Draft Non-cancer Human Health Hazard Assessment for Diisononyl Phthalate (DINP) Technical Support Document for the Draft Risk Evaluation CASRNs: 28553-12-0 and 68515-48-0 May 2024

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PUBLIC RELEASE DRAFT

May 2024

EPA Document #EPA-740-D-24-010

May 2024

United States	Office of Chemical Safety and

Environmental Protection Agency	Pollution Prevention

Draft Non-cancer Human Health Hazard Assessment for
Diisononyl Phthalate (DINP)

Technical Support Document for the Draft Risk Evaluation

CASRNs: 28553-12-0 and 68515-48-0

(Representative Structure)

May 2024

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27 TABLE OF CONTENTS

28	SUMMARY	9

29	1 INTRODUCTION	11

30	1.1 Human Epidemiologic Data: Approach and Preliminary Conclusions	11

31	1.2 Laboratory Animal Findings: Summary of Existing Assessments from Other Regulatory

32	Organizations	13

33	1.3 Laboratory Animal Data: Approach and Methodology	16

34	1.3,1 Approach to Identifying and Integrating Laboratory Animal Data	16

35	1,3,2 New Literature Identified and Hazards of Focus for DINP	18

36	2 TOXICOKINETICS	19

37	2.1 Oral Route	19

38	2.2 Inhalation Route	21

39	2.3 Dermal Route	21

40	2.4 Summary	22

41	3 HAZARD IDENTIFICATION	23

42	3.1 Developmental and Reproductive Toxicity	23

43	3.1.1 Summary of Available Epidemiological Studies	23

44	3,1,2 Summary of Laboratory Animals Studies	23

45	3.1.2.1 Developing Male Reproductive System	24

46	3.1.2.1.1 Summary of Studies Evaluating Effects on the Developing Male Reproductive

47	System	24

48	3.1.2.1.2 Mode of Action for Phthalate Syndrome	29

49	3.1.2.2 Other Developmental and Reproductive Outcomes	30

50	3.1.2.3 Conclusions on Reproductive and Developmental Toxicity	41

51	3.2 Liver Toxicity	42

52	3.3 Kidney Toxicity	44

53	3.4 Neurotoxicity	51

54	3.5 Cardiovascular Health Effects	58

55	3.6 Immune Sy stem Toxi city	61

56	3.7 Musculoskeletal Toxicity	67

57	4 DOSE-REPONSE ASSESSMENT	69

58	4.1 Selection of Studies and Endpoints for Non-cancer and Threshold Cancer Health Effects	69

59	4.1.1 Non-cancer Oral Points of Departure for Acute Exposures	69

60	4,1,2 Non-cancer Oral Points of Departure for Intermediate Exposures	76

61	4.1.3 Non-cancer Oral Points of Departure for Chronic Exposures	81

62	4.2 Weight of Scientific Evidence	88

63	4.2.1 POD for Acute and Intermediate Durations	88

64	4,2,2 POD for Chronic Durations	89

65	5 CONSIDERATION OF PESS AND AGGEGRATE EXPOSURE	91

66	5.1 Hazard Considerations for Aggregate Exposure	91

67	5.2 PESS Based on Greater Susceptibility	91

68	6 POINTS OF DEPARTURE USED TO ESTIMATE RISKS FROM DINP EXPOSURE	99

69	REFERENCES	101

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Appendix A EXISTING ASSESSMENTS FROM OTHER REGULATORY AGENCIES OF

DINP	116

Appendix B SUMMARY OF LIVER TOXICITY STUDIES	120

Appendix C FETAL TESTICULAR TESTOSTERONE AS AN ACUTE EFFECT	137

Appendix D SUMMARY OF EPIDEMIOLOGY STUDIES ON REPRODUCTIVE

OUTCOMES	138

Appendix E BENCHMARK DOSE ANALYSIS OF LINGTON ET AL. (1997)	 141

E.l Background	141

E.2 Summary of BMD Modeling Approach	141

E,3 Summary of BMD Modeling Results	142

E.4 Continuous Endpoints	143

E.4.1 Relative Liver Weight - Terminal Sacrifice	143

E.4.1.1 Male F344 Rats	143

E.4.1.2 Female F344 Rats	148

E.4.2 Serum ALT - Male F344 Rats	151

E.4.2.1 6-Month Sacrifice	151

E.4.2.2 18-Month Sacrifice	156

E,5	Dichotomous Endpoints	161

E.5.1 Focal Necrosis in the liver	161

E.5.1.1 Male F344 Rats	161

E.5.1.2 Female F344 Rats	167

E.5.2 Spongiosis hepatis in the liver - Male F344 Rats	172

E.5.3 Sinusoid Ectasia in the Liver Male F344 Rats	177

Appendix F CALCULATING DAILY ORAL HUMAN EQUIVALENT DOSES AND

HUMAN EQUIVALENT CONCENTRATIONS	182

F.l	DINP Non-cancer HED and HEC Calculations for Acute and Intermediate Duration
Exposures	183

F.2 DINP Non-cancer HED and HEC Calculations for Chronic Exposures	184

LIST OF TABLES

Table 1-1. Summary of DINP Non-cancer PODs Selected for Use by other Regulatory Organizations. 14

Table 2-1. Absorption and Excretion Summary of DINP	20

Table 2-2. Metabolites of DINP Identified in Urine from Rats and Humans after Oral Administration. 21
Table 3-1. Summary of DINP Studies Evaluating Effects on the Developing Male Reproductive System

	25

Table 3-2. Summary of DINP Studies Evaluating Effects on Reproduction and Development	31

Table 3-3. Mean Percent of Fetuses in Litter with Skeletal Variations (Waterman et al., 1999)	 35

Table 3-4. Incidence of Visceral, Skeletal, and Soft Tissue Variations (Hellwig et al., 1997)	 36

Table 3-5. Mean Measured Doses (mg/kg-day) from the One-Generation Study of DINP in SD Rats

(Waterman et al., 2000; Exxon Biomedical, 1996a)	37

Table 3-6. F1 Offspring Postnatal Body Weight (Grams) from the One-Generation Study of

Reproduction in SD Rats (Waterman et al., 2000; Exxon Biomedical, 1996a)	38

Table 3-7. Mean Measured Doses (mg/kg-day) from the Two-Generation Study of DINP in SD Rats

(Waterman et al., 2000; Exxon Biomedical, 1996b)	39

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Table 3-8. F1 and F2 Offspring Postnatal Body Weight (Grams) from the Two-Generation Study of

Reproduction in SD Rats (Waterman et al., 2000; Exxon Biomedical, 1996b) 39

Table 3-9. Incidence and Severity of Selected Non-neoplastic Lesions in the Kidneys of Male and

Female F344 Rats Fed DINP for 2 Years (Covance Labs, 1998c) 48

Table 3-10. Summary of Study Evaluating Cardiovascular Outcomes 60

Table 4-1. Summary of NASEM (2017) Meta-Analysis and BMD Modeling for Effects of DINP in Fetal

Testosterone 72

Table 4-2. Dose-Response Analysis of Selected Developmental Studies Considered for Deriving the

Acute Non-cancer POD 74

Table 4-3. Dose-Response Analysis of Selected Studies Considered for Deriving the Intermediate Non-

ancer POD 79

Table 4-4. Summary of BMD Model Results from Lington et al. (1997) 83

Table 4-5. Dose-Response Analysis of Selected Studies Considered for Deriving the Chronic Non-

cancer POD 85

Table 5-1. PESS Evidence Crosswalk for Biological Susceptibility Considerations 93

Table 6-1. Non-cancer HECs and HEDs Used to Estimate Risks 100

LIST OF FIGURES

Figure 1-1. Overview of DINP Human Health Hazard Assessment Approach 17

Figure 2-1. Postulated DINP Metabolism in Humans (Koch and Angerer, 2007) 19

Figure 3-1. Hypothesized Phthalate Syndrome Mode of Action Following Gestational Exposure 29

Figure 4-1. Dose-Response Array of Studies Considered for Deriving the Acute Duration Non-cancer

POD 71

Figure 4-2. Dose-Response Array of Studies Considered for Deriving the Intermediate Duration Non-

cancer POD 78

Figure 4-3. Dose Response Array of Studies Considered for Considered for Deriving the Chronic Non-
cancer POD 84

LIST OF APPENDIX TABLES

TableApx A-l. Summary of Peer Review, Public Comments, and Systematic Review for Existing

Assessments of DINP 116

Table Apx B-l. Summary of Liver Effects Reported in Animal Toxicological Studies Following Short-

Term Exposure to DINP 123

Table Apx B-2. Summary of Liver Effects Reported in Animal Toxicological Studies Following

Subchronic Exposure to DINP 127

Table Apx B-3. Incidence of Selected Non-neoplastic Hepatic Lesions in F344 Rats Exposed to DINP

for 24 Months (Lington et al., 1997) 130

Table Apx B-4. Incidence of Selected Hepatic Lesions in F344 Rats Exposed to DINP in the Diet for 2

Years (Covance Labs, 1998c) 131

Table Apx B-5. Overall Incidence of Selected Tumors in Male and Female Sprague Dawley Rats

Exposed to DINP for 2 Years (Bio/dynamics, 1987) 133

Table Apx B-6. Incidence of Selected Non-neoplastic Lesions in B6C3F1 Mice Exposed to DINP in the

Diet for 2 Years (Covance Labs, 1998b) 134

Table Apx B-7. Summary of Liver Effects Reported in Animal Toxicological Studies Following

Chronic Exposure to DINP 135

Table Apx E-l. Summary of Benchmark Dose Modeling Results from Selected Endpoints in Male and
Female F344 Rats Following 2-Year Exposure to DINP (Lington et al., 1997) 142

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TableApx E-2. Dose-Response Modeling Data for Relative Liver Weight at Terminal Sacrifice in Male

F344 Rats Following 2-Year Exposure to DINP (Lington et al., 1997) 143

Table Apx E-3. Summary of Benchmark Dose Modeling Results for Relative Liver Weight at Terminal
Sacrifice in Male F344 Rats Following 2-Year Exposure to DINP (Constant Variance)

(Lington et al., 1997) 144

Table Apx E-4. Dose-Response Modeling Data for Relative Liver Weight at Terminal Sacrifice in

Female F344 Rats Following 2-Year Exposure to DINP (Lington et al., 1997) 148

Table Apx E-5. Summary of Benchmark Dose Modeling Results for Relative Liver Weight at Terminal
Sacrifice in Female F344 Rats Following 2-Year Exposure to DINP (Non-Constant

Variance) (Lington et al., 1997) 149

Table Apx E-6. Dose-Response Modeling Data for Serum ALT Levels in Male F344 Rats Following 6-

Month Exposure to DINP (Lington et al., 1997) 151

Table Apx E-7. Summary of Benchmark Dose Modeling Results for Serum ALT Levels in Male F344
Rats Following 6-Month Exposure to DINP (Non-constant Variance) (Lington et al.,

1997) 152

Table Apx E-8. Dose-Response Modeling Data for Serum ALT Levels in Male F344 Rats Following

18-Month Exposure to DINP (Lington et al., 1997) 156

Table Apx E-9. Summary of Benchmark Dose Modeling Results for Serum ALT Levels in Male F344
Rats Following 18-Month Exposure to DINP (Non-constant Variance) (Lington et al.,

1997) 157

Table Apx E-10. Dose-Response Modeling Data for Focal Necrosis of the Liver in Male F344 Rats

Following 2-Year Exposure to DINP (Lington et al., 1997) 161

Table Apx E-l 1. Summary of Benchmark Dose Modeling Results for Focal Necrosis of the Liver in

Male F344 Rats Following 2-Year Exposure to DINP (Lington et al., 1997) 162

Table Apx E-12. Dose-Response Modeling Data for Focal Necrosis of the Liver in Female F344 Rats

Following 2-Year Exposure to DINP (Lington et al., 1997) 167

TableApx E-13. Summary of Benchmark Dose Modeling Results for Focal Necrosis of the Liver in

Female F344 Rats Following 2-year Exposure to DINP (Lington et al., 1997) 168

Table Apx E-14. Dose-Response Modeling Data for Spongiosis Hepatis of the Liver in Male F344 Rats

Following 2-Year Exposure to DINP (Lington et al., 1997) 172

TableApx E-15. Summary of Benchmark Dose Modeling Results for Spongiosis Hepatis of the Liver

in Male F344 Rats Following 2-Year Exposure to DINP (Lington et al., 1997) 173

Table Apx E-16. Dose-Response Modeling Data for Sinusoid Ectasia of the Liver in Male F344 Rats

Following 2-Year Exposure to DINP (Lington et al., 1997) 177

Table Apx E-17. Summary of Benchmark Dose Modeling Results for Sinusoid Ectasia of the Liver in
Male F344 Rats Following 2-Year Exposure to DINP (Lington et al., 1997) 178

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ABBREVIATIONS AND ACRONYMS

a2u-globulin

ACE

ADME

AGD

ALP

ALT

AST

AT1R

BMD

BMDL

CASRN

CPSC

DINP

ECB

ECHA

EFSA

eNOS

EPA

F344

GD

GLP

GSH

HEC

HED

IFN

Ig

IL

LABC

LOAEL

LOEL

MNG

MOA

MOE

MWM

NFkB

NICNAS

NOAEL

NOEL

Nrf2

NTP-CERHR

OCSPP

OECD

8-OH-dG

OPPT

PECO

PESS

PND

POD

Alpha 2u-globulin

Angiotensin converting enzyme

Absorption, distribution, metabolism, and excretion

Anogenital distance

Alkaline phosphatase

Alanine aminotransferase

Aspartate aminotransferase

Angiotensin-II type 1 receptor

Benchmark dose

Benchmark dose (lower confidence limit)

Chemical Abstracts Service registry number

Consumer Product Safety Commission (U.S.)

Diisononyl phthalate

European Chemicals Bureau

European Chemicals Agency

European Food Safety Authority

Endothelial nitric oxide synthase

Environmental Protection Agency (U.S.)

Fischer 344 (rat)

Gestation day

Good Laboratory Practice

Glutathione

Human equivalent concentration

Human equivalent dose

Interferon

Immunoglobulin

Interleukin

Levator ani-bulbocavernosus muscle
Lowest-ob served-adverse-effect level
Lowest-ob served-effect level
Multinucleated gonocytes
Mode of action
Margin of exposure
Morris Water Maze
Nuclear factor kappa B

National Industrial Chemicals Notification and Assessment Scheme

No-observed-adverse-effect level

No-observed-effect level

Nuclear factor erythroid 2-related factor 2

National Toxicology Program Center for the Evaluation of Risks to Human Reproduction

Office of Chemical Safety and Pollution Prevention

Organisation for Economic Co-operation and Development

8-Hydroxydeoxyguanosine

Office of Pollution Prevention and Toxics

Population, exposure, comparator, and outcome

Potentially exposed or susceptible subpopulations

Postnatal day

Point of departure

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247

PPARa

Peroxisome proliferator activated receptor alpha

248

ROS

Reactive oxygen species

249

SACC

Science Advisory Committee on Chemicals

250

SD

Sprague-Dawley (rat)

251

TNFa

Tumor necrosis factor alpha

252

TSCA

Toxic Substances Control Act

253

UF

Uncertainty factor

254

U.S.

United States

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ACKNOWLEDGEMENTS

This report was developed by the United States Environmental Protection Agency (U.S. EPA or the
Agency), Office of Chemical Safety and Pollution Prevention (OCSPP), Office of Pollution Prevention
and Toxics (OPPT).

Acknowledgements

The Assessment Team gratefully acknowledges the participation, input, and review comments from
OPPT and OCSPP senior managers and science advisors and assistance from EPA contractors SRC
(Contract Number 68HERH19D0022). Special acknowledgement is given for the contributions of
technical experts from EPA's Office of Research and Development (ORD), including Jeff Gift, and
Geoffrey Collin Peterson for their benchmark dose modeling support.

As part of an intra-agency review, the draft DINP Risk Evaluation was provided to multiple EPA
Program Offices for review. Comments were submitted by Comments were submitted by EPA's Office
of Air and Radiation (OAR.), Office of Children's Health Protection (OCHP), Office of General Counsel
(OGC), ORD, and Office of Water (OW).

Docket

Supporting information can be found in the public docket, Docket ID (EPA.-H.Q-OPPT-2018-0436).
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.

Authors: John Allran, Anthony Luz, Ashley Peppriell, Myles Hodge, Sailesh Surapureddi, Christelene
Horton, Brandall Ingle-Carlson, Collin Beachum (Branch Chief)

Contributors: Susanna Wegner, Abhilash Sasidharan

Technical Support: Mark Gibson, Hillary Hollinger

This report was reviewed and cleared by OPPT and OCSPP leadership.

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SUMMARY

This technical support document for diisononyl phthalate (DINP) summarizes the non-cancer hazards
associated with exposure to DINP and identifies the proposed points of departure (PODs) to be used to
estimate risks from DINP exposures in the draft risk evaluation of DINP. EPA summarizes the cancer
hazards associated with exposure to DINP in a separate technical support document, the Draft Cancer
Human Health Hazard Assessment for Diisononyl Phthalate (DINP) ( 2024a).

EPA identified developmental, liver, and kidney toxicity as the most sensitive and robust non-cancer
hazards associated with oral exposure to DINP in experimental animal models (Section 3.1 through 3.3).
Liver, kidney, and developmental toxicity were also identified as the most sensitive and robust non-
cancer effects following oral exposure to DINP by existing assessments of DINP—including those by
the U.S. Consumer Product Safety Commission ( 3. 2014). Health Canada (ECCC/HC. 2020).

European Chemicals Agency (ECHA. 2013b). European Food Safety Authority ( ), and the

Australian National Industrial Chemicals Notification and Assessment Scheme (NICNA.S. 2015). EPA
is proposing a point of departure (POD) of 49 mg/kg-day (human equivalent dose [HED] of 12 mg/kg-
day) to estimate non-cancer risks from oral exposure to DINP for acute and intermediate durations of
exposure in the draft risk evaluation of DINP. The proposed POD was derived through meta-regression
analysis and benchmark dose (BMD) modeling of fetal testicular testosterone data from two prenatal
exposure studies of rats by the National Academies of Sciences, Engineering, and Medicine (NA.SEM.
2017). The POD of 49 mg/kg-day is the 95 percent lower confidence limit of the BMD associated with a
benchmark response (BMR) of 5 percent.

As discussed further in Sections 4.1.1 and 4.1.2, several additional acute and intermediate duration
studies of DINP provide similar, although less-sensitive, candidate PODs, which further support EPA's
proposal to use the selected POD of 12 mg/kg-day for decreased fetal testicular testosterone production.
The Agency has performed % body weight scaling to yield the HED and is applying the animal to
human extrapolation factor {i.e., interspecies extrapolation; UFa) of 3x and an within human variability
extrapolation factor {i.e., intraspecies extrapolation; UFh) of 10x. Thus, a total uncertainty factor (UF)
of 30x is applied for use as the benchmark margin of exposure (MOE). Based on the strengths,
limitations, and uncertainties discussed Section 4.2.1, EPA has robust overall confidence in the
proposed POD based on fetal testicular testosterone for use in characterizing risk from exposure
to DINP for acute and intermediate exposure scenarios. For purposes of assessing non-cancer risks,
the selected POD is considered most applicable to women of reproductive age, pregnant women, and
infants. Use of this POD to assess risk for other age groups {e.g., older children and adult males) is
conservative.

EPA is proposing a no-observed-adverse-effect level (NOAEL) of 15 mg/kg-day (HED of 3.5 mg/kg-
day) from a high quality 2-year study of rats based on liver toxicity to estimate non-cancer risks from
oral exposure to DINP for chronic durations of exposure in the draft risk evaluation of DINP. More
specifically, liver toxicity in the key study (Lington et ai. 1997; Bio/dynamics. 1986) was characterized
by increased liver weight, increased serum alanine aminotransferase (ALT), aspartate aminotransferase
(AST), alkaline phosphatase (ALP), and histopathological findings {e.g., focal necrosis, spongiosis
hepatis). EPA considers the observed liver effects to be adverse and relevant for extrapolating human
risk from chronic exposures ( 1002a). As discussed further in Sections 4.1.1 through 4.1.3,

several additional studies of DINP provide similar, although less-sensitive, candidate PODs, which
further support EPA's decision to use the selected POD of 3.5 mg/kg-day for chronic exposures. The
Agency has performed 3/4 body weight scaling to yield the HED and is applying the animal to human
extrapolation factor {i.e., interspecies extrapolation; UFA) of 3/ and an within human variability
extrapolation factor {i.e., intraspecies extrapolation; UFH) of 10x. Thus, a total UF of 30x is applied for

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use as the benchmark MOE. Overall, based on the strengths, limitations, and uncertainties discussed in
Section 4.2.2, EPA has robust overall confidence in the proposed POD based on hepatic outcomes
for use in characterizing risk from exposure to DINP for chronic exposure scenarios.

No data were available for the dermal or inhalation routes that were suitable for deriving route-specific
PODs. Therefore, EPA used the acute/intermediate and chronic oral PODs to evaluate risks from dermal
exposure to DINP. Differences in absorption will be accounted for in dermal exposure estimates in the
draft risk evaluation for DINP. For the inhalation route, EPA extrapolated the oral HED to an inhalation
human equivalent concentration (HEC) using a human body weight and breathing rate relevant to a
continuous exposure of an individual at rest (	94). The oral HED and inhalation HEC values

selected by EPA to estimate non-cancer risk from acute/intermediate and chronic exposure to DINP in
the draft risk evaluation of DINP are summarized in Table ES-1 and Section 6.

EPA is soliciting comments from the Science Advisory Committee on Chemicals (SACC) on charge
questions and comments from the public for the upcoming SACC meeting.

Table ES-1. Non-cancer HECs and HEDs Uset

to Estimate

iisks

Exposure
Scenario

Target
Organ

System

Species
(Sex)

Duration

POD

(mg/kg-
day)

Effect

HEC
(mg/m3)
[ppm]

HED

(mg/
kg-day)

Benchmark
MOE

Reference

Acute and
Intermediate

Development

Rat

5 to 14 days

throughout

gestation

BMDLs
= 49"

i fetal testicular
testosterone

63
[3.7]

12

UFa= 3
UFh=10
Total UF=30

(NASEM.
2017)

Chronic

Liver

Rat

2 years

NOAEL
= 15

t liver weight,
t serum
chemistry,
histopathology''

19

[1.1]

3.5

UFa= 3
UFh=10
Total UF=30

(Lington et
al. 1997;
Bio/dynamics
6)c

HEC = human equivalent concentration; HED = human equivalent dose; POD = point of departure; MOE = margin of exposure;
BMDL = benchmark dose lower limit; UF = uncertainty factor; NOAEL = no observable adverse effect level

" The BMDL^ was derived by NASEM (2017) through meta-regression and BMD modeling of fetal testicular testosterone data from
two studies of DINP with rats (Boberg et al, 2011; Hannas et al, 2011). R code supporting NASEM's meta-regression and BMD
analysis of DINP is publicly available through GitlTiib.

4 Liver toxicity included increased relative liver weight, increased serum chemistry {i.e., AST, ALT, ALP), and histopathologic
findings (e.g., focal necrosis, spongiosis hepatis) in F344 rats following 2 years of dietary exposure to DINP (Lington et al, 1997;
Bio/dynamics, 1986).

c The Lington study presents a portion of the data from a larger good laboratory practice (GLP)-certified study by Bio/dynamics

(1986).

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1 INTRODUCTION

On May 24, 2019, EPA received a request, pursuant to 40 CFR 702.37, from ExxonMobil Chemical
Company, through the American Chemistry Council's High Phthalates Panel (ACC HPP. 2019). to
conduct a risk evaluation for diisononyl phthalate (DINP) (CASRNs 28553-12-0 and 68515-48-0)
(Docket ID: EPA-HQ-OPPT-2018-0436). EPA determined that these two CASRNs should be treated as
a category of chemical substances as defined in 15 U.S.C § 2625(c). On August 19, 2019, EPA opened a
45-day public comment period to gather information relevant to the requested risk evaluation. EPA
reviewed the request (along with additional information received during the public comment period) and
assessed whether the circumstances identified in the request constitute conditions of use under 40 CFR
702.33, and whether those conditions of use warrant inclusion within the scope of a risk evaluation for
DINP. EPA determined that the request meets the applicable regulatory criteria and requirements, as
prescribed under 40 CFR 702.37. The Agency granted the request on December 2, 2019, and published
the draft and final scope documents for DINP in August 2020 and 2021, respectively (I v << \

2020).

Following publication of the final scope document, one of the next steps in the TSCA risk evaluation
process is to identify and characterize the human health hazards of DINP and conduct a dose-response
assessment to determine the toxicity values to be used to estimate risks from DINP exposures. This
technical support document for DINP summarizes the non-cancer hazards associated with exposure to
DINP and proposes toxicity values to be used to estimate non-cancer risks from DINP exposures. EPA
summarizes the cancer hazards associated with exposure to DINP in a separate technical support
document, the Draft Cancer Human Health Hazard Assessment for Diisononyl Phthalate (DINP) (U.S.
EPA. 2024aY

Over the past several decades the human health effects of DINP have been reviewed by several
regulatory and authoritative agencies, including the: U.S. Consumer Product Safety Commission (U.S.
CPSC); Health Canada; U.S. National Toxicology Program Center for the Evaluation of Risks to Human
Reproduction (NTP-CERHR); European Chemicals Bureau (ECB); European Chemicals Agency
(ECHA); European Food Safety Authority (EFSA); the Australian National Industrial Chemicals
Notification and Assessment Scheme (NICNAS); The National Academies of Sciences, Engineering,
and Medicine (NASEM); and U.S EPA. EPA relied on information published in existing assessments by
these regulatory and authoritative agencies as a starting point for its human health hazard assessment of
DINP. Additionally, EPA considered new literature published since the most recent existing assessments
of DINP to determine if newer information might support the identification of new human health
hazards or lower PODs for use in estimating human risk. EPA's process for considering and
incorporating new DINP literature is described in the Draft Risk Evaluation for Diisononyl Phthalate
(DINP) - Systematic Review Protocol (also referred to as the Draft DINP Systematic Review Protocol)
(I E024b). EPA's approach and methodology for identifying and using human epidemiologic

data and experimental laboratory animal data is described in Section 1.1.

1.1 Human Epidemiologic Data: Approach and Preliminary Conclusions

To identify and integrate human epidemiologic data into the draft DINP Risk Evaluation, EPA first
reviewed existing assessments of DINP conducted by regulatory and authoritative agencies, as well as
several systematic reviews of epidemiologic studies of DINP published by U.S. EPA Integrated Risk
Information System (IRIS) program. Existing assessments reviewed by EPA are listed below. As
described further in Appendix A, most of these assessments have been subjected to peer-review and/or
public comment periods and have employed formal systematic review protocols.

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• Supporting documentation: Evaluation of epidemiologic studies on phthalate compounds and
their metabolites for hormonal effects, growth and development and reproductive parameters
(Health Canada. 2018b);

• Supporting documentation: Evaluation of epidemiologic studies on phthalate compounds and
their metabolites for effects on behaviour and neurodevelopment, allergies, cardiovascular
function, oxidative stress, breast cancer, obesity, and metabolic disorders (Health Canada.
2018a):

• Phthalate exposure and male reproductive outcomes: A systematic review of the human
epidemiological evidence (Radke et ah. 2018);

• Phthalate exposure andfemale reproductive and developmental outcomes: A systematic review
of the human epidemiological evidence (Radke et ah. 2019b);

• Phthalate exposure and metabolic effects: A systematic review of the human epidemiological
evidence (Radke et ah. 2019a); and

• Phthalate exposure and neurodevelopment: A systematic review and meta-analysis of human
epidemiological evidence (Radke et ah. 2020a).

Next, EPA sought to identify new population, exposure, comparator, and outcome (PECO)-relevant
literature published since the most recent existing assessments) of DINP by applying a literature
inclusion cutoff date. For DINP, the applied cutoff date was based on existing assessments of
epidemiologic studies of phthalates by Health Canada (2018a. b), which included literature up to
January 2018. The Health Canada (2018a. b) epidemiologic evaluations were considered the most
appropriate existing assessments for setting a literature inclusion cutoff date because those assessments
provided the most robust and recent evaluation of human epidemiologic data for DINP. Health Canada
evaluated epidemiologic study quality using the Downs and Black method (Downs i ck. 1998) and
reviewed the database of epidemiologic studies for consistency, temporality, exposure-response,
strength of association, and database quality to determine the level of evidence for association between
urinary DINP metabolites and health outcomes. New PECO-relevant literature published between 2018
to 2019 was identified through the literature search conducted by EPA in 2019, as well as references
published between 2018 to 2023 that were submitted with public comments to the DINP Docket (
HQ-QPPT-2018-0436). were evaluated for data quality and extracted consistent with EPA's Draft
Systematic Review Protocol Supporting TSCA Risk Evaluations for Chemical Substances (U.S. EPA.
2021a). Data quality evaluations for new studies reviewed by EPA are provided in the Draft Risk
Evaluation for Diisononyl Phthalate (DINP) - Systematic Review Supplemental File: Data Quality
Evaluation Information for Human Health Hazard Epidemiology ( 24d).

As described further in the Draft DINP Systematic Review Protocol ( 2024b). EPA considers

phthalate metabolite concentrations in urine to be an appropriate proxy of exposure from all sources—
including exposure through ingestion, dermal absorption, and inhalation. As described in thq Application
of US EPA IRIS systematic review methods to the health effects ofphthalates: Lessons learned and path
forward (Radke et ah. 2020b). from EPA's IRIS program, the "problem with measuring phthalate
metabolites in blood and other tissues is the potential for contamination from outside sources (Calafat et
ah. 2015). Phthalate diesters present from exogenous contamination can be metabolized to the
monoester metabolites by enzymes present in blood and other tissues, but not urine." Therefore, EPA
has focused its epidemiologic evaluation on urinary biomonitoring data; new epidemiologic studies that
examined DINP metabolites in matrices other than urine were considered supplemental and not
evaluated for data quality.

The Agency is proposing to use epidemiologic studies of DINP qualitatively; this proposal is consistent
with Health Canada, U.S. CPSC, ECHA, EFSA, and Australia NICNAS. The Agency did not use

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epidemiology studies quantitatively for dose-response assessment, primarily due to uncertainty
associated with exposure characterization. Primary sources of uncertainty include the source(s) of
exposure; timing of exposure assessment that may not be reflective of exposure during outcome
measurements; and use of spot-urine samples, which due to rapid elimination kinetics may not be
representative of average urinary concentrations that are collected over a longer term or calculated using
pooled samples. Additional uncertainty results from co-exposure to mixtures of multiple phthalates that
may confound results for the majority of epidemiologic studies, which examine one phthalate and one
exposure period at a time such that they are treated as if they occur in isolation (Shin et at.. 2019;
Aylward et at.. 2016). Conclusions from Health Canada (2018a. b) and U.S. EPA systematic review
articles (Radke et at.. 2020a; Radke et at.. 2019b; Radke et at.. 2019a; Radke et at.. 2018) regarding the
level of evidence for association between urinary DINP metabolites and each health outcome were
reviewed by EPA and used as a starting point for its human health hazard assessment. The Agency also
evaluated and summarized new epidemiologic studies identified by EPA's systematic review process to
use qualitatively during evidence integration to inform hazard identification and the weight of scientific
evidence (Shin et at.. 2019; Aylward et at.. 2016).

1.2 Laboratory Animal Findings: Summary of Existing Assessments from
Other Regulatory Organizations

The human health hazards of DINP have been evaluated in existing assessments by U.S. CPSC (2014.
2010). Health Canada (ECCC/HC. 2020; EC/HC. 2015). NTP-CERHR (2003). ECB (2003). ECHA
(201 ' 10 g/kg), dermal
(LD50 > 3g/kg), or inhalation (LC50 > 4.4 mg/L) routes of exposure. DINP only resulted in slight
irritation in primary skin and eye irritation studies in rabbits. Dermal sensitization studies with rodent
models (e.g., Buehler tests) indicate that DINP is not a dermal sensitizer. EPA identified no new
information that would change these conclusions; therefore, these hazards are not discussed further in
this draft hazard assessment.

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480 Table 1-1. Summary of DINP Non-cancer POPs Selected for Use by other Regulatory Organizations

Brief Study Description

TSCA Data
Quality

NOAEL/
LOAEL
(mg/kg-day)

Critical Effect

U.S. CPSC

(MM)

u

u ^

r \ ©J
W fSl

w

03 3

Uh ®

w d

NICNAS

(2012)

^ 2|
S3 ^
U ©1

y (S|J

Male and female F344 rats (110/sex/dose) fed diets containing
0, 300, 3,000, 6,000 ppm DINP (CASRN 68515-48-0) for two
years (equivalent to 15, 152, 307 mg/kg-day for males; 18,
184, 375 mg/kg-day for females) (GLP-compliant, non-
guideline studv) (Lington et aL 1997; Bio/dynamics, 1986)

High

15/152

t in absolute and relative
liver and kidney weight
with increase in
histopathological changes
(e.g., spongiosis hepatis)
and other signs of
hepatotoxicity

y/a

yb

•/c

^/d

•/e

Male and female F344 rats (70-85/sex/dose) administered 0,
500, 1500, 6000, 12,000 ppm in the diet for 104 weeks
(equivalent to 29, 88, 358, 733 mg/kg-day in males; 36, 108,
4422, 885 mg/kg-day in females) (GLP-compliant, adhered to
40 CFR Part 798 (S 798.330)) (Covance Labs. 1998c)

High

88/ 358

t Liver and kidney weight,
biochemical changes (f
serum ALT, AST), and
histopathological findings







^/d



Pregnant female SD rats (6/dose) gavaged with 0, 10, 100,
500, 1,000 mg/kg-day DINP on GDs 12-21. Dams were
allowed to give birth naturally, and then dams and pups were
sacrificed (non-euideline studv) (Li et al.. 2015)

Medium

10 (LOEL)/
100

(LOAEL)

t MNGs and Leydig cell
clusters/ aggregation











Hershberger assay: young (6-week old) castrated male SD rats
treated with testosterone propionate (0.4 mg/kg-day) were
gavaged with 0, 20, 100, 500 mg/kg-day DINP for 10 days and
then sacrificed (non-euideline studv) (Lee and Koo. 2007)

Medium

100/500

i absolute seminal vesicle
and LABC weights











Pregnant SD rats (8/dose) gavaged with 0, 50, 250, 500 mg/kg-
dav DIINP on GDs 12-19 (non-euideline studv) (Clewell et aL.

2013a)

High

50/ 250

Transient reduced fetal
testosterone level and
histopathological changes
(MNGs)

y/a



•/c

Jd

¦/e

Pregnant Wistar rats (16/dose) gavaged with 0, 300, 600, 750,
900 mg/kg-day DINP from GD 7 to PND 17 (non-guideline
studv) (Bobere et aL. 2011)

Medium

300/600

t Nipple retention

y/a





Jd



Pregnant Harlan SD rats (5-9/group) gavaged with 0, 500, 750,
1000, 1500 mg/kg-day DINP from GD 14 to 18 (non-guideline
studv) (Hannas et aL. 2011)

Medium

-/500

i fetal testicular
testosterone production

y/a





Jd



Pregnant SD rats (20-24/group) fed diets containing 0, 760,
3800, 11,400 ppm DINP from GD 12 to PND 14 (target doses:

Medium

250/750

i male pup AGD on PND
14

y/a









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Brief Study Description

TSCA Data
Quality

NOAEL/
LOAEL
(mg/kg-day)

Critical Effect

U.S. CPSC
( )

ECCC/HC

(2020)

¦^5 Q

Ifl vH|
U, O

3 GJ

t/l

< f5j

LJ ©1

a d

ECHA

(2013b)

0, 50, 250, 750 mg/kg-day; received doses: 56, 288, 720,
me/ke-dav on CDs 13-20) (non-euideline studv) (Clewell et
al.. 20Obi



50/250

i male pup body weight on
PND 14







¦/e



Male and female SD rats fed diets containing 0, 0.2, 0.4, 0.8%
(Received doses in units of mg/kg-day shown in Table 3-7)
DINP 10 weeks prior to mating, and throughout mating,
gestation and lactation continuously for two generations (GLP-
coniDliant. adhered to 40 CFR 798 (§ 798.4700)) (Waterman et
al.. 2000; Exxon Biomedical 1996b")

High

-/114-395

| F1 and F2 pup body
weight on PND7 and 21







•/e



CPSC = Consumer Product Safety Commission (U.S.); ECCC/HC = Environment and Climate Change Canada/Health Canada; ECHA = European Chemicals Agency;
EFSA = European Food Safety Authority; NICNAS = Australia National Industrial Chemicals Notification and Assessment Scheme; ALT = Alanine
aminotransferase; AGD = Anogenital distance; AST = Aspartate aminotransferase; LABC = Levator ani/bulbocavernosus; MNG = Multinucleated gonocytes; PND =
Post-natal day

" NOAELs from antiandroeenic endooints (i.e., nioDlc retention, fetal testosterone dinduction. MNGs) across several studies ((Clewell et al.. 2013a; Clewell et al..
2013b; Bobere et al. 2011; Hannas et al.. 2011)) were used bv U.S. CPSC to assien a NOAEL for developmental toxicity of 50 me/ke-dav based on antiandroeenic
endooints (see d. 98 of (U.S. CPSC. 2014)).

h NOAELs from Lineton et al. (1997) and Li et al. (2015) were used bv Health Canada to calculate MOEs for individual DINP exposure scenarios (see Table 9-58 of
(ECCC/HC. 2020)). NOAELs from Li et al. and Lee and Koo (2007) were used to estimate hazard quotients for DINP as part of the cumulative risk assessment (see
Tables F-5 throuehF-9 in (ECCC/HC. 2020)).

c NOAEL from Lington et al. (1997) was used by EFSA to derive a stand-alone tolerable daily intake (TDI) for DINP based on liver and kidney effects, while the
NOAEL from Clewell et al. (2013a) was used to establish a erouo-TDI for several nhthalates (e.s., DEHP. DBP. BBP. and DINP) based on developmental effects
related to a plausible common mechanism (i.e., reduced fetal testosterone).

¦'NIC AS derived a NOAEL for svstemic effects (liver and kidney toxicity) based on the results from two 2-year dietary studies of F344 rats (Covance Labs. 1998c;
Lineton et al. 1997). which were similar in desien and collectively suDDorted a NOAEL of 88 me/ke-dav. Similarly. NICNAS derived a NOAEL of 50 me/ke-dav for
fertility-related effects (i.e., reduced fetal testosterone) based on results from three studies (Clewell et al. 2013a; Bobere et al. 2011; Hannas et al. 2011) and a
NOAEL of 50 me/ke-dav for developmental effects (i.e., reduced dud weieht) based on results from two studies (Clewell et al. 2013b; Waterman et al.. 2000) (see
Table 7.1 in (NICNAS. 2012)).

• NOAELs used bv ECHA to calculate derived no effect levels (DNELs) (see Section 4.4.11.2 of (ECHA. 2013b)).

' Studies evaluated for data duality consistent with the Draft DINP Systematic Review Protocol (U.S. EPA. 2024b) and EPA's Draft Systematic Review Protocol (U.S.

EPA. 2021a).

481

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1.3 Laboratory Animal Data: Approach and Methodology

1.3.1 Approach to Identifying and Integrating Laboratory Animal Data

Figure 1-1 provides an overview of EPA's approach to identifying and integrating laboratory animal
data into the draft DINP Risk Evaluation. EPA first reviewed existing assessments of DINP conducted
by various regulatory and authoritative agencies. Existing assessments reviewed by EPA are listed
below. The purpose of this review was to identify sensitive and human relevant hazard outcomes
associated with exposure to DINP, and identify key studies used to establish PODs for estimating human
risk. As described further in Appendix A, most of these assessments have been subjected to external
peer-review and/or public comment periods but have not employed formal systematic review protocols.

•	Toxicity review of DiisononylPhthalate (DINP) (U.S. CPSC. 2010);

•	Chronic Hazard Advisory Panel on phthalates and phthalate alternatives (U.S. CPSC. 2014);

•	State of the science report: Phthalate substance grouping 1,2-Benzenedicarboxylic acid,
diisononyl ester; 1,2-Benzenedicarboxylic acid, di-C8-10-branchedalkyl esters, C9-rich
(Diisononyl Phthalate; DINP). Chemical Abstracts Service Registry Numbers: 28553-12-0 and
68515-48-0 (EC/HC. 2015);

•	Supporting documentation: Carcinogenicity ofphthalates - mode of action and human relevance
(Health Canada. 2015);

•	Screening assessment - Phthalate substance grouping (ECCC/HC. 2020);

•	NTP-CERHR monograph on the potential human reproductive and developmental effects of di-
isononyl phthalate (DINP) (NTP-CERHR. 2003);

•	European union risk assessment report: DINP (ECB. 2003);

•	Evaluation of new scientific evidence concerning DINP and DIDP in relation to entry 52 of
Annex XVII to REACH Regulation (EC) No 1907/2006 (EC	));

•	Committee for Risk Assessment (RAC) Opinion on the ECHA 's draft review report on
"Evaluation of new scientific evidence concerning DINP and DIDP in relation to entry 52 of
AnnexXVII to Regulation (EC) No 1907/2006 (REACH) " ECHA/RAC/A77-0-0000001412-86-
10/F (EC	a);

•	Committee for Risk Assessment (RAC) Opinion proposing harmonised classification and
labelling at EUlevel of 1,2-Benzenedicarboxylic acid, di-C8-10-branchedalkylesters, C9- rich;
[1] di- "isononyl"phthalate; [2] [DINP] (ECHA.. 2018);

•	Opinion of the scientific panel on food additives, flavourings, processing aids and materials in
contact with food (AFC) on a request from the commission related to di-isononylphthalate
(DINP) for use in food contact materials. (EFSA. 2005);

•	Update of the risk assessment of di-butylphthalate (DBP), butyl-benzyl-phthalate (BBP), bis(2-
ethylhexyl)phthalate (DEHP), di-isononylphthalate (DINP) and di-isodecylphthalate (DIDP) for
use in food contact materials (EFSA. );

•	Priority existing chemical assessment report no. 35: Diisononyl phthalate (NICNAS. 2012);

•	Application of systematic review methods in an overall strategy for evaluating low-dose toxicity
from endocrine active chemicals (NASEM. 2017);

•	Revised technical review of diisononyl phthalate (	305b); and

•	Technical review of diisononyl phthalate (Final assessment) (	1023c).

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Figure 1-1. Overview of DINP Human Health Hazard Assessment Approach

11 Any study that was considered for dose-response assessment, not necessarily limited to the study used for POD
selection.

h Extracted information includes PECO relevance, species, exposure route and type, study duration, number of
dose groups, target organ/systems evaluated, study-wide LOEL, and PESS categories.

EPA used the 2015 Health Canada assessment (EC/HC. 2015) as the key starting point for this draft
document. The Health Canada assessment included scientific literature up to August 2014, and
considered a range of human health hazards (e.g., developmental and reproductive toxicity, systemic
toxicity to major organ systems, genotoxicity, carcinogenicity) across all durations (i.e., acute, short-
term, subchronic, chronic) and routes of exposure (i.e., oral, dermal, inhalation). The EFSA (2019)
assessment was limited in scope (i.e., considered a limited range of human health hazards) and was not
subject to external peer-review, whereas the Health Canada (ECCC/HC. 2020) assessment did not
provide a specific literature inclusion cutoff date and the U.S. EPA (2023c) assessment did not describe
its approach to identifying literature. Therefore, EPA considered literature published between 2014 to
2019 further as shown in Figure 1-1. EPA first screened titles and abstracts and then full texts for
relevancy using PECO screening criteria described in the Draft DINP Systematic Review Protocol (U.S.
EPA. 2024b). EPA then identified PECO-relevant literature published since the most recent and
comprehensive existing assessment of DINP by applying a literature inclusion cutoff date from this
assessment.

Next, EPA reviewed new studies published between 2014 and 2019 and extracted key study information
as described in the Draft DINP Systematic Review Protocol (U.S. EPA. 2024b). Extracted information
included: PECO relevance; species tested; exposure route, method, and duration of exposure; number of
dose groups; target organ/systems evaluated; information related to potentially exposed or susceptible
subpopulations (PESS); and the study-wide lowest-observable-effect level (LOEL) (Figure 1-1).

New information for DINP was primarily limited to oral exposure studies, and study LOELs were
converted to HEDs by scaling allometrically across species using the 3A power of body weight (BW3/4)
for oral data, which is the approach recommended by U.S. EPA when physiologically based
pharmacokinetic models or other information to support a chemical-specific quantitative extrapolation is
absent (U.S. EPA. 2011b). EPA's use of allometric body weight scaling is described further in Appendix
F. EPA did not conduct data quality evaluations for studies with HEDs based on LOELs that were
greater than an order of magnitude of the lowest HED based on the lowest-observable-adverse-effect
level (LOAEL) across existing assessments because they were not considered sensitive for subsequent
POD selection. However, these studies were still reviewed and integrated into the hazard identification
process. Studies with HEDs for LOELs within an order of magnitude of the lowest LOAEL-based HED

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identified across existing assessments were considered sensitive and potentially relevant for POD
selection. These studies were further reviewed by EPA to determine if they provide information that
supports a human health hazard not identified in previous assessments or to determine if they contain
sufficient dose-response information to support a potentially lower POD than identified in existing
assessments of DINP.

Data quality evaluations for DINP animal toxicity studies reviewed by EPA are provided in the Draft
Risk Evaluation for Diisononyl Phthalate (DINP) - Systematic Review Supplemental File: Data Quality
Evaluation Information for Human Health Hazard Animal Toxicology ( 324c).

1,3.2 New Literature Identified and Hazards of Focus for DINP

As described in Section 1.3.1, EPA reviewed literature published between 2014 to 2019 for new
information on sensitive human health hazards not previously identified in existing assessments,
including information that may indicate a more sensitive POD. As described further in the Draft DINP
Systematic Review Protocol ( 2024b). EPA identified 13 new PECO-relevant studies that

provided information pertaining to 5 primary hazard outcomes, including reproduction/development,
neurological, cardiovascular, immune system, and the musculoskeletal system. Further details regarding
EPA's handling of this new information are provided below.

• Reproductive/Developmental. EPA identified six new studies evaluating reproductive/
developmental outcome (Chiang and Flaws. 2019; Neief et ai. 2019; Neier et ai. 2018; Setti
Ahmed et ai. 2018; Li et at . v edha et ai. I ) These new studies of DINP are discussed
further in Section 3.1.

• Neurotoxicity. EPA identified four new studies evaluating neurological outcomes, including two
that evaluate neurobehavioral outcomes (Ma et ai. 201 ^eng. 2015) and two that evaluate brain
weight (Neier et ai. 2018; Setti Ahmed et ai. 2018). Neurotoxicity is a new health outcome that
has not been seen in previous studies of DINP or discussed in existing assessments of DINP. The
neurologic effects of DINP are discussed further in Section 3.4.

• Cardiovascular. EPA identified one new study evaluating cardiovascular outcomes (Dene et ai.
2019). Results from Deng et ai provide evidence of a new health hazard associated with
exposure to DINP that has not been previously seen in studies of DINP. The cardiovascular
effects of DINP are discussed further in Section 3.5.

• Immune System EPA identified three new studies evaluating immune system effects (Kane et
ai. 2016; Wu et ai. 2015; Sadakane et ai. 2014). Results from these studies indicate that DINP
can have adjuvant-like effects on immune responses. The immune adjuvant effects of DINP are
discussed further in Section 3.6.

• Musculoskeletal. EPA identified one new study evaluating effects on the musculoskeletal system
(Hwane et ai. 2017). Results from Hwang et ai provide evidence of a new health hazard
associated with exposure to DINP that has not been previously seen in studies of DINP.
Musculoskeletal effects of DINP are discussed further in Section 3.7.

Based on information provided in existing assessments of DINP for liver, kidney, and developmental
effects in combination with new information identified by EPA that encompasses additional hazard
outcomes, the Agency focused its non-cancer human health hazard assessment on developmental
toxicity (Section 3.1); liver toxicity (Section 3.2); kidney toxicity (Section 3.3); neurotoxicity (Section
3.4); cardiovascular health effects (Section 3.5); immune system toxicity (Section 3.6); and
musculoskeletal toxicity (Section 3.7).

Genotoxicity and carcinogenicity data for DINP are summarized in EPA's Draft Cancer Human Health
Hazard Assessment for Diisononyl Phthalate (DINP) (U.S. EPA. 2024a).

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2 TOXICOKINETICS

2A Oral Route

Three experimental animal studies are available that provide useful data in evaluating absorption,
distribution, metabolism, and excretion (ADME) of DINP for the oral route. DINP is shown to be
predominantly metabolized in the liver in rodents, and urinary excretion is the primary route of
elimination for metabolites. In one of the few studies designed to investigate the metabolism of
phthalates in humans, a male volunteer (aged 63) was given a single oral dose of 1.27 mg of deuterium-
labeled DINP/kg bodyweight. DINP was found to be rapidly eliminated in a manner similar to rats
(Koch and Angerer. 2007). The postulated metabolic pathway of DINP in humans is shown in Figure
2-1. Results indicated that approximately 44 percent of the administered dose was recovered in urine
over 48 hours in the form of the following metabolites: (1) 20.2 percent as OFI-MINP (MHINP; based
on measured standard of 70H-MMeOP); (2) 10.7 percent as carboxy-MINP (MCiOP; based on
measured standard of 7-carboxy-MMeHP); (3) 10.6 percent as oxo-MINP (MOINP; based on measured
standard of 7oxo -MMeOP); and (4) 2.2 percent as MINP (Koch and Armerer. 2007).

D4-7o*o-MMeOP

R:

D4-OINP:
D+-MMCOP:
D4-70H- MMcOP:
D4-7o*o-MMcOP:
D4-7c»bo*v-MMcl IP:

iMvrnms I ;ilk) Ichain
IM-di-iso-nonylphtlu laic
l>4-mv»inM 4-rocthy loc t> I )pt> Aalutc

4-methyl-7-hydroxyociyI )phlhal-(4-mcthyl*7-oxoocty 1 )phlhaJatc
l>4 -mono «< 4- mcthy I• 7-oxy hepty 1 )ph dul.itc

Figure 2-1. Postulated DINP Metabolism in Humans (Koch and Angerer, 2007)

In Anderson et al, 20 volunteers were given two doses of DINP (0.78 mg and 7.3 mg) to examine its
metabolism and excretion. More than 33 percent of the labelled DINP was found as metabolites in urine
after 48 hours (Anderson et al., 2011).

Several studies investigated the toxicokinetics of DINP in animals. McKee et al. (2002) examined the
ADME of DINP in male and female F344 rats. Rats were administered single oral doses of 50 or 500
mg/kg 114C]DINP, and data on tissue distribution indicated that 2 to 4 hours following administration,

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the highest levels of radioactivity were found to be in the blood, liver, and kidneys. The distribution of
radiolabeled DINP to other tissues after 7 days of exposure, was gastrointenstinal (GI) tract (0.097
percent), fat (0.053 percent), muscle (0.024 percent), and other organs (<0.009 percent). No differences
in excretion were apparent in either sex at either dose. In the single dose studies, 50 percent of the
radioactivity was recovered in the urine and the remainder in the feces at the low dose; whereas at the
high dose, 35 to 40 percent of the radioactivity was excreted in the urine and the remainder in the feces,
suggesting an inverse relationship between dose level and absorption. In repeated dose studies, rats were
administered 50, 150, and 500 mg/kg-day [14C]DINP for 5 days, and excretion was evaluated (McKee et
at.. 2002). In the repeated dose studies, about 60 percent of the administered dose was excreted at all
doses, suggesting an elevation of esterase activity and more rapid conversion to monoester following
repeated treatment. The elimination (half-life) of absorbed [14C]DINP was about 7 hours.

In another study by Clewell et al. (2013a). pregnant Sprague-Dawley (SD) rats received 50, 250, and
750 mg/kg-day of DINP from gestation day (GD) 12 to 19 via oral gavage. The percentage of DINP
absorbed following oral exposure was lower at the higher doses of 750 mg/kg-day compared to the 250
mg/kg-day group. Additionally, Clewell et al. (2013a) characterized the metabolite disposition of DINP
in the fetus and demonstrated that MINP and its oxidative metabolites along with its glucuronidated
form (MINP-Gluc) were all present in the fetal plasma, testes, and amniotic fluid. MINP-Gluc was
present at higher concentrations in the fetal plasma than the maternal plasma (in contradiction with what
was observed with the other metabolites), indicating potential placental transfer of MINP-Gluc, or, more
likely, that conjugation could occur in the fetus by phase II detoxification enzyme systems. Because
these metabolites were localized in maternal plasma and MINP was present at similar concentrations as
MCiOP, it was suggested that (1) urinary clearance of both MINP and MINP-Gluc is limited, and (2)
these metabolites were poor predictors of plasma and tissue disposition for DINP.

A summary of different metabolites found in human and rat urine after oral administration of DINP is
presented in Table 2-1.

Table 2-1. Absorption and Excretion Summary of DINP

Species

Dose

Source

Absorption

Reference

Human

1.28 mg/kg

Urine

44% over 48 hours

(Koch a terer.

2007)

Human

0.78 and 7.3 mg/kg

Urine

33 ± 6.4% over 48 hours

(Anderson et al..
201 I)

Rat

50 mg/kg
500 mg/kg
50-500 mg/kg

Urine
Urine
Estimated
urine + bile

49% over 72 hours
39% over 72 hours
75% over 72 hours

(McKee et al.. 2002)



50, 150, or 500
mg/kg-day for 5
days

Urine
Estimated
urine + bile

56-62% over 24 hours, 62-64%
over 72 hours
90% over 72 hours



Rat
(non-
pregnant)

Single dose of 300
mg/kg

Urine

Mono(carboxy-isooctyl)phthalate

(MciOP) 82%

Other metabolites 18%

(Silva et al.. 2006)

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Silva et al. (2006) administered a single oral gavage dose of 300 mg/kg DINP to non-pregnant SD rats
and quantified the metabolites in urine daily for 4 days. MciOP accounted for 82 percent of the
identified metabolites, and the other metabolites constituted 18 percent. This study characterized the
different co- and co-1-oxidation metabolites found in urine and found that MciOP was the major urinary
metabolite recovered, while MINP and DINP were not found in significant amounts in the urine.

Based on the available data, EPA assumes an oral absorption of 100 percent for the draft DINP risk
evaluation.

Table 2-2. Metabolites of DINP Identified in Urine from Rats and Humans after Oral
Administration

Metabolite(s)

Abbreviation(s)

Reference(s) (Species)

Monoisobutyl phtlialate

MINP

(Anderson et al., .V1 () (human)
(Suzuki et al.. 2012) (human)

(Koch and Aneerer, 2007) (human)
(Calafat et al., 2006a) (rat)

Glucuronidated MINP

MINP-Glue

(Clewell et al., 2013a) (rat)

[mono-(4-methyl-7-carboxyheptyl)
phtlialate] representing:
Mono(carboxyisooctyl) phtlialate

[D4-7carboxy-MmeHP]
C02-MINP; MCIOP

(Anderson et al.. .'01 0 (human)
(Koch and Aneerer. 2007) (human)

[D4-mono-(4-methyl-7-
hydroxyoctyl) phthalate]
representing:

Mono(hydroxyisononyl) phthalate

[70H-MmeOP]
for OH-MINP; MHINP

(Anderson et al., .'01 () (human)
(Koch et al.. 2012) (human)

(Koch and Aneerer. 2007) (human)
(Silva et al., 2006) (rat)

[D4-mono-(4-methyl-7-
oxooctyl)phthalate] repre senting:
Mono(oxoisononyl) phthalate

[7oxo-MmeOP] for Oxo-MINP;
MOINP

(Anderson et al., .'01 () (human)
(Koch et al.. 2012) (human)

(Koch and Aneerer. 2007) (human)
(Silva et al., 2006) (rat)

Monocarboxylisononyl phthalate

cx-MINP

(Koch et al., 2012) (human)

Mono-carboxy-isooctyl phthalate

MCIOP (MCOP is sometimes
used to represent MCIOP)

(Silva et al., 2006) (rat)

Mono(carboxy-isoheptyl) phthalate

MciHpP

(Silva et al.. 2006) (rat)

Mono-(3-carboxypropyl) phthalate

MCPP

(Calafat et al.. 2006b; Calafat et al..
2006a) (rat)

Mono-n-octyl phthalate

Mil OP

(Calafat et al., 2006b) (rat)

Phthalic acid

PA

(McKee et al., 2002) (rat)

2.2	Inhalation Route

No controlled human exposure studies or in vivo animal studies are available that evaluate the ADME
properties of DINP for the inhalation route. Therefore, EPA is assuming 100 percent absorption via
inhalation. Similarly, ECHA concluded 75 percent absorption via inhalation for adults and 100 percent
for newborns and infants as a vulnerable subpopulation (EC U \ < l b, « i «1 „oQ3).

2.3	Dermal Route

In vivo and in vitro studies have shown that absorption of phthalates through rat and human skin
decreases as the length of the alkyl chain increases (Mint et al.. 1994; El si si et al.. 1989; Scott et al..
1987). Dermal absorption data specific to DINP are limited. EPA only identified one study directly

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Although specific data on DINP dermal absorption in humans is lacking, several regulatory agencies
(e.g., Danish EPA, ECHA, NICNAS) recognize that absorption of phthalates would likely be lower in
human skin than through rat skin. This observation is based on data from in vitro migration studies
conducted with DEHP and other phthalates. Notably, other regulatory agencies (e.g., Australia
NICNAS, ECHA) have reached similar conclusions regarding the low dermal absorption of DINP
(ECHA.. b, - lCNAS. :01:).

2.4 Summary

Toxicokinetic data indicates that orally administered DINP is rapidly metabolized in the gut to MINP
and distributed via blood to major tissues, particularly the liver and kidneys. DINP metabolites were
excreted in urine and to a lesser extent in feces. Repeated dosing did not result in accumulation of DINP
and/or its metabolites in blood and tissues but did result in increased formation and elimination of the
monoester oxidation products.

Tissue distribution patterns of DINP revealed that absorption from the GI tract was rapid after both
single and repeated oral dosing. DINP is then primarily hydrolyzed in the GI tract after oral
administration. DINP translocated from the GI tract via the blood rapidly to liver and kidney. The
metabolic profile suggests that DINP is recovered primarily as oxidized products and phthalic acid and
very little as the parent or the metabolite MINP, suggesting that DINP is rapidly metabolized in the GI
tract to the corresponding monoester with a second hydrolysis step in liver to phthalic acid.

DINP is primarily eliminated in urine following oral exposures. Available studies have reported that
more than 90 percent of [14C] DINP was eliminated over 72 hours, with the majority through urine and
to a minor extent through feces(Andcrson et ai. JO I I; ICoch and Angerer. 2007; Silva et ai. 2006;
McKee et ai. 2002). The total radioactivity recovered from the previously identified metabolites
combined was 33 ± 6.4 percent of the labeled DINP in urine over 48 hours. Metabolite half-lives were
estimated to be 4 to 8 hours with over 90 percent excreted in the first 24 hours of urine collection.

In contrast to absorption following oral exposure, dermal absorption of DINP in adult male F344 rats is
low, ranging from 2 to 4 percent of the applied dose when measured 7 days after application (McKee et
ai. 2002). This finding agrees with data from other in vivo and in vitro studies that show absorption of
phthalates through rat and human skin decreases as the length of the alkyl chain increases. The dermally
absorbed fraction is distributed to multiple tissues, including skin, GI tract, muscle, fat, and liver. The
recovery of radioactivity in feces and the GI tract suggests excretion of DINP or its metabolites in the
bile, which in turn suggests that after absorption, DINP undergoes a similar metabolic fate as orally
administered DINP.

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3 HAZARD IDENTIFICATION

EPA has developed detailed hazard characterization and mode of action (MOA) analysis for the effects
on fetal testicular testosterone and liver cancer, with an emphasis on liver effects leading to liver tumors.
Effects on fetal testicular testosterone are presented in Section 3.1.2.1. Non-cancer liver effects are
presented in Section 3.2, while liver cancer and EPA's MOA analysis of liver tumors is presented in
EPA's Draft Cancer Human Health Hazard Assessment for Diisononyl Phthalate (DINP) (U.S. EPA.
2024a). The scientific MOA analysis is presented in accordance with the EPA's Guidelines for
Carcinogen Risk Assessment ( 2005a) and the IPC S Mode of Action Framework (IPCS. 2007)

and includes a description of the state of the science with regards to key events, pathways of toxicity and
weight of evidence following the modified Bradford Hill criteria. Other hazards considered by EPA,
such as kidney, neurotoxicity, cardiovascular health effects, immune system toxicity, and
musculoskeletal toxicity, are presented in Sections 3.3 through 3.7.

3.1 Developmental and Reproductive Toxicity

3.1.1 Summary of Available Epidemiological Studies

EPA reviewed and summarized conclusions from previous assessment conducted by Health Canada
(2018b) and U.S. EPA's IRIS program, including systematic review articles by Radke et al. (2019b;
2018) that investigated the association between DINP exposure and male and female development and
reproductive outcomes. In the Health Canada (2018b) assessment, there were no studies that evaluated
the association between DINP and its metabolites and reproductive outcomes such as altered male
puberty, pregnancy complication and loss, uterine leiomyoma, sexual dysfunction in females, and age at
menopause. There was inadequate evidence for the association between DINP and its metabolites and
reproductive outcomes such as altered female puberty, changes in semen parameters, sexual dysfunction
in males, polycystic ovary syndromes, and sex ratios. There was also no evidence for the association
between DINP and its metabolites and reproductive outcomes such as gynecomastia, endometriosis and
adenomyosis. Overall, Health Canada found that the evidence could not be established for the
association between DINP and its metabolites and any reproductive outcomes, such as altered fertility.

In the conclusions from the IRIS systematic review articles by Radke et al. (2018). examining the
association between DINP male reproductive outcomes the authors found moderate evidence linking
DINP metabolites to lower testosterone levels. However, they could not find clear evidence linking
DINP and male reproductive outcomes such as AGD, time until pregnancy in males, and sperm
parameters due to a combination of low exposure levels {i.e., poor sensitivity) and data availability {i.e.,
fewer accessible studies). In terms of the association between female reproductive and developmental
outcomes and DINP, Radke et al. (2019b) found that the evidence was indeterminate.

EPA identified 11 new epidemiological studies published between 2018 and 2019 that were not
evaluated by Health Canada or either IRIS program systematic reviews. Eight of the available studies
were of medium quality and three were of low quality. Overall, conclusions of the 11 new studies were
consistent with that of Health Canada and the IRIS systematic review articles. EPA preliminarily
concluded that the existing epidemiological studies do not support quantitative dose-response
assessment, but rather provide qualitative support as part of weight of scientific evidence. Further
information on the new studies identified by the EPA can be found in Appendix D.

3.1.2 Summary of Laboratory Animals Studies

The developmental effects of exposure to DINP in experimental animal models have been evaluated as
part of several existing assessments. NTP-CERHR (2003). ECHA (2013b). EFSA (2019). Australia

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NICNAS (20121 Health Canada ffiC/HC. 2015) and U.S. CPSC (201 I, :0]0) have all consistently
concluded that oral exposure to DINP can cause developmental toxicity in experimental animal models.
Oral exposure to DINP has been shown to cause skeletal and visceral variations, reduced pup body
weight gain, and effects on the developing male reproductive system consistent with a disruption of
androgen action. Effects on the developing male reproductive system and other developmental and
reproductive toxicity are discussed in Sections 3.1.2.1 and 3.1.2.2, respectively.

3.1.2.1 Developing Male Reproductive System

EPA has previously considered the weight of scientific evidence and concluded that oral exposure to
DINP can induce effects on the developing male reproductive system consistent with a disruption of
androgen action (see EPA's Draft Proposed Approach for Cumulative Risk Assessment of High-Priority
and a Manufacturer-RequestedPhthalate under the Toxic Substances Control Act ( '023 a)).

Notably, EPA's conclusion was supported by the Science Advisory Committee on Chemicals (SACC)
( E023b). A summary of available studies evaluating effects on the developing male

reproductive system are provided in Section 3.1.2.1.1, while a brief MO A summary is provided in 0.
Readers are directed to see EPA's Draft Proposed Approach for Cumulative Risk Assessment of High-
Priority and a Manufacturer-Requested Phthalate under the Toxic Substances Control Act (

2023 a) for a more thorough discussion of DINP's effects on the developing male reproductive system
and EPA's MOA analysis. Effects on the developing male reproductive system are considered further
for dose-response assessment in Section 4.

3.1.2.1.1 Summary of Studies Evaluating Effects on the Developing Male
Reproductive System

Available studies (including 13 studies of rats) evaluating the antiandrogenic effects of DINP on the
male reproductive system are summarized below in Table 3-1.

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Table 3-1. Summary of DINP Stut

ies Evaluating Effects on the Developing Male Reproductive System

Brief Study Description

NOAEL/ LOAEL
(mg/kg-day)

Effect at LOAEL

Remarks

Pregnant SD rats (8/dose/ timepoint
evaluated) gavaged with 0 (corn oil
vehicle), 50, 250, 750 mg/kg-day DINP
(CASRN 68515-48-0) on GDs 12-19.
Dams sacrificed on GD 19 (2 hours post-
dosing) or GD 20 (24 hours post-dosing)
(Clewell et al. 2013a)

50/ 250

i fetal testicular
testosterone and
testicular

pathology (MNGs)

Maternal Effects

-1 (12%) absolute and relative maternal liver weight (>250 mg/kg-day)
Developmental Effects

-	i (50-65%) testicular testosterone on GD 19 (>250 mg/kg-day)

-	Testicular pathology on GD 20 (f MNGs [>250 mg/kg-day], Ley dig cell
aggregates [750 mg/kg-day])

Unaffected outcomes

-	Maternal body weight gain; terminal maternal body weight; fetal body weight;
male AGD (GD 20); testicular testosterone on GD 20; seminiferous tubule diameter
on GD 20

Pregnant SD rats (20-24 litters/dose) fed
diets containing 0, 760, 3,800, or 11,400
ppm DINP (CASRN 68515-48-0) on GD
12 through PND 14 (equivalent to: 56,
288, 720 mg/kg-day on GD 13-20 and
109, 555, 1,513 mg/kg-day on PND 2-
14). Dams allowed to deliver pups
naturally, and pups sacrificed on PNDs 49
or 50 (Clewell et al. 2013b)

56/ 288

i male pup body
weight on PND 14
and t incidence of
MNGs on PND 2

Maternal Effects

-	i body weight on GD 20, PND 2 and 14 (11,400 ppm)

-	i (30%) body weight gain on GD 10-20 (11,400 ppm)

-1 food consumption on GD 10-20 (11,400 ppm) and PND 2-14 (>3,800 ppm)
Developmental Effects

-	i (10-27%) male pup weight on PND 2 (720 mg/kg-day) and 14 (>288 mg/kg-
day)

-	Testicular pathology on PND 2 (f Leydig cell aggregates (720), MNGs (>288
mg/kg-day)

-| AGD on PND 14 (720 mg/kg-day)

-1 (10%) absolute LABC weight on PND 49-50 (720)

Unaffected outcomes

-	Live pups/litter; testicular testosterone (PND 49); PPS; AGD (PND 2, 49); NR
(PND 14, 49); absolute testis and epididymis weight (PND 2, 49); gubernacular
cord length (PND 49); male offspring body weight (PND 49); absolute testes,
epididymis, SV, ventral prostate, glans penis, Cowper's Glands weight (PND 49);
reproductive tract malformations (PND 49) (e.g., hypospadias, exposed os penis,
undescended testes, epididymal agenesis); testicular pathology (PND 49)

Pregnant Wistar rats (# of litters per dose
not stated) fed soy-free diets containing 0,
40, 400, 4000, or 20,000 ppm DINP
(CASRN 28553-12-0) from GD 15
through PND 21 and allowed to deliver
pups naturally [received doses, as
estimated bv (EC/HC. 2015): 2. 20. 200.
1000 me/ke-davl (Lee et al.. 2006a)

None/ 2

i male pup AGD,
| pup body
weight, i female
lordosis quotient

Maternal Effects

-	Not examined or reported
Developmental Effects

-	i male/female body weight on PND 1 (>2 mg/kg-day)

-	i male AGD on PND 1 (>2 mg/kg-day)

-	i frequency of mounts, intromissions, ejaculations in male rats (PNW 20) (only at
2 mg/kg-day, no dose-response)

-	i Lordosis quotient of females in PNW 20 (>2 mg/kg-day)

Unaffected outcomes

-	Serum testosterone and estradiol (PND 7); serum testosterone, luteinizing
hormone, follicle stimulating hormone, estradiol (PNW 20)

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Brief Study Description

NOAEL/ LOAEL
(mg/kg-day)

Effect at LOAEL

Remarks

Pregnant SD rats (5/dose) fed soy-free
diets containing 0, 400, 4,000, 20,000
ppm DINP (CASRN 28553-12-0) on GD
15 through PND 10 (equivalent to: 31,
307, 1,165 mg/kg-day on GD 15-20 and
66, 657, 2,657 mg/kg-day on PND 2-10)
(Masutomi et aL 2003)

66/ 657

i male body
weight on PND 27

Maternal Effects

-	i body weight gain and food consumption between GD 15-20 & PND 2-10
(20,000 ppm)

Developmental Effects

-j body weight gain between PND 2-10 (both sexes) (20,000 ppm)

-1 (18-43%) body weight on PND 27 for males (>4,000 ppm) and females (20,000

ppm)

-	i Absolute testes weight on PND 27 (20,000 ppm)

-	Testicular pathology on PND 77 (20,000 ppm) (i.e., vacuolar degeneration of
Sertoli cells, degeneration of meiotic spermatocytes at stage XIV, scattered cell
debris in ducts of epididymis)

Unaffected outcomes

-Number of live offspring; pup body weight (PND 2); AGD (PND 2); pup body
weight gain (PND 10-21); PPS; vaginal opening; absolute testes weight (PND 77)

Pregnant Wistar rats gavaged with 0 (corn
oil vehicle), 300, 600, 750, 900 mg/kg-
day DINP (CASRN 28553-12-0) on GD 7
through PND 17. Dams sacrificed on GD
21 (subgroup 1) or allowed to give birth
naturally and offspring sacrificed on PND
90 (suburouD 2) (Boberg et aL. 2016.
2011)

300/ 600

t MNGs in fetal
testis and j sperm
motility on PND
90

Maternal Effects

-	None

Developmental Effects

-	Testis pathology on GD 21 (t incidence of MNGs (>600 mg/kg-day); enlarged
diameter of seminiferous cords (>750); gonocytes with central location in chords
(>750))

-	i Testicular testosterone on GD 21 (600, no dose-response)

-	i male pup body weight on PND 13 (900)

-1 male pup AGD on PND 1 (900) and t male pup NR on PND 13 (>750)

-1 sperm motility on PND 90 (>600)

Unaffected Outcomes

-	Maternal body weight and weight gain; gestation length, post-implantation loss,
litter size, sex ratio, perinatal loss; testicular testosterone production (GD 21);
plasma testosterone and luteinizing hormone (GD 21); fetal birth weight; male and
female body weight (PND 90); absolute reproductive organ weight (PND 90) (e.g.,
testis, prostate LABC, SV, ovary, uterus); AGD or NR (PND 90); testis testosterone
(PND 90); SV, prostate, testis pathology

Pregnant Harlan SD rats (5-9/dose)
gavaged with 0, 500, 750, 1,000, or 1,500
mg/kg-day DINP (CASRNs 28553-12-0
and 68515-48-0 tested) on GDs 14-18.
Dams sacrificed on GD 18, approximately
2 hours oost-dosinu (Hannas et al.. 2011)

None/ 500

i fetal testicular

testosterone

production

Maternal Effects

-	None

Developmental Effects

-	i (30-69%) ex vivo fetal testicular testosterone production (>500 mg/kg-day, both
CASRNs)

-	i expression of StAR and Cyplla mRNA in fetal testes (>1,000 mg/kg-day, both
CASRNs)

Unaffected Outcomes

-	Dam mortality; dam body weight gain; litter size

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Brief Study Description

NOAEL/ LOAEL
(mg/kg-day)

Effect at LOAEL

Remarks

Pregnant Harlan SD rats gavaged with 0,
500, 750, 1,000, or 1,500 mg/kg-day
DINP (CASRNs 28553-12-0 and 68515-
48-0 tested) on GD 14-18. Dams
sacrificed on GD 18, approximately 2
hours oost-dosinu (Hannas et al. 2012)

NOEL/ LOEL:
None/ 500

i steroidogenic
gene expression in
the fetal testes

Maternal Effects

-	None

Developmental Effects

-1 mRNA expression of.SV.I/*', Cyplla, Cypllbl, Cypllb2, Hsd3b, Cypl7al,
Scarbl, Insl3, Dhcr7in the fetal testes (>500 mg/kg-day, both CASRNs)
Unaffected Outcomes

-	Dam mortality; dam body weight gain; litter size

Pregnant SD rats (5-8/dose) gavaged with
0 (corn oil vehicle), 250, or 750 mg/kg-
day DINP (CASRN not reported) on
embryonic days 13.5-17.5. Dams
sacrificed on embryonic day 19.5
(Adamsson et al.. 2009)

NOEL/ LOEL:
250/ 750

t GATA-4, Insl3,
P450scc mRNA in
the fetal testes

Maternal Effects

-	None

Developmental Effects

-1 Testicular mRNA expression of GATA-4, Insl3, P450scc (750 mg/kg-day)
Unaffected Outcomes

-	Plasma corticosterone; litter size; sex ratio; fetal body weight; testicular
testosterone; testicular mRNA expression of Star, 3/1-1ISD, SF-1; testicular protein
expression of StAR, P450scc, 3(3-HSD, androgen receptor; testicular pathology

Pregnant SD rats (14-19/dose) gavaged
with 0 (corn oil vehicle) or 750 mg/kg-day
DINP (CASRN 68515-48-0) from GD 14
through PND 3. Dams were allowed to
give birth naturally and mall offspring
were sacrificed between 3 to 7 months of
ase (Gray et al. 2000)

None/ 750

t male pup NR,

reproductive

malformations

Maternal Effects

-	i (10%) maternal weight gain to GD 21
Developmental Effects

-1 percent of males with areolas (22.4%) on PND 13

-	Reproductive malformations at 3-7 months: permanent nipples in 2/52 males from
2 litters, small and atrophic testes in 1/52 males; flaccid, fluid-filled in 1/52 males;
unilateral epididymal agenesis with hypospermatogenesis in 1/52 males
Unaffected outcomes

-	Maternal mortality; maternal weight gain to PND 3; male pup weight at birth;
PPS; absolute reproductive organ weight at 3-7 months (i.e., testes, LABC, SV,
glans penis, ventral prostate, epididymis, cauda epididymis, caput-corpus
epididymis); serum testosterone (3-7 months); male AGD (PND 2); reproductive
malformations at 3-7 months (hypospadias, cleft phallus, vaginal pouch, SV
agenesis, undescended testes, testis absent, abnormal gubernacular cord)

Pregnant Harlan SD rats (3-5/dose)
gavaged with 0 (corn oil vehicle) or 750
mg/kg-day DINP on GDs 14-18. Dams
sacrificed on GD 18, approximately 2
hours post-dosing. Study completed over
several blocks. Block 1 and 5 tested
CASRN 68515-48-0, Block 7 tested
CASRN 28553-12-0 (Fiiit et al. 2014)

None/ 750

i fetal testicular

testosterone

production

Maternal Effects

-	None

Developmental Effects

-	i (24-50% I across Blocks 1, 5, and 7) ex vivo fetal testicular testosterone
production

Unaffected Outcomes

-	Maternal weight gain, fetal viability (all blocks)

Pregnant Wistar rats (8/dose) gavaged
with 0 or 750 mg/kg-day DINP (CASRN

None/ 750

i fetal testicular
testosterone

Maternal Effects
- Not examined or reported

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Brief Study Description

NOAEL/ LOAEL
(mg/kg-day)

Effect at LOAEL

Remarks

28553-12-0) on GDs 7-21. Dams
sacrificed on GD 21 OBorcli et aL 2004)



content and
production

Developmental Effects

-1 ex vivo fetal testicular testosterone production and testicular testosterone content
(magnitude of effect not reported, only presented graphically
Unaffected Outcomes

- Plasma testosterone and luteinizing hormone

Hershberger assay: Testosterone
propionate-treated (0.4 mg/kg-day)
castrated immature (7 week old) male SD
rats were administered DINP via gavage
at 0, 20, 100, or 500 mg/kg-day for 10
davs (Lee and Koo. 2007)

NA

NA

-	i absolute SV (>20 mg/kg-day) (lacked dose-response) and LABC (500) weight
Unaffected Outcomes

-	Terminal body weight; absolute liver, kidney, adrenal; ventral prostate, Cowper's
gland; Glans penis weight

Pregnant SD rats (6/dose) gavaged with 0
(corn oil vehicle), 10, 100, 500, 1,000
mg/kg-day DINP (CASRN not provided)
on GD 12-21. Dams were allowed to give
birth naturally and then pups were
sacrificed (Li et aL. 2015)

None/10

i male pup body
weight and fetal
Leydig cell
aggregation

Maternal Effects

-	None

Developmental Effects

-	i male pup body weight (>10 mg/kg-day) (lacked dose-response)

-	i testicular testosterone (1,000)

-1 testis dysgenesis (>100)

-1 incidence of MNGs (>100)

-	Fetal Leydig cell aggregation (>10)

-1 testicular gene expression (Insl3 (>10), Lhcgr (>500), Star (>500), Cypllal
(>100), Hsd3bl (>100), Cypl7al (>100), Hsdl7b3 (1,000))

Unaffected outcomes

-	Gestation length; number of dams delivering litters; pups per litter; sex ratio; dam
body weight; male AGD

AGD = anogenital distance; GD = gestational day ; MNGs = multinucleated gonocytes; PND = postnatal day; PNW = postnatal week

801

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3.1.2.1.2 Mode of Action for Phthalate Syndrome

As shown in Figure 3-1. portions of an MOA for phthalate syndrome have been proposed to explain the
link between gestational or perinatal exposure to DINP and effects on the male reproductive system in
rats. The MOA has been described in greater detail in EPA's Draft Proposed Approach for Cumulative
Risk Assessment of High-Priority Phthalates and a Manufacturer-Re quested Phthalate under the Toxic
Substances Control Act (U.S. EPA. 2023a) and is described briefly below.

Chemical Structure
and Properties



Molecular Initiating
Event

=>

Cellular
Responses

¦=>

Organ
Responses



Adverse Organism
Outcomes

Phthalate
exposure during
critical window of
development

Fetal Male Tissue

-J, AR dependent
mRNA/protein
synthesis



Metabolism to
monoester &
transport to fetal
testes

"=0>

Unknown MIE

(not believed to be
AR or PPARa
mediated)



		

Key genes involved in the AOP \

for phthalate syndrome

Scarbl

Chcr7

Mvd

Ela3b

StAF

Ebp

Nsdhl

Insl3

Cypllal

Fdos

RGD1S64999 Lhcgr

Cypllbl

Hmqcr

Tm7sf2

Inha

Cypllb2

Hmgcsl

Cyp46ol

NrObl

Cypl 7al

Hsd3b

Ldlr

RhoxlO

CypSl

Fidil

Insigl

Wnt7a

¦=>

Testosterone
synthesis

\1/ Gene
expression

(INSL3, lipid
metabolism,
cholesterol and
androgen synthesis
and transport)



4> INSL3 synthesis

Fetal Leydig cell

Abnormal cell
apoptosis/
proliferation

(Nipple/areolae
retention, .J, AGD,

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

Suppressed
gubernacular cord
development

(inguinoscrotal phase)



J

<=>

4- Androgen-
dependent tissue
weights, testicular

pathology [e.g.,
seminiferous tubule

atrophy),
malformations (e.g.,
hypospadias), <1/
sperm production

Suppressed
gubernacular cord
development

(transabdominal
	phase)	J

&

<0

A

Impaired



V

fertility



1}



Undescended





testes



Figure 3-1. Hypothesized Phthalate Syndrome Mode of Action Following Gestational Exposure

Figure taken directly from ( J.S. EPA. 2023a) and adapted from (Conlev et al., 2021; Gray et al.. 2021;
Schwartz et aL 202.1; Howdeshell et al.. 2017).

AR = androgen receptor; INSL3 = insulin-like growth factor 3; MNG = multinucleated gonocyte; PPARa =
peroxisome proliferator-activated receptor alpha.

The MOA underlying phthalate syndrome has not been fully established; however, key cellular-, organ-,
and organism-level effects are generally understood (Figure 3-1). 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).

Exposure to DINP during the masculinization programming window (i.e., GDs 15.5 to 18.5 for rats;
GDs 14 to 16 for mice; gestational weeks 8 to 14 for humans) in which androgen action drives
development of the male reproductive system can lead to antiandrogenic effects on the male
reproductive system ( lacLeod et al.. 2010; Welsh et al., 2008; Carruthers and Foster. 2005). In vivo
pharmacokinetic studies with rats have demonstrated that monoester metabolites of DINP can cross the
placenta and be delivered to the target tissue, the fetal testes (Clewell et al.. 2013a; Clewell et al.. 2010).
Consistent with the MOA outlined in Figure 3-1, studies of DINP have demonstrated that exposure to
DINP during the masculinization programming window in rats can reduce mRNA levels of insulin-like

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growth factor 3 (INSL3), as well as genes involved in steroidogenesis in the fetal testes (Li et ai. 2015;
Hannas et ai. 2011; Adams son et al. 2009). Consistently, studies have also demonstrated that exposure
to DINP during the masculinization programming window can reduce fetal testicular testosterone
content and/or testosterone production (Li et al.. .^ I , Furr et al.. -01 I; i "lewell et al.. 2013a; Sobers et
al.. 2011; Hannas et al.. JO I I; »rch et al.. 2004). Exposure to DINP during the masculinization
programming window can also reduce male pup anogenital distance (AGD) and cause male pup nipple
retention (NR), which are two hallmarks of antiandrogenic substances; however effects on AGD and NR
are less consistently observed following oral exposure to DINP in rats (see Sections 3.1.3.3 and 3.1.3.4
of (	023a) for additional discussion). In contrast, exposure to DINP generally does not induce

severe reproductive tract malformations such as hypospadias and or cryptorchidism, but has been shown
to cause epididymal agenesis (Gray et al.. 2000). and a spectrum of other effects consistent with
phthalate syndrome, including increased numbers of multinucleated gonocytes (MNGs) (Li et al.. 2015;
Clewell et al.. 2013a; Clewell et al.. 2013b; Sobers et al.. 2011). fetal Leydig cell aggregation (Li et al..
201 \ t iewell et al.. 2013a; Clewell et al.. 2013b). and decrease sperm motility (Sobers ci al.. 201 I).

Based on available data, EPA previously concluded that the weight of scientific evidence demonstrates
that oral exposure to DINP can induce effects on the developing male reproductive system consistent
with a disruption of androgen action and the MOA outlined in Figure 3-1 (see EPA's Draft Proposed
Approach for Cumulative Risk Assessment of High-Priority and a Manufacturer-Requested Phthalate
under the Toxic Substances Control Act (U.S. EPA. 2023 a)). Notably, EPA's conclusion was supported
by the SACC (U.S. EPA. 2023b).

3.1.2.2 Other Developmental and Reproductive Outcomes

EPA has evaluated several oral exposure studies, including two prenatal developmental studies of rats
(Waterman et al.. 1999; Hellwis et al.. 1997). a one-generation study of reproduction of rats (Waterman
et al.. 2000; Exxon Biomedical. 1996a). and a two-generation study of reproduction of rats (Waterman
et al.. 2000; Exxon Biomedical. 1996b). EPA identified several studies published from 2015 to 2019
evaluating estrogenic potential (Sedha et al.. ), reproductive effects (Chiang and Flaws. 2019).
developmental effects (Neier et al.. 2018; Setti Ahmed et al.. 2018). and metabolic effects (Neier et al..
2019) of DINP in mice and rats treated in the perinatal period. No studies of development are available
for the dermal or inhalation exposure routes. Available studies are summarized in Table 3-2 and
discussed further below.

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Table 3-2. Summary of DINP Studies Evaluating Ef

'ects on Reproduction and Development

Brief Study Description

NOAEL/
LOAEL
(m«/k«-(lay)

Effect at
LOAEL

Remarks

Pregnant SD rats (23-25/dose) gavaged with
0 (corn oil vehicle), 100, 500, 1,000 mg/kg-
day DINP (CASRN 68515-48-0) on GDs 6-
15. Dams sacrificed on GD 21 (Waterman et

a.L 1999)

100/ 500"

t Skeletal
variations

Maternal Effects

-	i (13%) food consumption on GDs 6-9 (1,000 mg/kg-day)

-	i body weight gain on GDs 6-9, 6-15 (1,000)

Developmental Effects

-1 incidence of rudimentary lumbar (>500), supernumerary cervical ribs (1,000), renal
pelves (1,000)

Unaffected Outcomes

-	Maternal survival, clinical signs, resorptions, post-implantation loss, fetal viability,
fetal body weight, sex ratio, incidence of fetal malformations

Pregnant Wistar rats (10/dose) gavaged with
0 (corn oil vehicle), 40, 200, 1,000 mg/kg-
day DINP-1 (CASRN 68515-48-0) on GDs
6-15. Dams sacrificed on GD 20 (Hellwig et

al. 1997)

200/1000

t Skeletal
variations

Maternal Effects

-	i food consumption (1,000 mg/kg-day)

-	Clinical signs (vaginal haemorrhage and urine-smeared fur in one dam) (1000)
-1 (13%) relative kidney weight

Developmental Effects

-1 skeletal variations (rudimentary cervical and accessory 14th ribs) (1000)

Unaffected Outcomes

-	Survival; maternal body weight; uterus weight; relative liver weight; resorptions; post-
implantation loss; number of live fetuses per dam; fetal weight

Pregnant Wistar rats (10/dose) gavaged with
0 (corn oil vehicle), 40, 200, 1,000 mg/kg-
day DINP-2 (CASRN 28553-12-0) on GDs
6-15. Dams sacrificed on GD 20 (Hellwig et

al. 1997)

200/1000

t Skeletal
variations

Maternal Effects

-	Clinical signs (vaginal haemorrhage in one dam) (1,000)

Developmental Effects

-1 skeletal variations (rudimentary cervical and accessory 14th ribs) (1,000)

Unaffected Outcomes

-	Survival; food consumption; maternal body weight; uterus weight; relative liver and
kidney weight; resorptions; post-implantation loss; number of live fetuses per dam; fetal
weight

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Brief Study Description

NOAEL/
LOAEL
(mg/kg-day)

Effect at
LOAEL

Remarks

Pregnant Wistar rats (10/dose) gavaged with
0 (corn oil vehicle), 40, 200, 1,000 mg/kg-
day DINP-3 (CASRN 28553-12-0, resulting
from a different production line than DINP-2)
on GDs 6-15. Dams sacrificed on GD 20
(Hellwie et al. 1997)

200/1000

t incidence
of skeletal,
visceral, and
soft tissue
variations

Maternal Effects

-	i food consumption (1000 mg/kg-day)

-	i body weight gain from GD 6-15 (,1000)

-1 (11%) relative liver weight
Developmental Effects

-1 skeletal retardations (unossified or incompletely ossified sternebrae) (1,000)

-1 soft tissue variations (hydroureter) (1000

-1 skeletal variations (rudimentary cervical and accessory 14th ribs) (1,000)
Unaffected Outcomes

-	Survival; clinical signs; uterus weight; resorptions; post-implantation loss; number of
live fetuses per dam; fetal weight

Male and female SD rats (30/sex/dose) fed
diets containing 0, 0.5, 1.0, 1.5%DINP
(CASRN 68515-48-0) starting 10 weeks prior
to mating, through mating, gestation, and
lactation continuously for one generation.
Received doses in units of mg/kg-day shown
in Table 3-5. (Waterman et al. 2000; Exxon
Biomedical 1996a)

None/ 377

I F1 male
and female
body weight
on PNDs 0,
14,21

Parental (PI) Effects

-	i PI body weight (both sexes) (>1.0%)

-	i PI food consumption (both sexes) (>1.0%)

-1 absolute and relative liver weight (both sexes) (>0.5%)

-1 absolute and/or relative kidney weight (both sexes) (>0.5%)

-1 absolute testes, right epididymis, and ovary weight (1.5%)

Fertility Effects

-	None

OffsDri nu (Fl) Effects

-	i live births, j PND 4 survival, j PND 14 survival, j viability at weaning (all at 1.5%)

-	i male and female body weights on PND 0, 1, 14, 21 (>0.5%)

Unaffected Outcomes

-	Clinical signs (PI); survival (PI); reproductive indices (male mating, male/female
fertility, female fecundity, gestational indices); litter size; number of live/dead offspring
at birth; sex ratio

Male and female SD rats (30/sex/dose) fed
diets containing 0, 0.2, 0.4, 0.8% DINP
(CASRN 68515-48-0) starting 10 weeks prior
to mating, through mating, gestation, and
lactation continuously for two-generations.
Received doses in units of mg/kg-day shown
in Table 3-7. (Waterman et al. 2000; Exxon
Biomedical 1996b)

None/133

I F1 and F2
male and
female body
weight on
PNDs 7 and
21

Parental (PI. P2) Effects

-	i PI female body weight on PNDs 14 and 21 (0.8%)

-1 P2 male and female body weight (>0.4%)

-	i PI female food consumption during lactational period (0.8%)

-1 P2 male and female food consumption during premating, gestation, and lactational
periods (0.8%)

-1 relative and/or absolute liver weight for PI males and females (>0.4%) & P2 males
and females (0.8%)

-1 absolute kidney weight for PI males (>0.4%) and females (>0.2%) & P2 males
(0.8%)

-1 incidence of minimal to moderate cytoplasmic eosinophilia (both sexes in PI and P2)
(>0.2%)

-1 incidence of minimal to moderate dilation of the renal pelves for P2 males (>0.4%)

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Brief Study Description

NOAEL/
LOAEL
(mg/kg-day)

Effect at
LOAEL

Remarks







Fertility Effects

-	None

OffsDri nu (Fl. F2) Effects

-1 F1 male and female offspring body weight on PND 21 (>0.2%)

-	i F2 female offspring body weight on PND 7 (>0.2%)

Unaffected Outcomes

-	Clinical signs (PI, P2); survival (PI, P2); reproductive indices (male mating,
male/female fertility, female fecundity, gestational indices) (PI, P2); litter size (Fl, F2);
number of live/dead offspring at birth (Fl, F2); sex ratio (Fl, F2)

Uterotrophic Assay: 20 day old female Wistar
rats (6/group) were gavaged with 0
(untreated), 0 (corn oil vehicle), 276, 1380
mg/kg-day DINP (CASRN 68515-48-0), or
40 |ig/kg-dav diethylstilbesterol for 3 days.
Animals sacrificed 24 hours after dosing
(Sedha et al. 2015)

None/ 276

| body
weight gain

-	i body weight gain (>276 mg/kg-day)

-	Positive control gave anticipated results

Unaffected Outcomes

-	Uterine and pair ovary wet weight

Pubertal Assay: 20 day old female Wistar rats
were gavaged with 0 (untreated), 0 (corn oil
vehicle), 276, 1380 mg/kg-day DINP
(CASRN 68515-48-0), or diethylstilbesterol 6
Hg/kg-day diethylstilbesterol for 20 days
starting on PND 21. Animals were sacrificed
onPND 41 rSedha et al.. 2015)

None/ 276

| body
weight gain

-	i body weight gain (>276 mg/kg-day)

-	i (10-28%) relative and absolute ovary weight (1380 mg/kg-day)

-	Positive control gave anticipated results

Unaffected Outcomes

-	Absolute and relative uterine wet weight and vaginal weight; vaginal opening

Pregnant Wistar rats (36/dose) gavaged with
0 (corn oil vehicle) or 380 mg/kg-day DINP
(CASRN 68515-48-0) from GD 8 through

PND 30 (Setti Ahmed et al., 2018)

None/ 380

| pup body
weight gain

Maternal Effects

-	i food consumption during gestation (14-39%) and lactation (48-62%)

Developmental Effects

-	i pup body weight gain (54-56%) from PND 15 to 30

-	i small intestine weight

-	Villous atrophy in duodenum, ilium, jejumum (qualitative, no incidence data reported)
Unaffected Outcomes

-	Number of live pups per litter

CD-I female mice (4-12/dose) were gavaged
with 0 (corn oil vehicle), 0.02, 0.1, 20, or 200
mg/kg-day DINP (CASRN not provided) for
10 days and then mated with untreated males
immediately after, as well as 3 and 9 months
oost-dosinu (Chiang and Flaws. 2019)

200/ None

NA

Maternal Effects h

-	None definitively related to treatment
Developmental Effects h

-	None definitively related to treatment

Unaffected Outcomes (all timcDoints. unless otherwise noted)

-	Body weight; absolute ovary, uterine, liver weight; time to mating; fertility index;
gestational index; gestation length; litter size; pup weight; pup mortality; estrous
cyclicity (0, 9 months)

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Brief Study Description

NOAEL/
LOAEL
(mg/kg-day)

Effect at
LOAEL

Remarks

Female yellow agouti mice (resulting in 15-
17 litters/dose) were fed diets of 0 or 75
mg/kg feed DINP (equivalent to 0 or 15
mg/kg-day) from 2 weeks prior to mating
through weaning (PND21) with body and
organ weights were collected on PND21
(Neier et al. 2018)

None/15

t maternal
body weight
gain; f pup
body weight;
t pup

relative liver
weight

Maternal Effects
-1 body weight gains
Developmental Effects (PND21)

-1 pup body weight (both sexes)

-1 pup relative liver weight (females)

Unaffected Outcomes (PND21)

- Number of live pups per litter; maternal body weight; pup hepatic triglycerides; pup
gonadal fat, brain, spleen, and kidney weights; pup liver weight (male); pup Avy DNA
methylation

Female yellow agouti mice (17-21/group)
were fed diets of 0 or 75 mg/kg feed DINP
(equivalent to 0 or 15 mg/kg-day) from 2
weeks prior to mating through weaning
(PND21). 1 male and female pup/litter were
allowed to recover for 10 months (Neier et
al. 2019)

None/15

I birth rates;
t pup liver
masses;
altered pup
body

composition;
| glucose
tolerance

Maternal Effects

-	i birth rates

Developmental Effects (PND21)

-1 liver masses (males, p > 0.1)

-1 body fat (females, longitudinal 2-8 months)

-	i lean mass percentage (females, longitudinal 2-8 months)

-	i glucose tolerance (females, longitudinal 2-8 months)

Unaffected Outcomes (at 2 and 8 months unless noted)

-	Pup body weight across life course (PND21-10 months); pup physical activity; pup
food intake; pup energy expenditure; resting metabolic rate, respiratory exchange rate,
fat oxidation rate, glucose oxidation rate; pup plasma adipokines

" Waterman et al. originally identified a developmental NOAEL of 500 mg/kg-day DINP based on increased incidence of skeletal variations. However, a re-analysis of
the data by study sponsors using the generalized estimating equation approach to the linearized model supported a NOAEL of 100 mg/kg-day DINP. Results from the
statistical re-analvsis are reported in (NTP-CERHR 2003).

b The study authors in Chiang (2019) reported several statistically significant findings as related to treatment with DINP; however, EPA considered these differences to
be spurious and incidental to treatment because they were unrelated to dose, transient, and/or not adverse. These significant differences included: differences in estrous
cyclicity at 20 and 100 |ig/kg/day and 200 mg/kg-day DINP and fewer pregnant females at 20 |ig/kg/day at 3 months post-dosing; differences in estrous cyclicity at 100
Hg/kg-day and reduced time to mating at 100 |ig/kg-dav to 200 mg/kg-day DINP and increased percent males in litters at 100 |ig/kg-day and 20 and 200 mg/kg-day DINP
at 9 months post-dosing.

862

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In the first study, which adhered to EPA §798.4900 (40 CFR Part 798, 1985), Waterman et al.(1999)
gavaged pregnant SD rats (23 to 25 per dose) with 0, 100, 500, and 1,000 mg/kg-day DINP (CASRN
68515-48-0) on GDs 6 through 15. Maternal toxicity was limited to the high-dose group and included a
reduction in maternal body weight gain on GDs 6 through 9 and 6 through 15 (magnitude of effect not
reported), and a 13 percent decrease in food consumption on GDs 6 through 9. Food consumption and
bodyweight gain significantly increased after cessation of exposure between GDs 18 through 21 and
mean maternal body weight recovered to control levels by GD 21. No treatment-related effects on
maternal survival, clinical signs, resorptions, post-implantation loss, fetal viability, sex ratio, or fetal
body weight were observed. No malformations were observed at any dose. Fetal effects were limited to
treatment-related increases in skeletal and visceral variations, including increased incidence of renal
pelves at 1,000 mg/kg-day, rudimentary lumbar ribs at 500 and 1,000 mg/kg-day, and supernumerary
cervical ribs at 1,000 mg/kg-day (Table 3-3). EPA identified a developmental NOAEL of 100 mg/kg-
day DINP based on increased incidence of skeletal variations at 500 mg/kg-day and above and a
maternal NOAEL of 500 mg/kg-day based on reduced maternal weight gain and food consumption at
1000 mg/kg-day DINP.

Table 3-3. Mean Percent of

'etuses in Litter with Skeletal Variations (Waterman et ai. 1999V

Rudimentary lumbar ribs

Supernumerary cervical ribs

" Adapted from Tables 5 and 6 in (NTP-CERHR. 2003)

h * indicates P<0.05 and ** indicates p < 0.01. Skeletal variation data was re-analyzed by study sponsors using
the generalized estimating equation (GEE) approach to the linearized model to account for potential litter
effects. The statistical re-anah sis conducted bv studv sponsors is reported in (NTP-CERHR. 2003). Renal
pelves data could not be re-analyzed using the GEE methodology due to the zero incidence in the control. Renal
pelves data was re-analysed using two approaches, including a nested analysis that considered litter effects and
by changing one control fetus to affected and using the GEE approach. Both approaches provided similar
results (significant increase at 1,000 mg/kg-day).

In a second prenatal study, Hell wig et al. (1997) gavaged pregnant Wistar rats (10 per dose) with 0, 40,
200, and 1,000 mg/kg-day DINP on GDs 6 through 15. Three different formulations of DINP were
evaluated, including: DINP-1 (CASRN 68515-48-0, purity >99%), commercially available with the
alcohol moiety consisting of roughly equivalent amounts of 3,4-, 4,6-, 3,6-, 3,5-, 4,5-, and 5,6-
dimethylheptanol-1; DINP-2 (28553-12-0), with at least 95% of the alcohol components as alkyl-
substituted octanol or heptanol derived from //-butene; and DINP-3 (28553-12-0), resulting from a
different production line from DINP-2, with main alcohol components synthesized from //-isobutene,
resulting in >60% alkyl-substituted hexanols. The studies were Good Laboratory Practice (GLP)-
compliant and generally adhered to EPA §798.4900 (40 CFR pat 798, 1992), with the exception that 10
dams, instead of 20 were employed per dose group. For DINP-1, maternal toxicity was limited to the
high-dose group and included reduced food consumption (magnitude of effect not reported), clinical
signs (i.e., vaginal haemorrhage and urine smeared fur in one dam), and a 13 percent increase in relative
kidney (but not liver) weight. No treatment-related effects on maternal body weight, maternal survival,
resorptions, post-implantation loss, number of live fetuses per dam, or fetal weights were observed.

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Developmental effects were limited to the high-dose group and included a statistically significant
increase in the percent fetuses per litter with variations (35.3, 41.5, 29.5, and 58.4 percent across dose
groups). Variations showing dose-related increases included rudimentary cervical and accessory 14th
rib(s), and an apparent, non-statistically significant, increase in dilated renal pelves (Table 3-4). For
DINP-2, there was no statistically significant maternal toxicity that was treatment-related. One dam
given 1,000 mg/kg-day DINP-2 had vaginal hemorrhage on GD 14 and 15. No effects on food
consumption, maternal body weight, maternal survival, relative liver or kidney weight, resorptions, post-
implantation loss, number of live fetuses per dam, or fetal weights were observed. Study authors state
that "the only substance-related fetal effect was an increased incidence of a skeletal variation [accessory
14th rib(s)]" in the high-dose group, although the incidence of rudimentary cervical rib(s) also appeared
slightly increased (Table 3-4). Multiple malformations were observed in one high-dose fetus, including
globular-shaped heart, unilobular lung, hydrocephaly, dilation of aortic arch, and anasarca, which were
regarded as spontaneous and not treatment related by study authors. For DINP-3, maternal toxicity was
limited to the high dose group, and included reduced food consumption (magnitude of effect not
reported), decreased body weight gain from GDs 6 to 15, increased (11 percent) relative liver weight,
and a non-statistically significant increase (9 percent) in relative kidney weight. No effects on maternal
survival, resorptions, post-implantation loss, number of live fetuses per dam, or fetal weights were
observed. Developmental effects were limited to the high-dose group and included a statistically
significant increase in the percent fetuses per litter with variations (35.3, 29.6, 39.5, and 60.7 percent
across dose groups), including increased incidences of skeletal retardations (unossified or incompletely
ossified sternebrae), skeletal variations (rudimentary cervical and/or accessory 14th rib[s]) and soft tissue
variations (hydroureter, dilated renal pelvis) (Table 3-4). Additionally, study authors attributed some
soft tissue malformations (predominately affecting the urogenital tract) and skeletal malformations
(shortened and bent humerus and femur in a single fetus) in the high-dose group to be treatment-related.
Overall, similar developmental findings were observed for all three tested formulations of DINP and
support a developmental NOAEL of 200 mg/kg-day based on increased skeletal and visceral variations
at 1,000 mg/kg-day.

Table 3-4. Incidence of Visceral, Skeletal, and Soft Tissue Variations (I

lellwig et al..:

1997)"

DINP
Formulation

Number of Fetuses Evaluated and
Type of Fetal Variation

Control

40

(mg/kg-day)

200
(mg/kg-day)

1,000
(mg/kg-day)

DINP-1

No. fetuses (litters) evaluated

135 (9)

116(9)

111(8)

131 (10)

Rudimentary cervical rib(s)



2(1)

1

11(5)

Accessory 14th rib





2(2)

37 (10)

Dilated renal pelvis

12(7)

11(4)

8(4)

22 (9)

DINP-2

No. fetuses (litters) evaluated

135 (9)

116(9)

135(10)

141 (10)

Rudimentary cervical rib(s)



1

4(2)

10(5)

Accessory 14th rib





1

4(4)

DINP-3

No. fetuses (litters) evaluated

135 (9)

138 (10)

135 (9)

120 (9)

Rudimentary cervical rib(s)





2(1)

12(7)

Accessory 14th rib





9(5)

34 (8)

Sternebrae not ossified

6(3)

1

3(2)

26 (7)

Sternebrae incompletely ossified or
reduced in size

20 (7)

11(7)

16(6)

36(9)

Dilated renal pelvis

12(7)

15(8)

13(9)

20 (9)

Hydroureter

4(3)

5(3)

1

12(8)

"Table adapted from Tables 10, 12, and 14 in Hellwiu et al. (1997).

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DINP has also been evaluated in both one- and two-generation studies of reproduction, which were GLP
compliant and conducted in accordance with EPA Test Guidelines for Reproductive and Fertility Effects
(40 CFR Part 798, 1985) (Waterman et ai. 2000; Exxon Biomedical. 1996a. b). In the one generation
study, SD rats (30/sex/dose) were continuously administered dietary concentrations of 0, 0.5, 1.0, and
1.5 percent DINP (CASRN 68515-48-0) starting 10 weeks prior to mating, throughout mating, gestation,
and lactation. Mean received doses in units of mg/kg-day are shown in Table 3-5. PI males were
sacrificed following delivery of the last litter of F1 pups, while PI females were sacrificed at F1
weaning on postnatal day (PND) 21. No treatment-related clinical signs or effects on survival were
reported for PI males or females. Body weight was statistically significantly reduced in mid- and high-
dose males and females during the premating phase, and in mid- (5.3 to 15.3 percent decrease) and high-
dose (10.8 to 23.3 percent decrease) PI females during gestation and lactation. Similarly, food
consumption was significantly reduced in mid- (5.3 to 8.7 percent decrease) and high-dose (5.8 to 10.5
percent decrease) males and females during the premating phase, and in mid- (16.7 to 27.4 percent
decrease) and high-dose (11.6 to 42.2 percent decrease) PI females during gestation and lactation.
Treatment with DINP had no significant effects on any reproductive indices, including male mating,
male/female fertility, female fecundity, or gestational indices. Mean litter size, mean number of live and
dead offspring, and sex ratio were unaffected by treatment with DINP. At the high dose, treatment with
DINP significantly reduced percent live births (95.2 vs. 98.2 percent in controls), PND 4 survival (85.6
vs. 93.1 percent in controls), PND 14 survival (92.7 vs. 98.5 percent in controls), and viability at
weaning (87.2 versus 93.9 percent in controls). Male and female F1 offspring body weight was
significantly reduced in all treatment groups on PNDs 0 (7.9 to 11.5 percent) and continued to be
reduced, although not always statistically significantly, in all treatment groups for both sexes through
PND 21 (Table 3-6). Overall, this study supports a developmental LOAEL of 377 mg/kg-day (no
NOAEL identified), based on reduced F1 offspring body weight throughout the lactational period.

Table 3-5. Mean Measured Doses (mg/kg-day) from the One-Generation Study of DINP in SD
Rats (\XAierm,m ci ,tl. J!000; I won Uums			

Dose (%)

Premating - Males

Premating - Females

Gestation

Postpartum

0.5

301-591

363-624

377-395

490-923

1.0

622-1,157

734-1,169

741-765

1,034-1,731

1.5

966-1,676

1114-1,694

1087-1,128

1,274-2,246

"Adapted from Table 12 in Exxon Biomedical (1996a)

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956

957

958

959

960

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962

963

964

965

966

967

968

969

970

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Table 3-6. F1 Offspring Postnatal Body Weight (Grams) from the One-Generation Study of
Reproduction in SD Rats (Waterman et *1.. 2000; Ei.1.011 Biomedical l996aV*c	

F1 Offspring

Croup

Male

Female

PND 0

PND 1

PND 4

PND 7

PND 14

PND 21

PND 0

PND 1

PND 4

PND 7

PND 14

PND 21

0%

6.98

7.34

9.80

16.02

33.77

54.34

6.68

7.05

9.58

15.60

32.72

52.19

0.2%

6.49**

6.83

9.18

14.52

30.00**

48.94*

6.15**

6.52*

8.81

14.07

29.40**

47.77**

0.4%

6.42**

6.92*

9.12

14.00*

26.23**

39.93**

6.05**

6.49**

8.56*

13.24*

25.04**

38.13**

0.8%

6.27**

6.58**

8.19**

11.04**

20.18**

29.32**

5.91**

6.25**

7.84**

10.71**

19.31**

27.71**

Historical
Control

6.35-
7.02

6.68-
7.46

8.53-
11.43

13.64-
18.74

28.81-
36.73

44.89-
60.77

5.96-
6.74

6.30-
7.16

8.32-
11.05

13.33-
17.69

27.22-
35.74

42.39-
61.19

" Data from Table 4 in Waterman et al. (2000).

4 and '**' indicate the mean is significantly different from the control mean by p < 0.05 and p < 0.01, respectively.
c Historical control data reported to be from the laboratory conducting the study. Further details (e.g., number of studies data
collected from, timespan of studies) regarding the source of historical control data were not provided in (Exxon Biomedical,

1996a).

























In the two-generation study, SD rats (30/sex/dose) were continuously administered dietary
concentrations of 0, 0.2, 0.4, and 0.8 percent DINP (CASRN 68515-48-0) starting 10 weeks prior to
mating, throughout mating, gestation, and lactation continuously for two generations (Waterman et at..
2000; Exxon Biomedic; 3b). Mean received doses in units of mg/kg-day are shown in Table 3-7.
For the first parental generation (PI), no treatment-related clinical signs or effects on survival were
reported for PI animals. No significant effects on PI body weight were observed, except for a 6.7 to 7.8
percent decrease in high-dose female body weight on PNDs 14 and 21. Food consumption was
significantly reduced (9 percent) for high-dose females during the postnatal phase of the study but was
not reduced for males or females during other phases of the study. For the second parental generation
(P2), no treatment-related clinical signs or effects on survival were reported. At the start of the
premating period (six weeks after weaning), mean body weights for mid and high dose males and
females were lower than control. Females in the high-dose group had consistently lower body weight
gain compared to the control group during the premating (statistically significant for first 2 weeks),
gestation (not significant), and lactational (significant for PND 4 to 21) phases. Small (less than 8
percent), but statistically significant, decreases in food consumption were observed in high-dose males
and females during the premating period and in high-dose females during gestation (13 to 16 percent
decrease) and lactation (9 percent decrease). No treatment-related effects on any reproductive indices
were observed for either generations, including male mating, male/female fertility, female fecundity, or
gestational indices.

Similarly, gestation length, mean litter size, mean number of live and dead offspring, sex ratio, percent
live births, survival on PNDs 1, 4, 7, 14, and 21, and viability at weaning were unaffected by treatment
with DINP for both the F1 and F2 generations. F1 and F2 offspring body weight was significantly
reduced throughout PNDs 0 to 21 (Table 3-8). For F1 offspring, bodyweight was significantly reduced
6.8 percent for high-dose males on PND 0; 10 to 15 percent for mid- and high-dose males and females
on PNDs 7 and 14; and 8.9 to 19 percent for males and females on PND 21 in all dose groups. For F2
offspring, bodyweight was significantly reduced 14 to 17 percent for mid- and high-dose females on
PND 4; 14 to 19 percent for mid- and high-dose males and 10 to 21 percent for females in all dose
groups on PND 7; 12 to 21 percent for mid- and high-dose males and females on PND 14; and 12 to 22
percent for mid- and high-dose males and females on PND 21. Study authors state that the observed
body weight changes were within historical control ranges from the laboratory conducting the study and
that effects on body weight at 0.2 and 0.4 percent DINP "seem unrelated to treatment." However, no
information regarding the source of the historical control data is provided (e.g., number of studies, years

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study conducted, strain/species tested, and diet animals were maintained on were not reported), so it is
unclear if an appropriate historical control dataset was used. Overall, this study suggests a
developmental LOAEL of 0.2 percent DINP (equivalent to approximately 133 mg/kg-day during
gestation) for decrements in F1 and F2 body weight during lactation.

Table 3-7. Mean Measured Doses (mg/kg-day) from the Two-Generation Study of DINP in SD
Rats

Dose
(%)

PI Generation

P2 Generation

Prematinjj-
Malcs

Prematinjj-
Fcmalcs

Gestation

Postpartum

Prematin};-
Males

Prematinjj-
Females

Gestation

Postpartum

0.2

118-212

145-215

139-153

159-350

114-264

140-254

133-153

174-395

0.4

236-126

278-425

274-301

347-731

235-523

271-522

271-307

348-758

0.8

477-852

562-830

543-571

673-1,379

467-1,090

544-1,060

544-577

718-1541

" Adapted from Tables 12 and 32 in Exxon Biomedical C1996b)

Table 3-8. F1 and F2 Offspring Postnatal Body Weight (Grams) from the Two-Generation Study
of Reproduction in SD Rats (Waterman et ai. 2000: Exxon Biomedical. 1996b) ahc	

F1 Offspring

Group

Male

Female

PND 0

PND 1

PND 4

PND 7

PND 14

PND 21

PND 0

PND 1

PND 4

PND 7

PND 14

PND 21

0%

6.90

7.49

10.63

17.62

35.01

57.25

6.47

7.11

10.26

16.70

33.52

53.99

0.2%

6.78

7.39

10.26

16.44

33.28

51.40*

6.36

6.96

9.61

15.54

31.89

49.19*

0.4%

6.48

7.03

9.54

15.28**

30.43**

47.95**

6.16

6.67

9.24

14.21**

29.14**

45.63**

0.8%

6.43*

7.05

9.74

15.67*

29.66**

46.52**

6.08

6.70

9.36

15.03*

28.41**

44.68**

Histori-
cal

Control

6.35-
7.02

6.68-
7.46

8.53-
11.43

13.64-
18.74

28.81-
36.73

44.89-
60.77

5.96-
6.74

6.30-
7.16

8.32-
11.05

13.33-
17.69

27.22-
35.74

42.39-
61.19

l '2 iilTsprum

<)"„



"

|(K..

is.us

i" ()')

(¦: m

(.44

"|u

In 4S

1 ~ 4"

'5 X<>



0.2%

6.49

7.12

10.05

16.43

34.80

57.89

6.13

6.75

9.60

15.72*

33.64

55.50

0.4%

6.55

7.08

9.73

15.48**

32.51**

54.82**

6.11

6.59

9.05**

14.56**

31.22**

51.98**

0.8%

6.18

6.64

9.05

14.70**

29.88**

49.12**

5.92

6.41

8.68**

13.76**

28.20**

46.20**

Historic
al

Control

6.35-
7.02

6.68-
7.46

8.53-
11.43

13.64-
18.74

28.81-
36.73

44.89-
60.77

5.96-
6.74

6.30-
7.16

8.32-
11.05

13.33-
17.69

27.22-
35.74

42.39-
61.19

" Data from Tables 8 and 11 in Waterman et al. (2000).

h and '**' indicate the mean is significantly different from the control mean by p < 0.05 and p < 0.01, respectively.
c Historical control data reported to be from the laboratory conducting the study. Further details (e.g., number of studies data
collected from, timesoan of studies) reeardine the source of historical control data were not provided in (Exxon Biomedical.
1996b).

Ahmed et al. (2018) investigated the effects of DINP on development of the small intestine. Pregnant
Wistar rats (36 per dose) were gavaged with 0 (corn oil vehicle) or 380 mg/kg-day DINP from GD 8
through PND 30. Treatment with DINP reduced maternal food consumption 14 to 39 percent during
gestation and 48 to 62 percent during lactation (PNDs 1 to 21), however, it is unclear if reduced food
consumption led to reduced dam body weight, as this outcome was not reported. Pup body weight gain
was significantly reduced (54 to 56 percent) from PND 15 to 30. Study authors report that pup small
intestine weight was significantly reduced 41 percent by treatment with DINP, however, there are
apparent discrepancies between the text and tabular organ weight data (unclear if a statistical analysis as

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done on individual organs). Histologically, offspring small intestine (duodenal, jejunal and ileal
samples) showed villous atrophy following exposure to DINP, however, no incidence data is reported
(only representative photomicrographs are provided). Lactase, maltase, sucrase, and ALP activity in the
duodenum, ilium, and jejunum were also reported to be impacted by treatment with DINP on PND7,
PND15, and PND30. Although results from this study suggest that DINP has effects on the developing
small intestine in offspring exposed via maternal exposure during gestation and lactation, these effects
may be related to the substantial decreases in offspring body weight gain which may be secondary to
decreased maternal food consumption during gestation and lactation.

Chiang and Flaws (. ) gavaged adult CD-I female mice (4 to 12 per group) with 0 (corn oil vehicle),
0.02, 0.1, 20, or 200 mg/kg-day DINP (CASRN not provided) for 10 days and then evaluated effects on
organ weight, estrous cyclicity, and mating behavior with untreated male mice immediately after dosing,
as well as 3 and 9 months post-dosing. Treatment with DINP had no effect on body weight, absolute
ovary, uterine or liver weight at any timepoint. Three months post-dosing, females treated with 0.02 and
200 mg/kg-day spent significantly less time in proestrus and more time in metestrus and diestrus.
However, no dose-related effects on estrous cyclicity were observed immediately following dosing or
nine months post-dosing and the effects observed at three months appeared slight (magnitude of effect
not reported) and of uncertain toxicological significance. No adverse, dose-related, effects on time to
mating, fertility index, gestational index, gestation length, the number of females able to produce pups,
litter size, pup weight on PND20, pup mortality, sex ratio were observed at any timepoint. Several
parameters were statistically significantly altered (e.g., fertility index deceased at 0.02 mg/kg-day at 3
months [but not at higher doses or other timepoints], number of females able to produce pups decreased
at 0.02 mg/kg-day at 3 months and 20 mg/kg-day at 9 months [but not at higher doses]), however, these
findings were of uncertain toxicological significance, given the non-monotonic dose relationship and the
lack of mechanistic data from other studies supporting an effect of DINP on these endpoints.

Two perinatal exposure studies of DINP have also been conducted using the viable yellow agouti mouse
model (Neier et al. l , Meier et al. 2018). In in the first study, a/a dams were fed phytoestrogen-free
diets containing 0 of 75 mg/kg DINP (CASRN not reported) (equivalent to approximately 15 mg/kg-day
DINP) starting two weeks prior to mating with A1-7a males and continuing throughout gestation and
lactation until weaning on PND 21 (Neier et al.. 2018). The exact number of mating pairs per treatment
group is not provided in the 2018 study, however, 15 to 17 litters were produced for the control and
DINP treatment groups. Treatment with DINP had no effect on maternal body weight at PND21,
offspring sex ratio, mean pups per litter, pup mortality through PND21, or pup genotype. Body weight
was significantly increased 10 to 20 percent for females (of genotypes a/a and Avy/a) and 15 percent for
males (of genotype Avy/a) on PND21 in the DINP treatment group. Treatment with DINP correlated to
increased relative liver weight for female pups. There was no change in absolute or relative liver weight
(males only), gonadal fat, brain, spleen, or kidney weights in pups. Additionally, no change in pup DNA
methylation was observed. In summary, treatment with DINP showed modest decreases in pup body
weights and increased relative liver weight (females only).

The Neier et al. (2019) study followed the same dosing scheme with viable yellow agouti mouse dams
exposed from 2 weeks prior to mating through PND21 in diet at dosages of 0 and 75 mg/kg feed DINP
(equivalent to 0 and 15 mg/kg-day). A total of 17 control pairs and 21 DINP pairs were dosed to produce
a minimum of 15 litters per treatment group. The largest male and female from each litter (10 per sex
per dose) were fed a phthalate free diet until 10 months old; one male in the 15 mg/kg-day group died
during glucose gavage at 2 months. The DINP treatment group showed a decreased birth rate, a non-
significant increase in liver masses in males at 10 months (9.1 percent control vs. 33 percent treated).
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also evaluated metabolic effects through adulthood in mice exposed to DINP perinatally with
evaluations at 2, 8, and 10 months. The DINP treatment group showed altered body fat, lean mass
percentage in females longitudinally; however, these effects were not significant when accounting for
multiple comparisons. DINP treated females showed a moderate reduction in glucose tolerance
longitudinally driven by decreased glucose tolerance at two months that improved slightly at eight
months. There was no change in pup body weight across life, physical activity, or food intake.
Additionally, there was no alteration in energy expenditure, resting metabolic rate, respiratory exchange
rate, fat oxidation rate, glucose oxidation rate, or plasma adipokines. Overall, treatment with DINP
resulted decreased birth rate, as well as modest alterations to female pup body composition and glucose
tolerance, without corresponding alterations to diet, physical activity, or other markers for metabolic
activity.

Sedha et al. (2015) investigated the estrogenic potential of DINP in a three-day uterotrophic assay and a
20-day pubertal assay. For the uterotrophic assay, 20-day old female Wistar rats (6/dose/group) were
gavaged with 0 (corn oil vehicle), 276, or 1,380 mg/kg-day DINP (CASRN 68515-48-0) for three
consecutive days, while an additional group was treated with diethylstilbesterol (40 |ig/kg-day), which
served as the positive control. Body weight gain was reduced in both DINP treatment groups compared
to the control, however, treatment with DINP had no significant effect on uterine or paired ovary wet
weight, while the positive control increased ovary and uterus wet weight. For the pubertal assay, 20-day
old female Wistar rats were gavaged with 0 (corn oil vehicle), 276, or 1,380 mg/kg-day DINP and
diethylstilbesterol (6 |ig/kg-day) from PND 21 to sacrifice on PND 41. Body weight gain was
significantly reduced in all DINP treatment groups compared to the control. Absolute and relative
uterine wet weight and vaginal weight were unaffected by treatment with DINP, while relative and
absolute ovary weight was significantly reduced 10 to 28 percent by treatment with 1380 mg/kg-day
DINP. Timing of vaginal opening was unaffected by treatment with DINP. Collectively, results from
these assays indicate that DINP lacks estrogenic potential in vivo.

3.1.2.3 Conclusions on Reproductive and Developmental Toxicity

EPA previously proposed a MOA for male reproductive effects in rodents due to antiandrogenic activity
of DINP as part of a proposed approach for cumulative risk assessment of phthalates ( )23a).

which was supported by the SACC (I] S 1T \ 2023b). As outlined in Table 3-1, male reproductive
effects were observed in 13 rat studies with gestational or oral exposures. Collectively, these data
support EPA's conclusion that exposure to of pregnant female rodents to DINP during gestation results
in effects on male offspring consistent with androgen insufficiency.

An additional 12 developmental studies in mice and rats were included in the dataset covering a wide
developmental window. Available studies included a one-generation study of reproduction in rats
(Waterman et al. 2000; Exxon Biomedic; 3a) and two-generation study of reproduction in rats
(Waterman et al. 2000; Exxon Biomedic; 3b), and a uterotrophic assay in rats (Sedha et al. 2015).
along with multiple studies covering the pre-mating, gestation, and lactation periods. All studies were
limited to oral exposures in rodents.

The evidence for effects on the female endocrine system and reproduction is less clear than the evidence
supporting androgen insufficiency. The uterotrophic assay in rats showed decreased body weight gains,
but no change to uterine or paired ovary wet weight (Sedha ). In the pubertal assay, absolute

and relative uterine wet weight and vaginal weight were unaffected by treatment with DINP, while
relative and absolute ovary weight was significantly reduced at the high dose (1,380 mg/kg-day DINP).
Sexual maturation (time to vaginal opening) was unaffected by treatment with DINP. In the study by
Chiang and Flaws (. ) in which adult CD-I female mice were administered DINP via oral gavage and

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mated with untreated male mice, there were no adverse effects of treatment on body weight, weights of
the uterus or ovaries, time to mating, fertility index, gestational index, gestation length, the number of
females able to produce pups, litter size, pup weight on PND 20, pup mortality, or sex ratio. Several
parameters were significantly different from controls (e.g., decreases in fertility index and number of
females able to produce pups and differences in estrous cycle; however, these findings were of uncertain
toxicological significance, given the findings were often transient, and the non-monotonic dose
relationship and the lack of mechanistic data from other studies supporting an effect of DINP on these
endpoints. Collectively, results from these assays indicate that DINP lacks estrogenic potential in vivo,
and the results of in vitro receptor-binding assays (Kriieer et ai. 2008; Takeuchi et al. 2005; Roy et al.
2004) are consistent with the lack of effects in the uterotrophic and female pubertal assays in Sedha et
al. (20151

Skeletal variations (Waterman et a ); Hellwig et al.. 1997) and reduced body weights were
observed in rat pups across multiple studies (Setti Ahmed et al.. 2018; Sedha et al. 2015; Waterman et
al.. 2000; Exxon Biomedic 5a). Maternal body weights and food consumption were decreased in
several studies on rats (Setti Ahmed et al.. 2018; Waterman et al.. 1999; Hell wis et a I). The one
generation reproduction study showed decreased live births and postnatal survival (Waterman et al..
2000; Exxon Biomedical lvV)6a). One study also identified gastrointestinal effects, including reduced
small intestine weight and villous atrophy in duodenum, ileum, and jejumum, although these findings
are likely related to decreased offspring body weight gain, and secondary to decreased maternal food
consumption (Setti Ahmed et al.. 2018). Two studies of yellow agouti mice dosed with 15 mg/kg-day
DINP from 2 weeks prior to mating through lactation found increased pup body weights, altered body
compositions, and decreased glucose tolerances (Meier et al.. 2019; Neier et al.. 2018). as well as
decreased birth rates (Neier et al.. ). Although these data show different effects in mice and rats, the
low number of studies in mice make it difficult to confidently determine species sensitivity.

Oral exposure to DINP has consistently been shown to cause developmental effects in animal models as
illustrated by the studies described above and concluded by previous assessments by NTP-CERHR
(2003). ECHA (2013b). EFSA (2019). Australia NICNAS ( ), Health Canada (EC/HC. 2015) and
U.S. CPSC (20 I I, 2010). Therefore, EPA is considering developmental toxicity for dose-response
analysis in Section 4.

3.2 Liver Toxicity

The non-cancer health effects and carcinogenicity of DINP have been evaluated primarily in animal
toxicological studies; no human epidemiologic studies evaluating hepatic effects were identified by
EPA's review of existing assessments (primarily Health Canada (2018a)). Moreover, existing
assessments have consistently identified the liver as one of the most sensitive target organs following
oral exposure to DINP in experimental animal studies (ECCC/HC. 2020; EFSA. 2019; EC/I W _< < i ,
ECHA.. 2013b; NICNAS. 20I 2; I CPSC. 201^. i P \ , U ^ '03; NTP-CERHR. 200 , 1 \
C	)

EPA identified twenty-five animal toxicology studies that evaluated non-cancer effects on the liver
following short-term (>1 to 30 days), subchronic (>30 to 90 days), or chronic (>90 days) oral exposure
to DINP. Available studies include: 12 short-term oral studies (6 studies on rats, 5 studies on mice, 1
study on cynomolgus monkeys); 9 subchronic oral exposure studies (6 on rats, 1 on mice, 1 on beagle
dogs, and 1 on marmosets); 4 chronic 2-year oral exposure studies (3 on rats and 1 on mice); and one-
generation and two-generation studies of reproduction of rats that report non-cancer liver effects. More
detailed information on the available studies is provided in Appendix B, including information on
individual study design.

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Exposure to DINP resulted in adverse non-cancer effects on the liver across study designs. Adverse non-
cancer effects such as increased absolute and/or relative liver weight consistently coincided with
increased incidences of non-neoplastic lesions or changes in clinical chemistry parameters, indicative of
liver toxicity. Adverse non-cancer effects on the liver were primarily observed in rats and mice of both
sexes, although there was also evidence of hepatotoxicity from one study in beagles. Two studies in non-
human primates with dose ranges comparable to those in the rodent and beagle studies did not provide
evidence of non-cancer or pre-neoplastic effects on the liver following 14- (Push et at.. 2000) and 90-
day oral exposures to DINP (Hall et at.. 1999). Changes in liver weights, histopathology, and clinical
chemistry parameters in rodents coincided with mechanistic endpoints indicative of Peroxisome
proliferator activated receptor alpha (PPARa) activation, which is discussed further in EPA's Draft
Cancer Human Health Hazard Assessment for Diisononyl Phthalate (DINP) ( 24a).

In general, short term (9 of the 12 studies) and subchronic duration studies (9 of 9) consistently reported
increases in absolute and/or relative liver weight, sometimes in parallel with exposure-related
histopathological effects on the liver (e.g., hepatocellular hypertrophy) or coinciding with increases in
liver enzymes (e.g., ALT, AST, ALP), suggesting impaired liver function. These effects were generally
dose-dependent, observed in both sexes, and in multiple species, including rats, mice, and beagle dogs.
One 13-week study in marmoset monkeys reported non-statistically significant increases in liver weight,
but there was no dose-response, and the authors attribute the lack of statistical significance to high
variability and small sample size (Hall et at.. 1999). More detailed study information for short-term and
subchronic studies is available in Appendix B within TableApx B-l, and TableApx B-2, respectively.

Three chronic 2-year studies in rats (Covance Labs. 1998c; Lington et at.. 1997; Bio/dynamics. 1987)
and one in mice (Covance Labs. 1998b) consistently reported non-cancer liver effects, while all except
the Lington et al. (1997) study reported statistically significant increased incidences of liver tumors (i.e.,
hepatocellular adenomas and/or carcinomas). Non-cancer liver effects that were observed across these
four studies included consistent increases in liver weight that corresponded with histopathological
alterations (e.g., spongiosis hepatis, necrosis) and/or increases in serum enzyme levels or activity in both
sexes. An additional one- and two-generation study in rats by Waterman et al. (2000; Exxon Biomedical.
1996a) found increases in liver weight in the parental generation that coincided with minimal to
moderate cytoplasmic eosinophilia in the liver. More detailed study information for short-term and
subchronic studies is available in Appendix B within Table Apx B-l, and Table Apx B-2.

The NOAEL and LOAEL for non-cancer hepatic effects in F344 rats in Lington et al. (1997) were 15
and 152 mg/kg-day, respectively; based on a statistically significant increase in the incidence of
spongiosis hepatis in mid-dose male rats that was accompanied by increased absolute and relative liver
weights and changes in serum enzyme activities (i.e., increased ALT and AST). These effects are also
the basis for the LOAEL of 359 mg/kg-day (NOAEL of 88 mg/kg-day) in the Covance study (1998c) of
F344 rats. The incidence of spongiosis hepatis was dose-related and significantly increased in male rats
exposed to DINP in both studies. Moreover, a Histopathology Peer Review and Pathology Working
Group (EPL. 1999) independently evaluated the liver slides from rats chronically treated with DINP and
confirmed that the incidence of spongiosis hepatis was increased in male rats in each study.

Bio/dynamics (1987) also reported a significant increase incidence of spongiosis hepatis in male SD rats
of the two highest dose groups, and dose-related trends in both males and females. Detailed information
on lesion incidence is available in Appendix B within Table Apx B-7.

Conclusions on Non-cancer Liver Toxicity

Collectively, short-term, subchronic, and chronic studies of mice, rats, and beagles provide consistent
evidence that oral exposure to DINP can cause liver toxicity. The lowest non-cancer NOAEL identified

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by EPA was 15 mg/kg-day based on increased liver weight, increase serum ALT and AST, and
increased incidence of non-neoplastic lesions (e.g., spongiosis hepatis, enlargement, and granular and
pitted rough changes in hepatocytes, central vein dilation, enlarged, discoloured, congestion, oedema,
and narrowing sinusoidal with loose cytoplasm) in 2-year study of F344 rats (Lington etai. 1997). EPA
further considers liver toxicity for dose-response assessment in Section 4.

EPA summarizes the liver cancer associated with exposure to DINP in a separate technical support
document, the Draft Cancer Human Health Hazard Assessment for Diisononyl Phthalate (DINP) (U.S.
EPA. 2024a).

3.3 Kidney Toxicity

Kidney effects generally occur at higher doses than liver effects and occur inconsistently across study
designs and species (EFSA.. 2019; EC/HC. 2015; EC t P. b, NIC/HAS. 2012; U.S. CPSC. 2010;
EFSA.. 2005; ECB. 2003; NTP-CERHR. 20031

Humans

Although IRIS systematic review process identified five epidemiological studies that investigated the
association between DINP and renal effects, the evidence was deemed inadequate. Three of the five
studies had critical deficiencies in exposure measurement, and the other two studies were of low
confidence and had evidence of selection bias and reverse causality (Radke et al. 2019a).

EPA did not identify any new epidemiologic studies that examine the association between DINP and its
metabolites and/or biomarkers of kidney injury.

Laboratory Animals

Many experimental animal studies have evaluated the kidney toxicity of DINP following oral exposure.
Studies have evaluated the effects on kidney function (i.e., urinalysis parameters, serum BUN levels),
kidney weight, and histopathology. Seventeen studies are available that provide data on
histopathological effects of the kidney, 16 of which also provide data on absolute and/or relative kidney
weights. Six studies report changes in indices of kidney function such as serum BUN levels or urinalysis
parameters. One study was available for the dermal exposure route (Hazleton Laboratories. 1969). No
studies were available for the inhalation exposure route.

Short-Term (>l to 30 Days) Exposure Studies: EPA identified five short-term studies in rodent models
that provide data on the effects of DINP on the kidney (Ma et al.. 2014; Kwack et al.. 2010; Kwack et
al.. 2009; 5; Bio/dynamics. 1982a). An industry-sponsored study by Bio/dynamics (1982a)

exposed male Fischer 344 (F344) rats to 0 or 2 percent (equivalent to 1,700 mg/kg-day) DINP for one
week via feed and evaluated kidney weights and histopathology at study termination. Significant
increases in absolute (7.5 percent increase) and relative (12.2 percent increase) kidney weights were
observed in rats exposed to DINP. No abnormal histopathological findings were observed in the
kidneys. Another study in F344 rats reported similar findings (BIBRA. 1986). BIBRA (1986)
administered 0, 0.6, 1.2, 2.5 percent DINP for 21 days (equivalent to 0, 639, 1,192, 2,195 mg/kg-day
[males]; 0, 607, 1,198, 2,289 mg/kg-day [females]) in the diet to male and female rats and evaluated
kidney weights and histopathology at study termination. Dose-related increases in relative kidney
weights were observed in males and females at all dose levels; the LOAEL was 639 and 607 mg/kg-day
for males and females, respectively. No exposure-related histopathological findings were observed in the
kidneys.

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Not all short-term studies reported dose-related changes in kidney weights that coincide with other
effects of the kidney. Two studies in male SD rats reported no change in relative kidney weights and/or
no change in BUN or other urinalysis parameters (Kwack et ai. 2010; Kwack et al. 2009). while
another in B6C3F1 mice reported changes in weights without a consistent dose-related trend (Hazleton
Labs. 1991a). The studies by Kwack et al. exposed male SD rats to 0 or 500 mg/kg-day DINP via
gavage for 14 days (Kwack et al.. 2010; Kwack et al.. 2009). while the Hazleton study (1991a) exposed
mice to 0, 3,000, 6,000, or 12,500 ppm DINP for 4 weeks (equivalent to 0, 635, 1,377, 2,689, or 6,518
mg/kg-day [males]; 0, 780, 1,761, 3,287, or 6,920 mg/kg-day [females]). In the Hazleton study (1991a).
significant increases in relative kidney weight were observed at the highest dose in males (6,518 mg/kg-
day) and females (6,920 mg/kg-day), but significant decreases were observed at lower dose-levels in
both sexes, which was also true for absolute kidney weights. Nevertheless, the increased relative kidney
weights coincided with significant increased serum BUN levels in high-dose males and increased
incidences of tubular nephrosis in all high-dose males and females, supporting an exposure-related effect
on the kidney (Hazleton Labs. 1991a). In a study in which male Kunming mice were exposed to 0.2, 2,
20 or 200 mg/kg-day DINP for 14 days via gavage. Ma et al. (2014) reported significantly increased
incidences in histopathologic lesions of the kidney, including large reduction in tubular space and
extreme edema of epithelial cells in the glomeruli in animals exposed to the highest dose of DINP.
However, this publication only described these findings qualitatively in text and did not include
quantitative data on incidence or severity.

New Literature: EPA identified one new study published from 2015 through 2020 that provided data on
toxicological effects of the kidney following short term exposure to DINP. The developmental exposure
study by Neier et al. (2018) reported no change in absolute or relative kidney weights at PND21 in male
and female yellow agouti (Avy) mice offspring. In that study, dams were administered 0 or 75 ppm
DINP in the diet (equivalent to 15 mg/kg-day) beginning 2-weeks before mating and continuing through
PND21.

Subchronic (>30 to 90 days) Exposure Studies: EPA identified six dietary studies from existing
assessments that provide data on the toxicological effects of DINP on the kidneys following subchronic
oral exposure (Hazleton Labs 1001 h; Hio/dynamics. 1982b. c; Hazleton Labs. 1981; Hazleton
Laboratories. 1971; Hazleton Labs. 1971) and one gavage study in marmoset monkeys (Hall et al..
1999). These studies provided data across a range of species and strains as well as both sexes. Increases
in absolute and/or relative kidney weights and histopathological effects were reported in all of the
studies. (Hazleton Labs * U>, » tunics. 1982b. c; Hazleton Lab*- I'M; Hazleton Laboratories.

1971; Hazleton Labs. 1971). albeit the effects were sometimes attributable to decreased body weight.
Dose-related increases in absolute and/or relative kidney weights sometimes corresponded with
increased incidences of histopathological lesions or altered urine chemistry, but these trends were not
consistent across all studies.

A study by Bio/dynamics labs (1982b) exposed F344 rats to 0, 0.1, 0.3, 0.6, 1.0, or 2.0 percent DINP for
13 weeks via feed (equivalent to 0, 77, 227, 460, 767, 1,554 mg/kg-day). Dose-dependent increases in
kidney weight were noted in males, where doses as low as 227 mg/kg-day DINP resulted in increased
absolute (9.7 percent) and relative (21.9 percent) weights. The increase in kidney weight was
accompanied by a dose-dependent increase in dark brown discoloration in the kidney from 460 mg/kg-
day. A similar study from Bio/dynamics labs (1982c) exposed Sprague Davvley rats to 0.3 or 1.0 percent
DINP in the diet for 13 weeks (equivalent to 201 or 690 mg/kg-day [males]; 251 or 880 mg/kg-day
[females]). The authors reported dose-related increases in absolute and relative kidney weights in males
and females that corresponded with altered clinical chemistry parameters in males, most notably a dose-

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dependent decrease in triglycerides and increased calcium in high-dose males. The LOEL was 201 or
251 mg/kg-day for males or females, respectively.

These results were similar to three studies from Hazleton Labs (1991b. 19s I, r" I), each using a
different strain of rats. Hazleton Laboratories (1971) reported increases in absolute (9.3 to 17.6 percent
increases) and relative (14.4 to 25.5 percent increases) kidney weight in male and female albino rats of
the highest dose group (500 mg/kg-day). In that study, animals were exposed to 0, 50, 150, 500 mg/kg-
day DINP for 13 weeks. Hazleton Labs ( ) administered 0, 2,500, 5,000, 10,000, or 20,000 ppm
DINP via diet to CDF (F344)/CrlBr rats for 13 weeks (equivalent to 176, 354, 719, or 1545 mg/kg-day
[males]; 218, 438, 823, or 1,687 mg/kg-day [females]). Dose dependent increases in absolute and
relative kidney weights were observed in both sexes, which coincided with a dose-related increase in
granular casts and regenerative /basophilic tubules in the kidneys, beginning at 354 mg/kg-day in males.
Hazleton Laboratories (1981) administered 0, 1,000, 3,000, or 10,000 ppm DINP to SD rats via feed for
13 weeks (equivalent to 0, 60, 180, and 600 mg/kg-day). A LOAEL of 60 mg/kg-day was identified
based on an increased incidence of kidney lesions (focal mononuclear cell infiltration and
mineralization) in exposed males. Absolute and relative kidney weights were also increased in males
and females exposed to 600 mg/kg-day. Absolute weights increased 20 percent in males and 10.8
percent in females, while relative weight increased 17.7 percent in males and 13.7 percent in females.

Although there is ample evidence that the kidney is a target organ for DINP in rodents, the evidence is
less consistent and less numerous across other species, including dogs, monkeys and rabbits. Increased
kidney weights were observed in high-dose animals in a study of beagle dogs by Hazleton Laboratories
(1971). but were attributed to deceased body weight. In that study, animals were administered 0.125,
0.5, or 2 percent DINP in feed for 13 weeks (equivalent to 37, 160, or 2,000 mg/kg-day). The study also
reported increased incidences of tubular epithelial cell hypertrophy in high-dose (2,000 mg/kg-day)
males and females. Urinalysis parameters were comparable between control and test groups. In contrast,
a study in marmoset monkeys by Hall et al. (1999) did not observe any kidney effects. In that study,
male and female marmoset monkeys were exposed to 0, 100, 500, or 2,500 mg/kg-day DINP via gavage
for 13 weeks. No histological findings were exposure related, and there were no changes in kidney
weights. Similarly, no effects on the kidney were observed in a dermal study of New Zealand White
rabbits exposed to up to 2,500 mg/kg-day DINP for 6 weeks (Hazleton Laboratories. 1969).

New Literature: EPA identified one new study published from 2015 through 2020 that provided data on
toxicological effects of the kidney following subchronic exposure to DINP (Deng et al.. ). Deng et
al. (2019) exposed male C57/BL6 mice to 0, 0.15, 1.5 or 15 mg/kg-day DINP for 6 weeks via gavage.
The authors reported vacuoles and hyaline degeneration in the glomerulus of the kidney, as well as
smaller glomeruli and a thickened glomerular basement membrane, However, the authors do not specify
at which doses the effects were observed.

Chronic (>90 days) Exposure: EPA identified five rodent studies from existing assessments that provide
information on the toxicological effects of DINP on the kidney, including four studies following chronic
oral exposure to DINP (CASRN 685 15-48-0) (Covance Lal-v l 8b, c; Lington et al.. 1997). or DINP
(CASRN 71549-78-5)(Bio/dynamics. 1987). and one study following a one-or two-generation exposure
in SD rats (Waterman et al.. 2000). These studies provide data on absolute and/or relative kidney
weights, histopathology, and urinalysis measures that reflect kidney function {i.e., BUN levels).

Lington et al. (1997) and Covance Labs (1998c) evaluated kidney weights, urinalysis parameters, and
kidney histopathology in F344 rats following exposure to DINP for 2 years. Both studies observed
increases in kidney weights in the mid- and high-dose animals, but reported inconsistent results for

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urinalysis parameters and histopathology. Significant increases were observed in relative and absolute
kidney weights in males and females of the mid- and high-dose groups (i.e., 152 and 307 mg/kg-day
[males] or 184 and 375 mg/kg-day [females] at most time points {i.e., 6, 12, 18, and 24 months).
Moreover, relative kidney weight at study termination was increased 10 to 20 percent and 7 to 10
percent in males and females, respectively. In the 2-year study by Covance Labs (1998c). increased
relative kidney weights were observed in rats receiving dietary doses greater than 359 mg/kg-day for
males (over 25 percent increase) and 442 mg/kg-day for females (over 14 percent increase) at study
termination. Kidney weights in the recovery groups were comparable to the same-sex control values at
the end of the 26-week recovery period.

In Lington et al. (1997). there were no exposure-related changes in serum chemistry parameters such as
blood urea nitrogen (BUN). Some of the urine chemistry parameters were affected by DINP exposure in
males. Increased urine volume, potassium, and glucose were observed in high-dose (307 mg/kg-day)
males at most time intervals; potassium and glucose levels were also increased in mid dose males.
Excretion of renal epithelial cells was increased in high-dose males at 6 months, but not at other
timepoints. No urinalysis changes were observed in females. In contrast, Covance Labs (1998c) reported
increases in serum urea (BUN) levels in males and females from the two highest dose groups at multiple
timepoints during the study including study termination {i.e., weeks 26, 52, 78, and 104). BUN was
increased up to 32 percent over controls in the mid-dose (359 mg/kg-day [male] or 442 mg/kg-day
[female]), and 50 percent over controls at the high dose (733 mg/kg-day [male] or 885 mg/kg-day
[female]).

In Covance Labs (1998c). exposure-related increases in the severity of tubule cell pigment occurred in
the kidneys of males exposed to 733 mg/kg-day DINP (Table 3-9). At study termination, a dose-related
increase was observed in the incidence and severity of mineralization of the renal papilla in males at 359
and 733 mg/kg-day DINP as well as in the recovery group. Increased severity of tubule cell pigment was
observed at the two highest dose groups in both sexes (Table 3-9).

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Table 3-9. Incidence and Severity of Selected Non-neoplastic Lesions in the Kidneys of Male and
Female F344 Rats Fed DINP for 2 Years (C'ovance Labs. 1998c)	



Dose Group
mg/kg-day (ppm)

Control

29 M / 36 F
(500)

88 M /109 F
(1,500)

359 M / 442 F
(6,000)

733 Ml 885 F
(12,000)

Recovery"
637 M / 774 F
(12,000)

Number M/F
examined6

36/37

35/38

39/40

31/33

27/32

29/34

\lincrali/.nlion oI'ivikiI papilla (males)

Minimal

6

11

9

6

2

0

Slight

0

0

0

24

1

2

Moderate

0

0

0

0

22

27

Total

6

11

9

30

25

29

Tubule cell piemen! (males)

Minimal

24

21

18

0

0

0

Slight

10

12

21

23

7

26

Moderate

0

1

0

6

17

3

Moderately
severe

0

1

0

2

3

0

Total

34

35

39

31

27

29

Tubule cell piijmenl (Ivmak-s)

Minimal

22

27

34

4

0

1

Slight

14

10

5

27

21

33

Moderate

0

1

1

1

10

0

Moderately
severe

0

0

0

1

1

0

Total

36

38

40

33

32

34

Source: Table 10D on page 350 of Covance Labs 0998c)

M = Male; F = female

a The 12,000 ppm recovery group received 12,000 ppm DINP in the diet for 78 weeks, followed by a 26-week
recovery period during which the test animals received basal diet alone.
b Number examined at terminal sacrifice; does not include unscheduled deaths.

Bio/dynamics ( 7) also conducted a 2-year chronic dietary study in rats, albeit of a different strain
(SD), and noted significant increases in absolute and relative kidney weights in high-dose males at both
the interim (19 and 25 percent, respectively) and terminal (13 and 12 percent) timepoints. Kidney
weights of mid-dose group males (271 mg/kg-day) were increased by 11 percent, although this was not a
statistically significant change. In high-dose females (672 mg/kg-day), increased relative kidney weights
were observed (20 percent increase) at interim sacrifice as well as terminal sacrifice (14 percent
increase). Increased incidence of medullary mineral deposits in the kidney were observed in high-dose
males (25/70 treated vs. 3/70 controls). However, in females, incidences of renal medullary mineral
deposits at the high dose (15/70) were comparable to controls (14/70). No histopathological evaluation

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was conducted on samples from the low- or mid-dose groups, which limits the assessment of dose-
dependency and effect levels.

Waterman et al. (2000) assessed the potential kidney toxicity of DINP in one- and two-generation
studies conducted in SD rats. In the one-generation study, absolute and relative kidney weights in both
sexes were significantly increased at all doses, except in high-dose PI females, and generally in a dose-
related fashion. In the two-generation study, absolute kidney weights of PI males and females were
increased over controls at all DINP treatment levels. Although decreased mean body weights and body
weight gains were also observed in PI males and females for all doses, the changes in kidney weight are
not solely attributable to changes in body weight. Increased incidence of minimal to moderate renal
pelvis dilation was observed in F2 males of the two highest dose groups (0.4 and 0.8 percent, equivalent
to 741-796, 1087-1186 mg/kg-day). No changes were observed in the females; therefore, the authors
attributed the increased incidence of kidney lesions to induction of male rat-specific alpha 2u-globulin
(a2u-globulin).

In contrast to the studies in rats which consistently reported increases in relative and/or absolute kidney
weight, a study in male B6C3F1 mice reported decreased kidney weights (Covance Labs. 1998b). In that
study, male and female mice were exposed to 0, 1,500, 4,000, or 8,000 ppm DINP for 2 years via feed
(equivalent to 0, 276, 742, or 1,560 mg/kg-day). No effects were observed in females. In addition to the
weight changes in males, the authors reported significant increases in urine output, decreases in mean
urine osmolality; and decreased sodium, potassium, and chloride levels in male and female mice from
the 1,560 mg/kg-day dose group at 26, 52, 78, and 104 weeks. The study authors concluded that there
was no DINP-related change in glomerular filtration rate; however, they suggested that this pattern of
urinalysis findings may indicate a compromised ability to concentrate urine in the renal tubule
epithelium, as an increased incidence of chronic progressive nephropathy was observed in high-dose
females (1,888 mg/kg-day). The kidneys of 1,888 mg/kg-day females also had a granular pitted/rough
appearance. The effects of DINP on the kidney, including decreased kidney weights in males, were
partially attenuated in the recovery groups, which were evaluated 26-weeks after the end of exposure.
The reversibility of the kidney effects in the recovery groups was not as pronounced as that for liver
effects (Section 3.1). The incidences of chronic progressive nephropathy in female mice were
comparable to those of the control group upon termination, suggesting that nephropathy is reversible or
that exacerbation of this lesion halted when exposure to DINP was discontinued.

New Literature: EPA did not identify any new studies published from 2015 through 2019 that provided
data on toxicological effects of the kidney following chronic exposure to DINP.

Mechanistic Information

EPA identified two in vivo studies that provide data that may inform mechanisms of action of the
observed nephrotoxic effects of DINP (Ma et al.. 2014; Caldwell et al.. 1999). Mechanisms evaluated
include oxidative stress and male rat-specific a2u-globulin.

Ma et al. (2014) evaluated the contribution of oxidative stress to the aforementioned tissue lesions
observed in the kidneys of male Kunming mice, which were primarily observed at 200 mg/kg-day. In
that study, mice were exposed to 0.2, 2, 20, or 200 mg/kg-day DINP for 14 days via gavage, and
endpoints relevant to oxidative stress were evaluated in renal and hepatic tissue homogenates. Increases
in reactive oxygen species (ROS) and MDA, in parallel with decreases in glutathione (GSH) content,
were observed at 200 mg/kg-day DINP, indicative of oxidative stress. Some indices of oxidative stress
were observed at lower doses than those that resulted in kidney lesions. Indeed, the authors also reported
DNA-protein-crosslinks at 200 mg/kg-day and increases in 8-hydroxydeoxyguanosine (8-OH-dG) at 20

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and 200 mg/kg-day, which indicate oxidative damage to DNA. Levels of interleukin (IL)-l and tumor
necrosis factor alpha (TNFa) were also increased at 20 and 200 mg/kg-day, which would be consistent
with enhancement of an inflammatory response. The authors also evaluated the effect of combined
exposure of 200 mg/kg-day DINP and melatonin (50 mg/kg-day). Mice exposed to 200 mg/kg-day
DINP plus 50 mg/kg-day melatonin showed glomerular cell proliferation and milt renal tubule epithelial
cell edema, and attenuated indices of oxidative stress (ROS, GSH, MDA, DNA-protein-crosslinks, and
cytokine levels). These data indicate that melatonin can attenuate the oxidative stress that results from
exposure to DINP in mice, but not fully attenuate damage to renal tissue, and support an MOA where
oxidative stress may contribute to the toxicological effects of DINP on the kidney.

Caldwell et al. (1999) followed up on observations from Lington et al. (1997). that kidney tumors were
observed in male rats, but not female rats. The male-specific nature of the findings led them to evaluate
a mechanism of action involving the male rat-specific a2u-globulin. Tissue sections from male and
female F344 rats at the 12-month interim sacrifice were evaluated. In male rats, a dose-dependent
increase in a2u-globulin accumulation was observed in regions of the kidney where increased cell
proliferation was also observed. In parallel, tubular epithelial hypertrophy and tubular regeneration were
observed. a2u-globulin was not detected in the kidneys of female rats, and renal cell proliferation of
DINP-exposed female rats was comparable to controls. These results are consistent a mechanism where
a2u-globulin accumulation leads to kidney tissue damage, cell proliferation, and subsequent neoplastic
lesions of the kidney in male rats. The two-generation study by Waterman et al. (2000) also attributed
their observations of renal pelvis dilation in the kidney of F2 male rats to induction of a2u-globulin.
However, these effects are not regarded as relevant to humans (Swenberg and Lehm an - Mckeem an.
1999; U.S. EPA. 1991a). Kidney tumors and evidence for an a2u-globulin MOA are further discussed in
EPA's Draft Cancer Human Health Hazard Assessment for Diisononyl Phthalate (DINP) (U.S. EPA.
2024a).

Conclusions on Kidney Toxicity

Twenty studies in experimental animal models have evaluated toxicologic effects of DINP on the kidney
following short-term, subchronic, developmental, or chronic exposure to DINP. Findings were similar
across study designs, including increased absolute and/or relative kidney weights, and observed in both
sexes, but these data predominantly reflect rat studies, and the toxicological effects of DINP on the
kidney is less certain in other species.

Increases in absolute and/or relative kidney weight have been observed primarily in rat studies across
multiple study designs and often coincide with increased incidences of non-neoplastic lesions of the
kidney or altered urinalysis parameters. Indeed, increased kidney weights were reported in two short-
term studies in F344 rats ( 86; Bio/dynamics. 1982a). five subchronic studies in various

strains of rats (Hazletom Lahv I —! b, r% to namics. 1982b. c; Hazleton Labs. 19\ I, r" I), three
chronic studies in rats (Covance Labs. 1998c; Lington et al. rs° , ^to/dynamics. 1987) and one
developmental study in rats (Waterman et al.. 2000).

In the 2-year study conducted by Lington et al. (1997). increased relative kidney weights of male and
female rats were observed following exposure to dietary levels of 152 and 307 mg/kg-day (males) or
184 and 375 mg/kg-day (females). In the 2-year study reported by Covance Labs (1998c). increased
relative kidney weights occurred in rats receiving dietary doses greater than 359 mg/kg-day for males
and 442 mg/kg-day for females. Urinalysis findings from the chronic studies included significant
increases in urine output and corresponding decreases in electrolyte levels in high-dose males,
suggesting compromised ability to concentrate urine in the renal tubule epithelium. These effects
occurred at the same dosages that produced changes in kidney weights. In the Covance Labs (1998c)

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study, serum urea levels (a marker of kidney toxicity) were significantly increased in rats exposed to 359
mg/kg-day and higher during the second half of the study. Increases in urine volume and kidney lesions
were observed in the recovery group exposed to 733 mg/kg-day.

In many of the chronic studies, effects on the kidney generally occurred at doses equivalent to those
where effects on the liver were observed in rats (Covance Labs. 1998c; Lington et ai. 1997) and mice
(Covance Labs. 1998b). Moreover, the LOAELs ranged from 152 to 923 mg/kg-day which reflect
effects on both the liver and kidneys, including increases in absolute and relative kidney weight as well
as histopathologic findings in the kidney in two chronic studies of male rats (Covance Labs. 1998c;
Lington et at.. 1997). The NOAEL in the Lington study was 15 mg/kg-day (males) or 18 mg/kg-day
(females). However, in a third chronic exposure study in rats (Bio/dynamics. 1987). effects on the
kidney were observed, but not at the LOAEL, suggesting that the kidney may be less sensitive than the
liver to the effects of DINP.

The findings of increased kidney weight in rats were inconsistent with one study of mice, which
reported decreased absolute kidney weight in males (LOAEL of 276 mg/kg-day; NOAEL of 90 mg/kg-
day in males) (Covance Labs. 1998b). That study also reported chronic progressive nephropathy in
female mice of the high-dose group (1,888 mg/kg-day) but no effects in males (Covance Labs. 1998b).
The lack of coherence of effects (e.g., organ weight, histopathology data do not coincide in males or
females) is a limitation of this study.

The MOA of kidney toxicity is not currently known, and effects on the kidney are primarily observed in
one species (rats). Furthermore, kidney effects observed in the rat are less sensitive than effects on the
liver and on developmental outcomes. EPA is considering kidney toxicity for dose-response analysis in
Section 4.

3.4 Neurotoxicity

Humans

Health Canada (2018a) evaluated multiple studies that investigated the association between DINP
exposure and several behavioral and neurodevelopmental outcomes, including mental and psychomotor
neurodevelopment, behavioral and cognitive functioning (i.e., autism spectrum disorders, learning
disabilities, attention-deficit disorder, and attention-deficit hyperactivity disorder), neurological
function, and gender-related play behaviors. Across available studies of DINP, Health Canada
determined that the level of evidence for association between DINP and its metabolites and neurological
effects could not be established.

Radke et al. (2020a) evaluated the association between DINP and neurodevelopment and found that
there was no clear association between DINP and neurodevelopment. Three research studies examined
the relationship between DINP and cognition; however, two of the studies found no relationship and one
revealed an inverse relationship. As a result, the evidence supporting the relationship between DINP and
cognition is deemed inconclusive. Because of the limited number of studies examining this relationship,
the evidence linking DINP to motor ability is regarded as weak. The data supporting the link between
boys' behavior and DINP found no increased odds of ADHD with DINP exposure, and the authors
considered the evidence preliminary. Because of the inconsistent reports about the relationship between
DINP and newborn neurobehavior, the evidence was considered indeterminate. The inconsistent nature
of the currently available research renders the evidence for a connection between DINP and
autism/social impairment as unclear.

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New Literature: EPA identified eleven new studies (2 high quality and 9 medium quality), that evaluated
the association between urinary DINP and neurological effects. The first high-quality study, by Shin et
al. (2018). examined a subset of the of mother-child pairs from Markers of Autism Risk in Babies
Learning Early Signs (MARBLES) cohort to evaluate the association between exposure to DINP
metabolite (MCOP) and Autism spectrum disorder (ASD) and non-typical development (Non-TD).
Among mothers who did not take prenatal vitamins, prenatal MCOP exposure during mid to late
pregnancy was associated with higher risk of non-TD (vs. typical development) [MCOP RRR =1.86
(95% CI: 1.01, 3.39)]. Among mothers who did take prenatal vitamins, prenatal MCOP exposure during
mid-to-late pregnancy was associated with lower risk of autism spectrum disorder (versus typical
development) [MCOP RRR = 0.49 (95% CI: 0.27, 0.88)]. There was an association in multinominal
logistic regression of MCOP during 2nd trimester and ASD (vs. TD) among mothers who took prenatal
vitamins [RRR = 0.41 (95% CI: 0.21, 0.79)].

Another high quality cross-sectional study, by Jankowska et al. (2019b). conducted from a subset of the
Polish Mother and Child Cohort (REPRO PL), examined the association between Child behavioral and
emotional problems at age 7 years, child cognitive and psychomotor development and DINP exposure.
Negative associations in peer relationship problems were noted for sum DINP metabolites, and lower
Intelligence and Development Scales (IDS) scores were generally positively associated with higher
phthalate concentrations.

The first medium quality prospective analysis, by Balalian et al. (2019). of maternal prenatal and child
age 3, 5 and 7 postnatal DINP metabolite (MCOP) exposures with motor skills at age 11 as assessed by
the short form of the BOT-2 were selected from participants in an ongoing longitudinal birth cohort
study of mothers and newborns conducted by the Columbia Center for Children's Environmental Health
(CCCEH). MCOP measured at age 3 was inversely associated with BOT-2 total, fine motor, and gross
motor composite scores among boys. In linear regression models, a 1 log-unit increase in age 3 MCOP
was associated with lower total [beta: -3.08 995% CI: -5.35, -0.80)], fine motor [beta: -1.64 (95% CI:
-3.16, -0.12)], and gross motor [beta: -1.44 (95% CI: -2.60, -0.28)] composite scores in boys.
Comparisons of the 4th versus 1st quartiles of age 3 MCOP were also associated with all three outcomes
in boys [(Q4 vs. Q1 total composite score [beta: -7.47 (95% CI: -12.60, -2.34)]; fine motor composite
score [beta: -4.18 (95% CI: -7.51, -0.85)]; gross motor composite score [beta: -3.29 (95% CI -6.06,
-0.52)]. No significant associations were found between MCOP at age 3 and outcomes in girls. There
were no significant association for sex differences at age 3. There were also no significant associations
between prenatal MCOP and outcomes in either girls or boys. There were no significant associations
between MCOP measured at ages 5 or 7 and outcomes in either girls or boys.

A medium quality study, by Li et al. ( ), used data from children in the Cincinnati Health Outcomes
and Measures of the Environment (HOME) cohort to analyze associations between DINP metabolites
(MCOP, MCNP) and child cognition measured at ages 5 and 8 years. The pattern of associations for
MCOP and MCNP measures was heterogeneous (p < 0.20 for MCNP), and no adjusted associations
reached significance. Associations between child IQ scores and urinary MCOP measured at different
ages were not statistically significant and were heterogeneous (positive and negative). For exposure at
age 3 years, when associations with several other phthalate metabolites were significantly inverse,
adjusted beta for MCOP = -1.2 (95% CI: -3.2, 0.9)

Another medium quality cohort study, by Tanner et al. (2020). examined mother-child pairs from the
Swedish Environmental Longitudinal Mother and Child, Asthma and Allergy (SELMA) study and the
association between prenatal urinary DINP metabolite (MHiDP, MCNP, MHiNP, MOiNP, MCiOP)
exposure and child IQ at age 7 years. Since this is a mixtures analysis, the DINP metabolites of interest

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were not directly analyzed as they were only above the threshold of concern in sensitivity analyses using
positive weights.

A medium quality prospective cohort study, by Jankowska et al. (2019a), evaluated the association
between prenatal and postnatal (age 2 years) OH-MINP and child behavior, cognition, and psychomotor
development at age 7 years. The study included a subset of mother-child pairs from the Polish Mother
and Child Cohort. There were no statistically significant associations between prenatal or postnatal OH-
MINP and any of the study outcomes. There was also no clear pattern of associations with behavioral
outcomes, and associations with cognitive and psychomotor scores were generally weakly negative.
oxo-MINP was measured, but associations with outcomes were not analyzed, as detection rates were
less than 70 percent (56 and 65 percent for pre- and postnatal measures, respectively).

A medium quality cohort study by Hyland et al. (2019) analyzed associations between prenatal DINP
metabolites and neurodevelopment in live singletons in Center for the Health Assessment of Mothers
and Children of Salinas (CHAMACOS), a birth cohort of low-income Mexican American children in
Salinas, California. Associations between IQ scores and MCOP were shown only for combined sexes,
and not significant.

A medium quality longitudinal cohort study, by Jacobson et al. (2021). used data from the NYU
Children's Health and Environment Study, to evaluate urinary DINP metabolites (MCiOP, MINP) levels
in pregnant women and assessed the association with postnatal and postpartum depression following
delivery. There were no significant associations for the Edinburgh Postnatal Depression Scale (EPDS)
score or postpartum depression for sum DINP phthalates.

A medium quality study, by Dzwilewski et al. (2021). used data from a subset of participants in the
Illinois Kids Development Study (IKIDS) to evaluate associations between prenatal exposure to DINP
metabolites (MINP, MCOP, MONP), and infant cognition assessed at 7-8 months of age. The authors
presented results of analyses using the sum of 2 (DINP2) or 3 (DINP3) metabolites, and MONP
individually. Associations varied by infant sex and by the set of images used in testing. DINP2 was
associated with longer processing time for image set 2, and DINP3 with longer processing time among
males viewing set 2. DINP2 and DINP3 had weak negative associations with visual recognition memory
(novelty preference). Urinary EDINP2 metabolites (MINP and MCOP) was associated with significant
increases in average information processing speed (run duration) among infants administered set 2
images. DINP2 was also associated with a non-significant decrease in visual recognition memory
(novelty preference). Urinary EDINP3 metabolites (MINP, MCOP, and MONP) were associated with
significant increases in average information processing speed (run duration) among male infants
administered set 2 images. DINP3 was also associated with a non-significant decrease in visual
recognition memory (novelty preference) overall, while MONP was associated with a non-significant
increase in novelty preference among infants administered set 2 image. DINP3 was associated with a
non-significant decrease in visual attention (time to familiarization) for set 2.

A medium quality case-cohort study, by Kamai et al. (2021). nested in the Norwegian Mother and Child
Cohort (MoBa) analyzed the association between prenatal DINP measured in spot urines at about 17
weeks' gestation and ADHD at age 3 years. DINP was non-linearly associated with increased odds of
preschool ADHD. Results of multivariate logistic regression found an association between increasing
DINP quintile 2 vs. quintile 1, OR = 2.04 (95% CI: 1.2-3.33; includes adjustment for DEHP).

The final medium quality study, a population-based nested case-control study by Engel et al. (2018).
assessed the association of DINP metabolites and ADHD in children of at least 5 years of age of mothers

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within the Norwegian Mother and Child Cohort (MoBa). The authors reported no association of ADHD
with sumDINP metabolites. In Bayesian logistic regression models, there was no association [OR =
0.85, (95% CI: 0.61, 1.15) with log sum of DINP and ADHD. Associations with individual DINP
metabolites were also not significant.

Laboratory Animals

A limited number of experimental animal studies have evaluated the neurotoxicity of DINP following
oral exposure. Existing assessments of DINP have not drawn human health hazard conclusions on the
neurotoxicity of DINP, but have evaluated effects on behavior, brain weight, and/or brain histopathology
0 v 3c: ECCC/HC. 201:0:1 i PSC j'lLii -HA. 2013b: NIC/HAS. 2012: ECB. 2003).

Only three rodent studies ((Boberg et ai. m , U , et al , I . Tuyere et a I JO I I)) are

available that are specifically designed to evaluate neurotoxicity. Remaining studies evaluated brain
weight and/or brain histopathology. These included three subchronic exposure duration studies and three
chronic studies, as well as six developmental exposure studies {i.e., one- or two-generation studies of
reproduction, perinatal, postnatal, or peri-and-postnatal exposure studies). No studies are available for
the dermal or inhalation exposure routes.

One developmental study in Wistar rats (Sobers et al.. 1 ) reviewed in existing assessments (.

CPSC. 2014: NICNAS. 2012) provides data on behavior, including an evaluation of learning and
memory following DINP exposure. Boberg et al. (2011) exposed pregnant Wistar rats to 300, 600, 750,
or 900 mg/kg-day DINP via oral gavage daily from GD 7 through PND 17 and evaluated several
neurobehavioral endpoints on male and female offspring at later timepoints. Behavioral examinations
included those of motor activity levels at PNDs 27 through 28, Morris Water Maze (MWM) at 2 to 3
months of age, sweet preference at 4 months, and radial arm maze performance at 5 to 7 months of age.
The MWM test is used to evaluate learning and memory. In this test, animals are placed in a circular
pool of water and required to escape from water onto a hidden platform using spatial memory.

No changes were observed in motor activity levels and radial arm maze performances in male or female
offspring exposed to DINP during development. An increase in saccharin intake in the sweet preference
test was observed in female offspring of the 750 mg/kg-day group; however, this effect was not dose-
dependent, and the study authors concluded that it may be a chance finding. In the MWM test, dose-
dependent improvements in swim length and latency were observed on the first day of memory testing,
with significantly shorter swim length and latency in the 900 mg/kg-day females. The study authors
asserted that performance in the MWM test is sexually dimorphic, and concluded that DINP affected
spatial learning, as female offspring performed better than controls and similarly to control males in the
MWM, indicating masculinization of behavior in DINP exposed females. However, the effects were no
longer apparent on the second day of memory testing or when the platform was moved to a new position
in the maze. Performance was unaffected by exposure to DINP in males. Notably, the male reproductive
parameters were affected at a lower dose than the apparent effects on learning in memory in females,
with: increased MNGs and decreased sperm motility at 600 mg/kg-day and above; increased nipple
retention at 750 mg/kg-day and above; and decreased AGD at 900 mg/kg-day.

Several rodent studies were identified in existing assessments that provide data on absolute and/or
relative brain weight following exposure to DINP. These include three chronic studies (Covance Labs.
1998a. c; Lington et al.. 1997) and two developmental studies (Masutomi et al.. 2003). In general,
changes in absolute and/or relative brain weight were not observed or were only observed at the highest
doses tested in both males and females. No changes in brain index {i.e., relative brain weight) were
observed in male Kunming mice exposed to 1.5, 15, or 150 mg/kg-day DINP for 9 days via gavage
(Peng. 2015). Similarly, no changes were observed in relative and/or absolute brain weight of: B6C3F1.

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mice exposed to up to 8,000 ppm DINP in feed for two years (equivalent to 1,600 mg/kg-day) (Covance
Labs. 1998b); F344 rats exposed for 2 years to up to 12,000 ppm (equivalent to 733 mg/kg-day in males;
885 mg/kg-day in females) (Covance Labs. 1998c); or up to 0.6 percent (equivalent to 307 mg/kg-day in
males; 375 mg/kg-day in females) (Lington et al. 1997). In contrast, changes in brain weight were
observed in one perinatal exposure study (Masutomi et al.. 2003). In Masutomi et al. (2003). maternal
SD rats were fed test diets containing 0, 400, 4,000, or 20,000 ppm DINP from GD 15 through PND 10
(equivalent to 31, 307, or 1,164 mg/kg-day during gestation and 66, 657, or 2,657 mg/kg-day during
lactation). Significant decreases in absolute brain weight were observed in male (12.9 percent) and
female (11.1 percent) rat pups from the highest dose group at PND 27, while significant increases in
relative brain weight were observed in males (53.5 percent) and females (46 percent), which likely
reflects decreased terminal body weight at PND 27 in the highest dose group in both males and females.
Body weight gain of male and female pups was decreased as well.

Data from existing assessments on the histopathological effects on the brain following DINP exposure
have been reported. Identified literature includes one short-term exposure duration study (Midwest
Research Institu 1) and three chronic studies (Covance Labs. 1998b. c; Lington et al.. 1997). In
general, there were no exposure-related histopathological findings in the 28-day exposure study by the
Midwest Research Institute (1981) nor in the chronic exposure studies in mice (Covance Labs. 1998b)
and rats (Covance Labs. 1998c; Lington et al.. 1997).

New Literature: Four new studies were identified by EPA that had not been reviewed in existing
assessments (Neief et al.. 2018; Setti Ahmed et al.. 2018; Ma et al. JO I i^'ng. l ), which provide
data on neurobehavioral outcomes, brain weights, and brain histopathology following exposure to DINP.
Results of Ma et al. (2015) and Peng et al. (2015) were not fully evaluated in the 2020 Health Canada
Screening Assessment, (ECCC/HC. 2020). and are therefore considered new literature.

Two short-term exposure duration studies in male Kunming mice (Ma et al.. 201 \ Peng. 2015) are
available that provide data on behavior. Impaired learning and memory following DINP exposure was
observed consistently across the two short-term studies. Peng et al. (2015) and Ma et al. (2015) have
similar study designs. Peng et al. ( ) exposed mice to 1.5, 15, or 150 mg/kg-day DINP daily via oral
gavage for 9 days, while Ma et al. (2015) exposed mice to 0.2, 2, 20, or 200 mg/kg-day DINP daily via
oral gavage for 14 days. In both studies, the authors evaluated the effect of DINP on learning and
memory using the MWM test. In both studies, escape latency {i.e., time it took mice to locate submerged
escape platform) was evaluated throughout the exposure period ("training period"), and memory was
evaluated on the last day of exposure ("probe trial") following one day of no testing (a "forget" period).
Each study also investigated the combined effect of DINP and an antioxidant; these endpoints are
discussed in the mechanistic section. Mice were euthanized 24 h after the last DINP exposure, at which
point brain tissue was harvested for histological examination as well as various non-apical measures of
oxidative stress and inflammation (discussed in Mechanistic Information section).

In both Ma et al. (2015) and Peng et al. (2015) escape latency in the MWM test was reduced in each
exposure group at the end of the training periods compared to the first day. Escape latency was increased
in all groups exposed to DINP compared to controls, indicating impaired learning in DINP groups. Peng
et al. (2015) reported decreased retention time in the target quadrant in the MWM test during the probe
trial, indicative of impaired memory. Similarly, Ma et al. (2015) reported decreased time and number of
entries into the target quadrant in the MWM test during the probe trial, indicative of impaired memory.
In addition to MWM, Ma et al. (2015) conducted an open field test to evaluate locomotor activity.
Decreased time and number of entries into the central area were observed for mice exposed to 200
mg/kg-day DINP, which the authors attributed to anxiety-like behavior.

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Four new rodent studies were identified that provide data on absolute and/or relative brain weight
following exposure to DINP, three of which were oral exposure studies. These include one short-term
exposure duration study (Pens. 2015). and two developmental studies (Neier et ai. 2018; Setti Ahmed et
ai. 2018). In general, changes in absolute and/or relative brain weight were not observed, with the
exception of one study weight in yellow agouti (Avy) mice, where biologically significant {i.e., greater
than 10 percent change) changes in brain weight were observed at the highest doses tested in male mice,
which may be exposure-related. No changes in brain index {i.e., relative brain weight) were observed in
male Kunming mice exposed to 1.5, 15, or 150 mg/kg-day DINP for 9 days via gavage (Per ).

Ahmed et al. (2018) observed similar results. In that study, pregnant Wistar rats (36 dams/group) were
exposed to 0 or 380 mg/kg-day DINP via oral gavage beginning on GD 8 and continuing up to PND 30.
Interim sacrifices were conducted on PND 7, PND 15, and PND 21. Brain weight was determined at
interim and terminal timepoints. No changes were observed in absolute brain weights (relative brain
weights not reported) at PND 7, PND 15, or PND 30. Body weight was significantly reduced in pups
exposed to DINP at PND 15 and PND 30. In contrast to the findings of Ahmed et al. (2018). a
developmental study by Neier et al. (2018) reported changes in relative brain weight in yellow agouti
(Avy) mice fed diets containing 5 ppm (equivalent to 15 mg/kg-day) DINP from 2-weeks prior to mating
until weaning. The authors reported absolute and relative brain weights in PND 21 offspring. Decreased
relative brain weights were observed in PND 21 males only, and no changes in absolute weights were
observed. Increased terminal body weights were observed for females, but not males, at PND 21,
indicating that brain weight is decreased in males even when adjusted for body weight. Although it is
likely this observation is exposure-related, uncertainty exists due to the use of the yellow agouti (Avy)
mouse model in the Neier study.

New data on the histopathological effects on the brain following DINP exposure have been reported.
Identified literature includes two short-term exposure duration studies (Ma et al.. 2015; Peng. 2015).
which both reported histopathological alterations in the pyramidal cells of the CAi region of the
hippocampus following short-term exposure to DINP via gavage. Ma et al. (2015) reported damaged
pyramidal neurons in the 20 and 200 mg/kg-day dose groups. Peng et al. ( ) reported that with
increasing DINP exposure, the arrangement of hippocampal cells became more disordered, cells
swelled, and apical dendrites shortened or disappeared. Limitations of the histopathological dataset from
both studies include qualitative presentation of data that lacks incidence or severity information.

Mechanistic Information

EPA identified five in vivo studies and one in vitro study that provide data that may inform mechanisms
of the observed neurological effects of DINP. Three of the in vivo studies investigated mechanisms
involving oxidative stress in mouse models (Duan et al.. 2018; Ma et al.. 1 i 'eng. 2015). The
aforementioned studies by Peng et al. (2015) and Ma et al. (2015) exposed male Kunming mice to DINP
via oral gavage daily for 9 days or 14 days and evaluated several endpoints related to oxidative stress.
Both studies observed increases in ROS, decreases in superoxide dismutase activity, decreases in GSH
content, increases in inflammatory cytokines, and increases in caspase-3 levels, activity, or staining
intensity at the highest dose (200 mg/kg-day) (Ma et al.. 2015) or two highest doses (15 and 150 mg/kg-
day) (Peng. 2015). Ma et al. also reported increases in DNA-protein-crosslinks at 200 mg/kg-day and
increases in 8-OH-dG at 20 and 200 mg/kg-day, indicating oxidative damage to DNA. Although Ma et
al. did not quantify histopathological changes observed in the hippocampus (Section 3.4), they
quantified immunohistochemistry staining of glial fibrillary acidic protein, in addition to caspase-3 in
the hippocampus CAi region and cerebral cortex. Staining intensity of caspase-3 and glial fibrillary
acidic protein was increased at 200 mg/kg-day in both regions of the brain and increased in the cerebral
cortex at the 20 mg/mg-day dose.

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Both studies also evaluated the combined effects of the highest tested dose of DINP in addition to
vitamin E or melatonin {i.e., 150 mg/kg-day + 50 mg/kg-day vitamin E (Pens. 2015); or 200 mg/kg-day
+ 50 mg/kg-day melatonin (Ma et at.. )). Mice exposed to 200 mg/kg-day DINP plus 50 mg/kg-day
melatonin had less caspase-3 and glial fibrillary acidic protein staining than DINP alone, indicating that
melatonin can rescue the increase in caspase-3 and glial fibrillary acidic protein expression that follows
DINP exposure. The addition of melatonin was also sufficient to attenuate the effects consistent with an
oxidative stress response (i.e., increases ROS, DNA-protein-crosslinks, 8-OH-dG, cytokines; decreases
in superoxide dismutase activity and GSH content), implying that DINP induces oxidative stress in the
cerebral cortex which contributes to neuronal damage (Ma et ai. 2015). Similarly, Peng et al (2015)
observed that combined exposure of DINP + vitamin E, which has antioxidant properties, attenuated
effects consistent with an oxidative stress response, implying that the observed effects were consequent
to a pro-oxidant cellular environment in the brain.

In Duan et al (2018). specific pathogen free male Balb/c mice were divided into several groups designed
to evaluate the impact of DINP on an allergic response to an ovalbumin (OVA) antigen. The authors
also investigated the modulatory effect of melatonin, which they state has antioxidant properties, as well
as the role of nuclear factor kappa B (NFkB) signalling and oxidative stress using an inhibitor of NFkB,
Dehydroxymethylepoxyquinomicin (DHMEQ). DINP exposure exacerbated effects consistent with an
oxidative stress response in brain homogenates (i.e., increase in ROS levels and decreases in superoxide
dismutase activity). DINP also increased IL-1B and IL-17 levels in brain homogenates as well as nerve
growth factor (NGF) staining in the prefrontal cortex; all of which were attenuated by the combined
exposure of DINP + metalonin or DINP+ DHMEQ, suggesting that the inflammation is mediated by a
pro-oxidant environment and activation of NFkB signalling. Other endpoints in this study included:
brain histopathology of pyramidal cells in the prefrontal cortex, and immunohistochemistry staining in
the prefrontal cortex for eosinophil cationic proteins, nuclear factor erythroid 2-related factor 2 (Nrf2),
NFkB. Limitations include lack on quantitative results for histopathology.

The other identified study provides a diverse set of data evaluating sexually dimorphic gene expression
in relation to effects on sexual behavior in rodentsU.cc et al.. 2006b). Lee et al. (2006b) investigated the
effects of perinatal exposure to DINP on expression of sex-steroid-regulated genes in the hypothalamus
of offspring and sexual behaviors as adults. Pregnant rats were administered 40, 400, 4,000, or 20,000
ppm DINP in the diet from GD 15 through PND 21. At PND 7, male and female pups were sacrificed,
and the hippocampus was dissected from brains to quantify expression of sexually dimorphic genes such
as granulin (grri) and pi30. After maturation, the authors evaluated and sexual behaviors (e.g., lordosis,
copulatory behavior), reproductive endpoints (e.g., estrus cycles, serum levels of estradiol, LH, and
FSH); these data are discussed in detail in Section 3.1. In male PND 7 pups, there was no change in
hypothalamic grn expression, and a non-monotonic dose response was observed in pi30, but expression
was increased at all dose levels. In females, grn was increased in the 40 and 400 ppm, and 20,000 ppm
exposure groups, and no change was observed in pi30. While the increased pi30 expression in males
coincided with impaired male sexual behavior (i.e., decreased copulatory behavior), serum hormone
levels (i.e., testosterone, FSH, LH) were not changed. The authors suggest that DINP may act on regions
of the hypothalamus that alter sexual behavior, but not gonadotropin secretion, to influence sex-specific
adult behavior.

Conclusions on Neurotoxicity

Fifteen studies in experimental animal models have evaluated neurotoxicological endpoints (i.e.,
behavior, brain weight, or histopathology) following exposure to DINP. However, only three of these
were specifically designed to evaluate behavioral neurotoxicity, which typically may provide insight

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into more sensitive effects of DINP and supplement the neurobehavioral data from the epidemiological
database.

Two short-term duration exposure studies with similar designs in male Kunming mice (Ma et ai. 2015;
Pens. 2015) provide consistent evidence for impaired learning and memory following DINP exposure
for 9 or 14 days, with parallel perturbations in the pyramidal cells of the hippocampus at doses up to 200
mg/kg-day. The developmental exposure study by Bob erg et al. (2011) exposed rats to doses up to 900
mg/kg-day from GD7 to PND17 and conducted behavioral examinations at later timepoints. No
evidence of impairment was observed in males or females (2 to 3 months for MWM; radial arm maze
performance at 5 to 7 months). One consideration regarding the study design in Bob erg et al. (2011) is
that a considerable amount of time had elapsed between the cessation of exposure and time of outcome
evaluation, which could make it more difficult to detect an exposure-related effect {i.e., bias towards the
null), and this difference makes a direct comparison to the studies by Ma (2015) and Peng (. )
challenging. However, this design also helps determine the extent to which perinatal exposures influence
behavior later in life. Nevertheless, discordant results across these studies may reflect study design
differences that influence the degree to which the received dose influences the test animals. Moreover,
Ma et al., (2015) and Peng et al., (2015) exposed adult male Kunming mice and measured outcomes in
adults, while Bob erg et al. (2011) exposed pregnant rats and evaluated outcomes in the offspring. In
addition to the inconsistent findings across study designs, a limitation of the behavioral dataset is the
relative lack of studies that consider outcomes in both sexes, especially given the fact that performance
in the MWM test is sexually dimorphic.

Although histopathological alterations were observed in the pyramidal cells of the hippocampus in two
independent short-term exposure duration studies by Ma et al., (2015) and Peng et al., (2015). these
studies were limited by the lack of quantitative data and were inconsistent with findings of the 28-day
exposure study by the Midwest Research Institute ( [) as well as all the chronic exposure studies in
mice (Covance Labs. 1998b) and rats (Covance Labs. 1998c; Lington et al.. 1997). Strengths of the
dataset include coherence with the behavioral datasets from the Ma et al. (2015) and Peng et al. (2015)
studies; pyramidal cells of the hippocampus are involved in learning and memory, and the mechanistic
dataset from these studies provides evidence of biological plausibility via a mechanism involving ROS
damage by DINP to the pyramidal neurons. Limitations of the dataset include lack of quantitative results
for incidence and severity of histopathology effects and lack of chronic exposure studies with
histopathology of neural tissues.

Overall, available laboratory animal studies provide some evidence that DINP may cause behavioral
effects in rodents. Although some uncertainty exists, EPA considered neurotoxicity further for dose-
response analysis in Section 4. Specifically, neurobehavioral endpoints from Ma et al., (2015) and Peng
et al., (2015) were further considered.

3.5 Cardiovascular Health Effects

Humans

Health Canada (2018a) evaluated multiple studies that investigated the association between phthalate
exposure and several cardiovascular outcomes and/or associated risk factors {i.e., cholesterol, diastolic
and systolic blood pressure, HDL-cholesterol, LDL-cholesterol, and blood glucose levels), however only
two studies directly looked at evidence of an association between DINP and/or its metabolites and
cardiovascular effects. A cross-sectional study of good quality by Trasande et al. (2014) looked at
albumin/creatinine ratio (ACR), a biomarker of endothelial dysfunction and increased risk of CVD in
children and adolescents found that there was inadequate evidence for an association between ACR and
MCOP in children and adolescents (Health Canada. 2018a).

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New Literature: EPA identified three new medium qualtiy studies that evaluated the association between
urinary DINP levels of metabolite and cardiovascular effects. The first medium quality study, a
prospective birth cohort study, by Heggeseth et al. Q ), used data from the Center for the Health
Assessment of Mothers and Children of Salinas (CHAMACOS) cohort to assess the association between
prenatal urinary DINP measurements and BMI trajectories throughout childhood. The authors did not
report any significant results, however, functional principal components analysis found that MCOP was
an explanatory variable in variation of BMI trajectories among girls.

Another medium quality study, a cross-sectional and longitudinal analysis, by Diaz Santana et al.
(2019). of participants from a nested case-control included using data from the Worn ens Health
Initiative (WHI) evaluated the association between overweight and obesity as well as weight change and
DINP exposure. The study found no significant results in cross-sectional analyses by quartile of
exposure. However, there was significant association across quartiles with MCOP and overweights as
well as obese, with p-trend <0.001 and p-trend = 0.001 respectively.

Finally a medium quality study, a longitudinal cohort study, by Zettergren et al. (2021). examined
associations between DINP metabolites (MHiNP, MOiNP, MCiOP) and obesity measures through age
24y in a subset of participants in the Swedish Abbreviation for Children, Allergy, Milieu, Stockholm,
Epidemiology (BAMSE) cohort. The study found significant associations between increases in DINP
metabolites at age 4y and obesity measures obtained at ages 8 and above. Urinary MHiNP, MOiNP,
MCiOP and DINP measures were significantly associated with an increased odds of overweight at ages
8, 16, and 24 years, and with higher BMI [beta = 1.60 (95% CI: 0.37-2.84) , waist circumference [beta =
4.42 (95% CI: 1.35-7.49)], body fat percent [beta = 2.65 (95% CI: 0.52-4.77)], and trunk fat percent
[beta = 2.70 (0.33-5.07)] at 24 years. The cross-sectional association between DINP metabolites and
obesity at age 4 were not significant.

Laboratory Animals

A limited number of experimental animal studies have evaluated the cardiovascular effects of DINP
following oral exposure. Existing assessments of DINP have not drawn human health hazard
conclusions on the cardiotoxicity of DINP. Nevertheless, data are available on the effects of DINP on
blood pressure, heart rate, other indicators of adverse cardiac events, heart weight and/or heart
histopathology (I c- « ^ \ :023c; Ml ^ V. i PSC. 2010: ECB. 2003). Only one study was

available that was specifically designed to evaluate cardiotoxicity (Deng et al.. 2019). Remaining studies
evaluated heart weight and/or heart histopathology (Kwack et al.. 2009: Lington et al.. 1997:
Bio/dynamics. 1982b: Midwest Research Institm [))• No studies are available for the dermal or
inhalation exposure routes.

Three studies of varying study designs were identified that provide data on the effect of DINP exposure
on heart rate, blood pressure, or other indicators of adverse cardiac events, including levels of total
cholesterol and triglycerides. An subchronic duration study by Deng et al. (2019) investigated the
mechanisms associated with increased blood pressure following exposure to DINP. Groups of C57/BL6
mice were administered 0, 0.15, 1.5 or 15 mg/kg-day DINP via oral gavage daily for 6 weeks. At study
termination, systolic blood pressure, diastolic blood pressure, mean blood pressure, and heart rate were
measured. Additionally, blood samples were collected for measurements of serum nitric oxide levels and
levels of angiotensin converting enzyme (ACE), angiotensin-II type 1 receptor (AT1R), and endothelial
nitric oxide synthase (eNOS), were evaluated via immunohistochemistry staining. Increased systolic,
diastolic, and mean blood pressure was observed in mice of the two highest dose groups (1.5 and 15
mg/kg-day). Immunohistochemistry of the aorta showed increased staining intensity of ACE and AT1R

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as well as decreased staining intensity of eNOS and nitric oxide. These latter endpoints are discussed
more in detail in the mechanistic section.

Two additional studies are available that provide data on changes in triglycerides and cholesterol
following short-term duration exposure (Kwack et al. 2009) to DINP. Kwack et al (2009) exposed male
SD rats to 0 or 500 mg/kg-day DINP daily for 4 weeks via oral gavage and evaluated several
cardiovascular outcomes including serum levels of total cholesterol and triglycerides. Serum
triglycerides were significantly increased (50 percent increase compared to controls), while no change
was observed in serum total cholesterol.

Four studies were identified that provide data on the effect of DINP on heart weight, including one
short-term exposure duration study in male SD rats (Kwack et al.. 2009). one short-term study in male
and female F344 (Bio/dynamit 2b), and one chronic study in male and female F344 rats (Lington
et al.. 1997). In general, no statistically or biologically significant {i.e., more than 10 percent change)
exposure-related changes were observed in absolute or relative heart weight across study designs.

Two studies were identified that report histopathology of the heart and/or aorta following exposure to
DINP. The subchronic study in male mice by Deng et al. (2 ) also evaluated histopathology of the
heart and aorta. Lesions were observed in the high-dose group (15 mg/kg-day), including ventricular
wall thickening and cardiomyocyte hypertrophy. In contrast, the study by the Midwest Research
Institute (1981) did not observe discernable lesions in the heart at study termination. In this study, male
and female F344 rats were exposed to 0, 0.2, 0.67, or 2 percent DINP for 28-days via feed (estimated
doses: 0, 150, 500, 1,500 mg/kg-day [males]; 0, 125, 420, 1,300 mg/kg-day [females]). A limitation of
these studies is that histopathology was reported qualitatively.

Table 3-10. Summary of Study Evaluating Cardiovascular Outcomes

Brief Study
Description
(Reference)

NOAEL/
LOAEL
(mg/kg-day)

Effect at LOAEL

Comments

C57/BL6 mice (males
only); oral gavage; 0,
0.15, 1.5, 15 mg/kg-
day; 6 weeks; with or
without induction of
hypertension (Deng et
al. 2019) (Deng et al.,
2019)

0.15/15

t in systolic, diastolic,
and mean blood pressure;
ventricular wall
thickening &
cardiomyocyte
hypertrophy;

immunohistochemistry of
aorta showed |ACE &
AT1R& jeNOS&NO.

15 mg/kg-dav: t Heart Rate and diastolic blood
pressure. Pathological changes in the heart,
aorta, and kidney

Kidnev histopathology (qualitative only): Study
authors also state that "DINP exposure and
DEXA treatment could both induce vacuoles
and hyaline degeneration in the glomerulus as
compared to the saline group. We also found
that DINP exposure resulted in smaller
glomeruli and a thickened glomerular basement
membrane, and that ACEI effectively inhibited
these lesions." Doses at which this occurred are
not stated.

Mechanistic Information

EPA identified one in vivo study (Dene et al.. 2019) that provides data that may inform mechanisms of
the observed cardiovascular effects of DINP. The mouse study by Deng et al. (2019) investigated
mechanisms associated with increased blood pressure following exposure to DINP. Groups of C57/BL6
mice were exposed to 0, 0.15, 1.5, or 15 mg/kg-day DINP daily for 6 weeks via gavage. Parallel groups

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of mice also received a subcutaneous injection of 1 mg/kg-day dexamethasone to induce hypertension
and/or 5 mg/kg-day of an ACE inhibitor via gavage in addition to the highest dose of DINP. In addition
to the evaluations of blood pressure described above, the authors measured serum nitric oxide (NO)
levels and determined levels {i.e., staining intensity) of ACE, AT1R, and eNOS in the aorta via
immunohistochemistry staining. The authors observed increased staining intensity of ACE and AT1R as
well as decreased staining intensity of eNOS in the aorta using immunohistochemistry following
exposure to 1.5 or 15 mg/kg-day DINP (AT1R and eNOS) or all doses (ACE). Co-exposure of
15 mg/kg-day DINP and dexamethasone resulted in similar changes in expression of ACE, AT1R, and
eNOS. Co-exposure of dexamethasone +15 mg/kg-day DINP + the ACE-inhibitor did not fully
attenuate the changes. Serum levels of NO were decreased following DINP exposure (all doses) as well
as with co-exposure to dexamethasone and/or the ACE inhibitor. Given the aforementioned increases in
systolic, diastolic, and mean blood pressure, in mice of the two highest dose groups (1.5 and 15 mg/kg-
day), these results provide some evidence to support a mechanism whereby DINP acts through the ACE
pathway to increase blood pressure.

Conclusions on Cardiovascular Health Effects

The database of studies in experimental animals that has evaluated cardiovascular toxicity and
associated risk factors following exposure to DINP is limited and findings were generally inconsistent
across study designs and species. Only one subchronic study was available that was specifically
designed to evaluate cardiotoxicity (Dene et al. ). Limitations of the study included failure to
consider both sexes and reporting deficiencies, including the qualitative reporting of histopathology
data. Nevertheless, the consistency across endpoints within Deng et al. (2019). including increased
blood pressure and histopathological effects in the aorta suggest that DINP may be toxic to the
cardiovascular system. Mechanistic data from the same study suggest the underlying mechanism for
these effects involves the ACE pathway.

Overall, there is limited evidence that DINP can elicit cardiotoxicity in experimental laboratory animals;
only one study in one species of one sex evaluates cardiovascular outcomes. Additionally, the clinical
implications, or relevance to humans, is uncertain for cardiovascular effects of DINP. Due to these
limitations and uncertainty, EPA is not further considering cardiotoxicity for dose-response analysis.

3.6 Immune System Toxicity

Humans

Health Canada (2018a) evaluated multiple studies that investigated the association between urinary
metabolite and immunological outcomes. Across available studies of DINP, Health Canada found that
there was limited or inadequate evidence for association between DINP and its metabolites and
immunological outcomes.

New Literature: EPA identified three new studies (two medium quality studies and one low quality) that
evaluated the association between DINP and its metabolites and immune/allergy outcomes. The first
medium quality study, a prospective birth cohort, by Soomro et al. (2018). of the Etude des
Determinants pre et postnatals du developpement de la sante de l'Enfant (EDEN) study measured
maternal urinary DINP metabolites and their association with eczema diagnosed at ages 1-5 in boys, and
with elevated serum IgE at age 5 years. Results for the main effect association between DINP metabolite
and elevated IgE were described only as not significant for MCOP. There were no significant
associations found with DINP metabolites and elevated serum IgE (>60 IU/mL). However, multivariate
logistic regression of MCOP and odds of diagnosed eczema was only significant for age 5, OR = 1.60
(95% CI: 1.16, 2.23). There was a significant association found in multivariate logistic regression of
MCOP and association with early onset eczema (first 2 years of life), OR = 1.29 (95% CI: 1.04, 1.60), p

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< 0.05, and late-onset (age 3-5 years) eczema, OR = 1.63 (95% CI: 1.20, 2.21), p < 0.05. There was also
a significant association in Cox proportional hazard model of MCOP and ever diagnosed with eczema,
HR= 1.09 (95% CI: 0.95, 1.25), p = 0.05.

Another medium quality study, a cross-sectional study, by Ait Bamai et al. (2018). that used data from
Hokkaido study on Environment and Children's Health examined the association between DINP and
eczema within the past 12 months. Logistic regression of DINP (j_ig/g dust) exposure on eczema found
significant gene-environmental interaction with FLG mutation, OR total = 1.17 (95% CI: 0.91, 1.52; p =
0.039). No other significant associations were found between eczema and DINP exposure.

Finally, a low-quality study, a cohort study, by Wan et al. (2021). that used data from the Kingston
Allergy Birth Cohort (KABC) examined the association between skin prick testing and DINP exposure.
The authors did not find any statistically significant results in adjusted logistic regression models for
DINP exposure relation to allergic sensitization.

Laboratory Animals

A limited number of studies are available that have been evaluated for the toxicological effects of DINP
on the immune system. Available studies have provided data on the adjuvant properties of DINP; an
adjuvant is a substance that can enhance immune responsiveness without itself being an antigen. ECB
(2003) summarized the irritation and sensitization data and determined that DINP is a very slight skin
and eye irritant, with effects reversible in short time. The U.S. CPSC ( ) concluded that"in vivo
studies in guinea pigs suggest that DINP is not a skin sensitizer"; however, "in vivo studies in mice show
that DINP or other o-DAP's may augment an antigen mediated IL-4, IgE, and/or IgGl reaction." These
finding suggest that DINP may potentiate allergic and/or asthmatic responses.

The database of studies from existing assessments that evaluate the immune adjuvant effects of DINP is
limited to two studies (Koike et al.. 2010: Iroai et al.. 2006). which investigate the effects of DINP on
atopic dermatitis and skin sensitization.

Koike et al. (2010) investigated the effect of DINP on atopic dermatitis resulting from contact with a
dust mite allergen. Male NC/NgaTndCrlj mice were injected intradermally on the ventral side of their
right ears with saline or extract of the dust mite, Dermatophagoides pteronyssinus (Dp) on study days 0,
3, 5, 8, 10, 12, 15, and 17. On study days 2, 5, 9, and 16, DINP was administered via intraperitoneal (i.p)
injection dose levels: 0, 0.15, 1.5, 15, or 150 mg/kg-day. The authors evaluated several endpoints
including histopathology of the ears, protein expression (from ear homogenates) of Thi-type versus Tin-
type cytokines , as well as chemokines such as eotaxin, eotaxin-2, and thymic stromal lymphopoietin
(TSLP), via ELISA. DINP exposure significantly increased ear thickening and macroscopic features of
the ears from 4 and 6 days after the first injection of Dp. However, no dose-dependent effects of DINP
were observed. Animals exposed to 15 mg/kg-day DINP + Dp had more skin lesions when compared to
animals exposed to Dp or saline (no Dp). Histopathological evaluation of the ears showed that while Dp
had increased infiltration of eosinophils into the skin lesions when compared with saline controls, 15
mg/kg-day DINP + Dp potentiated the infiltration of eosinophils into the skin lesion (compared to Dp)
in parallel with increased mast cell degranulation. Alterations in cytokine levels were observed in the
ears of animals exposed to Dp (compared to saline), including increased IL-4, -5, and -13 and decreased
interferon-y (IFN-y). There was a decrease in expression of IFN-y, eotaxin and eotaxin-2, and increased
expression of TSLP were also observed in the ears of mice exposed to DINP, compared to those exposed
to Dp + vehicle. These data suggest that DINP aggravates allergic dermatitis-like skin lesions caused by
the Dp antigen. To evaluate the adjuvant capacity of DINP for immunoglobulin (Ig) production, the
authors also measured serum levels of anti-DP-IgGi, IgE, as well as histamine release. Intradermal

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injection of Dp increased the levels of Dp-specific IgGl, total IgE, and histamine levels in serum
compared to saline alone. Exposure to DINP significantly increased histamine levels in serum compared
to saline alone. However, no significant changes in serum levels of Dp-specific IgGl, total IgE, or
histamine were observed in groups exposed to DINP compared to Dp. Collectively, these data support
that DINP is not an adjuvant in an atopic dermatitis mouse model.

Imai et al. (2006) investigated whether different phthalate esters (including DINP) have adjuvant effects
on skin sensitization using FITC as a sensitizer. Female CD-I (ICR) and BALB/c mice were used for
this skin sensitization study. Experimental groups include having multiple phthalates mixed with
acetone at a 1:1 ratio and the control group with acetone alone. ICR mice were epicutaneously sensitized
with FITC dissolved in an acetone solution containing one of various phthalate esters, including DINP.
The applications on the forelimbs were repeated on day 7 and on day 14; ear thickness and ear swelling
were measured. There were no significant differences in ear thickness/swelling between the DINP
treated group compared to the acetone control group. Similar results with DINP were confirmed using
BALB/c mice. Twenty-four hours following skin sensitization, draining lymph node cells were
examined for FITC fluorescence by means of flow cytometry. Mice sensitized with FITC in acetone
containing DINP did not show consistent ear-swelling response. DINP also showed no significant
increase in the FITC-positive cell number in the draining lymph nodes. These data suggest that DINP
does not act as an adjuvant in a FITC skin sensitization model in mice.

New Literature: EPA identified two new studies that investigated the effects of DINP exposure on
atopic dermatitis (Wu et al.. 2015; Sadakane et al.. 2014).

Wu et al. (2015) investigated the effects of DINP on allergic dermatitis (AD) in a FITC-induced allergic
dermatitis model and the role of oxidative stress and inflammatory factors in skin lesions of the model
mice and characterize the mechanism involved in the DINP. Additionally, uncovering the protective role
of melatonin (MT) on AD and exploring its mechanism as an antioxidant. Forty-nine male Balb/c mice
were divided randomly into seven groups: control, melatonin (30 mg/kg-day) 3 h after saline skin
exposure, 0.5 percent FITC-sensitized group (FITC), 1.4 mg/kg-day) DINP skin exposure+0.5 percent
FITC-sensitized group (FITC + DINP1.4), 14.0 mg/kg-day DINP skin exposure+0.5 percent FITC-
sensitized group (FITC+DINP 14), 140.0 mg/kg-day DINP skin exposure+0.5 percent FITC sensitized
group (FITC+DINP 140), and MT (30 mg/kg-day) 3 h after 140.0 mg/kg-day DINP skin exposure
combined with 0.5 percent FITC sensitized group (FITC+DINP 140.0+MT). The mice were exposed for
40 days, then given saline or FITC on days 41 and 42. Sensitization was terminated on day 47 to
measure ear thickness. This experiment was terminated on day 48 and blood samples were collected to
measure IgE levels and immunohistochemistry were conducted on the sections from the right ear for
TSLP, p-STAT3, p-STAT5, p-STAT6, NF-kB, and p65. Markers of oxidative stress, including ROS,
MDA, GSH, along with cytokines, IL-4 and IFN-y, were evaluated from the ear tissue.

The highest concentration of DINP (140 mg/kg-day) with FITC significantly increased the number of
infiltrating inflammatory cells when compared with the FITC exposed only group. Moreover, the
pathological alterations and the number of infiltrating inflammatory cells were alleviated in the
FITC+DINP 140+MT group as compared with the FITC+DINP 140 group. Ear swelling and bilateral
ear weight were significantly altered in all FITC-immunized groups. Dermal DINP exposure
significantly increased ear swelling and bilateral ear weight when compared to the group exposed to
FITC only, and this adverse effect was potentiated. Also, when MT was added, diminished the DINP-
induced ear swelling and the bilateral ear weight when compared to the same concentration of DINP
without MT. FITC alone and all concentrations of FITC+DINP exposure significantly enhanced serum T
IgE levels, at all concentrations. The highest dose of DINP (140 mg/kg) exposure drastically elevated
serum T-IgE levels compared with the FITC-sensitization only group. Further, T-IgE levels in the FITC

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+ DINP 140 group significantly decreased when compared to the FITC+DINP 140+MT group.
Compared with the FITC only group, co-exposure with any concentration of DINP induced a significant
increase in IL-4, IL-5 and a resulting skew in the ratio of IL-4 to IFN-y. These adverse effects
exacerbated by DINP were concentration-dependent. However, MT alleviated the DINP-induced effects,
suggesting that DINP is associated with Th2 cytokine expression by FITC-mediated allergic
inflammation. Their results of histopathological examinations and measurements of ear swelling as well
as immunological and inflammatory biomarkers (total-immunoglobulin IgE and Th cytokines) supported
their conclusion that high doses of DINP may aggravate atopic dermatitis.

Lastly, Sadakane et al. (2014) is another study identified by the EPA that investigated the role of DEHP
and DINP on atopic dermatitis at doses lower than the NOAEL for humans based on chronic liver
toxicity (15 mg/kg-bw-day); however, the DINP-specific effects will be focused on here. Previous
studies have uncovered that DINP in low doses have been shown to cause aggravation of atopic
dermatitis-like skin lesions (ADSLs) in mouse models. In this study, 120 male NC/Nga mice were used
in this experiment, out of which 60 mice each were used to investigate the effect DINP on AD. From the
60 mice used for DINP exposure, they were placed in 5 groups of 12 (1 for saline vehicle control and 4
experimental groups). Animals in the experimental groups were exposed to the allergen,
Dermatophagoidespteronyssinus (Dp), by subcutaneous injection of 5 mg of dissolved in 10 mL of
saline in the ventral side of the right ear for 2 to 3 days a week (a total of 8 times) under anesthesia.
Animals in the experimental DINP groups were exposed to the allergen and treated with 0 (Dp+vehicle),
6.6 (Dp+DINP 6.6), 131.3 (Dp+DINP 131.3), or 2,625 (Dp+DINP 2,625) jig/animal of Dp. In the
experimental groups, mice were orally administered DINP dissolved in O.lmL of olive oil 5 days before
the first injection of the allergen. Control group animals (saline + vehicle and Dp+vehicle groups) were
not exposed to DINP and only given 0.1 mL of olive oil only orally.

Twenty-four hours following Dp injections, skin disease symptomatology and ear thickness were
evaluated and scored for symptom of skin dryness and eruption, edema, crusting and erosion. Also, the
clinical scores of the Dp+DINP 6.6 and Dp+DINP 131.3 groups began increasing when compared with
the Dp+vehicle group from day 16, the Dp+DINP 131.3. The Dp+DINP 131.3 group had a higher (not
significant) wound score compared with the Dp+vehicle group while the Dp+DINP 2,625 did not
change. Statistical tests revealed no significant differences between DINP treated groups and the control
at any doses to contribute to ASDLs. The dorsal skin of the Dp-treated groups with or without DINP
exposure exhibited epidermal and dermal thickening, eosinophil accumulation and mast cell
degranulation. The eosinophil counts of both DP+DINP treatments increased but not significantly.
However, oral exposure to DINP did not increase the eotaxin levels. Exposure to DINP modestly
increased mean total IgE levels. The rank of mean skin scores with specific DINP doses (Dp+DINP
131.34 > Dp+DINP 6.64 > Dp+DINP2,625 > Dp+vehicle) was found to be strongly positively
correlated with the number of eosinophils, the number of severely degranulated mast cells, and
moderately positively correlated with the total number of mast cells. In conclusion, at doses lower than
the NOAEL, DINP increases the allergic response in animal AD models, but the other concentrations of
DINP slightly aggravates allergen-induced ADSL production.

Mechanistic Information

EPA identified seven studies that describe the mechanism of action for the adverse immunological
effects of DINP (Yun-Ho et al.. 2019; Duan et al.. 2018; Kane et al.. 2017; Kane et al.. 2016; Chen et
al.. 201 ^ Koike et al.. 2010; Lee et al.. 2004).

Aforementioned Koike et al. Q ) not only conducted experiments in mice, but also evaluated the
adjuvant effects of DINP on bone-marrow-derived dendritic cells or splenocytes in vitro. Bone-marrow-

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derived dendritic cells and splenocytes were exposed to DINP for 24 hours at concentrations of 0
(control), 0.1 [xM, 1 [xM, 10 [xM, and 100 [xM. At 100 [xM, DINP exposure for 24 hours led to
significantly increased the production of Th2 chemokines, TARC/CCL17 and MDC/CCL22, in bone-
marrow-derived dendritic cells when compared with control (0 [xM DINP). However, Thi cytokine IL-
12p40 was not detected in any bone-marrow-derived dendritic cell culture. Moreover, DINP also
significantly increased the expression of the chemokine receptors CCR7, CXCR4, MHC class II, CD80,
and CD86 on bone-marrow-derived dendritic cells compared with controls. DINP exposure for 24 hours
significantly increased IL-4 production from splenocytes compared with controls. After 72-hours of
exposure to DINP in the presence of Dp, there was a significant increase in proliferation of splenocytes
at 0.001 to 1 [xM and decreased proliferation at 10 [xM compared with controls. These results show that
DINP augmented IL-4 production and Dp-stimulated proliferation of splenocytes to suggest that DINP
does aggravate AD-like skin lesions related to Dp through TSLP-related activation of dendritic cells and
by direct or indirect activation of other immune cells.

Kang et al. C ) examined the effects of DINP exposure on the development of allergies and the
underlying mechanisms. Male Balb/c mice were gavaged with 2, 20, or 200 mg/kg-day DINP for 21
days, then sensitized with either saline or 0.5 percent FITC (in 1:1 acetone/DBP) on days 22 and 23 via
dermal application to shaved skin. On day 28, the mice received a 0.5 percent FITC challenge (or saline)
to the right ear, and saline or vehicle (1:1 acetone/DBP) to the left ear and the baseline ear thickness was
measured. On day 29, the study was terminated, and blood samples were collected to determine IgE
levels. Immunohistochemistry staining was performed on the sections from the right ear to visualize the
localization and staining intensity of TSLP, p-STAT3, p-STAT5, p-STAT6, NF-kB, and p65. The
authors also evaluated ROS, MDA, and GSH levels in the ear tissue as well as levels of the cytokines,
IL-4 and IFN-y. In mice administered DINP+FITC, there was an increase in the number of infiltrating
inflammatory immune cells in their ear tissue. Dose-dependent, significant increases in IL-4 and IL-5
were observed in all groups exposed to FITC+DINP. In contrast, there was a dose-dependent decrease in
IFN-y, which increased the IL-4/IFN-y ratio, showing DINP only increases Th2-specific cytokines.
However, no significant pathological changes were observed in the ears of mice exposed to DINP alone,
but the ears of mice from the FITC only group showed inflammatory cell infiltration into the skin.
Additionally, to uncover the pathway of these adverse effects, treatment with FITC+DINP200, and
pyrollidine dithiocarbamate (PDTC), a well-known inhibitor of NF-kB, markedly reduced the ear
swelling when compared to the FITC+DINP200 exposed group. Further, bilateral ear weight decreased
significantly when the FITC+DINP-immunized groups were treated with PDTC. There was an increase
in ROS and MDA levels and a decrease GSH levels were observed in FITC+200 mg/kg-day DINP
exposure groups compared to FITC alone, but PDTC reversed those effects. The adverse pathological
effects observed in higher dose groups were attenuated with PDTC treatment, which suggest that the
adverse effects are facilitated by the NF-kB signalling pathway. Results support that DINP aggravates
FITC-induced allergic contact dermatitis through exacerbating increased MDA and ROS accumulation,
IL-4 and IL-5 production, while also decreasing GSH and IFN-y, which then activates the NF-kB
pathway. Following activation, TSLP expression and activation is increased, causing increased
production of STATs 3, 5, and 6 to ensue.

A subsequent study by Kang et al. (2017) expanded on the previously mentioned underlying
mechanisms of DINP and the role of TRP cation channel, subfamily A, member 1 (TRPA1) on the NF-
kB pathway. In this allergic dermatitis model, male BALB/c mice were gavaged with saline (control) or
DINP (2, 20, 200 mg/(kg-d) from day 1 to 21. On days 22 and 23 mice were smeared with saline or 0.5
percent FITC on their backs to sensitize them, then on day 28, mice are given saline or FITC on their
right ear. Following sensitization, skin lesions showed enhanced levels of IgGl, IL-6, IL-13, and
TRPA1 expression with DINP potentiating these levels. To determine the role of TRPA1 and NF-kB for

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allergic dermatitis, on days 22, 23, and 28, mice were injected with HC-030031, a TRPA1 antagonist,
and NF-kB inhibitor, PDTC. Blocking NF-kB inhibited TRPA1 expression; however, TRPA1
antagonism did not have any effect on NF-kB or TSLP expression. These findings suggest that TRPA1
is dependent on NF-kB activation and TSLP expression for DINP aggravated allergic dermatitis.

Similarly, Lee et al. (2004) examined the effects of DINP on IL-4 production in CD4+ T-cells and the
associated mechanisms. BALB/c mice were injected with Keyhole limpet hemocyanin in alum adjuvant
twice at 7-day intervals while being i.p injected with 2 or 5 mg/kg of DINP every other day. Lymph
node cells were harvested and cultured from these mice after 7 days of treatment and used to measure
IL-4 and IFN-y. DINP was shown to enhance IL-4 production in lymph node cells, which originated
from CD4+ T-cells in a concentration dependent manner and increase IgE serum levels in vivo.
Additionally, DINP exposure also increased IL-4 gene promotion activity in Phorbol-12-myristate-13-
acetate stimulated EL4 T-cells. IL-4 gene promoter contains multiple binding sites to nuclear factor of
activated T-cells (NF-AT), and DINP was shown to potentiate IL-4 production via enhancing PI and P4
binding site activity on NF-AT. These study results support that DINP augments the allergic response of
IL-4 production in CD4+T-cells via increased NF-AT binding activity.

Moreover, Chen et al. (2015) investigated how DINP exposure during gestation and lactation affects the
allergic response of pups and the role of the PI3K/Akt pathway. Female Wistar rats are treated with 0, 5,
50, and 500 mg/kg-day from GD 7 to PND 21. On PND 22, 23, and 37, pups were sensitized with
ovalbumin (OVA). Then, protein expression and production of cytokines associated with PI3K/Akt were
measured. In the 50 mg/kg-day DINP group, pups displayed significantly increased lung resistance (RI)
when compared to the controls. Moreover, all DINP-treated groups had significantly increased
eosinophil infiltration into the airways when compared to the control group, as indicated by
immunohistochemistry. Pups exposed to 50 mg/kg-day DINP had increased Akt phosphorylation, NF-
kB translocation, and increased Th2 cytokine (IL-13) expression, while having decreased Thi cytokine
(INF-r) expression, when compared to the vehicle control group. These results suggest DINP aggravates
the OVA-induced response and enhances expression of the PI3K/Akt pathway and NF-kB translocation.

Next, a neuroinflammation mouse asthma model study by Duan et al. (2018) exposed, via i.p injection,
groups of Balb/c mice (8 mice/group): 1) Saline only group (control); 2) Ovalbumin (OVA) only group);
3) OVA and formaldehyde (lmg/m3, 5h/day) exposure (OVA+FA group); 4) OVA and DINP (20
mg/kg-day) exposure (OVA+DINP group); 5) OVA and formaldehyde (lmg/m3, 5h/day) plus DINP (20
mg/kg-day) exposure (OVA+FA+DINP group); 6-9) melatonin (lOmg/kg-day) blocking groups
(OVA+MT group, OVA+FA+MT group, OVA+DINP+MT group, OVA+FA+DINP+MT group); 10-
13) were Dehydroxymethylepoxyquinomicin (DHMEQ; a NF-kB inhibitor) (lOmg/kg-day) NF-kB
blocking groups (OVA+DHMEQ group, OVA+FA+DHMEQ group, OVA+DINP +DHMEQ group,
OVA+FA+DINP+DHMEQ group). Following 18 days of exposure and 7 days of sensitization, allergic
asthma symptoms (eosinophilic catatonic protein) levels and mucus secretion), markers of oxidative
stress (ROS fluorescence, superoxide dismutase, and Nrf2 levels), cytokines (IL-ip and IL-17), and NF-
kB signaling were measured in the brain. Exposure to DINP increased eosinophilic catatonic protein
levels and the number of mucus secreting cells in the airway of the mice with OVA sensitization.
Additionally, DINP exposure increased levels of IL-ip, IL-17, and NGF levels in the brain and
increased NF-kB activation in the pre-frontal cortex. Moreover, DINP exposure increased ROS
fluorescence in the brain, Nrf2, and decreased superoxide dismutase. Results of this study indicate that
DINP promotes neuroinflammation through potentiating oxidative stress and NF-kB signal pathway
activation in this mouse asthma model.

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Lastly, another asthma mouse model study identified by EPA is Yun-Ho et al. (2019). They investigate
the role of TLR4 and HMGB1 in the mechanisms of DINP-induced asthma. In this study, female
C57BL/6 mice were i.p injected with 50mg/kg 1 DINP for a week to sensitize them and then challenged
with saline or DINP on days 19, 21, and 23. During the challenge, mice were injected in their tail vein
with either 3 mg/kg 1 TAK-242 (TLR4 inhibitor) or 10 mg/kg 1 anti-HMGBl antibody, respectively, on
each day of the challenge. DINP significantly increased airway hyperresponsiveness, number of
infiltrating cells in bronchoalveolar fluid, numbers of inflammatory cells in blood, pulmonary fibrosis,
mucus production, Th2 cytokine production (IL-4, IL-5, IL-13), and lung cell apoptosis. In contrast,
adding the TLR4 inhibitor or anti-HMGBl antibody following DINP exposure reduces airway
hyperresponsiveness, reduced production of IL-4, IL-5, and IL-13 cytokines, and number of
inflammatory cells in the airway. Therefore, this study supports that HMGB1 and TLR4 signalling
pathways both contribute to DINP-induced asthma and inhibiting them significantly reducing several
biological markers of asthma.

Conclusions on Immune System Toxicity

There are multiple animal toxicity studies that support the adjuvant effects of DINP exposure on the
immune response in dermatitis models and in vitro experiments (Koike et al.. 2010; Iroai et al.. 2006).
Koike et al. (2010) stated that DINP exposure did not aggravate serum levels of IgGl, IgE, and
histamine levels in vivo. Further, Imai et al. (2006) concluded that DINP is not considered a skin
sensitizer based on no significant increase in the FITC-positive cell number in the draining lymph nodes.
Additionally, there were three new studies that all support that DINP aggravates atopic dermatitis via
causing oxidative stress and NF-kB cellular pathway activation (Kane et al.. 2016; Wu et al.. 2015;
Sadakane et al.. 2014). Similarly, EPA identified six mechanistic studies that support DINP enhancing
NF-kB signalling, TSLP transcription, NF-AT, PI3K/Akt, TLR4, and HMGB1 in allergic dermatitis,
atopic dermatitis, and asthma mouse models (Yun-Ho et al.. 2019; Duan et al.. 2018; Kane et al.. 2017;
Kane et al.. 2016; Chen et al.. JO I et al.. 2004). Overall, available studies provide evidence that
DINP augments the inflammatory responses in several sensitization models and the underlying
mechanisms. Specifically, there are several studies that demonstrate DINP's role in potentiating ROS
production, TSLP transcription, PI3K/Akt, TLR4, and NF-kB pathway activation, and Th2 cytokine
production in allergic dermatitis, neuroinflammation, and asthma in animal models.

Although available studies of laboratory animals provide evidence for immune adjuvant effects of DINP
in sensitized animals, EPA is not further considering these effects for dose-response assessment or for
use in extrapolating human risk. Available studies evaluate the adjuvant properties of DINP in
experimental rodent models pre-sensitized by exposure to other compounds (e.g., FITC, ovalbumin).
While these studies may be useful for hazard identification for a specific population (pre-sensitized
individuals), the fact that the outcome evaluated in these studies requires prior exposure to another
chemical precludes its broader applicability.

3.7 Musculoskeletal Toxicity

Humans

Four epidemiologic studies, three cross-sectional and one cohort study examined the association
between DINP urinary levels of metabolites and bone mineral density, Osteoporosis and Vitamin D in
adults, however the evidence was considered inadequate due to inconsistent results (Health Canada.
2018a).

New Literature: EPA considered new studies published since Health Canada's assessment (i.e., studies
published from 2018 to 2019); however, no new studies were identified that evaluated musculoskeletal
injury for DINP and/or its metabolites.

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Laboratory Animals

Hwang et al. (2017) was the only study that investigated the relationship between DINP and osteopenia,
which is characterized by bone loss and deterioration of bone structure leading to fractures. DINP (2, 20,
or 200 mg/kg-day) was administered via intraperitoneal injection to 8-week-old female C3H/HeN
ovariectomized (OVX) mice (5 animals/group), including: a sham-operated control group injected with
PBS; a vehicle treated OVX group injected with PBS; and three DINP groups of 2, 20, or 200 mg/kg-
day. The vehicle and DINP were administered for 6 weeks, and the body weights were recorded weekly.
There was significant increase in body weights of OVX mice compared to sham control mice 6 weeks
after OVX surgery. DINP also significantly increased body weight compared to sham control mice.
DINP-treated mice had significantly reduced uterus weight and decreased tibia and femur lengths. Tibia
weights were decreased in OVX mice and in the DINP-treated mice. However, no differences were
noted in femur weights among the groups. DINP treatment of the normal mice increased the inorganic
phosphorus release. Lactate dehydrogenase was unaffected by OVX or DINP treatments.

Further, tartrate-resistant acid phosphatase activity (bone resorption marker) was significantly increased
in both OVX mice and in the mice treated with 200 mg/kg-day DINP at a similar magnitude over
controls. Bone ALP activity was lower than sham controls in the OVX mice and in the DINP mice
treated with 2 and 20 mg/kg-day; however, bone ALP activity in mice treated with 200 mg/kg-day DINP
was comparable to sham controls, indicating that these decreases were not dose-related. Further, the
microarchitecture of the femur and tibia were affected by OVX and DINP. The bone volume, tissue
volume, bone volume/tissue volume ratio, bone surface, bone surface/tissue volume ratio, trabecular
thickness, and trabecular number were all reduced, while the trabecular pattern factor, structure model
index, and trabecular separation were increased in the DINP-treated mice, although these differences
were not as substantial as in the OVX mice compared to sham controls. Similarly, the bone mineral
density of the femur and tibia was dose-dependently decreased in the DINP-treated mice, but not
decreased to the extent noted in the OVX mice, compared to the sham controls. The authors concluded
that these results indicate that DINP contributes to an increased risk of osteopenia via destruction of the
microarchitecture and enhancement of osteoclastic activity, although it is difficult to conclude as the
mechanism of action is currently unknown.

Conclusions on Musculoskeletal Toxicity

Four epidemiological studies and one study in experimental animals have provided data on the
associations between exposure to DINP and musculoskeletal outcomes such as osteoporosis or
osteopenia. The human evidence was considered inadequate due to inconsistent results across study
designs and not further evaluated by EPA. The animal evidence suggests that DINP can reduce bone
mineral density in female mice. Overall, there is limited evidence that DINP can elicit musculoskeletal
toxicity in experimental laboratory animals; only one study in one species of one sex evaluates
musculoskeletal outcomes. Additionally, the clinical implications, or relevance to humans, is uncertain
given the limitations of the epidemiologic database. Due to these limitations and uncertainty, EPA is not
further considering musculoskeletal toxicity for dose-response analysis.

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4 DOSE-REPONSE ASSESSMENT

EPA is considering four non-cancer hazard endpoints related to liver, kidney, neurological and
developmental toxicity for dose-response analysis as described in the following sections. These hazard
endpoints were selected for dose-response analysis because EPA has the highest confidence in these
hazard endpoints for estimating risk to human health in the non-cancer sections. The effects for liver,
kidney, and developmental effects were consistently observed across multiple rodent species and
durations of exposure and occurred in a dose-related manner. EPA considered liver and developmental
effects observed in experimental animal models to be relevant for estimating risk to human health. Other
non-cancer hazard endpoints considered by EPA {i.e., cardiovascular toxicity (Section 3.5), immune
system toxicity (Section 3.6), and musculoskeletal toxicity (Section 3.7) were not considered for dose-
response analysis due to limitations in the number of studies, unknown MOA and uncertainties that
reduce EPA's confidence in using these endpoints for estimating risk to human health.

EPA considered two approaches, including a NOAEL/LOAEL approach, and benchmark dose modeling
for liver effects and benchmark dose modeling of developmental effects performed by NASEM (2017).
EPA considered NOAEL and LOAEL values from oral toxicity studies in experimental animal models.
Acute, intermediate, and chronic non-cancer NOAEL/ LOAEL and BMDL values identified by EPA are
discussed further in Sections 4.1.1, 4.1.2 and 4.1.3, respectively. As described in Appendix F, EPA
converted oral PODs derived from animal studies to human equivalent doses (HEDs) using allometric
body weight scaling to the three-quarters power (	). In the absence of dermal toxicology

studies, EPA used the oral HED to assesses risks from dermal exposures. Differences in dermal and oral
absorption are corrected for as part of the dermal exposure assessment. In the absence of inhalation
studies, EPA performed route-to-route extrapolation to convert oral HEDs to inhalation human
equivalent concentrations (HECs) (Appendix F).

4.1 Selection of Studies and Endpoints for Non-cancer and Threshold
Cancer Health Effects

EPA considered the suite of oral animal toxicity studies for adverse liver, kidney, neurological and
developmental effects identified during hazard identification (Section 3) when determining non-cancer
PODs for estimating risks for acute, intermediate, and chronic exposure scenarios, as described in
Sections 4.1.1, 4.1.2 and 4.1.3, respectively. EPA assessed relevant non-cancer health effects in these
studies based on the following considerations:

•	Exposure duration;

•	Dose range;

•	Relevance (e.g., what species was the effect in, was the study directly assessing the effect, is the
endpoint the best marker for the toxicological outcome?);

•	Uncertainties not captured by the overall quality determination;

•	Endpoint/POD sensitivity; and

•	Total uncertainty factors (UFs).

The following sections provide comparisons of the above attributes for studies and hazard outcomes
relevant to each of these exposure durations and details related to the studies considered for each
exposure duration scenario.

4.1.1 Non-cancer Oral Points of Departure for Acute Exposures

EPA considered 12 developmental toxicity studies with endpoints relevant to acute exposure duration
(I	), summarized in Table 4-2. The endpoints considered relevant to acute exposure

durations include skeletal and visceral variations, and effects on the developing male reproductive

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system consistent with a disruption of androgen action during the critical window of male reproductive
development in rats. These studies were subjected to dose-response analysis to select the study and
endpoint most appropriate to derive the POD for acute hazard. The dose-response array for these studies
is depicted graphically in Figure 4-1. Although these studies entailed exposure durations that exceeded a
single day, EPA considered endpoints from these developmental toxicity studies for which there is
evidence that they can result from a single exposure day during a critical window of development during
gestation. For example, several studies have demonstrated that a single dose of DBP, which is
toxicologically similar to DINP (! ; S 1 T \ 2023a. b), during the critical window of development {i.e.,
GDs 15.5 to 18.5) is sufficient to disrupt fetal testicular testosterone production and steroidogenic gene
expression. Although analogous single dose studies are not available for DINP, studies of DBP support
the conclusion that effects on the developing male reproductive system may occur following acute,
single dose exposures in rodent models (see Appendix C for further justification).

In two prenatal developmental toxicity studies (Waterman et ai. 1999; Hellwie et ai. 1997). an
increased incidence of fetal skeletal variations (e.g., rudimentary/supernumerary cervical or lumbar ribs)
and urogenital variations (Hellwie et; 7) were observed following exposure during GDs 6 to 15.
rudimentary/supernumerary cervical or lumbar ribs) and urogenital variations were observed following
exposure during GDs 6 through 15. However, the doses at which fetal visceral and skeletal variations
occurred (500 and 1,000 mg/kg-day) were higher than doses in other developmental toxicity studies in
which more sensitive effects of androgen insufficiency were observed. Therefore, EPA did not select
these studies and endpoints because they do not provide the most sensitive robust endpoint for an acute
POD.

The remaining 10 developmental toxicity studies considered by EPA resulted in effects on the
developing male reproductive system consistent with a disruption of androgen action during the critical
window of development. EPA identified this hazard in the Draft Proposed Approach for Cumulative
Risk Assessment of High-Priority Phthalates and a Manufacturer-Requested Phthalate under the Toxic
Substances Control Act ( 323a) and concluded that the weight of scientific evidence

indicates that DINP can induce effects on the developing male reproductive system consistent with a
disruption of androgen action and rat phthalate syndrome. Notably, EPA's conclusion was supported by
the SACC ( 2023b). The exposure durations for these 10 studies ranged from initiation of

dosing at implantation through the day prior to expected parturition (i.e., GDs 7 to 21) as employed in
most guideline studies, to more narrow windows of exposure during gestation in which the phthalate-
specific effects on male rodent offspring are known to occur (e.g., GDs 14 to 18) or extended to
encompass the perinatal period (e.g., GDs 14 to PND3). Observed effects included decreased
steroidogenic gene expression in the fetal testes, decreased fetal testicular testosterone, decreased AGD,
increased NR, effects on fetal Ley dig cells, increased incidence of MNGs, and decreased sperm motility.
LOAELs for these effects ranged from 100 to 1,165 mg/kg-day.

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I HED LOAEL

A HED NOAEL

• Other Doses Tested

4- fetal testicular testosterone production; GD 14-18; rats; Hannas, 2011,
788239"

t MNGs (GD 21); ^sperm motility (PND 90); GD 7-17; rats; Boberg, 2011,
806135

4- male pup body weight; 1s fetal Leydig cell clusters/aggregates; 4,
testicular mRNA levels of Insl3; GD 12-21; rats; Li, 2015, 2807612°

i" testicular mRNA levels of P450scc, GATA-4, and Insl3; GD 13.5-17.5; rats;
Adamsson, 2009, 679859

4- fetal testicular testosterone content and production; GD 7-21; rats; Borch,
2004, 673587

4- maternal body weight gain; male pup nipples/areola retention; testes
malformations; GD 14 - PND 3; rats; Gray, 2000, 678742

4- Fetal testis testosterone production; GD 14-18; rats; Furr, 2014, 2510906

incidence of rudimentary cervical & accessory lumbar ribs; urogenital &
skeletal variations; GD 6-15; rats; Hellwig, 1997, 674193

1s skeletal variations (total skeletal variations and rudimentary lumbar ribs);
GD 6-15; rats; Waterman, 1999, 680201

4 male pup weight (PND 14), 1" MNGs (PND 2) in the testes; GD 12 - PND
14; rats; Clewell, 2013,1325348

•T" incidence of MNGs, 4' fetal testes testosterone (2 hrs post final dose); GD
12-19; rats; Clewell, 2013,1325350

A

A—

¦ •

10	100

HED (mg/kg-day)

1000

Figure 4-1. Dose-Response Array of Studies Considered for Deriving the Acute Duration Non-
cancer POD

Notes: | = statistically significant increase in response compared to controls; [ = statistically significant decrease
in response compared to controls; M = males; F= females; GD = Gestational Day; PND = Postnatal Day; MNGs =
multinucleated gonocytes; HED = human equivalent dose; NOAEL = No observable adverse effect level; LOAEL
= lowest observable adverse effect level.

11 Study included in NASEM (2017) meta-regression analysis and BMD modeling.

In 2017, NASEM (2017) assessed experimental animal evidence for effects on fetal testicular
testosterone following in utero exposure to DINP using the systematic review methodology developed
by the National Toxicology Program's (NTP) Office of Health Assessment and Translation (OHAT).
Based on results from four studies of rats (Li et al.. 2015; Boberg et al.. 2011; Hannas et al.. 2011;
Adamsson et al.. 20091 NASEM found high confidence in the body of evidence and a high level of
evidence that fetal exposure to DINP is associated with a reduction in fetal testosterone in rats. NASEM
reported that the literature search was conducted on August 15, 2016, so it is not clear why another study
measuring decreased fetal testosterone (Clewell et al.. 2013a) was not included in the analysis. NASEM
further conducted meta-regression analysis and benchmark dose (BMD) modeling on decreased fetal
testicular testosterone production data from two medium-quality prenatal exposure studies of rats (Li et
al.. 2015; Hannas et al.. 2011). although no explanation was provided for the fact that results from the
studies by Adamson et al. (2009) and Li et al. (2015) were not presented in the BMD modeling
supporting the final meta-analysis. NASEM found a statistically significant overall effect and linear
trends in logio(dose) and dose, with an overall large magnitude of effect (greater than 50 percent) in its
meta-analysis for DINP (Table 4-1). Further BMD analysis determined BMDLs and BMDL40 values of
49 and 552 mg/kg-day, the 95 percent lower confidence limit of the BMD associated with a benchmark
response (BMR) of 5 and 40 percent, respectively (Table 4-1). EPA has higher confidence in the

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NASEM meta-analysis since it takes into account data from multiples studies. Using allometric body
weight scaling to the three-quarters power, EPA extrapolated an HED of 12 mg/kg-day from the BMDLs
of 49 mg/kg-day. A total UF of 30 was selected for use as the benchmark MOE (based on an
interspecies UF (UFa) of 3 and an intraspecies UF (UFh) of 10).

Table 4-1. Summary of NASEM (2017) Meta-Analysis and BMD Modeling for Effects of DINP in
Fetal Testosterone 11 b

Database
Supporting
Outcome

Confidence
in Evidence

Evidence of
Outcome

Heterogeneity
in Overall
Effect

Model with
Lowest AIC

BMD5 mg/kg-
day (95% CI)

BMD40 mg/kg-
day (95% CI)

" R code supporting NASEM's meta-regression and BMD analysis of DINP is publicly available through GitHub
(httDs://github.com/wachiuDhd/NASEM-2017-Endocrine-Low-DoseV

b NASEM (20.1.7) 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%."

While one of the studies considered in the NASEM meta-analysis (Li et ai. 2015) appears to
demonstrate similar effects on male offspring at lower doses than indicated in many of the other
developmental toxicity studies, EPA did not consider this study further as the sole study on which to
derive the POD because several areas of uncertainty reduced EPA's confidence in the results when
considered independently from the other studies in a meta-analysis. While dose-dependent increases in
testes dysgenesis and decreases in fetal testicular testosterone were noted, this study had limited
statistical power (n = 6). It is also unclear what the study authors considered the broad description of
"testes dysgenesis" to represent, although there is some indication that they are referring to seminiferous
tubule atrophy. Further, effects on male pup body weight were not dose-related, with an essentially flat
dose-response across doses spanning three orders of magnitude. A similar flat dose-response was noted
in the frequency distribution of cluster sizes of fetal Leydig cells, and this endpoint is of uncertain
adversity. Although this study supports EPA's conclusions regarding the endpoint for hazard
identification, there is too much uncertainty in the dose-response in this study to use it quantitatively for
determination of the acute POD.

Two additional developmental toxicity studies not included among the four studies considered in the
meta-analysis by NASEM (Clewell et al. 2013a; Clewell et ai. 2013b; Hamner Institutes for Health
Sciences. 2011) resulted in decreased fetal testosterone production and other effects on the developing
male reproductive system at similar doses (LOAELs from 250 to 307 mg/kg-day and NOAELs from 50
to 56 mg/kg-day) to the BMDLs of 49 mg/kg-day derived from the NASEM meta-analysis. Therefore,
these studies support the selection of the BMDLs of 49 mg/kg-day for the acute POD.

Although several other additional studies were identified for effects on the developing male reproductive
system and specifically for decreased fetal testicular testosterone, they were single doses studies (Furr et
al.. 2014; Borch et al.. 2004; Gray et al.. 2000) with an identified LOAEL of 750 mg/kg-day,
considerably higher than the LOAELs identified in the above studies.

In a dietary study by Lee et al. (2006a). decreased male pup AGD was reported at the lowest dose tested,
40 ppm (estimated to be approximately 2 mg/kg-day). However, several factors reduce EPA's
confidence in this study and its results. First, study authors did not report dam body weight, food intake,
or calculate received doses in units of mg/kg-day, so there is uncertainty related to the achieved doses in
the study. Further, the effect of DINP on male pup AGD normalized to the cube root of bodyweight was

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2481	slight (overall magnitude of effect not reported), and treatment with DBP (a more potent antiandrogen

2482	compared to DINP) at equivalent or higher doses had no effect on male pup AGD once normalized to

2483	the cube root of body weight. This calls into question the significances of the slight change in AGD

2484	observed for DINP. Given these uncertainties, EPA does not consider the study by Lee et al. (2006a)

2485	suitable for use as the acute POD.

2486

2487	EPA selected the BMDLs of 49 mg/kg-day (HED = 12 mg/kg-day) as the acute exposure duration POD

2488	because it is the most sensitive robust endpoint and is based on the NASEM meta-analysis, and it falls

2489	within the narrow range of the NOAELs in two additional developmental toxicity studies, providing

2490	support and confidence in both the effect and the dose at which it occurs.

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Table 4-2. Dose-Response Analysis of Selected I

>evelopmental Studies Considered for Deriving the Acute Non-cancer

POD

Study Details
(Species, Duration, Exposure Route/
Method, Doses |mg/kg-dav|)

Study POD/

Type
(mg/kg-day)

Effect

HED

(mg/kg-
day)

HEC
(mg/m3)
[ppm]

Uncertainty
Factors d

Reference

Wislar-Imamichi rats GD 15 to PND 21;
estimated doses (as reported bv (EC/HC,

2015)) 0, 2, 20, 200, 1,000 mg/kg-day;
(28 days)

LOEL= 2

|AGD & AG1, | in hypothalamic
granulin (grn, females) and pl30
(males) mRNA levels; reduced lordosis
quotient in females

0.473

2.57 [0.150J

UFa = 3
UFH= 10
UFl=10
Total UF = 300

(Lee et al.,
2006a)fe

Pregnant SD rats; oral gavage (corn oil);
0, 10, 100, 500, 1,000 mg/kg-day; GD
12-21

NOAEL= 10

i male pup body weight; | fetal Leydig
cell clusters/aggregates; j testicular
mRNA levels for Insl3

2.36

12.9 [0.75]

UFa = 3
UFh= 10
Total UF = 30

(Li et al., 2015)a

Meta-regression and BMD modeling of
fetal testicular testosterone in rats

BMDL5 = 49

i Fetal testicular testosterone

11.6

63.0 [3.68]

UFa = 3
UFh= 10
Total UF = 30

(NASEM. 2017Y

Pregnant SD rats; oral gavage; 0, 50,
250, and 750 mg/kg-day; GDs 12-19

NOAEL = 50

t incidence of MNGs, | fetal testes
testosterone (2 hours post final dose)

11.8

64.3 [3.76]

UFa = 3
UFh= 10
Total UF = 30

(Clewell et al.,
2013 a)

Pregnant SD rats; dietary; 0, 760, 3,800,
11,400 ppm (est. 56, 288, 720 mg/kg-day
on GDs 13-20; 109, 555, 1,513 mg/kg-
day on PNDs 2-14); GD 12-PND 14

NOAEL = 56

i male pup weight (PND 14), | MNGs
(PND 2) in the testes

13.2

72.1 [4.21]

UFa = 3
UFh= 10
Total UF = 30

(Clewell et al.,

2013b)

Pregnant SD rats; oral gavage; 0, 100,
500, and 1,000 mg/kg-day; GDs 6-15

NOAEL =
100

t skeletal variations (total skeletal
variations and rudimentary lumbar ribs)

23.6

129 [7.52]

UFa = 3
UFh= 10
Total UF = 30

(Waterman et al.
bw)

Pregnant Wistar rats; oral gavage; 0, 40,
200, and 1,000 mg/kg-day; GDs 6-15

NOAEL =
200

t incidences of rudimentary cervical
and accessory lumbar ribs; urogenital
and skeletal variations

47.3

257 [15.0]

UFa = 3
UFh= 10
Total UF = 30

(Hellwie et al..

Pregnant Wistar rats; oral gavage (corn
oil); 0, 300, 600, 750, 900 mg/kg-day;
GD 7-17

NOAEL =
300

t MNGs (GD 21); jsperm motility
(PND 90)

70.9

386 [22.6]

UFa = 3
UFh= 10
Total UF = 30

(Sobers et al.,

201 iy

Pregnant Harlan SD rats; Oral gavage
(corn oil); 0, 500, 750, 1,000, 1,500
mg/kg-day; GDs 14-18

LOAEL =
500

I fetal testicular testosterone production

118

643 [37.6]

UFa = 3
UFh=10
UFl = 10
Total UF = 300

(Hannas et al.,

201 iy

Pregnant SD rats; oral gavage (corn oil);
0, 750 mg/kg-day; GDs 14-18

LOAEL =
750

I Fetal testis testosterone production

177

965 [56.4]

UFa = 3
UFh= 10
UFl = 10
Total UF = 300

(Furr et al..

2014)

Page 74 of 184

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Study Details
(Species, Duration, Exposure Route/
Method, Doses |mg/kg-day|)

Study POD/

Type
(mg/kg-day)

Effect

HED

(mg/kg-
day)

HEC
(mg/m3)
|ppm|

Uncertainty
Factors''

Reference

Pregnant Wistar rats; oral gavage (peanut
oil); 0, 750 mg/kg-day; GDs 7-21

LOAEL =
750

I fetal testicular testosterone content
and production

177

965 [56.4]

UFa = 3
UFh=10
UFl = 10
Total UF = 300

(Borcli et al.,

2004)

Pregnant SD rats; oral gavage (corn oil);
0, 750 mg/kg-day; GD 14-PND 3

LOAEL =
750

I maternal body weight gain; | male
pup nipples/areola retention; testes
malformations (small, atrophic, flaccid,
fluid-filled, azoospermia, epididymal
agenesis)

177

965 [56.4]

UFa = 3
UFh = 10
UFl=10
Total UF = 300

(Grav et al..

2000)

Pregnant SD rats; oral gavage (corn oil);
0, 250, 750 mg/kg-day; embryonic day
13.5-17.5

NOEL = 250

t testicular mRNA levels of P450scc,
GATA-4, and Ins/3







(Adanisson et al,
2009)fl

" Studv considered as part of NASEM meta-analysis (NASEM, 2017). EPA did not consider this studv (Li et aL, 2015) further as the sole studv on which to
derive the POD because several areas of uncertainty (e.g., low statistical power with n=6, questionable dose-response and uncertain adversity among several
endpoints) reduced EPA's confidence in the results when considered independently from the other studies in a meta-analysis.

h Lee et al. (2006a) was not suitable for use to determine an acute POD due to uncertainties (e.g., reporting deficiencies for dam bodv weight and food
consumption for a dietary exposure study, and others described in the text).

' R code supporting NASEM's meta-regression and BMD analysis of DINP is publicly available through GitHub (httDs://github.com/wachiuDhd/NASEM-2017-
Endocrine-Low-Dose).

'' EPA used allometric bodv weight scaling to the three-auarters power to derive the HED. Consistent with EPA Guidance (U.S. EPA. 201 lb), the interspecies
uncertainty factor (UFa), was reduced from 10 to 3 to account remaining uncertainty associated with interspecies differences in toxicodynamics. EPA used a
default intraspecies (UFh) of 10 to account for variation in sensitivity within human populations due to limited information regarding the degree to which
human variability may impact the disposition of or response to DINP. EPA used a LOAEL-to-NOAEL uncertainty factor (UFl) of 10 to account for the
uncertainty inherent in extrapolating from the LOAEL to the NOAEL.

2492

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4.1,2 Non-cancer Oral Points of Departure for Intermediate Exposures

EPA considered 12 short-term (>1 to 30 days) oral exposure studies (6 of rats and 6 of mice) of DINP
for establishing the intermediate duration POD (Table 4-3). Figure 4-2 depicts the dose-response array
for available studies. Ultimately, EPA selected the acute POD (12 mg/kg-day) and benchmark MOE
(total UF of 30) identified in Section 4.1.1 to evaluate risk from intermediate exposures {i.e., ranging
from 1 to 30 days) to DINP.

The acute POD is more sensitive than many of the intermediate HEDs based on liver, kidney, or
developmental toxicity in rodents. As can be seen from Table 4-3 and Figure 4-2, of the 12 short-term
studies under consideration, 7 supported HEDs ranging from 15.6 to 401 (Kwack et al. 2009; Kaufmann

et ai. 2002; Smith et ai. 2000; Hazleton Lab s s ., s 36; Bio/dynamics. 1982a; Midwest

Research Institu j_). These studies are less sensitive than the acute POD (HED of 12 mg/kg-day).
Further, several of these studies are limited by poor dose selection and did not test doses low enough to
support NOAEL identification (Hazleton Lab s \ <, >.' ^ \ s 86; Midwest Research Institute PM)
or only tested a single high dose of DINP (Kwack et al.. 2009; Bio/dynamics. 1982a).

Five short-term studies (Ma et al.. .^ s , < »g. 2015; Ma et al.. 2014; Masutomi et al.. 2003; Smith et
al.. 2000) report HEDs based on NOAELs ranging from 2.0 to 10 mg/kg-day, indicating that they are
more sensitive than the HED that EPA selected for a POD. However, each of these studies had
uncertainties that reduced EPA confidence in their use quantitatively for a POD for intermediate
duration exposure.

Masutomi et al. (2003) supports a developmental NOAEL of 3 1 mg/kg-day (HED of 7.3 mg/kg-day)
based on reduced F1 male offspring body weight on PND 27. However, this study is limited by its small
sample size (n of 5 rats per dose group). Further, the biological significance of the effect on F1 male
body weight is unclear, as F1 male bodyweight was unaffected on PND 2, and no effect on F1 male
bodyweight gain was observed from PND 2 to PND 10 or PND 10 to PND 21, and by PND 77 F1 male
body weight had recovered to control levels. These limitations and uncertainties reduce EPA's
confidence in using the study by Masutomi et al. (2003) for the intermediate POD.

Three studies (Ma et al. JO I \ i^'ng. 1:01 \ Ma et al.. 2014) reported treatment-related effects on
endpoints indicating oxidative stress, but it is unclear if the apparent effects on neurotoxicity (Ma et al..
2015; Pens. 2015) reported in Section 3.4, and the findings in the liver and kidney (Ma et al.. 2014)
reported in Section 3.2 and 3.3 can be directly attributed to the oxidative stress and inflammatory
responses observed in the studies. Although there is some evidence showing protective effects of
antioxidants in mitigating the effects of treatment with DINP, there is not enough data to determine the
link to the apparent effects on neurotoxicity and on the liver and kidneys. This limitation is due, in large
part, to the lack of quantitative data on the incidence or severity of the histopathology findings in the
brain, liver, and kidney. These data were only described qualitatively, with representative micrographs
of control and high dose groups presented as images, which precludes their usefulness to set a POD.
Additional limitations in the two neurotoxicity studies are described below.

In the two neurotoxicity studies (Ma et al.. 2015; Peng. 2015). male Kunming mice were administered
DINP by oral gavage at doses up to 200 mg/kg-day, followed by swim trials in the Morris Water Maze
to determine effects on learning and memory, along with measurements of oxidative stress and
histopathology evaluation of the brain. However, EPA identified several deficiencies in the study
methods and reporting. First, both studies only report mean escape latency of each swimming trial over
the 7-day acquisition phase but provide no measure of variability. Neither study conducted statistical

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May 2024

analysis on escape latency times within a given trial, but instead conducted statistical analysis on the
average escape latency times across the 7 trials. Therefore, EPA is not able to determine whether there
is a significant interaction between treatment and time to determine if the learning curve was steeper in
the controls compared to the mice administered DINP. Second, path length provides another measure of
learning, with path length decreasing over the acquisition phase if learning is occurring. The North
American Free Trade Agreement (NAFTA) Technical Working Group on Pesticides (TWG) -
Developmental Neurotoxicity Study Guidance Document (\ v < < \ < I ) indicates that the mean path
length per trial should be reported, as this outcome is highly correlated with escape latency times. Both
studies (Ma et at.. 2015; Pens. 2015) report use of camera tracking and computer software (ANY-
Maze), which has the capabilities to determine path length. However, neither study reports the path
length numerically for the swimming trials, but instead only depict an image of the swim path for a
representative trial in the high dose and control groups. The lack of quantitative data on swim path
length precludes EPA's ability to discern whether any increase in swim time is due to actual deficits in
learning and memory, or if there is an increase in swim time due to general toxicity (i.e., swimming
more slowly). Neither study included performance controls.

Per the NAFTA guidance document, swim speed and cued-trials are two common performance controls
that can be used to rule out treatment-related visual and motor impairments that can confound
interpretation of cognitive deficits (e.g., longer latency times may be due to slower swim speeds, not
cognitive impairment). Third, for the probe trial, Ma et al. (2015) report both the target quarter retention
time and the number of entries into the target quadrant, which is consistent with the NAFTA guidance
document. There is a clear treatment related effect on target quadrant retention time; however, the
controls spent only -16 seconds in the target quadrant, which is only slightly above chance levels of 25
percent. NAFTA guidance states that controls must show an increase in percent time in the correct
quadrant that exceeds chance levels of 25 percent. For the probe trial by Peng (2015). target quadrant
retention time is reported, and controls spent approximately 25 seconds in the target quadrant, well
above chance levels of 25 percent, but the number of entries into the target quadrant was not reported.
Fourth, both of these studies (Ma et al.. 2015; Pens. ) reported alterations in pyramidal cells in
hippocampus at the high dose (150 and 200 mg/kg-day); however, no quantitative data were provided on
the incidence or severity of the histopathology findings; the data were only described qualitatively, with
representative micrographs of control and high dose groups presented as images. Taken together, these
uncertainties limit the utility of the neurological studies for use in determining an intermediate duration
POD.

Finally, one study (Smith et al.. 2000) provided a HED value for the NOAEL (10 mg/kg-day) in the
same range as the acute HED value (12 mg/kg-day). Smith et al. (2000) report treatment-related
increases in liver weights, hepatic peroxisomal beta oxidation (PBOX), and DNA synthesis,
accompanied by inhibition of gap junctional intercellular communication (GJIC), in male B6C3F1 mice
fed diets containing 6,000 ppm DINP (approximately 900 mg/kg-day) for up to four weeks. This study is
limited in dose selection, with only two treated groups with doses spanning a wide range between the
NOAEL in the low-dose group at 75 mg/kg-day and the LOAEL in the high dose at 900 mg/kg-day.
Therefore, EPA did not consider the dose selection to be refined enough or endpoints examined to be
comprehensive enough to establish a robust POD. However, the fact that the HED value from this study
aligns with the HED from the acute POD adds further support to EPA's selection of the acute POD to be
protective of intermediate exposure durations.

Page 77 of 184

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IHED LOAEL

AHEDNOAEL

• Other Doses Tested

4/ male body weight on PND 27; GD 7-21; rats; Masutomi, 2003,192872

Histopathological alterations in pyramidal cells in hippocampus; 14 days; mice (M); Ma,
2015, 3071020

Histopathological alterations in pyramidal cells in hippocampus; Impaired learning &
memory in MWM; 9 days; mice (M); Peng, 2015, 2919074

"T abs. and rel. liver weight; macroscopic liver observations; sb triglycerides; 7 days; rats
(M); Bio/dynamics, 1982,1987586

'T abs. and rel. liver weight; -f 11- and 12-hydroxylase activity, hypolipidemic effects; 21
days; rats (M/F); BIBRA, 1986,1325463

1" relative liver weight; clinical chemistry AST, ALP & triglycerides); 28 days; rats (M);
Kwack, 2009, 697382

T" abs. 8t rel. liver weight; 1" peroxisomal vol.& enzyme activity; T^hepatocyte proliferation;
4 weeks; mice (M); Kaufmann, 2002, 680011

1" incidence of hepatocytomegaly; enlarged & discolored livers; 4 weeks; mice (M);
Hazleton, 1991,1325464

1s rel. liver wt T" PBOX; Inhibition of GJIC; 2 weeks; rats (M); Smith, 2000, 667301

T1 rel. liver wt T* PBOX; Inhibition of GJIC; 4 weeks; rats (M); Smith, 2000, 667301
1" rel. liver wt "T PBOX; Inhibition of GJIC; 2 weeks; mice (M); Smith, 2000, 667301

T rel. liver wt PBOX; Inhibition of GJIC; 4 weeks; mice (M); Smith, 2000, 667301

Kidney

T* abs. and rel. kidney weight; 7 days; rats (M); Bio/dynamics, 1982,1987586

• •

0.01	0.1	1	10

HED (mg/kg-day)

100

1000

2587

2588	Figure 4-2. Dose-Response Array of Studies Considered for Deriving the Intermediate Duration

2589	Non-cancer POD

2590	Notes: j = statistically significant increase in response compared to controls; j = statistically significant decrease

2591	in response compared to controls

2592	Dev = developmental; Neuro = neurological; M = males; F = females; GD = gestational day; PND = postnatal

2593	day; ALT = alanine aminotransferase; AST= aspartate aminotransferase; ALP = alkaline phosphatase; HED =

2594	human equivalent dose; NOAEL = no-observed-adverse-effect-level; LOAEL = lowest-observed-adverse-effect-

2595	level.

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Table 4-3. Dose-Response Analysis of Select

ed Studies Considered for Deriving the Intermediate

Non-ancer POD

Target
Organ/
System

Study Details
(Species, duration, exposure route/
method, doses |mg/kg-day|)

Study POD/
Type (mg/kg-
day)

Effect

HED

(mg/kg)

HEC
(mg/m3)
[ppm ]

Uncertainty
Factors"

Reference

Neurotoxicity

Kunming mice (males only); oral
gavage; 0, 1.5, 15, 150 mg/kg-day; 9
days

NOAEL = 15

i body weight gain; impaired learning
& memory in Morris Water Maze;
oxidative stress & inflammation;
histopathological alterations in
pyramidal cells in hippocampus

1.99

10.9
[0.634]

UFa = 3
UFh=10
Total UF = 30

(Pens. 2015)

Neurotoxicity

Kunming mice (males only); oral
gavage; 0, 0.2, 2, 20, 200 mg/kg-day;
14 days

NOAEL = 20

Histopathological alterations in
pyramidal cells; oxidative stress &
inflammation

2.66

14.5
[0.845]

UFa = 3
UFh=10
Total UF = 30

(Ma et a.L
2015)

Liver and
Kidney

Kunming mice (males only); oral
gavage; 0, 0.2, 2, 20, 200 mg/kg-day;
14 days

NOAEL = 20

Markers of oxidative stress (t ROS,
| GSH, t MDA, t 8-OH-dG) &
inflammation (f IL-1, f TNFa)

2.67

14.5
[0.845]

UFa = 3
UFh = 10

Total UF = 30

(Ma et al.
2014)

Developmental

Pregnant SD rats; dietary; 0, 400,
4000, 20,000 ppm (est. 31-66, 307-
657, 1,165-2,657 mg/kg-day); GD 15
to PND 10

NOAEL = 31
(males);
66 (females)

i male body weight on PND 27

7.33

39.9 [2.33]

UFa = 3
UFh = 10

Total UF = 30

(Masutomi et

al.. 2003)

Liver

B6C3F1 mice (males only); dietary;
0, 500, 6000 ppm (est. 0, 75, 900
mg/kg-day); 2 and 4 weeks

NOEL = 75

Hepatic changes (t liver weight,
t PBOX, t DNA synthesis; Inhibition
of GJIC)

9.97

54.3 [3.17]

UFa = 3
UFh = 10

Total UF = 30

(Smith et al.
2000)

Liver

B6C3F1 mice (both sexes); dietary;
0, 500, 1500, 4000, 8000 ppm (est.
117, 350, 913, 1,860 mg/kg-day
[males]; 0, 167, 546, 1,272, 2,806
mg/kg-day [females]); 1 or 4 weeks

NOAEL=
117 (males)

t absolute and relative liver weight; t
peroxisomal volume, and peroxisomal
enzyme activity; t hepatocyte
proliferation in males

15.6

84.7 [4.95]

UFa = 3
UFh= 10
Total UF = 30

(Kaufmann et

al. 2002)

Liver

F344 rats (males only); dietary; 0,
1000, 12,000 ppm (est. 0, 100, 1200
mg/kg-day); 2 and 4 weeks

NOAEL = 100

Hepatic changes (t liver weight,
t PBOX, t DNA synthesis; Inhibition
of GJIC)

23.6

129 [7.52]

UFa = 3
UFh = 10

Total UF = 30

(Smith et al.
2000)

Liver &
Kidney

F344 rats (both sexes); dietary; 0, 0.2,
0.67, 2% (est. 150, 500, 1,500 mg/kg-
day [males]; 0, 125, 420, 1,300
mg/kg-day [females]); 28 days

LOEL=
125 (females)

t hepatic catalase and carnitine
acetyltransferase activity

29.6

161 [9.39]

UFa = 3
UFh= 10
UFl = 10

Total UF =
300

(Midwest
Research
Institute. 1981)

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Target
Orjjan/
System

Study Details
(Species, duration, exposure route/
method, doses |mjj/kjj-day|)

Study POD/
Type (m«/k«-
day)

Effect

HED

(ms/k$i)

HEC
(mj*/mJ)
[ppm]

Uncertainty
Factors"

Reference

Liver

B6C3F1 mice (both sexes); dietary;
0, 3000, 6000, 12,500 ppm (est. 635,
1,377, 2,689, 6,518 mg/kg-day
[males]; 780, 1761, 3,287, 6,920
mg/kg-day [females]); 4 weeks

LOAEL=
635 (males)

Enlarged and discolored livers;
t incidence of hepatocytomegaly

84.4

460 [26.8]

UFa = 3
UFh= 10
UFl = 10

Total UF =
300

(Hazleton Labs,

1991a)

Liver

SD rats (males only); oral gavage; 0,
500 mg/kg-day; 28 days

LOAEL = 500

i body weight gain; f relative liver
weight; clinical chemistry (f AST,
ALP & triglycerides)

118

643 [37.6]

UFa = 3
UFh= 10
UFl = 10

Total UF =
300

(Kwack et al..

2009)

Liver

F344 rats (both sexes); diet; 0, 0.6,
1.2, 2.5% (est. 639, 1192, 2,195
mg/kg-day [males]; 607, 1,198, 2,289
mg/kg-day [females]); 21 days

LOAEL=
607 (females)

t absolute and relative liver weight;
|11- and 12-hydroxylase activity,
hypolipidemic effects

144

781 [45.6]

UFa = 3
UFh= 10
UFl = 10

Total UF =
300

(BIBRA. 1986)

Liver and
Kidney

F344 rats (males only); dietary; 0, 2%
(est. 1,700 mg/kg-day); 7 days

LOAEL=
1,700

t absolute and relative liver and kidney
weight, macroscopic liver
observations, changes in clinical
chemistry

402

2,187
[128]

UFa = 3
UFh= 10
UFl = 10

Total UF =
300

(Bio/dvnamics.

1982a)

"EPA used allometric bodv weieht scaline to the three-auarters power to derive the HED. Consistent with EPA Guidance (U.S. EPA. 201 lb), the interspecies
uncertainty factor (UFA), was reduced from 10 to 3 to account remaining uncertainty associated with interspecies differences in toxicodynamics. EPA used a default
intraspecies (UFH) of 10 to account for variation in sensitivity within human populations due to limited information regarding the degree to which human variability may
impact the disposition of or response to DINP. EPA used a LOAEL-to-NOAEL uncertainty factor (UFL) of 10 to account for the uncertainty inherent in extrapolating
from the LOAEL to the NOAEL.

2597

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4,1.3 Non-cancer Oral Points of Departure for Chronic Exposures

EPA considered four 2-year chronic dietary studies (3 of rats, 1 of mice), six 13-week subchronic dietary
studies (4 of rats, and 1 each of mice and beagles), a one-generation study of reproduction of rats, and a
two-generation study of reproduction of rats for establishing the chronic POD (Table 4-5). Across one-
and two-generation studies of reproduction, reduced offspring bodyweight was the most sensitive effect,
while liver and kidney toxicity were the most sensitive effects observed across chronic and subchronic
studies, and these effects were considered for establishing the chronic POD. Figure 4-3 depicts the dose-
response array for available studies.

Across the one- and two-generation studies of reproduction (Waterman et ai. 2000; Exxon Biomedical.
1996a. b), both of which were GLP-compliant and adhered to available guidelines (40 CFR Part 798, §
798.4700), LOAELs for developmental effects were 377 mg/kg-day in the one-generation study based
on reduced male and female F1 offspring body weight on PNDs 0, 14, and 21; and 133 mg/kg-day in the
two generation study based on reduced F1 and F2 offspring body weight on PNDs 7 and 21. Neither
study tested sufficiently low doses to establish a developmental NOAEL. Further, there is some
uncertainty associated with the LOAEL from the two-generation study, as F1 offspring bodyweight
(both sexes) was reduced on PND21, while F2 offspring body weight was reduced only on PND 7 for
females (Table 3-8). More consistent effects on F1 and F2 offspring body weight were observed in the
mid-dose group. These sources of uncertainty reduce EPA's confidence in using the LOAEL of 133
mg/kg-day from the two-generation study as a chronic POD. Further, EPA identified more sensitive
PODs based on liver toxicity from subchronic and chronic studies that tested lower doses of DINP and
allowed for the identification of a NOAEL.

Across the six available subchronic studies, the lowest LOAELs for each of the tested species were 160
mg/kg-day in beagles (NOAEL = 37 mg/kg-day; HED = 23) based on increased absolute and relative
liver weight and increase serum ALT (Hazleton Laboratory [); 972 mg/kg-day in mice (NOAEL =
365; HED = 49 mg/kg-day) based on increased absolute and relative liver weight and histopathological
findings (e.g., necrosis) (Hazleton Labs JiHj2); and 60 mg/kg-day in SD rats (no NOAEL identified;
HED = 14 mg/kg-day) based on increased incidence of histopathological lesions in the kidney of male
rats (i.e., focal mononuclear cell infiltration and mineralization) (Hazleton Labs. 1981). LOAELs based
on liver and kidney toxicity from the remaining three subchronic studies of rats were less sensitive and
ranged from 176 to 227 mg/kg-day (Hazleton Lab s U>, < namics. 1982b. c). The study of
beagles was conducted prior to the establishment of GLP principles and OECD test guidelines, and
additionally only included four dogs per sex in each treatment group, so no statistical analysis was
performed due to the small sample size (Hazleton Laboratories. 1971). These limitations reduced EPA's
confidence in using the study to establish a chronic POD, and importantly, other subchronic and chronic
studies of rats provide more sensitive and health protective candidate PODs. Similarly, the one
subchronic study of mice (Hazleton Labs. 1992) provides a less sensitive candidate POD compared to
studies of rats. The lowest subchronic LOAEL of 60 mg/kg-day in rats comes from a study conducted
prior to the establishment of GLP principles and OECD test guidelines (Hazleton Labs. 1981). and did
not test sufficiently low doses to establish a NOAEL. Furthermore, EPA did not consider this study
sufficient for selection of a POD because it only reported effects on kidney in male rats which may be
related to a2u-globulin and not relevant for human health.

Across the four available 2-year dietary studies of rats and mice, the lowest LOAEL is 152 mg/kg-day
(NOAEL = 15 mg/kg-day; HED = 3.5 mg/kg-day) from a 2-year dietary study of F344 rats (Lineton et
ai. 1997; Bio/dynamics. 1986). The study by Lington et al. is GLP-compliant and received a high
overall study quality determination. Although the study does not explicitly state compliance with any

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testing guidelines, it generally follows the guidelines outlined by OECD Test Number 453 (Combined
Chronic Toxicity/Carcinogenicity Studies). At the LOAEL, a spectrum of dose-related effects consistent
with liver toxicity was observed in male and female rats, including treatment related increases in relative
liver weight, serum ALT, AST, and ALP, and histopathological findings {i.e., spongiosis hepatis,
sinusoid ectasia, hepatopathy associated with leukemia). One source of uncertainty associated with the
findings of Lington et al. results from spongiosis hepatis. The MOA underlying spongiosis hepatis is
unknown but is not believed to be related to peroxisome proliferation. Further, as discussed by ECHA
(2 ), spongiosis hepatis has been observed in the livers of some strains of rats and certain species of
fish {e.g., medaka), but is less common in mice, has not been observed in non-human primates or dogs,
and with the exception of two case reports, has not been described in humans. These findings raise some
uncertainty as to the human relevance of spongiosis hepatis (Karbe and Kerlin. 2002). However,
spongiosis hepatis co-occurred with other hepatic effects that are more clearly adverse and relevant for
use in human health risk assessment {e.g., increase liver weight, serum ALT, AST, ALP, focal necrosis).
Further supporting use of the LOAEL reported by Lington et al., similar hepatic effects {e.g., increased
relative liver weight, serum ALT, AST, and ALP, spongiosis hepatis, necrosis) have consistently been
reported in two other chronic dietary studies of DINP with F344 (Covance Labs. 1998c) and SD rats
(Bio/dynamics. 1987). albeit at slightly higher doses of DINP (Table 4-5).

Given the broad dose spacing between the NOAEL of 15 mg/kg-day and LOAEL of 152 mg/kg-day
identified in Lington et al. (1997). EPA attempted to refine the POD by conducting BMD modeling in
accordance with EPA's Benchmark Dose Technical Guidance ( ). Endpoints modeled

included relative liver weight at terminal sacrifice (both sexes); serum ALT at 6-and 18-month sacrifices
(males only); incidence of focal necrosis in the liver (both sexes); incidence of spongiosis hepatis (males
only); and incidence of sinusoid ectasia (males only). For each endpoint, multiple BMRs were modeled.
BMD modeling results are presented in Appendix E, and results for representative BMRs are presented
in Table 4-4. For dichotomous endpoints, BMDLio values ranged from 8.6 mg/kg-day for spongiosis
hepatis to 125 mg/kg-day for focal necrosis in male rats. BMDLio values for spongiosis (8.6 mg/kg-day)
in the liver and sinusoid ectasia in the liver (14 mg/kg-day) were less than the study NOAEL of 15
mg/kg-day, however, BMD/BMDL ratios were greater than 3 (ranging from 3.7 to 8.9), indicating
model uncertainty. For continuous endpoints, the BMDLio was 85 mg/kg-day for increased relative liver
weights for males, while no models adequately fit relative liver weight data for female rats. For increase
in serum ALT at 6 and 18 months, BMDLioo values were 87 and 134 mg/kg-day, respectively. A BMR
of 100 percent was selected for this endpoint since 2 to 3 fold changes in ALT are generally considered
biologically significant and outside the range of normal variation (Hall et al.. JO I J; 1, c. < i1 \ 2002a).
However, there is some uncertainty related to the BMR selection, so EPA also presents BMDLisd values
in Table 4-4, which is consistent with EPA's Benchmark Dose Technical Guidance ( ).

BMDLisd values for increased serum ALT at 6 and 18 months were 16 and 33 mg/kg-day, respectively.

Overall, calculated BMDLs shown in Table 4-4 ranged from 8.6 to 134 mg/kg-day, which is similar to
the study NOAEL and LOAEL values of 15 and 152 mg/kg-day. The wide variability in BMDLs and
uncertainty in several modelled outcomes {i.e., BMD/BMDL ratios >3) reduce EPA's confidence in
using the BMD modeling results for establishing a POD.

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Table 4-4. Summary of B

MD Model Results from Lingt

oil et al. (1997)

Endpoint

Sex

Selected Model

BMDisd /
BMDLisd
(mg/kg-day)

BMDio/
BMDLio
(mg/kg-day)

BMDioo/
BMDLioo
(mg/kg-day)

Dicholomous endpoints

Focal necrosis in the liver

Male

Logistic

-

159/125

-

Focal necrosis in the liver

Female

Log-Probit

-

222/34

-

Spongiosis hepatis in the
liver

Male

Log-Probit

-

32/8.6

-

Sinusoid ectasia in the liver

Male

Log-Probit

-

125/14

-

( onlinuous aid points

Relative Liver weight at
terminal sacrifice

Male

Linear, CV

242/196

106/85

-

Relative Liver weight at
terminal sacrifice

Female

None selected;
LOAEL (184 mg/kg-
day) was used

"

"

"

Serum ALT at 6-month
sacrifice

Male

Linear

23/16

-

125/87

Serum ALT at 18-month
sacrifice

Male

Power

63/33

-

179/134

Overall, EPA selected the NOAEL of 15 mg/kg-day (HED = 3.5 mg/kg-day) based on liver toxicity
observed in a 2-year dietary study of F344 rats (Lington et ai. 199 , «'t<> dynamics. 1986) as the chronic
POD for use in estimating non-cancer risk from exposure to DINP in the draft DINP risk evaluation.

This POD represents the most sensitive POD identified by EPA. Furthermore, the NOAEL of 15 mg/kg-
day supports the suite of effects occuring at 152 mg/kg-day in Lington et al. (1997). Consistently, other
regulatory bodies have selected the same chronic POD for use in quantifying risk from exposures to
DINP (ECCC/HC. 2020; 1 U'SC^M i) (U v \ , LCI \ . M !.). A total UF of 30 was
selected for use as the benchmark MOE (based on an interspecies UF (UFa) of 3 and an intraspecies UF
(UFh) of 10). Consistent with EPA guidance (2022. 2002b. 1993). EPA reduced the UFa from a value of
10 to 3 because allometric body weight scaling to the three-quarter power was used to adjust the POD to
obtain a HED (Appendix F).

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0)
>

>


 <

<

10	100	1000

HED (mg/kg-day)

10000

Figure 4-3. Dose-Response Array of Studies Considered for Considered for Deriving the Chronic
Non-cancer POD

Notes: | = statistically significant increase in response compared to controls; | = statistically significant decrease
in response compared to controls; M = males; F= females; PI = parental generation; PND = postnatal day; ALT =
alanine aminotransferase; AST aspartate aminotransferase; ALP = alkaline phosphatase; HED = human
equivalent dose; NOAEL = no-observed-adverse-effect-level; LOAEL = lowest-observed-adverse-effect-level.

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Table 4-5. Dose-Response Analysis of Seleci

ted Studies Considered for Deriving the C

ironic >

on-cancer POD

Target Organ/
System

Study Details
(Species, Duration, Exposure
Route/ Method, Doses |mg/kg-
day])

Study POD/

Type
(mg/kg-day)

Effect

HEP

(mg/kg)

HEC
(mg/m3)
[ppm]

Uneertainty
Factors "b

Rcfcrcncc(s)

Liver and
Kidney

F344 rals (bolli sexes); dietary; 0,
0.03, 0.3, 0.6% (est. 15, 152, 307
mg/kg-day [males]; 18, 184, 375
mg/kg-day [females]); 2 years

NOAF.I ,=
15 (males)
18 (females)

t absolute and relative liver and
kidney weight; t in serum ALT and
AST; histopathological alterations
(e.g., spongiosis hepatis, focal
necrosis)

3.55

19.3 [1.13]

UFa = 3
UFh = 10
Total UF =
30

(Lington et al. 1997;

Bio/dvnamics, 1986)

Liver

SD rats (both sexes); dietary; 0, 500,
5000, 10,000 ppm (est. 27, 271, 553
mg/kg-day [males]; 33, 331, 672
mg/kg-day [females]); 2 years

NOAF.I. = 27

t serum ALT, AST, ALP (males);
histopathological findings in the
liver (i.e., minimal-to-slight focal
necrosis, spongiosis hepatis)

6.38

34.7 [2.03]

UFa = 3
UFl = 10
Total UF =
30

(Bio/dvnamics, 1987)

Liver and
Kidney

F344 rats (both sexes); dietary; 0,
500, 1500, 6000, 12,000 ppm (est.
29, 88, 359, 733 mg/kg-day [males];
36, 109, 442, 885 mg/kg-day
[females]); 2 years

NOAEL =
88 (males)
109 (females)

t absolute and relative liver and
kidney weight; t in serum ALT,
AST, BUN; histopathological
findings in liver (e.g., spongiosis
hepatis) and kidney (e.g.,
mineralization of renal papilla,
pigment in tubule cells)

20.8

113 [6.61]

UFa = 3
UFh = 10
Total UF =
30

(Covance Labs.
1998c)

Liver and
Kidney

B6C3F1 mice (both sexes); dietary;
0, 500, 1500, 4000, 8000 ppm (est.
90, 276, 742, 1,560 mg/kg-day
[males]; 112,336,910, 1,888
mg/kg-day [females]); 2 years

NOAEL=
90 (males)
112 (females)

t absolute and relative liver weight,
histopathological changes in the
liver (EXAMPLES); j body weight
gain (females); t incidence of liver
masses and j absolute kidney
weight (males)

12.0

65.1 [3.80]

UFa = 3
UFh = 10
Total UF =
30

(Covance Labs.
1998b)

Developmental

Two-eeneration studv: SD rats

LOAEL = 133

| F1 and F2 offspring body weight
onPNDs7 and 21

31.4

171 [10.0]

UFa = 3
UFh = 10
UFl = 10
Total UF =
300

(Waterman et al.

(30/group) administered 0, 0.2, 0.4,
0.8% DINP in the diet continuously
starting 10 weeks prior to mating,
throughout mating, gestation and
lactation for two generations

2000; Exxon
Biomedical, 1996b)

Developmental

One generation studv: SD rats

LOAEL = 377

i male and female offspring body
weight on PND 0, 14, and 21

89.1

485 [28.3]

UFa = 3
UFh = 10
UFL=10

(Waterman et al.

(30/group); administered 0, 0.5, 1.0,
1.5% DINP in diet continuously
starting 10 weeks prior to mating

2000; Exxon
Biomedical 1996a)

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Target Orjjan/
System

Study Details
(Species, Duration, Exposure
Route/ Method, Doses |mjj/kjj-
day|)

Study POD/

Type
(mjj/kjj-day)

Effect

HED

(mjj/kjO

HEC
(mjj/m3)
[ppm]

Uncertainty

Factors "b

Refcrcncc(s)



and throughout mating, gestation
and lactation for one generation.









Total UF =
300



Liver

Beagle dogs (both sexes); dietary; 0,
0.125, 0.5, 2% (est. 37, 160, 2,000
mg/kg-day); 13 weeks

NOAEL = 37

t absolute and relative liver weight;
t serum ALT

23.0

125 [7.32]

UFa = 3
UFh = 10

Total UF =
30b

(Hazleton

Laboratories. 1971)

Liver

B6C3F1 mice (both sexes); dietary;
0, 1500, 4000, 10,000, 20,000 ppm
(est: 365, 972, 2,600, 5,770 mg/kg-
day); 13 weeks (Hazleton 1992)

NOAEL = 365

t absolute and relative liver weight;
liver histopathology (e.g., necrosis,
degeneration, hepatocyte
enlargement)

48.5

264 [15.4]

UFa = 3
UFh = 10
UFS= 10
Total UF =
300

(Hazleton Labs, 1992)

Liver and
Kidney

F344 rats (both sexes); dietary; 0,
0.1, 0.3, 0.6, 1.0, 2.0% (est. 0, 77,
227, 460, 767, 1,554 mg/kg-day); 13
weeks

NOAEL = 77

t absolute and relative liver and
kidney weight; j cholesterol level
(females)

18.2

99.1 [5.79]

UFa = 3
UFh = 10
UFS= 10
Total UF =
300

(Bio/dvnamics.

1982b)

Liver &
Kidney

F344 rats (both sexes); dietary; 0,
2500, 5000, 10,000, 20,000 ppm
(est. 176, 354, 719, 1545 mg/kg-day
[males]; 218, 438, 823, 1,687
mg/kg-day [females]); 13 weeks

LOAEL =
176 (males)
218 (females)

t kidney and liver weights

41.6

226 [13.2]

UFa = 3
UFh= 10
UFS= 10
UFl = 10
Total UF =
3000

(Hazleton Labs.
1991b)

Liver &
Kidney

SD rats (both sexes); dietary; 0, 0.3,
1.0% (est. 201, 690 mg/kg-day
[males]; 251, 880 mg/kg-day
[females]); 13 weeks

LOAEL=
201 (males)
251 (females)

t absolute and relative liver &
kidney weight accompanied by j in
triglycerides and altered urine
chemistry

47.5

259 [15.1]

UFa = 3
UFh= 10
UFS= 10
UFl = 10
Total UF =
3000

(B io/dv nam ics, 1982c)

Kidney

SD rats (both sexes); dietary; 0,
1000, 3000, 10,000 ppm (estimated:
0, 60, 180, 600 mg/kg-day); 13

LOAEL = 60
(males)

t incidence of histopathology
lesions in the kidney [i.e., focal

14.2

77.2 [4.51]

UFa = 3
UFh= 10

(Hazleton Labs. 1981)

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Target Organ/
System

Study Details
(Species, Duration, Exposure
Route/ Method, Doses |mg/kg-
day|)

Study POD/

Type
(mg/kg-day)

Effect

HED

(nig/kg)

HEC
(mg/m3)
[ppm]

Uncertainty

Factors "b

Refcrcncc(s)



weeks



mononuclear cell infiltration and
mineralization]; males only





UFS= 10
UFl = 10

Total UF =
3000



"EPA used allomctric bodv weieht scaline to the three-auarters power to derive the HED. Consistent with EPA Guidance (U.S. EPA. 201 lb), the interspecies
uncertainty factor (UFA), was reduced from 10 to 3 to account remaining uncertainty associated with interspecies differences in toxicodynamics. EPA used a default
intraspecies (UFH) of 10 to account for variation in sensitivity within human populations due to limited information regarding the degree to which human variability
may impact the disposition of or response to DINP. EPA used a LOAEL-to-NOAEL uncertainty factor (UFL) of 10 to account for the uncertainty inherent in
extrapolating from the LOAEL to the NOAEL.

b EPA considered applying a subchronic-to-chronic (UFS) of 10 for the intermediate (13-week) dog study under consideration for deriving a chronic POD. However,
retrospective analyses of 13-week and 1-year dog studies have shown that dog studies beyond 13-weeks do not have a significant impact on the derivation of chronic
PODs (Bishop et al, 2023; Dellarco et al, 2010; Box and Spielmann, 2005). Therefore, this a UFs was not used.

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4.2 Weight of Scientific Evidence

4,2,1 POD for Acute and Intermediate Durations

EPA has preliminarily concluded that the HED of 12 mg/kg-day (BMDLs of 49 mg/kg-day) from the
NASEM (2017) meta-regression of reduced fetal testicular testosterone in rats is appropriate for
calculation of risks for acute and intermediate exposure durations. A total UF of 30 was selected for use
as the benchmark MOE (based on an interspecies UF (UFa) of 3 and an intraspecies UF (UFh) of 10).
Consistent with EPA guidance (2022. 2002b. 1993). EPA reduced the UFa from a value of 10 to 3
because allometric body weight scaling to the three-quarter power was used to adjust the POD to obtain
a HED (Appendix F). EPA has robust overall confidence in the selected POD based on the following
weight of scientific evidence:

• DINP exposure resulted in treatment-related effects on the developing male reproductive system
consistent with a disruption of androgen action during the critical window of development in 13
studies of rats (Section 3.1.2.1). Observed effects included: reduced mRNA expression of INSL3
and genes involved in steroidogenesis in the fetal testes; reduced fetal testes testosterone content
and/or production; reduced male pup anogenital distance; increased male offspring nipple
retention; increased incidence of MNGs and fetal Ley dig cell aggregation; and decreased sperm
motility in adult rats exposed perinatally to DINP.

• EPA has previously considered the weight of scientific evidence and concluded that oral
exposure to DINP can induce effects on the developing male reproductive system consistent with
a disruption of androgen action (see EP A's Draft Proposed Approach for Cumulative Risk
Assessment of High-Priority and a Manufacturer-Requested Phthalate under the Toxic
Substances Control Act ( 023 a)). Notably, EPA's conclusion was supported by the
SACC (U.S. EPA. 2023b).

• The selected POD is based on meta-regression analysis of fetal testosterone data from two
studies of rats (Li et ai. 2015: Hannas et ai. ).

• Two additional developmental toxicity studies (Clewell et ai. 2013a: Clewell et ai. )
resulted in decreased fetal testosterone production and other effects on the developing male
reproductive system at similar doses (LOAELs from 250 to 307 mg/kg-day and NOAELs from
50 to 56 mg/kg-day) to the BMDLs of 49 mg/kg-day derived from the NASEM meta-analysis.
These studies support the selection of the BMDLs of 49 mg/kg-day for the acute and
intermediate duration PODs.

There are no studies conducted via the dermal and inhalation route relevant for extrapolating human
health risk. Therefore, EPA is using the oral HED of 12 mg/kg-day to extrapolate to the dermal route.
Differences in absorption will accounted for in dermal exposure estimates in the draft risk evaluation for
DINP.

EPA is also using the oral HED of 12 mg/kg-day to extrapolate to the inhalation route. EPA assumes
similar absorption for the oral and inhalation routes, and no adjustment was made when extrapolating to
the inhalation route. For the inhalation route, EPA extrapolated the daily oral HEDs to inhalation HECs
using a human body weight and breathing rate relevant to a continuous exposure of an individual at rest.
Appendix F provides further information on extrapolation of inhalation HECs from oral HEDs.

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4.2,2 POD for Chronic Durations

EPA has preliminarily concluded that the HED of 3.5 mg/kg-day (NOAEL of 15 mg/kg-day) from the 2-
year dietary study of F344 rats based on liver toxicity (Lington et a I _ I' »• }7; <1t< < < l\ namit1 l - 5) is
appropriate for calculation of risk for chronic exposure durations. A total UF of 30 was selected for use
as the benchmark MOE (based on an interspecies UF (UFa) of 3 and an intraspecies UF (UFh) of 10).
Consistent with EPA guidance (2022. 2002b. 1993). EPA reduced the UFa from a value of 10 to 3
because allometric body weight scaling to the three-quarter power was used to adjust the POD to obtain
a HED (Appendix F). EPA has robust overall confidence in the selected POD based on the following
weight of scientific evidence:

•	The NOAEL of 15 mg/kg-day (HED = 3.5 mg/kg-day) from the 2-year dietary study of F344 rats
(Lington etai. 1997; Bio/dynamics. 1986) represents the most sensitive POD identified by EPA
across the 12 relevant studies subjected to dose-response analysis, including four 2-year chronic
dietary studies (3 of rats, 1 of mice), six 13-week subchronic dietary studies (4 of rats, and 1 each
of mice and beagles), a one-generation study of reproduction of rats, and a two-generation study
of reproduction of rats.

•	This study received a high overall study quality determination and is GLP-compliant.

•	At the LOAEL, a spectrum of dose-related effects consistent with liver toxicity was observed in
male and female rats, including treatment related increases in relative liver weight, serum ALT,
AST, and ALP, and histopathological findings (i.e., spongiosis hepatis, focal necrosis, sinusoid
ectasia, hepatopathy associated with leukemia).

•	Given the relatively broad dose-spacing between the NOAEL (15 mg/kg-day) and the LOAEL
(152 mg/kg-day) in the principal study, EPA attempted to refine the POD by conducting BMD
modeling of relevant dose-related findings showing a substantial increase in magnitude over
controls, including: relative liver weight at terminal sacrifice (both sexes); serum ALT at 6-and
18-month sacrifices (males only); incidence of focal necrosis in the liver (both sexes); incidence
of spongiosis hepatis (males only); and incidence of sinusoid ectasia (males only). Calculated
BMDLs ranged from 8.6 to 125 mg/kg-day, which is similar to the study NOAEL and LOAEL
values of 15 and 152 mg/kg-day. The wide variability in BMDLs and uncertainty in several
modelled outcomes (i.e., BMD/BMDL ratios greater than 3) reduce EPA's confidence in using
the BMD modeling results for establishing a POD, and further affirm the use of the NOAEL for
establishing the POD.

•	The NOAEL of 15 mg/kg/day in Lington et al. (1997) also aligns with the BMD05 of 12
mg/kg/day for one of the more sensitive endpoints in this study, spongiosis hepatis, determined
by CPSC (2010). However, EPA considers it more appropriate to use the NOAEL of 15 mg/kg-
day instead of the BMD05 of 12 mg/kg-day because the NOAEL supports the suite of effects on
the liver occuring at 152 mg/kg-day instead of being based on the single effect of spongiosis
hepatis with its associated uncertainty regarding human relevance.

•	The endpoints indicative of liver toxicity on which the POD is based were robust in that they
were observed across species and durations.

o The remaining three chronic studies in rodents (Covance Labs. 1998b. c; Bio/dynamics.
1987) reported similar findings of liver toxicity (e.g., increased liver weights; clinical
chemistry changes such as increased ALT, AST, ALP; and histopathology findings such
as liver necrosis and spongiosis hepatis), with similar but less sensitive NOAELs ranging
from 27 to 112 mg/kg-day.

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• Consistently, other regulatory bodies have selected the same chronic POD (NOAEL 15 mg/kg-
day) for use in quantify risk from exposures to DINP (ECCC/HC. 2020; 1 c. t PSC. 2014)

(EFSA. 2019; ECHA.: ).

There are no studies conducted via the dermal and inhalation route relevant for extrapolating human
health risk. Therefore, EPA is using the oral HED of 3.5 mg/kg-day to extrapolate to the dermal route.
Differences in absorption will accounted for in dermal exposure estimates in the draft risk evaluation for
DINP.

EPA is also using the oral HED of 3.5 mg/kg-day to extrapolate to the inhalation route. EPA assumes
similar absorption for the oral and inhalation routes, and no adjustment was made when extrapolating to
the inhalation route. For the inhalation route, EPA extrapolated the daily oral HEDs to inhalation HECs
using a human body weight and breathing rate relevant to a continuous exposure of an individual at rest.
Appendix F provides further information on extrapolation of inhalation HECs from oral HEDs.

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5 CONSIDERATION OF PESS AND AGGEGRATE EXPOSURE

5.1	Hazard Considerations for Aggregate Exposure

For use in the risk evaluation and assessing risks from other exposure routes, EPA conducted route-to-
route extrapolation of the toxicity values from the oral studies for use in the dermal and inhalation
exposure routes and scenarios. Health outcomes that serve as the basis for acute, intermediate and
chronic hazard values are systemic and assumed to be consistent across routes of exposure. EPA
therefore concludes that for consideration of aggregate exposures, it is reasonable to assume that
exposures and risks across oral, dermal, and inhalation routes may be additive for the selected PODs in
Section 6.

5.2	PESS Based on Greater Susceptibility

In this section, EPA addresses subpopulations expected to be more susceptible to DINP exposure than
other populations. Table 5-1 presents the data sources that were used in the potentially exposed or
susceptible subpopulations (PESS) analysis evaluating susceptible subpopulations and identifies whether
and how the subpopulation was addressed quantitatively in the draft risk evaluation of DINP. EPA
identified a range of factors that may have the potential to increase biological susceptibility to DINP,
including lifestage, chronic liver or kidney disease, pre-existing diseases, physical activity, diet, stress,
and co-exposures to other environmental stressors that contribute to related health outcomes.

Regarding lifestage, exposure to DINP during the masculinization programming window {i.e., GDs 15.5
to 18.5 for rats; GDs 14 to 16 for mice; gestational weeks 8 to 14 for humans) can lead to antiandrogenic
effects on the male reproductive system (MacLeod et at.. 2010; Welsh et at.. 2008; Carruthers and
Foster. 2005). Animal studies demonstrating effects of DINP on male reproductive development and
other developmental outcomes provide direct evidence that gestation is a particularly sensitive lifestage.
EPA considered the sensitivity of this lifestage in its derivation of the POD for acute and intermediate
exposure duration based on reduced fetal testicular testosterone in rats after evaluation of the weight of
scientific evidence that DINP resulted in treatment-related effects on the developing male reproductive
system consistent with a disruption of androgen action during the critical window of development in 13
studies of rats. In humans, there is moderate evidence for the association between DINP and testosterone
and semen parameters, based on studies that found decreasing testosterone levels with increasing DINP
exposure (Radke et at.. 2018). Based on this evidence from animal and human studies, EPA has
identified two groups that may be more susceptible to DINP exposure due to lifestages:

•	Pregnant women/women of reproductive age, and

•	Male infants, male toddlers, and male children.

Animal evidence also demonstrates that the liver, kidneys, nervous system, cardiovascular system,
immune system, may be sensitive target organs. EPA is quantifying risks based on liver and
developmental toxicity in the draft DINP risk evaluation, and determining risk based these endpoints is
protective of the other hazards that occur at higher doses.

Regarding the factor of co-exposure, studies have demonstrated that co-exposure to DINP and other
toxicologically similar phthalates {e.g., DEHP, DBP, BBP) and other classes of antiandrogenic
chemicals {e.g., certain pesticides and pharmaceuticals that are discussed more in (	323a))

can induce effects on the developing male reproductive system in a dose-additive manner. EPA details
how it intends to evaluate risk to above-identified PESS from co-exposure to DINP and several other
toxicologically similar phthalates in its Draft Proposed Approach for Cumulative Risk Assessment of

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High-Priority Phthalates and a Manufacturer-Requested Phthalate under the Toxic Substances Control
Act (	2023a).

The effect of other factors on susceptibility to health effects of DINP is not known; therefore, EPA is
uncertain about the magnitude of any possible increased risk from effects associated with DINP
exposure for subpopulations that may be relevant to other factors.

For non-cancer endpoints, EPA used a default value of 10 for human variability (UFh) to account for
increased susceptibility when quantifying risks from exposure to DINP. The Risk Assessment Forum, in
A Review of the Reference Dose and Reference Concentration Processes (	32b). discusses

some of the evidence for choosing the default factor of 10 when data are lacking and describe the types
of populations that may be more susceptible, including different lifestages (e.g., of children and elderly).
U.S. EPA (2002b). however, did not discuss all the factors presented in Table 5-1. Thus, uncertainty
remains whether additional susceptibility factors would be covered by the default UFh value of 10
chosen for use in the draft DINP risk evaluation.

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CSS Evidence Crosswalk for Biological Susceptibility Considerations

Susceptibility
Category

Examples of
Specific
Factors

Direct Evidence this Factor
Modifies Susceptibility to DINP

Description of Interaction

Kcv Citations

Indirect Evidence of Interaction with Target
Organs or Biological Pathways Relevant to
DINP

Description of
Interaction

Key Citation(s)

Susceptibility Addressed in
Risk Evaluation?

Lifestage

Embryos/
fetuses/infants

Direct quantitative animal
evidence for developmental
toxicity (e.g., increased skeletal
and visceral variations,
decreased live births, decreased
offspring body weight gain, and
decreased offspring survival
with increased severity in the
second generation).

There is direct quantitative
animal evidence for effects on
the developing male
reproductive system consistent
with a disruption of androgen
action.

Heltwig et aL

997)

Waterman et al..

999)

Waterman et al..

000)

U.S. EPA,

023a)

U.S. EPA,

2023b)

Pregnancy/
lactating status

Rodent dams not particularly
susceptible during pregnancy
and lactation, except for effects
related to reduced maternal
weight gain, food consumption,
and increased organ weight
(liver and kidney), which
occurred at doses higher than
those that caused developmental
toxicity.

(Heltwig et aL,
.1.997)

(Waterman et aL.
.1.999)

Acute and intermediate duration
PODs for developmental
endpoints protective of effects
in offspring

Acute and intermediate duration
PODs for developmental
endpoints protective of effects
in dams

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Susceptibility
Category

Examples of
Specific
Factors

Direct Evidence this Factor
Modifies Susceptibility to DINP

Indirect Evidence of Interaction with Target
Organs or Biological Pathways Relevant to
DINP

Susceptibility Addressed in
Risk Evaluation?

Description of Interaction

Key Citations

Description of
Interaction

Key Citation(s)

Lifestage

Males of

reproductive

age

Increased testes, right
epididymis, liver, and kidney
weights. There was also
decreased food consumption.

(Waterman et al.

2000; Exxon

Biomedical

1996a)





Use of default lOx UFH

Children

Reduced rodent offspring
bodyweight gain between PNDs
1 to 21 was observed in one and
two-generation studies of
reproduction.

(Waterman et al..

2000; Exxon
Biomedical
1996a. b)





Acute and intermediate duration
PODs for developmental
endpoints protective of effects
of offspring bodyweight gain
Use of default lOx UFH

Elderly

No direct evidence identified







Use of default lOx UFH

Pre-existing
disease or
disorder

Health
outcome/
target organs

No direct evidence identified



Several preexisting
conditions may contribute
to adverse developmental
outcomes (e.g., diabetes,
high blood pressure,
certain viruses).

Individuals with chronic
liver and kidney disease
may be more susceptible
to effects on these target
organs

Viruses such as viral
hepatitis can cause liver
damage.

CDC (2023e)
CDC (2023 g)

Use of default lOx UFH

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Susceptibility
Category

Examples of
Specific
Factors

Direct Evidence this Factor
Modifies Susceptibility to DINP

Indirect Evidence of Interaction with Target
Organs or Biological Pathways Relevant to
DINP

Susceptibility Addressed in
Risk Evaluation?

Description of Interaction

Key Citations

Description of
Interaction

Key Citation(s)

Pre-existing
disease or
disorder

Toxicokinetics

No direct evidence identified



Chronic liver and kidney
disease are associated with
impaired metabolism and
clearance (altered
expression of phase 1 and
phase 2 enzymes,
impaired clearance),
which may enhance
exposure duration and
concentration of DINP.



Use of default lOx UFH

Lifestyle
activities

Smoking

No direct evidence identified



Smoking during
pregnancy may increase
susceptibility for
developmental outcomes
(e.g., early delivery and
stillbirths).

CDC (2023f)

Qualitative discussion in
Section 5.2 and this table

Alcohol
consumption

No direct evidence identified



Alcohol use during
pregnancy can cause
developmental outcomes
(e.g., fetal alcohol
spectrum disorders).

Heavy alcohol use may
affect susceptibility to
liver disease.

CDC (2023d)
CDC (2023a)

Qualitative discussion in
Section 5.2 and this table

Physical
activity

No direct evidence identified



Insufficient activity may
increase susceptibility to
multiple health outcomes.

Overly strenuous activity
may also increase
susceptibility.

CDC (2022)

Qualitative discussion in
Section 5.2 and this table

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Susceptibility
Category

Examples of
Specific

Direct Evidence this Factor
Modifies Susceptibility to DINP

Indirect Evidence of Interaction with Target
Organs or Biological Pathways Relevant to
DINP

Susceptibility Addressed in
Risk Evaluation?

Factors

Description of Interaction

Key Citations

Description of
Interaction

Key Citation(s)

Sociodemo-

Race/ethnicity

No direct evidence identified
(e.g., no information on
polymorphisms in DINP
metabolic pathways or diseases
associated race/ethnicity that
would lead to increased
susceptibility to effects of DINP
by any individual group).







Qualitative discussion in
Section 5.2 and this table

graphic status

Socioeconomic
status

No direct evidence identified



Individuals with lower
incomes may have worse
health outcomes due to
social needs that are not
met, environmental
concerns, and barriers to
health care access.

ODPHP (2023b)





Sex/gender

No direct evidence identified







Use of default lOx UFH

Nutrition

Diet

No direct evidence identified



Poor diets can lead to
chronic illnesses such as
heart disease, type 2
diabetes, and obesity,
which may contribute to
adverse developmental
outcomes. Additionally,
diet can be a risk factor for
fatty liver, which could be
a pre-existing condition to
enhance susceptibility to
DINP-induced liver
toxicity.

CDC (2023e)
CDC (2023b)

Qualitative discussion in
Section 5.2 and this table

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Susceptibility
Category

Examples of
Specific
Factors

Direct Evidence this Factor
Modifies Susceptibility to DINP

Indirect Evidence of Interaction with Target
Organs or Biological Pathways Relevant to
DINP

Susceptibility Addressed in
Risk Evaluation?

Description of Interaction

Key Citations

Description of
Interaction

Key Citation(s)

Nutrition

Malnutrition

No direct evidence identified



Micronutrient malnutrition
can lead to multiple
conditions that include
birth defects, maternal and
infant deaths, preterm
birth, low birth weight,
poor fetal growth,
childhood blindness,
undeveloped cognitive
ability.

Thus, malnutrition may
increase susceptibility to
some developmental
outcomes associated with
DINP.

CDC (2021)
CDC (2023b)

Qualitative discussion in
Section 5.2 and this table

Genetics/
epigenetics

Target organs

No direct evidence identified



Polymorphisms in genes
may increase
susceptibility to liver,
kidney, or developmental
toxicity.



Use of default lOx UFH

Toxicokinetics

No direct evidence identified



Polymorphisms in genes
encoding enzymes (e.g.,
esterases) involved in
metabolism of DINP may
influence metabolism and
excretion of DINP.



Use of default lOx UFH

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Susceptibility
Category

Examples of
Specific

Direct Evidence this Factor
Modifies Susceptibility to DINP

Indirect Evidence of Interaction with Target
Organs or Biological Pathways Relevant to
DINP

Susceptibility Addressed in
Risk Evaluation?

Factors

Description of Interaction

Key Citations

Description of
Interaction

Key Citation(s)



Built

environment

No direct evidence identified



Poor-quality housing is
associated with a variety
of negative health
outcomes.

ODPHP (2023a)

Qualitative discussion in
Section 5.2 and this table

Other

Social

environment

No direct evidence identified



Social isolation and other
social determinants (e.g.,
decreased social capital,
stress) can lead to negative
health outcomes.

CDC (2023c)
ODPHP (2023c)

Qualitative discussion in
Section 5.2 and this table

chemical and
nonchemical
stressors

Chemical co-
exposures

Studies have demonstrated that
co-exposure to DINP and other
toxicologically similar
phthalates (e.g., DEHP, DBP,
BBP) and other classes of
antiandrogenic chemicals (e.g.,
certain pesticides and
pharmaceuticals - discussed
more in (U.S. EPA, 2023a")") can
induce effects on the developing
male reproductive system in a
dose-additive manner.

See fU.S. EPA.
2023a) and (U.S.
EPA. 2023b")





Qualitative discussion in
Section 5.2 and this table and
will be quantitatively addressed
as part of the phthalate
cumulative risk assessment.

2879

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2880	6 POINTS OF DEPARTURE USED TO ESTIMATE RISKS FROM

2881	DINP EXPOSURE

2882	After considering hazard identification and evidence integration, dose-response evaluation, and weight

2883	of scientific evidence of POD candidates, EPA chose two non-cancer endpoints for the risk evaluation—

2884	one for acute and intermediate exposure scenarios and a second one for chronic scenarios (Table 6-1).

2885	HECs are based on daily continuous (24-hour) exposure, and HEDs are daily values.

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2886 Table 6-1. Non-cancer HECs and HEDs Used to Estimate Risks

Exposure
Scenario

Target
Organ
System

Species

Duration

POD

(mg/kg-
day)

Effect

HEC
(mg/mJ) [ppm]

HED
(mg/kg-day)

Benchmark
MOE

Reference

Acute and
Intermediate

Development

Rat

5 to 14 days

throughout

gestation

BMDL5 =

49 a

i fetal testicular
testosterone

63
[3.7]

12

UFa= 3

ufh=io

Total UF=30

(NASEM. 2017)

Chronic

Liver

Rat

2 years

NOAEL
= 15

t liver weight, |
serum chemistry,
histopathology b

19

[1.1]

3.5

UFa= 3

ufh=io

Total UF=30

(Lington et al.,

jOO";

Bio/dvn amies.
t«>So)

" The BMDL5 was derived by NASEM (2017) through meta-regression and BMD modeling of fetal testicular testosterone data from two studies of DINP with rats
(Bobere et al. 2011: Kaunas et al. 2011). R code suroortins NASEM's meta-reeression and BMD analysis of DINP is Diibliclv available throueh GitHub
(httDs://gi thub.com/wachiuDhd/NASEM-2017-Endocrine-Low-Dose).

h Liver toxicity included increased relative liver weight, increased serum chemistry (i.e., AST, ALT, ALP), and histopathologic findings (e.g., focal necrosis, spongiosis
heratis)) in F344 rats following 2 vears of dietary exposure to DINP (Lington et al. 1997; Bio/dynamics. 1986).

2887

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3578	Appendix A EXISTING ASSESSMENTS FROM OTHER REGULATORY AGENCIES OF

3579	DINP

3580	The available existing assessments of DINP are summarized in TableApx A-l, which includes details regarding external peer-review, public

3581	consultation, and systematic review protocols that were used.

3582

3583	Table Apx A-l. Summary of Peer Review, Public Comments, and Systematic Review for Existing Assessments of DINP	

Agency

Asscssmcnt(s) (Reference)

External

Peer-
Review?

Public
Consultation?

Systematic
Review Protocol
Employed?

Remarks

U.S. EPA (IRIS
Program)

Phthalate exposure and male reproductive
outcomes: A systematic review of the
human epidemiological evidence (Hadke et

a.L 2018)

Phthalate exposure and female
reproductive and developmental outcomes:
A systematic review of the human
epidemiological evidence dladke et aL

2019b)

Phthalate exposure and metabolic effects:
A systematic review of the human
epidemiological evidence dladke et aL,
2019a)

Phthalate exposure and neurodevelopment:
A systematic review and meta-analysis of
human epidemiological evidence (Hadke et
al.. 2020a).

No

No

Yes

-	Publications were subjected to peer-review prior to
being published in a special issue of Environment
International

-	Publications employed a systematic review
process that included literature search and
screening, study evaluation, data extraction, and
evidence synthesis. The full systematic review
protocol is available as a supplemental file
associated with each publication.

U.S. EPA

Technical review of diisononyl phthalate
(Final assessment) (U.S. EPA, 2023c)

No

Yes

No

-	Technical review of DINP was reviewed by two
internal EPA reviewers, but was not subjected to
external peer-review

-	Draft technical review of DINP was subjected to a
public review period. Public comments available
here: httos://www.reeulations.eov/docket/EPA-HO-

TRI-2022-0262/comments

U.S. CPSC

Toxicity review of Diisononyl Phthalate
(DINP) (U.S. CPSC. 2010)

Yes

Yes

No

- Peer-reviewed by panel of four experts. Peer-
review report available at:

Page 116 of 184

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A^cncv

Asscssmcnt(s) (Reference)

External

Peer-
Review?

Public
Consultation?

Systematic
Review Protocol
Employed?

Remarks



Chronic Hazard Advisory Panel on
Phthalates andPhthalate Alternatives
(IIS. CPSC. 20141







httDs://www.CDSC.gov/s3ts-Dubiic/Peer-Review-

Reno rt-Co m me nts.odf

-Public comments available at:

httDs://www.cDsc. eov/chao

-	No formal systematic review protocol employed.

-	Details regarding CPSC's strategy for identifying
new information and literature are provided on page

12 of (U.S. CPSC, 2014)

NASEM

Application of systematic review methods
in an overall strategy for evaluating low-
dose toxicity from endocrine active
chemicals fNASEM. 2017)

Yes

No

Yes

-	Draft report was reviewed by individuals chosen
for their diverse perspectives and technical expertise
in accordances with the National Academies peer-
review process. See Acknowledgements section of
(NASEM, 2017) for more details.

-	Employed NTP's Office of Heath Assessment and
Translation (OHAT) systematic review method

Health Canada

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 fEC/HC. 2015)

Supporting Documentation:
Carcinogenicity of Phthalates - Mode of
Action and Human Relevance (Health

Canada. 2015)

Supporting documentation: Evaluation of
epidemiologic studies on phthalate
compounds and their metabolites for
hormonal effects, growth and development
and reproductive parameters (Health
Canada. 2018b")

Yes

Yes

No (Animal
studies)

Yes

(Epidemiologic
studies)

-	Ecological and human health portions of the
screening assessment report (ECCC/HC, 2020) were
subject to external review and/or consultation. See
uase 2 of (ECCC/HC, 2020) for additional details.

-	State of the science rcDort (EC/HC, 2015) and
draft screening assessment report for the phthalate
substance group subjected to 60-day public
comment periods. Summaries of received public
comments available at:
httDs://www.canada.ca/en/health-
canada/services/chemical-substances/substance-
grouDings-initiative/Dhthalate.html#al

-	No formal systematic review protocol employed to
identify or evaluate experimental animal toxicology
studies.

-	Details regarding Health Canada's strategy for
identifying new information and literature are
orovided in Section 1 of (EC/HC, 2015) and

(ECCC/HC. 2020)

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A^cncv

Asscssmcnt(s) (Reference)

External

Peer-
Review?

Public
Consultation?

Systematic
Review Protocol
Employed?

Remarks



Supporting documentation: Evaluation of
epidemiologic studies on phthalate
compounds and their metabolites for
effects on behaviour and
neurodevelopment, allergies,
cardiovascular function, oxidative stress,
breast cancer, obesity, and metabolic
disorders (Health Canada, 2018a)







- Human epidemiologic studies evaluated using
Downs and Black Method (Health Canada. 2018a.

b)



Screening Assessment - Phthalate
Substance Grouping (ECCC/HC, 2020)









NICNAS

Priority existing chemical assessment
report no. 35: Diisononyl phthalate

(NICNAS. 2012)

No

Yes

No

-	NICNAS (2012) states "The report has been
subjected to internal peer review by NICNAS
during all stages of preparation." However, a formal
external peer-review was not conducted.

-	NICNAS (2012) states "Applicants for assessment
are given a draft copy of the report and 28 days to
advise the Director of any errors. Following the
correction of any errors, the Director provides
applicants and other interested parties with a copy
of the draft assessment report for consideration.

This is a period of public comment lasting for 28
days during which requests for variation of the
report mav be made." See Preface of (NICNAS.
2012) for more details.

-	No formal systematic review protocol employed.

-	Details regarding NICNAS's strategy for
identifying new information and literature are
provided in Section 1.3 of (NICNAS, 2012)

ECHA

Evaluation of New Scientific Evidence
Concerning DINP and DIDP in Relation to
Entry 52 of AnnexXVII to REACH
Regulation (EC) No 1907/2006 (ECHA.

2013b)

Yes

Yes

No

-	Peer-reviewed by ECHA's Committee for Risk
Assessment (ECHA, 2013a)

-	Subject to 12-week public consultation

-	No formal systematic review protocol employed..

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Asscssmcnt(s) (Reference)

External

Peer-
Review?

Public
Consultation?

Systematic
Review Protocol
Employed?

Remarks











- Details regarding ECHA's strategy for identifying
new information and literature are provided on
raees 14-15 of (ECHA. 2013b)

EFSA

Update of the Risk Assessment ofDi-
butylphthalate (DBF), Butyl-benzyl-
phthalate (BBP), Bis(2-
ethylhexyl)phthalate (DEHP), Di-
isononylphthalate (DINP) and Di-
isodecylphthalate (DIDP) for Use in Food
Contact Materials (EFSA, 2019)

No

Yes

No

- Draft report subject to public consultation. Public
comments and EFSA's response to comments are
available at:

httDs://doi.ore/10.2903/so.efsa.2019.EN-1747

-	No formal systematic review protocol employed.

-	Details regarding EFSA's strategy for identifying
new information and literature are provided on page
18 and Aroendix B of (EFSA. 2019)

NTP-CERHR

NTP-CERHR monograph on the potential
human reproductive and developmental
effects of di-isononyl phthalate (DINP)
^NTP-CERHR. 2003)

No

Yes

No

-	Report prepared by NTP-CERHHR Phthalates
Expert Panel and was reviewed by CERHR Core
Committee (made up of representatives of NTP-
participating agencies, CERHR staff scientists,
member of phthalates expert panel)

-	Public comments summarized in Appendix III of

(NTP-CERHR. 2003)

-	No formal systematic review protocol employed.

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Appendix B SUMMARY OF LIVER TOXICITY STUDIES

This Appendix contains more detailed information on the available studies described in the liver toxicity
hazard identification (Section 3.2), including information on individual study design.

Humans

No epidemiologic studies were identified by Health Canada (2018a) or by IRIS assessment that
examined the association between DINP and/or its metabolites and biomarkers of liver injury.

New Literature: EPA considered new studies published since Health Canada's assessment (Health
Canada. 2018a) {i.e., studies published from 2018 to 2019); however, no studies were identified that fall
within this date range and evaluated liver injury for DINP and/or its metabolites.

Laboratory Animals

Existing assessments have consistently identified the liver as one of the most sensitive target organs
following oral exposure to DINP in experimental animal studies (ECCC/HC. 2020; EFSA. 2019;

EC/HC. 2015; ECHA. b, - VTNAS. 2012; U.S. CPSC. 20 h'. { P \	N U'

CERHR. 200 , "i v i PSC. 2001). Short-term (>1 to 30 days), subchronic (>30 to 90 days) and chronic
(>90 days) exposure studies have reported significant liver effects. Available studies include: 11 short-
term oral studies (six studies on rats, four studies on mice, 1 study on cynomolgus monkeys); nine
subchronic oral exposure studies (six on rats, one on mice, one on beagle dogs, and one on marmosets)
and five chronic oral exposure studies (four on rats and one on mice) Available studies are summarized
in TableApx B-l, TableApx B-2,and TableApx B-7, and are discussed further below.

Considerations for Interpretation of Hepatic Effects: Consistent with previous guidances (Hall et at..
201 J; 1 c. i i1 \ . K)2a), EPA considered hepatocellular hypertrophy and corresponding increases in
liver size and weight to be adaptive non-adverse responses, unless accompanied by exposure-related,
biologically significant changes in clinical markers of liver toxicity {i.e., decreased albumin; or
increased alanine aminotransferase (ALT), aspartate aminotransferase (AST), alkaline phosphatase
(ALP), gamma glutamyltransferase, bilirubin, cholesterol) and/or histopathology indicative of an
adverse response {e.g., hyperplasia, degeneration, necrosis, inflammation). Further, phthalates, including
DINP, can induce peroxisome proliferation in the livers of mice and rats (Gorton et at.. 2018; Lapinskas
et at.. 2005; Valles et at.. 2003). and EPA considered evidence supporting a role for PPARa activation in
peroxisome-induced hepatic effects of DINP. For purposes of identifying study NOAEL and LOAEL
values, effects consistent with peroxisome proliferation and PPARa activation were also considered
relevant for setting the LOAEL.

Short-Term (>l to 30 Days) Exposure Studies: EPA evaluated 12 short-term exposure animal studies
from existing assessments that evaluated liver effects following oral exposure to DINP (Ma et at.. 2014;
Kwack et at.. 2010; Kwack et at.. 2009; Valles et at.. 2003; Kaufmaim et at.. 2002; Push et at.. 2000;
Smith et at.. 2000; Hilts AG. 1992; Hazleton Labv	t'U*k \ Ps86; Bio/dynamics. 1982a;

Midwest Research Institute. 1981). The database includes six studies in various strains of rat, three
studies in mice, and one study in monkeys. One short-term dermal exposure study in female B6C3F1
mice was identified (Butala et at.. 2004). These studies provide data on relative/and/or absolute liver
weights, histopathology, hepatic enzyme levels and/or activity {e.g., AST, ALT, and ALP), and other
parameters useful to determining the effects of DINP on the liver. These studies are summarized in
Table Apx B-l.

Eight of the available short-term studies reported increases in absolute and/or relative liver weights or
incidences of hepatocyte proliferation or other nonneoplastic lesions following oral exposure to DINP

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(Ma et al.. 2014; Kwack et al.. 2009; Valles et ai. 2003; Kaufmann et ai. 2002; Smith et al.. 2000; Hills
\ ^ rs°2; Hazleton Lahv l Li, \ l 986; Bio/dynamics. 1982a). These observations sometimes
coincided with increases in peroxisomal volume, peroxisomal beta oxidation, and activity of enzymes
such as palmitoyl-CoA oxidase, indicative of PPARa activation, which is discussed in further detail in
the mechanistic section.

The BIBRA (1986) study evaluated the ability of DINP to induce peroxisome proliferation in male and
female F344 rats fed 0, 0.6, 1.2, or 2.5 percent DINP in the diet for 21 days (equivalent to 0, 639, 1,192,
or 2,195 mg/kg-day [males] and 0, 607, 1,193, or 2,289 mg/kg-day [females]). Body weights were
significantly reduced in males (6 to 12 percent decrease) and in females (6 to 14 percent decrease) in a
time- and dose-dependent manner. Food intake was also significantly reduced (19 to 49 percent) in
males and females. Significant dose-dependent increases in absolute and relative liver weight were
observed in males and females beginning in animals from the low dose group (639 mg/kg-day in males;
607 mg/kg-day females). The effects observed on liver weight were considered exposure-related even
though terminal body weights were significantly reduced in males and in females in a dose-dependent
manner, and body weight gain was reduced in animals at the highest dose level. In parallel with the
increases in liver weights, the authors reported dose-dependent increases in cyanide-insensitive
palmitoyl-CoA oxidation levels in males and females of the mid- and high-dose groups, dose-dependent
increases in microsomal protein levels of males and females (all dose levels) and increases in lauric acid
11- and 12-hydroxylase activities in males of the low-dose group (639 mg/kg-day in males).
Hydroxylase activities were increased in high-dose females. The authors also reported decreases in total
cholesterol in males (9 to 24 percent) and females (14 to 24 percent), as well as dose-dependent
decreases in serum triglycerides in males (24 to 48 percent). However, dose-dependent increases in
serum triglycerides (24 to 26 percent) were observed in females. The inconsistency of effects between
sexes is source of uncertainty in the dataset. The authors also examined liver tissue via electron
microscopy and observed increases in peroxisome proliferation in males and females from the highest
exposure groups. However, these effects were not further quantitatively described, which is another
limitation of the dataset.

Data from BIBRA (1986) were consistent with Kwack et al. (2009). In the Kwack study, male SD rats
were administered 0 or 500 mg/kg-day DINP daily via gavage for 4 weeks. Increased relative liver
weight (45 percent) was observed, which coincided with perturbations in several clinical chemistry
parameters. Increases were observed in the serum levels of AST (32 percent), ALP (260 percent), and
triglycerides (53 percent). The observed effects were considered adverse because the liver weight
changes were accompanied by clinical chemistry markers of hepatoxicity. Interestingly, these results
were not wholly consistent with a study by the same authors with a shorter exposure duration (Kwack et
al.. 2010). In that study, male SD rats were again administered to 0 or 500 mg/kg-day DINP daily via
gavage for 2 weeks. Increases in AST levels (31 percent) and ALP (159 percent) were observed as well
as increases in serum triglycerides. There was no change in ALT levels and no significant change in
relative liver weight.

Several other studies reported increases in relative and/or absolute liver weight with concomitant
changes in other hepatic endpoints in B6C3F1 mice (Valles et al.. 2003; Kaufmann et al.. 2002; Smith et
al.. 2000; Hazleton Labs. 1991a) and/or F344 rats (Smith et al.. 2000; Hub \ ^	
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weeks exposure to 6,000 ppm DINP (equivalent to 900 mg/kg-day). The LOEL in each species was the
high-dose of DINP (1,200 mg/kg-day for rats, 900 mg/kg-day in mice). Valles et al. (2003) reported
similar findings in male and female B6C3F1 mice fed diets containing 0, 150, 1,500, 4,000, or 8,000
ppm of DINP (CASRN 68515-48-0) for 2 weeks. Relative liver weight was significantly increased in
both sexes at the two highest dose groups and in females at the mid dose-group. The percent change in
relative liver weight for the high dose group was 37 percent in males and over 50 percent in females.
The other statistically significant increases in females were less than 10 percent over controls, while
relative liver weight in males of the 4,000 ppm increased by almost 17 percent.

Two other studies (Kaufmann et al. 2002; Hazleton Lai	) reported similar findings at lower

doses after similar exposure durations {i.e., 4 weeks). In Kaufmann et al. (2002). male and female
B6C3F1 mice were exposed to 0, 500, 1500, 4000, or 8000 ppm DINP in the diet for 4 weeks
(equivalent to 0, 117, 350, 913, 1860 mg/kg-day [males]; or 0, 167, 546, 1272, or 2806 mg/kg-day
[females]). Significant increases in absolute and relative liver weight were observed in males and
females, which corresponded with increased peroxisomal volume and peroxisomal enzyme activity
(cyanide-insensitive palmitoyl-CoA) at doses as low as 350 mg/kg-day in males or 546 mg/kg-day in
females. The LOEL/NOEL was 350/117 mg/kg-day in males and 546/167 mg/kg-day in females.
Hazleton Labs (1991a) reported similar LOEL values for liver effects in males (635 mg/kg-day) and
females (780 mg/kg-day). That study exposed male and female B6C3F1 mice to 0, 3000, 6000, or
12,500 ppm DINP in the diet for 4 weeks (equivalent to 0, 635, 1,377, 2,689, or 6,518 mg/kg-day
[males]; 0, 780, 1761, 3,287, or 6,920 mg/kg-day [females]) and evaluated liver weights, histopathology,
and serum liver enzymes at study termination. Increases in absolute and relative liver weights were
observed in all male and female exposure groups except the low dose, and increased ALT activity was
observed in males and females from the high dose only. Additional findings included enlarged and
discolored livers, increased incidence of hepatocytomegaly (all male dose groups; all female dose
groups except low dose), and increased incidence of coagulative necrosis and/or separate chronic
inflammatory foci in high-dose males (6,518 mg/kg-day) and females (6,920 mg/kg-day) as well as
females of the 3,287 mg/kg-day group. Similar findings were reported in a study by Ma et al. (2014).
which administered 0.2, 2, 20 or 200 mg/kg-day DINP to male Kunming mice via oral gavage daily for
14 days. This study established a NOAEL at 20 mg/kg-day and a LOAEL at 200 mg/kg-day based on
significantly increased incidences of histopathologic lesions of the liver, including central vein dilation,
congestion, and narrowing of the sinusoid with loose cytoplasm in animals exposed to the highest dose
of DINP.

The findings that support liver toxicity in mice and the rat study by Smith et al. (2000) were consistent
with two additional rat studies. A study by the Midwest Research Institute (1981) fed male and female
F344 rats 0, 0.2, 0.67, or 2 percent DINP in the diet for 28 days (estimated doses: 0, 150, 500, 1,500
mg/kg-day [males]; 0, 125, 420, 1,300 mg/kg-day [females]). Increases in hepatic catalase and carnitine
acetyltransferase activity were observed in low dose males (150 mg/kg-day) and females (125 mg/kg-
day). Increases in absolute and relative liver weight were also observed in the mid dose males (500
mg/kg-day) and females (420 mg/kg-day) with no corresponding change in body weight. Additionally,
Bio/dynamics ( la) administered 0 or 1,700 mg/kg-day DINP in the diet to male rats for 1 week and
then evaluated liver weight, general appearance (i.e., macroscopic observation), and clinical chemistry
parameters, including serum ALP at study termination. At study termination, the treated animals had
increased absolute and relative liver weight, as well as increased body weight, and the authors noted
slight congestion in all lobes of the liver in animals exposed to DINP. No statistically or biologically
significant changes were observed for serum ALP levels. A 14-day study by Hiils AG (1992) exposed
female F344 rats to 0, 25, 75, 150, or 1,500 mg/kg-day and then evaluated liver weights, clinical
chemistry parameters, and histopathology at study termination, as well as activities of several

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microsomal enzymes. In general, effects were observed at the highest dose, including increases in
absolute and relative liver weight, and increases in EROD. A dose-dependent increase was observed in
lauric acid hydroxylase, beginning at 25 mg/kg-day. Of note, this study was not reasonable available to
EPA, and data reported on this study reflect those reported by Health Canada's Hazard Assessment

(EC/HC. 2Q15Y

Not all studies identified in existing assessments reported hepatic effects consistent with peroxisomal
beta-oxidation and/or PPARa activation. Indeed, one study in cynomolgus monkeys (Push et at.. 2000)
reported no effect on relative liver weights, histopathology, or serum chemistry parameters in monkeys
administered 0 or 500 mg/kg-day DINP daily via oral gavage for 14 days.

New Literature: EPA identified one new study published between 2015 and 2020 that provided data on
toxicological effects of the liver following short term exposure to DINP ('Meier et at.. 2018). The
developmental exposure study by Neier et al. (2018) evaluated absolute and relative liver weights as
well as hepatic triglyceride levels in PND21 male and female yellow agouti (Avy) mice. Dams were
administered 0 or 75 ppm DINP in the diet (equivalent to 15 mg/kg-day) beginning 2-weeks before
mating and lasting through PND21. Increased absolute (27.6 percent) and relative (15.5 percent) liver
weights were observed in exposed female offspring at PND21. No significant changes were observed in
males. No significant changes were observed in hepatic triglyceride levels, suggesting that differences in
liver weight were not attributed to increases in lipid accumulation in the liver in this study.

TableApx B-l. Summary of Liver Effects Reported in Animal Toxicological Studies Following

Short-Term Exposure to DI

NP

Brief Study Description
(Reference)

NOAEL/
LOAEL
(m«/k«-day)

Effect at LOAEL

Remarks

Kunming mice (males only);
gavage; 0, 0.2, 2, 20, 200
me/ke-dav: 14 davs (Ma et al.

2014)

20/200

Markers of oxidative
stress (t ROS, j GSH,
|MDA, t 8-OH-dG) &
inflammation (|IL-1, f
TNFa) at > 20 mg/kg-
day

Other liver effects:

Liver histopathology: t incidences of
edema (20 mg/kg-day); central vein
dilation, congestion, edema, &
narrowing sinusoidal with extremely
loose cytoplasm (200 mg/kg-day).

Considerations: BW not reported.

Limitations: Historatholoev
qualitative only (no incidence data or
statistical analysis); organ weight and
clinical chemistry not evaluated

F344 rats (females only);
gavage; 0, 25, 75, 150, 1,500
me/ke-dav: 14 davs (Hills AG.

1992)

25 (LOEL)

t lauric acid hydroxylase
(dose-dependent
beginning at 25 mg/kg-
day)

Other liver effects: t absolute and
relative liver weight at 1,500 mg/kg-
day; t liver microsomal enzyme
activities (pentoxyresorufin O-
desalkylase (PROD) and lauryl-CoA
oxidase) at 1,500 mg/kg-day

F344 rats (both sexes); dietary;
0, 0.2, 0.67, 2% (est. 150, 500,
1,500 mg/kg-day [males]; 0,
125, 420, 1,300 mg/kg-day
Ifcmalcsl): 28 davs (Midwest
Research Institute. 1981)

ND/125
(females)

ND/ 150
(males) (LOEL)

t in hepatic catalase and
carnitine

acetyltransferase activity

Other liver effects: t absolute and
relative liver weight (500 mg/kg-day
[males]; 420 mg/kg-day [females])

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Brief Study Description
(Reference)

NOAEL/
LOAEL
(mg/kg-day)

Effect at LOAEL

Remarks

B6C3F1 mice (both sexes);
dietary; 0, 500, 1500, 4000,
8000 ppm (est. 117,350,913,
1,860 mg/kg-day [males]; 0,
167, 546, 1,272, 2,806 mg/kg-
day [females]); 1 or 4 weeks
(Kaufmann et aL 2002)

117/350
(males)

167/546
(female)

t abs. and rel. liver
weight; t peroxisomal
volume, and
peroxisomal enzyme
activity; t hepatocyte
proliferation in males

Other liver effects:

Liver histopathology: t hepatocyte
proliferation in females at >1272
mg/kg-day.

Considerations: Multiple zones of the
liver examined for quantitative
measurement of hepatocyte
proliferation; BW not reported.

SD rats (males only); oral
gavage; 0, 500 mg/kg-day; 28
davs (Kwack et aL 2009)

ND/500

i body weight gain; f
relative liver weight;
clinical chemistry (f
AST, ALP &
triglycerides)

Considerations: J. bodv weisht sain
(-10%) in DINP exposed mice

F344 rats (both sexes); diet; 0,
0.6, 1.2, 2.5% (est. 639, 1192,
2,195 mg/kg-day [males]; 607,
1,198, 2,289 mg/kg-day
ffemalesl): 21 davs (BIBRA.
1986)

ND/639
(males)

ND/607
(females)

t absolute and relative
liver weight (abs.
increase in males: 136,
150, and 165%; rel.
increase in males: 136,
173, 232%; abs. increase
in females: 124, 164,
and 198%; rel. liver
weights in females: 131,
175,231%)
t 11-and 12-
hydroxylase activity,
hypolipidemic effects

Considerations: Bodv weishts and
food intake were significantly reduced
in males (6 to 12%) and in females (6
to 14% decrease). Food intake was
also significantly reduced (19 to 49%)
in males and females.

B6C3F1 mice (both sexes);
dietary; 0, 3000, 6000, 12,500
ppm (est. 635, 1,377, 2,689,
6,518 mg/kg-day [males]; 780,
1761, 3,287, 6,920 mg/kg-day
1 Females 1): 4 weeks (Hazleton
Labs. 1991a)

ND/635 (males)

ND/780

(females)

(LOEL)

Enlarged and discolored
livers; t incidence of
hepatocytomegaly

Other liver effects:
t incidence of coagulative necrosis
and/or separate chronic inflammatory
foci.

B6C3F1 mice (males only);
dietary; 0, 500, 6000 ppm (est.
0, 75, 900 mg/kg-day); 2 or 4
weeks (Smith et aL 2000)°

75 (NOEL)/
900 (LOEL)

t in relative liver weight
at 4 weeks

Other liver effects: t PBOX. t DNA
synthesis; inhibition of GJIC

Limitations: BW not rcDortcd

F344 rats (males only); dietary;
0, 1000, 12,000 ppm (est. 0, 100,
1200 mg/kg-day); 2 or 4 weeks

(Smith et aL. 2000)°

100

(NOEL)/ 1200
(LOEL)

t in relative liver weight
at 4 weeks

Other liver effects: t PBOX. t DNA
synthesis; inhibition of GJIC
Considerations: significant increases
in relative liver weight observed at 4-
week but not 2-week timepoint.
Limitations: onlv males were
evaluated.

F344 rats (males only); dietary;
0, 2% (est. 1,700 mg/kg-day); 7
davs (Bio/dvnamics, 1982a)

ND/1,700

t abs. and rel. liver
weight; macroscopic
liver observations;
changes in clinical
chemistry (J,
triglycerides)



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Brief Study Description
(Reference)

NOAEL/
LOAEL
(m«/k«-day)

Effect at LOAEL

Remarks

Cynomolgus monkeys (males
only); 0, 500 mg/kg-day; oral
eavaee: 14 davs (Push et al.

2000)

500/ND



No statistically or biologically
significant effects were observed

SD rats (male and female); 0 or
500 mg/kg-day; gavage; 14 days
(Kwack et al. 2010)

ND/500

t AST activity (31%), t
ALP (159%); | serum
triglycerides

Other liver effects: liver weishts.
serum biochemistry, and urinalysis

Considerations: No chanee in serum
ALT

" Dose equivalent calculated from 75 mg DINP/kg chow/day based on the assumption that pregnant and nursing female
mice weigh approximately 25g and eat approximately 5 g chow/day.

b Data for the Huls AG studv (1992) were not reasonably available to EPA; Data here reflect those reported bv Health
Canada's Hazard Assessment (EC/HC. 2015).

c Smith et al. (2000) evaluated two isomers of DINP: DINP-1 [CAS 68515-48-0] and DINP-A [CAS 71549-78-5], The
DINP-A isomer is outside the scope of the hazard evaluation; all results herein refer to the DINP-1 isomer.

Sub-chronic (>30 to 90 Days) Exposure Studies: EPA identified nine studies from existing assessments
that provide data on the toxicological effects of DINP on the liver following subchronic duration oral
exposure, including six studies in rats (Hazleton Laty l lb, ^ V t l \ , «11<> dynamics. 1982b. c;
Hazleton Lai 1_, 1971). one in mice (Hazleton Labs. 1992). one study in dogs (Hazleton
Laboratories. 1971). and one study in marmoset monkeys (Hall et at.. 1999). The available studies are
summarized in Table Apx B-2 and discussed further below. One dermal exposure study in New Zealand
white rabbits was also available (Hazleton Laboratories. 1969).

The lowest achieved dose across these rodent studies was 50 mg/kg-day and the highest was 5,770
mg/kg-day (Table_Apx B-2). All studies reported increases in absolute and/or relative liver weight,
sometimes in parallel with exposure-related histopathological effects on the liver (e.g., hepatocytic
hypertrophy), and sometimes coinciding with increases in liver enzymes (i.e., ALT, ALP), suggesting
impaired liver function. These data suggest that the liver is a target organ for DINP, which is consistent
with conclusions from previous assessments by regulatory agencies.

Hazleton Laboratories ( ) reported increased absolute and relative liver weights in both sexes at 500
mg/kg-day as well as exposure-related changes in liver histopathology in males (hepatocytic
hypertrophy throughout the panlobular section). In that study, albino rats were exposed to 0, 50, 150, or
500 mg/kg-day DINP for 13 weeks via diet. Two additional dietary exposure studies in rats by Hazleton
Labs (1991b. 1981) reported increased liver weights, and increased incidences of histopathological
lesions or altered clinical chemistry parameters that suggest liver toxicity. Consistent with the earlier
Hazleton study ( ), Hazleton Labs ( ) found evidence to suggest liver toxicity in F344 rats
exposed to 0, 2500, 5,000, 10,000 or 20,000 ppm DINP for 13 weeks via feed (equivalent to 0, 176, 354,
719, or 1,545 mg/kg-day [males]; 0, 218, 438, 823, or 1,687 mg/kg-day [females]). Increases in absolute
and relative liver weight were accompanied by hepatocellular enlargement in the highest treatment
group. The LOEL was 176 mg/kg-day in males and 218 mg/kg-day in females based on increased liver
weights.

Another study from Hazleton Labs (1981) administered 0, 1,000, 3,000, or 10,000 ppm DINP to male
and female albino rats for 13 weeks in feed (equivalent to 0, 60, 180, or 600 mg/kg-day). Exposure
related increases in absolute and relative liver weights were observed in males and females from the

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high dose groups (absolute weights: 33 percent increase in males, 23.3 percent increase in females;
relative liver weights: 30.2 percent increase in males; 33.3 percent in females). Unlike the other
Hazleton rat studies (1991b. 1971). exposure-related nonneoplastic lesions in the liver were not
observed, although hepatocellular degeneration was noted in two individual high-dose (600 mg/kg-day)
males. Moreover, the authors note that exposure-related changes in histopathology were limited to the
kidneys of high dose males. Dose-related decreases in several clinical chemistry parameters were
observed in both sexes, including total protein, globulin, and total bilirubin, apart from total bilirubin
from males of the mid-dose group (180 mg/kg-day). The decrease in globulin levels reached statistical
significance in mid- (180 mg/kg-day) and high-dose (600 mg/kg-day) females. Decreased bilirubin
reached statistical significance in high-dose males.

Two similarly designed studies in rats from Bio/dynamics (1982b. c) also reported increased absolute
and/or relative liver weight at similar doses in parallel with changes in clinical chemistry parameters. In
the first Bio/Dynamics study, male and female F344 rats were administered 0, 0.1, 0.3, 0.6, 1.0, or 2.0
percent DINP in diet for 13 weeks (equivalent to 0, 77, 227, 460, 767, or 1,554 mg/kg-day)

(Bio/dynami 2b). In the second study, male and female SD rats were administered 0.3 or 1.0
percent DINP in diet for 13 weeks (equivalent to 0, 201 or 690 mg/kg-day [males]; 0, 251 or 880 mg/kg-
day [females]) (Bio/dynamics. 1982c). In the first study, increased absolute and relative liver weights
and decreased cholesterol were observed in females exposed to 227 mg/kg-day (LOAEL)

(Bio/dynami 2b). Other effects included increases in ALT in the two highest doses in males (767
or 1,554 mg/kg-day) and highest dose in females. In the second study, increased relative liver weight
and decreased serum triglyceride levels were observed in males exposed to doses as low as 201 mg/kg-
day and females exposed to 251 mg/kg-day (LOEL), as well as at higher doses. These changes were
accompanied by a 49 or 53 percent increase in ALP (in males or females, respectively) and 31 percent
increase in ALT (males) in rats from the high dose groups. In both studies, terminal body weight was
decreased by at least 10 percent in high-dose males and females. In the SD rat study, terminal body
weight was also reduced in the low dose animals by 24 percent (males; 201 mg/kg-day) or over 15
percent (females; 25 1 mg/kg-day) (Bio/dynamics. 1982c).

An additional study from BASF ( 7) reported effects on clinical chemistry and other hepatic changes
related to hepatotoxicity with similar LOAELs to the Bio/dynamics studies. In that study, male and
female Wistar rats were fed 0, 3000, 10,000, or 30,000 ppm DINP in the diet for 13 weeks (equivalent to
0, 152, 512, 1,543 mg/kg-day [males]; 0, 200, 666, 2,049 mg/kg-day [females]). Decreased triglyceride
levels and peripheral fat deposits in hepatocytes were reported in low-dose male (152 mg-kg-day) and
female (200 mg/kg-day) rats. Increased absolute and relative liver weights were observed at 1,101
mg/kg-day [males] and 1214 mg/kg-day [females]), which are doses much higher than those in which
increased liver weights were observed in the two Bio/dynamics studies (1982b. c). The BASF study
(1987) was not reasonably available to EPA in English; it was identified from Health Canada's Hazard
Assessment (EC/HC. 2015) and therefore is not further considered.

One sub chronic duration study in mice provided evidence that the liver is a target of DINP (Hazleton
Labs. 1992). In that study, male and female B6C3F1 mice were administered 1500, 4000, 10,000, or
20,000 ppm DINP (equivalent to 365, 972, 2,600, or 5,770 mg/kg-day) in the diet for 13 weeks.

Increases in absolute and relative liver weight, as well as histopathologic effects such as hepatocyte
enlargement, liver degeneration, necrosis, and pigment in Kupffer cells as well as in the bile canaliculi
were observed in the 972 mg/kg-day group (LOAEL). One limitation of this study was the small sample
size, which results in limited statistical power to detect differences between treated groups and controls.

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Not all studies have consistently demonstrated the liver toxicity of DINP. Indeed, studies in non-rodent
species, including one study in beagle dogs (Hazleton Laboratories. 1971) and one study in marmoset
monkeys (Hall et al. 1999). have reported contrasting findings. In a study by Hazleton Laboratories
(1971). 0, 0.125, 0.5, 2 percent DINP was administered to beagles in the diet for 13 weeks (equivalent to
0, 37, 160, or 2,000 mg/kg-day). Increases in absolute and relative liver weights were observed at 160
mg/kg-day in males and 2,000 mg/kg-day in both sexes. Histopathologic changes were also observed,
including hepatocyte hypertrophy associated with decreased prominence of hepatic sinusoids at 2,000
mg/kg-day in both sexes. Serum ALT levels increased by 37 percent in males and 48 percent in females
from week 4 at 160 and 2,000 mg/kg-day. Dose-responsive increases in ALT levels were observed in
males (47, 32 and 60 percent increase) and females (48, 74, and 107 percent increase) at study
termination. Limitations of this study include the small sample size and lack of statistical analysis,
which increase uncertainty in the data from this study. Nevertheless, existing assessments of DINP have
supported NOAEL and LOAEL values of 37 and 160 mg/kg-day based on increased absolute and
relative liver weights accompanied with histopathological changes at the highest dose (2,000 mg/kg-
day) tested (EC/HC. 2015). or a LOAEL of 37 mg/kg-day with no NOAEL based on increase liver
weight and serum ALT (ECHA. 2013b; ECB. 2003). Additional limitations of this study include
reporting deficiencies, including the lack of statistical analyses and inconsistencies between text and
tables. These limitations increase uncertainty in the data from this study.

In contrast, a study in marmoset monkeys by Hall et al. (1999) did not observe any statistically
significant liver effects. In that study, male and female marmoset monkeys were administered 0, 100,
500, or 2,500 mg/kg-day DINP daily via oral gavage for 13 weeks. Exposure to DINP increased liver
weight in males, but the effect was not dose-dependent nor statistically significant at any dose, which the
authors attribute to low sample size and high variability.

New Literature: EPA did not identify any new studies published from 2015 through 2020 that provided
data on toxicological effects of liver following chronic exposure to DINP.

TableApx B-2. Summary of Liver Effects Reported in Animal Toxicological Studies Following
Subchronic Exposure to DINP			

Brief Study Description
(Reference)

NOAEL/LOAEL
(mg/kg-day)

Effect at LOAEL

Comments

Beagle dogs (both sexes);
dietary; 0, 0.125, 0.5, 2% (est.
37, 160, 2,000 mg/kg-day); 13
weeks (Hazleton Laboratories.
1971)

37/160

t abs. and rel. liver wt.;
t ALT activity

Other liver effects:

Hepatocytic hypertrophy associated
with decreased prominence of
hepatic sinusoids at 2000 mg/kg-day.
Hepatocytic cytoplasm varied from
fine granular to vacuolated
appearance.

Considerations: No NOAEL
established due to absence of
statistical analysis and some
inconsistencies in data reporting (i.e.,
text and tables in the study).

F344 rats (both sexes); dietary;
0,0.1,0.3,0.6, 1.0, 2.0% (est.
0, 77, 227, 460, 767, 1,554
mg/kg-day); 13 weeks
(B io/dvnamics. 1982b)

77/ 227

t abs. and rel. liver wt.;
i cholesterol (females)

Other liver effects: t ALT (males of
767 and 1,554 mg/kg-day males;
1,554 mg/kg-day females); j
cholesterol (227, 460, 767, and 1,554
mg/kg-day females)

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Brief Study Description
(Reference)

NOAEL/LOAEL
(mg/kg-day)

Effect at LOAEL

Comments







Considerations:

i BW gains in the 767 mg/kg-day
males, j terminal BW (> 10%) at
1554 mg/kg-day in both sexes.

Wistar rats (both sexes);
dietary; 0, 3000, 10,000,
30,000 ppm (est. 0, 152, 512,
1,543 mg/kg-day [males]; 0,
200, 666, 2,049 mg/kg-day
Ifcmalcsl): 13 weeks ((BASF.
1987) as cited by Health
Canada (EC/HC. 2015V)3

ND/152 (males)

ND/ 200
(females)

Clinical chemistry and
liver changes related to
hepatotoxicity
(J, triglyceride level and
i peripheral fat deposits
in hepatocytes)

Considerations:

I BW in males at 152 and 1543
mg/kg-day. Insufficient information
to discern if reported BW was
terminal or B W change.

F344 rats (both sexes); dietary;
0, 2500, 5000, 10,000, 20,000
ppm (est. 0, 176, 354, 719,
1545 mg/kg-day [males]; 0,
218, 438,

823, 1,687 mg/kg-day
Ifcmalcsl): 13 weeks (Hazleton

Labs. 1991b)

ND/176 (males)
ND/218 (females)

t liver weights

Other liver effects:

Hepatocellular enlargement at the
highest dose.

Considerations:

i BW gain at 1545 mg/kg-day in
males and females. j terminal BW
> 10%. (Body weight gains were
decreased in both sexes at 1545
mg/kg-day, along with decreases in
terminal body weight >10% relative
to controls).

SD rats (both sexes); dietary; 0,
1000, 3000, 10,000 ppm (est. 0,
60, 180, 600 mg/kg-day); 13
weeks

LOEL = 180

i total protein and
globulin levels (males)

Other liver effects: t liver weishts
(high dose (both sexes); j total
protein, and total bilirubin
Considerations: historatholoeical
findings limited to the kidney

SD rats (both sexes); dietary; 0,
0.3, 1.0% (est. 201, 690 mg/kg-
day [males]; 251, 880 mg/kg-
day [females]); 13 weeks

(B io/dvnaniics, 1982c)

ND/201 (males;
LOEL)

ND/251 (females;
LOEL)

I terminal body
weights in both sexes; t
abs. and rel. liver wt.
accompanied by j in
triglycerides.

Other liver effects: t ALP (males &
females) and t ALT (males) from the
high dose groups
Considerations:

I Terminal B W by 24% and 28% in
201 mg/kg-day and 690 mg/kg-day
males, respectively. j Terminal BW
by >15% and 31% in 25 lmg/kg-day
and 880 mg/kg-day females,
respectively.

Albino rats (both sexes);
dietary; 0, 50, 150, 500 mg/kg-
dav: 3 months (Hazleton Labs.
1971)

150 (NOEL)/500
(LOEL)

t abs. and rel. liver wt.
and t hepatocyte
hypertrophy

Considerations:

Slight non-significant j BW gain in
500 mg/kg-day males. BW gain
similar across all female groups.
Terminal BW within 10% of controls
for all male and female exposed
groups.

B6C3F1 mice (both sexes);
dietary; 0, 1500, 4000, 10,000,
20,000 ppm (est: 0, 365, 972,
2,600, 5,770 mg/kg-day); 13
weeks (Hazleton Labs. 1992)

365/972

t abs. and rel. liver wt;
hepatocyte
enlargement; other
histopathology in liver
[i.e., pigments in
Kupffer cells and bile

Considerations: J. BW sain and J.
terminal BW of males and females at
5770 mg/kg-day.

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Brief Study Description
(Reference)

NOAEL/LOAEL
(mg/kg-day)

Effect at LOAEL

Comments





canaliculi, liver
degeneration/ necrosis]



Marmoset (both sexes); 0, 100,
500, 2,500 mg/kg-day; oral
savase; 13 weeks (Hall et al.

1999)

500/ND

i body weight and
body weight gain

Considerations: J. relative liver
weight (males) but not dose-
dependent & did not reach statistical
significance

" The BASF studv (1987) was only available in German; EPA reports its use bv Health Canada's Hazard Assessment
(EC/HC. 2015).

Chronic (>90 days) Exposure: EPA identified five studies from existing assessments that provide
information on the toxicological effects of DINP on the liver, including two oral exposure studies
conducted in F344 rats (Covance Labs. 1998c; Lington et ai. 1997). one oral study in SD rats
(Bio/dynami 7), one oral exposure study conducted in B6C3F1 mice (Covance Labs. 1998b). and
a combined one and two generation study in SD rats (Waterman et at.. 2000; Exxon Biomedical. 1996a.
b). No chronic exposure data on DINP are available for humans or other primates. Available studies are
summarized in TableApx B-7.

Two studies in F344 rats reported similar findings, most notably of nonneoplastic lesions of the liver
including spongiosis hepatis (Covance Labs. 1998c; Lington et at.. 1997). Lington et al. (1997)
administered 0, 300, 3,000, or 6,000 ppm DINP to F344 rats in the diet for up to 24 months,
corresponding to mean daily intakes of 0, 15, 152, or 307 mg/kg-day in males and 0, 18, 184, or 375
mg/kg-day in females, respectively. Male and female rats in the mid- and high-dose groups had
statistically significant increases in absolute and relative liver weights throughout the exposure period
and study termination, where relative weight increased 19 to 31 percent in males and 16 to 29 percent in
females. Increases in liver weight corresponded with increases in liver enzyme levels. In males, dose-
related increases of 1.5- to 3-fold were observed in ALP, AST, and ALT activities of mid- and high-dose
groups throughout the study. No significant differences were observed in females. Increased incidences
of several non-neoplastic histopathological lesions were observed in the liver at 18 months, including
minimal to slight centrilobular to midzonal hepatocellular enlargement in high-dose males (incidence:
9/10 vs. 0/10 in controls) and females (10/10 vs 0/10 in controls). At study termination {i.e., 24 months),
dose-related increases were observed in the incidence of focal necrosis, spongiosis hepatis, sinusoid
ectasia, hepatocellular enlargement, and hepatopathy associated with leukemia (Table Apx B-3). The
study authors did not report statistical significance for any of the observed lesions. EPA conducted an
independent review of the incidences of spongiosis hepatis and hepatopathy associated with leukemia
and determined that these histopathology findings were significantly increased in mid- (152 mg/kg-day)
and high-dose (307 mg/kg-day) male rats (Table Apx B-3). Additionally at the high dose in the males,
the incidences of sinusoid ectasia, hepatocellular enlargement, and focal necrosis were significantly
increased over controls. In females, dose-related increases in the incidence of focal necrosis,
hepatopathy associated with leukemia, and hepatocellular enlargement were noted at study termination.
The independent statistical analysis determined that the incidences of hepatocellular enlargement and
hepatopathy associated with leukemia were significantly increased in high-dose females. The NOAEL
and LOAEL for non-cancer hepatic effects in this study were 15 and 152 mg/kg-day, respectively; both
are based on a statistically significant increase in the incidence of spongiosis hepatis in mid-dose male
rats that was accompanied by increased absolute and relative liver weights and changes in serum
enzyme activities.

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TableApx B-3. Incidence of Selected Non-neoplastic Hepatic Lesions in F344 Rats Exposed to
DINP for 24 Months (Liiigton et ai. 1997) 	

Lesion

Dose Group
mg/kg-day (ppm)

Control

15 M/18 F
(300)

152 M/184
(3,000)

307 M/375
(6,000)

Males"

Spongiosis hepatis

24/81
(29.6%)

24/80
(30%)

51/80*
(63.8%)

62/80*
(77.5%)

Hepatopathy
associated
with leukemia

22/81
(27.2%)

17/80
(21.3%)

34/80*
(42.5%)

33/80*
(41.3%)

Sinusoid ectasia

16/81
(19.8%)

16/80
(20.0%)

24/80
(30.0%)

33/80*
(41.3%)

Hepatocellular
enlargement

1/81
(1.2%)

1/80
(1.3%)

1/80
(1.3%)

9/80*
(11.3%)

Focal necrosis

10/81
(12.3%)

9/80
(11.2%)

16/80
(20.0%)

26/80*
(32.5%)

1 vi miles

l ocal necrosis

13/81
(16.0%)

11/81
(13.6%)

19/80
(23.8%)

21/80
(26.3%)

Spongiosis hepatis

4/81
(4.9%)

1/81
(1.2%)

3/80
(3.8%)

4/80
(5.0%)

Sinusoid ectasia

9/81
(11.1%)

4/81
(4.9%)

6/80
(7.5%)

10/80
(12.5%)

Hepatocellular
enlargement

1/81
(1.2%)

0/81
(0%)

0/80
(0%)

11/80*
(13.8%)

Source: Table 7 in Lington et al. (1997)

M = male; F = female

"Number of animals with lesion/total number of animals examined. Percent lesion incidence in parentheses.
* Statistically significant at p < 0.05 when compared to the control incidence using Fischer's Exact test;
statistical analysis performed by EPA.

Another 2-year study in F344 rats with comparable dose levels to Lington et al. (1997) provided data to
support the liver toxicity of DINP (Covar s. 1998c). In that study, DINP was administered to rats
at dietary concentrations of 500, 1,500, 6,000 or 12,000 ppm (equivalent to average daily doses of 29,
88, 359, or 733 mg/kg-day in males, and 36, 109, 442, or 885 mg/kg-day in females for 104 weeks.
Additional groups of male and female rats were given 12,000 ppm (637 and 774 mg/kg-day,
respectively) for 78 weeks and received basal diet only for the remainder of the study (26 weeks) to
evaluate the reversibility of DINP toxicity (recovery group). Increased absolute and relative liver
weights were observed in the two highest dose groups in males and females at multiple timepoints
throughout the study as well study termination. Relative liver weights were increased 35 to 61 percent in
males and 26 to 71 percent in females. There were no significant changes in absolute liver weights in the
recovery group at the end of the 26-week recovery period, suggesting a reversibility of liver
enlargement. Significant increases in activities of serum enzymes (AST and ALT) were also observed in
both sexes at the two highest doses at weeks 52, 78, and study termination. Serum liver enzyme
activities were also increased in the recovery group. Increases in palmitoyl-CoA oxidase activity were
observed in high dose male and female rats, which is further discussed in the mechanistic section below.

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Histological evidence of liver toxicity was observed in parallel with increases in liver weight and
alterations in serum enzyme activity. Incidences of select non-neoplastic lesions from the Covance study
are summarized in TableApx B-4. A dose-responsive increase in the incidence of spongiosis hepatis
was observed at doses as low as 359 mg/kg-day in males. Other lesions observed in males, such as
cytoplasmic eosinophilia, diffuse hepatocellular enlargement, pigment, and individual cell degeneration
or necrosis were generally observed at higher doses, suggesting spongiosis hepatis was the most
sensitive histopathological response to DINP. EPA's independent review determined that diffuse
hepatocellular enlargement was significantly increased in high-dose males and females at study
termination.

Table Apx B-4. Incidence of Selected Hepatic Lesions in F344 Rats Exposed to DINP in the Diet

for 2 Years (C ovan- < 1 'Us. 1998c)



Dose Group mg/kg-day (ppm)

Lesion



29 Ml

88 Ml

359 M/

733 Ml

Recovery" 637



Control

36 F

109 F

442 F

885 F

M/ 774 F





(500)

(1,500)

(6,000)

(12,000)

(12,000)







\la

OS





Spongiosis

5/55b

5/50

2/50

13/55*

21/55*

9/50

hepatis

(9.1%)

(10.0%)

(4.0%)

(23.6%)

(38.2%)

(18.0%)

Cytoplasmic

0/55

0/50

0/50

0/55

31/55*

0/50

eosinophilia

(0%)

(0%)

(0%)

(0%)

(56.4%)

(0%)

Diffuse

0/55

0/50

0/50

0/55

17/55*

0/50

hepatocellular

(0%)

(0%)

(0%)

(0%)

(30.9%)

(0%)

enlargement













Increased

1/55

0/50

1/50

0/55

7/55*

9/50

pigment

(1.8%)

(0%)

(2.0%)

(0%)

(12.7%)

(18.0%)

Individual cell

0/55

0/50

0/50

1/55

5/55*

0/50

degeneration/

(0%)

(0%)

(0%)

(1.8%)

(9.1%)

(0%)

necrosis













Ivimik-s

Spongiosis

0/55

0/50

0/50

1/55

2/55

0/50

hepatis

(0%)

(0%)

(0%)

(1.8%)

(3.6%)

(0%)

Cytoplasmic

0/55

0/50

0/50

0/55

35/55*

0/50

eosinophilia

(0%)

(0%)

(0%)

(0%)

(63.6%)

(0%)

Diffuse

0/55

0/50

0/50

0/55

33/55*

0/50

hepatocellular

(0%)

(0%)

(0%)

(0%)

(60.0%)

(0%)

enlargement













Increased

7/55

8/50

9/50

5/55

16/55*

10/50

pigment

(12.7%)

(16.0%)

(18.0%)

(9.1%)

(29.1%)

(20.0%)

Individual cell

0/55

0/50

0/50

0/55

0/55

0/50

degeneration/

(0%)

(0%)

(0%)

(0%)

(0%)

(0%)

necrosis













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Dose Group mg/kg-day (ppm)

Lesion



29 M/

88 Ml

359 M/

733 Ml

Recovery" 637



Control

36 F

109 F

442 F

885 F

Ml 174F





(500)

(1,500)

(6,000)

(12,000)

(12,000)

Source: Tables 10A and 10C in Covance Labs (1998c)

M = male; F = female

* = significantly different from control (p < 0.05) by Fisher's Exact test as performed by EPA.
a The 12,000 ppm recovery group received 12,000 ppm DINP in the diet for 78 weeks, followed by a 26-
week recovery period during which the test animals received basal diet alone.

h Number of animals with lesion/number of animals with livers examined; percentage is given in parentheses.
Incidence is sum of lesions observed in unscheduled deaths and at terminal sacrifice.

A third study in rats by Bio/dynamics (1987) provided data on liver weights, histopathology, and effects
on clinical chemistry parameters following chronic exposure to DINP. In that study, male and female
SD rats were administered 0, 500, 5,000, or 10,000 ppm DINP in the diet for up to 2-years (equivalent to
0, 27, 271, or 553 mg/kg-day in males and 0, 33, 331, or 672 mg/kg-day in females). Increased absolute
and relative liver weights were observed in high-dose males and females at the 12-month interim
sacrifice and study termination; all increases were between 14 and 34 percent. In the mid-dose females,
there were non-significant increases in absolute (14 percent) and relative (11 percent) liver weight at
interim sacrifice and absolute liver weight (15 percent) at terminal sacrifice, and a significant increase in
relative liver weight (16 percent) at terminal sacrifice. In mid-dose males, a nonsignificant increase of
11 percent was seen in the mid-dose group at interim sacrifice. Histopathological findings were
observed at lower doses than changes in liver weights. Increased incidences of spongiosis hepatis and
minimal-to-slight hepatic focal necrosis were observed in males from the mid-dose group (271 mg/kg-
day). The increases in liver weights and incidences of nonneoplastic lesions were attributed to the
administration of DINP. Incidences of select non-neoplastic lesions from the Bio/dynamics (1987) study
are summarized in Table Apx B-5.

In parallel with increases in liver weight and histopathological findings, changes in clinical chemistry
parameters were observed. Serum ALT was significantly increased in high-dose males at interim
sacrifices on months 6, 12, and 18 by 292, 203, and 232 percent, respectively. A non-statistically
significant increase of 218 percent was observed in males at study termination (24 months). Serum ALP
was significantly increased at months 6 and 12 in high-dose males by 88 and 76 percent, respectively.
Non-significant increases in AST were observed in males from the mid and high dose groups. In
females, non-significant increases in AST (63 percent) and ALT (89 percent) were observed at 6
months. Serum ALP was significantly increased in females of the high-dose group by 81 percent at 18
months, while a non-significant increase of 38 percent was observed at study termination. No exposure-
related changes in serum ALP were observed at earlier timepoints in this group or in females of the low-
or mid-dose groups. The increased serum AST, ALT, and ALP in treated males were for the most part
not statistically significant; however, these findings were considered treatment-related due to the
consistency with which they were noted in the treated males at most timepoints. The increased ALP in
females of the high-dose group at month 18 and month 24 is considered treatment-related and adverse.
However, the increased AST and ALT values in females of the high-dose group at month 6 were not
considered treatment-related due to their isolated occurrence in only one animal at only one timepoint.
Moreover, data from this animal were considered to be statistical outliers via the Grubb's outlier test.

Overall, the Bio/dynamics study (1987) supports a NOAEL of 27 mg/kg-day in male rats based on
treatment related increases in histopathologic lesions {i.e., spongiosis hepatis, focal necrosis) and
increases in serum ALT, AST, and ALP at the LOAEL of 271 mg/kg-day.

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TableApx B-5. Overall Incidence of Selected Tumors in Male and Female Sprague Dawley Rats Exposed
to DINP for 2 Years (Bio/dvnamics. 1987)	

Lesion

Dose Group
mg/kg-day (ppm)"

Control

27 MI 33 F
(500 ppm)

271 Ml 331F
(5,000 ppm)

553 M/ 672 F
(10,000 ppm)

Males

nb

70 (57)c

69 (57)

69 (59)

70 (59)

Hepatocellular carcinoma^

2

2

6

4

Neoplastic nodulefs)



5

lemales

6

5

n

70 (59)

70 (56)

70 (60)

70 (59)

Hepatocellular carcinoma

ot

0

5

7*

Neoplastic nodule(s)

1

1

5

2

Source: Aroendix K. Fieure 1. dd. 11 (uu. 426 of the studv rcDort PDF) (Bio/dvnamics. 1987).

Statistical significance for an exposed group indicates a significant pairwise test. Statistical significance for the vehicle
control group indicates a significant trend test.

M = males; F = females; ppm = parts per million

* Statistically significant (p < 0.05) from the control group by a two-tailed Fisher's exact test
t Statistically significant trend (p < 0.05) based on a Chi-square contingency trend test calculated for this review.
" Equivalent doses in mg/kg-day, administered doses in ppm

h Number of animals with tissue examined microscopically; includes all animals throughout the study; i.e., including the
interim sacrifice, the terminal sacrifice, and unscheduled deaths.

c Sample size excluding animals that died or were sacrificed early, which was used for performing statistical analysis for
hepatocellular carcinoma.

d Number of animals with lesion. Percent lesion incidence in parentheses.

One chronic study in mice by Covance Labs (1998b) was identified from existing assessments. Covance
Labs exposed male and female B6C3F1 mice to 500, 1,500, 4,000, or 8,000 ppm DINP for at least 104
weeks. These concentrations corresponded to average daily doses of 0, 90, 276, 742, and 1,560 mg/kg-
day in males and 0, 112, 336, 910, and 1,888 mg/kg-day in females. Evidence of liver toxicity was
observed in treated animals of both sexes. At interim sacrifice, significant increases were observed in
relative liver weights in mid-dose males (742 mg/kg-day) and females (910 mg/kg-day) and in high-dose
males (1,560 mg/kg-day). At study termination, significant increases were observed in absolute (13 to
33 percent increase) and relative (25 to 60 percent increase) liver weights in males exposed to 742 or
1,560 mg/kg-day DINP. Relative liver weight was also significantly increased 32 percent in the recovery
group. In females, increases in absolute liver weight (18 to 34 percent increase) and relative liver weight
(24 to 39 percent) were observed in females exposed to 910 or 1,888 mg/kg-day DINP, as well as in the
recovery groups. However, the responses were not statistically significant.

Exposure-related changes in serum chemistry profiles were also observed and supported the liver as a
target organ. AST and ALT activities were increased in high-dose males (1,560 mg/kg-day) and
recovery group males and females. Exposure-related increases in the serum levels of total protein,
albumin, and globulin were also observed in high-dose males. Increases in albumin and globulin were
also observed in recovery males.

Gross findings, including liver masses, occurred with greatest frequency at the 910 and 1,560 mg/kg-day
dose groups, as well as the recovery group. These masses corresponded to hepatocellular neoplasms or
involvement by lymphoma or histiocytic sarcoma and are discussed further in (	024a).

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Increased incidences of several nonneoplastic lesions were observed in the livers of high-dose males and
females, including cytoplasmic eosinophilia, diffuse slight to moderate hepatocellular enlargement, and
slight to moderate pigment (TableApx B-6). These changes were also observed in the recovery group,
but generally at lower incidences than in the high-dose groups. No other statistically significant or dose-
related nonneoplastic lesions of the liver were observed in the Covance study (1998b). Liver weights in
recovery group animals were comparable to those of controls, and histological evidence of liver
enlargement was not observed in the male or female recovery groups. The incidences of non-neoplastic
lesions in the recovery groups were decreased at study termination relative to the high-dose groups, but
in most cases were significantly greater than the control values. These data suggest that DINP-induced
liver toxicity was partially reversed in the recovery groups.

EPA identified a LOAEL value from the Covance study (1998b) of 742 mg/kg-day in males and 910
mg/kg-day in females based on increased incidence of liver masses in males, and increased absolute and
relative liver weights, and decreased absolute and relative kidney weights (Section 3.3). ANOAEL of
276 mg/kg-day in males or 336 mg/kg-day in females was identified based on non-cancer and cancer
effects.

Table Apx B-6. Incidence of Selected Non-neoplastic Lesions in B6C3F1 Mice Exposed to DINP in
the Diet for 2 Years (Covance Labs. 1998b)	



Dose Group
mg/kg-day (ppm)

Lesion

Control

90 M
112 F
(500)

276 M
336 F
(1,500)

742 M
910 F
(4,000)

1,560 M
1,888 F
(8,000)

Recovery''
1,560 M
1,888 F
(8,000)

NhiL-s

Diffuse hepatocellular
enlargement

0/55fl
(0%)

1/50
(2.0%)

1/50
(2.0%)

2/50
(4.0%)

45/55*
(81.8%)

10/50*
(20.0%)

Increased cytoplasmic
eosinophilia

0/55
(0%)

0/50
(0%)

0/50
(0%)

0/50
(0%)

52/55*
(94.5%)

10/50*
(20.0%)

Pigment

0/55
(0%)

0/50
(0%)

0/50
(0%)

0/50
(0%)

49/55*
(89.1%)

6/50*
(12.0%)

Ivnuk-s

Diffuse hepatocellular

enlargement

0,55
(0%)

0.51
(0%)

0,50
(0%)

1.50
(2.0%)

52/55*
(94.5%)

6,50*
(12.0%)

Increased cytoplasmic
eosinophilia

0/55
(0%)

0/51
(0%)

0/50
(0%)

0/50
(0%)

53/55*
(81.8%)

6/50*
(12.0%)

Pigment

1/55
(1.8%)

1/51
(2.0%)

2/50
(4.0%)

2/50
(4.0%)

41/55*
(74.5%)

3/50
(6.0%)

Source: Tables 11A and 11C in Covance Labs (1998b).

M = male; F = female

* = significantly different from control (p < 0.05) by Fisher's Exact test performed by Syracuse Research
Corporation.

" Number of animals with lesion/total number of animals examined; percent incidence of lesion in parentheses.
Incidences are sum of unscheduled deaths and lesions observed at terminal sacrifice.

b The 8,000 ppm recovery group received 8,000 ppm for 78 weeks, followed by a 26-week recovery period during
which the test animals received basal diet alone.

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Waterman et al. (2000) assessed the potential toxicity of DINP in one- and two-generation studies
conducted in SD rats. In the one-generation study, male and female animals were administered 0.5, 1.0,
or 1.5 percent DINP in the diet for 10 weeks prior to mating and lasting throughout the mating period.
The females were subsequently exposed throughout gestation and lactation until PND 21. Mean received
doses in units of mg/kg-day are shown in Table 3-5. Parental body weight gain was significantly
reduced at the 1.0 and 1.5 percent dose groups in both sexes during the premating phase and in females
during gestation and lactation. Absolute liver weights in both sexes were significantly increased at all
doses, except in PI females at the 1.5 percent level.

For the two-generation study, male and female SD rats were fed DINP at dietary concentrations of 0.0,
0.2, 0.4, or 0.8 percent for 10 weeks before mating and for an additional 7 weeks, through mating,
gestation, and lactation continuously for two-generations. Mean received doses in units of mg/kg-day
are shown in Table 3-7. Absolute liver weights of PI males and females were increased over controls at
all DINP treatment levels. Minimal to moderate increases in cytoplasmic eosinophilia were observed in
all males and females from all dose groups of parents in both generations.

TableApx B-7. Summary of Liver Effects Reported in Animal Toxicological Studies Following
Chronic Exposure to DINP 			

Brief Study Description
(Reference)

NOAEL/ LOAEL
(mjj/kjj-day)

Effect at LOAEL

Remarks

F344 rats (both sexes); dietary; 0,
0.03, 0.3, 0.6% (est. 0, 15, 152, 307
mg/kg-day [males]; 0, 18, 184, 375
mg/kg-day [females]); 2 years
(Lington et al, 1997)

15/152
(males)

18/184
(females)

t abs. and rel. liver
weight; t in serum
ALT, AST; | non-
neoplastic lesions
(e.g., focal necrosis,
spongiosis hepatis)



SD rats (both sexes); dietary; 0, 500,
5000, 10,000 ppm (est. 0, 27, 271,
553 mg/kg-day [males]; 0, 33, 331,
672 mg/kg-day [females]); 2 years
(B io/dvnamics, 1987)

27/271 (males)

t serum ALT, AST,
ALP (males); t
spongiosis hepatis; t
hepatic focal necrosis

Other liver effects: t absolute and
relative liver weight (both sexes); t
serum ALP (females); t incidence
of hepatocyte necrosis at low- and
high-doses (males)

GLP-compliant study, non-guideline





Considerations: J. BW sains in
females (672 mg/kg-day); no
change in terminal B W in males; t
food consumption for females at
multiple timepoints during study
(672 mg/kg-day)

Male and female SD rats
(30/sex/dose) fed diets containing 0,
0.5, 1.0, 1.5% DINP (CASRN
68515-48-0) starting 10 weeks prior
to mating, through mating, gestation,
and lactation continuously for one
generation (received doses in units of
mg/kg-day shown in Table 3-5)
(Waterman et al., 2000; Exxon
Biomedical 1996a).

ND/ LOEL = 301

t absolute and
relative liver weight
for PI and P2 males
and females; |
incidence of minimal
to moderate
cytoplasmic
eosinophilia



Male and female SD rats
(30/sex/dose) fed diets containing 0,
0.2, 0.4, 0.8% DINP (CASRN
68515-48-0) starting 10 weeks prior
to mating, through mating, gestation,





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Brief Study Description
(Reference)

NOAEL/ LOAEL
(mg/kg-day)

Effect at LOAEL

Remarks

and lactation continuously for two-
generations. Received doses in units
of mg/kg-day shown in Table 3-7.

(Waterman et aL 2000; Exxon
Biomedical 19965).







B6C3F1 mice (both sexes); dietary;
0, 500, 1500, 4000, 8000 ppm (est. 0,
90, 276, 742, 1,560 mg/kg-day
[males]; 0, 112,336,910, 1,888
mg/kg-day [females]); 2 years

Recovery study; 0, 1,377 [males]; 0,
1,581 [females]); diet; 78 weeks,
followed by 26 weeks recovery.

(Covance Labs. 1998b)

GLP-compliant and adhere to EPA
guidelines (40 CFR Part 798.330)

276/742
(males)

336/910
(females)

t abs. liver weight,
histopathological
changes in the liver
and i body weight
gain) (females); (t
incidence of liver
masses (males)

Significant neoplastic findinss: t
hepatocellular carcinoma; t
incidence of total liver neoplasms
(combined carcinomas and
adenomas)

Considerations:
i mean body weights in males
(>742 mg/kg-day) and females
(>336 mg/kg-day)

F344 rats (both sexes); dietary; 0,
500, 1500, 6000, 12,000 ppm (est. 0,
29, 88, 359, 733 mg/kg-day [males];
0, 36, 109, 442, 885 mg/kg-day
[females]); 2 years

Recovery study: 0, 637 mg/kg-day
[males]; 0, 774 mg/kg-day
[females]); diet; 78-week exposure,
followed by 26 week recovery period

(Covance Labs. 1998c)

GLP-compliant and adhere to EPA
guidelines (40 CFR Part 798.330)

88/359
(males)

109/442
(females)

t abs. and rel. liver
wt.; t in serum ALT
and AST;
histopathological
findings in liver.

Significant neoplastic findinss
t incidence of mononuclear cell
leukemia; t in hepatocellular
carcinoma; t in combined
hepatocellular carcinoma and
adenoma (See (U.S. EPA. 2024a)
for further discussion)

Limitations:

Did not report results for statistical
analyses of lesion incidence data

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Appendix C FETAL TESTICULAR TESTOSTERONE AS AN
ACUTE EFFECT

No studies of experimental animal models are available that investigate the antiandrogenic effects of
DINP following single dose, acute exposures. However, there are studies of dibutyl phthalate (DBP)
available that indicate a single acute exposure during the critical window of development {i.e., GD14-
19) can reduce fetal testicular testosterone production and disrupt testicular steroidogenic gene
expression. Two studies were identified that demonstrate single doses of 500 mg/kg DBP can reduce
fetal testicular testosterone and steroidogenic gene expression. Johnson et al. (2012; ) gavaged
pregnant SD rats with a single dose of 500 mg/kg DBP on GD 19 and observed reductions in
steroidogenic gene expression in the fetal testes three (Cypl7al) to six {Cypllal, StAR) hours post-
exposure, while fetal testicular testosterone was reduced starting 18 hours post-exposure. Similarly,
Thompson et al. (2005) reported a 50 percent reduction in fetal testicular testosterone 1-hour after
pregnant SD rats were gavaged with a single dose of 500 mg/kg DBP on GD 19, while changes in
steroidogenic 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. Additionally, 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 nipple retention,
and increase the frequency of reproductive tract malformations and testicular pathology in adult rats that
received two doses of DBP during the critical window.

In summary, studies of DBP provide evidence to support use of effects on fetal testosterone as an acute
effect. However, the database is limited to just a few studies of DBP that test relatively high (500 mg/kg)
single doses of DBP. Although there are no single dose studies of DINP that evaluate anti androgenic
effects on the developing male reproductive system, there are four studies that have evaluated effects on
fetal testicular testosterone production and steroidogenic gene expression following daily gavage doses
of 500 to 1,500 mg/kg-day DINP on GDs 14 to 18 (5 total doses) (Gray et al.. 2021; Furr et al.. JO I I;
Hannas et al.. 2012; Hannas et al.. 2011)—all of which consistently report anti androgenic effects at the
lowest dose tested (500 mg/kg-day).

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Appendix D SUMMARY OF EPIDEMIOLOGY STUDIES ON
REPRODUCTIVE OUTCOMES

Radke et al. (2018) report the results of a systematic review that evaluated the association between
DINP and male reproductive outcomes. In examining the relationship between DINP exposure and
AGD, the authors found that there is little evidence linking DINP to AGD. The combination of low
exposure levels {i.e., poor sensitivity) and data availability {i.e., fewer accessible studies) may account
for the weaker evidence of an association between AGD and DINP. When evaluating the relationship
between DINP exposure and sperm parameters, the author determined that the association was moderate
due to the morphology's consistency across studies. In examining the association between DINP and the
time until pregnancy in males, the authors did not report a relationship for DINP and the evidence was
deemed inconclusive due to the small number of studies and narrow range of exposure. Finally, when
examining the relationship between DINP metabolite (MINP or MCiOP) exposure and testosterone, the
authors found that there is moderate evidence linking DINP metabolites to lower testosterone levels.

Another systematic review by Radke et al. (2019b) evaluated the association between DINP and female
reproductive and developmental outcomes and also found no clear evidence of association due to
inadequate sensitivity in the available data. When examining the relationship between DINP exposure
and pubertal development the authors found that there was no association linking DINP and pubertal
development and the strength of the evidence was deemed indeterminate. Study evaluations of the
relationship between DINP and a woman's time to pregnancy found that the evidence of an association
between fecundity and exposure to DINP was deemed indeterminate due to lack of the evidence of
relationship for the key fecundity outcomes. The authors also found that in studies that measured the
relationship between DINP and spontaneous abortion, there was no association between early loss and
total loss. Thus, the evidence for an association between DINP and spontaneous abortion was deemed
indeterminate. Finally, when evaluating the association between DINP and gestational duration, the
authors found slight evidence for the association between DINP exposure and preterm birth, however
while there was modest increase in the odds of preterm birth with higher DINP exposure the association
was not statistically significant. In summary there was indeterminate evidence linking DINP and female
reproductive and developmental outcomes.

EPA identified 11 new studies (8 medium quality and 3 low quality) that evaluated the association
between DINP metabolites and developmental and reproductive outcomes. The first medium quality
study, a longitudinal cohort study, by Berger et al. (2018). using data from Center for Health Assessment
of Mothers and Children of Salinas (CHAMACOS) cohort examined prenatal urinary DINP levels and
the association with timing of puberty milestones (thelarche, menarche, pubarche, gonadarche) in
children. The authors found an association between pubarche and menarche age increased in "normal"
weight girls per log2 increase in MCOP. The authors also found gonadarche and pubarche age decreased
in all obese boys. There was not significant a significant association between thelarche age increased in
all girls per log2 increase in MCOP.

A medium quality birth cohort study, by Philipat et al. (2019). Etude des Determinants pre et postnatals
du developpement et de la sante de l'Enfant (EDEN) cohort, evaluated associations between DINP
metabolites (MCOP, MCNP) and a set of outcomes measured at birth (birth weight, placental weight,
placental-to-birth weight ratio). MCNP and MCOP were both associated with lower placental-to-birth
weight ratio; MCNP was additionally associated with lower placental weight. MCOP was associated
with lower placental-to-birth weight ratio (PFR) in multipollutant elastic net penalized regression
models. MCOP was not associated with birth weight or placental weight based on elastic net regression
models.

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A medium quality cross-sectional pilot study, by Zota et al. (2019). included a racially diverse
population of premenopausal women within the Fibroids Observational Research on Genes and the
Environment (FORGE) study presenting to a university gynecology clinic and undergoing either
hysterectomy or myomectomy for symptomatic uterine fibroids to examine the potential associations
between urinary DINP biomarkers and two measures of fibroid burden (uterine volume and fibroid size).
Higher urinary concentrations of MCOP and MCNP were significantly associated with odds of greater
uterine volume. In multivariate logistic regression analyses, each log-unit increase in MCOP was
significantly associated with 2.1 (95% CI: 1.2-3.5) times increased odds of greater uterine volume, and
each log-unit increase in MCNP was associated with 2.8 (95% CI: 1.2-3.5) times increased odds of
greater uterine volume, p < 0.05. Results from additional multivariate linear regression analyses of
urinary phthalate exposure on percent increase in uterine volume were positive but not significant.
Results from multivariate logistic regression analysis of urinary DINP exposure on odds of fibroid size
increase for MCOP were non-significant. Results from additional multivariate linear regression analyses
of urinary MCOP phthalate exposure on percent increase in fibroid size (cm) were also non-significant.

A medium quality cross-sectional study, by Chang et al. (2019). evaluated the association between sex
hormone levels (luteinizing hormone (LH), follicle-stimulating hormone (FSH), sex hormone binding
globulin (SHBG), inhibin B, dehydroepiandrosterone (DHEA), dehydroepiandrosterone sulfate (DHEA-
S), androstenedione (AD), estrone (El), estradiol (E2), total testosterone (TT), free testosterone (FT),
dihydrotestosterone (DHT), dihydrotestosterone/total testosterone ratio, estradiol/total testosterone ratio,
estradiol/estrone ratio), Oxidative stress/Inflammation [(malondialdehyde (MDA), inducible nitric oxide
synthetase (iNOS), 8-hydroxy-2'-deoxyguanosine (8-OHdG)] and benign prostatic hyperplasia (prostate
specific antigen (PSA), prostate volume) and DINP exposure. There were significant positive
associations between the outcomes, FSH, Inhibin B, DHEA, iNOS and MINP with regression
coefficients of 0.91 (95% CI: 0.85, 0.98), 0.90 (95% CI: 0.83, 0.97), 1.58 (95% CI: 1.40, 1.79) and 1.61
(95%) CI: 1.29, 2.03) respectively, p < 0.05. Multivariate regression coefficients showed significant
results for FHS, Inhibin B, iNOS and DHEA, but showed nonsignificant results for LH, SHBG, DHEA-
s, AD, El, E2, TT, FT, DHT, MDA, 8-OHdG, PSA, and prostate volume.

A medium quality study, by Mustieles et al. ( ), used data from a small cohort of subfertile couples
in the Environment and Reproductive Health (EARTH) study to analyze the association between
paternal and maternal preconception urinary DINP metabolites (MCOP), as well as maternal prenatal
DINP metabolites, and measures of placental weight. The authors did not find any significant
association between paternal and maternal preconception urinary phthalates, as well as maternal prenatal
phthalates, and measures of placental weight and MCOP.

A medium quality cohort, by Machtinger et al. (2018). examined the association between urinary
concentrations of DINP with intermediate and clinical in vitro fertilization (IVF) outcomes. There was
an association (adjusted means) between urinary MCOP concentration and intermediate outcomes of
assisted reproduction (total oocytes and mature oocytes) [total oocytes T2 = 10.2 (95% CI: 9.3, 11.2), T2
vs. T1 < 0.05; mature oocytes T2 = 8.4 (95% CI: 7.6, 9.3) T2 vs. T1 < 0.05], However, there was no
significant association (adjusted means) between urinary MCOP concentration and intermediate
outcomes of assisted reproduction (fertilized oocytes, top quality embryos). While there was an
association (adjusted means) between urinary MINP concentration and intermediate outcomes of
assisted reproduction (total oocytes) [total oocytes T2 = 9.2 (95% CI: 8.2, 10.2), T2 vs. T1 < 0.05]; there
was not an association (adjusted means) between urinary MINP concentration and intermediate
outcomes of assisted reproduction (mature oocytes, fertilized oocytes, top quality embryos).

Associations between MOiNP or MONP and intermediate outcomes of assisted reproduction (total
oocytes, mature oocytes, fertilized oocytes, top quality embryos) and live birth following assisted

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reproduction were all non-significant for T2, T3 versus T1 intermediate outcomes and for p-trend of live
birth.

A medium quality case-control study, by Lee et al. (2020). assessed the relationship between uterine
fibroids and DINP metabolite concentrations. The authors did not find any statistically significant
associations between uterine fibroids and DINP metabolite concentrations. The authors did find
associations between cases and controls for OH-MINP concentrations (p-value: 0.042) as mono(4-
methyl-7-hydroxyoctyl) phthalate (OH-MINP) concentrations were significantly higher in the cases than
controls, but it was not statistically significant.

A medium quality occupational short longitudinal study, by Henrotin et al. (2020), observed the three-
day changes in levels of total and free testosterone and oxidized MINP exposure in male factory
workers. A significant inverse association was found between the decrease in serum total testosterone
(TT) concentrations between T1 and T2 and an increase in urinary OXO-MINP. There was no
significant associations observed for total testosterone and models for OH-MINP, or CX-MINP. No
significant associations were noted for free testosterone and oxo-MINP, OH-MINP, or CX-MINP.
Bivariate analyses of sexual health scales (IIEF-5 and ADAM) between DINP exposed and non-exposed
groups: No association was observed between the level of urinary oxo-MINP concentrations and FSH,
LH, index of aromatase activity (ratio of total testosterone to estradiol (TT/E2). No association was
observed between the level of urinary OXO-MINP concentrations and bone turnover biomarkers (P1NP,
CTX).

The first low qualtiy study, a case control study, by Durmaz et al. (2018), examined the association
between DINP metabolites (MINP, MHiNP, MOiNP, MCiOP) and serum luteinizing hormone (LH),
plasma follicle stimulating hormone (FSH) and serum estradiol in non-obese girls aged 4 to 8 years with
premature thelarche. DINP metabolites (MINP, MHiNP, MOiNP, MCiOP and their sum) measured in
spot urine samples were compared among cases and controls. Spearman correlations with uterine
volumes, ovarian volume and pubic hair growth varied but were largely weak, negative and/or not
significant, with some significant positive correlation for the association between MCiOP, MINP and
pubic hair growth, rho = 0.440, p = 0.002 and rho = 0.480, p = 0.000, respectively. Thyroid hormone
levels had largely negative Spearman correlations with DINP metabolites, however MCiOP had a
significant negative correlation with fT4 (rho = -0.335, p = 0.041). Spearman correlations between
DINP metabolites (MCiOP, MiNP, MHiNP, MOiNP, SumDiNP) and BMI and weight were positive and
significant.

A low quality case-control study, by Moreira Fernandez et al. (2019). of women in Brazil evaluated the
association between one DINP metabolite (MINP) and endometriosis. The authors found that there was
a positive but non-significant association for the relationship between MINP and endometriosis (OR=2.5
[95% CI: 0.46, 13.78]).

A final low quality study, a case-control study, by Liao et al. (2018). examined associations between
exposure to one DINP metabolite (MINP) measured in urine samples and recurrent pregnancy loss
among women in Taiwan. The MINP samples was below the limit of detection. The highest sample was
70.4 ng/mL in controls (detection rate 2.6 percent) and 1.43 ng/mL in cases (detection rate 2.9 percent).

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Appendix E BENCHMARK DOSE ANALYSIS OF LINGTON ET AL.
(1997)

E.l Background

OCSPP requested that CPHEA run benchmark dose (BMD) models that are available in EPA's
Benchmark Dose Software version 3.3.2 (BMDS 3.3.2), to estimate risk from DINP for select endpoints
from a chronic exposure study (Lington et al. s , ^ >vnamics. 1986) using specified benchmark
response (BMR) levels. The specific endpoints and BMRs provided by OCSPP for analysis are:

1.	Liver weight relative to bodyweight at terminal sacrifice (males and females)

o BMR: 1 control SD, 5%, 10%, 25%

2.	Serum ALT at 6- and 18-month sacrifices (males only)

o BMR: 1 control SD, 10%, 20%, 100% {i.e., 2x)

3.	Incidence of focal necrosis in the liver (males and females)

o BMR: 5%, 10%

4.	Incidence of spongiosis hepatis in the liver (males only)

o BMR: 5%, 10%

5.	Incidence of sinusoid ectasia in the liver (males only)

o BMR: 5%, 10%

While BMD and BMDL values are provided for all of the BMRs, this report provides detailed model run
outputs for only the models that were run using the standard BMRs generally recommended by EPA for
these endpoints, 10 percent relative deviation from the control mean (10 percent RD) for the
dichotomous endpoints and organ weight change and 1 standard deviation change from the control mean
(1 SD). Detailed modeling results for all standard noncancer models are provided for all six endpoints
using all of the BMRs requested by OCSPP in separately delivered BMDS Excel output files.

E.2 Summary of BMD Modeling Approach

All standard BMDS 3.3.2 dichotomous and continuous models that use maximum likelihood (MLE)
optimization and profile likelihood-based confidence intervals were used in this analysis. Standard
forms of these models (defined below) were run so that auto-generated model selection
recommendations accurately reflect current EPA model selection procedures EPA's benchmark Dose
Technical Guidance (	). BMDS 3.3.2 models that use Bayesian fitting procedures and

Bayesian model averaging were not applied in this work.

Standard BMDS 3.3.2 Models Applied to Continuous Endpoints:

•	Exponential 3-restricted (exp3-r)

•	Exponential 5-restricted (exp5-r)

•	Hill-restricted (hil-r)

•	Polynomial Degree 3-restricted (ply3-r

•	Polynomial Degree 2-restricted (ply2-r)

•	Power-restricted (pow-r)

•	Linear-unrestricted (lin-ur)

Standard BMDS 3.3.2 Models Applied to Dichotomous Endpoints:

•	Gamma-restricted (gam-r)

•	Log-Logistic-restricted (lnl-r)

•	Weibull-restricted (wei-r)

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•	Dichotomous Hill-unrestricted (dhl-ur)

•	Logistic (log)

•	Log-Probit-unrestricted (lnp-ur)

•	Probit (pro)

General Model Options Used for Individual Endpoint Analyses:

•	Risk Type: Extra Risk

•	Preferred Continuous Endpoint BMRs

o Relative Liver Weight: 0.1 (10%)
o Serum ALT: 1 Standard Deviation (1 SD)

•	Preferred Dichotomous Endpoint BMR: 0.1 (10%)

•	Confidence Level: 0.95

•	Background response: Estimated

•	Model Restrictions: Restrictions for BMDS 3.3.2 models are defined in the BMDS 3.3.2 User
Guide and are applied in accordance with EPA BMD Technical Guidance (U.S. EPA. 2012).

Model Selection:

The preferred model for the BMD derivations was chosen from the standard set of dichotomous and
continuous models listed above. The modeling restrictions and the model selection criteria facilitated in
BMDS 3.3.2, and defined in the BMDS User Guide, were applied in accordance with EPA BMD
Technical Guidance (	) for noncancer endpoints.

With respect to the continuous endpoints, responses were first assumed to be normally distributed with
constant variance across dose groups. If no model achieved adequate fit to response means (BMDS Test
4 p > 0.1) and response variances (BMDS Test 2 p > 0.05) under that assumption, models that assume
normal distribution with non-constant variance, variance modeled as a power function of the dose group
mean (	), were considered. If no model achieved adequate fit to response means and

variances (BMDS Test 2 p > 0.05) under that assumption, a BMD/BMDL was not derived, and a
LOAEL was selected as POD for the endpoint.

E.3 Summary of BMD Modeling Results

TableApx E-l. Summary of Benchmark Dose Modeling Results from Selected Endpoints in Male

Section

Endpoint

Sex

Selected
Model"

BMDio
(mg/kg-d)

BMDLio
(mg/kg-d)

E.4

Continuous endpoints

E.4.1.1

Relative Liver weight at terminal sacrifice

Male

Linear, CV

106

85.0

E.4.1.2

Relative Liver weight at terminal sacrifice

Female

LOAEL (184 mg/kg-day)

E.4.2.1

Serum ALT at 6-month sacrifice

Male

Linear, NCV

12.5

8.68

E.4.2.2

Serum ALT at 18-month sacrifice

Male

Power, NCV

37.2

17.4

i: ?

Didiolomous 1 indpoinls

E.5.1.1

Focal necrosis in the liver

Male

Logistic

159

125

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Section

Endpoint

Sex

Selected
Model"

BMDio
(mg/kg-d)

BMDLio
(mg/kg-d)

E.5.1.2

Focal necrosis in the liver

Female

Log-Probit

222

34.3

E.5.2

Spongiosis hepatis in the liver

Male

Log-Probit

31.9

8.57

E.5.3

Sinusoid ectasia in the liver

Male

Log-Probit

125

14.4

" As described in Section 2, BMDs for noncancer endpoints were derived from the standard set of models as defined in
the EPA BMD technical guidance and as identified in BMDS 3.3.2 as defaults. Since the standard approach gave
adequate results for all endpoints, non-standard models were not considered for BMD derivations.

CV = constant variance model; NCV = non-constant variance model

4277 E.4 Continuous Endpoints

4278	E.4.1 Relative Liver Weight - Terminal Sacrifice

4279	E.4.1.1 Male F344 Rats

4280

4281	Table Apx E-2. Dose-Response Modeling Data for Relative Liver Weight at Terminal Sacrifice in

Male F344 Rats Following 2-Year Exposure to D

NP (Lington et al., 1997)

Dose (mg/kg-dav)

Number per Group

Mean

Standard Deviation

0

61

0.032

0.006

15

54

0.034

0.008

152

50

0.038

0.008

307

51

0.042

0.008

4283

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TableApx E-3. Summary of Benchmark Dose Modeling Results for Relative Liver Weight at Terminal Sacrifice in Male F344 Rats

Models "

Restriction''

I5MR = 10%

1' Value

AIC

I5MDS
Recommends

HMDS Recommendation

Notes

I5MR = 5%

I5MR = 1 SI)

I5MR = 25%

15M1)

I5MDL

HMD

I5MDL

I5MI)

I5MDL

I5MI)

I5MDL

Exponential 3

Restricted

116.26

95.59

0.3786

-1497.4
98773

Viable -
Alternate

Modeled control response
std. dev. >1.5 actual
response std. dev.

59.51

48.93

248.94

206.95

272.19

223.80

Exponential 5

Restricted

79.84

36.41

0.3253

-1496.4
71899

Viable -
Alternate



37.70

16.38

218.32

131.93

248.11

147.52

Hill

Restricted

154.16

151.00

NA

-1488.6
14597

Questionable

Residual at control > 2
d.f.=0, saturated model
(Goodness of fit test cannot
be calculated)

85.09

83.34

303.22

296.39

340.22

333.23

Polynomial
Degree 3

Restricted

36.76

10.37

NA

-1495.3
18631

Questionable

BMD/BMDL ratio > 3
d.f.=0, saturated model
(Goodness of fit test cannot
be calculated)

16.01

4.92

272.09

29.48

283.55

31.16

Polynomial
Degree 2

Restricted

88.20

49.76

0.3087

-1496.4
03289

Viable -
Alternate



42.54

23.75

225.74

141.55

254.52

155.99

Power

Restricted

106.22

85.08

0.4626

-1497.8
97726

Viable -
Allemale



53.11

42.54

241.06

195.89

265.55

212.69

1.incur

1 nrc>lrielcd

HNi.44

X4.%

M.4(.2~

I4T.X
25

\ inlilc

Recommended

1 OHC-I \l<

50.59

42.54

241.50

195.75

266.10

211.11

AIC = Akaike information criterion; BMD = benchmark dose; BMDL = benchmark dose lower limit; NA = not applicable.

" Selected Model (bolded and shaded gray); residuals for doses 0, 15, 152, and 307 mg/kg-day were -0.8549, 0.7132, 0.4739, and -0.2682, respectively.
b Restrictions defined in the BMDS 3.3 User Guide.

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4287

Model Summary with BMR of 0.1 Rel. Dev. for the BMD and 0.95
Lower Confidence Limit for the BMDL

0 0455334



























U.U4J J JJt

H 041^334









































¦

0.0375334

















0.0335334

n rm ^^a



































U.Uj 1 j j

0.0295334

















-43	7	57	107	157	207	257	307

MG/KG-DAY

	Frequentist Exponential Degree 3 Estimated Probability	Frequentist Exponential Degree 5 Estimated Probability

Frequentist Hill Estimated Probability		Frequentist Polynomial Degree 3 Estimated Probability

	Frequentist Polynomial Degree 2 Estimated Probability 	Frequentist Power Estimated Probability

	Frequentist Linear Estimated Probability	• Data

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4288

Selected Frequentist Linear Model with BMR of 0.1 Added Risk for
the BMD and 0.95 Lower Confidence Limit for the BMDL

f) 044SQQ7













U.UttJ J J /

0 049SQQ7













U.Uft ujj/















U.U4UJ33 /















U.UjOj /
Ct n2CCQQ7







(

>







n 034^007





























n n^?qqQ7



<

















U.UjZ jjj /

C

n n^n^qQ7

~

















U.UjU J" J /

n mcic;QQ7

















0.0265997















-43	7	57	107	157	207	257	307

MG/KG-DAY

	Estimated Probability 	 Response at BMD • Data 	BMD 	BMDL

4289

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Results for Selected Model - Linear, CV (Unrestricted) - Rel. Dev., BMR = 0.1

User Input

Model Results

Benchmark Ddse

BMD

105.440033

BMDL

34.95359659

BMD'J

139.9032525

AIC

-1497.897925

Test 4 P-value

0.452657772

D.O.F.

2

Model Parameters

# of Parameters

3

Variable

Estimate

£

0.032S14937

beta

3.08295E-05

alpha

5.54312E-05

Goodness of Fit



Dose

Size

Estimated
Median

Cale'd
Median

Observed
Mean

Estimated
5D

Calcd
SD

Observed
5D

Scaled
Residual

0

61

0.032814937

0.032

0.032

0.00744522

O.OD6

0.006

-0.854892965

15

54

0.0332773S

0.034

0.034

0.00744522

O.OOS

O.OOS

0.713230418

152

50

0.037501021

0.038

0.038

0.00744522

0.008

O.OOS

0.47390353

307

51

0.042279594

0.042

0.042

0.00744522

O.OOS

0.008

-0.268185348

Likelihoods of Interest



Model

Log Likelihood*

# of Parameters

AIC

A1

752.7197303

5

-1495.43946

A2

755.9925165

8

-1495.98503

A3

752.7197303

5

-1495.43946

fitted

751.9489626

3

-1497.S9793

R

726.8720033

2

-1449.74401

Tests of Interest



Test

-2*LoEUikelihocd Ratio)

Test ^

p-value

1

58.24102629

6

<0.0001

2

6.545572396

3

0.0878S246

3

6.545572396

3

0.0S788246

4

1.541535303

2

0.46265777

4290

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E.4.1.2 Female F344 Rats

Table Apx E-4. Dose-Response Modeling Data for Relative Liver Weight at
Terminal Sacrifice in Female F344 Rats Following 2-Year Exposure to DINP

Dose
(mg/kg-day)

Number per
Group

Mean

Standard Deviation

0

65

0.031

0.005

18

57

0.032

0.007

184

48

0.036

0.008

375

53

0.04

0.007

4296

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TableApx E-5. Summary of Benchmark Dose Modeling Results for Relative Liver Weight at Terminal Sacrifice in Female F344

StiiiHliml Models "

Restriction''

I5MR = 10%

1' Virtue

AIC

HMDS Recommends

HMDS Recommendation

Notes

HMR = 5%

I5MR = 1 SI)

I5MR = 25%

I5MI)

ISM 1)1 y

15M1)

HMDL

15M1)

HMDL

15M1)

HMDL

Exponential 3

Restricted

143.27

118.57

0.2610

-1596.49

Questionable

Non-constant variance test
failed (Test 3 p-value <
0.05)

Modeled control response
std. dev. >1.5 actual
response std. dev.

73.34

60.66

268.59

219.51

335.42

277.61

Exponential 5

Restricted

86.77

35.03

0.3336

-1596.24

Questionable

Non-constant variance test
failed (Test 3 p-value <
0.05)

39.99

15.51

199.97

114.18

309.91

167.83

Hill

Restricted

135.95

99.63

NA

-1592.96

Questionable

Non-constant variance test
failed (Test 3 p-value <
0.05)

d.f.=0, saturated model
(Goodness of fit test
cannot be calculated)

69.29

48.44

263.02

194.84

338.00

256.96

Polynomial Degree 3

Restricted

72.04

14.45

NA

-1594.31

Questionable

Non-constant variance test
failed (Test 3 p-value <
0.05)

BMD/BMDL ratio > 3
d.f.=0, saturated model
(Goodness of fit test
cannot be calculated)

31.23

6.76

207.53

28.21

350.14

44.06

Polynomial Degree 2

Restricted

91.72

58.72

0.3068

-1596.13

Questionable

Non-constant variance test
failed (Test 3 p-value <
0.05)

44.59

27.86

204.48

123.24

308.82

189.00

Power

Restricted

131.94

106.23

0.3428

-1597.04

Questionable

Non-constant variance test
failed (Test 3 p-value <
0.05)

65.97

53.08

257.01

205.66

329.86

265.74

Linear

Unrestricted

128.47

105.83

0.3429

-1597.04

Questionable

Non-constant variance test
failed (Test 3 p-value <
0.05)

62.63

53.11

256.89

204.62

329.42

264.54

AIC = Akaike information criterion; BMD = benchmark dose; BMDL =benchmark dose lower limit; NA = not applicable.
" No selected model due to inadequate fit of constant or non-constant variance models.
b Restrictions defined in the BMDS 3.3 User Guide.

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4300

Model Summary with BMR of 0.1 Rel. Dev. for the BMD and 0.95
Lower Confidence Limit for the BMDL

n HAA1 7A3

n f)4?1743

















U.Utt l/tJ

n D401743

















u.utui / to
n 03R1743



















U.UJOl / to

n A3C1 "7/ia



















n H3/11 ~7A 3





















0.0321743
D 0301743











































u.uoui / to

n mai 7A3























5 25 75 125 175 225 275 325 375

MG/KG-DAY

Frequentist Exponential Degree 3 Estimated Probability Frequentist Exponential Degree 5 Estimated Probability

Frequentist Hill Estimated Probability 	Frequentist Polynomial Degree 3 Estimated Probability

Frequentist Polynomial Degree 2 Estimated Probability Frequentist Power Estimated Probability
Frequentist Linear Estimated Probability • Data

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4301	E.4.2 Serum ALT - Male F344 Rats	

4302	E.4.2.1 6-Month Sacrifice

4303

4304	Table Apx E-6. Dose-Response Modeling Data for Serum ALT Levels in Male F344

Rats Following 6-Month Exposure to DIN

P (Lineton et aL. 1997)

Dose (mg/kg-dav)

Number per Group

Mean

Standard Deviation

0

10

37

8

15

10

38

7

152

10

81

52

307

10

128

145

4306

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TableApx E-7. Summary of Benchmark Dose Modeling Results for Serum ALT Levels in Male F344 Rats Following 6-Month

Models "

Restriction 6

15MR = 10%

1> Value

AIC

HMDS Recommends

HMDS
Recommendation

Notes

HMR= 1 SI)

HMR = 20%

HMR = 100%

HMD

HMDL

HMD

HMDL

HMD

HMDL

HMD

HMDL

Exponential 3

Restricted

20.05

15.84

0.0692

382.00

Questionable

Goodness of fit p-
value <0.1
Modeled control
response std. dev.
>1.5 actual response
std. dev.

40.15

28.50

38.35

30.29

CF

CF

Exponential 5

Restricted

CF

CF

CF

CF

Unusable

BMD computation
failed

124.58

27.19

CF

CF

CF

CF

Hill

Restricted

19.94

9.12

NA

382.16

Questionable

d.f.=0, saturated
model (Goodness of
fit test cannot be
calculated)

34.15

16.39

CF

CF

123.97

90.11

Polynomial Degree 3

Restricted

40.68

11.16

NA

380.67

Questionable

BMD/BMDL ratio >
3 d.f.=0, saturated
model (Goodness of
fit test cannot be
calculated)

55.33

20.32

56.49

22.31

134.04

98.56

Polynomial Degree 2

Restricted

13.99

0

0.1351

380.89

Unusable

BMD computation
failed; lower limit
includes zero
BMDL not estimated

26.33

14.94

27.79

16.84

132.49

87.19

Power

Restricted

18.76

9.26

0.2143

380.20

Viable - Alternate



32.59

16.63

33.74

18.51

131.87

91.22

1 iiu;ii

1 nre-lricled

12.52

S.(.S

0.3050

3~').03

\ ialile Recommended

1 i m e»l VIC

23.42

15.50

25.04

17.37

125.20

86.83

AIC = Akaike information criterion; BMD = benchmark dose; BMDL = benchmark dose lower limit; NA = not applicable; CF = computation failed
" Selected model (bolded and shaded gray); residuals for doses 0, 15, 152, and 307 were 0.5396, -0.7686, 0.1084, 0.0955, respectively.

1 Restrictions defined in the BMDS 3.3 User Guide.

4309

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Model Summary with BMR of 0.1 Rel. Dev. for the BMD and 0.95
Lower Confidence Limit for the BMDL

1 39 9fiQQR















IjZ.ZDjjO
119 9fiQQR















11Z.ZD330

Q9 9P1QQR9















79 9P1QQR9







J,







CO OCOOQO















32.269982



















-43	7	57	107	157	207	257	307

MG/KG-DAY

— Frequentist Exponential Degree 3 Estimated Probability	Frequentist Hill Estimated Probability

Frequentist Polynomial Degree 3 Estimated Probability 	Frequentist Polynomial Degree 2 Estimated Probability

	 Frequentist Power Estimated Probability		Frequentist Linear Estimated Probability

• Data

4310

Page 153 of 184

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Frequentist Linear Model with BMR of 0.1 Added Risk for the
BMD and 0.95 Lower Confidence Limit for the BMDL

193 opi7f;Q





¦

¦





O

J.Z j.OO/ UI7

103.86769

Q2 QC7CQC





























OJ.OO / UOD















UJ.OU / uou
/I3 QC7CQC t















^O.OD/OOD

















Z j.Ou / uou















J.OO / OOJ j

43

-16.13231



57 107 157 207 257 30

MG/KG-DAY

	Estimated Probability 	 Response at BMD • Data 	BMD 	BMDL

4311

Page 154 of 184

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Results for Selected Model - Linear, NCV (Unrestricted) - Rel. Dev., BMR = 0.1
User Input	

Model



Data



Dependent
Variable

mg/kg-day

Independe



nt



Variable



Total # of



Observatio



n

4

Info



Model

Frequentist Linear,
NCV

Dataset
Name

Male F344 Rats
Serum ALT 6mon

Formula

M[dose] = g + bl
*dose

Var[i] = alpha
*meanfil A rho

Options



Risk



Type

Rel. Dev.

BMR

0.1

Confiden



ce Level

0.95

Distributi



on

Normal

Variance

Non-Constant

Model Results

Benchmark Dose

BMD

12.51986155

BMDL

8.683091255

BMDU

12.77902268

AIC

379.0287425

Test 4 P-value

0.304955816

D.O.F.

2

Model Parameters

# of Parameters

4

Variable

Estimate

8

35.85553524

beta

0.286389228

rho

4.902699939

alpha

1.07545E-06

Goodness of Fit



Dose

Size

Estimated
Median

Calc'd
Median

Observed
Mean

Estimated
SD

Calc'd
SD

Observed
SD

Scaled
Residual

0

10

35.85553524

37

37

6.7074289

8

8

0.539568203

15

10

40.15137365

38

38

8.85168002

7

7

-0.768581876

152

10

79.38669783

81

81

47.0696879

52

52

0.108386302

307

10

123.7770281

128

128

139.825984

145

145

0.095505923

Likelihoods of Interest



Model

Log Likelihood*

# of Parameters

AIC

A1

-228.508524

5

467.017048

A2

-184.1836225

8

384.367245

A3

-184.3267829

6

380.653566

fitted

-185.5143713

4

379.028743

Page 155 of 184

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4313

4314

4315

4316

PUBLIC RELEASE DRAFT
May 2024

E.4.2.2 18-Month Sacrifice

TableApx E-8. Dose-Response Modeling Data for Serum ALT Levels in Male F344

Dose (mg/kg-dav)

Number per Group

Mean

Standard Deviation

0

9

42

10

15

10

39

7

152

10

69

39

307

10

128

126

4317

Page 156 of 184

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4318 TableApx E-9. Summary of Benchmark Dose Modeling Results for Serum ALT Levels in Male F344 Rats Following 18-Month

Exposure to DI>

P (Non-constant Variance) (Line!

ton et al., 1997)

Models "

Restriction 6

BMR = 10%

1> Value

AIC

BMDS Recommends

BMDS
Recommendation

Notes

BMR = 1 SI)

BMR = 20%

BMR = 100%

HMD

BMDL

BMD

BMDL

BMD

BMDL

BMD

BMDL

Exponential 3

Restricted

28.31

19.66

0.0433

371.30

Questionable

Goodness of fit p-value
<0.1

Modeled control
response std. dev. >1.5
actual response std. dev.

56.70

37.76

52.87

37.61

191.28

143.00

Exponential 5

Restricted

103.76

21.91

NA

370.80

Questionable

BMD/BMDL ratio > 3;
d.f.=0, saturated model
(Goodness of fit test
cannot be calculated)

113.99

40.10

113.67

39.87

154.96

134.70

Hill

Restricted

61.57

28.62

NA

371.00

Questionable

d.f.=0, saturated model
(Goodness of fit test
cannot be calculated)

CF

CF

82.15

46.68

182.90

133.66

Polynomial Degree 3

Restricted

63.43

20.61

NA

370.94

Questionable

BMD/BMDL ratio > 3
d.f.=0, saturated model
(Goodness of fit test
cannot be calculated)

85.51

40.83

84.98

40.09

200.71

131.37

Polynomial Degree 2

Restricted

29.49

14.27

0.0428

371.32

Questionable

Goodness of fit p-value
<0.1

56.99

28.32

55.73

28.45

210.39

132.17

1'inwr'

Reminded



r. 4?

0.0')25

3~0.04

(Jiieolhinnlile

( ilMldlK'NN 0|* Ml |)-

\:ilur 0.1

62.51

33.36

59.71

33.45

179.20

134.31

Linear

Unrestricted

20.06

12.52

0.0655

370.67

Questionable

Goodness of fit p-value
<0.1

40.61

24.79

40.11

25.04

200.56

125.22

AIC = Akaike information criterion; BMD = benchmark dose; BMDL = benchmark dose lower limit; NA = not applicable
a Selected Model is bolded and shaded gray; residuals for doses 0, 15, 152, and 307 were 0.7610,-0.6609, -0.2070, and 0.0131, respectively.
b Restrictions defined in the BMDS 3.3 User Guide.

c Despite p < 0.1, the Power model fit would pass at p > 0.05, the variance model passed p > 0.05, and visual fit of model to data is still adequate for BMD calculation.

4320

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Model Summary with BMR of 0.1 Rel. Dev. for the BMD and 0.95
Lower Confidence Limit for the BMDL

-43

130

110

107	157

MG/KG-DAY

307

Frequentist Exponential Degree 3 Estimated Probability
Frequentist Hill Estimated Probability
Frequentist Polynomial Degree 2 Estimated Probability
¦Frequentist Linear Estimated Probability

Frequentist Exponential Degree 5 Estimated Probability

	Frequentist Polynomial Degree 3 Estimated Probability

	Frequentist Power Estimated Probability

O Data

4321

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4322

Frequentist Power Model with BMR of 0.1 Added Risk for the
BMD and 0.95 Lower Confidence Limit for the BMDL

Estimated Probability 	 Response at BMD • Data 	 BMD 	BMDL

Page 159 of 184

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Results for Selected Model - Power, NCV (Restricted) - Rel. Dev., BMR = 0.1
User Input	

Model



Data



Dependent
Variable

mg/kg-day

Independe



nt



Variable



Total # of



Observatio



n

4

Info



Model

Frequentist Power,
NCV

Dataset
Name

MaleF344Rats_S eru
m ALT 18mon

Formula

M[dose] = g + v *
dose A n
Var[i] = alpha *
mean[il A rho

Options



Risk



Type

Rel. Dev.

BMR

0.1

Confiden



ce Level

0.95

Distributi



on

Normal

Variance

Non-Constant

Model Results

Benchmark Dose

BMD

37.19126348

BMDL

17.45080887

BMDU

37.96112263

AIC

370.0444752

Test 4 P-value

0.092488008

D.O.F.

1

Model Parameters

# of Parameters

5

Variable

Estimate

8

39.8382544

V

0.019980069

n

1.464367921

rho

4.643124981

alpha

2.69559E-06

Goodness of Fit



Dose

Size

Estimated
Median

Calc'd
Median

Observed
Mean

Estimated
SD

Calc'd
SD

Observed
SD

Scaled
Residual

0

9

39.8382544

42

42

8.5216504

10

10

0.761030608

15

10

40.89222207

39

39

9.05422294

7

7

-0.66087743

152

10

71.14361683

69

69

32.7473294

39

39

-0.207000441

307

10

127.4742711

128

128

126.82257

126

126

0.013108871

Likelihoods of Interest



Model

Log Likelihood*

# of Parameters

AIC

A1

-217.2980126

5

444.596025

A2

-178.4089743

8

372.817949

A3

-178.6069741

6

369.213948

fitted

-180.0222376

5

370.044475

4324

Page 160 of 184

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4325	E.5 Dichotomous Endpoints

4326	E.5.1 Focal Necrosis in the liver

4327	E.5.1.1 Male F344 Rats

4328

4329	Table Apx E-10. Dose-Response Modeling Data for Focal Necrosis of the Liver in Male

4330	F344 Rats Following 2-Year Exposure to DINP (?	^ A )	

Dose (mg/kg-dav)

Number per Group

Incidence

0

81

10

15

80

9

152

80

16

307

80

26

4331

Page 161 of 184

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4332

4333

PUBLIC RELEASE DRAFT
May 2024

TableApx E-ll. Summary of Benchmark Dose Modeling Results for Focal Necrosis of the Liver in Male F344 Rats Following 2-Year

Models "

Restriction ''

HMR = 10%

P Value

AIC

HMDS Recommends

HMDS Recommendation

Notes

HMR = 5%

HMD

liMDL

HMD

BMDL

Dichotomous Hill

Restricted

154.87

48.90

NA

305.83

Questionable

BMD/BMDL ratio > 3
d.f.=0, saturated model
(Goodness of fit test cannot be
calculated)

132.94

18.97

Gamma

Restricted

161.40

85.98

0.7925

303.85

Viable - Alternate



100.26

41.86

Log-Logistic

Restricted

160.91

78.23

0.7930

303.85

Viable - Alternate



100.39

37.06

Multistage Degree 3

Restricted

162.13

85.74

0.7420

303.89

Viable - Alternate



94.76

41.74

Multistage Degree 2

Restricted

162.13

85.74

0.7420

303.89

Viable - Alternate



94.76

41.74

Multistage Degree 1

Restricted

126.33

84.11

0.8212

302.17

Viable - Alternate



61.50

40.94

Weibull

Restricted

161.48

85.94

0.7832

303.86

Viable - Alternate



98.74

41.84

1 ."ni-lir

1 nreNlriiled

I5S.52

124.50

O.'MI"

30l.')0

\ iiilile Recommended

l.owe-l \l(

88.34

69.47

Log-Probit

Unrestricted

159.84

46.47

0.8230

303.83

Viable - Alternate

BMD/BMDL ratio > 3

104.60

12.63

Probit

Unrestricted

153.31

118.45

0.9368

301.91

Viable - Alternate



83.82

64.96

Quantal Linear

Unrestricted

126.33

84.11

0.8212

302.17

Viable - Alternate



61.50

40.95

AIC = Akaike information criterion; BMD = benchmark dose; BMDL = benchmark dose lower limit; NA = not applicable

"Selected Model is bolded and shaded gray; residuals for doses 0, 15, 152 and 307 were 0.2347, -0.2546, 0.0189 and 0.0007, respectively.

bRestrictions defined in the BMDS 3.3 User Guide.

4334

Page 162 of 184

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Model Summary with BMR of 10% Extra Risk for the BMD and 0.95
Lower Confidence Limit for the BMDL

CD
U
£
CD

¦g

u
£

0.5
0.45
n a





\J. *T

0.35

n 3 i





n 9^:



U.Zj



0.15

0.05

-43

57

107	157

mg/kg-day

207

257

307

Frequentist Dichotomous Hill
Estimated Probability

Frequentist Gamma Estimated
Probability

Frequentist Log-Logistic Estimated
Probability

• Frequentist Multistage Degree 3
Estimated Probability

Frequentist Multistage Degree 2
Estimated Probability

¦	Frequentist Multistage Degree 1
Estimated Probability

¦	Frequentist Weibull Estimated
Probability

¦	Frequentist Logistic Estimated
Probability

¦	Frequentist Log-Probit Estimated
Probability

¦	Frequentist Probit Estimated
Probability

4335

Page 163 of 184

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Frequentist Logistic Model with BMR of 10% Extra Risk for the
BMD and 0.95 Lower Confidence Limit for the BMDL

i

0.9
0.8
0.7
0.6
0.5
0.4
0.3
0.2
0.1
0

^^Estimated Probability

Response at BMD
O Data

	BMD

BMDL

57

107	157

mg/kg-day

207

257

307

Page 164 of 184

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May 2024

Results for Selected Model - Logistic (Unrestricted) - Extra Risk, BMR = 0.1

User Input

Info



Model

Logistic

Dataset Name

Male F344 Rats-

Formula

P[dose] =
l/[l+exp(-a-b*de

Options







Risk Type

Extra Risk

Model Data

BMR

0.1

Dependent Variable

mg/kg-day

Confidence
Level

0.95

Independent Variable

Incidence

Total # of Observation

4

Background

Estimated





Model Results

Benchmark Dose

BMD

158.52

BMDL

124.56

BMDU

239.50

AIC

301.50

P-value

0.94

D.O.F.

2.00

Chi3

0.12

Slope Factor

158.52

Model Parameters

# of Parameters

2

Variable

Estimate

a

-2.0393

b

0.00426

Goodness of Fit



Dose

Estimated
Probability

Expected

Observed

Size

Seated
Residual

0

0.115134137

9.32586507

ID

81

0.2347

15

0.121808403

9.744672223

9

80

-0.2546

152

0.199154436

15.932354SS

16

80

0.0189

307

0.324963347

25.99710772

26

80

0.0007













Analysis of Deviance









Model

Log Likelihood

# of Parameters

Deviance

Test dX

P Value

Full Model

-14S.SS97738

4

-



NA

Fitted Model

-148.950072

2

0.12059642

2

0.9414837

Reduced Model

-156.0920707

1

14.4045939

3

0.0024031

4337

Page 165 of 184

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May 2024

User Input

Info



Model

Logistic

Dataset Name

Male F344 Rats-

Formula

P[dose] =
l/[l+exp(-a-b"dc

Options







Risk Type

Extra Risk

Model Data

BMR

0.1

Dependent Variable

mg/kg-day

Confidence
Level

0.95

Independent Variable

Incidence

Total it of Observation

4

Background

Estimated





Model Results

Benchmark Dose

BMD

158.52

BMDL

124.56

BMDU

239.50

AIC

301:90

P-value

0.94

D.O.F.

2.00

Chi3

0.12

Slope Factor

158.52

Model Parameters

# of Parameters

2

Variable

Estimate

a

-2.0393

b

0.00426

Goodness of Fit



Dose

Estimated
Probability

Expected

Observed

Size

Scaled
Residual

0

0.115134137

9.32586507

10

81

0.2347

15

0.121808403

9.744672223

9

SO

-0.2546

152

0..199154436

15.93235483

16

SO

0.0189

307

0.324963S47

25.99710772

26

80

0.0007













Analysis of Deviance









Model

Log Likelihood

# of Parameters

Deviance

Test cJX.

P Value

Full Model

-14S.S897738

4

-



NA

Fitted Model

-148.950072

2

0.12059642

2

0.9414837

Reduced Model

-156.0920707

1

14.4045939

3

0.0024031

4339

4340

4341

4342

4343

4344

4345

Page 166 of 184

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May 2024

4346	E.5.1.2 Female F344 Rats

4347

4348	Table Apx E-12. Dose-Response Modeling Data for Focal Necrosis of the Liver in

Female F344 Rats Following 2-Year Exposure to

DINP (Liiigtoi! et al, 1997)

Dose (mg/kg-dav)

Number per Group

Incidence

0

81

13

18

81

11

184

80

19

375

80

21

4350

Page 167 of 184

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4351

4352

PUBLIC RELEASE DRAFT
May 2024

TableApx E-13. Summary of Benchmark Dose Modeling Results for Focal Necrosis of the Liver in Female F344 Rats Following 2-

Models "

Restriction b

BMR = 10%

P
Value

AIC

BMDS Recommends

BMDS Recommendation Notes

BMR = 5%

BMD

BMDL

BMD

BMDL

Dichotomous Hill

Restricted

179.57

19.90

NA

323.73

Questionable

BMD/BMDL ratio > 3

d.f.=0, saturated model (Goodness of

fit test cannot be calculated)

148.09

7.87

Gamma

Restricted

247.12

136.68

0.7185

320.19

Viable - Alternate



120.31

66.54

Log-Logistic

Restricted

239.78

125.46

0.7335

320.15

Viable - Alternate



113.58

59.43

Multistage Degree 3

Restricted

247.12

136.68

0.7185

320.19

Viable - Alternate



120.31

66.53

Multistage Degree 2

Restricted

247.12

136.68

0.7185

320.19

Viable - Alternate



120.31

66.54

Multistage Degree 1

Restricted

247.12

136.68

0.7185

320.19

Viable - Alternate



120.31

66.54

Weibull

Restricted

247.12

136.68

0.7185

320.19

Viable - Alternate



120.31

66.54

I .ogistic

Unrestricted

275.16

179.48

0.6509

320.39

Viable - Alternate



148.92

98.02

l.iiSi-Pnihil

I nreslriiled

222.0N

34.30

0.4N0')

322.03

Viable - Recommended

Limes 1 BMDL
BMD/BMDL ralio>3

96.76

0.90

Probit

Unrestricted

271.03

173.31

0.6617

320.36

Viable - Alternate



144.53

93.23

Quantal Linear

Unrestricted

247.12

136.68

0.7185

320.19

Viable - Alternate



120.31

66.54

AIC = Akaike information criterion; BMD = benchmark dose; BMDL = benchmark dose lower limit; NA = not applicable.

" Selected Model is bolded and shaded gray; residuals for doses 0, 18, 184 and 375 were 0.3259, -0.4779, 0.3508 and -0.1977, respectively.
h Restrictions defined in the BMDS 3.3 User Guide.

4353

Page 168 of 184

-------
4354

0.4

0.35

0.3

0.1

0.05

PUBLIC RELEASE DRAFT

May 2024

Model Summary with BMR of 10% Extra Risk for the BMD and 0.95
Lower Confidence Limit for the BMDL

Frequentist Dichotomous Hill
Estimated Probability

Frequentist Gamma Estimated
Probability

Frequentist Log-Logistic Estimated
Probability

¦	Frequentist Multistage Degree 3
Estimated Probability

Frequentist Multistage Degree 2
Estimated Probability

¦	Frequentist Multistage Degree 1
Estimated Probability

¦	Frequentist Weibull Estimated
Probability

¦	Frequentist Logistic Estimated
Probability

¦	Frequentist Log-Probit Estimated
Probability

¦	Frequentist Probit Estimated
Probability

0

-25 25 75 125 175 225 275 325 375

mg/kg-day

4355

Page 169 of 184

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May 2024

4356

Female F344 Relative Liver Weight vs mg/kg-day; LogProbit model
with BMR of 10% Extra Risk for the BMD and 0.95 Lower Confidence

Limit for the BMDL

0.35

0.1
0.05

0

-25	25	75	125	175	225	275	325	375

mg/kg-day

Estimated Probability
Response at BMD
O Data
BMD
BMDL

Page 170 of 184

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May 2024

Results for Selected Model - ^4)gPyqbi^ (Unrestricted) - Extra Risk, BMR = 0.1

User Input

Info



Model

Log-Pro bit

Dataset
Name

Female F344 Rats - focal necrosis

Formula

Pfdose] =

g+(l-E) * KmMDOT[atb*LDSfDose))

Options

Risk Type

Extra Risk

BMR

0.1

Confidence
Level

0.95

Background

Estimated

Model Data



Dependent Variable

mg/kg-day

Independent Variable

Incidence

Total # of Observation

4

Model Results

Benchmark Dose

BMD

222.0606266

BMDL

34.3001408

BMDU

Infinity

AIC

322.0-314517

P-value

0.48 QUI 731

D.O.F.

1

Chi2

0.496782444

Model Parameters

# of Parameters

3

Variable

Estimate

Background (g)

0.147649782

a

-3.644150287

b

0.437272073

Goodness of Fit



Dose

Estimated
Probability

Expected

Observed

Size

Scaled
Residual

0

0.147649782

11.95963234

13

81

0.3258509

18

0.155022564

12.55682771

11

81

-0.477945

184

0.221220007

17.69760055

19

80

0.3508162

375

0.2723415S

21.7873264

21

80

-0.197738













Analysis of Deviance









Model

Log Likelihood

# of Parameters

Deviance

Test .it

P Value

Full Modell

-157.7653174

4

-



NA

Fitted Model

-158.0157259

3

0.50081701

1

0.4791414

Reduced Model

-160.5735074

1

5.61638012

3

0.1318411

4357

4358

Page 171 of 184

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4359

PUBLIC RELEASE DRAFT
May 2024

4360

4361

4362

4363

E.5.2 Spongiosis hepatis in the liver - Male F344 Rats

Table Apx E-14. Dose-Response Modeling Data for Spongiosis Hepatis of the Liver

Dose (mg/kg-dav)

Number per Group

Incidence

0

81

24

15

80

24

152

80

51

307

80

62

4364

Page 172 of 184

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May 2024

4365	TableApx E-15. Summary of Benchmark Dose Modeling Results for Spongiosis Hepatis of the Liver in Male F344 Rats Following 2-

4366	Year Exposure to DINP (Lington et ai. 1997)					

Models"

Restriction*

BMR = 10%

P Value

AIC

BMDS Recommends

BMDS Recommendation Notes

BMR = 5%

BMD

BMDL

BMD

BMDL

Dichotomous Hill

Restricted

53.05

9.92

1

394.27

Viable - Alternate

BMD/BMDL ratio > 3

37.76

4.81

Gamma

Restricted

26.33

20.77

0.8496

390.93

Viable - Alternate



12.82

10.11

Log-Logistic

Restricted

30.45

11.96

0.7322

392.47

Viable - Alternate



17.20

5.67

Mutlistage Degree 3

Restricted

26.33

20.77

1

-9999

Unusable

AIC not estimated

12.82

10.11

Mutlistage Degree 2

Restricted

26.33

20.77

1

-9999

Unusable

AIC not estimated

12.82

10.11

Mutlistage Degree 1

Restricted

26.33

20.77

0.8496

390.93

Viable - Alternate



12.82

10.11

Weibull

Restricted

26.33

20.77

0.8496

390.93

Viable - Alternate



12.82

10.11

Logistic

Unrestricted

42.42

35.87

0.6349

392.50

Viable - Alternate



21.74

18.35

Lo^-Prohil

I nresl riiied

3I.NN

8.57

O.NI37

3')2.37

Viable - Recommended

Limes i BMDL; BMD/BMDL ralio >3

20.08

4.03

Probit

Unrestricted

42.55

36.41

0.6037

392.70





21.70

18.55

Quantal Linear

Unrestricted

26.33

20.77

0.8496

390.93





12.82

10.11

AIC = Akaike information criterion; BMD = benchmark dose; BMDL = benchmark dose lower limit
" Selected Model is bolded; residuals for doses 0,15, 152, and 307 were 0.1279, -0.1656, 0.0941, and -0.0539, respectively.
4 Restrictions defined in the BMDS 3.3 User Guide.

4367

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Model Summary with BMR of 10% Extra Risk for the BMD and
0.95 Lower Confidence Limit for the BMDL

—1—

n q













U.J

n £













U.o
n 7















u. /
n F>













¦

u.o
n c:















n A

















n 5

















U.3
n i

















n 1













U. -L

0













-43	7	57	107	157	207	257	307

MG/KG-DAY

	 Frequentist Dichotomous Hill Estimated Probability 	 Frequentist Gamma Estimated Probability

Frequentist Log-Logistic Estimated Probability	Frequentist Multistage Degree 3 Estimated Probability

	 Frequentist Multistage Degree 2 Estimated Probability	Frequentist Multistage Degree 1 Estimated Probability

	Frequentist Weibull Estimated Probability		Frequentist Logistic Estimated Probability

Frequentist Log-Probit Estimated Probability	Frequentist Probit Estimated Probability
	Frequentist Quantal Linear Estimated Probability	• Data

4368

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4369

Frequentist Log-Probit Model with BMR of 10% Extra Risk for the
BMD and 0.95 Lower Confidence Limit for the BMDL

1

n q

0.8
0.7



0.6

rt c



n A -r- T ^0^0000"^

0.3 i—
n ? -L

r





n 1





U. -L

0





-43	7	57	107	157	207	257	307

MG/KG-DAY

	 Estimated Probability 	 Response at BMD • Data 	BMD 	BMDL

4370

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Results for Selected Model - LogProbit (Unrestricted) - Extra Risk, BMR = 0.1









Model



Info



Options



Data



Model

Log-Probit



Risk





Dependent



Dataset
Name

Male F344



Type

Extra Risk



Variable

mg/kg-day

Rats spongiosis



BMR

0.1



Independe



hepatis



Confiden





nt





P[dose] = g+(l-g) *



ce Level

0.95



Variable

Incidence

Formula

CumNorm(a+b*Log(
Dose))



Backgrou
nd

Estimated



Total # of
Observatio











n

4

Vlodel Results

Benchmark Dose

BMD

31.87966632

BMDL

8.566931336

BMDU

77.63938389

AIC

392.3657526

P-value

0.813651618

D.O.F.

1

Chi2

0.055562904

Model Parameters

# of Parameters

3

Variable

Estimate

Background (g)

0.288658724

a

-4.003497521

b

0.786242291

Goodness of Fit



Dose

Estimated
Probability

Expected

Observed

Size

Scaled
Residual

0

0.288658724

23.38135661

24

81

0.1279398

15

0.310314502

24.82516015

24

80

0.1656122

152

0.629151263

50.33210107

51

80

0.094143

307

0.780322211

62.4257769

62

80

-0.053889

Analysis of Deviance





Log

#of



Test



Model

Likelihood

Parameters

Deviance

d.f.

P Value

Full Model

-193.1328632

4

-

-

NA

Fitted Model

-193.1828763

3

0.10002618

1

0.7517982

Reduced Model

-222.4986873

1

58.6316221

3

0.7517982

4371

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4372	E.5.3 Sinusoid Ectasia in the Liver Male F344 Rats

4373

4374	Table Apx E-16. Dose-Response Modeling Data for Sinusoid Ectasia of the Liver

4375	in Male F344 Rats Following 2-Year Exposure to DINP (?	^ ^ A )

Dose (mg/kg-dav)

Number per Group

Incidence

0

81

16

15

80

16

152

80

24

307

80

33

4376

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TableApx E-17. Summary of Benchmark Dose Modeling Results for Sinusoid Ectasia of the Liver in Male F344 Rats Following 2-

Models "

Restriction b

BMR = 10%

P
Value

AIC

BMDS Recommends

BMDS Recommendation

Notes

BMR=5%

BMD

BMDL

BMD

BMDL

Dichotomous Hill

Restricted

126.62

19.59

NA

374.75

Questionable

BMD/BMDL ratio > 3
d.f.=0, saturated model
(Goodness of fit test cannot
be calculated)

79.29

7.58

Gamma

Restricted

121.73

68.52

0.9441

372.76

Viable - Alternate



66.95

33.36

Log-Logistic

Restricted

122.39

58.96

0.9572

372.75

Viable - Alternate



69.06

27.93

Multistage Degree 3

Restricted

118.39

68.47

0.9930

370.77

Viable - Alternate



60.57

33.33

Multistage Degree 2

Restricted

118.39

68.47

0.9930

370.77

Viable - Alternate



60.57

33.33

Multistage Degree 1

Restricted

104.19

68.30

0.9746

370.80

Viable - Alternate



50.72

33.25

Weibull

Restricted

121.20

68.51

0.9372

372.76

Viable - Alternate



65.82

33.35

Logistic

Unrestricted

128.86

97.30

0.9836

370.78

Viable - Alternate



68.24

51.73

l.iiSi-Pnihil

I nreslriiled

125.23

14.42

O.'WI 1

372.75

Viable - Recommended

Limesi BMDI.
BMD/BMDI. ralio >3

76.52

2.40

Probit

Unrestricted

125.62

93.71

0.9883

370.77

Viable - Alternate



65.79

49.29

Quantal Linear

Unrestricted

104.19

68.30

0.9746

370.80

Viable - Alternate



50.72

33.25

AIC = Akaike information criterion; BMD = benchmark dose; BMDL =benchmark dose lower limit; NA = not applicable.
" Selected Model is bolded; residuals for doses 0,15, 152 and 307 were -0.0075, 0.0082, -0.0013 and 0.0007, respectively.
4 Restrictions defined in the BMDS 3.3 User Guide.

4379

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Model Summary with BMR of 10% Extra Risk for the BMD and 0.95
Lower Confidence Limit for the BMDL

0.5

0

-43	7	57	107	157	207	257	307

mg/kg-day

^^—Frequentist Dichotomous Hill
Estimated Probability

Frequentist Gamma Estimated
Probability

Frequentist Log-Logistic Estimated
Probability

^^—Frequentist Multistage Degree 3
Estimated Probability

^^—Frequentist Multistage Degree 2
Estimated Probability

^^—Frequentist Multistage Degree 1
Estimated Probability

^^—Frequentist Weibull Estimated
Probability

^^—Frequentist Logistic Estimated
Probability

^^—Frequentist Log-Probit Estimated
Probability

^^—Frequentist Probit Estimated
Probability

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Male F344 Relative Liver Weight vs mg/kg-day; LogProbit model
with BMR of 10% Extra Risk for the BMD and 0.95 Lower Confidence

Limit for the BMDL

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Results for Selected Model - LogProbit (Unrestricted) - Extra Risk, BMR = 0.1

wwWvavwwvvvw-

User Input







Options









Info

Risk Type

Extra Risk



Model Data

Model

Log-Probit

BMR

0.1

Dependent Variable

rng/kg-day

Dataset Name

Sinusoid Ectasia -

Confidence
Level

0.95

Independent Variable

Incidence

Formula

Pjdose] = g+(l-g)

Total # of Observation

4





Background

Estimated





Model Results

Benchmark Dose

BMD

125.23

BMDL

14.42

BMDU

247.62

AIC

372.75

P-value

0.99

D.O.F.

1.00

Chi2

0.00

Model Parameters

# of Parameters



Variable

Estimate

g

0.197861854

a

-4.343490179

b

0.73743943

Goodness of Fit



Dose

Estimated
Probability

Expected

Observed

Size

Scaled
Residual

0

0.197861854

16.02681018

16

SI

-0.0075

15

0.199634872

15.97073978

16

SO

0.00S2

152

0.300063561

24.00543434

24

SO

-0.0013

307

0.412461541

32.99692324

33

SO

0.0007

Analysis of Deviance



Model

Log Likelihood

of Parameters

Deviance

Test slX.

P Value

Full Model

Full Model

-133.3755714

4



-

Fitted Model

Fitted Model

-183.3756339

3

0.00012493

1

Reduced Model

Reduced Model

-139.500S934

1

12.2506439

3

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4398

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4404

4405

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4409

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Appendix F CALCULATING DAILY ORAL HUMAN

EQUIVALENT DOSES AND HUMAN EQUIVALENT
CONCENTRATIONS

For DINP, all data considered for PODs are obtained from oral animal toxicity studies in rats, mice, or
beagles. Because toxicity values for DINP are from oral animal studies, EPA must use an extrapolation
method to estimate HEDs. The preferred method would be to use chemical-specific information for such
an extrapolation. EPA identified one study reporting a physiologically based pharmacokinetic model for
DINP based on humanized liver mice (Miura et ai. 2018). Since the study made use of genetically
modified animals and has not been validated by the Agency, it was not considered fit-for-purpose or
used to calculate HEDs. EPA did not locate other DINP information to conduct a chemical-specific
quantitative extrapolation. In the absence of such data, EPA relied on the guidance from U.S. EPA
(2 ), which recommends scaling allometrically across species using the three-quarter power of body
weight (BW3/4) for oral data. Allometric scaling accounts for differences in physiological and
biochemical processes, mostly related to kinetics.

For application of allometric scaling in risk evaluations, EPA uses dosimetric adjustment factors
(DAFs), which can be calculated using EquationApx F-l.

EquationApx F-l. Dosimetric Adjustment Factor

U.S. EPA (2 ), presents DAFs for extrapolation to humans from several species. However, because
those DAFs used a human body weight of 70 kg, EPA has updated the DAFs using a human body
weight of 80 kg for the DINP risk evaluation (	|). EPA used the body weights of 0.025,

0.25, and 12 kg for mice, rats and dogs, respectively, as presented in U.S. EPA (2 ). The resulting
DAFs for mice, rats, and dogs are 0.133, 0.236, and 0.622, respectively.

Use of allometric scaling for oral animal toxicity data to account for differences among species allows
EPA to decrease the default intraspecies UF (UFa) used to set the benchmark MOE; the default value of
10 can be decreased to 3, which accounts for any toxicodynamic differences that are not covered by use
of BW3 4 Using the appropriate DAF from Equation Apx F-l, EPA adjusts the POD to obtain the HED
using Equation Apx F-2:

Equation Apx F-2. Daily Oral Human Equivalent Dose

Where:

DAF

BWa
BWh

Dosimetric adjustment factor (unitless)

Body weight of species used in toxicity study (kg)

Body weight of adult human (kg)

HEDDaily — PODDauy X DAF

Where:

HEDnaily
P ODDaily

DAF

Human equivalent dose assuming daily doses (mg/kg-day)
Oral POD assuming daily doses (mg/kg-day)

Dosimetric adjustment factor (unitless)

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4441

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4444

4445

4446

4447

4448

4449

4450

4451

4452

4453

4454

4455

4456

4457

4458

4459

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For this draft risk evaluation, EPA assumes similar absorption for the oral and inhalation routes, and no
adjustment was made when extrapolating to the inhalation route. For the inhalation route, EPA
extrapolated the daily oral HEDs to inhalation HECs using a human body weight and breathing rate
relevant to a continuous exposure of an individual at rest, as follows:

EquationApx F-3. Extrapolating from Oral HED to Inhalation HEC

wrrn r BWh

HECDaily, continuous ~ H E D Daily X (	)

in^ * EiU£

Where:

HECoaily continuous	=	Inhalation HEC based on continuous daily exposure (mg/m3)

HEDoaiiy	=	Oral HED based on daily exposure (mg/kg-day)

BWh	=	Body weight of adult humans (kg) = 80

IRr	=	Inhalation rate for an individual at rest (m3/hr) = 0.6125

EDc	=	Exposure duration for a continuous exposure (hr/day) = 24

Based on information from U.S. EPA (201 la). EPA assumes an at rest breathing rate of 0.6125 nrVhr.
Adjustments for different breathing rates required for individual exposure scenarios are made in the
exposure calculations, as needed.

It is often necessary to convert between ppm and mg/m3 due to variation in concentration reporting in
studies and the default units for different OPPT models. Therefore, EPA presents all PODs in
equivalents of both units to avoid confusion and errors. Equation Apx F-4 presents the conversion of the
HEC from mg/m3 to ppm.

Equation Apx F-4. Converting Units for HECs (mg/m3 to ppm)

mg 24.45
X ppm = Y —5- x

m3 MW

Where:

24.45 = Molar volume of a gas at standard temperature and pressure (L/mol), default
MW = Molecular weight of the chemical (MW of DINP = 418.62 g/mol)

F.l DINP Non-cancer HED and HEC Calculations for Acute and
Intermediate Duration Exposures

The acute and intermediate duration non-cancer POD is based on a BMDLs of 49 mg/kg-day, and the
critical effect is decreased fetal testicular testosterone. The BMDLs was derived by NASEM (2017)
through meta-regression and BMD modeling of fetal testicular testosterone data from two studies of
DINP with rats (Bobere et al. JO I I; liannas et al. 2011). R code supporting NASEM's meta-regression
and BMD analysis of DINP is publicly available through GitHub). This non-cancer POD is considered
protective of effects observed following acute and intermediate duration exposures to DINP. EPA used
Equation Apx F-l to determine a DAF specific to rats (0.236), which was in turn used in the following
calculation of the daily HED using Equation Apx F-2:

mq mq
11.6 		— = 49-	— X 0.236

kg — day kg — day

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EPA then calculated the continuous HEC for an individual at rest using EquationApx F-3:

mq mq 80 kq
63-0 —j = 11.6-	x(	^	)

m	kg day 0.6125 * 24 hr

hr

Equation Apx F-4 was used to convert the HEC from mg/m3 to ppm:

mq 24.45

3.68 ppm = 63.0 —- x	

HH	m3 418.62

F.2 DINP Non-cancer HED and HEC Calculations for Chronic Exposures

The chronic duration non-cancer POD is based on a NOAEL of 15 mg/kg-day, and the critical effect is
liver toxicity (i.e., increased relative liver weight, increased serum chemistry (AST, ALT, ALP),
histopathologic findings (e.g., focal necrosis, spongiosis hepatis)) in F344 rats following two years of
dietary exposure to DINP (Lington etai. 1997; Bio/dynamics. 1986). EPA used Equation Apx F-l to
determine a DAF specific to rats (0.236), which was in turn used in the following calculation of the daily
HED using Equation Apx F-2:

mq mq
3.55 		— = 15-	— x 0.236

kg — day kg — day

EPA then calculated the continuous HEC for an individual at rest using Equation Apx F-3:

mq mq 80 kq
19.3-|= 3.55 		x(	^	)

m	kg day 0.6125 * 24 hr

hr

Equation Apx F-4 was used to convert the HEC from mg/m3 to ppm:

mg 24.45
1.13 ppm = 19.3 —7 x ^ ^ ^
m3 418.62

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