systematic review process to determine if multiple phthalates are
present at occupational sites and/or releasing facilities leading to exposure to workers or fenceline
communities (discussed in Section 6.4.3). The following sections will discuss EPA's proposed approach
for assessing cumulative risk for workers and cumulative exposure resulting from facility releases. This
includes EPA's proposed approach for the following:

•	Identifying sites with potential for release and cumulative occupational exposure.

•	Quantifying cumulative worker exposure at sites and/or for specific COUs where multiple high-
priority and manufacturer-requested phthalates are anticipated to be in use.

•	Quantifying cumulative phthalate releases at sites and/or for specific COUs where multiple high-
priority and manufacturer-requested phthalates are anticipated to be in use.

6.4.2.1 Data Needs for Releases and Cumulative Occupational Exposure Assessment

An engineering assessment is typically comprised of three primary elements, including: (1) facility
estimates, which provide a basis for the scope of release and exposure and estimates; (2) environmental
release estimates, which estimate the quantity and release frequency of the chemical into the
environment; and (3) occupational exposure estimates, which estimates the chemical exposure to
workers and occupational non-users (ONUs) at a facility; and EPA plans to utilize available Agency
data and any additional data sources for a cumulative assessment.

Common sources used to complete occupational exposure and environmental release assessments are
listed below and are summarized for the high-priority and manufacturer-requested phthalates in Table
6-7:

•	Chemical Data Reporting (CDR), to which import and manufacturing sites producing the
chemical at or above a specified threshold must report.

•	Toxics Release Inventory (TRI), to which facilities handling a chemical covered by the TRI
program at or above a specified threshold must report.

•	Discharge Monitoring Report (DMR), a periodic report required of National Pollutant Discharge
Elimination System (NPDES) permitted facilities discharging to surface waters.

•	National Emissions Inventory (NEI), a compilation of air emissions of criteria pollutants, criteria
precursors and hazardous air pollutants from point and non-point source air emissions.

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Resource Conservation and Recovery Act (RCRA)Info, to which small and large quantity
generators of hazardous waste and treatment, storage and disposal facilities must report.

National Institute of Occupational Safety and Health: Health Hazard Evaluation (NIOSH HHE),
a compilation of voluntary employee, union, or employer requested evaluations of health hazards
present at a given workplace.

Occupational Safety and Health Administration: Chemical Exposure Health Data (OSHA
CEHD), a compilation of industrial hygiene samples taken when OSHA monitors worker
exposures to chemical hazards.

Table 6-7. Available EPA Program and Common Source Data for Each Phthalate

Chemical

CDR

TRI

DMR

NEI

RCRAInfo

NIOSH HHE

OSHA CEHD

DEHP

~

~

~

~

~

~

~

DBP

~

~

~

~

~

~

~

BBP

~

X

~

X

X

X

X

DIBP

~

X

X

X

X

X

X

DCHP

~

X

X

X

X

~

X

DINP

~

X

X

X

X

~

X

S Indicates data available
x Indicates no data available

As can be seen from Table 6-7, the six phthalates selected for release and cumulative occupational
exposure assessment represent a challenge due to the limited data available from EPA programs. Only
two of the six phthalates are reported to TRI and NEI and only three have recent DMR data. This leaves
large data gaps in assessing environmental releases and occupational exposures for certain phthalates
and certain COUs that will require alternative methods and data sources to fill.

Additionally, while EPA program data is useful for identifying the presence of multiple phthalates at a
single site and quantifying cumulative release and exposure, further data from literature and other
sources will be necessary to fully understand how the phthalates are being used at each site. Data needs
to complete the cumulative assessment include the following:

•	Chemical reaction pathways and functionality: Some chemicals can be substituted for others
based on cost or availability at processing and use sites. It is anticipated that this may be the case
with some of the phthalates. Understanding the chemical reaction pathway of each phthalate
within each COU and the functionality it provides the end product will allow for the
determination of likelihood that phthalates would be manufactured/processed/used at the same
facility; exist in a process at the same time; or be used as occasional replacements for each other.

•	Facility operating schedules: A site may manufacture, process, or use multiple phthalates over
the span of a year, but not at the same time, such as if the site runs separate campaigns using
different phthalates for different products. Detailed process information and data on how
frequently a worker is exposed to different phthalates (i.e., same day, consecutive days, or more
sporadically) can help inform more accurate cumulative exposure profiles and define how
cumulative exposures are compared to benchmarks for acute, sub-chronic, or chronic health
effects.

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•	Facility process descriptions and worker tasks: At sites where multiple phthalate exposures
happen as part of the same operation, it is important to understand whether the phthalates are
used simultaneously, consecutively, or separately and whether these tasks are performed by the
same worker or workers within the same process units.

•	Market data on phthalate manufacturers and processors: The EPA programmatic data
provides insight into producers and users of selected phthalates at quantities above the reporting
thresholds, but market data may enable a more complete understanding of all sites,
manufacturers through end-users, working with multiple phthalates. This could include supplier
and retailer data, phthalate market use data, and other data that could help generate a list of
companies that manufacture or process multiple phthalates.

•	End user product data and compatibility: An inventory of which phthalates are found in
various similar products will be important for understanding the potential for end-user
cumulative exposures. An end user, such as for the application of paints, coatings, adhesives, and
sealants, may utilize multiple similar products for a given task that could contain different
phthalates. Information that can inform such an inventory include SDS, product use data from
various sources, or general COU process information.

This represents EPA's initial assessment of data needs and potential data sources. Data needs and source
may change during the cumulative assessment of the six phthalates in the phthalate cumulative chemical
group.

6.4.2.2 Co-exposure Resulting from TSCA Occupational COUs (Step 7 in Conceptual
Model [Figure 2-1])

This section reviews the potential data sources that exist to quantify sites with potential cumulative
releases or exposures to the selected phthalates that may lead to cumulative exposures to workers or to
fenceline communities (discussed further in Section 6.4.3). Cumulative occupational exposure to the six
high-priority and manufacturer-requested phthalates results from multiple phthalates being handled at
the same site. There are several considerations for estimating cumulative risk for workers, including
considerations of

•	Multiple phthalate direct exposure: the worker is directly exposed to two or more of the high-
priority and manufacturer-requested phthalates in their job, but there are no indirect exposures
from additional phthalates.

•	Phthalate direct exposure + indirect workplace exposure: the worker is directly exposed to
one or more of the selected phthalates in their job and indirectly exposed to additional phthalates
that are present in the workplace but may not be directly part of their job.

ONU exposures may follow the same mechanisms as described above but are often difficult to quantify
due to lack of data. At minimum, a cumulative exposure from all operations contributing to an indirect
exposure in a workplace setting can be established to represent a workplace ONU exposure. Depending
on the quality of meta data provided, a documented ONU exposure to a given phthalate may be a
sufficient estimate of the indirect workplace exposure to that phthalate.

6.4.2.2.1 EPA Program Data for Identifying Sites to Determine Co-exposure

EPA will use its program data as a starting point for identifying sites with cumulative release and
exposure potential. EPA program data include CDR, TRI, DMR, NEI, and RCRAInfo listed above and
described in further detail in Appendix D.

There may be significant challenges with using EPA programmatic data to identify sites with cumulative
release and exposure potential. In CDR, for example, facilities may claim data as confidential business

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information or not known or reasonably ascertainable. It should still be possible to identify sites with
multiple phthalates, but these claims may make it challenging for EPA to use information claimed CBI,
such as number of workers or production volume. Additionally, the CDR, TRI, NEI programs, as well as
aspects of the RCRAInfo program, all have reporting thresholds such that sites below this threshold do
not report to the program. Furthermore, the TRI, NEI, and RCRAInfo programs only require reporting
for specific chemicals and only a limited number of phthalates report to these programs. Facilities not
reporting to these programs will need to be identified via other means, such as those described below.

6.4.2.2.2	NIOSH HHE, OSHA CEHD, and Other Literature Sources Data for
Identifying Sites to Determine Co-exposure

NIOSH HHEs, OSHA CEHD, and results from systematic review can be used in a similar way to
determine the presence of multiple phthalates at a given site. The HHEs and CEHD represent NIOSH
and OSHA inspections at specific sites and typically include monitoring data. Systematically reviewed
literature may also contain data identifying specific sites or may contain broader data such as process
information or product usage data that can inform the potential for unidentified sites to handle multiple
phthalates. These systematically reviewed literature sources include EPA generic scenarios (GSs) and
emission scenario documents (ESDs), which provide general information for a specific industry or

cou.

The site information from these sources will be compared to sites in the other referenced datasets,
including CDR, TRI, DMR, NEI, RCRA Info. For example, if a site inspected by NIOSH is reported to
have DBP on site in the HHE and that same site reported to TRI for DEHP, it may be assumed that both
DBP and DEHP are used at that site. There are several aspects to consider before drawing this
conclusion:

•	Temporal: The two datasets must be reasonably close in time period—a site's operations can
evolve over time along with the chemicals used and products manufactured.

•	Non-detects: Some monitoring regimes test for a spectrum of chemicals regardless of whether
they are expected to be present at the site; therefore, the presence of a non-detect in a report does
not necessarily mean the presence of that chemical at the site.

•	Biomonitoring data: Health studies may include urinary biomonitoring data that describes
potential exposures to multiple phthalates in the workplace. The study must be clear if the tested
metabolites are unique to a given phthalate present in order to be used in determining overlap
with other phthalates.

6.4.2.2.3	Product Information Data for Identifying Sites to Determine Co-exposure

A compilation of all phthalate-containing products and the companies producing them may be useful in
determining sites that process multiple phthalates. SDSs or ingredient lists for products can inform if the
products contain one or more of the high-priority and manufacturer-requested phthalates. The collected
product data may indicate if a site produces multiple phthalate-containing products and may potentially
inform end-user sites with cumulative exposures to the selected phthalates.

6.4.2.2.4	Identifying Additional Unknown Sites with Release and Exposure
Potential to Determine Co-exposure

As discussed previously, there are limitations to the data sources listed above such that they may not be
sufficient to fully identify all sites with cumulative release and exposure potential. To account for this
uncertainty, EPA plans to evaluate the potential for additional unknown/unidentified sites to have
cumulative release and exposure potential. For example, an exposure assessor could expand the analysis
of CDR, TRI, NEI, and DMR to include additional phthalates that are not part of the cumulative
assessment to gain an understanding of how many phthalates a facility might use in a given COU. These

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methods would not identify specific sites with cumulative release and exposure but could be used to
identify the potential for additional unknown sites to have multiple phthalates present beyond the ones
identified using the above data sources.

Table 6-8 summarizes each unique COU and the applicability of these COUs to the selected phthalates,
according to the published scope documents for BBP (U.S. EPA. 2020a). DBP (U.S. EPA. 2020d).
DCHP (U.S. EPA 2020e). DEHP (U.S. EPA 2020b). DffiP (U.S. EPA. 2020c). and DINP (U.S. EPA
2021c). COUs will be used for each of the six phthalates to inform the potential for an unknown site
within a given COU to handle multiple phthalates. For example, hydraulic fracturing is only applicable
to DEHP; therefore, it is unlikely that a hydraulic fracturing site will handle multiple phthalates.
However, it is important to note that there may be instances where separate COUs could exist within the
same facility. For example, DIBP is the only phthalate that is used in textile finishing, but both DEHP
and BBP are used in textile dyeing and DINP and DBP are used in textiles, apparel, and leather
manufacturing, all of which could be co-located with the same site. Each COU should be carefully
considered when evaluating the potential for a site to have cumulative releases and exposures from
multiple COUs and phthalates.

Table 6-8: Conditions of Use for Each High-Priority and Manufacturer-Requested Phthalate

Condition of Use (COU)

DEHP

DINP

DIBP

DBP

BBP

DCHP

Manufacturing

X

X

X

X

X

X

Repackaging

X

X

X

X

X

X

Processing as a reactant

X

X



X

X

X

Incorporation into formulation, mixture, or reaction
product

X

X

X





X

Industrial processing (not including formulation)







X





Plastics compounding

X

X

X

X

X

X

Plastics converting

X

X

X

X

X

X

Use in hydraulic fracturing

X











Application of finishing agents



X









Textiles, apparel, and leather manufacturing



X



X





Application of paints, coatings, adhesives, and sealants

X

X

X

X

X

X

Use of laboratory chemicals

X

X

X

X

X

X

Use of automotive care products

X

X





X



Use of ink, toner, and colorant products (e.g., printing)



X



X

X

X

Use of cleaning and furnishing care products



X



X

X



Use of hydraulic fluids



X









Textile dyeing

X







X



Textile finishing





X







Use of air care products



X









Application of finishing agents in cellulose film production











X

Use of fuels and related products





X







Manufacturing of plastic foam products



X









Soldering and welding

X











Castings









X



Use of flush fluids



X









Textiles, apparel and leather manufacturing

X











Explosives manufacturing







X





Use of chemiluminescent light sticks







X





Use of inspection penetrant kits







X





Use of lubricants







X





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Condition of Use (COU)

DEHP

DINP

DIBP

DBP

BBP

DCHP

N/A - Fabrication or use of final product/articles

X

X

X

X

X

X

Recycling

X

X

X

X

X

X

Waste handling, treatment, and disposal

X

X

X

X

X

X

6.4.2.2.5 Workplace Monitoring Data for Determining Co-Exposure

Inhalation monitoring data that is identified during evaluation of the NIOSH HHEs, OSHA CEHDs,
literature, and other sources will be utilized for the cumulative occupational exposure assessment. Data
from monitoring activities at sites known or expected to have multiple phthalates are preferred, as these
data may represent both direct exposures and indirect workplace exposures. The frequency of this data,
however, is anticipated to be low.

Without monitoring data from sites with multiple phthalates, monitoring data for a given phthalate will
be used within a given COU to establish direct exposure to that phthalate in that COU. In some
instances, there may be no monitoring data, in which case surrogate data from other COUs or from other
similar phthalates may be used. Understanding the chemical reaction pathways and use patterns of each
phthalate will help to gauge which COUs or phthalate monitoring data is best suited for use as a
surrogate. In the absence of suitable surrogate data within the selected phthalates, there may be some
literature sources, such as GSs and ESDs, with surrogate data that can be used.

Also, area monitoring data from various workplace locations or ONU monitoring data may be used to
establish indirect workplace exposures. Surrogate data as described above may also be used if available.

6.4.2.3 Combining Exposure to Workers to Estimate Cumulative Risk (Steps 8 and 9
in Conceptual Model [Figure 2-1])

With direct and indirect workplace exposures estimated via the methods above, exposure estimates can
be combined to estimate cumulative exposures as each occupational scenario requires. This task will
rely heavily on establishing worker relationships with the chemicals to which they are exposed and any
operational and exposure details provided within the monitoring data assessed in this section. Risks from
individual phthalates across various exposure routes, presented in the individual risk evaluations, will be
combined based on the available evidence of co-exposure to determine a cumulative risk across relevant
phthalates. EPA also plans to develop generic occupational exposure estimates for the unknown sites
with cumulative occupational exposure potential identified according to the methodology in Section
6.4.2.2.4.

As previously stated, EPA is defining cumulative occupational exposure as being exposed to more than
one high-priority and manufacturer-requested phthalate directly on the job and/or being exposed to a
single or multiple phthalates directly on the job in addition to being indirectly exposed to levels of
phthalates in the workplace. Cumulative occupational exposure can be used to determine cumulative risk
to relevant phthalates occurring in the workplace during the 8-hour workday. Workers may have
exposures to multiple phthalates through sources occuring outside the workday. Additional sources of
exposure to multiple high-priority and manufacturer-requested phthalates are captured in EPA's
proposed non-attributable and non-TSCA exposure estimates (Section 6.3.2). Therefore, to estimate
cumulative exposure to workers, EPA is proposing to combine the non-attributable and non-TSCA
exposure (as defined in Section 6.3.2) with the cumulative exposure from the workplace to determine a
total cumulative exposure for workers using Equation 6-3 and as shown in Figure 2-1.

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Equation 6-3. Example estimation of cumulative exposure to occupational subpopulations

Cumulative exposure to occupational subpopulations (expressed as index chemical equivalents) = Non-
attributable exposure + Non-TSCA exposure + Cumulative occupational exposure

Determining reasonable cumulative exposure scenarios may involve considering the likelihood of co-
exposure, the possibility of double counting, and of over- or under-estimating exposures. The estimates
for the non-attributable and non-TSCA portion of Equation 6-2 may differ between a worker and a
consumer. Therefore, exposures that comprise each category {i.e., TSCA, non-attributable, non-TSCA)
may be adjusted to limit the possibility of "double counting." For example, exposure from a non-
attributable source of indoor air, as determined through residential estimates, may need to be adjusted
for a worker who spends 8 hours at the workplace and less than 24 hours in the home.

Because EPA is proposing to use an RPF approach (see Section 4.3.3), exposure from each individual
phthalate identified as part of the cumulative occupational exposure will be scaled to the potency of an
index chemical and expressed as index chemical equivalents, which will then be summed with non-
attributable and non-TSCA exposure (described in Section 6.3.2 and also expressed as IC equivalents) to
estimate a cumulative exposure for each occupational subpopulation (expressed as index chemical
equivalents).

6.4.2.4 Estimating Cumulative Risk for Workers (Step 10 in Conceptual Model
[Figure 2-1])

To estimate cumulative risk for each specific occupational exposure scenario, an MOE (ratio of index
chemical POD to occupational cumulative exposure estimate (expressed as IC equivalents) calculated
using Equation 6-3) would be calculated for comparison to the benchmark MOE {i.e., the total
uncertainty factor associated with the assessment) (described in Section 4.3.2). The lower the MOE
(margin between the toxicity effect level and the exposure dose), the more likely a chemical is to pose a
risk.

6,4,3 General Population (Fenceline Communities) Exposures and Risk

This section describes EPA's proposed approach for building cumulative exposure scenarios for
fenceline communities, who are part of the general population. Generally, an assessment of cumulative
exposure and risk for a general population may include components of the proposed approach described
for consumers, workers, and fenceline communities as they are all part of the general population.
Cumulative exposure to a general population may include exposures from multiple phthalates from
TSCA, non-attributable, and non-TSCA sources of exposures which are reasonably combined to
determine cumulative risk.

The proposed approach focuses on assessing cumulative exposure and risk for the fenceline
communities as there are additional consideration for determining cumulative exposure and risk from
single or multiple facility releases. Additionally, fenceline communities may have a higher exposure due
to proximity to facilities and may be considered a highly exposed subpopulation within the general
population assessment conducted in the individual risk evaluations. Risk evaluations for the phthalates
will provide exposure estimates for oral, dermal, and inhalation exposures; these exposure estimates are
not available at this time and fenceline communities have not yet been identified as having higher
exposures.

As stated previously in Section 5, EPA proposes to focus its CRA for phthalates on groups that may be
more susceptible to phthalate syndrome which include pregnant women/women of reproductive age, and
male infants, male toddlers, and male children who may be impacted by exposure from TSCA releases,

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but the proposed approach will be presented as applicable generally to a fenceline community. This
involves the following steps for estimating cumulative risk for fenceline communities as outlined in
EPA's conceptual model (Figure 2-1):

•	Step 6. Identifying major pathways of exposure: Determining the major pathways of exposure
from TSCA COUs (see green box in Step 4 of conceptual model (Figure 2-1); completed in
individual risk evaluations), non-attributable, and non-TSCA sources. This step would be
completed after exposures are estimated for the various pathways of exposure and is dependent
on the magnitude of those estimates. Major pathways may vary by relevant population and may
also vary by phthalate. Identification of major pathways of exposure to relevant populations may
require sensitivity analysis for determining inclusion of a pathway into a cumulative estimate.
Description of this process is not detailed in this document as it will be dependent on the
identified pathways.

•	Step 7. Determining co-exposure: Determining likelihood of co-exposure across TSCA COUs,
non-attributable sources, and non-TSCA sources (Section 6.4.3.2).

•	Step 8. Convert exposures to IC equivalents: Phthalate exposure from each individual
phthalate is scaled to the potency of an index chemical using RPFs and expressed in units of
index chemical equivalents (Section 6.4.3.3).

•	Step 9. Estimating cumulative exposure: Combining cumulative exposure fenceline
communities, the relevant non-attributable cumulative exposure, and the non-TSCA cumulative
exposure to estimate cumulative exposure in a reasonable manner (Section 6.4.3.3).

•	Step 10. Estimating cumulative risk: A cumulative MOE is calculated for comparison to the
benchmark MOE (total uncertainty factor associated with the assessment).

EPA's Draft TSCA Screening Level Approach for Assessing Ambient Air and Water Exposures to
Fenceline Communities, Version 1.0, (U.S. EPA. 2022) [hereinafter referred to as EPA's Draft Fenceline
Approach], defines fenceline communities as

Members of the general population that are in proximity to air emitting facilities or a
receiving waterbody, and who therefore may be disproportionately exposed to a chemical
undergoing risk evaluation under TSCA section (6). For the air pathway, proximity goes
out to 10,000 meters from an air emitting source. For the water pathway, proximity does
not refer to a specific distance measured from a receiving waterbody, but rather to those
members of the general population that may interact with the receiving waterbody and
thus may be exposed.

Fenceline communities may have greater exposure to a chemical from being near an emitting source or
interacting with a receiving waterbody. EPA's Draft Fenceline Approach focused on single chemical
exposures for fenceline communities (	22). There are unique considerations for estimating

cumulative risk to fenceline communities, including the estimation of cumulative environmental releases
from facilities and combining exposure across facility releases, COUs, and relevant pathways to
estimate cumulative risk. These considerations are discussed in the following sections.

6.4.3.1 Data Needs for General Population/Fenceline Community Exposure
Assessment

Data needs for conducting environmental release assessments that are utilized for determining the major
pathways of exposure and determining the potential for co-exposure to multiple phthalates to fenceline
communities were discussed previously in Sections 6.4.2.1 to 6.4.2.2, as well as Appendix D. Briefly,
these data sources may include

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•	EPA programmatic data (i.e., CDR, TRI, DMR, NEI, and RCRAInfo);

•	Monitoring data from other agencies such as NIOSH HHEs and OSHA CEHD;

•	Systematically reviewed literature that include EPA generic scenarios (GSs) and emission
scenario documents (ESD) to provide general information for a specific industry or COU; and

•	Surrogate release data from other COUs, other similar phthalates, or other release sites not
included in the potential cumulative release.

6.4.3.2 Co-exposure Resulting from TSCA COUs (Step 7 in Conceptual Model [Figure
	2-11)	

Fenceline communities may be exposed to more than one high-priority and manufacturer-requested
phthalate due to TSCA COUs. This may occur when

•	a single facility releases more than one phthalate to the ambient air or receiving waterbodies;

•	multiple TSCA facilities in close proximity release more than one phthalate to ambient air or
receiving waterbodies; and

•	a fenceline community is near one or more facilities releasing phthalates but is also being
exposed through consumer or occupational COUs.

These fenceline communities near one or more facilities releasing individual phthalates will be
identified in individual risk evaluations. Additionally, fenceline communities may be exposed to the
high-priority and manufacturer-requested phthalates through non-attributable and non-TSCA sources
(described in Section 6.3.2). Because fenceline communities may be exposed to more than one phthalate
undergoing TSCA risk evaluation due to facility releases, non-attributable sources, and non-TSCA
sources, EPA is proposing to evaluate fenceline communities for cumulative risk (Figure 2-1). Sections
6.4.3.2.1 to 6.4.3.2.2 describe the process for identifying the fenceline communities that may be near a
single facility releasing more than one phthalate or near multiple TSCA facilities releasing one or more
phthalates.

6.4.3.2.1 Using Reported Release Data to Determine Co-exposure

EPA programmatic data, including TRI, DMR, NEI, and RCRAInfo, will be utilized to quantify
environmental releases to the media listed in Table 6-9. Among the four EPA programs in Table 6-9,
there is overlap for each release media. However, release estimates for a given site may vary by
program. At each site where there is potential for cumulative release of phthalates, release estimates
from each programmatic database will be cataloged for each phthalate. In many instances, the program
data will not cover all potential releases for a given site due to the limited coverage of the selected
phthalates in these programs (Table 6-7). In these instances, relevant release data from literature sources
may be utilized. For COUs where no environmental release or estimation data from literature exists,
modeling approaches will be considered as described in Section 6.4.3.2.2.

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Table 6-9. Media of Release Covered by EPA Programs

Media of Release Covered in Dataset

EPA Program

TRI

DMR

NEI

RCRAInfo

Air (fugitive and stack)

~

X

~

X

Surface Water

~

S

X

X

POTW/WWTP

~

X

X

~

Landfill

~

X

X

~

Incineration

~

X

X

~

Energy Recovery

~

X

X

~

Injection

~

X

X

~

Reclamation/Recycling

~

X

X

~

Other

~

X

X

~

S Indicates data available
x Indicates no data available

Pertinent site location data such as site address and latitude and longitude, as well as wastewater
discharge data including receiving water bodies and publicly owned treatment works (POTWs) or
wastewater treatment plants (WWTP), allows for analysis of site proximity.

6.4.3.2.2 Surrogate Release Data for Determining Co-exposure

In the absence of suitable release data, EPA will rely on modeling to estimate cumulative environmental
releases. Specifically, EPA will first consider surrogate release data from other COUs, other similar
phthalates, or other release sites not included in the potential cumulative release or exposure selection
specified in Section 6.4.2.2. Understanding the chemical reaction pathways and use patterns of each
phthalate will help to gauge which COUs or phthalate release data is best suited for use as a surrogate. In
the absence of suitable surrogate data within the selected phthalates, EPA will consider using surrogate
data for other similar phthalates or chemicals from literature sources, if available.

In the absence of surrogate data for cumulative environmental releases, mathematical modeling
approaches may be utilized. EPA may incorporate the use of Monte Carlo simulation to vary release
calculation input parameters to estimate central tendency and high-end releases. Frequently used
literature sources will be used to inform the input parameter distributions.

6.4.3.3 Combining Exposure to General Population (Fenceline Communities) to

Estimate Cumulative Risk (Steps 8 and 9 in Conceptual Model [Figure 2-1])

There are multiple ways that EPA may consider combining exposures to estimate phthalate cumulative
risk to fenceline communities. These possible approaches are described below.

1. Cumulative Exposure to Air Releases

With phthalate environmental releases estimated via the described methods above, release estimates may
be combined to estimate cumulative releases from single sites and adjacent sites. EPA proposes to
follow tiered methodologies described in EPA's Draft Fenceline Approach (U.S. EPA 2022) for
estimating ambient air and water concentrations of the high-priority and manufacturer-requested
phthalates resulting from facility releases in the individual risk evaluations. In considering combining
exposures for cumulative risk, as stated in Section 6.4.3.2 cumulative environmental exposure potential
of the six high-priority and manufacturer-requested phthalates can result from multiple phthalates being
released at the same site or two different sites that handle more than one phthalate being located adjacent

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to one another. If cumulative releases are identified from a single site, ambient air concentrations may be
aggregated as appropriate across COUs for a single phthalate and the resulting inhalation risk from
individual phthalates may be combined to determine cumulative risk from facility air releases using
Equation 6-4.

Equation 6-4. Example estimation of cumulative exposure from single facility releases to air

Cumulative exposure (air) (expressed as index chemical equivalents) =

Phthalate A inhalation exposure (facility #1) +

Phthalate B inhalation exposure (facility #1) + ...

2.	Cumulative Exposure to Surface Water

Similarly, as appropriate, surface water concentrations in a receiving water body resulting from a single
facility release may be aggregated across COUs for individual phthalates and the resulting dermal and
incidental ingestion risk from individual phthalates may be combined to determine cumulative risk from
use of the receiving water body using Equation 6-5.

Equation 6-5. Example estimation of phthalate cumulative exposure from single facility
releases to surface water

Cumulative exposure (surface water) (expressed as index chemical equivalents) =

Phthalate A incidental ingestion (facility #1) + Phthalate A dermal exposure (facility #1) +
Phthalate B incidental ingestion (facility #1) +

Phthalate B incidental dermal exposure (facility #1) + ...

In Equation 6-4 and Equation 6-5, "Phthalate A" and "Phthalate B" could be any of the six
toxicologically similar phthalates under consideration. Because EPA is proposing to use an RPF
approach (Section 4.3.3), phthalate exposure from each facility release would be scaled to the potency of
an index chemical and expressed as index chemical equivalents, and them summed to estimate
cumulative exposure.

3.	Cumulative Exposure to Air and Water from Multiple Facilities

For scenarios where multiple sites that handle more than one phthalate are located adjacent to one
another, cumulative exposure can be calculated similarly as above using Equation 6-6 for releases to air
and Equation 6-7 for releases to surface water.

Equation 6-6. Example estimation of cumulative exposure from multiple facility releases to
air

Cumulative exposure (air) (expressed as index chemical equivalents) =

Phthalate A inhalation exposure (facility #1) +

Phthalate B inhalation exposure (facility #2) + ...

Equation 6-7. Example estimation of cumulative exposure from multiple facility releases to
surface water

Cumulative exposure (surface water) (expressed as index chemical equivalents) =

Phthalate A incidental ingestion (facility #1) + Phthalate A dermal exposure (facility #1) +
Phthalate B incidental ingestion (facility #2) +

Phthalate B incidental dermal exposure (facility #2) + ...

There are unique challenges associated with estimating cumulative risk from exposure to more than one
high-priority and manufacturer-requested phthalate for the drinking water pathway. For example,
concentrations at the point of actual drinking water intake are difficult to estimate because of the

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potential for transport, dilution, and treatment of drinking water from a given distance away from the
receiving waterbody. Depending on the available data and methods, cumulative risk from drinking water
attributable to TSCA releases may be included as appropriate. Drinking water concentrations may also
be included as a non-attributable exposure if it is not able to be attributed to a TSCA COU.

4.	Cumulative Exposure to Fenceline Communities Who Are Not Consumers or Workers

Because non-attributable and non-TSCA risk as described in Section 6.3.2, may include exposure
pathways such as ambient air, drinking water, and surface water, that may lead to exposures potentially
not as high as those for fenceline communities estimated using Equation 6-2, EPA may not utilize the
same non-attributable and non-TSCA risk used for consumers or workers. Instead, EPA may consider
major exposure pathways to relevant phthalates separately for fenceline communities. Determining
reasonable cumulative exposure scenarios may involve considering the likelihood of co-exposure, the
possibility of double counting, and of over- or under-estimating exposures. This may mean that the
intake estimates for ambient air, drinking water, and surface water which are determined based on
facility releases for fenceline communities are combined with risk from other major relevant exposure
pathways comprising a unique non-attributable and non-TSCA risk for fenceline communities which
does not already include ambient air, drinking water, and surface water, but does include major
identified pathways that may include dust (non-attributable) and diet (non-TSCA) to help avoid double
counting as shown in Equation 6-8. As stated previously, estimates of cumulative exposure for different
lifestages may differ based on exposure factors and interaction with different sources of exposure
leading to potentially different estimates in all exposure categories with some phthalates being more or
less impactful for different lifestages.

Equation 6-8. Example estimation of cumulative exposure to fenceline populations who are
not consumers or workers

Cumulative exposure to fenceline subpopulations (expressed as index chemical equivalents) =
Non-attributable exposure (not including ambient air and surface water) + Non-TSCA exposure +
Cumulative facility exposure (including ambient air and surface water)

5.	Cumulative Exposure to Fenceline Communities Who May Also Be Consumers and
Workers

Additionally, individuals who are part of the fenceline communities may be consumers and workers
living near the facilities. For these instances, additional combinations of exposure should be considered.
For example, cumulative exposure for an individual living near a TSCA facility who is also a consumer
of TSCA COUs and works in a facility handling TSCA COUs, would require consideration of exposures
from facility releases near the home, workplace exposure over the 8-hour workday, and exposure from
the use of consumer products at the home. EPA proposes that for these individuals, cumulative exposure
could include the cumulative occupational exposure to TSCA COUs as discussed in Section 6.4.2,
exposure from consumer TSCA COUs as discussed in Section 6.4.1, and the estimated cumulative
exposure to fenceline populations presented in Equation 6-8 to determine an estimated cumulative
exposure to fenceline populations who are also consumers and workers as shown in Equation 6-9.

Equation 6-9. Example estimation of cumulative exposure to fenceline populations who are
also consumers and workers

Cumulative exposure to fenceline/consumer/occupational (expressed as index chemical equivalents)
= Non-attributable exposure (not including ambient air and surface water) + Non-TSCA exposure
+ Cumulative facility exposure (including ambient air and surface water) + Cumulative
occupational exposure + Consumer COU exposure

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Combining exposures for these populations may require additional data or evidence not already covered
in this document to determine the major pathways that contribute to cumulative exposure (Step 6 in
conceptual model) to those individuals and support the likelihood of co-exposure to multiple phthalates
from the various pathways of exposure (Step 7 in conceptual model). Based on the outlined approaches
for the various populations {i.e., consumers [Section 6.4.1], workers [Section 6.4.2], fenceline
communities [Section 6.4.3]), reasonable combinations of exposure may be considered, as data allows.

EPA has not identified a proposed methodology, data sources, or lines of evidence to fully develop the
cumulative fenceline assessment. In the absence of data or evidence, assumptions may be necessary to
determine reasonable combinations of exposure for identified populations, which involve considering
the likelihood of co-exposure, the possibility of double counting, and of over- or under-estimating
exposures. EPA will be soliciting comments from the SACC and the public on this issue.

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7 SUMMARY OF PROPOSED APPROACH AND NEXT STEPS

This document describes EPA's proposed approach for assessing high-priority and manufacturer-
requested phthalates for cumulative risk to human health under TSCA. This document was prepared
based on the principles of CRA described in EPA's Draft Proposed Principles of CRA Under TSCA. As
discussed in Section 2, there are two primary considerations for grouping chemical substances for
inclusion in a CRA, including (1) toxicologic similarity and (2) evidence of co-exposure over a relevant
timeframe. To determine which high-priority and manufacturer-requested phthalates are toxicologically
similar, EPA reviewed data for seven key outcomes associated with phthalate syndrome {i.e., fetal
testicular gene expression and testosterone, decreased AGD, NR, hypospadias, seminiferous tubule
atrophy, and MNG formation). Based on the weight of evidence, EPA proposes that DEHP, BBP, DBP,
DIBP, DCHP, and DINP, but not DIDP, are toxicologically similar and induce effects on the developing
male reproductive system consistent with phthalate syndrome (Section 3.1.7). To determine if the U.S.
population is co-exposed to multiple phthalates EPA reviewed NHANES urinary biomonitoring data and
TSCA industrial, commercial and consumer use data (Section 3.2). Available biomonitoring data
demonstrate that the U.S. population is co-exposed to multiple phthalates, including DEHP, BBP, DBP,
DIBP, DINP, and DIDP, while co-exposure to DCHP is anticipated to occur through various industrial,
commercial and consumer uses under TSCA. These data qualitatively demonstrate that humans are co-
exposed to DEHP, BBP, DBP, DIBP, DCHP, DINP, and DIDP. EPA's proposed approach for
quantifying phthalate co-exposure is outlined in Section 6. Based on evidence of toxicologic similarity
and co-exposure, EPA is proposing to group DEHP, BBP, DBP, DIBP, DCHP and DINP for CRA under
TSCA (Step 1 in conceptual model [Figure 2-1]).

As discussed in Section 4.1, NRC presented two options for assessing the risks of phthalate syndrome
following exposures to phthalates, including assessing the syndrome as a whole and focusing on the
most sensitive effect associated with the syndrome (NRC. 2008). EPA identified a number of challenges
associated with addressing phthalate syndrome as a whole and therefore EPA is proposing to address
phthalate syndrome by focusing on the most sensitive effect (Section 4.1.3).

As described in EPA's Draft Proposed Principles of CRA under TSCA and in Section 4.2 of this
document, several additivity approaches can be used to assess multiple chemical substances for
cumulative risk to human health—including dose addition, response addition, and integrated addition as
well as approaches that account for toxicologic interactions (	XX), 1986). EPA is proposing

to rely upon a default assumption of dose addition when conducting CRAs for toxicologically similar
chemicals under TSCA. As described in Section 3.1.7, EPA considers there to be sufficient evidence to
conclude that DEHP, BBP, DBP, DIBP, DCHP, and DINP are toxicologically similar. Therefore, EPA
is proposing to assess these six phthalate for cumulative risk to human health under an assumption of
dose addition. In further support of this, EPA identified multiple in vivo phthalate mixture studies that
provide empirical evidence to support the use of dose additivity models and EPA's proposal to use dose
addition is consistent with the recommendations of the NRC (2008).

EPA is considering the applicability of two component-based, dose additive approaches, including the
HI and RPF approaches. Based on currently available data, EPA considers there to be sufficient data
available to support RPF derivation for the six toxicologically similar TCSA phthalates (Section 4.3.3).
To support RPF derivation, EPA considered the strengths and uncertainties associated with the dataset
for each of the seven evaluated key outcomes (Section 4.4.1). Given the strengths and uncertainties
associated with the datasets for each key outcome, EPA is proposing several options to derive RPFs
based on gestational {i.e., reduced fetal testicular testosterone content and reduced testicular
steroidogenic gene expression) and postnatal outcomes {i.e., reduced AGD, NR, seminiferous tubule
atrophy, and hypospadias) (Section 4.4.2).

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EPA will conduct consumer, occupational, and general population exposure assessments for each
individual phthalate risk evaluation. The key human populations considered in these exposure
assessments include consumers, workers, and the general population, including fenceline communities.
Within these populations, there are susceptible subpopulations with greater susceptibility to phthalate
syndrome based on lifestages, including pregnant women, women of reproductive age, and male infants,
toddlers, and children (Step 2 in conceptual model [Figure 2-1]). These groups are the focus of EPA's
CRA (Section 5).

To estimate cumulative exposure for subpopulations with increased susceptibility to phthalate syndrome
based on lifestages, EPA is proposing to consider exposures resulting from TSCA COUs (Section 6.2),
as well as non-attributable and non-TSCA exposures (Section 6.3.2). Prior to the development of the
phthalate CRA, exposure scenarios for TSCA COUs will be completed in individual phthalate risk
evaluations (Steps 3 to 4 in conceptual model). EPA is proposing to include non-attributable and non-
TSCA exposures as part of the phthalate CRA because certain non-attributable (e.g., dust) and non-
TSCA (e.g., dietary) pathways are anticipated to be major contributors to phthalate exposure that
contribute to cumulative risk (Section 6.2.2).

EPA is considering two approaches for estimating non-attributable and non-TSCA phthalate exposure,
including a scenario-based approach (Section 6.3.2.1) and a reverse dosimetry based approach (Section
6.3.2.2). The scenario-based approach involves estimating non-attributable and non-TSCA exposure to
populations of interest based on the concentrations of phthalates in various media, food, and other
sources using population specific exposure factors (e.g., inhalation rate, dietary intake, body weight,
etc.) (Section 6.3.2.1).The reverse dosimetry approach involves estimating aggregate exposure for each
individual phthalate from human urinary biomonitoring data for metabolites unique to each individual
parent phthalate as reported in nationally representative datasets, such as NHANES (Section 6.3.2.2).
Because the reverse dosimetry approach does not distinguish between routes or pathways of exposure
and does not allow for source apportionment, it provides an estimate of total non-attributable phthalate
exposure (Section 6.3.2.2). As described in Section 6.3.2.5, EPA is proposing to estimate non-
attributable and non-TSCA exposure for DEHP, BBP, DBP, DIBP, DCHP, and DINP from major
exposure pathways using a scenario-based approach (Step 5 in conceptual model), while the reverse
dosimetry approach, which does not allow for source apportionment, may be used to help characterize
phthalate exposure and serve as a comparator for scenario-based intake estimates.

As shown in EPA's draft conceptual model (Figure 2-1), EPA is proposing to assess consumers (Section
6.4.1), workers (Section 6.4.2), and general population/fenceline communities (Section 6.4.3) for
cumulative risk from exposure to DEHP, BBP, DBP, DIBP, DCHP, and DINP through TSCA COUs.
EPA proposes to identify major pathways of exposure and likelihood of co-exposure to multiple
phthalates through various pathways for combining to estimate cumulative exposure to identified
susceptible subpopulations based on lifestages (Steps 6 to 7 in conceptual model). To estimate
cumulative exposure to consumers (Section 6.4.16.4), EPA proposes to combine the non-attributable and
non-TSCA exposure with exposure from individual consumer COUs. To estimate cumulative exposure
to workers (Section 6.4.2), EPA proposes to combine the non-attributable and non-TSCA exposure with
cumulative occupational exposure from TSCA COUs in a work setting. For cumulative exposure to
fenceline communities (Section 6.4.3), EPA proposes estimating cumulative exposures from single or
multiple facility releases to ambient air and/or water and combining with non-attributable and non-
TSCA exposures. Because EPA is proposing to use an RPF approach (Section 4.3.3), exposure from
individual phthalates for each exposure scenario will be scaled to the potency of an index chemical and
expressed as index chemical equivalents (Step 8 in conceptual model), which will then be summed to
estimate cumulative exposure for each exposure scenario (Step 9 in conceptual model). Cumulative risk

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4335	may then be estimated using a margin of exposure (MOE) approach (Section 4.3.3) (Step 10 in

4336	conceptual model).

4337

4338	EPA is soliciting comments from the SACC on charge questions and comments from the public for the

4339	SACC meeting scheduled on May 8-11, 2023.

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REFERENCES

A.CC HPP. (2019a). Manufacturer request for risk evaluation Di-isodecyl Phthalate (DIDP). American

Chemistry Council High Phthalates Panel.

ACC HEP. (2019b). Manufacturer Request for Risk Evaluation Di-isononyl Phthalate (DINP).
(730R19001). American Chemistry Council High Phthalates Panel :: ACC HPP

https://nepis.epa.gov/exe/ZvPIJRL.cgi?Dockev=Pl 00YEGF.txt
Adamssoii \ c.alonen. \ l\tranko. J; Toppari. J. (2009). Effects of maternal exposure to di-

isononylphthalate (DINP) and l,l-dichloro-2,2-bis(p-chlorophenyl)ethylene (p,p'-DDE) on
steroidogenesis in the fetal rat testis and adrenal gland. Reprod Toxicol 28: 66-74.
http://dx.doi.ore	Teprotox.2009.03.002

Adham. IM; Em men. JM; En gel. W. (2000). The role of the testicular factor INSL3 in establishing the
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Ahbab. IV rlas. N. (2015). Influence of in utero di-n-hexyl phthalate and dicyclohexyl phthalate on
fetal testicular development in rats. Toxicol Lett 233: 125-137.
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Ahmad. R; Gautam. AK; Vernu \ Sedha. S: Kum.u c. (2014). Effects of in utero di-butyl phthalate
and butyl benzyl phthalate exposure on offspring development and male reproduction of rat.
Environ Sci Pollut Res Int 21: 3156-3165. http://dx.doi.	jil-x

Albert. O; Jegois (2014). A critical assessment of the endocrine susceptibility of the human testis to
phthalates from fetal life to adulthood [Review], Hum Reprod Update 20: 231-249.
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Anderson \\ \ t 1 Hird. S: letTery. J: Scotter. MI. (2011). A twenty-volunteer study using

deuterium labelling to determine the kinetics and fractional excretion of primary and secondary
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49: 2022-2029. http://dx.doi.oiv 10 101 /jjit _ 01 I 0 01 >

Anderson. WAC: Castle. L; Scotter. Ml; Massey. RC: Springall. C. (2001). A biomarker approach to
measuring human dietary exposure to certain phthalate diesters. Food Addit Contam 18: 1068-
1074. http://dx.doi.org/IO. i 080/026520301100501 I '<

Andrade. V randc S\\ , i ,dsness. CE; Gericke. C; Grote. K; Golombiewski. A; Sterner-Kock. A;

Chahoud. I. (2006a). A dose response study following in utero and lactational exposure to di-(2-
ethylhexyl) phthalate (DEHP): Reproductive effects on adult male offspring rats. Toxicology
228: 85-97. http://dx.doi.oi	'i.tox.2006.08.020

Andrade. V randc S\\ , i ,dsness. CE; Grote. K; Golombiewski \ c.u;rner-Kock \ i iiahoud. I.

(2006b). A dose-response study following in utero and lactational exposure to di-(2-ethylhexyl)
phthalate (DEHP): Effects on androgenic status, developmental landmarks and testicular
histology in male offspring rats. Toxicology 225: 64-74.
http://dx.doi.ore	.tox.2006.05.007

Arbuckle. TE; Davis. K; Marro. L; Fisher. M; Legrand. M; Leblan.' \ * ;audreai> < t < >ster. WG;

Choeurng. V; Fraser. WD. (2014). Phthalate and bisphenol A exposure among pregnant women
in Canada—results from the MIREC study. Environ Int 68: 55-65.
http://dx.doi.ore 10 101 | .envint.2014.0_ 010
Arbuckle. TE; Fisher. M; Macpherson. S; Lam 1 h<» /encher < ^blan»% \ llaus-n ^ t ^Hey. M;
Ayotte. P; Neisz unsay. T; Tawagi. G. (2016). Maternal and early life exposure to
phthalates: The Plastics and Personal-care Products use in Pregnancy (P4) study [Supplemental
Data], Sci Total Environ 551-552: 344-356. http://dx.doi.oiv 10 101 'j.scitotenv .01 02.022
Arzuaga. X; Walker. T; Yost. EE; Radke. EG; Hotchkiss. AK. (2020). Use of the Adverse Outcome
Pathway (AOP) framework to evaluate species concordance and human relevance of Dibutyl
phthalate (DBP)-induced male reproductive toxicity. Reprod Toxicol 96: 445-458.

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Ashley-Martin. J; Dodd1- < \ t buckle. TE; Lanphear H, Muck ^ ^ ter. WG; Ayotte. P; Zidek. A;
Asztalos. E; Bouchard. MF; Kuhle. S. (2021). Urinary phthalates and body mass index in
preschool children: The MIREC Child Development Plus study. Int J Hyg Environ Health 232:
113689. http://dx.doi.ore/10 101 /i.iiheh..^. I l l '89
Aso. S: Ehara. H; Miyata. K; Hosvuvama. S: Shiraishi. K; Umano. T; Minobe. Y. (2005). A two-

generation reproductive toxicity study of butyl benzyl phthalate in rats. J Toxicol Sci 30: S39-58.
http://dx.doi.ore	ts.30.S39

AT SDR. (2022). Toxicological profile for di(2-ethylhexyl)phthalate (DEHP) [ATSDR Tox Profile],

(CS274127-A). Atlanta, GA. https://www.atsdr.cdc.eov/ToxProfiles/tp9.pdf
Aylw	Hays. SM; Zidek. A. (2016). Variation in urinary spot sample, 24 h samples, and longer-

term average urinary concentrations of short-lived environmental chemicals: implications for
exposure assessment and reverse dosimetry. J Expo Sci Environ Epidemiol 27: 582-590.
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Barlow. NJ: Mclnty	ster. PM. (2004). Male reproductive tract lesions at 6, 12, and 18 months

of age following in utero exposure to di(n-butyl) phthalate. Toxicol Pathol 32: 79-90.
http://dx.doi.orE 30/01926230490265894
Beves	ht. CS; Furr. JR; Sampson. H; Wilson. VS: Mclntvre. BS; Foster. PMD; Travlos.

G; Gray. LE. Jr. (2014). Simvastatin and Dipentyl Phthalate Lower Ex Vivo Testicular
Testosterone Production and Exhibit Additive Effects on Testicular Testosterone and Gene
Expression Via Distinct Mechanistic Pathways in the Fetal Rat. Toxicol Sci 141: 524-537.
http://dx.doi.ore ?3/toxsci/kful49
Biesterbc'1 <\\ 1 Hidzina. T; Delmaar. CI; Bakker. MI; Russel. t \ on Goetz. N: Scheepers. PT;

Roeleveld. N. (2013). Usage patterns of personal care products: important factors for exposure
assessment. Food Chem Toxicol 55: 8-17. http://dx.doi.ore/10.1016/j.fct.1'0 i J I I 01 I
Birkhgj. M; Nellemann. C; Jarfelt. K; Jacobs	.ersen. HR; Dalgaard. M; Vineeaard. AM.

(2004). The combined antiandrogenic effects of five commonly used pesticides. Toxicol Appl
Pharmacol 201: 10-20. http://dx.doi.oi ^ 10 101 i/j.taap.-OO I 0 I 01
Blessineer. TP; Euline. SY; Warn 1 ,m. KA; Cai. C; Klinefelter. G: Saillenfait. AM. (2020).

Ordinal dose-response modeling approach for the phthalate syndrome. Environ Int 134: 105287.
http://dx.doi.ore 10 101 i.envint.-^M ..87
Blystone. CR; Kissline. GE; Bishc	tpin. RE; Wolfe. GW: Foster. PM. (2010). Determination of

the di-(2-ethylhexyl) phthalate NOAEL for reproductive development in the rat: importance of
the retention of extra animals to adulthood. Toxicol Sci 116: 640-646.
http://dx.doi.ore 10 10°'< toxsci/kln I I
Bobere. J; Christiansen. S; Axelstad. M; Ktedat. TS; Vineeaard. A eaard. M; Nellemann. C; Hass.
U. (2011). Reproductive and behavioral effects of diisononyl phthalate (DINP) in perinatally
exposed rats. Reprod Toxicol 31: 200-209. http://dx.doi.ore/10.1016/j .reprotox.2010 I I 001
Bobere. J; Metzdorff. S; Wortzieer. R; Axelstad. M; Brokken. L; Vineeaard. AM; Dal eaard. M;
Nellemann. C. (2008). Impact of diisobutyl phthalate and other PPAR agonists on
steroidogenesis and plasma insulin and leptin levels in fetal rats. Toxicology 250: 75-81.
http://dx.doi.ore	.tox.2008.05.020

Boekelheide. K; Kleymenova. E; Liu. K; Swanson. C; Gaido. KW. (2009). Dose-dependent effects on
cell proliferation, seminiferous tubules, and male germ cells in the fetal rat testis following
exposure to di(n-butyl) phthalate. Microsc Res Tech 72: 629-638.
http://dx.doi.ore/10.1002/iemt.20684
Borch < \\elstad. M; Vineeaji*! \ VI; Dal eaard. M. (2006a). Diisobutyl phthalate has comparable
anti-androgenic effects to di-n-butyl phthalate in fetal rat testis. Toxicol Lett 163: 183-190.
http://dx.doi.ore 10 101 i.toxlet.lW ? 10 0JO

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Borch. J; iadefoeed. O; Hass. U; Vinggaard. AM. (2004). Steroidogenesis in fetal male rats is reduced
by DEHP and DINP, but endocrine effects of DEHP are not modulated by DEHA in fetal,
prepubertal and adult male rats. Reprod Toxicol 18: 53-61.

http://dx.doi.ore/10.1016/i .reprotox.200 '< 10 01 I
Borch. J; Metzdorff. SB; Ytnggaai	okken. L; Dalgaard. M. (2006b). Mechanisms underlying

the anti-androgenic effects of diethylhexyl phthalate in fetal rat testis. Toxicology 223: 144-155.
http://dx.doi.org 10 101 |.tox.200 0 '< 01
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Corning Hazleton Inc. (1996). Initial submission: Oncogenicity study in rats of di(2-ethylhexyl)

phthalate including hepatocellular proliferation and biochemical analysis, w/TSCA cover sheet
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8EHQ-1196-13805. TSCATS/444561). Kingsport, TN: Eastman Kodak.

Cutty. M; Thuillier. R; Li. W; Wane. Y; Martine elles. D; Benjamin. C; Triantafilou. K; Zirkin. B;
Papadopoulos. V. (2008). In utero exposure to di-(2-ethylhexyl) phthalate exerts both short-term
and long-lasting suppressive effects on testosterone production in the rat. Biol Reprod 78: 1018-
1028. http://dx.doi.org/10.1095/biotrepro lQVa440a
Do. RP; Stahlhut. RW; Ponzi. D: Voro Saal. FS; Taylor. J A. (2012). Non-monotonic dose effects of in
utero exposure to di(2-ethylhexyl) phthalate (DEHP) on testicular and serum testosterone and
anogenital distance in male mouse fetuses. Reprod Toxicol 34: 614-621.
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Postal. LA; Chapin. RE; Stefanski. SA; Harris. MW; Schw	(1988). Testicular toxicity and

reduced Sertoli cell numbers in neonatal rats by di(2-ethylhexyl) phthalate and the recovery of
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Drab an den Driesche. S; Scott. HM; Hutchison. GR; Seckl. JR; Sharpe. RM. (2009).

Glucocorticoids amplify dibutyl phthalate-induced disruption of testosterone production and
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http://dx.doi.ore/10.1210/en.2009-0700
EC/HC. (2015a). Proposed approach for cumulative risk assessment of certain phthalates under the

Chemicals Management Plan. Gatineau, Quebec: Environment Canada, Health Canada.

EC/HC. (2015b). State of the science report: Phthalate substance grouping 1,2-Benzenedicarboxylic
acid, diisononyl ester; 1,2-Benzenedicarboxylic acid, di-C8-10-branched alkyl esters, C9-rich
(Diisononyl Phthalate; DINP). Chemical Abstracts Service Registry Numbers: 28553-12-0 and
68515-48-0. Gatineau, Quebec, https://www.ec.ec.ca/ese~
ees/default.asp?lane=En&n=47F58AA5~l
EC/HC. (2015c). State of the science report: Phthalate substance grouping: Medium-chain phthalate
esters: Chemical Abstracts Service Registry Numbers: 84-61-7; 84-64-0; 84-69-5; 523-31-9;
5334-09-8; 16883-83-3; 27215-22-1; 27987-25-3; 68515-40-2; 71888-89-6. Gatineau, Quebec:
Environment Canada, Health Canada. https://www.ec.ec.ca/ese-ees/4D84519	^8B~

5B3E5/SoS Phthalates%20%28Medium-chain%29 EN.pdf
EC/HC. (2015d). State of the science report: Phthalate substance grouping: Short-chain phthalate esters:
1,2-Benzenedicarboxylic acid, dimethyl ester (DMP). Chemical Abstracts Service Registry
Number: 131-11-3. Gatineau, Quebec: Environment Canada, Health Canada.
https://www.ee. ec.ca/ese-ees/default.asp?lane=En&	94-1

EC/HC. (2015e). State of the science report: Phthalates substance grouping: Long-chain phthalate esters.
1,2-Benzenedicarboxylic acid, diisodecyl ester (diisodecyl phthalate; DIDP) and 1,2-
Benzenedicarboxylic acid, diundecyl ester (diundecyl phthalate; DUP). Chemical Abstracts
Service Registry Numbers: 26761-40-0, 68515-49-1; 3648-20-2. Gatineau, Quebec:

Environment Canada, Health Canada, https://www.ec.ec.ca/ese~
ees/default.asp?lane=En&n=D3FB0F30-l
ECCC/HC. (2020). Screening assessment - Phthalate substance grouping. (En 14-393/2019E-PDF).
Environment and Climate Change Canada, Health Canada.

https://www.canada.ca/en/environment-climate-change/services/evaluating-existing-
substances/screenine-assessment-phthalate-substance-eroupine.html

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ECH.A. (201 1). Annex XV restriction report: Proposal for a restriction, version 2. Substance name:
bis(2-ehtylhexyl)phthlate (DEHP), benzyl butyl phthalate (BBP), dibutyl phthalate (DBP),
diisobutyl phthalate (DIBP). Copenhagen, Denmark: Danish Environmental Protection Agency ::
Danish EPA, https://echa.europa.eu/documents/) ' l . * l l _ >-45c2-bf48-8890876f.t rs
ECHA. (2013). Evaluation of new scientific evidence concerning DINP and DIDP in relation to entry 52
of Annex XVII to REACH Regulation (EC) No 1907/2006. Helsinki, Finland.
http://echa.europa.eu/documents/101 .	 I* \\ .H . e. DG; Borghoff. S; Johnson. U I Hall. SJ; Boekelheide. K
(2007). Fetal mouse phthalate exposure shows that gonocyte multinucleation is not associated

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with decreased testicular testosterone. Toxicol Sci 97: 491-503.
http://dx.doi.ore ?3/toxsci/kfm049
Graham. PR. (1973). Phthalate ester plasticizers-why and how they are used. Environ Health Perspect 3:

3-12. http://dx.doi.org/10.12S9/ehp.73Q33
Grande. SW; Andra	Isness. CE; Grote. K; Chahoud. I. (2006). A dose-response study

following in utero and lactational exposure to di(2-ethylhexyl)phthalate: effects on female rat
reproductive development. Toxicol Sci 91: 247-254. http://dx.doij	3/toxsci/kfj 128

Girh 1 . Hnilow. N: Howdeshell. K; Ostbv. J; Furr. J: Gray. C. (2009). Transgenerational effects of Di
(2-ethylhexyl) phthalate in the male CRL:CD(SD) rat: added value of assessing multiple
offspring per litter. Toxicol Sci 1 10: 41 1 -425. http://dx.doi.	>3/toxsci/kfpl09

Gray. LE. Jr.: Lambright. CS; Conlev. JM; Evans. N: Furr. JR; Hannas. BR; Wilson. VS: Sampson. H;
Foster. P (2021). Genomic and Hormonal Biomarkers of Phthalate-lnduced Male Rat
Reproductive Developmental Toxicity Part II: A Targeted RT-qPCR Array Approach That
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Variability of urinary concentrations of phthalate metabolites during pregnancy in first morning
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automated counting procedure for phthalate-induced testicular multinucleated germ cells.

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Spade [all. SJ; Saffarini. C; Huse. SM; McDonnell. EV; Boekelheide. K. (2014). Differential
response to abiraterone acetate and di-n-butyl phthalate in an androgen-sensitive human fetal
testis xenograft bioassay. Toxicol Sci 138: 148-160. http://dx.doi.ore/10.1093/toxsci/kft266
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public/Estimated%2QPhthalate%2QExposure%2Qand%2QRisk%2Qto%2QWomen%2Qof%2QRepr
oductive%2QAge%2Qas%2QAssessed%2QUsing%2Q2Q 13%202014%20NHA.NES%20Biomonito
rin g%20Data. pdf

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risks/eeneral-principles-performine-aeereeate-exposure
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09/docum ents/can cer guidelines final 3-25-05.pdf
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Report], (EPA/600/R-05/062F). Washington, DC.

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Potential options and methods for evaluating the cumulative hazard associated with six selected
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U.S. EPA. (2019a). Guidelines for human exposure assessment [EPA Report], (EPA/100/B-19/001).
Washington, DC: Risk Assessment Forum, https://www.epa.gov/sites/production/files/2020-
01/documents/ guidelines for human exposure assessment finaC If
U.S. EPA. (2019b). Proposed designation of butyl benzyl phthalate (CASRN 85-68-7) as a high priority
substance for risk evaluation.

(2019c). Proposed designation of Di-Ethylhexyl Phthalate (DEHP) (1,2-Benzene-
dicarboxylic acid, 1,2-bis (2-ethylhexyl) ester) (CASRN 117-81-7) as a high-priority substance
for risk evaluation. Washington, DC: Office of Pollution Prevention and Toxics.
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, oposeddesignation 08 22 V1 [\< !f
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Priority Substance for Risk Evaluation. https://www.epa.gov/sites/production/files/2Q19-
08/docum ents/dii sobuty Iphth alate 84-69-5 high-priority proposeddesignation082:

U.S. EPA. (2019e). Proposed designation of Dibutyl Phthalate (CASRN 84-74-2) as a high-priority

substance for risk evaluation. U.S. Environmental Protection Agency, Office of Chemical Safety
and Pollution Prevention, https://www.epa.gov/sites/prodiiction/files/2019-
08/documents/dibutv Iphth alate 84-74-2 high-priority proposeddesignation 082 '< r" pi If
U.S. EPA. (2019f). Proposed designation of dicyclohexyl phthalate (CASRN 84-61-7) as a high-priority
substance for risk evaluation (pp. 1-21). U.S. Environmental Protection Agency, Office of
Chemical Safety and Pollution Prevention, https://www .regulations, gov/docum es IPA-HQ-
QPPT-2018-0504-0009
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benzenedicarboxylic acid, 1-butyl 2-(phenylmethyl) ester); CASRN 85-68-7 [EPA Report],
(EPA-740-R-20-015). Washington, DC: Office of Chemical Safety and Pollution Prevention.
https://www.epa.gOv/sites/defaiilt/files/2020-09/documents/casm 85-68-
tyl benzyl phthalate finalscope.pdf
U.S. EPA. (2020b). Final scope of the risk evaluation for di-ethylhexyl phthalate (1,2-

benzenedicarboxylic acid, l,2-bis(2-ethylhexyl) ester); CASRN 117-81-7 [EPA Report], (EPA-
740-R-20-017). Washington, DC: Office of Chemical Safety and Pollution Prevention.
https://www.epa.gov/sites/defaiilt/files/2020-Q9/dociiments/ca5	di-

ethylhexyl phthalate final scope.pdf
U.S. EPA. (2020c). Final scope of the risk evaluation for di-i sobutyl phthalate (1,2-benzenedicarboxylic
acid, l,2-bis(2-methylpropyl) ester); CASRN 84-69-5 [EPA Report], (EPA-740-R-20-018).
Washington, DC: Office of Chemical Safety and Pollution Prevention.
https://www.epa.gov/sites/defaiilt/files/2020-09/documents/casm 84-69-5 di-
isobutyl phthalate final scope.pdf
U.S. EPA. (2020d). Final scope of the risk evaluation for dibutyl phthalate (1,2-benzenedicarboxylic

acid, 1,2-dibutyl ester); CASRN 84-74-2 [EPA Report], (EPA-740-R-20-016). Washington, DC:
Office of Chemical Safety and Pollution Prevention.
https://www.epa.gov/sites/default/files/2020-09/documents/cas
2 dibutyl phthalate final scope 0 |;df
U.S. EPA. (2020e). Final scope of the risk evaluation for dicyclohexyl phthalate (1,2-

benzenedicarboxylic acid, 1,2-dicyclohexyl ester); CASRN 84-61-7 [EPA Report], (EPA-740-R-
20-019). Washington, DC: Office of Chemical Safety and Pollution Prevention.
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;yctohexyt phthalate final scope.pdf
U.S. EPA. (2020f). Use report for butyl benzyl phthalate (BBP) - 1,2-Benzenedicarboxylic acid, 1- butyl
2(phenylmethyl) ester (CAS RN 85-68-7). (EPA-HQ-OPPT-2018-0501-0035). Washington, DC:

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2018-0501-0035

U.S. EPA. (2020g). Use report for di-ethylhexyl phthalate (CAS RN 117-81-7). (EPA-HQ-OPPT-2018-

0433-0024).	Washington, DC: U.S. Environmental Protection Agency.

https://www.reeiilations.eov/dociiment/EPA-HQ-Q]	0433-0024

U.S. EPA. (2020h). Use report for di-isobutyl phthalate (CAS RN 84-69-5). (EPA-HQ-OPPT-2018-

0434-0029).	Washington, DC: U.S. Environmental Protection Agency.

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U.S. EPA. (2020i). Use report for dibutyl phthalate (DBP) - (1,2-Benzenedicarboxylic acid, 1,2- dibutyl
ester) (CAS RN 84-74-2). (EPA-HQ-OPPT-2018-0503-0023). Washington, DC: U.S.
Environmental Protection Agency. https://www.reeiilations.eov/dociiment/EPA-HQ-QPPT-
2018-0503-0023

I v ii \ (2020j). Use report for dicyclohexyl phthalate (CAS RN 84-61-7). (EPA-HQ-OPPT-2018-
0504-0030). Washington, DC: U.S. Environmental Protection Agency.

https://www.reeiilations.eov/dociiment/EPA-HQ-Q]	0504-0030

U.S. EPA. (2021a). About the Exposure Factors Handbook [Website],

https://www.epa.eov/expobox/aboiit-exposiire-factors-handbook
1 v M \ (2021b). Final scope of the risk evaluation for di-isodecyl phthalate (DIDP) (1,2-

benzenedicarboxylic acid, 1,2-diisodecyl ester and 1,2-benzenedicarboxylic acid, di-C9-ll-
branched alkyl esters, ClO-rich); CASRN 26761-40-0 and 68515-49-1 [EPA Report], (EPA-740-
R-21-001). Washington, DC: Office of Chemical Safety and Pollution Prevention.
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final-scope.pdf

U.S. EPA. (2021 c). Final scope of the risk evaluation for di-isononyl phthalate (DINP) (1,2-benzene-
dicarboxylic acid, 1,2-diisononyl ester, and 1,2-benzenedicarboxylic acid, di-C8-10-branched
alkyl esters, C9-rich); CASRNs 28553-12-0 and 68515-48-0 [EPA Report], (EPA-740-R-21-
002). Washington, DC: Office of Chemical Safety and Pollution Prevention.

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final-scope.pdf

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1,2-diisononyl ester, and 1,2-benzenedicarboxylic acid, di-C8-10-branched alkyl esters, C9-rich)
(CASRN 28553-12-0 and 68515-48-0). (EPA-HQ-OPPT-2018-0436-0035). Washington, DC:
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2018-0436-0035

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analysing the relevance of a non-cancer mode of action for humans. Geneva, Switzerland: World
Health Organization.

http://www.who.int/ipcs/methods/harmonization/areas/cancer mode.pdf?ua=l
Wilson. VS; Lambright v	.< UHln < \\<<<«! t Hetu > I •! (2004). Phthalate ester-

induced gubernacular lesions are associated with reduced insl3 gene expression in the fetal rat
testis. Toxicol Lett 146: 207-215. http://dx.doi.oi^ 10 101 /i.toxlet.200 '< 0° 01.

Wine. RN; Li. LH; Barnes. LH; Gulati. DK; Chapin. RE. (1997). Reproductive toxicity of di-n-
butylphthalate in a continuous breeding protocol in Sprague-Dawley rats. Environ Health
Perspect 105: 102-107. http://dx.doi.ore/10.1289/ehp.97105102
Wohlfahrt-V Main. KM; Skakkebaek. NE. (2009). Testicular dysgenesis syndrome: foetal origin
of adult reproductive problems [Review], Clin Endocrinol 71: 459-465.
http://dx.doi.ore	:265.2009.03545.x

Wormuth. M; Scheringer. I enweider. M; Huneerbuhler. K. (2006). What are the sources of
exposure to eight frequently used phthalic acid esters in Europeans? Risk Anal 26: 803-824.
http://dx.doi.ore	>924.2006.00770.x

Xiao-Fene. Z; Nai-Qiang. O; line. Z; Zi. g. Z. (2009). Di (n-butyl) phthalate inhibits testosterone
synthesis through a glucocorticoid-mediated pathway in rats. Int J Toxicol 28: 448-456.
http://dx.doi.ore	3342596

Yamasaki. K; Okuda. H; Takeuchi. T; Minot (2009). Effects of in utero through lactational

exposure to dicyclohexyl phthalate and p,p'-DDE in Sprague-Dawley rats. Toxicol Lett 189: 14-
20. http://dx.doi.ore/10.1016/i.toxlet.2009.04.023
Yost. EE; Euline. SY; Weaver <\ r%evei h  \ sineer. T;
Dishaw. L; Hotchkiss. A; Makris. SL. (2019). Hazards of diisobutyl phthalate (D1BP) exposure:
A systematic review of animal toxicology studies [Review], Environ Int 125: 579-594.
http://dx.doi.ore	.envint.2018.09.038

Zhane. Y; Jiang. X; Chen. B. (2004). Reproductive and developmental toxicity in F1 Sprague-Dawley
male rats exposed to di-n-butyl phthalate in utero and during lactation and determination of its
NOAEL. Reprod Toxicol 18: 669-676. http://dx.doi.ore/10.1016/j.reprotox.2004.04.009
Zki I i\ l\\ \»driaii.n HH, \kiiu. MS; Choi. EK; Tsunekawa. N; Kanai. Y; Kurohmaru. M.

(2010). Effects of di-iso-butyl phthalate on testes of prepubertal rats and mice. Okajimas Folia
Anat Jpn 86: 129-136. http://dx.doi.ore/10.2535/ofai.86.129

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APPENDICES

Appendix A Phthalate Cumulative Risk Assessment Initiatives

This appendix briefly summarizes the approaches used by U.S. CPSC (Appendix A.l), Health Canada
(Appendix A.2), Danish EPA (Appendix A.3), Australia NICNAS (Appendix A.4), and EFSA
(Appendix A. 5) to assess phthalates for cumulative risk. TableApx A-l provides a summary of high-
priority and manufacturer-requested phthalates included in the CRAs conducted by each regulatory
body. PODs used in previous phthalate CRAs are summarized in Appendix A.6.

Table Apx A-l. Summary of Phthalates Included in Previous CRAs

Regulatory Agency

DEHP

DIBP

DBP

BBP

DC H P

DINP

DIDP

U.S. CPSC

V

V

V

V

-

V

X

Health Canadaa

V

V

V

V



V

X

NICNAS6

V

-

V

V

-

V

-

EFSA

V

-

V

V

-

V

X

Danish EPA

V



V

V

-

-

-

"Health Canada included 16 phthalates in

their CRA, including six high-priority and manufacturer-requested

phthalates and 10 phthalates not being considered under TSCA.









Australia NICNAS has conducted five phthalate CRAs, which in addition to the listed phthalates, have also

included di(methoxyethyl) (DMEP), dimethyl (DMP), and diethyl (DEP) phthalates.





^Included in the CRA















x Excluded from the CRA; studies indicate no effects consistent with phthalate syndrome.



- Not considered as part of CRA planning.











A.l United States Consumer Product Safety Commission

In their report to the U.S. CPSC, the Chronic Hazard Advisory Panel (CHAP) on Phthalates and
Phthalate Alternatives assessed five phthalates (i.e., BBP, DBP, DIBP, DEHP, DINP) for cumulative
risk (US CPSC. 2014). In considering the best available approach for phthalates, the CHAP concluded
that experimental data on combination effects of phthalates from multiple studies provide strong
evidence that dose addition produces good approximations of mixtures effects. Male developmental and
reproductive effects occurring via an antiandrogenic MOA served as the basis of the CRA. Three sets of
anti androgenic PODs were selected, including (1) the anti androgenic PODs published by Kortenkamp
and Faust (2010); (2) PODs derived based on reduced fetal testosterone data published in Hannas et al.
(2011); and (3) PODs identified by the CHAP via a de novo literature review. The PODs based on data
from Hannas et al. ( ) were derived using relative potency assumptions. DEHP was used as the
index chemical. The NOAEL for DEHP-induced testosterone modulated effects was 5 mg/kg/day, and
DIBP, DBP and BBP were approximately equipotent (i.e., RPFs for DIBP, DBP, and BBP were all 1),
while DINP was 2.3 times less potent than DEHP (DINP NOAEL =11.5 mg/kg/day DEHP equivalent
units). The three sets of PODs used by the CHAP are summarized in TableApx A-3.The CHAP
considered including DIDP in the CRA but concluded that there is no evidence that DIDP causes
phthalate syndrome-related effects in experimental models, so DIDP was excluded from the analysis.

For the exposure assessment, the CHAP assessed cumulative exposure for various groups, including
women of reproductive age, pregnant women, and infants (2 to 36 months), using a reverse dosimetry
approach with human urinary biomonitoring data. CDC's NHANES urinary biomonitoring data from the
2005 to 2006 cycle was used to estimate cumulative exposure of pregnant women in the general U.S.
population. Infants are not included in the NHANES study design, so urinary biomonitoring data from

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mother/infant pairs reported in the Study for Future Families (Sathyanaravana et ai. 2008b;
Sathyanaravana et al. 2008a). which is a multicenter pregnancy cohort study, was used to estimate
exposure to infants and provided a second measure of cumulative exposure for pregnant women. In their
2015 report, CHAP revised their original analysis using the 2005/2006 NFLANES data based on updates
the CDC made to data files, including demographic and phthalates data, available to the public and
included newer analysis for 2007/2008, 2009/2010, and 2011/2012 cycles of NFLANES data Q1S
CPSC. 2015). In their 2017 report, U.S. CPSC analyzed the 2013/2014 cycle of NHANES data (

CPSC. 2017). In both updated reports, U.S. CPSC stated that analysis for pregnant women were not
updated because NHANES datasets following the 2005/2006 cycle did not include an oversampling of
pregnant women leading to a small sample size unsuitable for statistical analysis (CDC. 2013b; NCHS.
2012). Reverse dosimetry was used to calculate daily intake values for each parent phthalate using the
methodology published by Koch et al. (2007).

CPSC also explored a scenario-based method for determining aggregate exposure from all pathways. To
estimate total intake, the CHAP grouped sources and scenarios into the following categories: diet,
prescription drugs, toys, child-care articles, personal care products, indoor environment, and outdoor
environment. The total exposure from phthalates was assessed for each residue dataset, food
categorization scheme, and population (infant, toddler, children, teen, adult) using a deterministic
approach, calculating average and 95th percentile total exposure values. Although their scenario-based
modeled exposure estimates were not used to estimate cumulative risk, they concluded that their
scenario-based modeled estimates were in general agreement with the daily intake values derived from
biomonitoring data used for calculating cumulative risk (U .$ CPSC. 2014). Estimated phthalate
exposure by individual exposure scenario for women is shown in Figure Apx A-l.

To characterize risk, the CHAP applied the hazard index (HI) approach. His were calculated for each
individual based on their own unique phthalate urinary exposure profile, which is in contrast to the
standard HI approach in which population-level exposure statistics {i.e., mean, median, 95th percentile)
are used. Based on each individual's exposure profile, His were calculated using the three different sets
of PODs described above and summarized in Table Apx A-3. Based on 2005 to 2006 NHANES data,
approximately 9 to 10 percent of pregnant women had HI values >1.0 indicating risk, while <1 to 4
percent of women and 4 to 5 percent of infants in the SFF dataset had HI values >1.0. In all cases, the
HQ for DEHP contributed the most to the calculated His, while HQs for DINP, DBP, and BBP were
approximately similar, and DIBP consistently had the smallest HQs. CPSC found in their cumulative
assessment that the main sources of phthalate exposure for pregnant women/women of reproductive age
were food, beverage, and drugs via direct ingestion. Exposure to infants to phthalates was also primarily
through diet but exposure to DINP also occurred through mouthing of toys and teethers, and dermal
contact with personal care products (U.S. CPSC. 2014).

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Source

DEP

DBP

DIBP

BBP

DNOP

DEHP

DINP

DIDP

ave. 0.95

ave. 0.95

ave. 0.95

ave. 0.95

ave. 0.95

ave. 0.95

ave. 0.95

ave. 0.95

Total

1.8
E+01

4.0
E+02

2.9
E-01

5.7
E-00

1.5
E-01

5.0
E-01

1.1

E+OO

2.6
E-00

1.7
E-01

2.1
E+01

1.6
E+00

5.6
E+00

5.1
E+00

3.3
E+01

3.2
E+00

1.2
E-01

Diet

9.3
E-02

3.6
E-01

7.8
E-02

2.3
E-01

1.3
E-01

4.6
E-01

1.6
E-01

2.5
E-01

1.3
E-01

3.6
E-01

1.4

E+00

4.9
E+OO

4.8
E+00

1.5
E+01

3.2
E+00

9.3
E+00

Drags*

1.4
E+01

3.7
E+02





























Personal
care, dermal

































Shampoo

U
E-02

6.5
E-02





























Soap / bodv
wash

2.3
E-02

4.1
E-02





























Lotion

5.0
E-02

1.8
E-01





























Deodorant

7.4
E-01

1.9
E+01



























Perfume

2.8
E+00

6.2
E+00





























Nail polish

• 4
£-03

1.5
E-02

1.7
E-01

5,4
E+00

























Hair spray

4.7
E-02

1.4
E-01





























Personal care,
inhalation1































Deodorant

5.1
E-02

1.3
E+00





























Perfume

2.0
E-01

4.2
E-01



























Hair spray

6.2
E-03

1.8
E-02





























Dermal,
PVCc

































Toys'*

















8.0
E-03

8.0
E-03

8.0
E-03

8,0
E-03

6,7
E-03

6.7
E-03

1.1
E-03

II

E-03

Furniture*

















0.0
E+00

2.0
E+01





0.0
E+00

1-7
E+01

0.0
E+00

2.9
E+OO

Gloves













2.3
E-01

2.3
E-Ol

3.3
E-02

3.3
E-02

3.3
E-02

3.3
E-02

2.8
E-02

2.S
E-02

4.7
E-03

4.7
E-03

Household-
dermal'

































Paint'
lacquer













5.4
E-04

1,5
E-03









2.5
E-05

0.0
E+00





Adhesive













1.0
E-03

3.6
E-03

















Household,
inhalation'

































Air

freshener,
spray"

1.1
E-01

3.6
E-01

1.6
E-05

2.0
E-05

























Aii-

freshener,
liquid

1.5
E-02

4.0
E-02

9.2
E-06

2.4
E-05

6.8
E-06

9.8
E-06





















Paint, spray'

Indirect
ingestion













6.6
E-01

2.0
E+OO









1.5
E-01

3.1
E-01





































Dust

3.4
E-03

4.3
E-03

1.1
E-02

IS

E-02

12

E-03

2.0
E-03

5.0
E-02

II

E-01





2.0
E-01

3.4
E-01

5.2
E-02

4.0
E-01

1.4
E-02

4.4
E-02

Soil





9.3
EJ*5

4.3
E-05





1.6
E-06

6.9
E-06

3.5
E-06

1.1
E-05

7,2
E-05

3.1
E-04

2.1
E-05

8.1
E-05





Itihalatiou.
ail-































Indoor air

9.5
E-02

2.4
E-01

3.3
E-02

7.4
E-02

1.8
E-02

4 4

E-02

3.8
E-03

8.9
E-03

5.9
E+05

5.9
E-05

1.5
E-02

2.9
E-02









Outdoor air

14

E-03

3.8
E-03

8.4
E-05

3.6
E-04

S.6
E-05

2.6
E-04

7 2
E-05

1_2
E-04

S.4
E-06

8.4
E-06

4.8
E-04

2.9
E-03







Adult toys8

















3.8
E-04

8.0
E-02

1.9
E-04

2.6
E-01









5281	FigureApx A-l. Estimated Phthalate Exposure by Individual Exposure Scenario for Women

5282	Adapted from Table El-Sl in (U.S. CPSC. 2014).

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A.2 Health Canada

In June 2020, Health Canada and Environment and Climate Change Canada published their final
screening assessment of 28 phthalates (ECCC/HC. 2020). Phthalates were divided into three subgroups
{i.e., short-, medium-, and long-chain) based on the length of the carbon backbone in the ester side-
group. Based on a structure activity relationship analysis, Health Canada concluded that there was
evidence that medium-chain phthalates (but not short- (EC/HC. 2015d) or long-chain (EC/HC. 2015eV)
are capable of eliciting effects on the developing male reproductive system through a common MOA
{i.e., through a disruption of androgen action) (EC/HC. 2015c; Health Canada. 2015). Based on this
finding, Health Canada assessed 16 medium-chain phthalates for cumulative risk to the general
Canadian population.9

Combined exposure for three sensitive subpopulations {i.e., pregnant women/women of childbearing
age, infants, and children) was assessed using two approaches. First, occurrence data for environmental
media {e.g., dust, air, drinking water, soil, etc.) and food was used to estimate daily intake values.
Second, human urinary biomonitoring data for phthalate metabolites was used to estimate daily intake
values for parent phthalates using reverse dosimetry.

Health Canada estimated daily intake for BBP, DBP, DEHP, and DINP through routes of exposure to
environmental media and food for the general population, including ambient air, indoor air, drinking
water, food and beverages, breast milk, soil, and dust. Exposure scenarios included detailed assumptions
regarding daily intake of each phthalate via each route of exposure and separate estimates by varying
age groups and categories. Additional pathways were assessed for specific populations and phthalates,
including dermal and inhalation (aerosol) exposure to personal care products for adults and infants, and
DIBP and DINP in children's toys and articles. However, based on regulatory status of products or
determinations of minimal exposures, not all pathways of exposure were included in total intake
estimates. This is shown in Appendix D of (ECCC/HC. 2020) (see Tables D-1 a, D-2a, D-3a, D6) where
central tendency and upper bound estimates of exposure through ambient air, indoor air, drinking water,
food and beverages, soil, and dust are combined for each phthalate to estimate aggregate exposure.
Health Canada then generated distributions of phthalate exposure using a probabilistic exposure
assessment, randomly selecting phthalate concentrations for each food from the matched sources.
Exposure estimates from each food were summed for each individual, and a distribution of exposure was
generated for all respondents. This process was then iterated 500 times to model the variability of the
distribution of exposures.

Health Canada also utilized urinary biomonitoring data for six phthalates {i.e., DEHP, DBP, DINP,
DIBP, BBP, DCHP) to estimate daily intake values using reverse dosimetry (ECCC/HC. 2020). To do
this, Health Canada utilized urinary biomonitoring data from several sources, including the Canadian
Health Measures Survey (Cycle 1 (2007 to 2009) and 2 (2009 to 2011) data); U.S. CDC NHANES
(2009 to 2010 survey data); the Maternal Infant Research on Environmental Chemicals study which
includes urinary biomonitoring data from 2008 to 2011 for approximately 2,000 Canadian women
during their first trimester of pregnancy (Arbuckle et ai. 2014); the Maternal Infant Research on
Environmental Chemicals - Child Development Plus study, which includes urinary biomonitoring data

9 Medium-chain phthalates assessed by Health Canada included six high-priority and manufacturer-requested phthalates
(DIBP, DCHP, DINP, BBP, DBP, DEHP) and 10 phthalates not undergoing risk evaluation at EPA, including: butyl
cyclohexyl phthalate (BCHP, CASRN 84-64-0), dibenzyl phthalate (DBzP, CASRN 523-31-9), cyclohexyl isobutyl phthalate
(CHIBP, CASRN 5334-09-8), benzyl 3-isobutyryloxyl-l-isopropyl-2,2-dimethylpropyl phthalate (B84P, CASRN 16883-83-
3), benzyl isooctyl phthalate (BIOP, CASRN 27215-22-1), bis(methylcyclohexyl)phthalate (DMCHP, CASRN 27987-25-3),
benzyl octyl phthalate (B79P, CASRN 68515-40-2), diisoheptyl phthalate (DIHepP, CASRN 71888-89-6), diisooctyl
phthalate (DIOP, CASRN 27554-26-3), and dihexyl ester phthalate (DnHP, CASRN 84-75-3).

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from 2013 to 2015 for approximately 200 Canadian children aged 2 to 5 years (Ashley-Martin et ai.
2021); and the Plastics and Personal Care Product Use in Pregnancy survey, which includes
biomonitoring data from 2009 to 2010 for 80 mother-infant pairs from Ottawa Canada (Arbuckle et ai.
2016V

PODs based on antiandrogenic effects on the developing male reproductive system were selected for
each phthalate based on both in utero exposure and prepubertal/pubertal exposure studies. PODs based
on in utero exposure were used to characterize risk for pregnant women/women of childbearing age and
infants, while PODs based on prepubertal/pubertal exposure were used to characterize risk for children.
When phthalate-specific PODs could not be derived, Health Canada used read-across from structurally
similar phthalates to fill data gaps. PODs for high-priority and manufacturer-requested phthalates
assessed by Health Canada are shown in TableApx A-3, while PODs for other phthalates not being
assessed under TSCA. are summarized in Table F-5 of ECCC/HC- (2020).

To characterize cumulative risk, Health Canada used the HI approach. His were calculated for pregnant
women/women of childbearing age, infants, and children based on 95th percentile daily intake values
estimated using human urinary biomonitoring data and occurrence data from environmental media and
food. For pregnant women/women of childbearing age His were 0.34 (environmental occurrence) and
0.49 (biomonitoring), with HQs for DEHP (36 to 61 percent) and DINP (34-55 percent) being the
largest contributors to the His. For infants, His were 0.83 (environmental occurrence) and 0.37
(biomonitoring), with HQs for DEHP (68 to 69 percent), DINP (14 to 25 percent), and DBP (3.6 to 14
percent) being the largest contributors. Finally, for children, His were 0.60 (environmental occurrence)
and 0.54 (biomonitoring), with HQs for DEHP (67 to 88 percent), DBP (9.1 to 29 percent), and DINP
(1.6 to 2.8 percent) being the largest contributors. Based on these results, Health Canada concluded that
phthalates do not currently pose a cumulative risk to the general population in Canada.

A.3 Danish EPA

In 2011, the Danish EPA submitted a proposal for restrictions on four phthalates {i.e., DEHP, BBP,
DBP, DIBP) under Annex XV of REACH (Registration, Evaluation, Authorisation and Restriction of
Chemicals) (EC	). At the time of the proposal, all four of the assessed phthalates had already

been classified under REACH as Category IB reproductive toxicants (presumed human reproductive
toxicant) with adverse effects on male sexual differentiation during the developmental process {i.e.,
anti androgenic effects). To support the proposal for restrictions, combined exposure from DEHP, BBP,
DBP, and DIBP from "articles intended for use indoors and articles that may come into direct contact
with the skin or mucous membranes" were assessed for cumulative risk to human health.

When assessing the four phthalates, Danish EPA relied upon assumptions of dose addition and similar
MO A. Selected PODs were based on adverse effects on the developing male reproductive system that
were associated with an anti androgenic MOA. PODs selected for use in the CRA are shown in
Table Apx A-3. For the exposure assessment, combined exposure to DEHP, BBP, DBP and DIBP was
estimated for three groups, including 2-year old children, 6- to 7-year old children, and adults. Danish
EPA considered a number of exposures routes, including exposure to phthalate containing articles {e.g.,
erasers, sandals, sex toys), indoor dust, indoor air, and food. Cumulative exposure to phthalates was also
assessed using human urinary biomonitoring data of phthalate metabolites, which was converted into a
daily intake value for each parent phthalate using reverse dosimetry (Koch et at.. 2007). Low median
{i.e., the lowest calculated median value), high median {i.e., highest calculated median value), and
realistic worst-case scenario {i.e., 95th percentile value) exposure estimates were derived and used to
characterize cumulative risk.

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To assess risk, the risk characterization ratio (RCR) approach was used. The RCR approach is analogous
to the HI approach, in which RCRs are calculated for each chemical in the mixture of interest {i.e., RCR
= exposure ^ derived no effect level (DNEL))10 and then summed to calculate a cumulative RCR. If the
cumulative RCR exceeds 1.0, then risk is considered not to be controlled for the chemicals being
assessed. RCR values were generally >1.0 for 2-year old and 6- to 7-year old children in both the high-
median and 95th percentile exposures groups based on both biomonitoring data and exposure data for
combined articles, food, and indoor dust and air, while adult RCR values exceeded 1.0 only in the 95th
percentile exposure groups based on biomonitoring data and combined exposure to articles, food, and
the indoor environment. Based on these results, Danish EPA concluded that "for a large part of the
population the risk is not sufficiently controlled and the exposure to DEHP, DBP, BBP, and DIBP
should be reduced."

A.4 Australia NICNAS

Australia NICNAS has issued Priority Existing Chemical (PEC) Assessment Reports for DINP
(NICNAS. 20121 DBP (NICNAS. 20131 di(methoxyethyl) phthalate (DMEP) (NICNAS. 20 Hal
dimethyl phthalate (DMP) (NICNAS. 2014bl and BBP (NICNAS. 201 Sal As part of each PEC
assessment NICNAS assessed cumulative risk for a limited number of phthalates, populations, and
exposure scenarios. Table Apx A-2 provides a summary of the phthalates, exposure scenarios, and
critical health effects assessed in each PEC report.

CRAs conducted by NICNAS relied upon assumptions of dose addition, no toxicologic interactions, and
a similar MOA for each health outcome considered. Systemic effects {i.e., enlarged liver and/or kidney)
were assessed as part of the CRA for DINP (NICNAS. ^ ), but not other phthalates. Fertility-related
effects {i.e., reduced testes weight and/or testosterone) and developmental effects {i.e., reduced pup
weight) were assessed as part of the CRAs reported in all five phthalate PEC Assessment Reports. PODs
selected by NICNAS are shown in Table_Apx A-3. As can be seen from Table_Apx A-2, a limited
number of phthalates were included in each CRA and exposure assessments focused on exposure of a
single population group, 6-month old infants, due to use of phthalate containing plasticizers in toys and
child-care articles and use of phthalate containing lotions and other cosmetics. To characterize risk from
exposure to multiple phthalates, a cumulative margin of exposure approach was used. Cumulative
MOEs were compared to a benchmark MOE of 100. Cumulative MOEs for all assessed exposures
scenarios were >100, indicating an adequate margin of safety.

10 DNELs are analogous to oral reference doses or inhalation reference concentrations calculated by the EPA IRIS program,
i.e., DNELs and RfDs/RfCs are calculated by dividing a POD by a set of uncertainty factors.

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Table Apx A-2. Summary of Australia NICNAS Cumulative Phthalate Assessments



PEC No. 35: DINP

(NICNAS. 2012)

PEC No. 36: DBP

(NICNAS. 2013)

PEC No. 37: DMP

(NICNAS. 2014b)

PEC No. 38: DMEP

(NICNAS. 2014a)

PEC No. 40: BBP

(NICNAS. 2015a)

Phthalates
Included

- DINP, DEHP, DEP

- DBP, DINP, DEHP,
DEP

- DMP, DBP, DINP,
DEHP, DEP

- DMEP, DINP, DEHP,
DEP

- BBP, DINP, DEP

Critical Health
Effect(s)

-	Systemic toxicity

-	Developmental

-	Fertility-related

-	Developmental

-	Fertility-related

-	Developmental

-	Fertility-related

-	Developmental

-	Fertility-related

-	Developmental

-	Fertility-related

Evaluated
Population(s)

- Infants (6-months
old)

- Infants (6-months old)

- Infants (6-months
old)

- Infants (6-months old)

- Infants (6-months old)

Assessed
Exposure
Scenarios

-	Exposure to DINP
in toys and child-
care articles + DEP
in cosmetics

-	Exposure to a
mixed plasticiser
(42% DINP, 1%
DEHP) in toys and
child-care articles

-	Exposure to a
mixed plasticizer
(DINP/DEHP) in
toys and child-care
articles + DEP in
cosmetics

-	Exposure to a mixed
plasticizer (0.5% DBP,
41.5% DINP, 1%
DEHP) in toys and
childcare articles

-	Exposure to a mixed
plasticiser

(DBP/DINP/DEHP) in
toys and childcare
articles + DEP in
lotions for children

-	Exposure to a mixed
plasticiser (42.5%
DINP, 0.5% DMP) in
toys + 0.5% DEP or
0.5% DMP in
cosmetics

-	Exposure to a mixed
plasticiser (41.5%
DINP, 0.5% DMP,
1% DEHP) in toys +
0.5% DEP or 0.5%
DMP in cosmetics

-	Exposure to a mixed
plasticiser (41.5%
DINP, 0.5% DBP,
1% DEHP) in toys +
0.5% DMP in
cosmetics

-	Exposure to a mixed
plasticiser (42.5%

DINP, 0.5% DMEP) in
toys + 0.5% DEP or
0.5% DMP in cosmetics

-	Exposure to a mixed
plasticiser (41.5%

DINP, 0.5% DMEP, 1%
DEHP) in toys + 0.5%
DEP or 0.5% DMP in
cosmetics

-	Exposure to a mixed
plasticizer (42.5% DINP
+ 0.5% BBP) in toys +
0.5% DEP (or 0.5%
DMP) in cosmetics

-	Exposure to a mixed
plasticizer (41.5% DINP
+ 0.5% BBP + 1%
DEHP) in toys + 0.5%
DEP (or 0.5% DMP) in
cosmetics

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5428

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PUBLIC COMMENT DRAFT - DO NOT CITE OR QUOTE
A.5 European Food Safety Authority

In response to a request from the European Commission to update its 2005 risk assessments of DBP,
BBP, DEHP, DINP, and DIDP, the EFSA Panel on Food Contact Materials, Enzymes, and Processing
Aids (CEP Panel) established a group tolerable daily intake (TDI) value for DBP, BBP, DEHP, and
DINP (EFSA. 2019). The group-TDI was derived using an RPF approach in which DEHP served as the
index chemical. Using this approach, EFSA relied upon assumptions of dose additivity, no toxicological
interactions, and a common MOA (i.e., reduction in fetal testosterone). Effects on the developing male
reproductive system were selected as the key health outcome for deriving a group-TDI. RPFs were
derived based on PODs summarized in Table Apx A-3, which are based on a spectrum of effects
associated with phthalate syndrome (i.e., the critical effect for BBP was decreased anogenital distance,
while the critical effect for DBP was decreased spermatocyte development), instead of a single health
effect. Derived RPFs were 1.0 for DEHP (index chemical), 0.1 for BBP, 5.0 for DBP, and 0.3 for DINP,
and the group-TDI was 50 |ig/kg-d DEHP equivalent units. DIDP was not included in the group-TDI,
because EFSA concluded that DIDP does not induce reproductive effects involving a reduction in fetal
testosterone.

Dietary exposure to the five phthalates was assessed using data on the levels of occurrence of phthalates
in food from the EFSA Chemical Occurrence database and scientific literature. Dietary exposure was
estimated for a variety of populations, including infants (<12 months), toddlers (>12 to <36 months),
children (>3 to <10 years), adolescents (>10 to <18 years), adults (>18 to <65 years), elderly (>65 to
<75 years), very elderly (>75 years), pregnant women, and lactating women. To characterize risk,
estimates of combined dietary exposure to DEHP, DBP, BBP and DINP, which ranged from 0.9 to 7.2
and 1.6 tol 1.7 |ig/kg/day for mean and high-end consumers of all age groups, were compared to the
group TDI. In the worst-case scenario, dietary exposure contributed up to 23 percent of the group-TDI.

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5430

5431

5432

A.6 PODs Used in Previous Phthalate CRAs

Table Apx A-3. Summary of POPs for High-Priority and Manufacturer-Requested Phthalates Considered in Previous CRAs

Country
(Reference)

Hazard
Type

Point of Departure Used in Cumulative Risk Assessments"''

BBP

DBP

DEHP

DIBP

DCHP

DINP

Australia

(2015a,
2014a. b.
2013. 2012)

Systemic





28.9 mg/kg/d
(NOAEL, t
kidney weight)

(Corning
Hazleton Inc..





88 mg/kg/d
(NOAEL, t liver &
kidney weight)
(Lington et aL.

1997)

1996)



Reproductive

10 mg/kg/d2
(NOAEL, |
testosterone in
fetal testes)
(Lehmann et aL
2004)

10 mg/kg/d
(NOAEL, |
testosterone in fetal
testes) (Lehmann et
aL. 2004)

4.8 mg/kg/d

(NOAEL, I testes

weight,

seminiferous

tubule atrophy in

F1& F2)

(Therlmmune

Research

Corporation.

2004)





50 mg/kg/d
(NOAEL, |
testosterone in fetal
testes) (Boberg et
aL. 2011; Hannas et
aL. 2011)

Developmental

50 mg/kg/d
(NOAEL, | birth
weight in both
sexes) (Aso et
aL. 2005; Tvl et
aL. 2004; Naeao
et al.. 2000)

50 mg/kg/d
(NOAEL, | pup
weight) (Zhang et al.

2004)

46 mg/kg/d
(NOAEL, | pup
weight)
(Therlmmune

Research
Co roo ration.
2004)





50 mg/kg/d
(NOAEL, | pup
weight) (Waterman

et al, 2000)

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Country

Hazard

Point of Departure Used in Cumulative Risk Assessments"6

(Reference)

Type

BBP

DBP

DEHP

DIBP

DCHP

DINP

Canada

(ECCC/HC.

2020)

Antiandrogenic

(in utero
exposure)

50 mg/kg/d
(NOAEL, |
AGD at birth in
F2 males) (Aso
etal..2005:Tvl
et al.. 2004;
Nasao et al..
2000)

10 mg/kg/d
(NOAEL, I testicular
testosterone, fertility
effects, altered
spermatocyte
development, J,
tubular & interstitial
cell populations,
altered seminiferous
tubule morphometry)
(Ahmad et al.. 2014;
Boekelheide et al..
2009; Lehmann et al..
2004)

4.8 mg/kg/d
(NOAEL, Small
and/or aplastic
epididymis, TP,
other RPS effects
inFl &F2)
(Therlmmune
Research
Coroo ration.
2004)

125 mg/kg/d
(NOAEL, | AGD,
|NR, effects on
fertility, J,
testosterone in
fetal testes) (Furr
etal.. 2014;
Saillenfait et al..
2008)

10 mg/kg/d
(LOAEL, | AGD,
TP, | resorptions)
(Lietal.. 2016)

10 mg/kg/d
(LOEL, MNGs,
Leydig cell
aggregation) (Li et
al.. 2015a)



Antiandrogenic

(pre-pubertal

exposure)

500 mg/kg/d
(LOEL, I sperm
count (30%), I
sperm motility,
|BW gain, |
relative liver
weieht) (Kwack
et al.. 2009)

10-50 mg/kg/d
(LOEL, delayed
spermatogenesis, J,
AGD) (Moodv et al..
2013; Xiao-Fens et
al.. 2009)

10 mg/kg/d
(NOAEL, |
absolute &
relative testis
weight)
(Dostal et al..
1988)

300 mg/kg/d
(NOAEL, TP)
(Zhu et al.. 2010)

18 mg/kg/d
(NOAEL, |
spermatid head
counts, testicular
atrophy, J, BW
gain, I food
consumption in
F1 males)
(Hoshino et al..
2005)

500 mg/kg/d
(LOEL, I absolute
seminal vesicle and
LABC wt) (Lee and
Koo. 2007)

Denmark

(ECHA.

2011)

Antiandrogenic

50 mg/kg/d
(NOAEL, |
AGD inFl & F2
buds) (Tvl et al..
2004)

2 mg/kg/d
(LOAEL, |
spermatocyte
development on PND
21) (Lee etal.. 2004)

4.8 mg/kg/d
(NOAEL, I testes
weight, testicular
atrophy inFl &
F2 males)
(Therlmmune
Research
Coroo ration.
2004)

125 mg/kg/d
(NOAEL, | AGD,
tNR) (Saillenfait
et al.. 2008)





EFSA

(EFSA. 2019)

Antiandrogenic

50 mg/kg/d
(NOAEL, |
AGD inFl &F2
buds) (Tvl et al..
2004)

2 mg/kg/d
(LOAEL, |
spermatocyte
development on PND
21) (Leeetal.. 2004)

4.8 mg/kg/d
(NOAEL, I testes
weight, testicular
atrophy inFl &
F2 males)
(Therlmmune
Research





50 mg/kg/d
(NOEL, Transient J,
in fetal testosterone,
MNGs) (Clewell et
al.. 2013a)

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Country
(Reference)

Hazard
Type

Point of Departure Used in Cumulative Risk Assessments"6

BBP

DBP

DEHP

DIBP

DCHP

DINP









Coroo ration.
2004)







United States
(U.S. CPSC.
2014)

Case 1

Antiandrogenic
PODs from
(KortcnkaniD
and Faust. 2010)

66 mg/kg/d
(BMDL, | fetal
testosterone
synthesis, as
reported by
(NRC. 2008))

20 mg/kg/d
(BMDL, | fetal
testosterone synthesis,
as rcDortcd bv (NRC.
2008))

3 mg/kg/d
(NOAEL, |NR)
(Christiansen et
al.. 2009)

40 mg/kg/d
(BMDL, | fetal
testosterone
synthesis, as
reported by
(NRC. 2008))



750 mg/kg/d
(LOAEL, | fetal
testosterone
svnthesis) (Borch et
al.. 2004; Grav et
al.. 2000)

Case 2

Antiandrogenic
PODs from
(Hannas et al..
2011)

5 mg/kg/d

(NOAEL,

testosterone

modulated

effects)

5 mg/kg/d

(NOAEL, testosterone
modulated effects)

5 mg/kg/d
(NOAEL,
testosterone
modulated effects)

5 mg/kg/d
(NOAEL,
testosterone
modulated effects)



11.5 mg/kg/d
(NOAEL,
testosterone
modulated effects)

Case 3

Antiandrogenic
PODs from de
novo CPSC
review

50 mg/kg/d
(NOAEL, |NR,
jAGD) (Tvl et
al.. 2004)

50 mg/kg/d
(NOAEL, |NR,
J. AGD) (Zhane et al..
2004; Mvlchreest et
al.. 2000)

5 mg/kg/d
(NOAEL,
j spermatocytes &
spermatids,
reproductive tract
malformations,
delayed vaginal
opening)

(Blvstone et al..
2010; Andrade et
al.. 2006a: Grande
et al.. 2006)

125 mg/kg/d
(NOAEL, jAGD)
(Saillenfait et al..
2008)



50 mg/kg/d
(NOAEL, |NR)
(Bobers et al.. 2011)

" Some CRAs included phthalates not currently undergoing risk evaluation at EPA. PODs for these phthalates are not shown.

h NICNAS concluded that fetal testosterone changes are not well characterized for BBP in available studies but considered BBP to be equivalent to DBP in reducing
fetal testosterone. Therefore, NICNAS used the DBP NOAEL for reduced fetal testosterone for BBP.

AGD = anogenital distance; BW = body weight; LOAEL = lowest-observed-adverse-effect level; MNG = multinucleated gonocytes; NOAEL = no-observed-adverse-
effect level; NR = nipple retention; PND = postnatal day; RPS = rat phthalate syndrome; TP = testicular pathology

5433

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PUBLIC COMMENT DRAFT - DO NOT CITE OR QUOTE

Appendix B Additional Toxicity Information

B.l Dose-Response Data for Effects on Fetal Testicular Gene Expression
and Testosterone Production

Dose Response studies I Positis v Fblh.il.itcs





TPROD

NrObl

S«ar

Cypllal

Cvpllb2

3

s

Cy*17»l|

Lbcg,

Sf jrbl

l*U3

M..-|

Cvpllbl

RbotlO

W«7a

UkM

DF.IIP

Positive

100

(2-3

0.86

0.55

068

0.50

0.75

0.59

0.36

0.57

0.65

0.67

0.23

0.80

0.54

1.00

HARLAN SD



300

26.5

0.69

0.19

029

0.44

0J5

0.23

032

0.22

0.23

0.48

0.07

0.83

0.63

0.43





*00

11.1

0 64

0 16

019

0.71

029

0.07

0.13

0 15

0.17

034

0.02

088

0.34

0.34





900

S.9

0.53

019

0.19

0.80

0.32

0.07

0.14

0.16

0.17

0.65

0.03

0.66

o.r

0.28

DBP

PoiHhc

1

129

1.01

1.01

or

1.06

1.05

0.31

1.39

1.13

0.32

0.98

0.39

1.01

0.75

0.32

HARLAN SD



10

99.1

1.03

0.97

1.03

0.86

1.04

1.05

137

1.00

0.97

0.S3

0.35

0.93

080

l.U





33

92.2

1.16

1.56

1-3S

0.91

1.16

1.06

1.34

1.41

0.36

1.23

1.14

1.08

1.36

1.05





50

35.5

1.20

106

124

160

0.95

1 00

1.39

1.27

033

1.07

1.35

133

1.61

093





100

t7J

0.95

060

0.89

0.7|

0.76

0.79

130

0 70

0.77

0.72

049

030

0.76

0.79





300

23.2

0.32

0.41

058

0.43

0.69

031

1.05

049

0.40

0.36

0.19

0.91

1.36

0.53





?50

li.7

0.54

0.23

0.21

0.80

0.40

0.13

0.23

0 13

0.24

0.71

0.04

039

0.63

049

BBP

Positive

11

10S3

1.10

0.73

0.78

0.55

0.99

1.11

1.19

0.55

1.17

0.59

0.44

1.09

0.81

2.01

HARLAN SD



33

89.0

0.73

0.70

0.55

032

0.68

OSS

0.68

0.64

063

0.53

0J1

0.66

0.44

0.33





100



0.S1

037

0.61

0.65

0.44

0.67

0.55

0.39

0.59

0.57

0.26

058

0.67

0.73





300

33.9

0.91

0J2

0.46

0.41

0.43

0J5

0.57

0.33

043

0.61

0.10

0.92

1.01

033





600

24.6

0.64

0.22

0.27

0.42

032

0.16

0.20

0.17

0.21

033

0.05

0.59

0.89

037





900

15.4

0.55

0.20

0.22

0.15

0J3

0.11

0.19

0.15

023

0.72

0.02

0.61

0.72

0.36

DiBP

Positive

100

K>

OSS

0.89

0.99

1.34

0.93

0.99

1.30

0.93

103

0.91

1.07

0.90

0.76

0-87

HARLAN SD



200

n.i

0.65

0.63

0.71

061

0.53

0.61

063

048

062

0.75

0.41

0.93

0.5

055





300

J4.2

1.13

0.33

0.42

104



0-39

062



0.4-

069

0 14

0.-9

Ml







soo

O.I

0.95

040

0.47

045

0.53

043

066

0.33

0 45

063

0.12

1.12

086

• ?4





M0

IU

0.31

0.30

0.2S

0.54

04,

0J1

050

0.26

0.31

0.62

0 09

0.33

O.V

0M





rw

16

OSS

0.45

053

0.^6

0.53

051

066

048

0 50

079

041

0.34

1 00

0.69





900

U

0.34

0.26

0 24

1.14

0.40

0.13

0J3

0.24

0.30

0.79

008

036

048



DINP

Positive

500

703

0.76

0.54

0.71

066

0.60

0.64

0.90

0.52

0.73

0.31

031

0.94

1.02



HARLAN SD



750

63.1

0J2

0.46

0.71

0.76

0.63

0.56

0.35

0.49

0.79

0.32

0.26

0.39

0.93

0."6





1000

43.1

0.66

033

0.54

054

0.46

0.40

0.64

034

056

0.63

0.15

0.^

0.73

0.63





1500

31.0

0.65

0.27

039

0.53

0J9

030

0.62

0.25

0.51

0.64

0.06

0."*6

0.89

0.63

DCHP

Positive

33

74.6

0.S6

0.87

1.00

1.33

0.79

0.63

0.20

0.69

OSS

0.91

0.97

1.02

1.14

1.06

HARLAN SD



100

34-9

0.61

or

0.39

208

0.36

0.24

0.16

0.32

0.51

0.85

0 26

1.02

084

0.41





300

26.5

0.63

0.22

0.27

2.57

0.24

0.12

0.09

0.18

0.36

0.64

0.07

0.35

0.97

0.34





600

20.7

0.60

0.21

0.25

1.40

0.26

0.U

0.16

0.13

0.28

0.64

0.06

1.16

1.14

042





900

24.2

0.4S

0.16

0.18

1.06

0.25

0.11

0.13

0.13

0.22

033

0.03

0.96

1.09

0.34

DIDP

Negative

WO

120

1.01

tu

1.24

LSI

1.14

1.26

LSI

1.24

L2.

1.24

L2J

UN

1.4!

1.0)

CRSD



no

114

1.10

1.11

1.11

0*

0.9/

1.11

U!

1.02

L2

l.OS

Of

on

1.04

1.2S





1000

lOS

tu

LIS

1.17

2.09

104

11«

1.6

1.1

1-0C

1.24

l.O!

1.1

i.m

1 IS





JSOO

101

1 08

1.21

1.10

Lit

10*

1.14

lJi

l.U

l-2(

1.16

1.01

1.01

1.4J

108

FigureApx B-l. Dose-Response Data from Gray et al. (2021).

Figure adapted from Gray et al. (2021).

Doses are in units of mg/kg/day. Ex vivo fetal testicular testosterone production (TPROD) presented as percent
control. mRNA values are presented as fold change versus control. Values highlighted in yellow are statistically
significantly different from controls.

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5445	B.2 DEHP Study Summaries

5446

Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])

(Grav et al.. 2021)

Harlan SD rat; oral/gavage; GD 14-
18; 0, 100, 300, 600, 900 mg/kg/d;
GD 18

-	1 ex vivo fetal testes testosterone production (100)

-	i fetal testicular expression of Insl3 (300) & steroidogenic
genes (e.g., StAR (300), Cypllal (300), Cypllb2 (300),
Cvpl 7al (300), Dhcr7 (300), Cvpllbl (100), Hsd3b (300),
Scarbl (300))

CRSD rat; oral/gavage; GD 14-18;
0, 100, 300, 600, 900 mg/kg/d; GD
18

-	1 ex vivo fetal testes testosterone production (300)

-	i fetal testicular expression of Insl3 (300) & steroidogenic
genes (e.g., StAR (300), Cypllal (600), Cypllb2 (600),
Cvpl 7a1 (600), Dhcr7 (600), Cvpllbl (300), Hsd3b (600),
Scarbl (600))

(Hannas et al..
2011)

SD rats; gavage; GD 14-18; 0, 100,
300, 500, 625, 750, 875 mg/kg/d;
GD 18

-	1 ex vivo fetal testes testosterone production (300)

-	i fetal testicular expression of Ins!3 (625), StAR (500),
Cvpl la (500)

Wistar rats; gavage; GD 14-18; 0,
100, 300, 500, 625, 750, 875
mg/kg/d; GD 18

-	1 ex vivo fetal testes testosterone production (300)

-	1 fetal testicular expression of Iris/3 (500), StAR (500),
Cvpl la (500)

(Furret al.. 2014)

SD rats; oral/gavage; GD 14-18; 0,
100, 300, 600, 900 mg/kg/d; GD 18

- 1 ex vivo fetal testes testosterone production (100)

(Howdeshell et
al.. 2008)

SD rats; oral/gavage; GD 8-18; 0,
100, 300, 600, 900 mg/kg/d; GD 18

- 1 ex vivo fetal testes testosterone production (300)

(Wilson et al..
2004)

SD rats; oral/gavage; GD 14-18; 0,
750 mg/kg/day; GD 18

-	| Testicular In si3 mRNA (750/GD 18)

-	i Testicular testosterone production (750/GD 18)

(Saillenfait et al..
2009a)

SD rats; oral/gavage; GD 12-21, 0,
625 mg/kg/d; PND 70-84

- t Reproductive malformations (small penis, cleft prepuce,
hypospadias, cleft phallus with exposed os penis, vaginal
pouch, undescended testes) (625/PND 70-84)

SD rats; oral/gavage; GD 12-21, 0,
500 mg/kg/d; PND 1-120

-	| AGD (500/PND 1)

-	NR (500/PND 12-14, 70-78, 111-120)

-	Reproductive tract malformations (cleft prepuce,
hypospadias, cleft phallus with exposed os penis,
undescended testes (unilateral and bilateral), underdeveloped
testes, malformed epididymis, absent SV or prostate)
(500/PND 70-78, 111-120)

Unaffected outcomes

-	PPS

(Saillenfait et al..
2013)

SD rats; oral/gavage; GD 12-19; 0,
50, 625 mg/kg/d; GD 19

-	i fetal testicular testosterone (50)

-	i fetal testicular expression of steroidogenic genes (SR-B1
(50), StAR (50), P450scc (625), P450cl 7 (625), 3/3-HSD
(625))

(Soade et al..
2018)

SD rats; oral/gavage; GD 17-21; 0,
750 mg/kg/d; GD 21

-	1 ex vivo fetal testicular testosterone production (750)

-	t Incidence of MNGs (750)

(Borch et al..
2004)

Wistar rats; oral/gavage; GD 7-21;
0, 300, 750 mg/kg/d; GD 21, PNDs
3-190

-	1 ex vivo fetal testes testosterone production and testosterone
content (300/GD 21)

-	i plasma testosterone (750/GD 21)

-	t luteinizing hormone (750/GD 21)

-	| AGD (750/PND 3)

-	NR (750/PND 13)

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Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])





Unaffected outcomes

- Serum testosterone (PND 22, 190) or testicular testosterone
content (PND 190)

(Parks et al..
2000)

SD rats; oral/gavage; GD 14-PND
2; 0, 750 mg/kg/day; GD 17, 18, 20,
PND2

-	i Testicular testosterone production (750/GD 17, 18, 20,
PND 2)

-	1 Testicular testosterone (750/GD 17, 18, PND 2)

-	1 absolute testis weight (750/GD 20, PND 2)

-	| AGD (750/PND 2)

-	Testicular pathology (Leydig cell hyperplasia, Leydig cell
aggregation, f # of gonocytes in seminiferous cord, MNGs)
(750/PND 2)

(Borch et al..
2006b)

Wistar rats; oral/gavage; GD 7-21;
0, 10, 30, 100, 300, mg/kg/d; GD 21

-	1 ex vivo fetal testes testosterone production and testosterone
content (300/GD 21)

-	i steroidogenic gene (SR-B1 (300), StAR (100), P450scc
(300)) and InsB (300) expression

-	Testicular pathology (t gonocytes (100), MNGs (100),
vacuolization of Sertoli cells (300), Leydig cell clustering
(300)

Unaffected outcomes

-	Plasma testosterone (GD 21)

(Cultv et al..
2008)

SD rats; oral/gavage; GD 14-PND
0; 0, 234, 469, 700, 750, 938, 1250
mg/kg/d; PND 21 or 60

-	| AGD (1250/PND 60)

-	Cryptorchidism (938/PND 60)

-	t Leydig cell volume (234/PND 60)

-	i serum testosterone (234)

Unaffected outcomes

-	Testis weight (PND 21, 60); germ cell volume (PND 60)

SD rats; oral/gavage; GD 14-PND
0; 0, 117, 234, 469, 938 mg/kg/d;
GD 20 or PND 3

-	1 ex vivo fetal testicular testosterone production (117/GD 20)

-	1 ex vivo fetal testicular testosterone production (938/PND 3)

SD rats; oral/gavage; GD 14-PND
0; 0, 234, 469, 938 mg/kg/d; GD
19, PND 3, 21, 60

-	1 InsB and steroidogenic (Cvpllal, Cvpl7al) gene
expression (469 /GD 20)

-	| mRNA expression of Cvpllal, Cypl7al, Ins3 reported at
PNDs 3, 21, 60

Unaffected outcomes

-	Star mRNA (GD 19, PNDs 3, 21, 60)

(Vo et al.. 2009)

SD rats; oral/gavage; GD 11-21; 0,
10, 100, 500 mg/kg/d; GD 21, PND
1,63

-	i serum testosterone (500/GD 21)

-	NR (500/PND 13)

-	Sperm parameters (concentration (500/PND 63), viability
(500/PND 63), motility (10/PND 63)

-	Hypospadias (500/PND 63)

-	Cryptorchidism (500/PND 63)

Unaffected outcomes

-	Serum testosterone (PND 63); AGD (PND 63); Testis,
epididymis, prostate weight (PND 63)

(Lin et al.. 2008)

Long-Evans rats; oral/gavage; GDs
2-20; 0, 10, 100, 750;GD21

-	| AGD (750)

-	i testicular testosterone (750)

-	1 mRNA expression of steroidogenic (Scarbl (750), St. IN
(750), Cvplla (100), Cypl9 (100)) genes and InsB (750)
mRNA '

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Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])





-	i absolute testes weight (100)

-	Ley dig cell aggregation/Increased # of Ley dig cells per
cluster (10)

(Martino-Andrade
et al.. 2008)

Wistar rats; oral/gavage; GDs 13-
21; 0, 150 mg/kg/d; GD 21, PND
13, PND 90

-	i absolute SV weight (150/PND 90)

Unaffected outcomes

-	Testicular testosterone, seminiferous cord diameter, incidence
of MNGs, AGD (GD 21)

-	NR (PND 13)

-	PPS

-	Testis, epididymis, prostate, LABC weight (PND 90)

-	# spermatids/testis (PND 90)

(Jarfelt et al..
2005)

Wistar rats; oral/gavage; GD 7-
PND 17; 0, 300, 750; PND 3, 13,
33, 190

-	| AGD (300/PND 3)

-	NR (300/PND 13)

-	Sperm parameters (# sperm per cauda epididymis (severely
reduced in 3 males at 300 mg/kg); sperm motility (severely
reduced in 3 males at 300 and 2 males at 750 mg/kg) (neither
effect was statistically significant)

-	i absolute weight of paired testes (750/PND 22); ventral
prostate (300/PND 190); LABC (300/PND 190)

-	Reproductive malformations (small testis or lack of one testis
(300/PND 22, 190), small/malformed epididymis (300/PND
22, 190), malformed SVs (300/PND 22, 190), Cryptorchid
testis (750/PND 22, 190), hypospadias (300/PND 22)

-	Testicular pathology (300/PND 22, 190)

Unaffected outcomes

-	Absolute epididymis (PND 22, 190), prostate (PND 22), SV
(PND 22, 190), paired testis (PND 190), LABC (PND 22)

(Gray et al.. 2000)

SD rats; oral/gavage; GD 14-PND
3; 0, 750 mg/kg/d; PND 2-mature
adults (3-7 months of age)

-	| AGD (750/PND 2)

-	NR (750/PND 13)

-	Permanent nipples (750/3-7 months)

-	i absolute testes, LABC, SV, ventral prostate, glans penis,
epididymis, cauda epididymis, caput-corpus epididymis
weight (750/3-7 months)

-	Incomplete PPS due to genital malformations (750)

-	Reproductive tract malformations (cleft phallus, hypospadias,
vaginal pouch, SV and epididymal agenesis, fluid filled testis,
small testis, testis absent, abnormal gubernaculum) (750/3-7
months)

-	Undescended testes (750/3-7 months)

-	Testicular pathology (e.g., MNGs)

Unaffected outcomes

-	Mean age at PPS; serum testosterone (3-7 months)

(Moore et al..
2001)

SD rats; oral/gavage; GD 9-PND
21; 0, 375, 750, 1,500 mg/kg/d;
PND 1-112

-	| AGD (750/PND 1)

-	NR (375/PND 14) (% litters with males with NR)

-	Incomplete PPS (375)

-	t Litters with undescended testes (750/PND 21)

-	i absolute testes (750/PND 21, 63), epididymis (750/PND 21,
63. 105), glans penis (750/PND 21, 63, 105)

-	i epididymal sperm number (750/PND 63)

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Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])





-	Anterior prostate agenesis (750)

-	i masculine sexual behavior (J, incidence of mounting)
(1500/PND 77)

(Howdeshell et
al.. 2007)

SD rats; oral/gavage; GD 14-18; 0,
500 mg/kg/d; PND 3, PND 14,
adult (7-11 months of age)

-	| AGD (500/PND 3)

-	NR (500/PND 14, adult)

-	1 absolute LABC weight (500/adult)

-	Low incidence of hypospadias, testes and epididymal
malformations, SV, vas deferens and gubernacular agenesis
reported (500/adult; not statistically significant)

Unaffected outcomes

-	Absolute glans penis, ventral prostate, SV, testes, epididymis
weight (500/adult)

(Grav et al.. 2009)

SD rats; oral/gavage; GD 8-PND
17; 0, 11, 33, 100, 300 mg/kg/d;
PND 2, PND 13, adult (7 months of
age)

-	| AGD (300/PND 2)

-	NR (300/PND 13, adult)

-	i Absolute glans penis, ventral prostate, LABC, Cowper's
gland, epididymis weight (300/Adult); SV (100/adult)

-	t incidence of testicular pathologies and malformations such
as hypospadias (11/adult)

Unaffected outcomes

-	Age at PPS; Serum testosterone (adult)

SD rats; oral/gavage; GD 8-PND
64; 0, 11, 33, 100, 300 mg/kg/d;
PND 64

-	i Absolute ventral prostate, SV, LABC, Coper's gland,
epididymis weight (300/PND 64)

-	i Epididymal sperm count (300/PND 64)

-	Delayed PPS (300)

-	t incidence of testicular pathologies and malformations such
as hypospadias (11/PND 64)

Unaffected outcomes

-	Serum testosterone (PND 64)

(Li et al.. 2013)

SD rat; oral/gavage; GD 12-19; 0,
500, 750, 1000 mg/kg/d; PND 1,
30, 60

-	| AGD (500/PND 1)

-	1 penile length (750/PND 30)

-	Hypospadias (500/PND 1, 60)

(Christiansen et
al.. 2010)

Study (S) 1

Wistar rat; oral/gavage; GD 7-PND
16; 0, 10, 30, 100, 300, 600, 900
mg/kg/d; PND 1, 12, 16

-	| AGD (10/PND 1; SI) (100/PND 1; S2) (10/PND 1;
combined)

-	NR (10/PND 12; SI) (none/PND 12; S2) (10/PND 12;
combined)

-	Mild external genital dysgenesis (100/PND 16; SI) (3/PND
16; S2) (3/PND 16; combined)

-	i absolute testis weight (600/PND 16; SI & combined);
ventral prostate (30/PND 16; SI) (10/PND 16; combined);
LABC (10/PND 16; SI & S2 & combined)

-	1 Diameter of seminiferous tubules (300/PND 16; SI &
combined)

-	Testicular pathology (J, germ cells, focal Leydig cell
hyperplasia) (300, PND 16; SI)

Unaffected outcomes

-	Hypospadias (S1 or S2)

Study (S) 2

Wistar rat; oral/gavage; GD 7-PND
16; 0, 3, 10, 30, 100 mg/kg/d; PND
1, 12, 16

(Andrade et al..
2006b)

Wistar rats; oral/gavage; GD 6-
PND 21; 0.015, 0.045, 0.135, 0.405,

-	t absolute testis weight (5/PND 22)

-	| AGD (405/PND 22)

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Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])



1.215, 5, 15, 45, 135, 405 mg/kg/d;
PND 1, 13, 22, 33

-	NR (405/PND 13)

-	Delayed PPS (15)

-	Testicular pathology (e.g., MNGS, acute interstitial
hemorrhage/loosening of connective tissue, J, germ cell
differentiation in seminiferous tubules) (135/PND 1, 22)

Unaffected outcomes:

-	Seminiferous tubule diameter, testes descent, epididymis
weight, testicular testosterone (PND 1), malformations of
external genitalia (e.g., hypospadias)

(Blvstone et al..
2010;

Therlmmune
Research
Co monition.
2004)

SD rats; oral/diet; 3-generation
(continuous breeding protocol); 1.5
(control), 10, 30, 100, 300, 1000,
7500, 10000 ppm (eq. 0.12, 0.78,
2.4, 7.9, 23, 77, 592, 775 mg/kg/d
(F0); 0.09, 0.48, 1.4, 4.9, 14, 48,
391, 543 mg/kg/d (Fl); 0.1, 0.47,
1.4, 4.8, 14, 46, 359 mg/kg/d (F2))

-	| AGD (7500 ppm/PND 1 (Fl, F2, F3))

-	NR (7500 ppm/PND 12-13 (F3))

-	Delayed PPS (7500 ppm/Fl, F3; 10 ppm/F2)

-	Delayed testes descent (7500 ppm/Fl, F3; 30 ppm/F2)

-	i absolute and/or relative cauda, epididymis, testis weight
(7500 ppm/adult (Fl, F2, F3))

-	Gross necropsy findings (small or aplastic testis, SV,
epididymis or cauda) (300 ppm/adult (Fl, F2))

-	Testicular pathology (seminiferous tubule atrophy, failure of
sperm release; sloughed epithelial cells, residual bodies in
epididymis) (7500 ppm/adult (Fl, F2))

-	i spenn/cauda (or mg cauda) & J, spermatid/testes (or mg
testes) (7500 ppm/adults (Fl, F2, F3))

-	i Pregnancy index (10,000 ppm/Fl (no F2 litters produced);
7500/F2)

(Pocar et al..
2012)

CD-I mice; oral/diet; GD 0.5-PND
21; 0, 0.05, 5, 500* mg/kg/d; PND
42

*Only 1 out of 10 high-dose dams
produced a litter. This dose group
was excluded from most analyses

-	i absolute testes weight at 0.05, but not 5 mg/kg/d (PND 42)

-	i absolute SV weight (0.05/PND 42)

-	i Sperm count and sperm viability (0.05/PND 42)

-	1 testicular cvp!9al mRNA (5/PND 42)

Unaffected outcomes

-	AGD (PND 42); testicular &4.R, CYP17al mRNA (PND 42)

(Liu et al.. 2008)

C57BL/6 mice; oral/gavage; el 2-
17; 0, 100, 200, 500 mg/kg/d; el9

-	| AGD (100)

-	i urethra length (200)

-	Hypospadias (100)

(Do et al.. 2012)

CD-I mice; oral/gavage; GD 9-18;
0, 0.0005, 0.001, 0.005, 0.5, 50, 500
mg/kg/d; GD 18

-	i absolute testes weight (50)

-	t serum testosterone (0.0005)
Unaffected outcomes

-	AGD, testicular testosterone

(Gaido et al..
2007)

C57B1/6J; oral/gavage; GD 15-17;
0, 1000 mg/kg/d MEHP; GD 17 (8-
hours post dosing)

Unaffected outcomes
- Testicular testosterone

C57B1/6J; oral/gavage; GD 14-16;
0, 500 mg/kg/d MEHP; GD 17 (24-
hours post dosing)

Unaffected outcomes
- Testicular testosterone

" Study design comprises species/strain, exposure route, exposure duration, doses, and timing of evaluation.
AGD = anogenital distance; e = embryonic day; GD = gestation day; LABC = levator ani/bulbocavernosus muscle; NR =
nipple retention; PPS = preputial separation; PND = postnatal day; PNW = postnatal week; PS = phthalate syndrome; SD =
Sprague-Dawley; SV = seminal vesicle

5447

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5448	B.3 BBP Study Summaries

5449

Reference

Study Design"

Reported Phthalate Syndrome-Related Effects

(LOAEL/LOEL (mg/kg/day)/timing of evaluation (if different
than listed under Study Design))

(Ema et al.. 2003)

Wistar rats; oral/gavage; GD 15-17;
0 , 167, 250, 375 mg/kg/dMBP;
GD 21

-	| AGD (250)

-	Cryptorchidism (250)

(Grav et al.. 2021)

Harlan SD rat; oral/gavage; GD 14-
18; 0, 11,33, 100,300, 600, 900
mg/kg/d; GD 18

-	1 ex vivo fetal testes testosterone production (100)

-	i fetal testicular expression of Insl3 (33) and steroidogenic
genes (StAR (100), Cvpllal (33), Cypllb2 (33), Cypl7al
(300), Dhcr7 (11), Cypllbl (11 \Hsd3b (100\Scarbl (33))



Charles River SD rat; oral/gavage;
GD 14-18; 0, 100, 300, 600, 900
mg/kg/d; GD 18

-	1 ex vivo fetal testes testosterone production (300)

-	i fetal testicular expression of Ins!3 (600) and steroidogenic
genes (StAR (600), Cvpllal (600), Cypl7al (600), Dhcr7
(900), Cypllbl (600), Hsd3b (900), Scarbl (600))

(Howdeshell et
al.. 2008)

SD rats; oral/gavage; GD 8-18; 0,
100, 300, 600, 900 mg/kg/d; GD 18

- 1 ex vivo fetal testes testosterone production (300)

(Furret al.. 2014)

SD rats; oral/gavage; GD 14-18; 0,
100, 300, 600, 900 mg/kg/d (Block
36) orO, 11,33, 100 mg/kg/d
(Block 37); GD 18

-	Block 36: I ex vivo fetal testes testosterone production (100)

-	Block 37: No effect on testosterone

(Grav et al.. 2000)

SD rats; oral/gavage; GD 14-PND
3; 0, 750 mg/kg/d; PND 2-mature
adults (3-7 months of age)

-	| AGD (750/PND 2)

-	NR (750/PND 13)

-	Permanent nipples (750/3-7 months)

-	i absolute testes, LABC, SV, ventral prostate, glans penis,
epididymis, cauda epididymis, caput-corpus epididymis
weight (750/3-7 months)

-	Incomplete PPS due to genital malformations (750)

-	Reproductive tract malformations (cleft phallus, hypospadias,
vaginal pouch, SV and epididymal agenesis, fluid filled testis,
small testis, testis absent, abnormal gubernaculum) (750/3-7
months)

-	Undescended testes (750/3-7 months)

Unaffected outcomes

-	Mean age at PPS; serum testosterone (3-7 months)

(Soade et al..
2018)

SD rats; oral/gavage; GD 17-21; 0,
750 mg/kg/d; GD 21

-	1 ex vivo fetal testicular testosterone production (750)

-	t Incidence of MNGs (750)

(Wilson et al..
2004)

SD rats; oral/gavage; GD 14-18; 0,
750 mg/kg/day; GD 18

-	| Testicular In si3 mRNA (750/GD 18)

-	i Testicular testosterone production (750/GD 18)

(Ahmad et al..
2014)

Albino rats; oral/gavage; GD 14-
21; 0, 4, 20, 100 mg/kg/d; PND 5,
25, 75

-	i 17(3-HSD activity (trend/PND 75)

-	i serum testosterone (100/PND 75)

-	i absolute epididymis & prostate weight (100/PND 75)

-	i cauda epididymal sperm count, I sperm motility, t sperm
abnormalities (100/PND 75)

Unaffected outcomes

-	AGD (PND 5, 25); testis descent; testis & SV weight (PND
75)

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Reference

Study Design"

Reported Phthalate Syndrome-Related Effects

(LOAEL/LOEL (mg/kg/day)/timing of evaluation (if different
than listed under Study Design))

(Naeao et al..
2000)

SD rats; oral/gavage; 2-generation;
0, 20, 100, 500 mg/kg/day

-	| AGD (500/F1 PND 0)

-	i serum testosterone (500/F0 & F1 adults)

-	i absolute testes & epididymis weight (500/F1 PND 22)

-	i absolute testes, epididymis, ventral prostate weight (500/F1
adults)

-	Testicular pathology (J, spermatocytes in seminiferous
tubules (500/F1 PND 22); atrophy of seminiferous tubules
(500/F1 adults); j germ cells in seminiferous tubule (500/F1
adults); edema, interstitium (500/F1 adults); decreased sperm
in epididymis, with cell debris (500/F1 adults)

-	Delayed PPS (500/F1)

Unaffected outcomes

Mating, fertility, delivery indices (F0, Fl); gestation length (F0,
Fl); absolute reproductive organ weight (testes, epididymides,
ventral prostate, SV; F0 adults); absolute SV weight (Fl
adults); testicular pathology (F0); sperm motility and
concentration (F0, Fl adults); serum testosterone (Fl PND 22);
hypospadias (Fl), cryptorchidism (Fl)

(Aso et al.. 2005)

Crj:CD(SD)IGS rats; oral/gavage;
2-generation; 0, 100, 200, 400
mg/kg/day

-	Low rate for completed PPS (400/F1)

-	i absolute epididymis (400/F0 adults; 200/F1 adults) & SV
(400/F1 adults) weight

-	t incidence of small testes (400/F1 adult), softening of testes
(100/F1 adult); t incidence of small or hypoplastic
epididymides (400/F1 adult)

-	Testicular pathology (e.g., Ley dig cell hyperplasia (400/F0 &
400/F1 adults), diffuse atrophy of testicular seminiferous
tubules (400/F1 adults); j spermatozoa in epididymides
(400/F0; 100/F1 adults), j germ cells in epididymal lumen
(100/F1 adults), bilateral or unilateral partial aplasia or
unilateral aplasia of epididymides (400/F1 adults))

-	| AGD (100/F2 pups)

Unaffected outcomes

-	Estrous cyclicity, mating index, days required for mating,
gestation length, # implantations, fertility index, delivery
index, gestation index, # of pups delivered, # of sperm in
testis and epididymis, epididymal sperm motility or
morphology (F0 and Fl parents); serum hormones (FSH, LH,
testosterone, estradiol (F0 and Fl parents); absolute testis and
ventral prostate weight (Fl adults); AGD (Fl pups)

(Tvl et al.. 2004)

CD rats; oral/diet; 2-generation; 0,
750, 3750, 11,250 ppm (eq. 0, 50,
250, 750 mg/kg/d)

-	i Mating and fertility indices (750/F1)

-	i epididymal sperm concentration & motility (750/F1 adults)

-	i absolute testes, epididymis, prostate, SV weight (750/F1
adult)

-	1 absolute testes (250/F1 weanlings (PND 21); 750/F2
weanlings (PND 21)) and epididymis weight (750/F1
weanlings (PND 21))

-	| AGD (250/F1 and F2 at PND 0)

-	NR (750/F1 and F2 at PND 11-13)

-	Delayed PPS (750/F1)

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Reference

Study Design"

Reported Phthalate Syndrome-Related Effects

(LOAEL/LOEL (mg/kg/day)/timing of evaluation (if different
than listed under Study Design))





-	Gross malformations (undescended testes) (750/F1 pups
(PND 4))

-	Gross malformations (missing epididymis (whole or part),
epididymis reduced in size, missing testes, testes reduced in
size, and undescended testis(es) (750/F1 weanlings (PND
21))

-	Gross malformations (hypospadias, missing reproductive
organ or portion(s) of organs and/or abnormal organ size
and/or shape) (750/F1 adults)

-	Gross malformations (missing SVs, missing epididymides)
(750/F2 pups (PND 4))

-	Testicular pathology (epididymal aspennia, testis dilation,
seminiferous tubule degeneration & atrophy) (750/F1 adult)

Unaffected outcomes

-	Mating, fertility, gestation, pregnancy indices (F0);
gestational and pregnancy indices (Fl); absolute testes,
epididymis, prostate, SV weight (F0); epididymal sperm
concentration and motility (F0 adults)

" Study design comprises species/strain, exposure route, exposure duration, doses, and timing of evaluation.
AGD = anogenital distance; GD = gestation day; LABC = levator ani/bulbocavernosus muscle; NR = nipple retention; PPS
= preputial separation; PND = postnatal day; PNW = postnatal week; PS = phthalate syndrome; SD = Sprague-Dawley; SV
= seminal vesicle

5450	B.4 DBP Study Summaries

5451

Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])

(Mvlchreest et al..
1998)

SD rats; oral/gavage; GD 3-PND
20; 0, 250, 500, 750 mg/kg/d;
PND 1, adults (PND 100)

-	| AGD (500/PND 1)

-	i absolute reproductive organ weight at PND 100 (testis
(500), epididymis (750), SVs (500), prostate (750))

-	t Reproductive malformations at PND 100 (e.g., hypospadias
(250), nonscrotal testes (250), epididymal
dysgenesis/agenesis (250), SV agenesis (500))

-	Testicular pathology at PND 100 (e.g., degeneration and
atrophy of seminiferous tubules (250))

(Mvlchreest et al..
1999)

SD rats; oral/gavage; GD 12-21;
0, 100, 250, 500 mg/kg/d; PND 1,
14, adults (3 months of age)

-	| AGD (250/PND 1)

-	NR (250/PND 14)

-	Delayed PPS (100)

-	i absolute reproductive organ weight in adults (testis,
epididymis, SV (500))

-	Reproductive malformations in adults (e.g., hypospadias
(500), prostate agenesis (500), epididymal
dysgenesis/agenesis (250))

-	Cryptorchidism (250/adults)

-	t Testicular pathology in adults (e.g., degeneration of
seminiferous epithelium (250), interstitial cell hyperplasia or
adenoma (500))

(Mvlchreest et al..
2000)

SD rats; oral/gavage; GDs 12-21;
0, 0.5, 5, 50, 100, 500 mg/kg/d;

-	| AGD (500/PND 1)

-	NR (100/PND 14)

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Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])



PND 1-sexual maturity (PND
100-120)

-	i reproductive organ weight (testis, epididymis, prostate,
LABC) (500/sexual maturation)

-	Reproductive tract malformations (absent or malformed
epididymis & vas deferens; hypospadias; exposed os penis)
(500/sexual maturation)

-	Testicular pathology (seminiferous tubule degeneration, focal
interstitial hyperplasia, adenoma) (500/sexual maturation)

Unaffected outcomes

-	PPS; absolute vas deferens and SV weight

(Mvlchreest et al..
2002)

SD rats; oral/gavage; GD 12-21;
0, 500 mg/kg/day; GDs 14, 16, 18,
21

-	i fetal testicular testosterone (500/GD 18, 21)

-	Testicular pathology (Leydig cell hyperplasia (500/GD 16,
18, 21), testis atrophy (500/GD 18, 21), reduced epididymal
ducts (500/GD 21), MNGs (500/GD 21))

(Lehmann et al..
2004)

SD rats; oral/gavage; GDs 12-19;
0,0.1, 1, 10, 50, 100, 500
mg/kg/day; GD 19

-	i testicular mRNA & protein expression of Insl3 (500), StAR
(50), P450scc (50), CYP17 (500), SR-B1 (50))

-	i fetal testicular testosterone (50)

(Wilson et al..
2004)

SD rats; oral/gavage; GD 14-18;
0, 750 mg/kg/day; GD 18

-	| Testicular Ins/3 mRNA (750/GD 18)

-	i Testicular testosterone production (750/GD 18)

(Soadc et al.. 2018)

SD rats; oral/gavage; GD 17-21;
0, 750 mg/kg/d; GD 21

-	1 ex vivo fetal testicular testosterone production (750)

-	t Incidence of MNGs (750)

(Howdeshell et al..
2008)

SD rats; oral/gavage; GDs 8-18; 0,
33, 50, 100, 300, 600; GD 18

- 1 ex vivo fetal testicular testosterone production (300)

(Furretal.,2014)

SD rats; oral/gavage; GDs 14-18;
0, 33, 50, 100, 300; GD 18

- 1 ex vivo fetal testicular testosterone production (100)

(Grav et al.. 2021)

Harlan SD rat; oral/gavage; GD
14-18; 0, 1, 10, 33, 50, 100, 300,
750 mg/kg/d; GD 18

-	1 ex vivo fetal testes testosterone production (100)

-	i fetal testicular expression of Insl3 (100) & steroidogenic
genes (e.g., StAR (100), Cypllal (300), Cypl7al (100),
Dhcr7 (750), Cvpllbl (100), Hsd3b (100), Scarbl (100))

(Drake et al.. 2009)

Wistar rats; oral/gavage; el3.5-
21.5; 0, 100, 500 mg/kg/day;
adults (>12 weeks of age)

-	| AGD (500/adult)

-	Cryptorchidism (500/adult)

-	Hypospadias (500/adult)

-	i penis length (500/adult)

-	i absolute testis and ventral prostate weight (500/adult)

Wistar rats; oral/gavage; el3.5-
16.5; 0, 500 mg/kg/day; adults
(el7.5)

-	i Testicular testosterone (500/el7.5)

-	1 Testicular Cypllal and Star mRNA (500/el7.5)

(Martino-Andrade
et al.. 2008)

Wistar rats; oral/gavage; GDs 13-
21; 0, 100, 500 mg/kg/d; GD 21,
PND 13, PND 90

-	i testicular testosterone (500/GD 21)

-	t Seminiferous cord diameter (500/GD 21)

-	t incidence of MNGs (500/GD 21)

-	| AGD (500/GD 21)

-	t NR (500/PND 13)

Unaffected outcomes

-	PPS

-	Testes, epididymis, prostate, SV, LABC weight (PND 90)

-	# of spermatids/testis (PND 90)

(Strove et al..
2009)

CD rats; oral/diet; GD 12-19; 0,
112, 582 mg/kg/d (received

- i Testicular testosterone (500/GD 19; 100/GD20)

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Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])



doses); GD 19 (4-h post-dosing) or
GD 20 (24-h post-dosing)

-	1 Cvpllal, Cypl7al, Scarbl, Star mRNA (100/GD 19); {
Cvpllal, Cypl 7a 1, Scarbl mRNA (500/GD 20)

-	| AGD (500/GD 19, 20)

-	Ley dig cell aggregation (100/GD 19, 20), f seminiferous cord
diameter (100/GD 19, 20), MNGs (100/GD 19, 20)

(Kuhl et al.. 2007)

SD rats; oral/gavage; GD 18; 0,
100, 500 mg/kg/d; GD 19

-	i Testicular testosterone (500)

-	i mRNA expression of St4R, SR-B1, Cvpllal, Cypl 7 (100)

(Ema et al.. 1998)

Wistar rats; oral/diet; GD 11-21;
0, 331, 555, 661 mg/kg/d; GD 21

-	| AGD (555)

-	t incidence of internal malformations, including undescended
testes (555)

(Mahood et al..
2007)

Wistar rats; oral/gavage; GD 13.5—
20.5; 0, 4, 20, 100, 500 mg/kg/d;
GD 21.5

-	i Testicular testosterone (100)

-	i absolute testes weight (500)

-	t incidence of MNGs (100)

-	Changes in Ley dig cell distribution (i.e., j # of total Ley dig
cell clusters, t occurrence of medium and large Leydig cell
clusters) (100)

Wistar rats; oral/gavage; GD 13.5—
21.5; 0 , 4, 20, 100, 500 mg/kg/d;
PND90

-	Increased incidence of infertility (i.e., male produce offspring
with untreated females) (500)

-	t incidence of cryptorchidism (500)

-	t incidence of Sertoli cell only tubules in cryptorchid testes
(100) and increased incidence of Sertoli cell only tubules in
scrotal testes (20)

-	i absolute testes weight (500)

(Barlow et al..
2004)

SD rats; oral/gavage; GD 12-21;
0, 100, 500 mg/kg/d; PND 1, 13,
90, 180,370 or 540

-	t incidence of gross lesions in testes (lesions included
atrophy, enlarged, or absent organ), vas deferens (absent
organ), SVs (small, malformed or absent lobes), prostate
(small or absent), penis (hypospadias of varying severity)
(500, PND 180, 370, 540)'

-	Testicular pathology (testicular dysgenesis, germ cell
degeneration, rete testes) (500/PND 180, 370, 540)

-	| AGD (500/PND 1, PND 180)

-	NR (100/PND 13) (500/PND 180)

(MacLeod et al..
2010)

Wistar rats; oral/gavage; el3.5-
15.5; 0, 500; el7.5

- i Testicular testosterone content (31% reduction) (500)

Wistar rats; oral/gavage; el3.5-
20.5; 0, 500; e21.5

-	i Testicular testosterone content (60% reduction) (500)

-	| AGD (500)

Wistar rats; oral/gavage; el3.5-
21.5; 0, 100, 500; PND 25

-	i absolute SV (500), ventral prostate (100), and testis (500)
weight

-	i penis length (500)

-	| AGD (500)

(Li et al.. 2009)

Wistar rats; oral/diet; GD 6-PND
28; 0,0.037, 0.111,0.333, l%(eq.
to 0, 31, 94, 291, 797 (GD 6-21);
0, 55, 165, 486, 1,484 (PND 0-
15); 0, 47, 140, 433, 1,283 ( PND
16-28) mg/kg/d); PND 1- 28

-	| AGD (291/PND 1)

-	i relative testes weight (797/PND 28)

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Reference

Study Design "

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [m^/k^/dav]/Timin
-------
PUBLIC COMMENT DRAFT - DO NOT CITE OR QUOTE

Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])



Wistar rats; oral/gavage; el3.5-
21.5; 0, 4, 20, 100, 500 mg/kg/d;
PND4

Unaffected outcomes

-	Germ cell aggregation

-	Sertoli-genn cell interactions (qualitative imaging)

(Gaido et al.. 2007)

C57B1/6J mice; oral/gavage; GD
14-16; 0, 1500 mg/kg/d DBP; GD
17 (24-hpost final dose)

Unaffected outcomes
- Testicular testosterone

C57B1/6J mice; oral/gavage; GD
14-16; 0, 1,000 mg/kg/d MBP;
GD 17 (24-hpost final dose)

Unaffected outcomes
- Testicular testosterone

C3H/HeJ mice; oral/gavage; GD
15-17; 0, 1,000 mg/kg/d MBP;
GD 17 (8-hpost final dose

Unaffected outcomes
- Testicular testosterone

C57B16 mice; oral/gavage; GD
16-18; 0, 250, 500 mg/kg/d DBP;
GD 19

-	t Seminiferous cord diameter (250/GD 19)

-	t MNGs & t nuclei/MNG (250/GD 19)

CD-I mice; oral/gavage; GD 18; 0,
500 mg/kg DBP; GD 18 (2, 4, and
8 hours post dosing)

Unaffected outcomes

- Steroidogenesis related genes (Scarbl, St. IR, Cvpllal,
Cypl7al, Dhcr7) (microarray experiment)

CD-I mice; oral/gavage; GD 14-
17; 0, 250 mg/kg/d DBP; GD 17 (2
h post final dose)

Unaffected outcomes

- mRNA expression of genes involved in cholesterol and lipid
homeostasis and steroidogenesis was not decreased
(Microarray experiment).

(Lee et al.. 2004)

CD(SD)IGS rats; oral/feed; GD
15-PND 21; 0, 20, 200, 2000,
10000 ppm (eq. to 2-3, 14-29,
148-291, 712-1,372 mg/kg/d);
PND 2, 14, 21;PNW 11,20

-	| AGD (712/PND 2)

-	NR (712/PND 14)

-	i relative testes weight (712/PND 21)

-	Testicular pathology (J, spermatocyte development (2/PND
21), Leydig cell aggregation (148/PND 21), J, ductular cross
sections of epididymal duct (148/PND 21), J, germ cell
development (148/PNW 11)

Unaffected outcomes

-	PPS; relative epididymis (PND 21, PNW 11, 20), testes
(PNW 11, 20), prostate (PNW 11, 20), SV (PNW 11, 20)
weight

(Howdeshell et al..
2007)

SD rats; oral/gavage; GD 14-18;
0, 500 mg/kg/d; PND 3, PND 14,
adult (7-11 months of age)

-	| AGD (500/PND 3)

-	1 absolute LABC weight (500/adult)

-	Low incidence of testes and epididymal malformations, vas
deferens and gubernacular agenesis (500/adult; not
statistically significant)

-	Testicular degeneration (500/adults)

Unaffected outcomes

-	NR (500/PND 14, adult); absolute testes, glans penis, ventral
prostate, SV, epididymis weight (500/adult); hypospadias

(Jiane et al.. 2007)

SD rats; oral/gavage; GD 14-18;
0, 250, 500, 750, 1000 mg/kg/d;
PND 1, 7, 35, 70

-	i serum testosterone (250/PND 70)

-	| AGD (500/PND 1)

-	Hypospadias (500)

-	Cryptorchidism (250)

-	i relative testis & epididymis weight (500/PND 70)

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Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])

(Clewell et al..
2013b)

SD rats; oral/feed; GD 12-PND
14; 0, 7600 ppm (eq. 642-1,138
mg/kg/d); PND 2, 14, 49-50

-	| AGD (642/PND 2, 14)

-	NR (642/PND 14, 49-50)

-	Testicular pathology (MNGs, Leydig cell aggregates, t # of
gonocytes) (642/PND 2)

-	1 absolute and relative LABC & SV weight (642/PND 49-
50)

-	Reproductive tract malformations (incomplete epididymis,
flaccid epididymis, enlarged testis (unilateral); PND 49-50)

-	Testicular pathology (MNGs, Leydig cell aggregation, f # of
gonocytes; PND 2) (tubular/rete dilation, atrophic tubules,
MNGs; PND 49-50)

Unaffected outcomes

-	AGD (PND 49-50); testis testosterone (PND 49-50),
gubernacular cord length (PND 49); absolute glans penis,
cowpers gland weight, testis, epididymis, ventral prostate
weight (PND 49-50)





(Zhane et al..
2004)

SD rats; oral/gavage; GD 1-PND
21; 0, 50, 250, 500 mg/kg/day;
PND 4, 21, 70

-	| AGD (250/PND 4)

-	Cryptorchidism (500/PND 21)

-	Testicular pathology (testicular atrophy, epididymal agenesis
(250/PND 70)

-	i absolute epididymis weight (250/PND 70)

-	Altered sperm parameters at PND 70 (J, epididymal sperm #
(500), I motility (250), J, sperm heads per testis (250)

(Wine et al.. 1997:
NTP. 1995)

CD SD rats; oral/feed; 2-
generation (continuous breeding
protocol); 0, 0.1, 0.5, 1.0% (eq. 53,
256, 509 mg/kg/d for males and
80, 385, 794 mg/kg/d for females)

-	i mating, pregnancy, and fertility indices (509-794/F1)

-	i absolute testis, relative SV & prostate weight (509/F1)

-	i epididymal sperm number & total spermatid heads in the
testis (509/F1)

-	Testicular pathology inFl (degeneration of seminiferous
tubules (256), interstitial cell hyperplasia (509),
underdeveloped epididymis (509), apparent sperm content
reduction (509)

Unaffected outcomes

-	Mating, pregnancy and fertility indices (F0); epididymal
sperm motility and percent abnormal (Fl)

(Hieuchi et al..
2003)

Dutch-Belted rabbits; oral/gavage;
GD 15-29; 0, 400 mg/kg/d; PNW
6-25

-	Undescended testes in 1/17 pups (400/PNW 12); in same pup
malformed prepuce, hypospadias, hypoplastic seminal vesicle
and prostate, and agenesis of bulbourethral gland also
observed

-	i paired testes (400/PNW 12) & accessory sex gland weight
(400/PNW 12, 25)

-	Altered sperm parameters (J, ejaculate volume, j sperm
concentration j total sperm/ejaculate, j morphologically
normal sperm, f acrosome-nuclear defects) (400/PNW 22-
24)

-	t incidence of histopathological changes in the seminiferous
epithelium (400/PNW 25)

-	i Serum testosterone (400/PNW 6)

Unaffected outcomes

-	Epididymal weight (PNW 12, 25); agenesis of epididymides;
mating ability (PND 22-24); sperm parameters (daily sperm
production, caput epididymal sperm reserve, cauda





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Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])





epididymal sperm reserve) (PNW 22-24); serum testosterone
(PNW 12, 25)

(McKinnell et al..
2009)

Marmoset; oral/gavage; GW 7-15;
0, 500 mg/kg/d MBP; PND 1-5
(birth, n=6 offspring) or 18-21
months of age (adult, n = 5)

-	t Incidence of clusters of undifferentiated germ cells in 2 out
of 6 animals (400/birth)

Unaffected outcomes

-	Reproductive tract malformations (hypospadias,
cryptorchidism, small testes/impaired spermatogenesis, focal
testicular dysgenesis)

-	Plasma testosterone (birth)

-	Absolute testis weight

-	# of germ cells/testis, germ cell proliferation or
differentiation, # Sertoli cells/testis, germ cell: Sertoli cell
ratio

-	MNGs

-	Germ cell # and proliferation, Sertoli cell #, germ: Sertoli cell
ratio





" Study design comprises species/strain, exposure route, exposure duration, doses, and timing of evaluation.
AGD = anogenital distance; e = embryonic day; GD = gestation day; LABC = levator ani/bulbocavernosus muscle; NR =
nipple retention; PPS = preputial separation; PND = postnatal day; PNW = postnatal week; PS = phthalate syndrome; SD =
Sprague-Dawley; SV = seminal vesicle

5452	B.5 DIBP Study Summaries

5453

Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])

(Howdeshell et al..
2008)

SD rats; oral/gavage; GD 8-18; 0,
100, 300, 600, 900 mg/kg/d; GD
18

- 1 ex vivo fetal testes testosterone production (300)

(Hannas et al..
2011)

SD rats; oral/gavage; GD 14-18;
0, 100, 300, 600, 900 mg/kg/d; GD
18

-	1 ex vivo fetal testes testosterone production (300)

-	i expression of steroidogenic genes in fetal testes (St. IR
(300), Cyplla ( 100))

(Grav et al.. 2021)

Harlan SD rats; oral/gavage; GD
14-18; 0, 100, 200, 300, 500, 600,
750, 900 mg/kg/d; GD 18

-	1 ex vivo fetal testes testosterone production (300)

-	i fetal testicular expression of Ins!3 (300) and steroidogenic
genes (e.g., Star (200), Cvpllal (300), Hsd3b (200), Scarbl
(200) Cypl 7a1 (200)))

Charles River SD rats;
oral/gavage; GD 14-18; 0, 100,
300, 600, 900 mg/kg/d; GD 18

-	1 ex vivo fetal testes testosterone production (300)

-	i fetal testicular expression of Ins!3 (300) and steroidogenic
genes {e.g., Star (600), Cvpllal (600), Hsd3b (600), Scarbl
(600), Cypl 7a1 (600))

(Borch et al..
2006a)

Wistar rats; oral/gavage; GD 7-19;
0, 600 mg/kg/d; GD 19

-	i fetal testicular testosterone content & ex vivo testicular
testosterone production (600, not statistically significant)

-	| AGD (600)

-	Testicular pathology (Leydig cell clusters, j staining intensity
of StAR in Leydig cells) (600)

Wistar rats; oral/gavage; GD 7-21;
0, 600 mg/kg/d; GD 20/21

-	i fetal testicular testosterone content & ex vivo testicular
testosterone production (600)

-	| AGD (600)

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Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])





- Testicular pathology (Leydig cell clusters, Sertoli cell
vacuolization, central localization of gonocytes, MNGs, J,
staining intensity of StAR and P450scc in Leydig cells) (600)

(Bobere et al..
2008)

Wistar rats; oral/gavage; GD 7-19;
0, 600 mg/kg/d; GD 19

- 1 testicular mRNA expression of SR-B1, StAR, P450scc,
Cvpl 7, InsB (600)

Wistar rats; oral/gavage; GD 7-21;
0, 600 mg/kg/d; GD 21

- 1 testicular mRNA expression of SR-B1, StAR, P450scc,
Cvpl 7, InsB (600)

(Saillenfait et al..
2006)

SD rats; oral/gavage; GDs 6-20; 0,
250, 500, 750, 1000 mg/kg/d; GD
21

-	t incidence ectopic testis (750)

-	Unilateral or bilateral undescended testes (500)

-	t incidence of ureter variations (1,000)

-	t degree of transabdominal testicular migration in relation to
the bladder (500)

(Saillenfait et al..
2008)

SD rats; oral/gavage; GDs 12-21;
0, 125, 250, 500, 625 mg/kg/d;
PND 1-122

-	| AGD (250/PND 1)

-	NR (250/PND 12-14 & PND 76-122)

-	Delayed PPS (500)

-	Reproductive tract malformations ((hypospadias, exposed os
penis, nonscrotal testes) (500/PND 76-122)

-	Underdeveloped or absent testis and/or epididymis (250/PND
76-122); cleft prepuce (625/PND 76-122))

-	i absolute weight of testes (625/PND 76-122), epididymis
(500/PND 76-122), SV (500/PND 76-122), prostate '
(250/PND 76-86; 500/PND 111-122)

-	Testicular pathology (epididymal oligospermia or
azoospermia (250/PND 76-86), interstitial cell hyperplasia
(500/PND 76-86), tubular necrosis (250/PND 76-86), tubular
atrophy/hypoplasia (250/PND 76-86))

(Saillenfait et al..
2017)

SD rats; oral/gavage; GDs 13-19;
0, 250 mg/kg/d; GD 19

-	i fetal testes testosterone production (250)

-	| AGD (250)

-	i expression of steroidogenic genes (Hmg-CoAR, Hmg-CoAS,
StAR, SR-B1, P450cl 7) in fetal testes (250)

Unaffected outcomes

-	mRNA expression of P450scc, 3B-HSD in fetal testes;
external malformations

(Wane etal.. 2017)

ICR mice; dietary; GD 0-21; 0,
450 mg/kg/d; PND 21-80

-	i testosterone in serum & testes (450/PND 21)

-	i absolute testes weight (450/PND 21)

-	i mRNA and protein levels of steroidogenic genes (450/PND
21 & 80)

Unaffected outcomes

-	AGD (PND 21); absolute epididymal weight (PND 80);
testosterone in serum & testes (PND 80); sperm concentration
& motility (PND 80)

ICR mice; dietary; GD 0-PND 21;
0,450 mg/kg/d; PND 21-80

-	i testosterone in serum & testes (450/PND 21 & 80)

-	i absolute testes weight (450/PND 21)

-	i mRNA and protein levels of steroidogenic genes (450/PND
21 & 80)

-	1 Sperm concentration & motility (450/PND 80)

Unaffected outcomes

-	AGD (PND 21); absolute epididymal weight (PND 80)

Page 196 of 209


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Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])

" Study design comprises species/strain, exposure route, exposure duration, doses, and timing of evaluation.
AGD = anogenital distance; GD = gestation day; LABC = levator ani/bulbocavernosus muscle; NR = nipple retention; PPS
= preputial separation; PND = postnatal day; PNW = postnatal week; PS = phthalate syndrome; SD = Sprague-Dawley; SV
= seminal vesicle

5454	B.6 DCHP Study Summaries

5455

Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])

(Hoshino et al..
2005)

SD rats; oral/feed; 2-generation
study; 0, 240, 1200, 6000 ppm
(equivalent to 16, 80, 402 & 18,
90, 457 mg/kg/d for F0 and F1
males; 21, 104, 511 and 21, 107,
534 mg/kg/d for F0 and F1
females)

-	| AGD (511/F1 pups (PND 4)) (107/F2 pups (PND 4))

-	NR (511/F1 pups (PND 14)) (107/F2 pups (PND 12))

-	Soft small sized testes (457/F1 adults)

-	i absolute prostate weight (16/F1 adults)

-	Testicular pathology (seminiferous tubule atrophy (90/F1
adults)

-	i spermatid head counts in the testes (90/F1 adults)

Unaffected outcomes

-	Mating index, fertility index, gestation length, gestation
index, birth index (F0 and F1 mating pairs); pup viability
index or physical development (pinna unfolding, incisor
eruption, eye opening) (Fl, F2 pups); sperm motility (F0, F1
adults), cauda epididymal sperm count (F0, Fl adults),
abnormal or tailless sperm (F0, Fl adults), and spermatid
head counts in testis (F0 adults); serum hormone
(testosterone, FSH, LH) levels (F0, Fl adult males); absolute
prostate weight (F0 adults); testicular pathology (F0 adults);
PPS (Fl)

(Saillenfait et al..
2009b)

SD rats; oral/gavage; GD 6-20; 0,
250, 500, 750 mg/kg/d; GD 21

-	| AGD (250)
Unaffected outcomes

-	Testes descent

(Yamasaki et al..
2009)

SD rats; oral/gavage; GD 6-PND
20; 0, 20, 100, 500 mg/kg/d; PND
4-70

-	| AGD (500/PND 4)

-	NR (500/PND 13)

-	Delayed PPS (500)

-	1 relative prostate & LABC weight (500/PND 70)

-	1 testicular germ cells (500/PND 70)

-	t incidence of hypospadias (500/PND 49-70)
Unaffected outcomes

-	Relative testis, epididymis, SV weight (PND 70)

(Furret al.. 2014)

SD rats; oral/gavage; GD 14-18;
0, 100, 300, 600, 900 mg/kg/d
(Block 23) or 0, 33, 100, 300
mg/kg/d (Block 33); GD 18

-	Block 23: I ex vivo fetal testes testosterone production (100)

-	Block 33: J, ex vivo fetal testes testosterone production (100)

(Grav et al.. 2021)

Harlan SD rats; oral/gavage; GD
14-18; 0, 33, 100, 300, 600, 900
mg/kg/d; GD 18

-	1 ex vivo fetal testes testosterone production (33)

-	i fetal testicular expression of Ins!3 (100) and steroidogenic
genes (e.g., Star (100), Cypllal (100), Hsd3b (100), Scarbl
(100), Cypl 7a 1 (100))) '

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Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])



Charles River SD rats;
oral/gavage; GD 14-18; 0, 100,
300, 600, 900 mg/kg/d; GD 18

-	1 ex vivo fetal testes testosterone production (100)

-	i fetal testicular expression of Ins!3 (100) and steroidogenic
genes (e.g., Star (100), Cypllal (300), Hsd3b (100), Scarbl
(100), Cvpl7al (100))

(Ahbab and Barlas.
2015)

Wistar albino rats; oral/gavage;
GD 6-19; 0, 20, 100, 500 mg/kg/d;
GD 20

-	1 immunohistochemical staining for 3(3-HSD (20)

-	i serum testosterone (100)

-	| AGD (20)

-	Testicular pathology (atrophic seminiferous tubules (20), J,
germ cells in tubules (20), Sertoli cell only tubules (100),
detached cells from tubular wall (20), MNGs (100))

-	| number of medium and large Leydig cell clusters (20))

(Li et al.. 2016)

SD rats; oral/gavage; GD 12-21;
0, 10, 100, 500 mg/kg/d; GD 21.5

-	1 testes mRNA for Star (10), Scarbl (500), hsd3bl (10),
hsdl 7b3 (10), Insl3 (100), Lhcgr (1000)

-	i testicular testosterone (100)

-	| AGD (100)

-	Testicular pathology (focal testis dysgenesis (100), MNGs
(100)); I Leydig cell size, cytoplasmic size, nuclear size (10);
t # & size of Leydig cell clusters (10))

" Study design comprises species/strain, exposure route, exposure duration, doses, and timing of evaluation.
AGD = anogenital distance; GD = gestation day; LABC = levator ani/bulbocavernosus muscle; NR = nipple retention; PPS
= preputial separation; PND = postnatal day; PNW = postnatal week; PS = phthalate syndrome; SD = Sprague-Dawley; SV
= seminal vesicle

5456	B.7 DINP Study Summaries

5457

Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])

(Borch et al.. 2004)

Wistar rats; oral/gavage; GD 7-21;
0, 750 mg/kg/d; CAS no. 28553-
12-0; GD 21

-	J, Testicular testosterone (750)

-	1 ex vivo testicular testosterone production (750)
Unaffected outcomes

-	Plasma testosterone

(Furret al.. 2014)

Harlan SD rats; oral/gavage; GD
14-18; 0, 750 mg/kg/d; CASRN
not reported (Block 1 and 5
reported to use DINP from Exxon;
Block 7 reported to use DINP from
BASF); GD 18 (2 hpost final
dose)

- 1 ex vivo fetal testes testosterone production (750) (effect
reported for Blocks 1, 5, and 7 studies)

(Hannas et al..
2011)

SD rats; oral/gavage; GD 14-18; 0,
500, 750, 1,000, 1,500 mg/kg/d;
CASRN 28553-12-0 & 68515-48-
0; GD 18

-	1 ex vivo fetal testes testosterone production (500)

-	i expression of steroidogenesis genes in fetal testes (i.e.,
StAR (1000), Cvplla (1000))

(Grav et al.. 2021)

Harlan SD rat; oral/gavage; GD
14-18; 0, 500, 750, 1000, 1500
mg/kg/d; GD 18

-	1 ex vivo fetal testes testosterone production (500)

-	i fetal testicular expression of Insl3 (500) and steroidogenic
genes (e.g., Star (500), Cvpllal (500), Cvpllb2 (1000),
Hsd3b (500), Scarbl (500))

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Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])

(Adamsson et al..
2009)

SD rats; oral/gavage; el3.5-17.5;
0, 250, 750 mg/kg/d; CASRN not
reported; el9.5

-	t Testicular mRNA expression of GATA-4 (750) andInsB
(750)

Unaffected outcomes

-	Testicular testosterone, testicular pathology, testicular protein
expression of StAR, P450scc, 3(3-HSD, AR; testicular mRNA
expression of Star, P450scc, 3fi-HSD, SF-1

(Bobere et al..
2011)

Wistar rats; gavage; GD 7-PND
17; 0, 300, 600, 750, 900 mg/kg/d;
CASRN 28553-12-0; GD 21, PND
13, PND 90

-	i Testicular testosterone (600/GD 21 (no dose-response))

-	| AGD (900/PND 13)

-	NR (750/PND 13)

-	Testicular pathology (MNGs (600/GD 21); enlarged diameter
of seminiferous chords (750/GD 21), gonocytes with central
location in chords (750/GD 21))

-	1 sperm motility (600/PND 90)

Unaffected outcomes

-	Serum testosterone (GD 21), ex vivo testicular testosterone
production (GD 21); testes testosterone (PND 90); NR (PND
90); AGD (PND 90); reproductive organ weight (PND 90);
testicular pathology (PND 90)

(Clewell et al..
2013a)

SD rats; oral/gavage; GD 12-19; 0,
50, 250, 750; CASRN 68515-48-0;
GD 19 (2 h post-dosing) or GD 20
(24 h post-dosing)

-	i Testicular testosterone (250/GD 19)

-	Testicular pathology (MNGs (250/GD 20), Leydig cell
aggregates (750/GD 20))

Unaffected outcomes

-	Testicular testosterone (GD 20); AGD (GD 20)

(Clewell et al..
2013b)

SD rats; oral/feed; GD 12-PND
14; 0, 760, 3,800, 11,400 ppm (est.
0, 56, 288, 720 mg/kg/d); CASRN
68515-48-0; PND 2, 14 or 49

-	| AGD (720/PND 14)

-	Testicular pathology (t Leydig cell aggregates (720/PND 2),
MNGs (288/PND 2))

Unaffected outcomes

-	Testicular testosterone (PND 2, 49); AGD (PND 2, 49); NR
(PND 14, 49); absolute testis and epididymis weight (PND 2,
49); absolute testes, epididymis, SV, ventral prostate, glans
penis, LABC, Cowper's Glands weight (PND 49); testicular
pathology (PND 49); PPS; reproductive tract malformations
(e.g., hypospadias, exposed os penis, undescended testes,
epididymal agenesis) (PND 49)

(Masutomi et al..
2003)

SD rats; diet; GD 15-PND 10; 0,
400, 4,000, 20,000 ppm (est. 31-
66, 307-657, 1165-2,657
mg/kg/d); CASRN 28553-12-0;
PNDs 2, 27, or 77

-	i Absolute testes weight (20,000/PND 27)

-	Testicular pathology (degeneration of Sertoli cells
(20,000/PND 77), degeneration of stage XIV meiotic
spermatocytes (20,000/PND 77), scattered cell debris in
ducts of epididymis (20,000/PND 77)))

Unaffected outcomes

-	AGD (PND 2); PPS; absolute testes weight (PND 77)

(Li et al.. 2015a)

SD rats; oral/gavage; GD 12-21; 0,
10, 100, 500, 1,000 mg/kg/d;
CASRN not provided; GD 21.5

-	i testicular testosterone (1,000)

-	i testicular gene expression (InsB (10), Lhcgr (500), Star
(500), Cvpllal (100\Hsd3bl (100), Cvpl7al (100),
Hscll 7b3 (1,000))

-	Testicular pathology (focal testis dysgenesis (100); MNGs
(100); clusters of Leydig cells (10))

Unaffected outcomes

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Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])





- AGD

(Gray et al.. 2000)

SD rats; oral/gavage; GD 14-PND
3; 0, 750 mg/kg/d; CASRN 68515-
48-0; PND 2-mature adults (3-7
months of age)

-	NR in 2/52 male pups (750/PND 13)

-	Permanent nipples (750/3-7 months)

-	Reproductive malformations (small and atrophic testes, fluid
filled testes lacking sperm, epididymal agenesis)

Unaffected outcomes

-	Serum testosterone levels, PPS, AGD (PND 2), reproductive
malformations (hypospadias, cleft phallus, vaginal pouch, SV
agenesis, undescended testes, testis absent, abnormal
gubernacular cord (3-7 months)), reproductive organ weight
(i.e., testes, LABC, SC, glans penis, ventral prostate,
epididymis, cauda epididymis, caput-corpus epididymis)

(Waterman et al..
2000)2

SD rats; oral/feed; 1-generation
study (10 wks prior to mating-
PND 21); 0, 0.5, 1.0, 1.5% (est. 0,
360-923, 734-1731, 1,087-2,246
mg/kg/d); CASRN 68515-48-0

-	t absolute testes and epididymis (left only) weight in P0
(1.5%)

Unaffected outcomes

-	Reproductive indices (e.g., mating index, fertility index,
gestation index, birth index, sex ratio); absolute epididymis,
prostate, SV weight inPO; repro

(Waterman et al..
2000)3

SD rats; oral/feed; 2-generation
study; 0, 0.2, 0.4, 0.8% (est. 133-
153, 271-307, 543-577 mg/kg/d
during gestation); CASRN 68515-
48-0

Unaffected outcomes (both generations)

- Reproductive indices (e.g., mating index, fertility index,
gestation index, birth index, sex ratio); absolute testes,
epididymis, prostate and SV weight; testicular pathology

" Study design comprises species/strain, exposure route, exposure duration, doses, and timing of evaluation.
AGD = anogenital distance; e = embryonic day; GD = gestation day; LABC = levator ani/bulbocavernosus muscle; NR =
nipple retention; PPS = preputial separation; PND = postnatal day; PNW = postnatal week; PS = phthalate syndrome; SD =
Sprague-Dawley; SV = seminal vesicle

5458	B.8 DIDP Study Summaries

5459

Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])

(Hellwie et al..
1997)

Wistar rats; oral/gavage; GD 6-15;
0, 40, 200, 1,000 mg/kg/d; GD 21

- No phthalate syndrome-related effects reported (exposure did
not cover critical window)

(Waterman et al..
1999)

SD rats; oral/gavage; GD 6-15; 0,
100, 500, 1,000 mg/kg/d; GD 21

- No phthalate syndrome-related effects reported (exposure did
not cover critical window)

(Hannas et al..
2012)

SD rats; oral/gavage; GD 14-18;
0, 500, 750, 1,000, 1,500 mg/kg/d;
GD 18

Unaffected outcomes

- Ex vivo testes testosterone production; steroidogenic gene
expression

(Furretal.,2014)

SD rats; oral/gavage; GD 14-18;
0, 500, 750, 1,000, 1,500 mg/kg/d;
GD 18

Unaffected outcomes
- Ex vivo testes testosterone production

(Gray et al.. 2021)

Charles River SD rat; oral/gavage;
GD 14-18; 0, 300, 750, 1,000,
1,500 mg/kg/d; GD 18

Unaffected outcomes

-	Ex vivo testes testosterone production

-	Steroidogenic gene expression in the fetal testes

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Reference

Study Design"

Reported Phthalate Syndrome-Related Effects
(LOAEL/LOEL [mg/kg/day]/Timing of Evaluation [if
Different than Listed under Study Design])

(Hushka et al..
2001)

SD rats; oral/feed; 2-generation
study; 0, 0.2%, 0.4%, 0.8% (Study
A) '

Unaffected outcomes (both generations)

- Reproductive indices (i.e., mating, fertility, gestation and
birth index); reproductive organ (i.e., prostate, testes,
epididymis, SV) weight; sperm parameters (sperm count,
motility, morphology); testicular pathology, gross external
abnormalities

(Hushka et al..
2001)

SD rats; oral/feed; 2-generation
study; 0, 0.02, 0.06, 0.2%, 0.4%
(Study B)

Unaffected outcomes (both senerations)

- Reproductive indices (i.e., mating, fertility, gestation and
birth index); AGD (PND 0); NR (PND 12-13); age at PPS,
gross external abnormalities

" Study design comprises species/strain, exposure route, exposure duration, doses, and timing of evaluation.
AGD = anogenital distance; GD = gestation day; LABC = levator ani/bulbocavernosus muscle; NR = nipple retention; PPS
= preputial separation; PND = postnatal day; PNW = postnatal week; PS = phthalate syndrome; SD = Sprague-Dawley; SV
= seminal vesicle

5460

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Appendix C

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Methodology for Preliminary Dose-Response Modeling

C.l General Approach

•	Data from the peer-reviewed literature and cited in summary tables for each key outcome
throughout Section 3.1.3 were combined to produce a single dose response curve for each high-
priority and manufacturer-requested phthalate for several phthalate syndrome effects {i.e.,
decreased AGD, nipple/areolae retention, testicular pathology, and hypospadias).

•	Studies included in the analysis all exposed the pregnant rat to the phthalate ester during the male
programing window (at a minimum). Studies included those that administered the chemical daily
from oral gavage on a mg/kg/day basis and those that administered the chemical in the diet on a
ppm basis. Studies used Charles River Sprague-Dawley, Wistar and other rat strains. Phthalate
ester effects including testicular histology and hypospadias used only data from adult male rat
offspring.

•	Effects modeled included in male rat offspring included decreased AGD, percent of males per
litter with retained nipples, percent with testicular histopathological lesions {i.e., seminiferous
tubule atrophy), and percent with hypospadias.

•	Data were not modeled for phthalate syndrome effects when no data were available for one of
the five high-priority phthalate esters {i.e., DEHP, BBP, DIBP, DBP and DCHP).

•	Similar models have been constructed for DINP when effects are equal to or exceed 15 percent
of control, but some of the data are not yet peer-reviewed and were not included in the
preliminary dose-response analysis.

•	Data were fit using GraphPad Prism 8 software to four parameter logistic regression models
(4PL). For each effect the top and bottom of the curve was constrained, as appropriate (described
in more detail below). Since an RPF approach for CRA is being proposed, the slope was
constrained to "shared value for each dataset." This improves the confidence in the ED50 value
and 95 percent confidence intervals (95 percent CI) and is a biologically plausible approach
because available data indicate that these phthalate esters share a common MOA.

€.2 Anogenital Distance (AGD)

C.2.1 Calculation of Individual Phthalate Ester AGD Dose-Response Models

AGD data from studies published in the literature was entered by study into a GraphPad Prism data file
for each individual phthalate ester. All the studies reported AGD for males and some reported the female
offspring AGD as well. The age at AGD measurement ranged from late fetal life to 4 days of age,
although most studies measured AGD at 1 to 2 days of age. The data for each study was normalized with
the control male AGD being 100 percent (the top of the curve) and control female being 0 percent (the
bottom of the curve). If the female AGD was not reported then the bottom of the curve was assigned a
value of 50 percent of the control male AGD, a value that consistently seen in studies with phthalates
and other chemicals. The data from all the studies was combined into a single data set, sorted by dose (in
units of mg/kg/day if administered by oral gavage to the dam daily or estimated mg/kg/day if
administered in ppm in the diet).

The normalized data for each phthalate ester was entered into a single GraphPad Prism file and 4PL
models were run with the bottom constrained to 0 percent, the top constrained to a shared value less than

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110 percent, and the slope constrained to a shared value for each phthalate ester, and the ED50 value and
95 percent CI for each PE was estimated.

C.3 Nipple/Areolae Retention in 13 to 14 Day Old Infant Male Rats

C.3.1 Calculation of Individual Phthalate Ester Dose-Response Models

Nipple/areolae data from studies published in the literature was entered by study into a GraphPad Prism
data file for each individual phthalate ester. All the studies reported nipple retention for males as percent
of males/litter showing any retained areolae/nipples (irrespective if a male had 1 or 12 areolae). Data on
the number of nipples {i.e., 1-12) per male were not reported for all five phthalate esters, and this
measure of nipple retention was not used for dose-response modeling. The data from all the studies was
combined into a single data set, sorted by dose (in units of mg/kg/day if administered by oral gavage to
the dam daily or estimated mg/kg/day if administered in ppm in the diet).

The normalized data for each phthalate ester was entered into a single GraphPad Prism file and 4PL
models were run with the bottom constrained to a shared bottom greater than 0 percent, the top
constrained to 100 percent, and the slope constrained to a shared value for each PE, and the ED50 value
and 95 percent CI for each phthalate ester was estimated.

€.4 Testicular Pathology - Seminiferous Tubule Atrophy

C.4.1 Calculation of Individual Phthalate Ester Dose-Response Models

In the phthalate syndrome, histopathology of the testis and epididymis often occur concurrently along
with gross malformations of these tissues and these effects are among the more sensitive effects
resulting from in utero phthalate ester exposure. One of the most frequently reported effects following
phthalate ester exposure is seminiferous tubular atrophy/agenesis of the testis in adult male rats (it can
be with uni- or bilateral). Histopathology of the epididymis is less frequently reported. Only
histopathology scores greater than 1 (minimal effect or single tubule affected) were used in the initial
dose-response analysis. Abnormal testis differentiation can result from direct testicular effects of the
phthalate ester on the endocrine and paracrine environment disrupting seminiferous tubular
development, Leydig cell differentiation and vasculature differentiation of the testis during fetal and
neonatal life. Histopathological alterations of the testes also can also result from indirect in utero effects
post-puberty as a result of excessive fluid back pressure in the testis caused by epididymal abnormalities
that prevent fluid and sperm flow from the testis and from testis nondescent associated with
gubernacular abnormalities (spermatogenesis does not occur in such testes being temperature sensitive).
The data from all the studies was combined into a single data set, sorted by dose (in units of mg/kg/day
if administered by oral gavage to the dam daily or estimated mg/kg/day if administered in ppm in the
diet).

The normalized data for each phthalate ester was entered into a single GraphPad Prism file and 4PL
models were run with the bottom constrained to shared bottom greater than 0 percent, the top
constrained to 100 percent, and the slope constrained to a shared value for each PE, and the ED50 value
and 95 percent CI for each phthalate ester was estimated.

C.5 Hypospadias

In the phthalate syndrome, malformations of external genitalia are one of the least sensitive effects in
some rat strains. All the studies reported incidence of hypospadias for adult F1 males following
gestational exposure to each of the individual phthalate esters. The data from all the studies was

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5545	combined into a single data set, sorted by dose (in units of mg/kg/day if administered by oral gavage to

5546	the dam daily or estimated mg/kg/day if administered in ppm in the diet).

5547

5548	The normalized data for each phthalate ester was entered into a single GraphPad Prism file and 4PL

5549	models were run with the bottom constrained to 0 percent, the top constrained to 100 percent, and the

5550	slope constrained to a shared value for each PE and the ED50 value and 95 percent CI for each phthalate

5551	ester was estimated.

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Appendix D Occupational Exposure Assessment

EPA Program Data for Identifying Sites

•	CDR: CDR can be used to identify manufacturing and import sites that handle multiple
phthalates. All six phthalates are reported in CDR and represent all import and manufacturing
sites producing the chemical at or above a specified threshold. Because CDR reporting is done
on a site-by-site basis, any site reporting more than one of the designated phthalates is a site with
cumulative exposure and release potential. Both 2016 and 2020 CDR, as well as record relevant
data, can be used for each instance of phthalate production including the number of workers,
chemical concentration, and volume of chemical.

•	TRI, DMR. and NEI: These EPA release datasets can be used similarly to the CDR data for
determining sites with potential for cumulative exposure and release. Only a portion of the
selected phthalates are required to report to some of these programs, however, limiting the
dataset utility. Nonetheless, the datasets can be useful for determining cumulative releases
between the reporting chemicals. North American Industry Classification System code and other
facility reporting parameters will be used to assign COUs to EPA release datasets.

•	RCRAInfo: This dataset is split into multiple modules, with the two main modules of interest
for release assessment being the E-manifest module and the Biennial Report module. All
hazardous waste shipments are reported in the E-manifest module and represent movement from
the hazardous waste generator to treatment, storage, or disposal facilities. E-manifest reporting
does not have a reporting threshold. The Biennial Report is an annual summary of hazardous
waste that is generated at a facility, including the quantity and nature of the waste and its
disposition {i.e., recycling, treatment, storage, or disposal). Only Large Quantity Generators
(LQC) are required to submit a Biennial Report, but LQCs are defined by the overall volume of
hazardous waste and is not chemical specific like CDR, TRI, or NEI and therefore the dataset
may provide EPA with a better understanding of some sites handling smaller quantities of
phthalates. Like many other EPA programs, only a portion of the selected phthalates are required
to report, limiting the overall utility of RCRAInfo for release assessments.

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Appendix E Glossary of Key Terms

Additivity (U.S. EPA. 2007b. 2000): "when the effect of the combination of chemicals can be estimated
directly from the sum of the scaled exposure levels (dose addition) or of the responses (response
addition) of the individual components."

Aggregate exposure (40 CFR § 702.33): "means the combined exposures to an individual from a single
chemical substance across multiple routes and across multiple pathways."

Best available science (	L33): "means science that is reliable and unbiased. Use of best

available science involves the use of supporting studies conducted in accordance with sound and
objective science practices, including, when available, peer reviewed science and supporting studies and
data collected by accepted methods or best available methods (if the reliability of the method and the
nature of the decision justifies use of the data). Additionally, EPA will consider as applicable:

(1)	The extent to which the scientific information, technical procedures, measures, methods,
protocols, methodologies, or models employed to generate the information are reasonable for and
consistent with the intended use of the information;

(2)	The extent to which the information is relevant for the Administrator's use in making a decision
about a chemical substance or mixture;

(3)	The degree of clarity and completeness with which the data, assumptions, methods, quality
assurance, and analyses employed to generate the information are documented;

(4)	The extent to which the variability and uncertainty in the information, or in the procedures,
measures, methods, protocols, methodologies, or models, are evaluated and characterized; and

(5)	The extent of independent verification or peer review of the information or of the procedures,
measures, methods, protocols, methodologies or models."

Biomonitoring (U.S. EPA. 2019a): "measures the amount of a stressor in biological matrices."

Category of chemical substances (	2625(c)(2)(A)): "means a group of chemical substances

the members of which are similar in molecular structure, in physical, chemical, or biological properties,
in use, or in mode of entrance into the human body or into the environment, or the members of which
are in some other way suitable for classification as such for purposes of [TSCA], except that such term
does not mean a group of chemical substances which are grouped together solely on the basis of their
being new chemical substances."

Chemical substance (	02(2)): "means any organic or inorganic substance of a particular

molecular identity, including—(i) any combination of such substances occurring in whole or in part as a
result of a chemical reaction or occurring in nature, and (ii) any element or uncombined radical. Such
term does not include—(i) any mixture, (ii) any pesticide (as defined in the Federal Insecticide,
Fungicide, and Rodenticide Act [7 U.S.C. 136 et seq.]) when manufactured, processed, or distributed in
commerce for use as a pesticide, (iii) tobacco or any tobacco product, (iv) any source material, special
nuclear material, or byproduct material (as such terms are defined in the Atomic Energy Act of 1954 [42
U.S.C. 2011 et seq.] and regulations issued under such Act), (v) any article the sale of which is subject
to the tax imposed by section 4181 of the Internal Revenue Code of 1986 [26 U.S.C. 4181] (determined
without regard to any exemptions from such tax provided by section 4182 or 4221 or any other
provision of such Code) and any component of such an article (limited to shot shells, cartridges, and
components of shot shells and cartridges), and (vi) any food, food additive, drug, cosmetic, or device (as

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5633

5634

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5640

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such terms are defined in section 201 of the Federal Food, Drug, and Cosmetic Act [21 U.S.C. 321])
when manufactured, processed, or distributed in commerce for use as a food, food additive, drug,
cosmetic, or device."

Condition of use (COU) (15 U.S.C. § 2602(4)): "means the circumstances, as determined by the
Administrator, under which a chemical substance is intended, known, or reasonably foreseen to be
manufactured, processed, distributed in commerce, used, or disposed of."

Consumer exposure (40 CFR § 711.3): Human exposure resulting from consumer use. This exposure
includes passive exposure to consumer bystanders.

Consumer use (	): "means the use of a chemical substance or a mixture containing a

chemical substance (including as part of an article) when sold to or made available to consumers for
their use."

Cumulative risk (	): "The combined risks from aggregate exposures to multiple agents

or stressors."

Cumulative risk assessment (CRA) (	03): "An analysis, characterization, and possible

quantification of the combined risks to health or the environment from multiple agents or stressors."

Dose additivity (	)07b. 2003. 2000): "when each chemical behaves as a concentration or

dilution of every other chemical. The response of the combination of chemicals is the response expected
from the equivalent dose of an index chemical (the chemical selected as a basis for standardization of
toxicity of components in a mixture). The equivalent dose is the sum of component doses scaled by their
toxic potency relative to the index chemical."

Fenceline exposure: General population exposures occuring in communities near facilities that emit or
release chemicals to air, water, or land with which they may contact.

Index chemical (	XX)): "The chemical selected as the basis for standardization of toxicity of

components in a mixture. The index chemical must have a clearly defined dose-response relationship."

Integrated addition: a hybrid additivity approach that incorporates both dose addition and response
addition for dichotomous endpoints, thus, producing a mixture estimate that is the probabilistic risk of
the adverse endpoint of concern.

Margin of exposure (MOE) (U.S. EPA. 2002): "a numerical value that characterizes the amount of
safety to a toxic chemical-a ratio of a toxicological endpoint (usually a NOAEL [no observed adverse
effect level]) to exposure. The MOE is a measure of how closely the exposure comes to the NOAEL."

Mixture (s 1	10)): "means any combination of two or more chemical substances if the

combination does not occur in nature and is not, in whole or in part, the result of a chemical reaction;
except that such term does include any combination which occurs, in whole or in part, as a result of a
chemical reaction if none of the chemical substances comprising the combination is a new chemical
substance and if the combination could have been manufactured for commercial purposes without a
chemical reaction at the time the chemical substances comprising the combination were combined."

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Mode of Action (MOA) (	2000): "a series of key events and processes starting with

interaction of an agent with a cell, and proceeding through operational and anatomical changes causing
disease formation."

Non-TSCA exposure: exposure that can be attributed to specific activities that are excluded from the
TSCA definition of "chemical substance," under TSCA Section 3(2), such as a pesticide, food, food
additive, drug, cosmetic, or medical device.

Occupational exposure: Exposure to a chemical substance by industrial or commercial employees.

Occupational non-users (ONU): Employed persons who do not directly handle the chemical substance
but may be indirectly exposed to it as part of their employment due to their proximity to the substance.

Pathways (40 CFR § 702.33): "means the mode through which one is exposed to a chemical substance,
including but not limited to: Food, water, soil, and air."

Point of departure (POD) (	302): "dose that can be considered to be in the range of

observed responses, without significant extrapolation. A POD can be a data point or an estimated point
that is derived from observed dose-response data. A POD is used to mark the beginning of extrapolation
to determine risk associated with lower environmentally relevant human exposures."

Potentially exposed or susceptible subpopulations (PESS) (	02(12)): "means a group of

individuals within the general population identified by the Agency who, due to either greater
susceptibility or greater exposure, may be at greater risk than the general population of adverse health
effects from exposure to a chemical substance or mixture, such as infants, children, pregnant women,
workers, or the elderly."

Reasonably available information (40 CFR § 702.33): "means information that EPA possesses or can
reasonably generate, obtain, and synthesize for use in risk evaluations, considering the deadlines
specified in TSCA section 6(b)(4)(G) for completing such evaluation. Information that meets the terms
of the preceding sentence is reasonably available information whether or not the information is
confidential business information, that is protected from public disclosure under TSCA section 14."

Response addition (U.S. EPA. 2007b. 2003. 2000): "When the toxic response (rate, incidence, risk, or
probability of effects) from the combination is equal to the conditional sum of component responses as
defined by the formula for the sum of independent event probabilities. For two chemical mixtures, the
body's response to the first chemical is the same whether or not the second chemical is present."

Routes (40 CFR § 702.33): "means the particular manner by which a chemical substance may contact
the body, including absorption via ingestion, inhalation, or dermally (integument)."

Sentinel exposure (40 CFR § 702.33): "means the exposure from a single chemical substance that
represents the plausible upper bound of exposure relative to all other exposures within a broad category
of similar or related exposures."

Stressor (U ,S. EPA. 2019a): "Any chemical, physical or biological entity that induces an adverse
response."

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5721	Toxicologic interactions (	307b. 2000): "Any toxic responses that are greater than or less

5722	than what is observed under an assumption of additivity."

5723

5724	Weight of the scientific evidence (40 CFR § 702.33): "means a systematic review method, applied in a

5725	manner suited to the nature of the evidence or decision, that uses a pre-established protocol to

5726	comprehensively, objectively, transparently, and consistently, identify and evaluate each stream of

5727	evidence, including strengths, limitations, and relevance of each study and to integrate evidence as

5728	necessary and appropriate based upon strengths, limitations, and relevance."

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