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EPA Document# EPA-740-D-24-027
December 2024
United States Office of Chemical Safety and
Environmental Protection Agency Pollution Prevention
Draft Non-cancer Human Health Hazard Assessment for
Diisobutyl Phthalate (DIBP)
Technical Support Document for the Draft Risk Evaluation
CASRN: 84-69-5
December 2024
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28 TABLE OF CONTENTS
29 KEY ABBREVIATIONS AND ACRONYMS 5
30 SUMMARY 6
31 1 INTRODUCTION 9
32 1.1 Human Epidemiologic Data: Approach and Preliminary Conclusions 9
33 1.2 Laboratory Animal Data: Summary of Existing Assessments, Approach, and Methodology .... 11
34 1.2.1 Summary of Existing Assessments 11
35 1.2.2 Approach to Identifying and Integrating Laboratory Animal Data 14
36 1.2.3 New Literature Identified and Hazards of Focus for DIBP 16
37 2 TOXICOKINETICS 17
38 2.1 Oral Route 17
39 2.2 Inhalation Route 17
40 2.3 Dermal Route 17
41 3 NON-CANCER HAZARD IDENTIFICATION 19
42 3.1 Effects on the Developing Male Reproductive System 19
43 3.1.1 Summary of Available Epidemiological Studies 19
44 3.1.1.1 Previous Epidemiology Assessment (Conducted in 2019 or Earlier) 19
45 3.1.1.1.1 Health Canada (2018b) 20
46 3.1.1.1.2 Radkeetal. (2019b; 2018) 21
47 3.1.1.1.3 NASEM (2017) 22
48 3.1.1.1.4 Summary of the Existing Assessments of Male Reproductive Effects 23
49 3.1.1.2 EPA Summary of New Studies (2018 to 2019) 23
50 3.1.2 Summary of Laboratory Animals Studies 26
51 3.1.2.1 Developing Male Reproductive System 32
52 3.1.2.2 Other Developmental Outcomes 33
53 4 DOSE-RESPONSE ASSESSMENT 35
54 4.1 Selection of Studies and Endpoints for Non-cancer Health Effects 36
55 4.2 Non-cancer Oral Points of Departure for Acute, Intermediate, and Chronic Exposures 36
56 4.2.1 Studies Considered for the Non-Cancer POD 36
57 4.2.2 Options Considered by EPA for Deriving the Acute Non-Cancer POD 41
58 4.2.2.1 Option 1. NOAEL/LOAEL Approach 41
59 4.2.2.2 Option 2. Application of a Data-Derived Adjustment Factor 41
60 4.2.2.3 Option 3. BMD Analysis of Individual Fetal Testicular Testosterone Studies 42
61 4.2.3 POD Selected for Acute, Intermediate, and Chronic Durations 42
62 4.3 Weight of The Scientific Evidence Conclusion: POD for Acute, Intermediate, and Chronic
63 Durations 46
64 5 CONSIDERATION OF PESS AND AGGEGRATE EXPOSURE 48
65 5.1 Hazard Considerations for Aggregate Exposure 48
66 5.2 PESS Based on Greater Susceptibility 48
67 6 POINTS OF DEPARTURE USED TO ESTIMATE RISKS FROM DIBP EXPOSURE,
68 CONCLUSOINS, AND NEXT STEPS 56
69 REFERENCES 57
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70 APPENDICES 67
71 Appendix A Existing Assessments of DIBP 67
72 Appendix B Fetal Testicular Testosterone as an Acute Effect 71
73 Appendix C Calculating Daily Oral Human Equivalent Doses and Human Equivalent
74 Concentrations 72
75 C. 1 DIBP Non-cancer HED and HEC Calculations for Acute, Intermediate, and Chronic
76 Duration Exposures 73
77 Appendix D Considerations for Benchmark Response (BMR) Selection for Reduced Fetal
78 Testicular Testosterone 75
79 D.l Purpose 75
80 D.2 Methods 75
81 D.3 Results 76
82 D.4 Weight of Scientific Evidence Conclusion 77
83 Appendix E BENCHMARK DOSE MODELING OF FETAL TESTICULAR
84 TESTOSTERONE 80
85 E.l BMD Model Results (Gray et al. 2021) 81
86 E.2 BMD Model Results (Howdeshell et al. 2008) 84
87 E.3 BMD Model Results (Hannas et al. 2011) 88
88
89 LIST OF TABLES
90 Table 1-1. Summary of DIBP Non-cancer PODs Selected for Use by other Regulatory Organizations. 12
91 Table 3-1. Summary of Scope and Methods Used in Previous Assessments to Evaluate the Association
92 Between DIBP Exposure and Male Reproductive Outcomes 19
93 Table 3-2. Summary of Epidemiologic Evidence of Male Reproductive Effects Associated with
94 Exposure to DIBP 21
95 Table 3-3. Summary of Studies of DIBP Evaluating Developmental and Reproductive Outcomes 27
96 Table 4-1. Summary of NASEM (2017) Meta-Analysis and BMD Modeling for Effects of DIBP in Fetal
97 Testosterone ab 38
98 Table 4-2. Overall Analyses of Rat Studies of DIBP and Fetal Testosterone (Updated Analysis
99 Conducted by EPA) 39
100 Table 4-3. Benchmark Dose Estimates for DIBP and Fetal Testosterone in Rats 39
101 Table 4-4. Summary of Dichotomous BMD Analysis of Data from Saillenfait et al. (2008) by Blessinger
102 et al. (2020)° 40
103 Table 4-5. Dose-Response Analysis of Selected Studies Considered for Acute, Intermediate, and
104 Chronic Exposure Scenarios 44
105 Table 5-1. PESS Evidence Crosswalk for Biological Susceptibility Considerations 49
106
107 LIST OF FIGURES
108 Figure 1-1. Overview of DIBP Human Health Hazard Assessment Approach 15
109 Figure 3-1. Hypothesized Phthalate Syndrome Mode of Action Following Gestational Exposure 32
110
111 LIST OF APPENDIX TABLES
112 Table Apx A-l. Summary of Peer-review, Public Comments, and Systematic Review for Existing
113 Assessments of DIBP 67
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TableApx D-l. Comparison of BMD/BMDL Values Across BMRs of 5%, 10%, and 40% with PODs
and LOAELs for Apical Outcomes for DEHP, DBP, DIBP, BBP, DCHP, and DINP .... 79
Table Apx E-l. Summary of BMD Model Results for Decreased Ex Vivo Fetal Testicular Testosterone
81
Table Apx E-2. Ex Vivo Fetal Rat Testicular Testosterone Data (Gray et al. 2021) 81
Table Apx E-3. BMD Model Results Ex Vivo Fetal Testicular Testosterone (Gray et al. 2021) 82
Table Apx E-4. Ex Vivo Fetal Rat Testicular Testosterone Data (Howdeshell et al. 2008) 84
Table Apx E-5. BMD Model Results Ex Vivo Fetal Testicular Testosterone (Howdeshell et al. 2008). 86
Table Apx E-6. Ex Vivo Fetal Rat Testicular Testosterone Data (Hannas et al. 2011) 88
Table Apx E-7. BMD Model Results Ex Vivo Fetal Testicular Testosterone (All Dose Groups) (Hannas
etal. 2011) 90
Table Apx E-8. BMD Model Results Ex Vivo Fetal Testicular Testosterone (Highest Dose Group
Removed) (Hannas et al. 2011) 91
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KEY ABBREVIATIONS AND ACRONYMS
ADME
Absorption, distribution, metabolism, and excretion
BMD
Benchmark dose
BMDL
Benchmark dose lower bound
BMR
Benchmark response
CASRN
Chemical abstracts service registry number
CPSC
Consumer Product Safety Commission (U.S.)
BBP
Butyl -b enzy 1 -phthal ate
DBP
Dibutyl phthal ate
DEHP
Di-ethylhexyl phthalate
DIBP
Diisobutyl phthalate
DINP
Di-isononyl phthalate
ECHA
European Chemicals Agency
EPA
Environmental Protection Agency (U.S.)
GD
Gestational day
HEC
Human equivalent concentration
HED
Human equivalent dose
LOAEL
Lowest-observable-adverse-effect level
LOEL
Lowest-observable-effect level
MOA
Mode of action
MOE
Margin of exposure
NICNAS
National Industrial Chemicals Notification and Assessment Scheme
NOAEL
No-observed-adverse-effect level
OECD
Organisation for Economic Co-operation and Development
OPPT
Office of Pollution Prevention and Toxics
PBPK
Physiologically based pharmacokinetic
PECO
Population, exposure, comparator, and outcome
PESS
Potentially exposed or susceptible subpopulations
PND
Postnatal day
POD
Point of departure
RPF
Relative Potency Factor
SACC
Science Advisory Committee on Chemicals
SD
Sprague-Dawley
TSCA
Toxic Substances Control Act
UF
Uncertainty factor
U.S.
United States
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ACKNOWLEGEMENTS
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, review, and input from EPA OPPT
and OSCPP senior managers and science advisors.The Agency is also grateful for assistance from the
following EPA contractors for the preparation of this draft technical support document: ICF (Contract
No. 68HERC23D0007); and SRC, Inc. (Contract No. 68HERH19D0022). Special acknowledgement is
given for the contributions of technical experts from EPA's Office of Research and Development (ORD)
including Justin Conley, Earl Gray, and Tammy Stoker.
As part of an intra-agency review, this technical support document was provided to multiple EPA
Program Offices for review. Comments were submitted by EPA's Office of General Counsel (OGC) and
ORD.
Docket
Supporting information can be found in the public docket, Docket ID EPA-HQ-QPPT-2018-0434.
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: Collin Beachum (Management Lead), Brandall Ingle-Carlson (Assessment Lead), Myles
Hodge (Human Health Hazard Assessment Lead), Anthony Luz (Human Health Hazard Discipline
Lead), Christelene Horton (Human Health Hazard Assessor)
Contributors: Azah Abdallah Mohamed, Devin Alewel, John Allran, Rony Arauz Melendez, Sarah Au,
Lillie Barnett, Jone Corrales, Maggie Clark, Daniel DePasquale, Lauren Gates, Amanda Gerke, Annie
Jacob, Ryan Klein, Sydney Nguyen, Ashley Peppriell, Brianne Raccor, Maxwell Sail, Joe Valdez, Leora
Vegosen, Susanna Wegner
Technical Support: Kelley Stanfield, Hillary Hollinger, S. Xiah Kragie
This draft technical support document was reviewed and cleared for release by OPPT and OCSPP
leadership.
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SUMMARY
This technical support document is in support of the TSCA Draft Risk Evaluation for Diisobutyl
Phthalate (DIBP) (U.S. EPA. 2024i). This document describes the use of available information to
identify the non-cancer hazards associated with exposure to DIBP and the points of departure (PODs) to
be used to estimate risks from DIBP exposures in the draft risk evaluation of DIBP. Environmental
Protection Agency (EPA, or the Agency) summarizes the cancer and genotoxicity hazards associated
with exposure to DIBP in the Draft Cancer Raman Health Hazard Assessment for Di(2-ethylhexyl)
Phthalate (DEHP), Dibutyl Phthalate (DBP), Diisobutyl Phthalate (DIBP), Butyl Benzyl Phthalate
(BBP) and Dicyclohexyl Phthalate (DCHP) (U.S. EPA. 2025a).
EPA identified effects on the developing male reproductive system as the most sensitive and robust non-
cancer hazard associated with oral exposure to DIBP in experimental animal models (Section 3.1).
Existing assessments of DIBP also identified effects on the developing male reproductive system as the
most sensitive and robust non-cancer effect following oral exposure to DIBP. Existing assessments
included those by the U.S. Consumer Product Safety Commission (U.S. CPSC. 2014. 2011). Health
Canada (ECCC/HC. 2020; EC/HC. 2015b). European Chemicals Agency (2017a. b), and the Australian
National Industrial Chemicals Notification and Assessment Scheme (NICNAS. 2008a). as well as a
systematic review by Yost et al., (2019). which drew conclusions consistent with those of the
aforementioned regulatory bodies. EPA also considered epidemiologic evidence qualitatively as part of
hazard identification and characterization. However, epidemiologic evidence for DIBP was not
considered further for dose response analysis due to limitations and uncertainties in exposure
characterization (discussed further in Section 1.1). Use of epidemiologic evidence qualitatively is
consistent with phthalates assessment by Health Canada, U.S. CPSC, NICNAS, and ECHA.
As discussed further in Section 3.1.2, EPA identified 13 oral exposure studies (11 of rats, 2 of mice) that
have investigated the developmental and reproductive effects of DIBP following gestational and/or
perinatal exposure to DIBP (Gray et al.. 2021; Pan et al.. 2017; Saillenfait et al.. 2017; Wang et al..
2017; Sedha et al.. 2015; Furr et al.. 2014; Hannas et al.. 2012; Hannas et al.. 2011; Howdeshell et al..
2008; Saillenfait et al.. 2008; BASF. 2007; Borch et al.. 2006; Saillenfait et al.. 2006). No one- or two-
generation reproduction studies of DIBP are available for any route of exposure. Across available
studies, the most sensitive developmental effects identified by EPA include effects on the developing
male reproductive system consistent with a disruption of androgen action and the development of
phthalate syndrome. EPA is proposing a POD of 24 mg/kg-day (human equivalent dose [HED] of 5.7
mg/kg-day) based on phthalate syndrome-related effects on the developing male reproductive system
(i.e., decreased fetal testicular testosterone) to estimate non-cancer risks from oral exposure to DIBP for
acute, intermediate, and chronic durations of exposure in the draft risk evaluation of DIBP. The
proposed POD was derived from benchmark dose (BMD) modeling of ex vivo fetal testicular
testosterone data and supports a 95 percent lower confidence limit on the BMD associated with a
benchmark response (BMR) of 5 percent (BMDLs) of 24 mg/kg-day (Gray et al.. 2021).
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 3x and a 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 4.3, EPA reviewed the weight of the scientific evidence
and has robust overall confidence in the proposed POD based on decreased fetal testicular
testosterone for use in characterizing risk from exposure to DIBP for acute, intermediate, and
chronic exposure scenarios. The applicability and relevance of this POD for all exposure durations
(acute, intermediate, and chronic) is described in the introduction to Section 4 and additionally in
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Section 4.2 and Appendix B. For purposes of assessing non-cancer risks, the selected POD is considered
most applicable to women of reproductive age, pregnant women, and male infants. Use of this POD to
assess risk for other age groups (e.g., older children, adult males, and the elderly) is considered to be
conservative and appropriate for a screening level assessment for these other age groups.
No data are available for the dermal or inhalation routes that are suitable for deriving route-specific
PODs. Therefore, EPA is using the acute/intermediate/chronic oral PODs to evaluate risks from dermal
and inhalation exposure to DIBP. For the dermal route, differences in absorption are being accounted for
in dermal exposure estimates in the draft risk evaluation for DIBP. For the inhalation route, EPA is
extrapolating the oral HED to an inhalation human equivalent concentration (HEC) per EPA's Methods
for derivation of inhalation reference concentrations and application of inhalation dosimetry (U.S.
EPA. 1994) using the updated human body weight and breathing rate relevant to continuous exposure of
an individual at rest provided in EPA's Exposure factors handbook: 2011 edition (U.S. EPA. 201 lb).
Table ES-1 and Section 6 summarize EPA's selection of the oral HED and inhalation HEC values used
to estimate non-cancer risk from acute/intermediate/chronic exposure to DIBP in the draft risk
evaluation of DIBP.
EPA is soliciting comments from the Science Advisory Committee on Chemicals (SACC) and the public
on the non-cancer hazard identification, dose-response and weight of evidence analyses, and the selected
POD for use in risk characterization of DIBP.
Table ES-1. Non-cancer HED and HEC Used to Estimate Ris
ks
Exposure
Scenario
Target Organ
System
Species
Duration
POD
(mg/kg-
day)
Effect
HED
(mg/
kg-day)
HEC
(mg/m3)
[ppm]
Benchmark
MOE
Reference
Acute,
intermediate,
chronic
Developmental
toxicity
Rat
4 days
during
gestation
(GDs 14-
18)
BMDLS=
24
I ex vivo
fetal
testicular
testosterone
production
5.7
30.9
[2.71]
UFa= 311
ufh=io
Total UF=30
Grav et al..
2021
HEC = human equivalent concentration; HED = human equivalent dose; MOE = margin of exposure; NOAEL = no-observed-adverse-
effect level; POD = point of departure; UF = uncertainty factor
"EPA used allometric bodv weight scaling to the three-quarters power to derive the HED. Consistent with EPA Guidance (U.S. EPA,
201 let. the UFa was reduced from 10 to 3.
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1 INTRODUCTION
In December 2019, EPA designated diisobutyl phthalate (DIBP) (CASRN 85-69-5) as a high-priority
substance for risk evaluation following the prioritization process as required by Section 6(b) of the
Toxic Substances Control Act (TSCA) and implementing regulations (40 CFR 702). The Agency
published the draft and final scope documents for DIBP in 2020 (U.S. EPA. 2020a. b). 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 DIBP and conduct a dose-response assessment to
determine the toxicity values to be used to estimate risks from DIBP exposures. This technical support
document summarizes the non-cancer human health hazards associated with exposure to DIBP and
proposes non-cancer toxicity values to be used to estimate risks from DIBP exposures. Cancer human
health hazards associated with exposure to DIBP are summarized in EPA's Draft Cancer Raman Health
Hazard Assessment for Di (2-e thy Ihexyl) Phthalate (DEHP), Dibutyl Phthalate (DBP), Diisobutyl
Phthalate (DIBP), Butyl Benzyl Phthalate (BBP) and Dicyclohexyl Phthalate (DCHP) (U.S. EPA.
2025a).
Over the past several decades, the human health effects of DIBP have been reviewed by several
regulatory and authoritative agencies, including: the U.S. Consumer Product Safety Commission (U.S.
CPSC); Health Canada; the European Chemicals Agency (ECHA); the Australian National Industrial
Chemicals Notification and Assessment Scheme (NICNAS); and The National Academies of Sciences,
Engineering, and Medicine (NASEM). EPA relied on information published in these assessments as a
starting point for its human health hazard assessment of DIBP. Additionally, EPA considered new
literature published since the most recent existing assessments of DIBP 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 DIBP literature is
described in the Draft Systematic Review Protocol for Diisobutyl Phthalate (DIBP) (U.S. EPA. 2024k).
EPA's approach and methodology for identifying and using human epidemiologic data and experimental
laboratory animal data is described in Sections 1.1 and 1.2
1.1 Human Epidemiologic Data: Approach and Preliminary Conclusions
To identify and integrate human epidemiologic data into the draft DIBP Risk Evaluation, EPA first
reviewed existing assessments of DIBP conducted by regulatory and authoritative agencies, as well as
several systematic reviews of epidemiologic studies of DIBP published by Radke et al., in the open
literature. Although the authors (i.e., Radke et al.) are affiliated with the U.S. EPA's Center for Public
Health and Environmental Assessment, the reviews do not reflect EPA policy. Existing epidemiologic
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 of varying structure and scope. The assessments and open literature
used as a baseline in this risk evaluation are listed below.
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):
Application of systematic review methods in an overall strategy for evaluating low-dose toxicity
fi'om endocrine active chemicals ^NASEM. 2017);
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Phthalate exposure and male reproductive outcomes: A systematic review of the human
epidemiological evidence (Radke et al.. 2018);
Phthalate exposure andfemale reproductive and developmental outcomes: A systematic review
of the human epidemiological evidence (Radke et al.. 2019b);
Phthalate exposure and metabolic effects: A systematic review of the human epidemiological
evidence (Radke et al.. 2019a); and
Phthalate exposure and neurodevelopment: A systematic review and meta-analysis of human
epidemiological evidence (Radke et al.. 2020a).
Next, EPA sought to identify new population, exposure, comparator, and outcome (PECO)-relevant
literature published since the most recent existing assessment(s) of DIBP by applying a literature
inclusion cutoff date. For DIBP, 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 a robust and the most recent evaluation of human epidemiologic data for DIBP. Health Canada
evaluated epidemiologic study quality using the Downs and Black method (Downs and Black. 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 DIBP 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 DIBP Docket
(https://www.regulations.gov/docket/EPA-HQ-QPPT-2018-0434). and these studies 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. 2021). Data quality evaluations for new studies
reviewed by EPA are provided in the Data Quality Evaluation Information for Raman Health Hazard
Epidemiology for Diisobutyl Phthalate (DIBP) (U.S. EPA. 2024c).
As described further in the Draft Systematic Review Protocol for Diisobutyl Phthalate (DIBP) (U.S.
EPA. 2024k). EPA considers phthalate metabolite concentrations in urine to be an appropriate proxy of
exposure from all sourcesincluding exposure through ingestion, dermal absorption, and inhalation. As
described in the Application of US EPA IRIS systematic review methods to the health effects of
phthalates: Lessons learned and path forward (Radke et al.. 2020b). the "problem with measuring
phthalate metabolites in blood and other tissues is the potential for contamination from outside sources"
(Calafat et al.. 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 DIBP metabolites in matrices other than urine were considered supplemental and not
evaluated for data quality.
EPA used epidemiologic studies of DIBP qualitatively. This is consistent with Health Canada, U.S.
CPSC, and ECHA. EPA did not use 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. The majority of epidemiological studies introduced additional
uncertainty by considering DIBP in isolation and failing to account for confounding effects from co-
exposure to mixtures of multiple phthalates (Shin et al.. 2019; Aylward et al.. 2016). Conclusions from
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Health Canada (2018a. b), NASEM (2017) and systematic review articles by Radke et al. (2020a;
2019b; 2019a; 2018) regarding the level of evidence for association between urinary DIBP 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 (as described in the Draft Systematic Review Protocol for Diisobutyl
Phthalate (DIBP) (U.S. EPA. 2024k)) to use qualitatively during evidence integration to inform hazard
identification and the weight of scientific evidence (Shin et al.. 2019; Aylward et al.. 2016).
1.2 Laboratory Animal Data: Summary of Existing Assessments,
Approach, and Methodology
1.2.1 Summary of Existing Assessments
The human health hazards of DIBP have been evaluated in existing assessments by the U.S. CPSC
(2014. 2011). Health Canada (ECCC/HC. 2020; EC/HC. 2015b). ECHA (2017a. b), and Australia
NICNAS (2016. 2008a. b). These assessments have consistently identified toxicity to the developing
male reproductive system as the most sensitive and robust outcome for use in estimating human risk
from exposure to DIBP. The PODs from these assessments are shown in Table 1-1.
Additionally, a recent systematic review of animal toxicology studies of DIBP was published by Yost et
al. (2019) in the open literature. Although the authors (i.e., Yost et al.) are affiliated with the U.S. EPA's
Center for Public Health and Environmental Assessment, the review does not reflect EPA policy.
Consistent with existing assessments of DIBP by regulatory bodies, Yost et al. (2019) concluded that
there was: robust evidence that DIBP causes male reproductive toxicity and robust evidence that DIBP
causes developmental toxicity. Additionally, Yost et al. (2019) concluded that there was slight evidence
for female reproductive toxicity and effects on the liver, and indetermincmt evidence for effects on
kidney. However, for these hazards, evidence was "limited by the small number of studies, experimental
designs that were suboptimal for evaluating outcomes, and study evaluation concerns such as incomplete
reporting of methods and results."
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394 Table 1-1. Summary of DIBP Non-cancer POPs Selected for Use by other Regulatory Organizations
Brief Study Description
NOAEL/
LOAEL
(mg/kg-
day)
Critical Effect
o
w
u
in
a.
U
t/5
P
ECCC/HC (2020)
NICNAS (2008a)
3
03
(N
O
w
<
X
u
H
3
O
w
<
X
u
UJ
Pregnant Sprague-Dawley rats (20-22 pregnant
rats/dose) gavaged with 0, 250, 500, 750, 1000
mg/kg-day DIBP on GD 6-20 (non-guideline
study) (Saillenfait et al 2006)
250/500
I fetal body weight (both sexes);
t incidence of cryptorchidism
Pregnant Sprague-Dawley rats (11-14 dams/dose)
gavaged with 0, 125, 250, 500, 625 mg/kg-day of
DIBP on GD 12-21 (non-guideline study)
(Saillenfait et al 2008)
125/250
I AGD, NR, testicular pathology
(degeneration of seminiferous
tubules) and oligo-/azoospermia
in epididymis)
None/ 125
Testicular pathology
(degeneration of seminiferous
tubules) and oligo-/azoospermia
in epididymis)
Pregnant rats (6-8/group) exposed to 0, 20, 200,
2000, or 10,000 ppm DBP via diet from GDI5 -
PND21 (equivalent to 0, 1.5, 14, 148, 712 mg/kg-
day [males]; 0, 3, 29, 291, 1372 mg/kg-day
[females]). F1 evaluated at PND14, PND21, &
PNW 8-11 (non-suideline study) (Lee et al 2004)6
None/ 2.5b
Reduced spermatocyte
development (PND 21) and
mammary gland changes
(vacuolar degeneration, alveolar
hypertrophy) in adult male
offspring6
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Brief Study Description
NOAEL/
LOAEL
(mg/kg-
day)
Critical Effect
o
(N-
u
in
a.
U
t/5
P
o
(N
O
(N
u
u
u
S3
00
o
o
w
in
<
Z
u
aj
(N
O
w
<
B
U
H
X>l
03
O
w
<
B
U
H
11ECHA (2012a. b) considered the study by Saillenfait et al. to support a LOAEL of 125 mg/kg-day based on increased incidence of testicular
pathology.
h ECHA (2017a. b) concluded "Few reproductive toxicity studies have been published on [DIBP] compared to DEHP and DBP. No two-generation
studies are available and the substance has not been tested at doses <100 mg/kg bw/d. Current data suggest that DIBP could have similar effects to
DBP, if studied at lower dose levels. If the potency difference between DIBP and DBP, as a very rough estimate of the observed effects in Saillenfait
et al. (2008) (type of effects seen at 500 and 625 mg/kg bw-day, corresponding to a difference of 25%). is extrapolated from the high dose area to the
lower dose area, an estimated LOAEL for DIBP would be 25% higher than the current LOAEL for DBP (2 mg/kg bw-day). Available information is
shown in Table B7. A LOAEL for DIBP of 2.5 mg/kg bw-day is selected for use in the current combined risk assessment
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1.2.2 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 DIBP Risk Evaluation. EPA first reviewed existing assessments of DIBP 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 DIBP, and while these authoritative sources identified a broader pool of
studies to inform hazard identification, EPA only selected those studies used quantitatively for dose-
response analysis in prior assessments for further consideration in 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 diisobutylphthalate (DiBP, CASRN84-69-5) (U.S. CPSC. 2011);
Chronic Hazard Advisory Panel on Phthalates and Phthalate Alternatives (with appendices)
(U.S. CPSC. 2014);
State of the science report: Phthalate substance grouping: Medium-chain phthalate esters:
Chemical Abstracts Service Registry Numbers: 84-61-7; 84-64-0; 84-69-5; 523-31-9; 5334-09-
8; 16883-83-3; 27215-22-1; 27987-25-3; 68515-40-2; 71888-89-6 (EC/HC. 2015b):
Screening assessment - Phthalate substance grouping (ECCC/HC. 2020);
Existing chemical hazard assessment report: Diisobutyl phthalate (NICNAS. 2008a);
Phthalates hazard compendium: A summary of physicochemical and human health hazard data
for 24 ortho-phthalate chemicals (NICNAS. 2008b);
C4-6 side chain transitional phthalates: Human health tier II assessment (NICNAS. 2016);
Committee for Risk Assessment (RAC) Opinion on an Annex XV dossier proposing restrictions
on four phthalates (ECHA. 2012b);
Committee for Risk Assessment (RAC) Committee for Socio-economic Analysis (SEAC):
Background document to the Opinion on the A nnex XV dossier proposing restrictions on four
phthalates (ECHA. 2012a);
Opinion on an Annex XV dossier proposing restrictions on four phthalates (DEHP, BBP, DBP,
DIBP) (ECHA. 2017b);
Annex to the Background document to the Opinion on the Annex XV dossier proposing
restrictions on four phthalates (DEHP, BBP, DBP, DIBP) (ECHA. 2017a);
Application of systematic review methods in an overall strategy for evaluating low-dose toxicity
fi'om endocrine active chemicals (NASEM. 2017); and
Hazards of diisobutyl phthalate (DIBP) exposure: A systematic review of animal toxicology
studies (Yost et al.. 2019).
Next, EPA sought to identify new PECO-relevant literature published since the most recent existing
assessment(s) of DIBP by applying a literature inclusion cutoff date. Along with existing assessments,
EPA used the systematic review in the open literature by Yost et al. (2019) as the starting point for this
draft document (publicly available at https://www.ncbi.nlm.nih.gov/pmc/articles/PMC8596331/). The
systematic review by Yost et al. employed a systematic review protocol and included scientific literature
up to July 2017. Further, Yost et al. (2019) considered a range of human health hazards (e.g.,
developmental toxicity, male and female reproductive toxicity, liver and kidney toxicity, and cancer)
across all durations (i.e., acute, intermediate, subchronic, chronic) and routes of exposure (i.e., oral,
dermal, inhalation). Likewise, Yost et. al reached similar conclusions related to the human health
hazards of DIBP, as other assessments by U.S. CPSC, Health Canada, NICNAS, ECHA, and NASEM.
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EPA considered literature published between 2017 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 Systematic Review Protocol for Diisobutyl Phthalate (DIBP) (U.S. EPA. 2024k). EPA then
identified PECO-relevant literature published since the most recent and comprehensive existing
assessment of DIBP by applying a literature inclusion cutoff date from this assessment.
Figure 1-1. Overview of DIBP 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.
Next, EPA reviewed and extracted key study information from those new studies including: 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 DIBP was limited to one new oral exposure study, and the study LOEL was
converted to an HED by allometric scaling 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 (PBPK) or other information to support a chemical-specific quantitative
extrapolation is absent (U.S. EPA. 2011c). EPA's use of allometric body weight scaling is described
further in Appendix C.
Effects on the developing male reproductive system are a focus of EPA's DIBP hazard assessment.
Therefore, EPA also considered literature identified outside of the 2019 TSCA literature search that was
identified through development of EP A's Draft Proposed Approach for Cumulative Risk Assessment of
High-Priority Phthalates and a Manufacturer-Requested Phthalate under the Toxic Substances Control
Act (U.S. EPA. 2023a). which focused on the developing male reproductive system.
Data quality evaluations for DIBP animal toxicity studies are provided in the Data Quality Evaluation
Information for Human Health Hazard Animal Toxicology for Diisobutyl Phthalate (DIBP) (U.S. EPA.
2024b). Notably, Yost et al. (2019) included data quality evaluations, which are documented and
publicly available in the Health Assessment Workspace Collaborative (HAWC)
(https://hawc.epa.gov/assessment/497/). As described further in the Draft Systematic Review Protocol
for Diisobutyl Phthalate (DIBP) (U.S. EPA. 2024k). EPA relied on the data quality evaluations
completed by Yost et al. (2019). which were imported from HAWC to Distiller and are included in the
Data Quality Evaluation Information for Human Health Hazard Animal Toxicology for Diisobutyl
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Phthalate (DIBP) (U.S. EPA. 2024b). Further, as described in the Draft Systematic Review Protocol for
DiisobutylPhthalate (DIBP) (U.S. EPA. 2024k). OPPT harmonized its draft TSCA systematic review
protocol for human health animal toxicology and epidemiologic study data quality evaluations with the
process described in the IRIS Systematic Review Handbook (U.S. EPA. 2022). Therefore, the data
quality evaluations completed by Yost et al. (2019) are reflective of the harmonized TSCA data quality
evaluation process.
1.2.3 New Literature Identified and Hazards of Focus for DIBP
In its review of literature published between 2017 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, EPA identified two new PECO-relevant studies that provided information
pertaining to one primary hazard outcome (i.e., reproductive/developmental toxicity) (Gray et al.. 2021;
Pan et al.. 2017). These new studies of DIBP are discussed further in Section 3.1.2. Based on
information provided in existing assessments of DIBP for developmental and reproductive effects in
combination with new information identified by EPA, the Agency focused its non-cancer human health
hazard assessment on developmental and reproductive toxicity (Section 3.1).
Further, EPA reviewed and supports the conclusions of the systematic review and hazard identification
for DIBP published by Yost et al. (2019). EPA did not identify any new literature that would change the
conclusions of Yost et al. (2019) pertaining to slight evidence for female reproductive effects and liver
effects and indeterminant evidence for kidney effects. Therefore, EPA did not further characterize these
non-cancer hazards in this assessment or carry them forward to dose-response assessment in Section 4.
Genotoxicity and carcinogenicity data for DIBP are summarized in EPA's Draft Cancer Raman Health
Hazard Assessment for Di (2-e thy Ihexyl) Phthalate (DEHP), Dibutyl Phthalate (DBP), Diisobutyl
Phthalate (DIBP), Butyl Benzyl Phthalate (BBP) and Dicyclohexyl Phthalate (DCHP) (U.S. EPA.
2025a).
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2 TOXICOKINETICS
2.1 Oral Route
No in vivo studies of experimental animal models are available that have evaluated the absorption,
distribution, metabolism, and excretion (ADME) properties of DIBP for the oral exposure route.
One intentional human dosing study is available that investigates urinary elimination of DIBP (Koch et
al.. 2012). In this study, an individual volunteer (36-year-old male, 87 kg) was administered a single oral
dose of 60 |ig/kg deuterium-labelled DIBP (5.001 mg total), and urine samples were collected up to 48
hours following dosing (Koch et al.. 2012). Three urinary metabolites of DIBP were detected:
Monoisobutyl phthalate (MIBP); 20H-MIBP; and 30H-MIBP. MTBP was the primary urinary
metabolite of DIBP (70 to 71 percent of excreted DIBP over 24 to 48 hours), while 20H-MIBP
(approximately 19 percent of excreted DIBP over 24 to 48 hours) and 30H-MIBP (0.7 percent of
excreted DIBP over 24 to 48 hours) were minor urinary metabolites. After 24 hours, 90.27 percent of the
administered dose was recovered in urine, indicating DIBP is absorbed across the gastrointestinal tract
and urine is the primary elimination route. After 48 hours, 90.84 percent was recovered. Peak urinary
metabolite concentrations occurred 2.83 hours post-dosing. Urinary elimination half-lives were similar
for MIBP (3.9 hours), 20H-MIBP (4.2 hours) and 30H-MIBP (4.1 hours), indicating rapid absorption
and urinary elimination. Fecal and biliary excretion were not investigated in this study.
MIBP has been measured in human milk in the United States (Hartle et al.. 2018). Korea (Kim et al..
2020; Kim et al.. 2018; Kim et al.. 2015). Italy (Del Bubba et al.. 2018; Latini et al.. 2009). Germany
(Fromme et al.. 2011). Taiwan (Lin et al.. 2011). Switzerland (Schlumpf et al.. 2010). and Sweden
(Hogberg et al.. 2008). indicating that absorbed DIBP can partition into human milk. Furthermore,
because human biomonitoring data reflects recent aggregate exposure, it cannot quantitatively be
attributed to a specific route although it is assumed to predominately come from oral exposure; however,
exposure from the dermal and inhalation routes may also contribute.
For the draft DIBP risk evaluation, EPA will assume 100 percent oral absorption of DIBP. Notably,
other regulatory agencies have also assumed 100 percent oral absorption of DIBP (ECCC/HC. 2020;
ECHA. 2017a. b; EC/HC. 2015b; U.S. CPSC. 2014. 2011).
2.2 Inhalation Route
No controlled human exposure studies or in vivo animal studies are available that evaluate the ADME
properties of DIBP for the inhalation route. EPA will assume 100 percent absorption via inhalation for
the draft DIBP risk evaluation. Notably, ECHA (2017a. b) has also assumed 100 percent absorption via
the inhalation route for DIBP.
2.3 Dermal Route
No in vitro or controlled human exposure studies are available that evaluate the ADME of DIBP for the
dermal route.
One in vivo ADME study is available that indicates dermally absorbed DIBP is widely distributed to
tissues in rats (Elsisi et al.. 1989). Skin on the backs of male Fischer 344 (F344) rats was shaved one
hour before DIBP administration (rats with visual signs of abrasions were eliminated from the study).
Neat carbon-14 labelled DIBP (14C-DIBP) in an ethanol vehicle (30 to 40 mg/kg) was applied to a
circular area of the skin 1.3 centimeters in diameter, which represents a dose of 5 to 8 mg/cm2. Ethanol
was allowed to evaporate and then the application site was covered with a perforated circular plastic
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cup. Rats were then housed in metabolic cages for 7 days during which time urine and feces were
collected every 24 hours. Following 7 days of dermal exposure to 14C-DIBP, Elsisi et al. measured low
levels of radioactivity associated with 14C-DIBP in adipose tissue (0.11 percent of applied dose), muscle
(0.22 percent of applied dose), skin (0.2 percent of applied dose) and other tissues (less than 0.5 percent
of applied dose found in the brain, lung, liver, spleen, small intestine, kidney, testis, spinal cord, and
blood). Thirty-five percent of the applied dose was recovered from the skin at the application site, while
six percent was recovered from the plastic cap. Total recovery of the applied dose was 93 percent. After
24 hours of exposure, approximately 6 percent of the applied dose was recovered in urine, while
approximately 1 percent was recovered in feces. After seven days, approximately a total of 51 percent of
the applied dose was excreted in urine and feces.
As described in Section 2.3 of the Draft Consumer and Indoor Exposure Assessment for Diisobutyl
Phthalate (DIBP) (U.S. EPA. 2025b). the rate of transport of DIBP across the dermal barrier is
considered flux-limited, rather than delivery limited. The physicochemical properties of DIBP (high
molecular weight, large size, and low solubility in water) impede its ability to cross the dermal barrier,
limiting the rate of flux independent of the concentration on the skin. Therefore, to estimate dermal
exposures to DIBP for workers, consumers and the general population, EPA used a flux-limited dermal
absorption approach for liquids and solid articles. Dermal absorption data from the study by Elsisi et al.
(1989) was used to estimate a steady-state flux of neat DIBP of 2.43 x 10"2 mg/cm2/hr, which is
considered representative for dermal contact with liquids or formulations containing DIBP. No empirical
data exists to estimate flux-limited dermal absorption of DIBP from solid articles. Therefore, EPA used
a framework based on physical chemical properties of DIBP to estimate flux for solid articles. Briefly,
EPA assumes that DIBP first migrates from the solid matrix to a thin layer of moisture on the skin
surface. Therefore, absorption of DIBP from solid matrices is considered limited by aqueous solubility
and is estimated using an aqueous absorption model, which is described further in Section 2.3.1 of the
Draft Consumer and Indoor Exposure Assessment for Diisobutyl Phthalate (DIBP) (U.S. EPA. 2025b).
Overall, EPA estimated an 8-hour time weighted average flux value of 1.7 x 10"4 mg/cm2/hr, which is
considered a representative value for worker dermal exposures to solids or articles containing DIBP. For
consumers, dermal flux values range from 1.51 x 10"4 to 9.62 x 10"4 mg/cm2/hr depending on model
input parameters used in the dermal models for different consumer exposure scenarios for solids. For
further information pertaining to EPA's dermal approach, see Section 2.3 of the Draft Consumer and
Indoor Exposure Assessment for Diisobutyl Phthalate (DIBP) (U.S. EPA. 2025b) and Appendix C.2.1 of
the Draft Environmental Release and Occupational Exposure Assessment for Diisobutyl Phthalate
(DIBP) (U.S. EPA. 2025c).
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3 NON-CANCER HAZARD IDENTIFICATION
3.1 Effects on the Developing Male Reproductive System
As discussed in Section 1.2, the effects on the developing male reproductive system has consistently
been identified as the most sensitive effects associated with oral exposure to DIBP in experimental
animal models in existing assessments of DIBP (ECCC/HC. 2020; ECHA. 2017a. b; NASEM. 2017;
EC/HC. 2015b; U.S. CPSC. 2014; ECHA. 2012a. b; U.S. CPSC. 2011; NICNAS. 2008a) as well as
prior systematic reviews (Yost et al.. 2019). EPA identified no new information through systematic
review that would change this conclusion. Therefore, EPA focused its non-cancer hazard
characterization on the developing male reproductive system. Evidence from epidemiological and
laboratory animal studies for developmental and reproductive outcomes is summarized in Sections 3.1.1
and 3.1.2, respectively.
3.1.1 Summary of Available Epidemiological Studies
3.1.1.1 Previous Epidemiology Assessment (Conducted in 2019 or Earlier)
EPA reviewed and summarized conclusions from previous assessments conducted by Health Canada
(2018b) and NASEM (2017). as well as systematic review articles by Radke et al. (2019b; 2018). that
investigated the association between exposure to DIBP and its metabolites and male and female
developmental and reproductive outcomes. As can be seen from Table 3-1, epidemiologic assessments
by Health Canada (2018b). NASEM (2017). and systematic review articles by Radke et al. (2019b;
2018) varied in scope and considered different developmental and reproductive outcomes. Further, these
assessments used different approaches to evaluate epidemiologic studies for data quality and risk of bias
in determining the level of confidence in the association between phthalate exposure and evaluated
health outcomes (Table 3-1). Section 3.1.1.1.1, Section 3.1.1.1.2, and Section 3.1.1.1.3 provide further
details on previous assessments of DIBP by Health Canada (2018b). Radke et al., (2019b; 2018) and
NASEM (2017). respectively, including conclusions related to exposure to DIBP and health outcomes.
Additionally, EPA also evaluated epidemiologic studies published after the Health Canada (2018b)
assessment as part of its literature search (i.e., published between 2018 and 2019) to determine if newer
epidemiologic studies would change the conclusions of existing epidemiologic assessments or provide
useful information for evaluating exposure-response relationship (Section 3.1.1.2).
Table 3-1. Summary of Scope and Methods Used in Previous Assessments to Evaluate the
Association Between DIBP Exposure and Male Reproductive Outcomes
Previous Assessment
Outcomes Evaluated
Method Used for Study
Quality Evaluation
Health Canada
(2018b)
Hormonal effects:
Sex hormone levels (e.g.,
testosterone)
Growth & Development:
AGD
Birth measures
Male infant genitalia (e.g.,
hypospadias/cryptorchidism)
Placental development and gene
expression
Preterm birth and gestational age
Downs and Black (Downs
and Black, 1998)
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Previous Assessment
Outcomes Evaluated
Method Used for Study
Quality Evaluation
Postnatal growth
DNA methylation
Reproductive:
Altered male puberty
Gynecomastia
Changes in semen parameters
Sexual dysfunction (males)
Sex ratio
Radke et al. (2018)
AGD
Hypospadias/cryptorchidism
Pubertal development
Semen parameters
Time to pregnancy
Testosterone
Timing of pubertal development
Approach included study
sensitivity as well as risk of
bias assessment consistent
with the study evaluation
methods described in (U.S.
EPA. 2022)
Radke et al. (2019b)
Pubertal development
Time to pregnancy (Fecundity)
Preterm birth
Spontaneous abortion
ROBINS-I (Sterne et al.,
2016)
NASEM (2017)
AGD
Hypospadias (incidence, prevalence,
and severity/grade)
Testosterone concentrations
(measured at gestation or delivery)
OHAT (based on GRADE)
(TSTTP. 2015)
Abbreviations: AGD = anogenital distance; ROBINS-I= Risk of Bias in Non-randomized Studies of
Interventions; OHAT = National Toxicology Program's Office of Health Assessment and Translation;
GRADE = Grading of Recommendations, Assessment, Development and Evaluation.
3.1.1.1.1 Health Canada (2018b)
Health Canada evaluated studies that looked at individual phthalates (or their metabolites) and health
outcomes and did not consider studies that only looked at summed exposure to multiple phthalates due
to the challenging nature of interpreting results for the sum of several phthalates. The outcomes that
were evaluated are listed in Table 3-1. To evaluate the quality of individual studies and risk of bias,
Health Canada (2018b) used the Downs and Black evaluation criteria (Downs and Black. 1998). which
is based on the quality of the epidemiology studies and the strength and consistency of the relationship
between a phthalate and each health outcome. The level of evidence for association of a phthalate and
each health outcome was established based on the quality of the epidemiology studies and the strength
and consistency of the association.
Health Canada (2018b) evaluated several studies that investigated the association between urinary
metabolites of DIBP and several developmental and reproductive outcomes. Health Canada concluded
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that there was some limited evidence of association1 for DIBP and several outcomes, including changes
in serum levels of sex hormones (e.g., follicle stimulating hormone, luteinizing hormone, testosterone),
increased sperm DNA damage and apoptosis, and changes in infant sex ratio at birth. For other health
outcomes, Health Canada concluded there was inadequate evidence of association1 (i.e., for changes in
thyroid and other miscellaneous hormones, changes in semen parameters, pregnancy complication and
loss, sexual dysfunction in males and females, and age at menopause). In addition, there was no
evidence of association1 based on lack of changes in AGD, birth weight, birth length, head
circumference, femur length, preterm birth, gestational age, altered male puberty, gynecomastia, time to
pregnancy, uterine leiomyoma, and polycystic ovary syndrome, or that the level of evidence of
association could not be established due to limitations in the available studies (i.e., for changes in
placental development, postnatal growth, altered female puberty, altered fertility).
3.1.1.1.2 Radke et al. (2019b: 2018)
Radke et al. conducted systematic reviews of male (Radke et al.. 2018) and female (Radke et al.. 2019b)
developmental and reproductive outcomes. These systematic review articles are considered herein.
Radke et al. (2018) evaluated the associations between DIBP or its metabolite (MIBP) and male
reproductive outcomes, including AGD and hypospadias/cryptorchidism following in utero exposures;
pubertal development following in utero or childhood exposures, and semen parameters, time to
pregnancy (following male exposure), and testosterone following adult exposures. Male reproductive
outcomes and level of confidence in the associations is listed in Table 3-2.
Data quality evaluation criteria and methodology used by Radke et al. (2018) were qualitatively similar
to those used by NASEM (2017) (i.e., OHAT methods) and Health Canada (2018b). Similar to NASEM
(2017) and Health Canada (2018b). most studies reviewed by Radke et al. (2018) relied on phthalate
metabolite biomarkers for exposure evaluation. Therefore, different criteria were developed for short-
chain (DIBP, DEP, DBP, BBP) and long-chain (DEHP, DINP) phthalates due to better reliability of
single measures for short-chain phthalates. Radke et al. (2018) used data quality evaluations to inform
overall study confidence classifications, and ultimately evidence conclusions of "Robust," "Moderate,"
"Slight," "Indeterminate," or "Compelling evidence of no effect." "Robust" and "Moderate" evidence of
an association is distinguished by the amount and caliber of data that can be used to rule out other
possible causes for the findings. "Slight" and "Indeterminate" describe evidence for which uncertainties
prevent drawing a causal conclusion in either direction.
Table 3-2. Summary of Epidemiologic Evidence of Male Reproductive Effects Associated with
Exposure to DIBP
Timing of Exposure
Outcome
Level of Confidence in Association
In utero
Anogenital distance
Slight
Hypospadias/cryptorchidism
Slight
In utero or childhood
Pubertal development
Indeterminate
1 Health Canada defines limited evidence as "evidence is suggestive of an association between exposure to a phthalate or its
metabolite and a health outcome; however, chance, bias or confounding could not be ruled out with reasonable confidence."
Health Canada defines inadequate evidence as "the available studies are of insufficient quality, consistency or statistical
power to permit a conclusion regarding the presence or absence of an association." Health Canada defines no evidence of
association as "the available studies are mutually consistent in not showing an association between the phthalate of interest
and the health outcome measured."
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Timing of Exposure
Outcome
Level of Confidence in Association
Adult
Semen parameters
Slight
Time to pregnancy
Slight
Testosterone
Moderate
Male Reproductive Outcomes Overall
Moderate
"Table from Figure 3 in Radke et al. (2018).
Similar to the conclusions of Health Canada, Radke et al. (2019b; 2018) found moderate evidence of an
association2 between exposure to DIBP and decreased testosterone levels in males, while evidence of an
association between exposure to DIBP and other male and female reproductive outcomes was found to
the slight {i.e., for decreased AGD, hypospadias and/or cryptorchidism, changes in semen parameters,
time to pregnancy [based on male exposure to DIBP]) or indeterminant {i.e., for male and female
pubertal development, spontaneous abortion, time to pregnancy [based on female exposure to DIBP]).
3.1.1.1.3 NASEM (2017)
NASEM (2017) included a systematic review of the epidemiological evidence of the associations
between exposure to various phthalates or their monoester or oxidative metabolites including DIBP, and
the following male reproductive outcomes 1) AGD measurements, 2) incidence, prevalence, and
severity/grade of hypospadias, and 3) testosterone concentrations measured at gestation or delivery. In
contrast to Health Canada (2018b). and Radke et al. (2018). NASEM (2017) relied on methodological
guidance from the National Toxicology Program's Office of Health Assessment and Translation
(OHAT) to assign confidence ratings and determine the certainty of the evidence to ultimately draw
hazard conclusions (NTP. 2015).
NASEM (2017) concluded that there was inadequate evidence to establish an association between
prenatal exposure to DIBP and hypospadias due to the limited number of studies and dissimilar matrices
utilized to evaluate them (urine and amniotic fluid). NASEM also concluded that there is inadequate
evidence to determine whether fetal exposure to DIBP is associated with a decrease in fetal testosterone
in males, given the various matrices used to measure testosterone (amniotic fluid, maternal serum, or
cord blood), the differences in timing of exposure (during pregnancy or at delivery), and the limited
number of studies. This conclusion is slightly different from those of Health Canada (2018b) and Radke
et al. (2019b; 2018). because they are looking at different life stages, each of which found limited and
moderate evidence, respectively, of an association between exposure to DIBP and decreased testosterone
levels in males. Radke et al. (2018) and Health Canada (2018b) considered the association between
exposure to DIBP and testosterone in children and adults while NASEM looked at fetal life stages.
NASEM also concluded that there was an inadequate level of evidence to determine an association
between DIBP (MIBP) and AGD, although there was moderate confidence in the evidence of
association based on three prospective cohort studies. However, NASEM also conducted a meta-analysis
2 Radke et al. (2019b: 2018) define Robust and Moderate evidence descriptors as "evidence that supports a hazard,
differentiated by the quantity and quality of information available to rule out alternative explanations for the results." Slight
and indeterminant evidence descriptors are defined as "evidence that could support a hazard or could support the absence of a
hazard. These categories are generally limited in terms of quantity or confidence level of studies and serve to encourage
additional research across the exposure range experienced by humans."
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of three studies (Jensen et al.. 2016; Swan et al.. 2015; Swan. 2008) and found that the available studies
do not support the association between DIBP exposure and decreased AGD (% change [95% CI] = -2.23
[-5.15, 0.70] [p = 0.13]). The AGD effect estimates in the NASEM (2017) meta-analyses are slope
estimates based on the assumption that exposure and effect have a monotonic dose-response
relationship. This conclusion is similar to the conclusions of Radke et al. (2018). who found slight
evidence of an association between DIBP exposure and decreased AGD.
3.1.1.1.4 Summary of the Existing Assessments of Male Reproductive Effects
Each of the three assessments discussed above provided qualitative support as part of the weight of
scientific evidence for the association between DIBP exposure and male reproductive outcomes. The
existing assessments and review article came to similar conclusion on the effect of exposure to DIBP
and male reproductive outcomes. Radke et al. (2018) concluded that there was a slight level of
confidence in the association between exposure to DIBP and AGD, while Health Canada (2018b) and
NASEM (2017) found inadequate evidence of an association. Further, Radke et al. (2018) found that
there was moderate evidence for the association between testosterone and exposure to DIBP, while
Health Canada (2018b) found that total testosterone (TT) and free testosterone (fT) had negative
associations (i.e., increase exposure to DIBP with decrease testosterone) in peri pubertal or adolescent
boys (6-12, 8-14 or 12-20 years) per IQR increase with exposure to DIBP and its metabolite MIBP, and
negative associations for total testosterone in adult males 17 to 52 years. Radke et al. (2018). also found
a slight level of confidence in the association between exposure to DIBP and
cryptorchidism/hypospadias, but this association was not consistent with the findings of Health Canada
(2018b) or NASEM (2017). The scope and purpose of the assessments by Health Canada (2018b).
systematic review articles by Radke et al. (2018). and the report by NASEM (2017) differ and may be
related to differences in quality evaluation and confidence conclusions drawn. Health Canada (2018b)
was the most comprehensive review, and considered prenatal and perinatal exposures, as well as
peripubertal exposures and multiple different outcomes. NASEM (2017) evaluated fewer
epidemiological outcomes than Health Canada (2018b) and systematic review articles by Radke et al.
(2018). but also conducted a second systematic review of the animal literature, which will be discussed
further in Section 4.The results of the animal and epidemiological systematic reviews were considered
together by NASEM (2017) to draw hazard conclusions. Each of the existing assessments covered above
considered a different number of epidemiological outcomes and used different data quality evaluation
methods for risk of bias. Despite these differences, and regardless of the limitations of the
epidemiological data, each assessment provides qualitative support as part of the weight of scientific
evidence.
3.1.1.2 EPA Summary of New Studies (2018 to 2019)
EPA also evaluated epidemiologic studies published after the Health Canada (2018b) assessment as part
of its literature search (i.e., published between 2018 and 2019). EPA identified 40 new epidemiologic
studies (24 developmental and 16 reproductive) that evaluated the association between urinary DIBP
and its metabolite (MIBP) and reproductive and developmental outcomes. Studies reporting a significant
association are discussed further below.
Further information (i.e., data quality evaluations and data extractions) on the new studies identified by
EPA can be found in:
Draft Data Quality Evaluation Information for Raman Health Hazard Epidemiology for
DiisobutylPhthalate (DIBP) (U.S. EPA. 2024c). and
Draft Data Extraction Information for Environmental Hazard and Human Health Hazard
Animal Toxicology and Epidemiology for Diisobutyl Phthalate (DIBP) (U.S. EPA. 2024a).
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In text below, EPA discussed the evaluation of the new studies by outcome that contribute to the weight
of scientific evidence.
Developmental Outcomes for Males
Twenty-four studies were evaluated for the association between DIBP and developmental outcomes
including birth measures, size trajectory, fetal loss, pubertal development, and gestational duration. Of
those studies, 1 was high confidence, 17 were of medium confidence and 6 were of low confidence.
There were only four studies with significant results, one high confidence study (Harlev et al.. 2019).
one medium confidence study (Burns et al.. 2022) and two low confidence study (Durmaz et al.. 2018;
Yang et al.. 2018). The remaining 20 studies evaluating developmental outcomes in males did not show
any significant results and are not discussed further in this document. However, further information for
these 20 studies can be found in the Draft Data Quality Evaluation Information for Raman Health
Hazard Epidemiology for Diisobutyl Phthalate (DIBP) (U.S. EPA. 2024c) and Draft Data Extraction
Information for Environmental Hazard and Human Health Hazard Animal Toxicology and
Epidemiology for Diisobutyl Phthalate (DIBP) (U.S. EPA. 2024a).
In the evaluation of pubertal development and DIBP exposure, one high confidence (Harlev et al.. 2019)
and one medium confidence (Burns et al.. 2022) study examined the relationship between exposure to
MIBP and pubertal onset and both reported increasing developmental delay in association with MIBP
exposure. The high confidence study (Harlev et al.. 2019) examined the relationship between prenatal
MIBP exposure and pubertal timing (thelarche, pubarche, menarche, gonadarche) among 159 boys and
179 girls enrolled in the CHAMACOS Study and found significant positive association between prenatal
MIBP exposure (measured via maternal urinary MIBP) and age at thelarche among girls in exposure
quartile 2 vs. quartile 1 [6.5 month mean shift in age at thelarche, 95% CI (1.0, 12.3)]. However, no
significant associations were found for Q2 or Q4 vs. Ql, and no significant associations were found for
other pubertal timing outcomes among girls or boys. The medium confidence study (Burns et al.. 2022)
examined the association between prepubertal MIBP exposure (assessed via urinary MIBP
concentrations) in relation to age at pubertal onset among 304 boys enrolled in the Russia Children's
Study. Puberal onset outcomes were defined as testicular volume greater than 3 mL, Tanner Genitalia
Stage greater than or equal to 2, and Tanner Pubarche Stage greater than or equal to 2. Significant
positive associations were found for all three outcomes. Significant mean delays in testicular growth
were found across all quartiles, as compared to Ql [Q2 vs Ql: 8.5 months, 95% CI (3.7, 13.5); Q3 vs
Ql: 6.4, 95% CI (1.1, 11.7); Q4 vsQl: 5.7(0.2, 11.1). Significant mean delays reaching a Tanner
Genital Stage > 2 were found for Q2 and Q3 vs Ql [Q2 vs Ql: 6.4 months, 95% CI (0.2, 12.6); Q3 vs
Ql: 7.2 (0.5, 13.8)] but not for Q4 vs Ql. Significant mean delays in reaching Pubarche Stage > 2c were
found for Q3 and Q4 vs Ql [Q3 vs Ql: 10.2 months, 95% CI (2.9, 17.5); Q4 vs Ql: 12.8, 95% CI (5.3,
20.3)], but not for Q2 vs Ql. Trend tests were only significant for increasing quartiles of MIBP exposure
for Pubarche Stage greater than or equal to 2c.
Other Developmental Outcomes
Other developmental outcomes such as body mass index (BMI) trajectories were also assessed. One
medium confidence study (Heggeseth et al.. 2019) and two low confidence study (Durmaz et al.. 2018;
Yang et al.. 2018) examined BMI trajectories in relation to MIBP exposure. Heggeseth et al. (2019)
(medium confidence) used growth mixture models and functional principal components analysis to
assess whether prenatal phthalate exposure helped explain variation in size trajectory among 162 boys
and 173 girls enrolled in Center for the Health Assessment of Mothers and Children of Salinas
(CHAMACOS) Study. The study, although no effect estimates were provided, found that urinary
concentrations of MIBP at greater than or equal to 1.7 ng/mL explains variation in BMI in boys. One
low confidence study (Yang et al.. 2018) examining BMI trajectories in relation to MIBP exposure
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among 239 children from Mexico City enrolled in the Early Life in Mexico to Environmental Toxicants
(ELEMENT) found, without reporting effect estimates, that exposure to the first tertile of MIBP
predicted the lowest BMI trajectory in infancy and early childhood but crossed over to predict the
highest BMI by age 14. The other low confidence study (Durmaz et al.. 2018) examined the relationship
between MIBP exposure and BMI and weight in 29 girls between the ages of 4 years and 8 years with
premature thelarche, from Antalya City, Turkey and found significant positive associations for both
weight (Spearman correlation coefficient: 0.742, p< 0.01) and BMI (Spearman correlation coefficient:
0.574, p = 0.002).
Reproductive Outcomes for Males
Five medium confidence studies evaluated the association between DIBP exposure and male
reproductive outcomes; however, only one (Wenzel et al.. 2018) found significant results.
Epidemiologic literature that identified male reproductive effects associated with DIBP exposure found
one medium confidence study (Wenzel et al.. 2018) of infants in Charleston, South Carolina that
reported a significant positive association between maternal urinary concentrations of MIBP and
anoscrotal distance in white infants only [Beta (95% CI) per unit increase in MIBP for anoscrotal
distance = 1.68 (0.09, 3.27)]. No other significant results were reported for other anthropometric
measurements or when results were not stratified by race/ethnicity. Studies on other male reproductive
effects such as anthropometric measures of male reproductive organs, sperm parameters, prostate and
male reproductive hormones found no significant associations. However, further information for these 5
studies can be found in the Draft Data Quality Evaluation Information for Raman Health Hazard
Epidemiology for Diisobutyl Phthalate (DIBP) (U.S. EPA. 2024c) and Draft Data Extraction
Information for Environmental Hazard and Human Health Hazard Animal Toxicology and
Epidemiology for Diisobutyl Phthalate (DIBP) (U.S. EPA. 2024a).
Reproductive Outcomes for Females
Eleven studies (1 high confidence, 9 medium confidence, 1 uninformative) evaluated the association
between DIBP exposure and female reproductive outcomes. Of those studies, two medium confidence
studies (Chin et al.. 2019: Machtinger et al.. 2018) and one low confidence studv(Durmaz et al.. 2018)
had significant results. The remaining eight studies evaluating reproductive outcomes in females did not
show any significant results and are not discussed further in this document. However, further
information for these 9 studies can be found in the Draft Data Quality Evaluation Information for
Human Health Hazard Epidemiology for Diisobutyl Phthalate (DIBP) (U.S. EPA. 2024c) and Draft
Data Extraction Information for Environmental Hazard and Human Health Hazard Animal Toxicology
and Epidemiology for Diisobutyl Phthalate (DIBP) (U.S. EPA. 2024a).
Female reproductive effects associated with DIBP exposure were identified in two medium confidence
studies (Chin et al.. 2019: Machtinger et al.. 2018) and one low confidence study (Durmaz et al.. 2018).
Chin et al. (2019) (medium confidence study) investigated North Carolina women without known
fertility issues and reported significantly increased odds of a shorter time between ovulation and
implantation [OR (95% CI) for early implantation = 2.09 (95% CI=1.18, 3.69)]. The other medium
confidence study (Machtinger et al.. 2018) examined women undergoing in vitro fertilization (IVF) in
Israel and reported a significantly reduced mean number of total oocytes in tertile 2 compared to tertile 1
of urinary MIBP in women undergoing a fresh IVF cycle [Mean difference (95%) CI for tertile 2 = 8.7
(7.9, 9.6)]. This study further reported a significantly reduced mean number of mature oocytes in both
tertiles 2 and 3 compared to tertile 1 of MIBP exposure [Mean difference (95% CI) for tertile 2 = 6.7
(6.0, 7.5); Mean (95% CI) for tertile 3 = 8.0 (7.2, 8.8)]. The mean number of fertilized oocytes was also
significantly reduced in tertile 2 compared to tertile 1 of MIBP exposure [Mean difference (95% CI) for
tertile 2 = 4.6 (4.0, 5.3)]. Women with higher MIBP exposure also had a significantly reduced mean
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number of top-quality embryos. This study also reported significantly reduced mean number of top-
quality embryos [Mean difference (95% CI) for tertile 2 = 2.0 (1.7, 2.5); Mean (95% CI) for tertile 3 =
2.2 (1.8, 2.7)]. The low confidence study (Durmaz et al.. 2018) conducted in Turkey reported a
significant unadjusted positive correlation between urinary MIBP concentrations and basal follicle
stimulating hormone (FSH) in girls with premature thelarche [Spearman correlation coefficient between
MIBP and basal FSH = 0.323, p-value = 0.045], Other studies that examined female reproductive
measures, such as anthropometric measures of female reproductive organs or fibroids, and association
with DIBP exposure found no significant association.
Conclusion
In conclusion, Health Canada (2018b) and NASEM (2017) found inadequate evidence of association
between DIBP and AGD while systematic review articles published by Radke et al. (2018) found slight
evidence of association with AGD. Moreover, new studies identified by EPA from 2018 to 2019 do not
alter the previous conclusions from Health Canada (2018b) and NASEM (2017). and systematic review
articles published by Radke et al. (2018). Although there is slight evidence of an association between
DIBP and AGD. the results for testosterone were measured at different life stages (i.e., fetal/infants to
adults) and causality could not be established, thus the overall evidence does not support an association
between DIBP and AGD or testosterone.
Furthermore, EPA preliminarily concludes that the existing epidemiological studies do not support
quantitative exposure-response assessment due to uncertainty associated with exposure characterization
of individual phthalates, including source or exposure and timing of exposure as well as co-exposure
confounding with other phthalates, discussed in Section 1.1. The epidemiological studies provide
qualitative support as part of the weight of scientific evidence.
3.1.2 Summary of Laboratory Animals Studies
EPA identified 13 oral exposure studies (11 of rats, 2 of mice) that have investigated the effects of DIBP
on the developing male reproductive system (Gray et al.. 2021; Pan et al.. 2017; Saillenfait et al.. 2017;
Wang et al.. 2017; Sedha et al.. 2015; Furr et al.. 2014; Hannas et al.. 2012; Hannas et al.. 2011;
Howdeshell et al.. 2008; Saillenfait et al.. 2008; BASF. 2007; Borch et al.. 2006; Saillenfait et al.. 2006).
No studies evaluating the developmental and/or reproductive toxicity of DIBP are available for the
inhalation or dermal exposure routes.
Available oral exposure studies of DIBP evaluating developmental and reproductive outcomes are
summarized in Table 3-3. Most of the available studies evaluate effects on the developing male
reproductive system consistent with a disruption of androgen action following gestational, perinatal, or
pre-pubertal oral exposures to DIBP. However, several studies are available that evaluate other
developmental outcomes (e.g., post-implantation loss, resorptions, fetal body weight, skeletal variations,
etc.). Effects on the developing male reproductive system and other developmental and reproductive
outcomes are discussed in Sections 3.1.2.1 and 3.1.2.2, respectively.
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Table 3-3. Summary of Studies of DII
P Evaluating Developmental and Reproductive Outcomes
Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
(Howdeshell
et al.. 2008)
Pregnant SD rats (5-8
dams/dose) gavaged with 0,
100, 300, 600, 900 mg/kg-
day DIBP onGDs8-18
100/300
I ex vivo testicular
testosterone production
(40%) on GD 18
Maternal Effects
- i Maternal body weight on GD 18 (>600 mg/kg-day) and weight gain (900)
Developmental Effects
-1 fetal morality (900 mg/kg-day)
- i # of live fetuses (900 mg/kg-day)
-1 total resorptions (900 mg/kg-day)
(Hannas et al..
2011)
Pregnant SD rats (3
dams/dose) gavaged with 0,
100, 300, 600, 900 mg/kg-
day DIBP on GDs 14-18
100/300
I ex vivo fetal testicular
testosterone production
(56%) and j expression
of steroidogenic genes in
fetal testes on GD 18
Maternal Effects
- None
Developmental Effects
-1 ex vivo fetal testicular testosterone production on GD 18 (>300 mg/kg-day)
-1 Fetal testis mRNA levels of StAR (>300 mg/kg-day) and Cvplla on GD 18
(>100 mg/kg-day)
Unaffected Outcomes
- Maternal mortality, clinical signs of toxicity, maternal body weight, litter size
(Saillenfait et
al.. 2008)
Pregnant SD rats (11-14
dams/dose) gavaged with 0,
125, 250, 500, 625 mg/kg-
day DIBP on GDs 12-21
None/125
t Testicular pathology
(degeneration of
seminiferous tubules) and
oligo-/azoospennia in
epididymis)
Maternal Effects
- None
Developmental Effects
-1 Male AGD (absolute) on PND 1 (>250 mg/kg-day)
-1 Male NR on PNDs 12-14 and at necropsy on PNW 11-12 or PNW 16-17
(>250 mg/kg-day)
- i Male pup weight on PND 1 and PND 21 (625 mg/kg-day)
-1 hypospadias (>500 mg/kg-day), cleft prepuce (625 mg/kg-day), exposed os
penis (>500 mg/kg-day), non-scrotal testes at necropsy (PNW 11-12 or 16-17)
(>500 mg/kg-day)
- Delayed PPS (>500 mg/kg-day)
-1 male offspring body weight on PNW 11-12 and PNW 16-17 (>500 mg/kg-
day)
- i absolute prostate weight on PNW 11-12 (>250 mg/kg-day) and 16-17 (>500
mg/kg-day); j absolute testis, epididymis, and SV weight on PND 11-12 and
16-17 (>500 mg/kg-day)
Unaffected Outcomes
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Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
- Maternal weight gain (GD 0-12, GD 12-21, PND 1-21); post-implantation
loss; % live pups; pup survival (PND 1-4, PND 4-21); female offspring body
weight on PND 4, 7, 14, 21
(Saillenfait et
al.. 2006)
Pregnant SD rats (20-22
pregnant rats/dose) gavaged
with 0, 250, 500, 750, 1000
mg/kg-day DIBP on GDs 6-
20
250/500
i fetal body weight (7%)
(both sexes); t incidence
of undescended testes
(unilateral or bilateral)
and degree of trans-
abdominal testicular
migration. J, Maternal
weight gain
Maternal Effects
-1 Maternal weight gain on GD 6-9 and GD 15-18 (>500 mg/kg-day)
Developmental Effects
-1 % Resorptions per litter (>750 mg/kg-day)
-1 % Post-implantation losses per litter (>750 mg/kg-day)
- i Number of live fetuses per litter (>750 mg/kg-day)
- i Fetal body weight (|7%) (both sexes) (>500 mg/kg-day)
-1 Total number of fetuses with external, visceral, and skeletal malformations
(>750 mg/kg-day)
-1 Total number of litters with visceral and skeletal malformations (>750
mg/kg-day)
-1 incidence of visceral and skeletal variations, including ectopic testis (>750
mg/kg-day), increased degree of trans-abdominal testicular migration (>500)
Unaffected Outcomes
- Maternal mortality; maternal food consumption; overall maternal weight gain
corrected for gravid uterine weight; % dead fetus per litter; sex ratio
(BASF. 2007)
Pregnant Wistar rats (22-
23/dose) administered diets
containing 0, 1000, 4000,
11,000 ppm DIBP on GDs 6-
20 (equivalent to 88, 363,
942 mg/kg-day)
Adhered to OECD TG 414;
GLP-compliant
363/942
i maternal food
consumption, J, maternal
body weight gain, J, fetal
body weight (5%);
skeletal variations
Maternal Effects
- i Maternal food consumption (approximately 5% below control across GD 6-
20) (942 mg/kg-day)
- i Maternal body weight gain (approximately 11% below control across GD 6-
20) (942 mg/kg-day)
Developmental Effects
- i Fetal (both sexes) body weight (approximately 5% below control) (942
mg/kg-day)
-1 skeletal variations, including incomplete ossification of sternebra and
unilateral ossification of sternebra (942 mg/kg-day)
Unaffected Outcomes
- Maternal mortality; no clinical signs; post-implantation loss; resorptions; # of
viable fetuses; sex ratio; external or visceral malformations or variations
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Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
(Saillenfait et
al.. 2017)
Pregnant SD rats (15-20
/dose) gavaged with 0 or 250
mg/kg-day DIBP on GDs 13-
19
None/250
I AGD, I testicular
testosterone (45%) and,
androstenedione (27%)
production; altered
mRNA expression of
steroidogenesis genes in
the testes
Maternal Effects
- None
Developmental Effects
- i AGD (normalized to cubic root of body weight)
- i (27-45%) ex vivo testis testosterone and androstenedione production
i gene expression in cholesterol and steroid synthesis in fetal testes (Hmg-
CoAR, Hmg-CoAS, SR-B1, StAR, P450cl7, 170-HSD)
Unaffected Outcomes
- Dam body weight gain; gravid uterine weight; post-implantation loss; # live
fetuses per litter; sex ratio; fetal body weight
(Wans et al..
2017)
Pregnant ICR Mice (15-18
offspring/dose) fed diets
containing 0 or 2.8 g
DIBP/kg diet (dry weight)
(equivalent to 450 mg/kg-
day) from GDs 0-21
(designated TC) or from
GDs 0 to PND 21
(designated TT)
None/450
i absolute testes weight
on PND 21; J, serum and
testes testosterone; j
expression of
steroidogenic genes in
testes; |sperm
concentration and
motility on PND 80
Maternal Effects
- None
Developmental Effects
- i absolute testes weight on PND 21 (TT group only)
- i serum and testes testosterone in PND 21 males (TC and TT groups)
- i serum and testes testosterone in PND 80 males (TT group only)
- i mRNA and protein expression of steroidogenic genes in testes of PND 21
and PND 80 males (e.g., Cypl 7a 1) (TC and TT groups)
- i sperm concentration and motility for PND 80 males (TT group only)
Unaffected Outcomes
- Maternal weight gain; litter size; fetal viability; PND 21 male offspring body
weight; offspring liver weight; AGD
(Hannas et al..
2012)
Pregnant SD rats (3/dose)
gavaged with 0 or 500
mg/kg-day DIBP on GDs 14-
18
None/500
I ex vivo fetal testicular
testosterone production
(-25%) on GD 18
Unaffected Outcomes
- Maternal body weight gain maternal liver weight, # of live fetuses.
Pregnant SD rats gavaged
with 0, 100, 300, 600, 900
mg/kg-day DIBP on GDs 14-
18
100/300
(LOEL)
i fetal testicular mRNA
levels of steroidogenic
genes
-1 mRNA expression levels of StAR, Cvpllal, Hsd3b, Cvpl7al, Scarbl,
InsB, Cypl lb 1 (>300)
(Borch et al..
2006)
Pregnant Wistar Rats
(6/dose) gavaged with 0 or
None/600
I ex vivo testes
testosterone production
Maternal Effects
- None
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Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
600 mg/kg-day DIBP on
GDs 7-19 or GDs 7-20/21
(96%), | AGD, t
testicular histopathology
Developmental Effects
- i Testes testosterone content on GD20/21 (effect on GD 19 not statistically
significant)
-1 ex vivo testis testosterone production on GD 20/21, but not GD 19
-1 absolute AGD on GD 19 and GD 20/21; J, AGD (normalized to cubic root
of body weight) on GD 20/21
- i fetal body weight on GD 19
-1 testicular pathology (Leydig cell clusters on GD 19 and GD 20/21), Sertoli
cell vacuolization MNGs, central localization of gonocytes on GD 20/21)
-1 immunohistochemistry staining for StAR and P450scc in Leydig cells
Unaffected Outcomes
- Maternal weight gain during pregnancy; litter size; fetal viability; number of
resorptions
(Furr et al..
2014)°
Pregnant SD rats (3-5/group)
gavaged 0 or 750 mg/kg-day
DIBP on GDs 14-18 (Block
2)
None/750
I ex vivo fetal testicular
testosterone production
(81%) on GD 18
Unaffected Outcomes
- Fetal viability on GDI8
- Dam body weight gain
Pregnant SD rats (3-4/group)
gavaged with 0 or 500
mg/kg-day DIBP on GDs 14-
18 (Block 14)
None/500
I ex vivo fetal testicular
testosterone production
(70%) on GD 18
Unaffected Outcomes
- Fetal viability on GDI8
- Dam body weight gain
Pregnant SD rats (2-4/group)
gavaged with 0 or 200
mg/kg-day DIBP on GDs 14-
18 (Block 30)
None/200
I ex vivo fetal testicular
testosterone production
(47%) on GD 18
Unaffected Outcomes
- Fetal viability on GDI8
- Dam body weight gain
(Sedha et al..
2015)
UtcrotroDhic Assav
Young female Wistar rats
(20 days old) (>6 mice/dose)
gavaged 0, 250, or 1250
mg/kg-day DIBP for 3 days
None/250
i Body weight gain
- i Body weight gain (>250 mg/kg-day)
- Lack of effect on uterus and ovary wet weight indicate DIBP lacks estrogenic
potential
Unaffected Outcomes
- No clinical signs; uterus and ovary wet weight
Pubertal Assav
Young female Wistar rats
(20 days old) (>6 mice/dose)
None/250
i Body weight gain
- i Body weight gain (>250 mg/kg-day)
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Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
gavaged 0, 250, or 1250
DIBP for 20 days (PND 21-
41)
- Lack of effect on reproductive organ weight and vaginal opening indicate
DIBP lacks estrogenic potential
Unaffected Outcomes
- Absolute and relative uterus, ovary, and vagina weight; vaginal opening
New Studies of DIBP Since Yost et al. (2019)
(Pan et al..
2017)
Young (6-8 week old) male
ICR mice (20/dose) fed diets
containing 0 or 2.8 g
DIBP/kg chow (equivalent
received dose of 450 mg/kg-
day) for 28 days
None/450
Sperm effects, [ serum &
testes testosterone, [
mRNA & protein levels
of steroidogenesis genes
- i Epididymal sperm concentration sperm motility, and progressiveness (450
mg/kg-day)
-1 Sperm malformation (450 mg/kg-day)
-j Serum and testis testosterone, j serum follicle stimulating hormone levels
(450 mg/kg-day)
- i mRNA and protein levels of steroidogenic genes in testes (e.g., P450cc,
StAR, 3fi-hsd)
Unaffected Outcomes
- Body weight gain; food intake; absolute and relative testes and epididymis
weight; serum levels of estradiol and luteinizing hormone
(Gray et al..
2021)°
Pregnant SD rats (3-4
dams/dose) were gavaged
with 0, 100, 300, 600, or 900
mg/kg-day DIBP on GDs 14-
18
100/300
I ex vivo testicular
testosterone production
(34%) on GD 18
-1 ex vivo testicular testosterone production on GD18; Block 67 (>300 mg/kg-
day)
- i mRNA expression of Phase I metabolism genes (e.g., Cypl lb 1. Cypl lal,
Cypl7al, ALDH2) (900 mg/kg-day)
- i mRNA expression of lipid signaling and cholesterol metabolism gene (>900
mg/kg-day)
Unaffected Outcomes
- Maternal liver weight (Block 19)
Abbreviations: [ = statistically significant decrease; t = statistically significant increase; NOAEL = No observed adverse effect level; LOAEL = Lowest observed adverse
effect level; GD = Gestational Day; PND = Postnatal Day AGD = Anogenital distance; GLP = Good Laboratory Practice; MNG = multinucleated gonocytes; NR = Nipple
Retention; PPS = preputial separation; SD = Sprague Dawley; SV = Seminal Vesicles; TT = pups exposed both prenatally and postnatally; TC = pups exposed prenatally
only
"These studies were conducted by EPA's Office of Research and Development (ORD).
882
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3.1.2.1 Developing Male Reproductive System
EPA previously developed a weight of scientific evidence analysis and concluded that oral exposure to
DIBP 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 (U.S. EPA. 2023 a)).
Notably, EPA's conclusion was supported by the Science Advisory Committee on Chemicals (SACC)
(U.S. EPA. 2023b). A brief summary of the MOA for phthalate syndrome and data available for DIBP
supporting this MOA is provided below in Figure 3-1. 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 (U.S. EPA. 2023 a) for a more thorough discussion of
DIBP'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.
As shown in Figure 3-1, a MOA for phthalate syndrome has been proposed to explain the link between
gestational or perinatal exposure to DIBP 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-Requested 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
o
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
l=>
Unknown MIE
(not believed to be
AR or PPARa
mediated)
4- Testosterone
synthesis
IT
Key genes involved in the AOP \
for phthalate syndrome
Scarbl Cher?
StAF Ebp
Cypllcl Fdps
Cypllbl Hmgcr
Cypllb2 Hmgcsl
Cypl7ol Hsd3b
CypSl Fldll
Mvd Ela3b
Nsdhl Insl3
RGD1564999 Lhcgr
Tm7sf2 inha
Cyp46al NrObl
Ldlr RhoxlO
Insigl Wnt7a
4/ Gene
expression
(INSL3, lipid
> metabolism,
cholesterol and
androgen synthesis
and transport)
17
4- INSL3 synthesis
Fetal Leydig cell
^>
Abnormal cell
apoptosis/
proliferation
(Nipple/areolae
retention, >1/ AGD,
Disrupted testis
tubules, Leydig cell
clusters, MNGs,
agenesis of
reproductive tissues)
Suppressed
gubernacular cord
development
{inguinoscrotal phase)
Suppressed
gubernacular cord
development
(transabdominal
phase)
4- Androgen-
dependent tissue
weights, testicular
pathology [e.g.,
seminiferous tubule
atrophy),
malformations (e.g.,
hypospadias), 4*
sperm production
<0
Impaired
s.
fertility
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.. 2021; Howdeshell et al.. 2017). AR = androgen receptor; INSL3 = insulin-like growth factor 3; MNG =
multinucleated gonocytes; PPARa = peroxisome proliferator-activated receptor alpha.
Phthalate syndrome is characterized by both androgen-dependent (e.g., reduced AGD, increased male
NR.) and -independent effects (e.g., germ cell effects) on the male reproductive system ( J.S. EPA.
2023a). The MOA underlying phthalate syndrome has not been fully established; however, key cellular-,
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organ-, and organism4evel 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 DIBP 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 (MacLeod et al.. 2010; Welsh et al.. 2008; Carruthers and Foster. 2005). Consistent
with the MOA outlined in Figure 3-1, seven studies (5 of rats, 2 of mice) of DIBP have demonstrated
that oral exposure to DIBP during the masculinization programming window can reduce mRNA and/or
protein expression of insulin-like growth factor 3 (INSL3), as well as genes involved in steroidogenesis
in the testes of rats (Gray et al.. 2021; Saillenfait et al.. 2017; Hannas et al.. 2012; Hannas et al.. 2011;
Borch et al.. 2006) and mice (Pan et al.. 2017; Wang et al.. 2017). Consistently, nine studies (7 of rats, 2
of mice) have also demonstrated that oral exposure to DIBP during the masculinization programming
window can reduce testicular testosterone content and/or testosterone production in rats (Gray et al..
2021; Saillenfait et al.. 2017; Furr et al.. 2014; Hannas et al.. 2012; Hannas et al.. 2011; Howdeshell et
al.. 2008; Borch et al.. 2006) and mice (Pan et al.. 2017; Wang et al.. 2017). Oral exposure of rats to
DIBP during the masculinization programming window has also been shown to reduce male pup
anogenital distance (AGD) in three studies (Saillenfait et al.. 2017; Saillenfait et al.. 2008; Borch et al..
2006) and cause male pup nipple retention (NR) in one study (Saillenfait et al.. 2008). which are two
hallmarks of anti androgenic substances (see Sections 3.1.3.3 and 3.1.3.4 of (U.S. EPA. 2023 a) for
additional discussion). Additional effects consistent with phthalate syndrome observed in mice and rats
following oral exposure to DIBP during the critical window of development include: reproductive tract
malformations (i.e., hypospadias, undescended testes, exposed os penis, cleft prepuce) in two studies of
rats (Saillenfait et al.. 2008; Saillenfait et al.. 2006); delayed preputial separation (PPS) in one study of
rats (Saillenfait et al.. 2008); testicular pathology in two studies of rats (e.g., degeneration of
seminiferous tubules, oligospermia, azoospermia, Leydig cell aggregation, Sertoli cell vacuolation,
multinucleated gonocytes) (Saillenfait et al.. 2008; Borch et al.. 2006); and decreased sperm
concentration and motility in two studies of mice (Pan et al.. 2017; Wang et al.. 2017).
Collectively, available studies consistently demonstrate that oral exposure to DIBP during the
masculinization programming window in rats and mice can disrupt androgen action, leading to a
spectrum of effects on the developing male reproductive system consistent with development of
phthalate syndrome. As noted above, this conclusion was supported by the Science Advisory Committee
on Chemicals (SACC) (U.S. EPA. 2023b) and readers are directed to 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) for additional discussion of DIBP's effects on the
developing male reproductive system and EPA's MOA analysis.
3.1.2.2 Other Developmental Outcomes
In addition to effects on the developing male reproductive system, other developmental effects (e.g.,
decreased fetal weight, decreased offspring body weight, resorptions, post-implantation loss, skeletal
variations) have been observed in experimental animal models following oral exposure to DIBP.
However, these effects occur at higher doses than those that result in effects on the developing male
reproductive system and frequently coincide with maternal toxicity (Table 3-3). Data supporting other
developmental effects of DIBP are discussed below.
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In a study that adhered to OECD test guideline 414, pregnant Wistar rats (22 to 23 per dose) were
administered diets containing 0, 1000, 4000, or 11,000 ppm DIBP (equivalent to 88, 363, or 942 mg/kg-
day) from GD 6 through 20 and then sacrificed on GD 20 (BASF. 2007). Maternal and developmental
effects were limited to the high-dose group and included a 5 percent decrease in maternal food
consumption as well as an 11 percent decrease in maternal bodyweight gain from GD 6 through 20, a 5
percent decrease in fetal body weight, and increased incidences of skeletal variations (e.g., incomplete
ossification of sternebra, unilateral ossification of sternebra). No significant increases in malformations
were observed. No developmental or maternal toxicity was observed in the low- or mid-dose groups.
In a second study, pregnant SD rats (20 to 22 per dose) were exposed to 0, 250, 500, 750, or 1000
mg/kg-day DIBP from GD 6 through 20 via gavage and then sacrificed on GD 21 (Saillenfait et al..
2006). Maternal effects were limited to a decrease in weight gain on GD 6 through 9 and GD 15 through
18 in dams treated with 500 mg/kg-day DIBP and above; however, dam body weight gain on GD 6
through 21 corrected for gravid uterine weight was unaffected. Developmental toxicity was observed at
500 mg/kg-day and above. Observed developmental effects included: increased resorptions and post-
implantation loss per litter and decreased live fetuses per litter at 750 mg/kg-day and above; increased
incidence of total number of fetuses and/or litters with external, visceral, and skeletal malformations at
750 mg/kg-day and above; and increased incidence of undescended testes and decreased fetal body
weight (both sexes) at 500 mg/kg-day and above.
Howdeshell et al. (2008) reported increased fetal mortality and total resorptions, and decreased numbers
of live fetuses in pregnant SD rats gavaged with 900 mg/kg-day DIBP from GDs 8 to GDI 8 and
sacrificed on GD 18. Additionally, Borch et al. (2006) reported reduced fetal body weight on GD 19 in
pregnant Wistar rats gavaged with 600 mg/kg-day DIBP on GD 7 through 19. In addition to decreased
fetal weight, decreased offspring body weight was observed following gestational exposures. Saillenfait
et al. (2008) reported reduced male offspring body weight on PND1, PND21 as well as PNW11 to 12
and PNW16 to 17 following gestational exposure to 500 to 625 mg/kg-day DIBP on GD 12 through 21.
Collectively, available studies provide consistent evidence that gestational exposure to DIBP can result
in a spectrum of developmental effects in addition to those of the developing male reproductive system.
However, effects on the developing male reproductive system (Section 3.1.2.1) occur at much lower
doses than the aforementioned other developmental effects. Therefore, effects on the developing male
reproductive system are the most sensitive to DIBP exposure and are consistent with a disruption of
androgen action and phthalate syndrome. Furthermore, the lowest LOAELs for effects on the developing
male reproductive system range from 125 to 300 mg/kg-day, while the lowest LOAELs for other
developmental outcomes range from 500 to 600 mg/kg-day (Table 3-3).
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4 DOSE-RESPONSE ASSESSMENT
EPA considered reproductive/developmental toxicity as the sole non-cancer hazard endpoint for dose-
response analysis. This hazard endpoint was selected for dose-response analysis because EPA has the
highest confidence in this hazard endpoint for estimating risk to human health; effects were consistently
observed across species and durations of exposure and occurred in a dose-related manner. Other non-
cancer hazard endpoints considered by EPA (i.e., liver and kidney toxicity) were not utilized for dose-
response analysis due to limitations and uncertainties that reduce EPA's confidence in using these
endpoints for estimating risk to human health. For toxicologically similar phthalates (i.e., DEHP, DBP,
BBP, DCHP), which include larger databases of animal toxicology studies including numerous well-
conducted subchronic and chronic toxicity studies, effects on the developing male reproductive system
consistent with a disruption of androgen action have consistently been identified by EPA as the most
sensitive and well-characterized hazard in experimental animal models. This is demonstrated by the fact
that the preliminary acute/intermediate/chronic PODs selected by EPA for use in risk characterization
for DEHP (U.S. EPA. 2024h). DBP (U.S. EPA. 2024f). BBP (U.S. EPA. 2024e). DCHP (U.S. EPA.
2024g) are all based on effects related to phthalate syndrome. According to previous assessments, liver
is a target organ following DIBP exposure (U.S. CPSC. 2011; NICNAS. 2008a); however, Health
Canada (2015b) concluded that DIBP has low systemic toxicity based on a limited number of repeated
oral dose toxicity studies. Additionally, a systematic review by Yost et al. (2019) stated that several
studies indicate dose dependent increases in liver weight following intermediate and chronic DIBP
exposure in rats and male mice (Wang et al.. 2017; Foster et al.. 1982; Oishi andHiraga. 1980;
University of Rochester. 1953). However, there are no available data on other hepatic endpoints, such as
clinical chemistry (e.g., ALT, ALT, bilirubin) and histology effects, following oral DIBP exposure. The
lack of such data reduces EPAs confidence in using effects on the liver as an endpoint from which to
derive a POD, because there is uncertainty about adversity without corroborating clinical chemistry or
histology (Hall et al.. 2012; U.S. EPA. 2002a). Likewise, effects on the kidney following exposure to
DIBP were evaluated by a limited number of studies, wherein inconsistencies across species were
observed, as summarized in previous assessments and publications (Yost et al.. 2019; ECHA. 2017b;
NICNAS. 2016; U.S. CPSC. 2011; NICNAS. 2008a). No new studies were identified that provided data
on hepatic or renal effects following exposure to DIBP were identified through the TSCA systematic
review process; therefore, EPA is in agreement with the conclusions of these previous assessments as
well as those of the systematic review by Yost et al. (2019) [as described previously in Section 1.2.3],
EPA used a BMD modeling approach for individual data sets of fetal testicular testosterone changes for
the dose-response analysis. EPA did consider NOAEL and LOAEL values from oral toxicity studies in
experimental animal models. The use of a NOAEL/LOAEL approach is supported by consistency across
several studies that have evaluated effects on the developing male reproductive system consistent with
phthalate syndrome that are similar and cluster around a single human equivalent dose (HED) NOAEL
value, which supports identification of a consensus NOAEL. For reduced fetal testicular testosterone in
rats, EPA conducted meta-analysis and benchmark dose modeling using the approach previously
published by NASEM (2017). which is further described in EPA's Draft Meta-Analysis and Benchmark
Dose Modeling of Fetal Testicular Testosterone for Di (2-ethylhexyl) Phthalate (DEHP), Dibutyl
Phthalate (DBP), Butyl Benzyl Phthalate (BBP), Diisobittyl Phthalate (DIBP), Dicyclohexyl Phthalate
(DCHP), andDiisononylPhthalate (U.S. EPA. 2024d). Acute, intermediate, and chronic non-cancer
NOAEL, LOAEL, and BMDLs values identified by EPA are discussed further in Section 4.2. As
discussed further in Section 4.2, EPA considers effects on the developing male reproductive system
consistent with a disruption of androgen action relevant for setting a POD for acute exposure durations.
However, because these acute effects are the most sensitive effects following exposure to DIBP, they are
also considered protective of intermediate and chronic duration exposures. As described in Appendix C,
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EPA converted oral PODs derived from animal studies to human equivalent doses (HEDs) using
allometric body weight scaling to the three-quarters power (U.S. EPA. 2011c). Species differences in
dermal and oral absorption are corrected for as part of the dermal exposure assessment (U.S. EPA.
2025c). In the absence of inhalation studies, EPA performed route-to-route extrapolation to convert oral
HEDs to inhalation human equivalent concentrations (HECs) (Appendix C).
4.1 Selection of Studies and Endpoints for Non-cancer Health Effects
EPA considered the suite of oral animal toxicity studies primarily indicating effects on the developing
male reproductive system consistent with phthalate syndrome when considering non-cancer PODs for
estimating risks for acute, intermediate, and chronic exposure scenarios as described in Section 4.2. EPA
considered the following factors during study and endpoint selection for POD determination from
relevant non-cancer health effects:
Exposure duration;
Dose range;
Relevance (e.g., considerations of species, whether the study directly assesses the effect, whether
the endpoint is the best marker for the toxicological outcome, etc);
Uncertainties not captured by the overall quality determination;
Endpoint/POD sensitivity; and
Total uncertainty factors (UFs). EPA considers the overall uncertainty with a preference for
selecting studies that provide a lower uncertainty (e.g., lower benchmark MOE) because
provides higher confidence (e.g., use of a NOAEL or BMDLs vs. a LOAEL with additional UFl
applied).
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.
4.2 Non-cancer Oral Points of Departure for Acute, Intermediate, and
Chronic Exposures
4.2.1 Studies Considered for the Non-Cancer POD
EPA considered 11 developmental toxicity studies (10 of rats, 1 of mice) with endpoints relevant to
acute, intermediate, and chronic exposure duration (U.S. EPA. 1996. 1991). summarized in Table 4-5.
Of the considered studies, all 11 evaluated gestational or perinatal exposures to DIBP. No one or two-
generation studies on the effects of DIBP on reproduction have been identified by EPA. Further, of the
11 studies considered, 5 only evaluated one exposure level of DIBP (i.e., did not evaluate dose-response
across multiple exposure levels) ranging from 200 to 750 mg/kg-day (Table 4-5) (Saillenfait et al.. 2017;
Wang et al.. 2017; Furr et al.. 2014; Hannas et al.. 2012; Borch et al.. 2006). Of the six remaining
studies considered, four tested doses as low as 100 to 125 mg/kg-day (Table 4-5) (Gray et al.. 2021;
Hannas et al.. 2011; Howdeshell et al.. 2008; Saillenfait et al.. 2008). however, no studies evaluating
effects on the developing male reproductive system consistent with a disruption of androgen action have
been conducted with DIBP that have evaluated doses below 100 mg/kg-day. Available studies
considered for dose-response are discussed further below.
As discussed in Sections 3.1.2.1 and 3.1.2.2, oral exposure to DIBP can cause effects on the developing
male reproductive system consistent with a disruption of androgen action and other developmental
effects (i.e., decreased fetal weight, resorptions, post-implantation loss, skeletal variations). Effects on
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the developing male reproductive system are more sensitive than other observed developmental effects.
This is demonstrated by the fact that the lowest LOAELs for effects on the developing male
reproductive system range from 125 to 300 mg/kg-day, while the lowest LOAELs for other
developmental outcomes range from 500 to 600 mg/kg-day (Table 3-3, Table 4-5). Therefore, EPA's
dose-response assessment in this section focuses on effects on the developing male reproductive system
consistent with a disruption of androgen action.
Although single dose studies evaluating the effects of DIBP on the developing male reproductive system
are not available, studies of the toxicologically similar phthalate dibutyl phthalate (DBP) have
demonstrated that a single exposure during the critical window of development can disrupt expression of
steroidogenic genes and decrease fetal testes testosterone. Therefore, EPA considers effects on the
developing male reproductive system consistent with a disruption of androgen action to be relevant for
setting a POD for acute, intermediate, and chronic duration exposures (see Appendix B for further
discussion). Notably, SACC agreed with EPA's decision to consider effects on the developing male
reproductive system consistent with a disruption of androgen action to be relevant for setting a POD for
acute durations during the July 2024 peer-review meeting of the DINP human health hazard assessment
(U.S. EPA. 2024m). Studies considered for dose-response assessment are summarized in Table 4-5.
Of the 11 developmental toxicity studies considered for dose-response, two studies (BASF. 2007;
Saillenfait et al.. 2006) were not considered further for dose-response analysis because of limitations and
other factors that increase uncertainty. In Saillenfait et al. (2006). rats were exposed to doses of DIBP
ranging from 250 to 1000 mg/kg-day on GD 6 through 20 via gavage. Decreased fetal body weight and
increased incidence of cryptorchidism were observed at 500 mg/kg-day. Based on these effects, EPA
identified a NOAEL of 250 mg/kg-day. Similarly, BASF (2007) conducted a dietary study of pregnant
Wistar rats in which animals were exposed to 88 to 942 mg/kg-day of DIBP from GDs 6 through 20. A
NOAEL of 363 mg/kg-day was identified based on decreases in fetal body weight, maternal food
consumption, and maternal body weight gain at 942 mg/kg-day. However, the doses at which
developmental effects were observed in these studies were higher than doses at which more sensitive
effects of androgen insufficiency (e.g., decreased fetal testicular testosterone) were observed in other
studies. Therefore, EPA did not select these studies and endpoints because they do not provide the most
sensitive robust endpoint for an acute/intermediate/chronic POD.
Seven studies reported across five publications (Saillenfait et al.. 2017; Wang et al.. 2017; Furr et al..
2014; Hannas et al.. 2012; Borch et al.. 2006) that exposed pregnant mice or rats to DIBP via gavage
have observed effects on the developing male reproductive system. However, experiments in each of
these studies only tested one dose level in addition to vehicle controls, and support LOAELs ranging
from 200 to 750 mg/kg-day DIBP. These studies do not allow for the identification of a NOAEL, which
increases the uncertainty in the data set. Ultimately, these studies were not further considered because
other developmental studies of DIBP are available that test more than one dose level, including doses
less than 200 mg/kg-day and support identification of more sensitive NOAELs.
In contrast, three studies of pregnant SD rats provide consistent evidence of dose-related reductions in ex
vivo fetal testicular testosterone production and support NOAEL and LOAEL values of 100 and 300
mg/kg-day, respectively (Table 4-5) (Gray et al.. 2021; Hannas et al.. 2011; Howdeshell et al.. 2008).
Notably, the magnitude of effect on ex vivo fetal testicular testosterone production was consistent across
tested doses in all three studies when measured on GDI8. For example, the response compared to the
control ranged from 95 to 110 percent at 100 mg/kg-day and 44 to 66 percent at 300 mg/kg-day. Across
the three studies, there is consistent evidence of no effect on ex vivo fetal testicular testosterone
production in rats dosed with 100 mg/kg-day DIBP.
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In 2017, NASEM (2017) assessed experimental animal evidence for effects on fetal testicular
testosterone following in utero exposure to DIBP using the systematic review methodology developed
by the National Toxicology Program's (NTP) Office of Health Assessment and Translation (OHAT).
Based on results from two studies of rats (Hannas et al.. 2011; Howdeshell et al.. 2008). NASEM found
high confidence in the body of evidence and a high level of evidence that fetal exposure to DIBP is
associated with a reduction in fetal testosterone in rats. NASEM further conducted a meta-regression
analysis and benchmark dose (BMD) modeling analysis on decreased fetal testicular testosterone
production data from the same two prenatal exposure studies of rats (Hannas et al.. 2011; Howdeshell et
al.. 2008). 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 DIBP.
BMD analysis determined BMDLs and BMDL40 values of 23 and 225 mg/kg-day, respectively, the 95
percent lower confidence limits of the BMDs associated with a benchmark response (BMR) of 5 and 40
percent (Table 4-1).
Table 4-1. Summary of NASEM (2017) Meta-Analysis and BMD Modeling for Effects of DIBP in
Fetal Testosterone ab
Database
Supporting
Outcome
Confidence
in Evidence
Evidence of
Outcome
Heterogeneity
in Overall
Effect
Model with
Lowest AIC
BMDs mg/kg-
day (95% CI)
BMD40 mg/kg-
day (95% CI)
2 rat studies
High
High
I2 > 60%
Linear
27 (23, 34)'
270 (225, 340)
0 R code supporting NASEM's meta-regression and BMD analysis of DIBP is publicly available through GitHub
(https ://github. com/wachiuphd/NASEM-2017 -Endocrine-Low-Dose).
b NASEM (2017) calculated BMD40s for this endpoint because "previous studies have shown that reproductive-tract
malformations were seen in male rats when fetal testosterone production was reduced by about 40%."
c EPA noted an apparent discrepancy in the NASEM (2017) report. In Table 3-26, NASEM (2017) notes that no BMD/BMDL
estimates could be generated at the 5% response level for DIBP because "the 5% change was well below the range of the data,
but it will be 10 times lower because a linear model was usedHowever, in Table C6-12 of the NASEM (2017) report,
BMD/BMDL estimates at the 5% response level are provided for DIBP for the best-fit linear model. In EPA's replicate analysis,
which is provided in EPA's Draft Meta-Analysis and BMD of Fetal Testicular Testosterone for DEHP, DBP, BBP, DIBP, and
DCHP (U.S. EPA. 2024d). identical BMD/BMDL estimates for the 5% response level were obtained. Therefore, BMD/BMDL
estimates at the 5% response level for DIBP are reported in this table.
Since EPA identified new fetal testicular testosterone data (Gray et al.. 2021) for DIBP, an updated
meta-analysis was conducted. Using the publicly available R code provided by NASEM
(https://github.com/wachiuphd/NASEM-2017-Endocrine-Low-Dose). EPA applied the same meta-
analysis and BMD modeling approach used by NASEM, with the exception that the most recent Metafor
package available at the time of EPA's updated analysis was used (i.e., EPA used Metafor package
Version 4.6.0, whereas NASEM used Version 2.0.0), and an additional BMR of 10 percent was
modelled. Appendix D provides justification for the evaluated BMRs of 5, 10, and 40 percent. Fetal rat
testosterone data from three studies was included in the analysis (Gray et al.. 2021; Hannas et al.. 2011;
Howdeshell et al.. 2008). Overall, the meta-analysis found a statistically significant overall effect and
linear trends in logio(dose) and dose, with an overall effect that is large in magnitude (>50% change)
(Table 4-2). There was substantial, statistically significant heterogeneity in all cases (I2>60%). The
statistical significance of these effects was robust to leaving out individual studies. The linear-quadratic
model provided the best fit (based on lowest AIC) (Table 4-2). BMD estimates from the linear-quadratic
model were 270 mg/kg-day [95% confidence interval: 136, 517] for a 40 percent change (BMR = 40%>)
and 55 mg/kg-day [NA, 266] for a 10 percent change (BMR = 10%>), although a BMDL10 could not be
estimated (Table 4-3). No BMD could be estimated for a 5 percent change (BMR = 5%). Further
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methodological details and results (e.g., forest plots, figures of BMD model fits) for the updated meta-
analysis and BMD modeling of fetal testicular testosterone data are provided in the Draft Meta-Analysis
and Benchmark Dose Modeling of Fetal Testicular Testosterone for Di(2-ethylhexyl) Phthalate (DEHP),
Dibutyl Phthalate (DBP), Butyl Benzyl Phthalate (BBP), Diisobiityl Phthalate (DIBP), Dicyclohexyl
Phthalate (DCHP), andDiisononylPhthalate (U.S. EPA. 2024d).
Table 4-2. Overall Analyses of Rat Studies of DIBP and Fetal Testosterone (Updated Analysis
Conducted by EPA)
Analysis
Estimate
Beta
CI,
Lower
Bound
CI,
Upper
Bound
P
value
Tau
I2
P value for
Heterogeneity
AICs
Primary Analysis
Overall
intrcpt
-82.21
-122.85
-41.56
0.000
68.02
96.52
0.000
130.45
Trend in loglO(dose)
loglO(dose)
-165.55
-205.47
-125.64
0.000
19.89
65.48
0.004
106.31
Linear in doselOO
doselOO
-18.48
-25.14
-11.81
0.000
60.86
96.92
0.000
120.04
LinearQuadratic in doselOO
doselOO
-19.18
-41.21
2.85
0.088
48.79
94.49
0.000
111.51*
LinearQuadratic in doselOO
I(dosel00A2)
0.09
-2.70
2.88
0.950
48.79
94.49
0.000
111.51
Sensitivity Analysis
Overall minus Gray et al. 2021
intrcpt
-82.31
-135.11
-29.52
0.002
71.76
96.96
0.000
87.28
Overall minus Hannas et al.
2011b
intrcpt
-69.98
-110.63
-29.34
0.001
55.43
95.94
0.000
83.66
Overall minus Howdeshell et al.
2008
intrcpt
-94.90
-151.74
-38.06
0.001
78.38
94.86
0.000
88.36
* Indicates lowest AIC.
Table 4-3. Benchmark Dose Estimates for DIBP and Fetal Testosterone in Rats
Analysis
Benchmark
Response
(BMR)
Benchmark
Dose
(BMD)
Confidence
Interval, Lower
Bound
Confidence
Interval, Upper
Bound
Linear in doselOO
5%
28
20
43
Linear in doselOO
10%
57
42
89
Linear in doselOO
40%
276
203
432
LinearQuadratic in doselOO*
5%
NA
NA
207
LinearQuadratic in doselOO*
10%
55
NA
266
LinearQuadratic in doselOO*
40%
270
136
517
* Indicates model with lowest AIC.
'NA' indicates a BMD or BMDL estimate could not be derived.
Since no BMDLs could be derived through the updated meta-analysis and BMD modeling analysis, EPA
modelled individual fetal testicular testosterone data from the three studies included in the updated meta-
analysis using EPA's BMD Software (BMDS version 3.3.2) (Gray et al.. 2021; Hannas et al.. 2011;
Howdeshell et al.. 2008). This analysis included the full suite of standard continuous models
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(Exponential, Hill, Polynomial, Power, Linear), compared to the meta-analysis that only included the
linear and linear-quadratic models. Further methodological details and results from this BMD analysis
are provided in Appendix E. As can be seen from Table Apx E-l, no models adequately fit the fetal
testicular testosterone data from Hannas et al. (2011). In contrast, BMDs and BMDLs values of 63 and
24 mg/kg-day were derived from the fetal testicular testosterone data reported in Gray et al. (2021)
based on the best fitting exponential 3 model (constant variance), while BMDs and BMDLs values of
103 and 52 mg/kg-day were derived from the fetal testicular testosterone data reported in Howdeshell et
al. (2008) based on the best fitting hill model (constant variance).
Lastly, Saillenfait et al. (2008) reported the results of oral exposure to 0, 125, 250, 500, or 625 mg/kg-
day y DIBP on GD 12 through 21 on F1 male offspring. Treatment-related effects at 250 mg/kg-day
DIBP and above include decreased F1 male AGD on PND 1, increased male nipple retention on PND 12
to 14 and PNW 11 to 12 or PNW 16 to 17, while more severe reproductive tract malformations (e.g.,
hypospadias, exposed os penis, nonscrotal testes) were observed at 500 mg/kg-day DIBP and above. In
the low dose group (125 mg/kg-day), low incidence of testicular pathology was observed in F1 males
from PNW 11 to 12, including oligospermia (low sperm) (incidence: 0/24, 1/20, 3/28, 2/22, 1/20),
azoospermia (no sperm present) (0/24, 1/20, 3/28, 10/22, 18/20), and tubular degeneration, which
showed evidence of increasing severity with dose. However, the study is limited due to a lack of
statistical analysis on the testicular pathology data and due to the small sample size (only two F1 males
were examined per litter). Although the incidence of testicular pathology at 125 mg/kg-day is low, EPA
considers the study to support a LOAEL of 125 mg/kg-day (no NOAEL identified) due to the severity of
the observed effects (i.e., reduced and/or absence of sperm in 2/20 adult F1 males). EPA considered
BMD modeling of data from Saillenfait et al. (2008). However, BMD modeling of data from Saillenfait
et al. (2008) has previously been published by EPA's Office of Research and Development (Blessmger
et al.. 2020). As can be seen from Table 4-4, the BMDs and BMDLs values for the more sensitive
outcomes evaluated by Saillenfait et al. (i.e., combined azoospermia and oligospermia) fall outside of
the range of measured tested doses.
Table 4-4. Summary of Dichotomous BMD Analysis of Data from Saillenfait et al. (2008) by
Blessinger et al. (2020)"
Endpoint
BMR
BMD
(mg/kg-day)
BMDL
(mg/kg-day)
Hypospadias
1% extra risk
401
242
Undescended testes
1% extra risk
342
194
Exposed os penis
1% extra risk
361
112
Areola or nipple retention
5% extra risk
317
205
Azoospermia or grade 2-5 oligospermia
5% extra risk
117
60
Tubular degeneration
5% extra risk
480
266
Sloughed cells
5% extra risk
112
56
11 Adapted from Table 6 in Blessinger et al. (2020). See Blessinger et al. for a description of the BMD
modeling approach. BMD modeling outputs from Blessinger et al. are available at:
httos://doi.ors/10.23719/1503702.
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4.2.2 Options Considered by EPA for Deriving the Acute Non-Cancer POD
In order to derive a non-cancer POD for DIBP, EPA considered three options, including:
Option 1. NOAEL/LOAEL approach to identify the highest NOAEL below the lowest LOAEL
(Section 4.2.2.1).
Option 2. Application of a data-derived adjustment factor based on differences in relative
potency to reduced fetal testicular testosterone (Section4.2.2.2).
Option 3. BMD modeling of fetal testicular testosterone (Section 4.2.2.3).
The strengths and limitations of each of the approaches considered by EPA to derive a non-cancer POD
for DIBP are discussed further below, while the POD EPA selected for the draft risk evaluation of DIBP
is discussed in Section 4.2.3.
4.2.2.1 Option 1. NOAEL/LOAEL Approach
Overall, EPA considers Saillenfait et al. (2008) to support a LOAEL of 125 mg/kg-day based on low
incidence of testicular histopathological findings. Three additional studies of fetal testicular testosterone
all support a NOAEL of 100 mg/kg-day (Gray et al.. 2021; Hannas et al.. 2011; Howdeshell et al..
2008). Each of these three studies gavaged pregnant SD rats with the same DIBP doses (0, 100, 300,
600, 900 mg/kg-day) on GDs 14-18 (Gray et al.. 2021; Hannas et al.. 2011) or GDs 8-18 (Howdeshell et
al.. 2008). For each of the three studies, ex vivo fetal testicular testosterone production was then
measured on GD 18, approximately 2 hours after the final dose of DIBP was administered. Results from
these studies did not observe any significant changes in ex vivo fetal testicular testosterone production at
100 mg/kg-day when measured on GDI8; however, at the 300 mg/kg-day DIBP dose, the response
compared to the control ranged from 44 to 66 percent, supporting a NOAEL of 100 mg/kg-day in these
studies (Gray et al.. 2021; Hannas et al.. 2011; Howdeshell et al.. 2008). Therefore, EPA considers these
4 studies to support a NOAEL for fetal testicular testosterone of 100 mg/kg-day and a LOAEL for
testicular histopathology of 125 mg/kg-day.
However, there are several lines of evidence that suggest a NOAEL of 100 mg/kg-day may be under-
protective, including:
The database of studies for DIBP is limited to 11 gestational or perinatal oral exposures studies,
5 of which tested a single high dose level of 200 to 750 mg/kg-day, while no studies have
evaluated doses below 100 mg/kg-day.
BMD modeling of testicular pathology data from Saillenfait et al. (2008) supports BMDLs
values of 56 to 60 mg/kg-day based on incidence of sloughed cells or combined
azoospermia/oligospermia (Table 4-4).
EPA's updated meta-analysis and BMD modeling analysis of fetal testicular testosterone
supported a BMDio of 55 mg/kg-day. No BMDLs could not be derived from the best-fitting
linear quadratic model as part of the updated analysis (Table 4-3).
4.2.2.2 Option 2. Application of a Data-Derived Adjustment Factor
EPA also considered differences in relative potency between toxicologically similar phthalates to derive
a data-derived adjustment factor. As discussed in EPA's Draft Technical Support Document for the
Cumulative Risk Analysis of Di(2-ethylhexyl) Phthalate (DEHP), Dibutyl Phthalate (DBP), Butyl Benzyl
Phthalate (BBP), Diisobutyl Phthalate (DIBP), Dicyclohexyl Phthalate (DCHP), andDiisononyl
Phthalate (DINP) Under the Toxic Substances Control Act (TSCA) (U.S. EPA. 20241). EPA has derived
a draft relative potency factor (RPF) of 0.53 for DIBP, based on its relative potency compared to the
index chemical, dibutyl phthalate (DBP), at reducing fetal testicular testosterone. The draft POD for the
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index chemical, DBP, is a BMDLs of 9 mg/kg-day derived from EPA's updated meta-analysis and BMD
modeling analysis of fetal testicular testosterone (U.S. EPA. 2024f 1). The draft POD of 9 mg/mg-day
for the index chemical (DBP) is approximately 11.1 times lower than the NOAEL of 100 mg/kg-day
identified for DIBP identified above in Section 4.2.2.1. In contrast, the draft RPF of 0.53 indicates that
the POD for DIBP should be approximately twice that of DBP, since DIBP is approximately half as
potent as DBP as reducing fetal testicular testosterone. Therefore, EPA considered adjusting the DIBP
NOAEL of 100 by a factor of 5.89 (i.e., (DIBP NOAEL DBP BMDLs) * RPFdibp), which would result
in an adjusted NOAEL of 17 mg/kg-day.
Notably, ECHA (2017a. b) employed a similar relative potency adjustment for DIBP. When deriving a
POD for DIBP for use in risk characterization, ECHA (2017a. b) stated:
"Few reproductive toxicity studies have been published on [DIBP] compared to DEHP and
DBP. No two-generation studies are available and the substance has not been tested at doses
<100 mgkg bw/d. Current data suggest that DIBP coirfd have similar effects to DBP, if studied
at lower dose levels. If the potency difference between DIBP and DBP, as a very rough
estimate of the observed effects in Saillenfait et al. (2008) (type of effects seen at 500 and 625
mg kg bw-day, corresponding to a difference of 25%), is extrapolatedfrom the high dose area
to the lower dose area, an estimatedLOAEL for DIBP would be 25% higher than the current
LOAEL for DBP (2 mg kg bw-day). Available information is shown in Table B7. A LOAEL for
DIBP of 2.5 mg kg bw-day is selectedfor use in the current combined risk assessment "
4.2.2.3 Option 3. BMD Analysis of Individual Fetal Testicular Testosterone Studies
Because no BMDLs could be derived via the updated meta-analysis and BMD analysis of fetal testicular
testosterone data, EPA modelled individual ex vivo fetal testicular testosterone production data sets
using EPA's BMD Software (Gray et al.. 2021; Hannas et al.. 2011; Howdeshell et al.. 2008). No
models adequately fit the Hannas et al. (2011) data set (TableApx E-l). In contrast, BMDs and BMDLs
values of 63 and 24 mg/kg-day were derived from the Gray et al. (2021) data set based on the best fitting
exponential 3 model, while BMDs and BMDLs values of 103 and 52 mg/kg-day were derived from the
Howdeshell et al. (2008) data set based on the best fitting Hill model (Table Apx E-l).The BMDLs of
52 mg/kg-day from Howdeshell et al. (2008) is similar to the derived BMDio of 55 mg/kg-day from
EPA's updated meta-analysis (Table 4-3) suggesting the BMDLs of 52 mg/kg-day is not appropriate for
use in human health risk characterization. Additionally, although the linear model in EPA's updated
meta-analysis did not provide the best-fit (i.e., the linear-quadratic model had a lower AIC), the linear
model did appear to adequately fit the data set and supports BMDs and BMDLs values of 28 and 20
mg/kg-day (Table 4-3). The BMDLs of 24 mg/kg-day from Gray et al. (2021) is similar to the BMDLs
of 20 mg/kg-day derived using the linear model in the updated meta-analysis. Although there is some
uncertainty because derived BMDLs estimates are below the lowest dose with empirical data (i.e., 100
mg/kg-day), EPA considers this BMD analysis to support a BMDLs of 24 mg/kg-day based on reduced
fetal testicular testosterone in the study by Gray et al. (2021).
4.2.3 POD Selected for Acute, Intermediate, and Chronic Durations
Considering the three options described above in Section 4.2.2,, EPA selected the BMDLs of 24 mg/kg-
day (option 3) based on reduced fetal testicular testosterone from the study by Gray et al. (2021). EPA
considered the POD derived from the BMD analysis of data in this study to have the least uncertainty
and highest confidence upon examination of the weight of evidence provided by the three options. This
POD is more sensitive than the NOAEL of 100 mg/kg-day (option 1), which is likely under-protective
due to the limited number of studies and lack of testing at doses lower than 100 mg/kg-day. EPA
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considers the BMDLs of 24 mg/kg-day (option 3) to be more appropriate than using the data-derived
adjustment factor based on relative potency to the index chemical, DBP, (Option 2) because the BMDLs
of 24 mg/kg-day relies on adequately modeled fetal testosterone production data specific to DIBP.
While the application of relative potency is appropriate and necessary for the cumulative risk assessment
across phthalates, the use of data exclusive to DIBP has less uncertainty for the individual risk from
DIBP. Using allometric body weight scaling to the three-quarters power, (U.S. EPA. 2011c). EPA
extrapolated an HED of 5.7 mg/kg-day from the BMDLs of 24 mg/kg-day. A total uncertainty factor of
30 was selected for use as the benchmark margin of exposure (based on an interspecies uncertainty
factor (UFa) of 3 and an intraspecies uncertainty factor (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 C).
EPA considered reducing the UFa further to a value of 1 based on apparent differences in
toxicodynamics between rats and humans. As discussed in Section 3.1.4 of EPA's Draft Proposed
Approach for Cumulative Risk Assessment of High-Priority Phthalates and a Manufacturer-Requested
Phthalate under the Toxic Substances Control Act (U.S. EPA. 2023 a). several explant (Lamb rot et al..
2009; Hallmark et al.. 2007) and xenograft studies (van Den Driesche et al.. 2015; Spade et al.. 2014;
Heger et al.. 2012; Mitchell et al.. 2012) using human donor fetal testis tissue have been conducted to
investigate the antiandrogenicity of mono-2-ethylhexyl phthalate (MEHP; a monoester metabolite of
DEHP), DBP, and monobutyl phthalate (MBP; a monoester metabolite of DBP) in a human model.
Generally, results from human explant and xenograft studies suggest that human fetal testes are less
sensitive than rat testes to the antiandrogenic effects of phthalates, however, effects on Sertoli cells and
increased incidence of MNGs have been observed in two human xenograft studies of DBP (van Den
Driesche et al.. 2015; Spade et al.. 2014; Heger et al.. 2012; Mitchell et al.. 2012). As discussed in
EPA's draft approach document (U.S. EPA. 2023a). the available human explant and xenograft studies
have limitations and uncertainties, which preclude definitive conclusions related to species differences
in sensitivity. For example, key limitations and uncertainties of the human explant and xenograft studies
include: small sample size; human testis tissue was collected from donors of variable age and by
variable non-standardized methods; and most of the testis tissue was taken from fetuses older than 14
weeks, which is outside of the critical window of development (i.e., gestational weeks 8 to 14 in
humans). Therefore, EPA did not reduce the UFa.
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1336 Table 4-5. Dose-Response Analysis of Selected Studies Considered for Acute, Intermediate, and Chronic Exposure Scenarios
Study Details
(Species, Duration, Exposure
Route/ Method, Doses [mg/kg-
day])
Study POD/
Type
(mg/kg-day)
Effect
HED
(mg/kg)
Uncertainty
Factors"b c
Reference(s)
Sprague-Dawley rats; GD 14-18;
oral/gavage; 0, 100, 300, 600, 900
BMDLs = 24
I ex vivo testicular testosterone production
(34%)
5.7
UFa = 3
UFh = 10
Total UF = 30
(Grav et al.. 2021)
Sprague-Dawley rats; GD 8-18;
0, 100, 300, 600, 900
BMDLs = 52
I ex vivo testicular testosterone production
(40%)
12.3
UFa = 3
UFh = 10
Total UF = 30
(Howdeshell et al..
2008)
Sprague-Dawley rats; GD 14-18;
oral/gavage; 0, 100, 300, 600, 900
NOAEL =
100
I ex vivo fetal testicular testosterone production
(56%); I expression of steroidogenic genes in
fetal testes
23.6
UFa = 3
UFh = 10
Total UF = 30
(Hannas et al.. 2011)
Sprague-Dawley rats; GD 12-21;
oral/gavage; 0, 125, 250, 500, 625
LOAEL =
125
Testicular pathology (degeneration of
seminiferous tubules and oligo-/azoospennia in
epididymis)
29.6
UFa = 3
UFh = 10
UFl = 10
Total UF = 300
(Saillenfait et al.. 2008)
Sprague-Dawley rats; GD 14-18;
oral/gavage; 0, 200 (Block 30)
LOAEL =
200
I ex vivo fetal testicular testosterone production
47.3
UFa = 3
UFh = 10
UFl = 10
Total UF = 300
(Furr et al.. 2014)
Sprague-Dawley rats; GD 6-20;
oral/gavage; 0, 250, 500, 750,
1000
NOAEL =
250
I fetal body weight (both sexes); t incidence of
cryptorchidism
59.1
UFa = 3
UFh = 10
Total UF = 30
(Saillenfait et al.. 2006)
Sprague-Dawley rats; GD 13-19;
oral/gavage; 0, 250
LOAEL =
250
J.AGD, I testicular testosterone &
androstenedione production, altered mRNA
expression of steroidogenesis genes in the testes
59.1
UFa = 3
UFh = 10
UFl = 10
Total UF = 300
(Saillenfait et al.. 2017)
ICR Mice; GD 0-21; oral/gavage;
0, 450
LOAEL =
450
I absolute testes weight on PND 21; J. serum
and testes testosterone; J. expression of
steroidogenic genes in testes; |sperm
concentration and motility on PND 80
59.8
UFa = 3
UFh = 10
UFl = 10
Total UF = 300
(Wane etal..2017)
Wistar Rat; oral/diet; 0, 88, 363,
942
NOAEL =
363
I maternal food consumption, [ maternal body
weight gain, [ fetal body weight
85.8
UFa = 3
UFh = 10
(BASF. 2007)
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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)
Uncertainty
Factors"b c
Reference(s)
Total UF = 30
Sprague-Dawley rats; GD 14-18;
oral/gavage; 0, 500 (Block 14)
LOAEL =
500
I ex vivo fetal testicular testosterone production
118
UFa = 3
UFh = 10
UFl = 10
Total UF = 300
(Furr et al.. 2014)
Sprague-Dawley rats; GD 14-18;
oral/gavage; 0, 500
LOAEL =
500
I ex vivo fetal testicular testosterone production
118
UFa = 3
UFh = 10
UFl = 10
Total UF = 300
(Hannas et al.. 2012)
Wistar Rat; GD 7-19 or 7-20/21;
oral/gavage; 0, 600
LOAEL =
600
I testes testosterone, [ AGD, | testicular
histopathology
142
UFa = 3
UFh = 10
UFl = 10
Total UF = 300
(Borch et al.. 2006)
Sprague-Dawley rats; GD 14-18;
oral/gavage; 0, 750 (Block 2)
LOAEL =
750
I ex vivo fetal testicular testosterone production
177
UFa = 3
UFh = 10
UFl = 10
Total UF = 300
(Furr et al.. 2014)
Abbreviations: j = statistically significant decrease; t = statistically significant increase; POD = Point of Departure; HED = Human equivalent Dose; UF =
uncertainty factor; NOAEL = No observed adverse effect level; LOAEL = Lowest observed adverse effect level; GD = Gestational Day; PND = Postnatal Day
AGD = Anogenital distance; BMD = benchmark dose.
"EPA used allometric bodv weieht scalins to the three-auarters rower to derive the HED. Consistent with EPA Guidance (U.S. EPA. 201 lc). the interspecies
uncertainty factor (UFA), was reduced from 10 to 3 to account remaining uncertainty associated with interspecies differences in toxicodynamics.
b 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 DIBP.
c 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.
J Two studies with similar desisns were included in the meta-analvsis bv NASEM (2017). each of which exrosed Suraeue-Dawlev rats (< 3 dams/dose) to 0.
100, 300, 600, 900 mg/kg-day DIBP during the masculinization programming window during gestational development.
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4.3 Weight of The Scientific Evidence Conclusion: POD for Acute,
Intermediate, and Chronic Durations
EPA considered BMD modelling from the study by Gray et al. to support a BMDLs of 24 mg/kg-day
(Gray et al.. 2021). EPA has preliminarily concluded that the HED of 5.7 mg/kg-day (BMDLs of 24
mg/kg-day) based on decreased fetal testicular testosterone production from the gestational exposure
study of rats by Gray et al. is appropriate for calculation of risk from acute, intermediate, and chronic
durations. A total uncertainty factor of 30 was selected for use as the benchmark margin of exposure
(based on an interspecies uncertainty factor (UFa) of 3 and an intraspecies uncertainty factor (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 C).
Given the limited database of studies for DIBP that have evaluated outcomes other that developmental
toxicity and effects on the developing male reproductive system. For toxicologically similar phthalates
(i.e., DEHP, DBP, BBP, DCHP), which include larger databases of animal toxicology studies including
numerous well-conducted subchronic and chronic toxicity studies, effects on the developing male
reproductive system consistent with a disruption of androgen action have consistently been identified by
EPA as the most sensitive and well-characterized hazard in experimental animal models. This is
demonstrated by the fact that the preliminary acute/intermediate/chronic PODs selected by EPA for use
in risk characterization for DEHP (U.S. EPA. 2024h). DBP (U.S. EPA. 2024ft. BBP (U.S. EPA. 2024e).
DCHP (U.S. EPA. 2024g) are all based on effects related to phthalate syndrome. EPA has robust
overall confidence in the selected POD based on the following weight of the scientific evidence:
EPA has previously considered the weight of scientific evidence and concluded that oral
exposure to DIBP 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-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).
DIBP 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 fetal testicular testosterone
content and/or testosterone production; reduced male pup anogenital distance; male pup nipple
retention; reproductive tract malformations (i.e., hypospadias, undescended testes, exposed os
penis, cleft prepuce); delayed preputial separation; testicular pathology (e.g., degeneration of
seminiferous tubules, oligospermia, azoospermia, Leydig cell aggregation, Sertoli cell
vacuolation, multinucleated gonocytes); decreased sperm concentration and motility.
The selected POD is a BMDLs based on reduced ex vivo fetal testicular testosterone production
in one gestational exposure studies of rats (Gray et al.. 2021).
Consistently, other regulatory and authoritative bodies have also concluded that DIBP induces
effects on the developing male reproductive system consistent with a disruption of androgen
action and phthalate syndrome and that these effects are relevant for estimating human risk
(ECCC/HC. 2020: ECHA. 2017a. b; U.S. CPSC. 2014: ECHA. 2012a. b; NICNAS. 2008a).
EPA considers effects on the developing male reproductive system consistent with a disruption
of androgen action to be relevant for setting a POD for acute, intermediate, and chronic duration
exposures, based on studies of the toxicologically similar phthalate dibutyl phthalate (DBP)
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which have demonstrated that a single exposure during the critical window of development in
rats can disrupt expression of steroidogenic genes and decrease fetal testes testosterone
production.
EPA did not identify any studies conducted via the dermal route relevant for extrapolating human health
risk. Therefore, EPA is using the oral HED of 24 mg/kg to extrapolate to the dermal route. EPA's
approach to dermal absorption for workers, consumers, and the general population is described in EPA's
Draft Environmental Release and Occupational Exposure Assessment for Diisobiityl phthalate (U.S.
EPA. 2025c).
EPA did not identify any inhalation studies of DIBP. Therefore, EPA is also using the oral HED of 24
mg/kg 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 C 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 likely to be more susceptible to DIBP 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 DIBP.
Although ample human epidemiologic data are available on health effects of DIBP (see Section 3.1.1),
EPA was unable to identify direct evidence of differences in susceptibility among human populations.
Animal studies demonstrating effects on male reproductive development and other developmental
outcomes provide direct evidence that gestation is a particularly sensitive lifestage. Evidence from
animal studies also suggests that the liver may also be a target organ; however, there is not enough
evidence to reliably inform specific health outcomes or to be used in risk quantification. Therefore, EPA
is quantifying risks including those for PESS based on reproductive and developmental toxicity in the
draft DIBP risk evaluation.
As summarized in Table 5-1, EPA identified a range of factors that may have the potential to increase
biological susceptibility to DIBP, including lifestage, pre-existing diseases, physical activity, nutritional
status, stress, and co-exposures to other environmental stressors that contribute to related health
outcomes. The effect of these factors on susceptibility to health effects of DIBP is not known; therefore,
EPA is uncertain about the directions and magnitude of any possible increased risk from effects
associated with DIBP exposure for relevant subpopulations.
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 DIBP. The Risk Assessment Forum, in
A Review of the Reference Dose and Reference Concentration Processes (U.S. EPA. 2002b). 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).
Although U.S. EPA (2002b) did not discuss all the factors presented in Table 5-1, EPA considers the
POD selected for use in characterizing risk from exposure to DIBP to be protective of effects on the
developing male reproductive system consistent with phthalate syndrome in humans.
As discussed in U.S. EPA (2023a). exposure to DIBP and other toxicologically similar phthalates (i.e.,
DEHP, DBP, BBP, DCHP, DINP) that disrupt androgen action during the development of the male
reproductive system cause dose additive effects. Cumulative effects from exposure to DIBP and other
toxicologically similar phthalates will be evaluated as part of U.S. EPA's cumulative risk assessment of
phthalates.
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Table 5-1. PESS Evidence Crosswalk for Biological Susceptibility Considerations
Susceptibility
Category
Examples of
Specific
Factors
Direct Evidence this Factor
Modifies Susceptibility to DIBP
Description of Interaction
Key Citations
Indirect Evidence of Interaction with
Target Organs or Biological Pathways
Relevant to DIBP
Description of
Interaction
Key Citation(s)
Susceptibility Addressed in Risk
Evaluation?
Embryos/
fetuses/infants
Lifestage
Direct quantitative animal
evidence for developmental
toxicity (e.g., increased skeletal
variations, decreased fetal body
weight, increased resorptions,
and post-implantation loss).
There is direct quantitative
animal evidence for effects on
the developing male
reproductive system consistent
with a disruption of androgen
action.
(U.S. EPA.
2023a)
(U.S. EPA.
2023b)
(Howdeshell et
al 2008)
(Hannas et al
2011)
(Wang et al..
2017)
(Saillenfait et al..
2008)
(BASF. 2007)
(Borch et al..
2006)
(Saillenfait et al..
2006)
Pregnancy/
lactating status
Rodent dams not particularly
susceptible during pregnancy
and lactation, except for effects
related to reduced maternal
weight gain and food
consumption evident only at
high concentrations.
(Howdeshell et
al.. 2008)
(Saillenfait et al..
2006)
(BASF. 2007)
POD selected for assessing risks
from acute, intermediate, and
chronic exposures to DIBP is
based on developmental toxicity
(i.e., reduced fetal testicular
testosterone production) and is
protective of effects on the fetus
and offspring.
POD selected for assessing risks
from acute, intermediate, and
chronic exposures to DIBP based
on developmental toxicity (i.e.,
reduced fetal testicular
testosterone production) is
protective of effects on dams
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Susceptibility
Category
Examples of
Specific
Direct Evidence this Factor
Modifies Susceptibility to DIBP
Indirect Evidence of Interaction with
Target Organs or Biological Pathways
Relevant to DIBP
Susceptibility Addressed in Risk
Evaluation?
Factors
Description of Interaction
Key Citations
Description of
Interaction
Key Citation(s)
Males of
reproductive
age
Consistent evidence of effects
on endpoints related to male
reproductive development in
rats and mice, including
steroidogenesis in the testes and
effects on sperm (i.e., decreased
concentration and motility,
increased malformation).
(Pan etal.. 2017)
POD selected for assessing risks
from acute, intermediate, and
chronic exposures to DIBP is
based on effects on male
reproductive development (i.e.,
reduced fetal testicular
testosterone production) is
expected to be protective of adult
male reproductive effects.
Children
Reduced rodent offspring body
weight gain between PNDs 1 to
21 was observed in three
gestational exposure studies.
(Saillenfait et al..
2008)
(Wans et al..
2017)
(BASF. 2007)
POD selected for assessing risks
from acute, intermediate, and
chronic exposures to DIBP based
on developmental toxicity (i.e.,
reduced fetal testicular
testosterone production) is
expected to be protective of
effects of offspring bodyweight
gain.
Use of default lOx UFH
Elderly
No direct evidence identified
Use of default lOx UFH
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Susceptibility
Category
Examples of
Specific
Direct Evidence this Factor
Modifies Susceptibility to DIBP
Indirect Evidence of Interaction with
Target Organs or Biological Pathways
Relevant to DIBP
Susceptibility Addressed in Risk
Evaluation?
Factors
Description of Interaction
Key Citations
Description of
Interaction
Key Citation(s)
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).
CDC (2023e)
CDC (2023 g)
Use of default lOx UFH
Pre-existing
disease or
disorder
Individuals with chronic
liver disease may be more
susceptible to effects on
these target organs.
Viruses such as viral
hepatitis can cause liver
damage.
Toxicokinetics
No direct evidence identified
Chronic liver disease is
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
DIBP.
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)
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Susceptibility
Category
Examples of
Specific
Direct Evidence this Factor
Modifies Susceptibility to DIBP
Indirect Evidence of Interaction with
Target Organs or Biological Pathways
Relevant to DIBP
Susceptibility Addressed in Risk
Evaluation?
Factors
Description of Interaction
Key Citations
Description of
Interaction
Key Citation(s)
Alcohol
consumption
No direct evidence identified
Alcohol use during
pregnancy can cause
adverse developmental
outcomes (e.g., fetal
alcohol spectrum
disorders).
Heavy alcohol use may
affect susceptibility to
liver disease.
CDC (2023d)
CDC (2023a)
Physical
activity
No direct evidence identified
Insufficient activity may
increase susceptibility to
multiple health outcomes.
Overly strenuous activity
may also increase
susceptibility.
CDC (2022)
Sociodemo-
graphic status
Race/ethnicity
No direct evidence identified
(e.g., no information on
polymorphisms in DIBP
metabolic pathways or diseases
associated race/ethnicity that
would lead to increased
susceptibility to effects of DIBP
by any individual group).
Socioeconomic
status
No direct evidence identified
Individuals with lower
incomes may have worse
health outcomes due to
social needs that are not
met, enviromnental
concerns, and barriers to
health care access.
ODPHP (2023b)
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Susceptibility
Category
Examples of
Specific
Factors
Direct Evidence this Factor
Modifies Susceptibility to DIBP
Indirect Evidence of Interaction with
Target Organs or Biological Pathways
Relevant to DIBP
Susceptibility Addressed in Risk
Evaluation?
Description of Interaction
Key Citations
Description of
Interaction
Key Citation(s)
Sex/gender
Male reproductive development
is a sex-specific endpoint and
consistent evidence indicates it
is the most sensitive effect
following gestational or early
life DIBP exposure.
See discussion in
Section 3.1.2.1.
POD selected for assessing risks
from acute, intermediate, and
chronic exposures to DIBP is
based on effects on male
reproductive development (i.e.,
reduced fetal testicular
testosterone production)
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
DIBP-induced liver
toxicity.
CDC (2023e)
CDC (2023b)
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Susceptibility
Category
Examples of
Specific
Direct Evidence this Factor
Modifies Susceptibility to DIBP
Indirect Evidence of Interaction with
Target Organs or Biological Pathways
Relevant to DIBP
Susceptibility Addressed in Risk
Evaluation?
Factors
Description of Interaction
Key Citations
Description of
Interaction
Key Citation(s)
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
DIBP.
CDC (2021)
CDC (2023b)
Target organs
No direct evidence identified
Polymorphisms in genes
may increase
susceptibility to liver or
developmental toxicity.
Use of default lOx UFH
Genetics/
epigenetics
Toxicokinetics
No direct evidence identified
Polymorphisms in genes
encoding enzymes (e.g.,
esterases) involved in
metabolism of DIBP may
influence metabolism and
excretion of DIBP.
Use of default lOx UFH
Other
chemical and
nonchemical
stressors
Built
enviromnent
No direct evidence identified
Poor-quality housing is
associated with a variety
of negative health
outcomes.
ODPHP (2023a)
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Susceptibility
Category
Examples of
Specific
Direct Evidence this Factor
Modifies Susceptibility to DIBP
Indirect Evidence of Interaction with
Target Organs or Biological Pathways
Relevant to DIBP
Susceptibility Addressed in Risk
Evaluation?
Factors
Description of Interaction
Key Citations
Description of
Interaction
Key Citation(s)
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)
Chemical co-
exposures
Studies have demonstrated that
co-exposure to DIBP and other
toxicologically similar
phthalates (e.g., DEHP, DBP,
BBP, DCHP, DINP) 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 (TJ.S. EPA.
2023a) and (U.S.
EPA 2023b)
Co-exposures will be
quantitatively addressed as part of
the phthalate cumulative risk
assessment and are not addressed
in the individual DIBP
assessment.
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6 POINTS OF DEPARTURE USED TO ESTIMATE RISKS FROM
DIBP EXPOSURE, CONCLUSOINS, AND NEXT STEPS
After considering hazard identification and evidence integration, dose-response evaluation, and weight
of the scientific evidence of POD candidates, EPA chose one non-cancer endpoint for use in determining
the risk from acute, intermediate, and chronic exposure scenarios (see Table ES-1). The critical effect is
disruption to androgen action during the critical window of male reproductive development (i.e., during
gestation), leading to a spectrum of effects on the developing male reproductive system consistent with
phthalate syndrome. Decreased fetal testicular testosterone was selected as the basis for the POD of 24
mg/kg-day (HED = 5.7 mg/kg-day) for acute, intermediate, and chronic durations. EPA has robust
overall confidence in the selected POD for acute, intermediate, and chronic durations. There are no
studies conducted via the dermal and inhalation route relevant for extrapolating human health risk. In the
absence of inhalation studies, EPA performed route-to-route extrapolation to convert the oral HED to an
inhalation human equivalent concentration (HEC) of 30.9 mg/m3 (2.71 ppm). EPA is also using the oral
HED to extrapolate to the dermal route. HECs are based on daily continuous (24-hour) exposure, and
HEDs are daily values.
The POD of 24 mg/kg-day (HED = 5.7 mg/kg-day) will be used in the Draft Risk Evaluation for DIBP
(U.S. EPA. 2024i) to estimate acute, intermediate, and chronic non-cancer risk. EPA summarizes the
cancer hazards of DIBP in a separate technical support document, Draft Cancer Raman Health Hazard
Assessment for Di(2-ethylhexyl) Phthalate (DEHP), Dibutyl Phthalate (DBP), Butyl Benzyl Phthalate
(BBP) andDicyclohexylPhthalate (DCHP) (U.S. EPA. 2025a).
EPA is soliciting comments from the Science Advisory Committee on Chemicals (SACC) and the public
on the non-cancer hazard identification, dose-response and weight of evidence analyses, and the selected
POD for use in risk characterization of DIBP.
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APPENDICES
Appendix A Existing Assessments of DIBP
The available existing assessments of DIBP are summarized in Table Apx A-l, which includes details regarding external peer-review, public
consultation, and systematic review protocols that were used.
Table Apx A-l. Summary of Peer-review, Public Comments, and Systematic Review for Existing Assessments of DIBP
Agency
Assessment(s) (Reference)
External
Peer-
Review?
Public
Consultation?
Systematic
Review
Protocol
Employed?
Remarks
U.S. EPA
(Publications by
the Center for
Public Health
and
Environmental
Assessment
[CPHEA] within
the Office of
Research and
Development
[ORD])
Phthalate exposure and male
reproductive outcomes: A
systematic review of the human
epidemiological evidence (Radke
et al.. 2018)
No
No
Yes
Phthalate exposure andfemale
reproductive and developmental
outcomes: A systematic review of
the human epidemiological
evidence (Radke et al.. 2019b)
Phthalate exposure and metabolic
effects: A systematic review of the
human epidemiological evidence
(Radke et al.. 2019a)
Phthalate exposure and
neurodevelopment: A systematic
review and meta-analysis of
human epidemiological evidence
(Radke et al.. 2020a).
Hazards of diisobutylphthalate
(DIBP Exposure): A systematic
- Publications were subjected to peer-
review prior to being published in a
special issue of the journal 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.
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Agency
Assessment(s) (Reference)
External
Peer-
Review?
Public
Consultation?
Systematic
Review
Protocol
Employed?
Remarks
review of animal toxicology
studies (Yost et al., 2019)
U.S. CPSC
Toxicity review of diisobutyl
phthalate (DiBP, CASRN 84-69-
5) (U.S. CPSC. 2011)
Yes
Yes
No
- Peer-reviewed by panel of four experts.
Peer review report available at:
httDs://www.cDsc.sov/s3fs-Dublic/Peer-
Chronic Hazard Advisory Panel
on Phthalates and Phthalate
Alternatives (U.S. CPSC, 2014)
Review-Report-Comments.pdf
-Public comments available at:
https://www.cpsc.sov/chap
- No formal systematic review protocol
employed.
- Details regarding CPSC's strategy for
identifying new information and literature
are provided on case 12 of (U.S. CPSC,
2014)
NASEM
Application of systematic review
methods in an overall strategy for
evaluating low-dose toxicity fi'om
endocrine active chemicals
mASEM. 2017)
Yes
No
Yes
- Draft report was reviewed by individuals
chosen for their diverse perspectives and
technical expertise in accordance 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:
Medium-chain phthalate esters:
Chemical Abstracts Service
Registry Numbers: 84-61-7; 84-
Yes
Yes
No (Animal
studies)
- Ecological and human health portions of
the screening assessment report
(ECCC/HC, 2020) were subiect to
external review and/or consultation. See
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Agency
Assessment(s) (Reference)
External
Peer-
Review?
Public
Consultation?
Systematic
Review
Protocol
Employed?
Remarks
64-0; 84-69-5; 523-31-9; 5334-
09-8; 16883-83-3; 27215-22-1;
27987-25-3; 68515-40-2; 71888-
89-6 (EC/HC. 2015b)
Supporting documentation:
Evaluation of epidemiologic
studies on phthalate compounds
and their metabolites for
hormonal effects, growth and
development and reproductive
parameters (Health Canada,
Yes
(Epidemiologic
studies)
case 2 of (ECCC/HC, 2020) for additional
details.
- State of the science report (EC/HC,
2015 a) and draft screening assessment
report for the phthalate substance group
subjected to 60-day public comment
periods. Summaries of received public
comments available at:
httos ://www. Canada, ca/ en/heal th-
canada/ servi ce s/chemi cal -
substances/substance-srouDinss-
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)
Screening Assessment - Phthalate
Substance Grouping (ECCC/HC,
2020)
initi ative/phthal ate. html#a 1
- 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 provided in Section 1 of
(EC/HC. 2015a) and (ECCC/HC. 2020)
- Human epidemiologic studies evaluated
usins Downs and Black Method (Health
Canada, 2018a, b)
NICNAS
Existing chemical hazard
assessment report: Diisobiityl
phthalate (NICNAS, 2008a)
No
Yes
No
- No details regarding peer-review are
provided.
- No formal systematic review protocol
employed.
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Agency
Assessment(s) (Reference)
External
Peer-
Review?
Public
Consultation?
Systematic
Review
Protocol
Employed?
Remarks
- No details regarding how NICNAS
identified literature for inclusion in its
assessment are provided.
ECHA
Committee for Risk Assessment
(RAC) Opinion on an Annex XV
dossier proposing restrictions on
four vhthalates (ECHA, 2012b)
Committee for Risk Assessment
(RAC) Committee for Socio-
economic Analysis (SEAC):
Background document to the
Opinion on the Annex XV dossier
proposing restrictions on four
phthalates (ECHA, 2012a)
Opinion on an Annex XV dossier
proposing restrictions on four
phthalates (DEHP, BBP, DBF,
DIBP) (ECHA. 2017b)
Annex to the Background
document to the Opinion on the
Annex XV dossier proposing
restrictions on four phthalates
(DEHP, BBP, DBP, DIBP)
(ECHA. 2017a)
Yes
Yes
No
- Peer-reviewed by ECHA's Committee
for Risk Assessment (RAC)
- Subject to public consultation
- No formal systematic review protocol
employed.
1952
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Appendix B Fetal Testicular Testosterone as an Acute Effect
No studies of experimental animal models are available that investigate the antiandrogenic effects of
DIBP following single dose, acute exposures. However, there are studies of its isomer, dibutyl phthalate
(DBP) available that indicate a single acute exposure during the critical window of development (i.e.,
GD 15.5 to GD 18.5 in rats) 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;
2011) 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.
Studies of DBP provide evidence to support use of effects on fetal testosterone and the developing male
reproductive system consistent with phthalate syndrome as an acute effect. However, the database is
limited to just a few DBP studies that test relatively high (500 mg/kg) single doses of DBP. Although
there are no single exposure studies of DIBP that evaluate anti androgenic effects on the developing male
reproductive system, there are three studies that have evaluated effects on fetal testicular testosterone
production and steroidogenic gene expression following daily gavage doses of 100 to 900 mg/kg-day
DIBP from GDs 14 to 18 (5 total doses) (Gray et al.. 2021; Furr et al.. 2014; Hannas et al.. 2012; Hannas
et al.. 2011). all of which consistently report anti androgenic effects at 300 mg/kg-day DIBP.
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Appendix C Calculating Daily Oral Human Equivalent Doses and
Human Equivalent Concentrations
For DIBP, all data considered for PODs are obtained from oral animal toxicity studies in rats or mice.
Because toxicity values for DIBP 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. However, no PBPK models or chemical-specific information was identified for DIBP to
support a quantitative extrapolation. In the absence of such data, EPA relied on the guidance from U.S.
EPA (2011c). which recommends scaling allometrically across species using the three-quarter power of
body weight (BW34) 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 C-l.
EquationApx C-l. Dosimetric Adjustment Factor
/BWa\1/4
Where:
DAF = Dosimetric adjustment factor (unitless)
BWa = Body weight of species used in toxicity study (kg)
BWh = Body weight of adult human (kg)
U.S. EPA (2011c). 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 DIBP risk evaluation (U.S. EPA. 2011a). EPA used the body weights of 0.025
and 0.25 kg for mice and rats, respectively, as presented in U.S. EPA (2011c). The resulting DAFs for
mice and rats are 0.133 and 0.236, 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 BW34. Using the appropriate DAF from EquationApx C-l, EPA adjusts the POD to obtain the HED
using Equation Apx C-2:
Equation Apx C-2. Daily Oral Human Equivalent Dose
Where:
HEDDaily
PODDaily
DAF
HEDDaily PODDaiiy X DAF
Human equivalent dose assuming daily doses (mg/kg-day)
Oral POD assuming daily doses (mg/kg-day)
Dosimetric adjustment factor (unitless)
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:
Equation Apx C-3. Extrapolating from Oral HED to Inhalation HEC
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jirn r BWh
HECDaily,continuous ~ H E D Daily ^ ^JR * ED ^
Where:
HECDaily, continuous = Inhalation HEC based on continuous daily exposure (mg/m3)
HEDDaiiy = 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 (2011a). EPA assumes an at rest breathing rate of 0.6125 m3/hr.
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. EquationApx C-4 presents the conversion of
the HEC from mg/m3 to ppm.
Equation Apx C-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 DIBP = 278.35 g/mol)
C.l DIBP Non-cancer HED and HEC Calculations for Acute,
Intermediate, and Chronic Duration Exposures
The acute, intermediate, and chronic duration non-cancer POD is based on a BMDLs of 24 mg/kg-day
and the critical effect is decreased fetal testicular testosterone. The POD was derived from three
gestational exposure studies of rats (Gray et al.. 2021: Hannas et al.. 2011: Howdeshell et al.. 2008).
This non-cancer POD is considered protective of effects observed following acute, intermediate, and
chronic duration exposures to DIBP. EPA used Equation Apx C-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 C-2:
mq mq
5.66 = 24- X 0.236
kg day kg day
EPA then calculated the continuous HEC for an individual at rest using Equation Apx C-3:
mq mq 80 kq
30.9 ^ = 5.66 2- x ( = )
m kg day 0.6125* 24 hr
hr
Equation Apx C-4 was used to convert the HEC from mg/m3 to ppm:
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ma 24.45
2068 2.71 ppm = 30.9 - X
m3 278.35 g/mol
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Appendix D Considerations for Benchmark Response (BMR) Selection
for Reduced Fetal Testicular Testosterone
D.l Purpose
EPA has conducted an updated meta-analysis and benchmark dose modeling (BMD) analysis of
decreased fetal rat testicular testosterone (U.S. EPA. 2024d). During the July 2024 Science Advisory
Committee on Chemicals (SACC) peer-review meeting of the draft risk evaluation of diisodecyl
phthalate (DIDP) and draft human health hazard assessments for diisononyl phthalate (DINP), the
SACC recommended that EPA should clearly state its rational for selection of benchmark response
(BMR) levels evaluated for decreases in fetal testicular testosterone relevant to the single chemical
assessments (U.S. EPA. 2024m). This appendix describes EPA's rationale for evaluating BMRs of 5, 10,
and 40 percent for decreases in fetal testicular testosterone. {Note: EPA will assess the relevant BMR for
deriving relative potency factors to be used in the draft cumulative risk assessment separately fi'om this
analysis.)
D.2 Methods
As described in EPA's Benchmark Dose Technical Guidance (U.S. EPA. 2012). "Selectinga BMR(s)
involves making judgments about the statistical and biological characteristics of the dataset and about
the applications for which the resulting BMDs BMDLs will be used. " For the updated meta-analysis and
BMD modeling analysis of fetal rat testicular testosterone, EPA evaluated BMR values of 5, 10, and 40
percent based on both statistical and biological considerations (U.S. EPA. 2024d).
In 2017, NASEM (2017) modeled BMRs of 5 and 40 percent for decreases in fetal testicular
testosterone. NASEM did not provide explicit justification for selection of a BMR of 5 percent.
However, justification for the BMR of 5 can be found elsewhere. As discussed in EPA's Benchmark
Dose Technical Guidance (U.S. EPA. 2012). a BMR of 5 percent is supported in most developmental
and reproductive studies. Comparative analyses of a large database of developmental toxicity studies
demonstrated that developmental NOAELs are approximately equal to the BMDLs (Allen et al.. 1994a.
b; Faustman et al.. 1994).
EPA also evaluated a BMR of 10 percent as part of the updated BMD analysis. BMD modeling of fetal
testosterone conducted by NASEM (2017) indicated that BMDs estimates are below the lowest dose
with empirical testosterone data for several of the phthalates (e.g., DIBP). As discussed in EPA's
Benchmark Dose Technical Guidance (U.S. EPA. 2012) "For some datasets the observations may
correspond to response levels far in excess of a selected BMR and extrapolation sufficiently below the
observable range may be too uncertain to reliably estimate BMDsBMDLs for the selected BMR. "
Therefore, EPA modelled a BMR of 10 percent because datasets for some of the phthalates may not
include sufficiently low doses to support modeling of a 5 percent response level.
NASEM (2017) also modeled a BMR of 40 percent using the following justification: "previous studies
have shown that reproductive-tract malformations were seen in male rats when fetal testosterone
production was reduced by about 40% ^Grav et al.. 2016; Howdeshell et al.. 2015)."
Further description of methods and results for the updated meta-analysis and BMD modeling analysis
that evaluated BMRs of 5, 10, and 40 percent for decreased fetal testicular testosterone are provided in
EPA's Draft Meta-Analysis and Benchmark Dose Modeling of Fetal Testicular Testosterone for Di (2-
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ethylhexyl) Phthalate (DEHP), Dibutyl Phthalate (DBP), Butyl Benzyl Phthalate (BBP), Diisobutyl
Phthalate (DIBP), andDicyclohexylPhthalate (DCHP) (U.S. EPA. 2024cT).
D.3 Results
BMD estimates, as well as 95 percent upper and lower confidence limits, for decreased fetal testicular
testosterone for the evaluated BMRs of 5, 10, and 40 percent are shown in TableApx D-l. BMDs
estimates ranged from 8.4 to 74 mg/kg-day for DEHP, DBP, DCHP, and DINP; however, a BMDs
estimate could not be derived for BBP or DIBP. Similarly, BMDio estimates ranged from 17 to 152 for
DEHP, DBP, DCHP, DIBP and DINP; however, a BMDio estimate could not be derived for BBP.
BMD40 estimates were derived for all phthalates (i.e., DEHP, DBP, DCHP, DIBP, BBP, DINP) and
ranged from 90 to 699 mg/kg-day.
In the mode of action (MOA) for phthalate syndrome, which is described elsewhere (U.S. EPA. 2023a)
and in Section 3.1.2 of this document, decreased fetal testicular testosterone is an early, upstream event
in the MOA that precedes downstream apical outcomes such as male nipple retention, decrease
anogenital distance, and reproductive tract malformations. Decreased fetal testicular testosterone should
occur at lower or equal doses than downstream apical outcomes associated with a disruption of androgen
action. Because the lower 95 percent confidence limit on the BMD, or BMDL, is used for deriving a
point of departure (POD), EPA compared BMDL estimates at the 5, 10, and 40 percent response levels
for each phthalate (DEHP, DBP, DCHP, DIBP, BBP, DINP) to the lowest identified apical outcomes
associated with phthalate syndrome to determine which response level is protective of downstream
apical outcomes.
Table Apx D-l provides a comparison of BMD and BMDL estimates for decreased fetal testicular
testosterone at BMRs of 5, 10, and 40 percent, the lowest LOAEL(s) for apical outcomes associated
with phthalate syndrome, and the POD selected for each phthalate for use in risk characterization. As
can be seen from Table Apx D-l, BMDL40 values for DEHP, DBP, DIBP, BBP, DCHP, and DINP are
all well above the PODs selected for use in risk characterization for each phthalate by 3X (for BBP) to
25 .4X (for DEHP). Further, BMDL40 values for DEHP, DBP, DIBP, BBP, and DCHP, but not DINP,
are above the lowest LOAELs identified for apical outcomes on the developing male reproductive
system. These results clearly demonstrate that a BMR of 40 percent is not appropriate for use in human
health risk assessment.
As can be seen from Table Apx D-l, BMDL10 values for DBP (BMDL10, POD, LOAEL = 20, 9, 30
mg/kg-day, respectively) and DCHP (BMDL10, POD, LOAEL = 12, 10, 20 mg/kg-day, respectively) are
slightly higher than the PODs selected for use in risk characterization and slightly less than the lowest
LOAELs identified based on apical outcomes associated with the developing male reproductive system.
This indicates that a BMR of 10% may be protective of apical outcomes evaluated in available studies
for both DBP and DCHP. BMDL10 values could not be derived for DIBP or BBP (Table Apx D-l).
Therefore, no comparisons to the POD or lowest LOAEL for apical outcomes could be made for either
of these phthalates at the 10 percent response level.
For DEHP, the BMDL10 is greater than the POD selected for use in risk characterization by 5X
(BMDL10 and POD = 24 and 4.8 mg/kg-day, respectively) and is greater than the lowest LOAEL
identified for apical outcomes on the developing male reproductive system by 2.4X (BMDL10 and
LOAEL = 24 and 10 mg/kg-day, respectively). This indicates that a BMR of 10 percent for decreased
fetal testicular testosterone is not health protective for DEHP. For DEHP, the BMDLs (11 mg/kg-day) is
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similar to the selected POD (NOAEL of 4.8 mg/kg-day) and the lowest LOAEL identified for apical
outcomes on the developing male reproductive system (10 mg/kg-day).
D.4 Weight of Scientific Evidence Conclusion
As discussed elsewhere (U.S. EPA. 2023a). DEHP, DBP, BBP, DIBP, DCHP, and DINP are
toxicologically similar and induce effects on the developing male reproductive system consistent with a
disruption of androgen action. Because these phthalates are toxicologically similar, it is more
appropriate to select a single BMR for decreased fetal testicular testosterone to provide a consistent
basis for dose response analysis and for deriving PODs relevant to the single chemical assessments. EPA
has reached the preliminary conclusion that a BMR of 5 percent is the most appropriate and health
protective response level for evaluating decreased fetal testicular testosterone when sufficient dose-
response data are available to support modeling of fetal testicular testosterone in the low-end range of
the dose-response curve. This conclusion is supported by the following weight of scientific evidence
considerations.
For DEHP, the BMDLio estimate is greater than the POD selected for use in risk characterization
by 5X and is greater than the lowest LOAEL identified for apical outcomes on the developing
male reproductive system by 2.4X. This indicates that a BMR of 10 percent is not protective for
DEHP.
The BMDL5 estimate for DEHP is similar to the selected POD and lowest LOAEL for apical
outcomes on the developing male reproductive system.
BMDLio estimates for DBP (BMDLio, POD, LOAEL = 20, 9, 30 mg/kg-day, respectively) and
DCHP (BMDLio, POD, LOAEL = 12, 10, 20 mg/kg-day, respectively) are slightly higher than
the PODs selected for use in risk characterization and slightly less than the lowest LOAELs
identified based on apical outcomes associated with the developing male reproductive system.
This indicates that a BMR of 10 percent may be protective of apical outcomes evaluated in
available studies for both DBP and DCHP. However, this may be a reflection of the larger
database of studies and wider range of endpoints evaluated for DEHP, compared to DBP and
DCHP.
NASEM (2017) modeled a BMR of 40 percent using the following justification: "previous
studies hcn'e shown that reproductive-tract malformations were seen in male rats when fetal
testosterone production was reduced by about 40% ^Grav et al.. 2016; Howdeshell et al.. 2015)."
However, publications supporting a 40 percent response level are relatively narrow in scope and
assessed the link between reduced fetal testicular testosterone in SD rats on GD 18 and later life
reproductive tract malformations in F1 males. More specifically, Howdeshell et al. (2015) found
reproductive tract malformations in 17 to 100 percent of F1 males when fetal testosterone on GD
18 was reduced by approximately 25 to 72 percent, while Gray et al. (2016) found dose-related
reproductive alterations in F1 males treated with dipentyl phthalate (a phthalate not currently
being evaluated under TSCA) when fetal testosterone was reduced by about 45 percent on GD
18. Although NASEM modeled a BMR of 40 percent based on biological considerations, there is
no scientific consensus on the biologically significant response level and no other authoritative
or regulatory agencies have endorsed the 40 percent response level as biologically significant for
reductions in fetal testosterone.
BMDL40 values for DEHP, DBP, DIBP, BBP, DCHP, and DINP are above the PODs selected for
use in risk characterization for each phthalate by 3X to 25.4X (Table Apx D-l). BMDL40 values
for DEHP, DBP, DIBP, BBP, and DCHP, but not DINP, are above the lowest LOAELs
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2203 identified for apical outcomes on the developing male reproductive system. These results clearly
2204 demonstrate that a BMR of 40 percent is not health protective.
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TableApx D-l. Comparison of BMD/BMDL Values Across BMRs of 5%, 10%, and 40% with PODs and LOAELs for Apical
Outcomes for DEHP, DBP, DIBP, BBP, DCHP, and DINP
Phthalate
POD (mg/kg-day) Selected for use
in Risk Characterization
(Effect)
Lowest LOAEL(s)
(mg/kg-day) for Apical
Effects on the Male
Reproductive System
BMDS
Estimate"
(mg/kg-day)
[95% CI]
BMDio
Estimate"
(mg/kg-day)
[95% CI]
BMD40
Estimate"
(mg/kg-day)
[95% CI]
Reference For Further
Details on the Selected
POD and Lowest
Identified LOAEL,
DEHP
NOAEL = 4.8
(t male RTM in F1 and F2 males)
10 to 15
(NR, | AGD, RTMs)
17 [11, 31]
35 [24, 63]
178 [122, 284]
(U.S. EPA. 2024h)
DBP
BMDL5 = 9
(J, fetal testicular testosterone)
30
(t Testicular Pathology)
14 [9, 27]
29 [20, 54]
149 [101,247]
(U.S. EPA. 2024f)
DIBP
BMDL5 = 24
(J, fetal testicular testosterone)
125
(t Testicular Pathology)
_b
55 [NA, 266f
279 [136, 517]
(U.S. EPA. 2024i)
BBP
NOAEL = 50
(phthalate syndrome-related effects)
100
(IAGD)
_b
_b
284 [150, 481]
(U.S. EPA. 2024e)
DCHP
NOAEL = 10
(phthalate syndrome-related effects)
20
(t Testicular Pathology)
8.4 [6.0, 14]
17 [12, 29]
90 [63, 151]
(U.S. EPA. 2024a)
DINP
BMDL5 = 49
(J, fetal testicular testosterone)
600
(J, sperm motility)
74 [47, 158]
152 [97, 278]
699 [539, 858]
(U.S. EPA. 2025d)
Abbreviations: AGD = anogenital distance; BMD = benchmark dose; BMDL = lower 95% confidence limit on BMD; CI = 95% confidence interval; LOAEL = lowest
observable-adverse-effect level; NOAEL = no observable-adverse-effect level; POD = point of departure; RTM = reproductive tract malformations
" The linear-quadratic model provided the best fit (based on lowest AIC) for DEHP, DBP, DIBP, BBP, DCHP, and DINP.
h BMD and/or BMDL estimate could not be derived.
2207
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Appendix E BENCHMARK DOSE MODELING OF FETAL
TESTICULAR TESTOSTERONE
EPA conducted benchmark dose (BMD) modeling of ex vivo fetal testicular testosterone data from three
gestational exposure studies of DIBP (Gray et al.. 2021; Hannas et al.. 2011; Howdeshell et al.. 2008).
The BMD modeling for continuous data was conducted with the EPA's BMD software (BMDS 3.3.2).
All standard BMDS 3.3.2 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 (U.S. EPA. 2012).
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)
EPA evaluated benchmark response (BMR) levels of 1 control standard deviation (1 SD) and 5, 10, and
40% relative deviation. Model fit was judged consistent with EPA's benchmark Dose Technical
Guidance (U.S. EPA. 2012). An adequate fit was judged based on the %2 goodness-of-fit p-value (p >
0.1), magnitude of the scaled residuals in the vicinity of the BMR, and visual inspection of the model fit.
In addition to these three criteria forjudging adequacy of model fit, a determination was made as to
whether the variance across dose groups was constant. If a constant variance model was deemed
appropriate based on the statistical test provided in BMDS (i.e., Test 2; p-value > 0.05 [note: this is a
change from previous versions of BMDS, which required variance p-value > 0.10 for adequate fit]), the
final BMD results were estimated from a constant variance model. If the test for homogeneity of
variance was rejected (i.e., p-value < 0.05), the model was run again while modeling the variance as a
power function of the mean to account for this nonconstant variance. If this nonconstant variance model
did not adequately fit the data (i.e., Test 3; p-value < 0.05), the data set was considered unsuitable for
BMD modeling. Among all models providing adequate fit, the lowest BMDL was selected if the
BMDLs estimated from different adequately fitting models varied >3-fold; otherwise, the BMDL from
the model with the lowest AIC was selected.
If no model adequately fit the data set using the approach described above, EPA removed the highest
dose group and modelled the data again using the approach described above.
Table Apx E-l summarizes BMD modeling results for reduced ex vivo fetal testicular testosterone data,
while more detailed BMD model results are provided in Appendices E.l through E.3.
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2251 TableApx E-l. Summary of BMD Model Results for Decreased Ex Vivo Fetal Testicular
2252 Testosterone
Data set
BMR
Best-Fit
Model
(Variance)
BMD
(mg/kg-
day)
BMDL
(mg/kg-
day)
Notes
Appendix
Containing
Results
(Grav et al.. 2021)
5%
Exponential
3 (Constant)
63
24
E.l
(Howdeshell et al.,
2008)
5%
Hill
(Constant)
103
52
E.2
(Hannas et al., 2011)
5%
No models
adequately fit the
data set
E.3
2253
2254 E.l BMD Model Results (Gray et al. 2021)
2255
Table Apx E-2. Ex Vivo Feta
Rat Testicular Testosterone Data (Gray et al. 2021)
Dose
(mg/kg-day)
N
(# of litters)
Mean
Standard
Deviation
Notes
0
3
7.972888889
1.465303963
Data for Block 67 rats reported in
Supplementary Data file associated with
(Gray et al., 2021)
100
3
7.727111111
1.105751094
300
3
5.247777778
1.429576563
600
2
2.082416667
0.659141371
900
2
1.705333333
0.145192592
2257
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Table Apx E-3. BM1
) Mode
Results Ex Vivo Fetal Testicular Testosterone (Gray et al. 2021
Models3
Restriction15
Variance
BMR = 5%
BMR = 10%
BMR = 40%
BMR = 1 SD
P Value
AIC
BMDS
Recommends
BMDS Recommendation
Notes
BMD
BMDL
BMD
BMDL
BMD
BMDL
BMD
BMDL
Exponentia
13
Restricted
Constant
63.12026
24.43019
106.4244
50.18702
334.6495
242.737
124.0844
57.8285
0.4081261
44.84273867
Viable -
Recommended
Lowest AIC
BMDL 3x lower than lowest
non-zero dose
Exponential
5
Restricted
Constant
117.9931
36.28242
161.5201
69.23603
329.678
257.1545
173.525
79.69645
0.9645746
45.05235319
Viable -
Alternate
BMD/BMDL ratio > 3
Hill
Restricted
Constant
159.6043
38.88005
195.2869
72.20498
326.5199
255.0622
205.6406
83.34761
0.8074988
45.10974788
Viable -
Alternate
BMD/BMDL ratio > 3
Polynomial
Degree 3
Restricted
Constant
50.44101
43.48323
100.882
86.96657
403.528
347.8659
143.1312
102.5143
0.1694486
46.08266114
Viable -
Alternate
Polynomial
Degree 2
Restricted
Constant
50.4376
43.48294
100.8752
86.96588
403.5008
347.8635
143.1079
102.5161
0.1694491
46.08265471
Viable -
Alternate
Power
Restricted
Constant
50.42151
43.48131
100.843
86.96262
403.3721
347.8505
143.04
102.5174
0.1694501
46.08264079
Viable -
Alternate
Linear
Unrestricted
Constant
50.42151
43.48125
100.843
86.96262
403.3721
347.8505
143.04
102.5173
0.1694501
46.08264079
Viable -
Alternate
AIC = Akaike information criterion; BMD = benchmark dose; BMDL =benchmark dose lower limit; NA = Not Applicable.
a Selected Model (bolded and shaded gray).
b Restrictions defined in the BMDS 3.3 User Guide.
2259
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12
Frequentist Exponential Degree 3 Model with BMR of 0.05 Added
Risk for the BMD and 0.95 Lower Confidence Limit for the BMDL
10
2260
2261
2262
2263
£
o
+-<
LO
3
u
Estimated Probability
Response at BMD
O Data
BMD
BMDL
100
200
300
400 500
mg/kg-d
600
700
800
900
User Input
Info
Model
frequentist Ewponential degree 3
Model Restriction
Restricted
Dataset Name
Gray et al. (2021) - DIBP Testosterone
User notes
[Add user notes here]
Dose-Response Modi
M[dose] = a" ewp(±11 (b1 dose)"d)
Variance Model
Var[i] = alpha
Model Options
BMR Type
Rel. Dev.
BMRF
0.05
Tail Probability
-
Confidence Level
0.95
Distribution Type
Normal
Variance Type
Constant
Model Data
Dependent Variable
mg/kg-d
Independent Variable
[Custom]
Total # of Observation
5
Adverse Direction
Automatic
Page 83 of 94
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Model Results
Benchmark Dose
BMD
63.12026361
BMDL
24.43018547
BMDU
136.7436224
AIC
44.84273867
Test4P-value
0.40812612
0.0. F.
2
Model Parameters
# of Parameters
4
Variable
Estimate
Std Error
Lower Conf
Upper Conf
a
8.176712147
0.50528127
7.186379
9.167045
b
0.00183526
2.35E-04
0.001375
0.002296
d
1.377943821
3.25E-01
0.741889
2.013999
log-alpha
-0.003820249
1.44E-01
-0.28666
0.279017
Goodness of Fit
Dose
Size
Estimated
Median
Calc'd
Median
Observed
Mean
Estimated
3D
Calc'd
3D
Observe
d 3D
Scaled
Residual
8.17671215
7.972889
7.972889
0.998092
1.4659
1.4659
-0.959707
100
7.42306676
7.727111
7.727111
0.998092
1.1058
1.10575
0.5276271
900
5.26924478
5.247778
5.247778
0.998092
1.4296
1.42958
-0.037253
600
2.60984715
2.082417
2.082417
0.998092
0.6591
0.65914
-0.747325
900
1.11027296
1.705333
1.705333
0.998092
0.1452
0.14519
0.8431514
Likelihoods of Interest
#of
Model
Log Likelihood"
Parameters
AIC
A1
-17.5251903
6
47.05038
A2
-13.06218817
10
46.12438
A3
-17.5251903
6
47.05038
fitted
-18.42136934
4
44.84274
R
-31.49407383
2
66.98815
' Includes additive constant of -11.9462. This constant was not included in the LL derivation prior to BMDS 3.0.
2264
Tests of Interest
Test
2'Log(Likeliho
od Ratio)
Test df
p-value
1
36.86377131
6
<0.0001
2
8.926004259
4
0.062976
3
8.926004259
4
0.062976
4
1.792358067
2
0.408126
2265
2266
2267
E.2 BMP Model Results (Howdeshell et al. 2008)
Table Apx E-4. Ex Vivo Feta
Rat Testicular Testosterone Data (Howdeshell et al. 2008)
Dose
(mg/kg-day)
N
(# of litters)
Mean
Standard
Deviation
Notes
0
5
5.7
0.290688837
Data from Table 6 in (Howdeshell et al.,
2008)
100
8
5.44
0.537401154
300
5
3.4
0.626099034
600
5
2.31
0.782623792
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Dose
(mg/kg-day)
N
(# of litters)
Mean
Standard
Deviation
Notes
900
2
2.09
1.286934342
2268
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Table Apx E-5. BM1
) Mode
Results Ex Vivo Fetal Testicular Testosterone (Howdes
iell et al. 2008)
Models3
Restriction15
Variance
BMR = 5%
BMR = 10%
BMR = 40%
BMR = 1 SD
P Value
AIC
BMDS
Recommends
BMDS Recommendation
Notes
BMD
BMDL
BMD
BMDL
BMD
BMDL
BMD
BMDL
Exponential
3
Restricted
Constant
37.18733
28.33677
74.74254
58.22005
345.5016
282.4184
82.49155
57.22709
0.0322184
57.43690601
Questionable
Goodness of fit p-value <0.1
BMDL 3x lower than lowest
non-zero dose
Modeled control response std.
dev. >1.5 actual response
std. dev.
Exponential
5
Restricted
Constant
101.588
45.31085
139.2085
77.87386
298.0878
246.6513
139.0035
75.64862
0.6618655
52.75773749
Viable -
Alternate
Modeled control response std.
dev. > 1.5 actual response
std. dev.
Hill
Restricted
Constant
102.9819
52.24216
136.2697
82.27878
297.6961
236.3744
135.9319
80.08333
0.9596039
52.56903757
Viable -
Recommended
Lowest AIC
Modeled control response
std. dev. >|1.5| actual
response std. dev.
Polynomial
Degree 3
Restricted
Constant
56.3685
49.44861
112.737
98.89673
450.9479
395.5888
149.7341
115.3485
0.0035229
62.15461672
Questionable
Goodness of fit p-value <0.1
Modeled control response std.
dev. >1.5 actual response
std. dev.
Polynomial
Degree 2
Restricted
Constant
56.36571
49.44887
112.7314
98.89766
450.9255
395.5908
149.7243
115.3486
0.0035229
62.15461883
Questionable
Goodness of fit p-value <0.1
Modeled control response std.
dev. >1.5 actual response
std. dev.
Power
Restricted
Constant
56.37483
49.44799
112.7497
98.89597
450.9986
395.5839
149.7562
115.3481
0.0035229
62.15461483
Questionable
Goodness of fit p-value <0.1
Modeled control response std.
dev. >1.5 actual response
std. dev.
Linear
Unrestricted
Constant
56.37483
49.448
112.7497
98.89599
450.9986
395.584
149.7562
115.3481
0.0035229
62.15461483
Questionable
Goodness of fit p-value <0.1
Modeled control response std.
dev. >1.5 actual response
std. dev.
AIC = Akaike information criterion; BMD = benchmark dose; BMDL =benchmark dose lower limit; NA = Not Applicable.
a Selected Model (bolded and shaded gray).
b Restrictions defined in the BMDS 3.3 User Guide.
2270
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Frequentist Hill Model with BMR of 0.05 Added Risk for the BMD
and 0.95 Lower Confidence Limit for the BMDL
Estimated Probability
Response at BMD
O Data
BMD
BMDL
mg/kg-d
2271
2272
User Input
Info
Model
frequentist Hill
Model Restriction
Restricted
Dataset Name
Howdeshell (£008) - OIBP Testosterone
User notes
[Add user notes here]
Dose-Response Modi
M[dose] = q + v"dose*n/(k*n + dose'n)
Variance Model
Var[i] = alpha
Model Options
BMR Type
Rel. Dev.
BMRF
0.05
Tail Probability
-
Confidence Level
0.95
Distribution Type
Normal
Variance Type
Constant
Model Data
Dependent Variable
mg/kg-d
Independent Variable
[Custom]
Total # of Observation
5
Adverse Direction
Automatic
2274
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Model Results
Benchmark Dose
BMD
102.981862
BMDL
52.24216337
BMDU
185.1065152
AIC
52.56903757
Test4P-value
0.959603873
D.O.F.
1
Model Parameters
# of Parameters
5
Variable
Estimate
Std Error
Lower Conl
Upper Conl
3
5.702356605
0.24846946
5.215365
6.189348
Y
-3.700369546
0.54289795
-4.76443
-2.63631
k
251.0918424
37.2929652
177.999
324.1847
n
2.786041669
1.15019349
0.531704
5.04038
alpha
0.321384979
2.92E-02
0.264127
0.378643
Goodness of Fit
Dose
Size
Estimated
Calc'd
Observed
Estimated
Calc'd
Observe
Scaled
Median
Median
Mean
3D
3D
d 3D
Residual
0
5
5.7023566
5.7
5.7
0.566908
0.2907
0.29069
-0.009295
100
8
5.43804842
5.44
5.44
0.566908
0.5374
0.5374
0.0097369
300
5
3.40266882
3.4
3.4
0.566908
0.6261
0.6261
-0.010527
bUU
b
Z. JUZZ 3bb4
Z.'Jl
Z.J1
U.bbbSUH
U. (bZb
0. {'6ZtiZ
0. U JU6Z04
900
2
2.10465068
2.09
2.09
0.566908
1.2869
1.28693
-0.036548
2275
2276
Likelihoods of Interest
Model
Log Likelihood"
#of
Parameters
AIC
A1
-21.28323604
6
54.56647
A2
-18.36479748
10
56.72959
A3
-21.28323604
6
54.56647
fitted
-21.28451878
5
52.56904
R
-46.73498878
2
97.46998
' Includes additive constant of -22.97346. This constant w a;
Tests of Interest
Test
2"Log(Likeliho
od Ratio)
Test df
p-ualue
1
56.74038261
8
<0.0001
2
5.83687712
4
0.211666
3
5.83687712
4
0.211666
4
0.002565492
1
0.959604
2277
2278
2279
E.3 BMP Model Results (Hannas et al. 2011)
Table Apx E-6. Ex Vivo Feta
Rat Testicular Testosterone Data (Hannas et al. 2011)
Dose
(mg/kg-day)
N
(# of litters)
Mean
Standard
Deviation
Notes
0
3
5.19
1.195115057
Data from Table 1 in (Hannas et al.,
2011)
100
3
5.7
0.294448637
300
3
2.27
1.420281662
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Dose
(mg/kg-day)
N
(# of litters)
Mean
Standard
Deviation
Notes
600
3
1.05
0.692820323
900
3
0.65
0.173205081
2280
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2281 Table Apx E-7. BMP Mode
Results Ex Vivo Fetal Testicular Testosterone (All Dose Groups) (Hannas et al. 2011)
Models"
Restrictionb
Variance
BMR = 5%
BMR = 10%
BMR = 40%
BMR = 1 SD
P Value
AIC
BMDS
Recommends
BMDS Recommendation
Notes
BMD
BMDL
BMD
BMDL
BMD
BMDL
BMD
BMDL
Exponential
3
Restricted
Constant
53.19681
17.80873
86.18426
36.58139
248.2968
173.5584
121.3679
58.47676
0.0435151
47.56574178
Questionable
Constant variance test failed
(Test 2 p-value < 0.05)
Goodness of fit p-value <0.1
BMDL 3x lower than lowest
non-zero dose
Exponential
5
Restricted
Constant
253.0245
60.82534
263.8179
89.98152
289.7556
204.3087
269.3968
109.1352
0.2886587
44.42231369
Questionable
Constant variance test failed
(Test 2 p-value < 0.05)
BMD/BMDL ratio > 3
Hill
Restricted
Constant
168.4655
62.74761
190.5248
161.2484
259.3454
187.7035
202.9592
104.7965
0.2934707
44.40007786
Questionable
Constant variance test failed
(Test 2 p-value < 0.05)
Polynomial
Degree 3
Restricted
Constant
44.5932
38.08458
89.1864
76.16924
356.7457
304.6767
191.342
137.0539
0.0060429
51.72768824
Questionable
Constant variance test failed
(Test 2 p-value < 0.05)
Goodness of fit p-value <0.1
Polynomial
Degree 2
Restricted
Constant
44.61656
38.08266
89.23315
76.16526
356.9326
304.661
191.5359
137.0482
0.0060428
51.72769929
Questionable
Constant variance test failed
(Test 2 p-value < 0.05)
Goodness of fit p-value <0.1
Power
Restricted
Constant
44.60136
38.08385
89.20272
76.1677
356.8109
304.6708
191.4181
137.0529
0.0060429
51.72768347
Questionable
Constant variance test failed
(Test 2 p-value < 0.05)
Goodness of fit p-value <0.1
Linear
Unrestricted
Constant
44.60135
38.08385
89.20271
76.16769
356.8109
304.6725
191.4181
137.0514
0.0060429
51.72768347
Questionable
Constant variance test failed
(Test 2 p-value < 0.05)
Goodness of fit p-value <0.1
Exponential
3
Restricted
Non-
Constant
27.99886
15.98305
53.77395
32.81671
224.9883
159.1025
150.5282
68.44409
0.1963944
44.46367306
Questionable
Non-constant variance test
failed (Test 3 p-value < 0.05)
BMD 3x lower than lowest
non-zero dose
BMDL 3x lower than lowest
non-zero dose
Exponential
5
Restricted
Non-
Constant
54.67479
14.05239
87.27415
28.92156
246.6568
142.7024
167.4876
74.84092
0.1504498
45.2760973
Questionable
Non-constant variance test
failed (Test 3 p-value < 0.05)
BMD/BMDL ratio > 3
BMDL 3x lower than lowest
non-zero dose
Hill
Restricted
Non-
Constant
89.27453
18.68063
119.5257
35.15881
243.712
156.1378
171.9022
88.60382
0.2773613
44.38838686
Questionable
Non-constant variance test
failed (Test 3 p-value < 0.05)
BMD/BMDL ratio > 3
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Models"
Restrictionb
Variance
BMR = 5%
BMR = 10%
BMR = 40%
BMR = 1 SD
P Value
AIC
BMDS
Recommends
BMDS Recommendation
Notes
BMD
BMDL
BMD
BMDL
BMD
BMDL
BMD
BMDL
BMDL 3x lower than lowest
non-zero dose
Polynomial
Degree 3
Restricted
Non-
Constant
52.39019
47.32479
104.7804
94.65019
419.1216
378.5982
500.2151
240.1444
0.0315801
48.04250119
Questionable
Non-constant variance test
failed (Test 3 p-value < 0.05)
Goodness of fit p-value <0.1
Modeled control response std.
dev. >1.5 actual response
std. dev.
Polynomial
Degree 2
Restricted
Non-
Constant
52.38006
47.32657
104.7601
94.65252
419.0404
378.615
499.0547
240.1565
0.0315806
48.04246336
Questionable
Non-constant variance test
failed (Test 3 p-value < 0.05)
Goodness of fit p-value <0.1
Modeled control response std.
dev. >1.5 actual response
std. dev.
Power
Restricted
Non-
Constant
51.70133
47.10326
103.4027
94.20652
413.6106
376.8261
434.4408
236.6546
0.0300659
48.15086454
Questionable
Non-constant variance test
failed (Test 3 p-value < 0.05)
Goodness of fit p-value <0.1
Modeled control response std.
dev. >1.5 actual response
std. dev.
Linear
Unrestricted
Non-
Constant
52.37973
47.32659
104.7594
94.65314
419.0377
378.6051
499.0851
240.1585
0.0315806
48.04246244
Questionable
Non-constant variance test
failed (Test 3 p-value < 0.05)
Goodness of fit p-value <0.1
Modeled control response std.
dev. >1.5 actual response
std. dev.
AIC = Akaike information criterion; BMD = benchmark dose; BMDL =benchmark dose lower limit; NA = Not Applicable.
" Selected Model (bolded and shaded gray).
b Restrictions defined in the BMDS 3.3 User Guide.
2282
2283
Table Apx E-8. BMP Mode
Results Ex Vivo Fetal Testicular Testosterone (Highest Dose Group Removed) (Hannas et al. 2011)
Models"
Restrictionb
Variance
BMR
= 5%
BMR
= 10%
BMR
= 40%
BMR
= 1 SD
P Value
AIC
BMDS
Recommends
BMDS Recommendation
Notes
BMD
BMDL
BMD
BMDL
BMD
BMDL
BMD
BMDL
Exponential
3
Restricted
Constant
256.2599
17.37356
266.7157
35.68542
291.1642
166.2802
276.4071
63.30013
0.0329607
41.77330522
Questionable
Goodness of fit p-value <0.1
BMD/BMDL ratio > 3
BMDL 3x lower than lowest
non-zero dose
Page 91 of 94
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Models"
Restrictionb
Variance
BMR = 5%
BMR = 10%
BMR = 40%
BMR = 1 SD
P Value
AIC
BMDS
Recommends
BMDS Recommendation
Notes
BMD
BMDL
BMD
BMDL
BMD
BMDL
BMD
BMDL
Exponential
5
Restricted
Constant
252.0464
55.61988
262.9164
83.58491
289.2605
194.5526
270.1764
108.3206
NA
39.79519328
Questionable
BMD/BMDL ratio > 3
d.f.=0, saturated model
(Goodness of fit test cannot
be calculated)
Hill
Restricted
Constant
244.6114
207.076
255.1814
217.9335
284.1949
174.1718
262.508
107.2089
0.450376
37.79519327
Viable -
Recommended
Lowest AIC
EPA Notes: Poor visual fit.
No model selected for this
data set.
Polynomial
Degree 3
Restricted
Constant
34.81702
28.89754
69.63404
57.79509
278.5362
231.18
134.3093
92.80279
0.0401532
41.65559526
Questionable
Goodness of fit p-value <0.1
BMDL 3x lower than lowest
non-zero dose
Polynomial
Degree 2
Restricted
Constant
34.81929
28.89742
69.63859
57.79485
278.5543
231.1794
134.3216
92.80369
0.0401532
41.65559589
Questionable
Goodness of fit p-value <0.1
BMDL 3x lower than lowest
non-zero dose
Power
Restricted
Constant
34.81603
28.89757
69.63207
57.79515
278.5283
231.1806
134.3029
92.80444
0.0401532
41.65559519
Questionable
Goodness of fit p-value <0.1
BMDL 3x lower than lowest
non-zero dose
Linear
Unrestricted
Constant
34.81604
28.89758
69.63208
57.79515
278.5283
231.1806
134.3029
92.80386
0.0401532
41.65559519
Questionable
Goodness of fit p-value <0.1
BMDL 3x lower than lowest
non-zero dose
AIC = Akaike information criterion; BMD = benchmark dose; BMDL =benchmark dose lower limit; NA = Not Applicable.
" Selected Model (bolded and shaded gray).
b Restrictions defined in the BMDS 3.3 User Guide.
2284
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Frequentist Hill Model with BMR of 0.05 Added Risk for the BMD
and 0.95 Lower Confidence Limit for the BMDL
Estimated Probability
Response at BMD
O Data
BMD
BMDL
600
2285
2286
User Input
Info
Model
frequentist Hill
Model Restriction
Restricted
Dataset Name
as (2011.1 - DIBP Testosterone (high dose removed)
User notes
[Add user notes here]
Dose-Response Modi
M[dose] = q + v"doseYif(k*n + dose"n)
Variance Model
Var[i] = alpha
Model Options
BMR Type
Rel. Dev.
BMRF
0.05
Tail Probability
-
Confidence Level
0.95
Distribution Type
Normal
Variance Type
Constant
Model Data
Dependent Variable
mgfkg-d
Independent Variable
[Custom]
Total # of Observation
4
Adverse Direction
Automatic
2288
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2289
2290
2291
Model Results
Benchmark Dose
BMD
244.6114226
EiMDL
207.075989
BMDU
256.7691509
AIC
37.79519327
Test 4 P-ualue 1
0.450375979
D.O.F.
1
Model Parameters
# of Parameters
5
Variable
Estimate
Std Error
Lower Coni
Upper Uonl
9
5.445000004
0.34186044
4.774966
6.115034
V
-4.395006471
0.59212145
-5.55554
-3.23447
k
284.475258
10.8295577
263.2497
305.7008
n
Bounded
NA
MA
NA
alpha
0.701212507
0.20072773
0.307793
1.094632
Goodness of Fit
Dose
Size
Estimated
Calc'd
Observed
Estimated
Calc'd
Observe
Scaled
Median
Median
Mean
SD
3D
d 3D
Residual
0
3
5.445
5.19
5.19
0.837384
1.1951
1.19512
-0.527444
100
3
5.44499997
5.7
5.7
0.837384
0.2944
0.29445
0.5274436
300
3
2.27000003
2.27
2.27
0.837384
1.4203
1.42028
-6.12E-08
bUU
J
iLwyyyyyr1
lUb
lUb
U.tiJ (by 4
U. bbLlHZ
b.Zbbt-Ub
Likelihoods of Interest
#of
Model
Leg Likelihood"
Parameters
AIC
A1
-14.6127439
5
39.22549
A2
-11.4128589
8
38.82572
A3
-14.6127439
5
39.22549
fitted
-14.89759663
4
37.79519
R
-26.01000707
2
56.02001
" Includes additive constant of -11.02726. This constant was not included in the LL derivation prior to BMDS 3.0.
Tests of Interest
Test
2'Log(Likeliho
od Ratio)
Test df
p-value
1
29.19429634
6
<0.0001
2
6.399769998
3
0.0937
3
6.399769998
3
0.0937
4
0.569705469
1
0.450376
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