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EPA Document# EPA-740-D-24-024
December 2024
United States Office of Chemical Safety and
Environmental Protection Agency Pollution Prevention
Draft Non-cancer Human Health Hazard Assessment for
Dibutyl Phthalate (DBP)
Technical Support Document for the Draft Risk Evaluation
CASRN: 84-74-2
December 2024
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28 TABLE OF CONTENTS
29 ACKNOWLEGEMENTS 7
30 SUMMARY 8
31 1 INTRODUCTION 11
32 1.1 Human Epidemiologic Data: Approach and Preliminary Conclusions 11
33 1.2 Laboratory Animal Findings: Summary of Existing Assessments, Approach, and
34 Methodology 13
35 1.2.1 Existing Assessments of DBP 13
36 1.2.2 Approach to Identifying and Integrating Laboratory Animal Data 16
37 1.2.3 New Literature Identified and Hazards of Focus for DBP 18
38 2 TOXICOKINETICS 20
39 2.1 Oral Route 20
40 2.2 Inhalation Route 22
41 2.3 Dermal Route 23
42 2.4 Additional Toxicokinetic Considerations 25
43 2.5 Summary 26
44 3 NON-CANCER HAZARD IDENTIFICATION 27
45 3.1 Effects on the Developing Male Reproductive System 27
46 3.1.1 Summary of Available Epidemiological Studies 27
47 3.1.1.1 Previous Epidemiology Assessment (Conducted in 2019 or earlier) 27
48 3.1.1.1.1 Health Canada (2018b) 28
49 3.1.1.1.2 Radkeetal. (2019b; 2018) 29
50 3.1.1.1.3 NASEM report (2017) 32
51 3.1.1.1.4 Summary of the Existing Assessments of Male Reproductive Effects 32
52 3.1.1.2 Summary of New Studies Identified by EPA (2018 through 2019) 33
53 3.1.2 Summary of Laboratory Animals Studies 34
54 3.1.2.1 Developing Male Reproductive System 35
55 3.1.2.2 Other Developmental and Reproductive Outcomes 36
56 3.1.3 Mode of Action for Phthalate Syndrome 52
57 3.2 New Literature Considered for Non-Cancer Hazard Identification 56
58 3.3 Summary 56
59 4 DOSE-REPONSE ASSESSMENT 58
60 4.1 Selection of Studies and Endpoints for Non-cancer Health Effects 58
61 4.2 Non-cancer Oral Points of Departure for Acute, Intermediate, and Chronic Exposures 59
62 4.3 Weight of Scientific Evidence: POD for Acute, Intermediate, and Chronic Durations 71
63 5 CONSIDERATION OF PESS AND AGGEGRATE EXPOSURE 73
64 5.1 Hazard Considerations for Aggregate Exposure 73
65 5.2 PESS Based on Greater Susceptibility 73
66 6 POINTS OF DEPARTURE USED TO ESTIMATE RISKS FROM DBP EXPOSURE,
67 CONCLUSIONS, AND NEXT STEPS 81
68 REFERENCES 82
69 APPENDICES 99
70 Appendix A Existing Assessments from Other Regulatory Agencies of DBP 99
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Appendix B New Literature Considered for Non-Cancer Hazards 103
B. 1 Reproductive and Developmental Effects 103
B.2 Neurotoxicity 109
B.3 Metabolic/Nutritional 113
B.4 Cardiovascular Health Effects 116
B.5 Immune adjuvant effects 117
Appendix C Fetal Testicular Testosterone as an Acute Effect 120
Appendix D Calculating Daily Oral Human Equivalent Doses and Human Equivalent
Concentrations 121
D. 1 DBP Non-cancer HED and HEC Calculations for Acute, Intermediate, and Chronic Duration
Exposures 122
Appendix E Considerations for Benchmark Response (BMR) Selection for Reduced Fetal
Testicular Testosterone 124
E.l Purpose 124
E.2 Methods 124
E.3 Results 125
E.4 Weight of Scientific Evidence Conclusion 126
LIST OF TABLES
Table 1-1. Summary of DBP Non-cancer PODs Selected for Use by Other Regulatory Organizations . 14
Table 2-1. Metabolites of DBP Identified in Urine from Rats and Humans after Oral Administration... 21
Table 3-1. Summary of Scope and Methods Used in Previous Assessments to Evaluate the Association
Between DBP Exposure and Male Reproductive Outcomes 28
Table 3-2. Summary of Epidemiologic Evidence of Male Reproductive Effects Associated with
Exposure to DBP (Radke et al., 2018) 30
Table 3-3. Summary of Studies Evaluating Effects on the Developing Male Reproductive System
Following In Utero Exposures to DBP 37
Table 3-4. Summary of Studies Evaluating Effects on the Developing Male Reproductive System
following Prepubertal and Pubertal Exposure to DBP 50
Table 4-1. Studies Being Considered for POD Selection 64
Table 4-2. Summary of Effects of Gestational Exposure to DBP on Testicular Testosterone Across
Select Studies 67
Table 4-3. Overall Analyses and Sensitivity Analyses of Rat Studies of DBP and Fetal Testosterone
(Updated Analysis Conducted by EPA) 69
Table 4-4. Benchmark Dose Estimates for DBP and Fetal Testosterone in Rats 70
Table 5-1. PESS Evidence Crosswalk for Biological Susceptibility Considerations 75
Table 6-1. Non-cancer HECs and HEDs Used to Estimate Risks for Acute, Intermediate, and Chronic
Exposure Scenarios 81
LIST OF FIGURES
Figure 1-1. Overview of DBP Human Health Hazard Assessment Approach 17
Figure 2-1. Proposed Metabolic Pathway of DBP Following Oral Exposure (From (ECB, 2004)) 22
Figure 3-1. Hypothesized Phthalate Syndrome Mode of Action Following Gestational Exposure 52
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LIST OF APPENDIX TABLES
TableApx A-l. Summary of Peer-review, Public Comments, and Systematic Review for Existing
Assessments of DBP 99
TableApx B-l. Summary of New Animal Toxicology Studies Evaluating Effects on the Developmental
and Reproductive System Following Exposure to DBP 106
Table Apx B-2. Summary of New Animal Toxicology Studies Evaluating Effects on the Nervous
System Following Exposure to DBP Ill
TableApx B-3. Summary of New Animal Toxicology Studies Evaluating Effects on Metabolism
Following Exposure to DBP 114
Table Apx B-4. Summary of New Animal Toxicology Study Evaluating Effects on the Cardiovascular
System Following Exposure to DBP 116
Table Apx B-5. Summary of New DBP Studies Evaluating Effects on the Immune System 118
Table Apx E-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 .127
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KEY ABBREVIATIONS AND ACRONYMS
ACE Angiotensin converting enzyme
ADME Absorption, distribution, metabolism and excretion
AGD Anogenital distance
ALP Alkaline phosphatase
ALT Alanine aminotransferase
ATSDR Agency for Toxic Substances and Disease Registry
CASRN Chemical Abstracts Service Registry Number
CI Confidence Interval
CPSC Consumer Product Safety Commission (U.S.)
BMD Benchmark Dose
BMDL Benchmark dose (lower confidence limit)
BBP Butyl-benzyl-phthal ate
DBP Dibutyl phthalate
DEHP Di-ethylhexyl phthalate
DIBP Di-isobutyl phthalate
DIDP Diisodecyl phthalate
DINP Di-isononyl phthalate
E2 P-estradiol
ECB European Chemicals Bureau
ECP Eosinophil Cationic Protein
ECHA European Chemicals Agency
EDSP Endocrine Disrupting Screening Program
EFSA European Food Safety Authority
EPA Environmental Protection Agency (U.S.)
EPM Elevated Plus Maze
F344 Fischer 344 rat
FSH Follicle Stimulating Hormone
FST Forced Swim Test
GD Gestation Day
GLP Good Laboratory Practice
GSH Glutathione
HEC Human equivalent concentration
HED Human equivalent dose
Ig Immunoglobulin
LH Luteinizing Hormone
IHC Immunohistochemistry
LOAEL Lowest-observed-adverse-effect level
LOEL Lowest-observed-effect level
MNG Multinucleated gonocytes
MOA Mode of action
MOE Margin of exposure
NASEM National Academies of Sciences, Engineering, and Medicine
NICNAS National Industrial Chemicals Notification and Assessment Scheme
NOAEL No-observed-adverse-effect level
NOEL No-observed-effect level
NTP National Toxicology Program
NTP-CERHR National Toxicology Program Center for the Evaluation of Risks to Human Reproduction
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OFT
Open Field Test
OCSPP
Office of Chemical Safety and Pollution Prevention
OECD
Organisation for Economic Co-operation and Development
OPPT
Office of Pollution Prevention and Toxics
OR
Odds Ratio
PBPK
Physiologically based pharmacokinetic
PND
Post-natal day
PECO
Population, exposure, comparator, and outcome
PESS
Potentially exposed or susceptible subpopulations
PND
Postnatal Day
PNW
Postnatal Week
POD
Point of departure
PPARa
Peroxisome proliferator activated receptor alpha
SACC
Science Advisory Committee on Chemicals
SD
Sprague-Dawley
TSCA
Toxic Substances Control Act
TST
Tail Suspension Test
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 Children's Health Protection
(OCHP), Office of General Counsel (OGC), and ORD.
Docket
Supporting information can be found in the public docket, Docket ID EPA-HQ-QPPT-2018-0503.
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), Mark Myer, Jennifer Brennan (Assessment Leads),
Anthony Luz (Human Health Hazard Discipline Lead), Ashley Peppriell, Christelene Horton (Human
Health Hazard Assessors)
Contributors: Azah Abdallah Mohamed, Devin Alewel, John Allran, Rony Arauz Melendez, Sarah Au,
Lillie Barnett, Maggie Clark, Jone Corrales, Daniel DePasquale, Lauren Gates, Amanda Gerke, Myles
Hodge, Annie Jacob, Ryan Klein, Sydney Nguyen, Brianne Raccor, Maxwell Sail, Joe Valdez, Leora
Vegosen, and Susanna Wegner
Technical Support: Mark Gibson, Hillary Hollinger, and 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 Dibutyl Phthalate
(DBP) (U.S. EPA. 2024m). This document describes the use of reasonably available information to
identify the non-cancer hazards associated with exposure to DBP and the points of departure (PODs) to
be used to estimate risks from DBP exposures in the draft risk evaluation of DBP. EPA summarizes the
cancer and genotoxicity hazards associated with exposure to DBP 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,
2024a).
EPA identified effects on the developing male reproductive system as the most sensitive and robust non-
cancer hazard associated with oral exposure to DBP in experimental animal models (Section 3.1).
Effects on the developing male reproductive system were also identified as the most sensitive and robust
non-cancer effect following oral exposure to DBP by existing assessments of DBP, including those by
the U.S. Consumer Product Safety Commission (U.S. CPSC. 2014. 2010). Health Canada (ECCC/HC,
2020). European Chemicals Agency (2017a. b, 2010; ECB. 2004). The European Food Safety Authority
(2019. 2005). the Australian National Industrial Chemicals Notification and Assessment Scheme
(NICNAS. 2013). the NTP, (NTP-CERHR, 2003b) the California EPA (OEHHA, 2007) and in other
assessments (NASEM. 2017). EPA also considered epidemiologic evidence qualitatively as part of
hazard identification and characterization. However, epidemiologic evidence for DBP 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 and U.S. CPSC.
As discussed further in Section 3.1, EPA identified 37 oral exposure studies (35 of rats, 2 of mice) that
investigated the developmental and reproductive effects of DBP following gestational and/or perinatal
exposure to DBP, including multi-generational studies of reproduction (Wine et al.. 1997; NTP. 1995).
However, there are limited data that evaluate the effects of DBP following inhalation or dermal
exposures. Data that evaluate chronic exposures via any route are limited to one study (NTP. 2021).
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
development of phthalate syndrome.
EPA is proposing a point of departure (POD) of 9 mg/kg-day (human equivalent dose [HED] of
2.1 mg/kg-day) based on phthalate syndrome-related effects on the developing male reproductive system
(decreased fetal testicular testosterone) to estimate non-cancer risks from oral exposure to DBP for
acute, intermediate, and chronic durations of exposure in the draft risk evaluation of DBP. The proposed
POD was derived from EPAs updated meta-analysis originally conducted by the NAS (NASEM, 2017)
and subsequent benchmark dose (BMD) modeling of decreased fetal testicular testosterone (ex vivo
testicular testosterone production or testicular testosterone content) in eight studies of rats exposed to
DBP during gestation (Gray et al., 2021; Furr et al., 2014; Johnson et al., 2011; Struve et al., 2009;
Howdeshell et al.. 2008; Martino-Andrade et al.. 2008; Johnson et al.. 2007; Kuhl et al.. 2007). The
BMDLs of 9 mg/kg-day (HED 2.1 mg/kg-day) is within the range of PODs (i.e., 1 to 10 mg/kg-day)
identified from other studies based on antiandrogenic effects on the developing male reproductive
system (Furr et al.. 2014; Moody et al.. 2013; Boekelheide et al.. 2009; Lee et al.. 2004). These studies
support the selection of the BMDLs of 9 mg/kg-day for the acute, intermediate, and chronic duration
PODs. The sole chronic study identified by EPA does not offer a more sensitive chronic POD; the NTP
(2021) identified a POD of 510 mg/kg-day (based on LOAEL; HED = 130 mg/kg-day).
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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 DBP 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 4.2 and Appendix C. For
purposes of assessing non-cancer risks, the proposed POD is considered most applicable to women of
reproductive age, pregnant women, and infants. Use of this POD based on sensitive male reproductive
effects are expected to be protective of effects in other age groups (e.g., older children, adult males, and
women above reproductive age) and appropriate for a screening level assessment for these other age
groups.
No data are reasonably 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 exposure to DBP. Differences in absorption will be accounted for in dermal exposure estimates
in the draft risk evaluation for DBP. For the inhalation route, EPA is extrapolating the oral HED to an
inhalation human equivalent concentration (HEC) using a human body weight and breathing rate
relevant to a continuous exposure of an individual at rest (U.S. EPA. 1994). The oral HED and
inhalation HEC values selected by EPA to estimate non-cancer risk from acute/intermediate/chronic
exposure to DBP in the draft risk evaluation of DBP are summarized in Table ES-1 and Section 6.
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
proposed POD for use in risk characterization of DBP.
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311 Table ES-1. Non-cancer HED and HEC Used to Estimate Risks
Exposure
Scenario
Target Organ
System
Species
Duration
POD
(mg/kg-
day)
Effect
HED
(mg/
kg-day)
HEC
(mg/m3)
[ppm]
Benchmark
MOE
References
Acute,
intermediate
, chronic
Effects on the
developing
reproductive
system
Rat
5 to 14 day
throughout
gestation
bmdl5 =
9
i fetal
testicular
testosterone
2.1
11.6
[1.02]
UFa= 311
ufh=io
Total UF=30
(Grav et al..
2021;
NASEM.
2017)fe
(U.S. EPA.
2024s,)c
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 body weight scaling to the three-quarters power to derive the HED. Consistent with EPA Guidance (U.S.
EPA. 201 lb), the UFa was reduced from 10 to 3.
b EPA conducted an updated BMD analysis of the meta-regression and BMD modeling of DBP and fetal testicular testosterone
data in rats published bv NASEM (2017). The updated analysis included eight total studies: seven studies from NASEM(Furr et
al.. 2014; Johnson et al.. 2011; Strove et al.. 2009; Howdeshell et al.. 2008; Martino-Andrade et al.. 2008; Johnson et al.. 2007;
Kuhl et al.. 2007). in addition to a more recent studv bv Grav et al. (2021).
c The updated meta-analysis and BMD modeling of fetal testicular testosterone data are provided in the Draft Meta-Analvsis
and Benchmark Dose Modeling of Fetal Testicular Testosterone for Di(2-ethylhexvl) Phthalate (DEHP), Dibutvl Phthalate
(DBP), Butvl Benzvl Phthalate (BBP), Diisobutvl Phthalate (D1BP), Dicvclohexvl Phthalate (DCHP), and Diisononvl Phthalate
("U.S. EPA. 2024e).
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1 INTRODUCTION
In December 2019, the United States Environmental Protection Agency (U.S. EPA or the Agency)
designated dibutyl phthalate (DBP) 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) (U.S. EPA. 2019). EPA published the draft and final scope
documents for DBP 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 DBP and conduct a dose-response assessment to determine the toxicity values to be
used to estimate risks from DBP exposures. This technical support document for DBP summarizes the
non-cancer hazards associated with exposure to DBP and proposes non-cancer toxicity values to be used
to estimate risks from DBP exposures. Cancer human health hazards associated with exposure to DBP
are summarized in EPA's Draft Cancer Raman Health Hazard Assessment for Di(2-ethylhexyl)
Phthalate (DEHP), Dibutyl Phthalate (DBP), Diisobutyl Phthalate (DIBP), Butyl Benzyl Phthalate
(BBP) andDicyclohexylPhthalate (DCHP) (U.S. EPA. 2024a).
Over the past several decades the human health effects of DBP have been reviewed by several
regulatory and authoritative agencies, including the: U.S. Consumer Product Safety Commission (U.S.
CPSC); U.S. Agency for Toxic Substances and Disease Registry (ATSDR); U.S. National Toxicology
Program Center for the Evaluation of Risks to Human Reproduction (NTP-CERHR); The National
Academies of Sciences, Engineering, and Medicine (NASEM); Health Canada; European Chemicals
Bureau (ECB); European Chemicals Agency (ECHA); European Food Safety Authority (EFSA); and the
Australian National Industrial Chemicals Notification and Assessment Scheme (NICNAS). EPA relied
on information published in existing assessments by these regulatory and authoritative agencies as a
starting point for its human health hazard assessment of DBP. Additionally, EPA considered new
literature published since the most recent existing assessments of DBP 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 DBP literature is described in the
Draft Systematic Review Protocol for Dibutyl Phthalate (DBP) (U.S. EPA. 2024o). 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, respectively.
1.1 Human Epidemiologic Data: Approach and Preliminary Conclusions
To identify and integrate human epidemiologic data into the draft DBP Risk Evaluation, EPA first
reviewed the conclusions of existing assessments of DBP conducted by regulatory and authoritative
agencies, as well as several systematic reviews of epidemiologic studies of DBP published by Radke et
al.; authors are affiliated with the U.S. EPA's Center for Public Health and Environmental Assessment.
Existing assessments reviewed by EPA are listed below. As described further in Appendix A, most of
these epidemiologic assessments have been subjected to peer review and/or public comment periods and
have employed formal systematic review protocols.
• 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 nearodevelopment, allergies, cardiovascular
function, oxidative stress, breast cancer, obesity, and metabolic disorders (Health Canada.
2018a);
• Phthalate exposure and male reproductive outcomes: A systematic review of the human
epidemiological evidence (Radke et al.. 2018);
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• 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); and
• Application of systematic review methods in an overall strategy for evaluating low-dose toxicity
fi'om endocrine active chemicals (NASEM. 2017).
EPA relies on conclusions from Health Canada (2018a. b) and systematic review publications in the
open literature from authors affiliated with EPAs Center for Public Health and Environmental
Assessment (2020a; 2019b; 2019a; Radke et al.. 2018) for interpretation of epidemiological studies
published prior to publication of those assessments. EPA also considered the conclusions from NASEM
(2017). OPPT reviewed new literature to evaluate whether new data alter conclusions of these previous
assessments. To do this, EPA identified new population, exposure, comparator, and outcome (PECO)-
relevant literature published since the most recent existing assessment of DBP. PECO-relevant literature
published since the most recent existing assessment(s) of DBP was identified by applying a literature
inclusion cutoff date from existing assessments of DBP. For DBP, 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 the assessments provided the most robust and recent evaluation of human epidemiologic data
for DBP. 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 DBP metabolites and health outcomes. New PECO-relevant
literature published between 2018 to 2019 that 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 DBP Docket (EPA-HQ-QPPT-2018-0503). 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 Draft Risk Evaluation for Dibutyl Phthalate - Systematic Review Supplemental File:
Data Quality Evaluation Information for Human Health Hazard Epidemiology (U.S. EPA. 2024e).
As described further in the Draft Systematic Review Protocol for Dibutyl Phthalate (DBP) (U.S. EPA.
2024o). EPA considers phthalate metabolite concentrations in urine to be an appropriate proxy of
exposure from all sources—including exposure through ingestion, dermal absorption, and inhalation. As
described in 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 DBP metabolites in matrices other than urine were considered supplemental and not evaluated
for data quality.
The Agency is proposing to use epidemiologic studies of DBP qualitatively due to the low confidence in
the level of evidence for association between urinary metabolites of DBP and health outcomes. This
proposal is consistent with the conclusions of Health Canada, U.S. CPSC, ECHA, EFSA, and Australia
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NICNAS. EPA reviewed the conclusions from Health Canada (2018a. b) and U.S. EPA systematic
review articles (Radke et al.. 2020a; Radke et al.. 2019b; Radke et al.. 2019a; Radke et al.. 2018) and
used the conclusions as a starting point for its human health hazard assessment. The Agency also
evaluated and summarized new epidemiologic studies identified by EPA's systematic review process to
use qualitatively during evidence integration to inform hazard identification and the weight of evidence.
The Agency 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. Additionally, the majority of epidemiological studies examine one phthalate and
one exposure period at a time such that they are treated as if they occur in isolation, which contributes
additional uncertainty due to co-exposure that may confound results for the majority of epidemiologic
studies (Shin et al.. 2019; Aylward et al.. 2016).
1.2 Laboratory Animal Findings: Summary of Existing Assessments,
Approach, and Methodology
1.2.1 Existing Assessments of DBP
The human health hazards of DBP have been evaluated in existing assessments by U.S. EPA (1987).
U.S. CPSC (2014. 2010). AT SDR (2001); NTP-CERHR (2003b); NASEM (2017). California OEHHA
(2007). Health Canada (ECCC/HC. 2020; EC/HC. 2015); ECB (2004). ECHA (2017a. b, 2010). EFSA
(2019. 2005). and Australia NICNAS (2013). These assessments have consistently identified male
reproductive development as the most sensitive outcome for use in estimating human risk from exposure
to DBP. The PODs from these assessments are shown in Table 1-1.
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433 Table 1-1. Summary of DBP Non-cancer POPs Selected for Use by Other Regulatory Organizations
Brief Study Description
NOAEL/
LOAEL
(mg/kg-
day)
Critical Effect
o
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Brief Study Description
NOAEL/
LOAEL
(mg/kg-
day)
Critical Effect
d
u
w
o
CNl
<
S/l
Ph
w
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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 DBP Risk Evaluation. EPA first reviewed existing assessments of DBP 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 DBP, and identify key studies used to establish PODs for estimating human risk. As
described further in 0, most of these assessments have been subjected to external peer review and/or
public comment periods.
• Integrated Risk Information System (IRIS), chemical assessment summary, dibutylphthalate;
CASRN84-74-2 (U.S. EPA. 1987);
• Toxicity review of di-n-butylphthalate (DBP) (U.S. CPSC. 2010);
• Chronic Hazard Advisory Panel on phthalate s and phthalate alternatives (U.S. CPSC. 2014);
• Toxicological profile for di-b-phthalate (ATSDR. 2001);
• NTP-CERHR Monograph on the Potential Raman Reproductive and Developmental Effects of
Di-n-Butyl Phthalate (DBP) (NTP-CERHR. 2003b);
• Application of systematic review methods in an overall strategy for evaluating low-dose toxicity
fi'om endocrine active chemicals (NASEM. 2017);
• Proposition 65 Maximum Allowable Dose Level (MADL) for reproductive toxicity for di(n-
butyl)phthalate (DBP) (OEHHA. 2007);
• 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. 2015);
• Supporting documentation: Carcinogenicity ofphthalate s - mode of action and human relevance
(Health Canada. 2015);
• Screening assessment - Phthalate substance grouping (ECCC/HC. 2020);
• European Union Risk Assessment Report: Dibutyl phthalate with addendum to the environmental
section (ECB. 2004);
• Evaluation of new scientific evidence concerning the restrictions contained in Annex XVII to
Regulation (EC) No 1907 2006 (REACH): Review of new available information for dibutyl
phthalate (DBP) CAS No 84-74-2 Einecs No 201-557-4 (ECHA. 2010);
• 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 phthalate s (DEHP, BBP, DBP, DIBP) (ECHA. 2017a);
• Opinion of the Scientific Panel on food additives, flavourings, processing aids and materials in
contact with food (AFC) related to di-Butylphthalate (DBP) for use in food contact materials
(EFSA. 2005);
• Update of the risk assessment of di-butylphthalate (DBP), butyl-benzyl-phthalate (BBP), bis(2-
ethylhexyl)phthalate (DEHP), di-isononylphthalate (DINP) and di-isodecylphthalate (DIDP) for
use in food contact materials (EFSA. 2019); and
• Priority existing chemical assessment report no. 36: Dibutyl phthalate (NICNAS. 2013).
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Figure 1-1. Overview of DBP 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.
Similar to the epidemiological analysis, EPA used the 2015 Health Canada assessment (EC/HC. 2015)
as a starting point for this draft document. EPA identified key quantitative studies used to support dose-
response analysis in other recent assessments and selected these key studies to inform evidence
integration and dose-response analysis in this hazard assessment. EPA assumes that previous
assessments effectively identified relevant key studies published prior to publication. EPA used
systematic review to identify additional studies for consideration in the assessment as detailed in the
Draft DBP Systematic Review Protocol (U.S. EPA. 2024o). The Health Canada assessment included
scientific literature up to August 2014, and considered a range of human health hazards (e.g.,
developmental and reproductive toxicity, systemic toxicity to major organ systems, genotoxicity) across
all durations (i.e., acute, intermediate (>1 to 30 days), subchronic (>30 to 90 days), chronic) and routes
of exposure (i.e., oral, dermal, inhalation). Therefore, EPA considered additional literature published
between 2014 to 2019 further as shown in Figure 1-1. EPA first screened titles and abstracts and then
full texts for relevancy using PECO screening criteria described in the Draft Risk Evaluation for Dibutyl
Phthalate - Systematic Review Protocol (U.S. EPA. 2024o).
Next, EPA reviewed PECO relevant new studies identified through this literature update published
between 2014 and 2019 and extracted key study information as described in the Draft DBP Systematic
Review Protocol (U.S. EPA. 2024o). Extracted information included: PECO relevance; species tested;
exposure route, method, and duration of exposure; number of dose groups; target organ/systems
evaluated; information related to potentially exposed or susceptible subpopulations (PESS); and the
study-wide lowest-observed-effect level (LOEL) Figure 1-1.
New information for DBP identified through systematic review was primarily limited to oral exposure
studies. Study LOELs were converted to HEDs based on LOELs by scaling allometrically across species
using the three-quarter power of body weight (BW3 4) for oral data, which is the approach recommended
by U.S. EPA when physiologically based pharmacokinetic (PBPK) models or other information to
support a chemical-specific quantitative extrapolation is absent (U.S. EPA. 2011b). EPA's use of
allometric body weight scaling is described further in Appendix D.
EPA conducted data quality evaluations for studies with HEDs based on LOELs that were within an
order of magnitude of the lowest HED based on the lowest-observed-adverse-effect level (LOAEL)
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across existing assessments. Studies with HEDs for LOELs within an order of magnitude of the lowest
LOAEL-based HED identified across existing assessments were considered sensitive and potentially
relevant for POD selection. These studies were further reviewed by EPA to determine if they provide
information that supports a human health hazard not identified in previous assessments or to determine
if they contain sufficient dose-response information to support a potentially lower POD than identified
in existing assessments of DBP. Although EPA did not conduct data quality evaluations for studies with
HEDs based on LOELs that were greater than an order of magnitude of the lowest LOAELs, these
studies were still reviewed and integrated into the hazard identification process.
Effects on the developing male reproductive system are a focus of EPA's DBP hazard assessment.
Therefore, EPA also considered literature identified outside of the 2019 TSCA literature searches that
was identified through development 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). As discussed further in the Draft Cancer Human 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.
2024a). no two-year bioassays were identified outside of EPA's 2019 literature searches or through
EPA's review of existing assessments of DBP. However, the Division of Translational Toxicology
(DTT) more recently published a technical report (i.e., two-year bioassays in mice and rats) (NTP.
2021). which was also considered by EPA in the development of this Draft TSD.
Data quality evaluations for DBP animal toxicity studies reviewed by EPA are provided in the Data
Quality Evaluation Information for Human Health Hazard Animal Toxicology for Dibutyl Phthalate
(DBP) (U.S. EPA. 2024d).
1.2.3 New Literature Identified and Hazards of Focus for DBP
As described in Section 1.2.2, and as described further in the Draft DBP Systematic Review Protocol
(U.S. EPA. 2024o). EPA reviewed literature published between 2014 to 2019 for new information on
sensitive human health hazards not previously identified in existing assessments, including information
that may indicate a more sensitive POD. As described further in the Draft DBP Systematic Review
Protocol (U.S. EPA. 2024o). EPA identified 63 new PECO-relevant animal toxicology studies that
provided information pertaining to various primary hazard outcomes, including:
reproduction/development, neurological, metabolic/nutritional, cardiovascular, and the immune system.
Twelve of these studies supported an HED based on a LOEL within an order of magnitude of the most
sensitive LOAEL of 2 mg/kg-day identified in recent hazard assessments of DBP (EFSA. 2019; ECHA.
2017a; OEHHA. 2007). Further details regarding EPA's handling of new information provided in these
12 studies are provided below. Information pertaining to the remaining 51 studies with HEDs based on
LOAELs greater than an order of magnitude of the most sensitive LOAEL of 2 mg/kg-day is available
in a supplemental file (U.S. EPA. 2024n).
• Reproductive/Developmental. EPA identified seven new studies evaluating reproductive/
developmental outcomes that provided potentially sensitive LOAELs (Xie et al.. 2019; Zhang et
al.. 2018a; Xie et al.. 2016; Ahmad et al.. 2015; de Jesus et al.. 2015; Sen et al.. 2015; Ahmad et
al.. 2014). These new studies of DBP are discussed further in Section B.l. Of the seven, only
Ahmad et al. (2014) and de Jesus et al. (2015) evaluated endpoints relevant to phtalate syndrome
(i.e., histopathology and/or organ weights of the male reproductive system, anogenital distance).
The others evaluated a range of endpoints including changes in the estrus cycle or serum
estradiol, progesterone, FSH, LH, number of ovarian follicles, reproductive organ weights (i.e.,
ovary and or uterus), and/or pup body weights (Xie et al.. 2019; Zhang et al.. 2018a; Xie et al..
2016; Ahmad et al.. 2015; Sen et al.. 2015).
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• Neurological. EPA identified three new studies evaluating neurotoxicity that provided
potentially sensitive LOAELs (Farzanehfar et al.. 2016; Yan et al.. 2016; Zuo et al.. 2014). These
new studies of DBP are discussed further in Section B.2.
• Nutritional/metabolic. EPA identified three new studies evaluating nutritional and/or metabolic
outcomes that provided potentially sensitive LAOELs (Maieed et al.. 2017; Ahmad et al.. 2015;
de Jesus et al.. 2015). These new studies of DBP are discussed further in Section B.3.
• Cardiovascular. EPA identified one new study evaluating cardiovascular outcomes that provided
potentially sensitive LAOELs (Xie et al.. 2019). This new study is discussed further in Section
B.4.
• Immune. EPA identified two new studies evaluating the immune adjuvant properties of DBP
that provided potentially sensitive LAOELs (Li et al.. 2014; Zuo et al.. 2014). These new studies
of DBP are discussed further in Section B.5.
The most sensitive and robust PODs selected from existing hazard assessments of DBP are based on
effects on the developing male reproductive system (EFSA. 2019; ECHA. 2017a; OEHHA. 2007).
Existing assessments have consistently shown that effects on other health outcomes (i.e., female
reproduction, neurological, cardiovascular, metabolic) are generally observed at higher dose levels than
developmental effects on male reproduction or are not supported by as robust databases of studies. This
is further supported by the new literature published from 2014 to 2019, as some of the lowest LOAELs
were identified for reproductive and developmental effects (Table Apx B-l) Therefore, the Agency
focused its non-cancer human health hazard assessment on toxicity to the male reproductive system
following developmental exposures (Section 3.1).
Genotoxicity and carcinogenicity data for DBP are summarized in EPA's Draft Cancer Raman Health
Hazard Assessment for Di (2-e thy Ihexyl) Phthalate (DEHP), Dibutyl Phthalate (DBP), Diisobiityl
Phthalate (DIBP), Butyl Benzyl Phthalate (BBP) and Dicyclohexyl Phthalate (DCHP) (U.S. EPA.
2024a). In sum, EPA has preliminarily concluded that DBP is Not Likely to Be Carcinogenic to Humans
based on a lack of carcinogenic activity in male and female mice as well as female rats, as reported in a
recent NTP Technical Report (NTP. 2021).
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2 TOXICOKINETICS
2.1 Oral Route
EPA identified several animal studies and 3 human studies (Koch et al.. 2012; Seckin et al.. 2009;
Anderson et al.. 2001) that evaluated the absorption, distribution, metabolism, and/or excretion (ADME)
of DBP following oral exposure. In humans and rodents, DBP undergoes hydrolysis to the bioactive
metabolite, mono-n-butyl phthalate (MBP) by non-specific lipases and esterases in the gastrointestinal
tract (Takahashi and Tanaka. 1989; White et al.. 1980; Lake et al.. 1977; Rowland et al.. 1977). and a
relatively small amount of the parent (DBP) reaches circulation (White et al.. 1980). Human and rodent
lipases show similar rates of conversion of DBP to MBP (Lake et al.. 1977). MBP is rapidly absorbed
and broadly distributed throughout the body with minimal bioaccumulation (Fennell et al.. 2004; Foster
et al.. 1983; Tanaka et al.. 1978). MBP can be excreted unchanged or undergo further oxidation to
produce more hydrophilic oxidative products. Alternatively, MBP can undergo phase II
biotransformation by glucuronosyltransferase, whereby MBP reacts with glucuronic acid to form
glucuronide conjugates, namely MBP-glucoronide (MBP-G. In rat serum, 80-90% of the total MBP is
free monobutyl phthalate and the remainder is MBP-G (ATSDR. 2001; Albro and Moore. 1974). This
differs from humans, where 25 to 30 percent of the total MBP in serum is free MBP and the remainder is
MBP-G (Silva et al.. 2003).
MBP can also be metabolized further through hydrolysis to phthalic acid, or through oxidation to
produce 3-hydroxybutyl phthalate (30H-MBP), 4-hydroxybutyl phthalate (40H-MBP), 3-ketobutyl
phthalate, or 4-carboxypropyl phthalate. Mono-carboxy-propyl phthalate (MCPP) has also been detected
in humans and rats exposed to DBP but is a minor metabolite (Table 2-1; Figure 2-1). In animals and
humans, MBP and MBP-glucuronide are the primary metabolites of DBP (ATSDR. 2001). MBP and
MBP-glucuronide are eliminated primarily in urine (ATSDR. 2001; Foster et al.. 1983). and to a smaller
extent in feces (Chang et al.. 2013; Fennell et al.. 2004; Saillenfait et al.. 1998). Enterohepatic
circulation has been reported (Tanaka et al.. 1978). A summary of different metabolites found in human
and rat urine after oral administration of DBP is presented in Table 2-1.
The elimination of various metabolites of DBP have been described in animals and humans. Since DBP
does not bioaccumulate, excretion may also be an indicator for absorption. Studies in pregnant and non-
pregnant animals have demonstrated that most (67 to 97 percent) of the administered dose is excreted in
urine within 24 hours (Chang et al.. 2013; Saillenfait et al.. 1998; Foster et al.. 1983; Tanaka et al..
1978). The elimination half time has been estimated to be approximately 3-hours in plasma of pregnant
rats given oral doses of 50 to 250 mg/kg-day DBP (Fennell et al.. 2004). Similarly, it has been reported
to be approximately 3.6 hours in rats given an intravenous injection of 30 mg/kg-day DBP (Chang et al..
2013). In humans, three separate studies suggest elimination of DBP between 73 and 92.2 percent within
24 hours (Koch et al.. 2012; Seckin et al.. 2009; Anderson et al.. 2001). In Anderson et al. (2001). 24
volunteers consumed DBP (255 or 510 |ig) administered in margarine spread on toast, and 73 percent of
the administered dose was excreted in 24 hours. In Seckin et al. (2009). 17 volunteers ingested a capsule
containing 3,600 |ig DBP, and 78 percent of the administered dose was excreted within 24 hours. In
Koch et al. (2012). one male volunteer ingested 60 (J,g/kg body weight of DBP, and 92.2 percent of the
dose was eliminated within 24 hours. Koch et al. (2012) and Seckin et al. (2009) also provide data on
elimination kinetics. In Koch et al., the elimination half-time for MBP was 2.6 hours, and approximately 6
hours for other metabolites, such as 30H-MBP and 40H-MBP (Koch et al.. 2012). Seckin et al. reported a
urinary elimination half-time for MBP of 6 hours in one individual who consumed a tablet containing
3,600 (ig DBP (Seckin et al.. 2009).
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There are species differences in some aspects of DBP biotransformation. There are species differences in
excretion of MBP or MBP-G, which suggest differences in metabolism of MBP. Indeed, the proportion of
MBP-G:MBP excreted differed in rats (1:1), guinea pigs (1.5:1), and hamsters (2.3:1) (Tanaka et al.. 1978).
Moreover, an in vitro study demonstrated that rates of MBP glucuronidation in pooled liver microsomes
differ across species, where microsomes from mice and rats had slower rates of MBP-G formation from
MBP substrate than human liver microsomes from rats with humanized livers (Miura et al.. 2019). The data
suggest that human liver microsomes have a faster rate of glucuronidation of MBP than mouse or rat
microsomes. Additionally, human biomonitoring studies (e.g., NHANES (Silva et al.. 2003)) and one
human ingestion study (Seckin et al.. 2009) have demonstrated that MBP in human urine and plasma is
conjugated with glucuronide. In contrast, most MBP is mostly unconjugated in rodents.
Table 2-1. Metabolites of DBP Identified in Urine from Rats and Humans after Oral
Administration
Urinary Metabolite
Abbreviation
Rat
Human"
Reference(s)
Mono-n-butyl phthalate
MBP
V
(Koch et al.. 2012) (human)
(Seckin et al.. 2009) (human)
(Anderson et al.. 2001) (human)
(Silva et al.. 2003) (human)11
(Foster et al.. 1983) (rat)
(Albro and Moore. 1974) (rat)
(Tanaka et al.. 1978) (rat. hamster)
(Clewell et al.. 2009) (rat
(Fennell et al.. 2004) (rat)c
Mono-n-butyl phthalate glucuronide
MBP-G
S
(Seckin et al.. 2009) (human)
(Silva et al.. 2003) (human)11
(Foster et al.. 1983) (rat)
mono-carboxy-propyl phthalate
MCPP
S
(Koch et al.. 2012) (human)
(Calafat et al.. 2006) (human a. rat)
(Albro and Moore. 1974) (rat)
3-hydroxybutyl phthalate
30H-MBP
s
(Koch et al.. 2012) (human)
(Albro and Moore. 1974) (rat)
4-hydroxybutyl phthalate
40H-MBP
s
(Albro and Moore. 1974) (rat)
(Tanaka et al.. 1978) (rat. hamster)
3-ketobutyl phthalate
-
s
ND
(Albro and Moore. 1974) (rat)
(Tanaka et al.. 1978) (rat)
4-carboxypropyl phthalate
-
s
ND
(Albro and Moore. 1974) (rat)
(Tanaka et al.. 1978) (rat)
Monobutanoic phthalic acid
-
s
ND
(Fennell et al.. 2004; General Motors.
1983a) (rat)c
Mono-n-hydroxybutylphthalate
-
s
ND
(Fennell et al.. 2004; General Motors.
1983a) (rat)
mono-1 -hydroxybutan-
2-one phthalic acid glucuronide
-
s
ND
(Fennell et al.. 2004) (rat)
Phthalic acid
PA
s
ND
(Fennell et al.. 2004; General Motors.
1983a) (rat)
(Albro and Moore. 1974) (rat)
ND = no data available
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Urinary Metabolite
Abbreviation
Rat
Human"
Reference(s)
"Metabolites detected as part of human urinary biomonitoring studies (Calafat et al.. 2006: Silva et al.. 2003) not controlled
exposure studies. Although biomonitoring studies do not distinguish between routes or pathways of exposure, urinary
metabolites are shown for comparison to urinary metabolites detected in rodent models.
h Reflects pup plasma concentrations on PND2 following exposure of dams from GD12 - PND14.
c Reflect maternal urine concentrations 24 hours after a single dose of 100 mg/kg DBP on GD20 in CD rats.
-COO(CH,)3CH;
-j-COO(CH,)3CH.
Di-ii-butylplithalate (DBP)
COOH
COOH
Phthalic add
i—COOH
COO(CH,),CH3
Monobiitylplithalate (MBP)
COO ghicuronide
COO(CH,)3CH3
MBP alucuronide
-COOH
-COO(CH,)2CHOHCH3
3-Hydroxy-butylphthalate
COOH
COO (CH^CHjOH
4-Hydroxy-butylphthalate
"OOH
-COO(CH,),COCH.
3 -Keto-butylphthalate
COOH
COO(CH,)XOOH
4-Carboxypropylphthalate
Figure 2-1. Proposed Metabolic Pathway of DBP Following Oral Exposure (From (ECB, 2004))
Note that metabolism of OH-MnBP into MCPP has been reported to occur in humans (Koch et al.. 2012) and rats
(Calafat et al.. 2006) but is not shown in the figure.
Based on the reasonably available data, which indicate DBP is readily absorbed and most of the
administered dose is eliminated in the urine within 24 hours following oral exposure in humans and rats,
EPA will assume an oral absorption of 100 percent for the draft risk evaluation. This is consistent
with assumptions used for adults and children in other existing assessments of DBP (EFSA. 2019;
ECHA. 2017a: NICNAS. 2013).
2.2 Inhalation Route
EPA identified two inhalation studies of rats, but each had several uncertainties that preclude their use to
inform the ADME of DBP following inhalation exposure in animals. However, inhalation studies from
other phthalates (i.e., DEHP and DIDP) exist and may be informative. No studies in humans were
available that evaluated the ADME properties of DBP for the inhalation route.
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Kawano (1980) and Walseth (1984) each provide data that DBP is absorbed and distributes to other
tissues following inhalation exposure. Kawano (1980) exposed male Wistar rats to aerosolized DBP via
a non-continuous whole body inhalation exposure. The target concentration was 50 mg/m3, and the
actual concentration of DBP ranged from approximately 45 to 60 mg/m3 over the course of 100 days, as
verified by gas chromatography. Animals were exposed 6 hours/day on weekdays, 3 hours/day on
Saturday, and rest day every Sunday. Changes in body and organ weight were observed, as well as
changes in liver enzyme levels and changes in white blood cell counts. Walseth et al. (1984) exposed
male Sprague Dawley rats to 0.5, 2.5, or 7 ppm DBP aerosols 6 hours/day for 5 days (equivalent to 5.7,
28.4, or 79.5 mg/m3). Rats were exposed via whole body inhalation in chambers, and exposure
concentrations were verified via gas chromatograph, but the data for the aerosol concentrations during
the experimental period were not provided. Changes in the activities of cytochrome P450 enzymes were
observed in liver and lung samples. No data are available on DBP metabolism following inhalation
exposure. CYPs and glucuronosyltransferases are included among the xenobiotic metabolizing enzymes
found in the respiratory tract, so it is feasible that metabolism of DBP to MBP and MBP-G occurs in the
lung. No data are available for elimination following inhalation exposure. Neither study characterized
the particle size distribution (e.g., no reporting of mass median aerodynamic diameter [MMAD] or
geometric standard deviation [GSD]), which is an important limitation. The systemic effects observed
by Kawano (1980) and Walseth (1984) may indicate that some absorption occurs in the lung. However,
these data are difficult to interpret due to the whole-body inhalation exposure method in both available
studies, including the potential for DBP deposited on the fur during whole body exposure and
subsequent grooming resulting in oral exposure. The aforementioned limitations (i.e., exposure method,
lack of presentation of MMAD or GSD) limit the ability to quantify the achieved dose from these two
whole-body inhalation studies.
Inhalation studies for other phthalates such as DIDP and DEHP exist (1991; General Motors. 1983b).
which may provide some insight into the toxicokinetic properties of DBP. In the DIDP exposure study,
rats were exposed to aerosolized 14C-DIDP (target concentration was 91 mg/m3; MMAD was 0.98 |im)
via head-only inhalation for 6 hours/day, 5 days/week for 2 weeks. In the DEHP exposure study, rats
were exposed to aerosolized 14C-DEHP (target concentration was 100 mg/m3; MMAD was 0.6 |im) via
head-only inhalation for 6 hours/day, 5 days/week for 2 weeks. In the DIDP study, absorption through
the lung was approximately 73 percent over 72 hours. In the DEHP study, absorption through the lung
was approximately 92 percent after 72 hours. Collectively, these studies of structurally similar
phthalates provide some indication that DBP can be expected to be readily absorbed through the lung.
No data from animal models are reasonably available for the inhalation route that are suitable for
deriving a route-specific POD. Therefore, EPA extrapolated the inhalation POD from the oral POD,
assuming similar absorption for the oral and inhalation routes, and no adjustment was made when
extrapolating between exposure routes. EPA will assume an inhalation absorption of 100 percent for
the draft risk evaluation. This is consistent with assumptions used in existing assessments (NICNAS.
2013).
2.3 Dermal Route
EPA identified two in vivo studies (Doan et al.. 2010; Elsisi et al.. 1989) and two in vitro/ex vivo studies
(Sugino et al.. 2017; Scott et al.. 1987) that evaluated the ADME properties of DBP following dermal
application. An additional study in humans was identified that provided data following dermal
application of skin cream containing several phthalates, including DBP (Janiua et al.. 2008).
Elsisi et al. (1989) provided data on the dermal absorption of eight phthalate diesters including DBP by
measuring the percentage of dose excreted in the urine and feces daily over the 7-day exposure.
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Radiolabeled DBP (14C-DBP) (5 to 8 mg/cm2) was applied to a circular area 1.3 centimeters in diameter
(1.3 cm2) on the shaved skin on the backs of male F344 rats, and the application site was covered with a
perforated circular plastic cap for seven days. Low levels (less than one percent for combined tissues) of
14C were found in adipose tissue, muscle, skin, and other tissues (i.e., brain, lung, liver, spleen, small
intestine, kidney, testis, spinal cord, and blood), suggesting DBP or its metabolites were systemically
distributed. In the first 24 hours, 11 percent of the administered dose of DBP was excreted in urine and 1
percent was excreted in the feces. DBP was excreted at a near constant rate of 10 to 12 percent every 24
hours. After 7 days of exposure, approximately 61 percent of the applied dose was recovered in urine or
feces. Based on the amount of radioactivity recovered from urine, feces, and other tissues, study authors
estimated that approximately 66 percent of the applied dose of 14C-DBP was absorbed over seven days.
The total recovery of the applied dose was 100 percent. Most of the applied dose was recovered in the
urine and feces, and 33 percent of the applied dose was recovered from skin at the application site. DBP
had a fast rate of excretion relative to other phthalates tested (i.e., DEHP, DIBP, BBP, DIDP, DEP,
DMP, and DHP), which may be related to its relatively low molecular weight and branched structure.
A more recent study in hairless guinea pigs (Doan et al.. 2010) also reported high dermal absorption of
DBP. Following a single dermal application via covered patch (3 x 3-centimeter square area; 9 cm2) of
an emulsion containing 1 mg/cm2 DBP, in vivo dermal absorption of DBP was estimated to be
approximately 62 percent of the applied dose after 24 hours (Doan et al.. 2010). The percent total
recovery was 92.9 percent after 24 hours. The major strength of the in vivo part of this study was that the
outcomes assessment method mostly agreed with guideline OECD 427 (OECD. 2004a). The study also
included an ex vivo experiment, where skin was excised from the guinea pigs (anatomical site of the
tissue collections was not specified) and radiolabeled DBP (1 mg/m2) was applied to a skin preparation.
A total of 56.3 percent of the administered dose was absorbed after 6 hours, and the percent total
recovery was 96.3 percent of the administered dose. Strengths of the ex vivo part of this study include
that the test system was un-occluded, the skin was washed prior to application, and overall, the study
complies with OECD guideline 428 (OECD. 2004b).
Scott et al. (1987) used epidermal membranes prepared from human abdominal skin and dorsal rat skin
to compare percutaneous absorption rates of four phthalates, including DBP and DEHP. The authors
also compared the permeability of human skin compared to rat skin. DBP is much more readily
absorbed in rat skin that in human skin (steady state absorption rate: 2.40 ± 0.63 [j,g/cm2/hr [human],
93.35 ± 0.94 [j,g/cm2/hr [rat]), which is related to the relatively higher permeability of rat skin
(permeability constant: 0.23 ± 0.06 x 10"5 cm/hr [human], 8.95 ± 0.09 10"5 cm/hr [rat]). A more recent ex
vivo study by Sugino et al. (2017) also noted species differences between humans and rats. That study
applied DBP to mounted skin membranes prepared from hairless rats or from human skin and evaluated
the permeability of the skin. After dermal application of DBP, esterases in the skin hydrolyze DBP to
MBP, which subsequently permeates the skin. The steady state permeability coefficients for MBP across
stripped skin (Kp) were 6.8 x 10"5 ± 2.2 x 10"5 cm/sec in rats and 7.2 x 10A"6 ± 1.1 x 10"6 cm/sec in
humans. This is equivalent to 0.245 cm/hr in rats and 0.026 cm/hr in humans, when adjusting for the BP
metabolite. These values reflect faster rates for the metabolite of DBP (i.e., MBP) than the rates for the
parent chemical described by Scott et al. (1987).
A study in humans provided data consistent with low dermal absorption of DBP following application of
a skin cream formulation (Janiua et al.. 2008). In that study, the authors applied 2 mg/cm2 of a control
cream (no added phthalates) or a cream with 2 percent (weight-to-weight) DBP (and other phthalates) to
the skin of participants (whole body topical application) for daily for 5 consecutive days. Urine was
collected via a 24-hour pooled collection method, and concentration of MBP was analyzed to estimate
absorption of DBP. The maximum dermal absorption in human participants in that study corresponded
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to approximately 6 percent of the applied dose of DBP. However, this study had significant limitations,
including very large inter-individual variability in absorption values and daily variations in values for
the same individual.
Although specific data on DBP dermal absorption in humans is limited to one study (Scott et al.. 1987).
several regulatory agencies (e.g., Danish EPA, ECHA, NICNAS) recognize that absorption of phthalates
would likely be lower in human skin than through rat skin. This observation is based on data from in
vitro migration studies conducted with DEHP and other phthalates. Notably, other regulatory agencies
(e.g., Australia NICNAS, ECHA) have reached similar conclusions regarding the low dermal absorption
of DBP (ECHA. 2013; NICNAS. 2012).
Details of the approach used by EPA to estimate exposure via the dermal exposure route for
occupational, consumer, and general population exposure assessments can be found in the Draft
Environmental Release and Occupational Exposure Assessment for Dibutyl Phthalate (DBP) (U.S. EPA.
2024f) and Draft Consumer and Indoor Dust Exposure Assessment for Dibutyl Phthalate (DBP) (U.S.
EPA. 2024b). In sum, EPA is proposing to use DBP dermal absorption data from the Doan et al. (2010)
study to estimate dermal absorption of liquid formulations of DBP. Using Doan (2010). EPA derived an
estimate of 56.3 percent absorption of 1 mg/cm2 of DBP over a day period (24 hours); the steady-state
flux of neat DBP is estimated as 2.35 x 10"2 mg/cm2/hr.
2.4 Additional Toxicokinetic Considerations
Transfer across the placenta
DBP and its metabolites can be transferred across the placenta to the fetus during gestation, including to
the fetal testis, which is the target organ of toxicity (Section 3.1) (Clewell et al.. 2009; Kremer et al..
2005; Fennell et al.. 2004; Saillenfait et al.. 1998). In pregnant Sprague-Dawley rats given a single oral
dose of 0.5 or 1.5 g/kg radiolabeled DBP (di-n-butyl[carboxyl-14C]phthalate) on GD14, radiolabel was
detected in the plasma, placenta, embryo, and amniotic fluid within a half hour (Saillenfait etal.. 1998).
MBP was the major metabolite in plasma and MBP-glucuronide was the minor metabolite from both
pregnant rats and the fetus. These findings were supported by additional studies, including that of Fennel
et al. (2004). who reported MBP and MBP-glucuronide in plasma of the dams exposed to DBP, as well
as the amniotic fluid and plasma of the fetus. Following exposure to 50 or 100 mg/kg DBP, the time to
reach maximum plasma concentration (Tmax) in the maternal plasma for MBP and MBP-glucuronide is
0.5 hour and 1 hour, respectively. The Tmax for fetal plasma is 1 and 4 hours, respectively. Kremer et al.
(2005) provide data that further support transfer of DBP metabolite across the placenta. Briefly, 50
mg/kg MBP was administered to pregnant rats via intravenous injection on GDI9. Levels of MBP-G
were higher in fetal plasma than maternal plasma 24 hours after dosing, which may imply that MBP-G is
diffusion limited from fetus to dam.
Inter-individual and intra- species considerations
Inter-individual and intra-species differences exist across various toxicokinetic parameters which may in
turn impact toxicity. Interspecies differences in DBP toxicity, including to the male reproductive system,
have been demonstrated (Gray et al.. 1982). which may reflect species-specific differences in
toxicokinetics. Some studies have demonstrated differences in toxicokinetics across species. For
instance, P-glucuronidase activity in testicular tissue was shown to be higher in rats than in hamsters
(Foster et al.. 1983). For dermal exposures to DBP, data from Scott et al. (1987) and Sugino et al. (2017)
demonstrate that there are large differences in the absorption rates of DBP between human and rodent
skin. These are important when considering the dermal absorption data provided by Elsisi et al. where
10 to 12 percent of DBP applied to rodent skin is absorbed and excreted every 24 hours. There are also
inter-individual ADME differences to account for, including age-related differences in metabolism of
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DBP in humans. For instance, the activity of glucuronosyltransferase differs between adults and infants,
where adult activity is higher and achieved at 6 to 18 months of age (Leeder and Kearns. 1997).
Additionally, toxicokinetic differences exist between males and females in general, and the paucity of
data comparing toxicokinetic parameters across the sexes represents an additional source of uncertainty
to be considered. Toxicokinetic factors that modify susceptibility to DBP are further discussed in
Section 5.
2.5 Summary
The majority of data pertaining to the absorption, distribution, metabolism, and excretion of DBP are of
oral exposure studies. Following oral exposure, DBP is hydrolyzed in the gut to the bioactive phthalate
monoester, MBP, and rapidly absorbed in the gastrointestinal tract. MBP is broadly distributed
throughout the body, and minimal bioaccumulation occurs. MBP and MBP-G are the predominant
metabolites in humans and rodents. Most of the administered dose of DBP is excreted in urine within 24
hours, and a small proportion is also eliminated in the feces. DBP and its metabolites can cross the
placenta to the developing fetus.
The reasonably available data on other routes of exposure are sparse, especially for inhalation. Studies
that do exist for dermal routes of exposure suggest dermal absorption of approximately 11 percent and
indicate a prerequisite for maximal absorption is hydrolysis of DBP to MBP by serine esterases in the
skin (Sugino et al.. 2017). Inter-individual and intra-species differences exist across various
toxicokinetic parameters (e.g., species differences in skin thickness affect dermal absorption; differences
in metabolism) which may in turn impact toxicity.
Given the toxicokinetic information available for DBP, EPA will assume an oral absorption of 100
percent and an inhalation absorption of 100 percent for the draft risk evaluation. The approach
EPA used to estimate exposure via dermal routes of exposure is covered in the Draft Environmental
Release and Occupational Exposure Assessment for Dibutyl Phthalate (DBP) (U.S. EPA. 2024f) and
Draft Consumer and Indoor Dust Exposure Assessment for Dibutyl Phthalate (DBP) (U.S. EPA. 2024b).
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3 NON-CANCER HAZARD IDENTIFICATION
As was stated in Section 1.2.3, EPA is focusing its hazard identification on effects on the developing
male reproductive system. EPA evaluated non-cancer effects across epidemiological studies cited in
existing assessments and from literature published between 2014 to 2019, and NTP (2021). Other
hazards considered by EPA but not used for point of departure derivation, such as neurotoxicity,
metabolic effects, cardiovascular toxicity, immune adjuvant effects that were evaluated as part of EPA's
further filtering process are presented in Section 3.1.3.
The sections below focus on hazard identification, characterization, and weight of evidence analysis of
on effects associated with the developing male reproductive system (3.1.2.1), which are the most
sensitive human health hazard outcomes associated with oral exposure to DBP in laboratory animals.
Several studies have also evaluated the effects of DBP exposure on the nervous system, cardiovascular
system, immune system, and metabolism. Although the data on the health effects on animals following
developmental exposures is abundant, data following chronic exposure durations to adult animals is
limited to one well-conducted NTP technical report (NTP. 2021) with 2-year studies in mice and rats
that provide far less sensitive LOAELs (i.e., above 500 mg/kg-day) than the developmental studies. In
the draft risk evaluation of DBP, effects on the developing male reproductive system form the basis of
the POD used for acute, intermediate, and chronic exposure scenarios.
3.1 Effects on the Developing Male Reproductive System
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 DBP 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). Sections 3.1.1.1.1, 3.1.1.1.2, and 3.1.1.1.3 provide further details on previous
assessments of DBP by Health Canada (2018b). Radke et al., (2019b; 2018) and NASEM (2017).
respectively, including conclusions related to exposure to DBP and health outcomes. Additionally, EPA
also evaluated new epidemiologic studies published after the Health Canada (2018b) assessment (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).
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876 Table 3-1. Summary of Scope and Methods Used in Previous Assessments to Evaluate the
877 Association Between DBP 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 (e.g., low birth weight)
• Male infant genitalia (e.g.,
hypospadias/cryptorchidism)
• Placental development and gene
expression
• Preterm birth and gestational age
• Postnatal growth
• DNA methylation
Reproductive:
• Altered male puberty
• Gynecomastia (i.e., the increase of male
breast glands in pubescent boys)
• Changes in semen parameters
• Sexual dysfunction (males)
• Sex ratio
Downs and Black (1998)
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)
(NTP. 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.
878
879 3.1.1.1.1 Health Canada (2018b)
880 Health Canada (2018b) considered 83 studies that evaluated the association between DBP and its
881 metabolites (MBP/MnBP) and reproductive outcomes. The outcomes that were evaluated are listed in
882 Table 3-1. Female reproductive outcomes were also evaluated by Health Canada (e.g., altered female
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puberty, pregnancy complications and loss, altered fertility and time to pregnancy, endometriosis and
adenomyosis, uterine leiomyoma, sexual dysfunction in females, polycystic ovary syndrome, age at
menopause).(2018b). Health Canada considered associations with prenatal, perinatal, and adult
exposures.
Health Canada evaluated studies that looked at individual phthalates (or their metabolites) and health
outcomes, due to the challenging nature of interpreting results for the sum of several phthalates. 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.
There was limited evidence1 for the association between DBP and its metabolites and sperm DNA
damage/apoptosis. uterine leiomyoma, and sex ratio at birth. There was inadequate evidence for the
association between DBP and its metabolites and sexual dysfunction in males and females, polycystic
ovary syndrome, and age at menopause. The level of evidence could not be established for the
association between DBP and its metabolites and altered fertility. There was no evidence for the
association between exposure to DBP and its metabolites and endometriosis and adenomyosis. All other
reproductive outcomes (i.e., altered male or female puberty, gynecomastia, pregnancy complication and
loss) did not have reported evidence of association with DBP and/or its metabolites.
Sixty-five studies were assessed by Health Canada (2018b) to evaluate the association between exposure
to DBP and growth and developmental outcomes. These studies evaluated outcomes such as AGD, birth
measures, male infant genitalia, placental development and gene expression, preterm birth and
gestational age, as well as postnatal growth and DNA methylation. There was inadequate evidence of
association for DBP and its metabolites and the following outcomes: birth measures, placental
development, preterm birth and gestational age, postnatal growth and postnatal DNA methylation. There
was no evidence of association for DBP and its metabolites and AGD. Health Canada (2018b) did not
report evidence of an association between exposure to DBP and altered development of male infant
genitalia (e.g., hypospadias and cryptorchidism).
The relationship between DBP and its metabolites and the human endocrine system was investigated in
48 studies by Health Canada (2018b). Effects on thyroid-related hormones, sex hormones, and other
hormones were the three categories used to evaluate the hormonal effects. The authors found that there
was limited evidence for association between MBP/MnBP with sex hormone levels (i.e., follicle
stimulating hormone, luteinizing hormone, testosterone, estradiol, prolactin, inhibin B, anti-Mullerian
hormone, androstenedione). There was inadequate evidence for association between MBP/MnBP and
thyroid-related hormones or growth hormone homeostasis.
3.1.1.1.2 Radke et al. (2019b: 2018)
Systematic reviews conducted by Radke et al. used in this assessment include male (2018) and female
(2019b) developmental and reproductive outcomes. Radke et al. (2018) evaluated the associations
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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between DBP or its metabolite (MBP) 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 (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 (DEP, DBP, DIBP, 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.
Radke et al. found the strongest inverse relationship between AGD and urinary MBP in a study reported
by Bornehag et al. (2014). Inverse associations were also observed by Swan et al. (2015) and Swan
(2008). the latter of which was statistically significant. An additional two birth cohort studies by Suzuki
et al. (2012) and Jensen et al. (2016) reported no association. Overall, Radke et al. (2018) concluded that
there was moderate evidence of an association between AGD and DBP exposure based on these five
studies. Inverse associations were found between exposure to DBP and its metabolites and sperm
parameters, including sperm concentration (8 of 12 studies) and sperm morphology (7 of 12 studies).
Three of the studies (Wang et al.. 2015; Liu et al.. 2012; Hauser et al.. 2006). found statistically
significant and monotonic dose-response associations with sperm concentration, while two found
statistically significant inverse associations with sperm motility (Axelsson et al.. 2015a; Hauser et al..
2006). The results for semen parameters were noted throughout the entire spectrum of exposures noted
in the research The studies with lower exposure levels were more likely to indicate an association than
those with higher levels. Ten studies assessed sperm morphology, six of which support an association
with DBP. Biological plausibility for the association between exposure to DBP and semen parameters is
provided by Jurewicz et al. (2013). who showed increased sperm aneuploidy with higher DBP exposure.
However, not all studies reported associations for all sperm parameters. Indeed, one investigation
spanning two studies found no association between DBP exposure and sperm apoptosis (Wang et al..
2016; You et al.. 2015). Overall, the evidence of an association between higher DBP exposure and lower
semen quality, specifically sperm concentration, was robust because it is consistent across many
medium confidence studies and shows dose-response associations.
Table 3-2. Summary of Epidemiologic Evidence of Male Reproductive Effects Associated with
Exposure to DBP (Radke et
al.. 2018)
Timing of Exposure
Outcome
Level of Confidence in Association
In utero
Anogenital distance
Moderate
Hypospadias/cryptorchidism
Slight
In utero or childhood
Pubertal development
Indeterminate
Adult
Semen parameters
Robust
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Timing of Exposure
Outcome
Level of Confidence in Association
Time to pregnancy
Moderate
Testosterone
Slight
Male Reproductive Outcomes Overall
Robust
Data for DBP are taken directlv from Fieure 3 in Radke et al. (2018)
Time to pregnancy following male exposure to DBP was evaluated by one study (Buck Louis et al..
2014). which reported statistically significant associations between higher exposure to the DBP
metabolite, MBP, and either a longer time to pregnancy or a lower fecundity. The evidence is deemed
moderate due to the high degree of confidence in the study and its coherence with semen parameters.
Ten studies (Axelsson et al.. 2015b; Chang et al.. 2015; Den Hond et al.. 2015; Pan et al.. 2015; Wang et
al.. 2015; Han et al.. 2014; Meeker and Ferguson. 2014; Jurewicz et al.. 2013; Meeker et al.. 2009a; Pan
et al.. 2006) are used to evaluate the relationship between exposure to DBP as measured by MBP and
testosterone. Results from five studies (Pan et al.. 2015; Wang et al.. 2015; Meeker and Ferguson. 2014;
Meeker et al.. 2009a; Pan et al.. 2006) show that higher exposure to DBP is associated with lower
testosterone levels. Only Pan et al. (2015). found a statistically significant association with DBP and
testosterone. There was no discernible pattern linking the observed relationships to the range or intensity
of exposure. Thus, Radke et al. (2018) regarded the evidence for the association between DBP and
testosterone as slight.
Radke et al. (2019b) also evaluated the association between DBP and its metabolite (MBP) and female
reproductive and developmental outcomes. Four studies (two using childhood exposure measurements,
two using prenatal exposure measurements) examined the association between pubertal development
and DBP. Later age at pubarche (for at least one measure) was reported by two studies (Wolff et al..
2014; Mouritsen et al.. 2013) following childhood exposure to DBP and its metabolites, but the findings
were inconsistent among the measure. One study (Watkins et al.. 2017) found inconsistent results for in
utero exposure in terms of age of menarche (a later age with higher MBP exposure) and pubic hair
stages (an earlier age with higher exposure), the latter of which also disagreed with the findings for
exposure during childhood. Overall, there is indeterminate evidence about the association between DBP
exposure on pubertal development due to inconsistencies and lack of coherence among associated
measures of puberty. In four studies (Machtinger et al.. 2018; Wu et al.. 2017; Hauser et al.. 2016;
Messerlian et al.. 2015). there were decreases in outcomes related to time to pregnancy in women
undergoing in vitro fertilization in at least one secondary outcome related to DBP. However, because
there was no association found for the primary fecundity outcomes (time to pregnancy and rate of
clinical pregnancy), evidence of a relationship between fecundity and exposure to DBP is deemed
indeterminate. Five studies serve as the basis for evaluating the evidence of an association between
spontaneous abortion and DBP exposure. A high confidence study by Jukic et al. (2016) reported
slightly higher odds ratios between MBP exposure levels and early pregnancy loss (tertile 2 OR
[95% CI] = 1.1 [0.47, 2.58]; tertile 3 OR [95% CI] = 1.12 [0.46, 2.74]). Toft et al. (2012). reported an
inverse association between exposure and clinical pregnancy loss and a monotonic increase in OR for
early loss. A case-control study by Mu et al. (2015) found an inverse relationship between quartiles 2
and 3 and quartile 1, but an increased OR for clinical loss for quartile 4 compared to quartile 1. The
associations that were reported were not statistically significant. Neither Yi et al. (2016) nor the high-
confidence study by Messerlian et al. (2016). found an association between exposure to DBP and
spontaneous abortion. The effect estimates for early loss were modest and not statistically significant,
while the results for clinical loss were inconsistent. Overall, due to the inconsistency among the high
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confidence studies, Radke et al. (2019b) concluded that there is slight evidence of association between
early spontaneous abortion and DBP exposure.
Radke et al. (2019b) also evaluated six pregnancy cohort studies (two being nested cohort studies within
a case-control design) that provided information on the associations between preterm birth and exposure
to DBP and its metabolites. Two studies examined gestational duration (Polanska et al.. 2016; Watkins
et al.. 2016) and the remaining studies examined preterm birth (Smarr et al.. 2015; Ferguson et al.. 2014;
Meeker et al.. 2009b). Three studies (Casas et al.. 2016; Ferguson et al.. 2014; Meeker et al.. 2009b)
found increased odds of preterm birth with increasing DBP exposure, including two high confidence
studies. Meeker et al. (2009b) reported high OR and both Ferguson et al. (2014) and Meeker et al.
(2009b) reported statistically significant results. Overall, Radke et al. (2019b) found moderate evidence
of an association between DBP exposure and preterm birth, despite some inconsistencies across studies.
3.1.1.1.3 NASEM report (2017)
NASEM (2017) also evaluated the associations between in utero exposure to DBP and male
reproductive outcomes. 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 DBP, 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 DBP 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 DBP is associated with a decrease in fetal testosterone
in males, given the various different 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. However, consistent with the conclusions of Radke et al. (2018) NASEM also
concluded that there was moderate evidence of association between DBP and AGD. The AGD effect
estimates in the meta-analysis NASEM (2017) (% change [95% CI] =-3.13 [-5.63, -0.64] [p = 0.04])
are slope estimates based on the assumption that exposure and effect have a monotonic dose-response
relationship.
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 link between DBP exposure and male reproductive outcomes. Radke et al.
(2018). and NASEM (2017) concluded that there was an association between exposure to DBP and
decreased AGD, while Health Canada (2018b) did not. The scope and purpose of the assessments by
Health Canada (2018b). and systematic review articles by Radke et al. (2018). and NASEM (2017)
differ from that of Health Canada related to their moderate confidence conclusions drawn for AGD,
which may be related to the different conclusions. Health Canada (2018b) was the most comprehensive
review, considering pre 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 (discussed further in 4.2). The results of the animal and epidemiological systematic
reviews were considered together by NASEM (2017) to draw hazard conclusions. Each of the existing
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assessments covered above considered a different number of epidemiological outcomes and used
different data quality evaluation methods for risk of bias. Despite these differences, each assessment
provides qualitative support as part of the weight of scientific evidence.
3.1.1.2 Summary of New Studies Identified by EPA (2018 through 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 45 new developmental
(26 studies) and reproductive (19 studies) epidemiology studies published between 2018 to 2019.
Fourteen of those studies were female reproductive outcomes (1 high confidence, 11 medium
confidence, 1 low confidence and 1 uninformative) and 5 medium studies were male reproductive
outcomes. Of the forty-five studies, five medium confidence studies evaluated male reproductive
outcomes and 26 studies evaluated male developmental outcomes (2 high confidence, 18 medium
confidence and 6 low confidence). Thirty-three studies found no association between exposure to DBP
or its metabolites and developmental and reproductive outcomes. In contrast, two medium confidence
male reproductive studies found a significant association between exposure to DBP or its metabolites
while seven male developmental studies (1 high confidence; 3 medium confidence; 3 low confidence)
found a significant association between exposure to DBP or its metabolites. Studies reporting an
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 Dibutyl
Phthalate (DBP) (U.S. EPA. 2024e)
• Draft Data Extraction Information for Environmental Hazard and Human Health Hazard
Animal Toxicology and Epidemiology for Dibutyl Phthalate (DBP) (U.S. EPA. 2024c).
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. A medium confidence study Arbuckle et al. (2018) reported a
significant positive association between prenatal first trimester urinary MBP and anopenile distance in
Canadian male infants at birth (beta [95% confidence interval] for the change in anopenile distance
(millimeters) per In- unit increase in MBP: 1.1689 [0.0207, 2.317]). One medium confidence study by
Zhang et al. (2018b) reported a significant positive association between maternal urinary MBP during
first, second and third trimesters and birth weight in male infants in the normal birth weight group (beta
[95% CI] for change in birth weight per unit increase in MBP = 10.438 [0.502, 2.0374]). Another
medium confidence study by Burns et al. (2022) reported a significant positive association between pre-
pubertal urinary MBP and pubertal onset (as measured by pubic hair development) in boys in the fourth
quartile of compared to the first quartile of MBP [beta (95% CI) for Q4 vs. Ql= 9.3 (1.5, 17.1)];
associations were positive but not significant for other quartiles, although the trend test was significant (
p-value = 0.03). No other indicators of pubertal development had significant results.
Developmental Outcomes for Females. A high confidence study by Bloom et al. (2019) reported a
significant negative association between maternal urinary MBP at gestational weeks 18 through 22 and
24 through 32 and odds of low birth weight in female infants only. A medium confidence study by
Arbuckle et al. (2018) reported a significant positive association between prenatal first trimester urinary
MBP and right-hand digit ratio (ratio of the lengths of the second and fourth finder digits of the right
hand) in female infants at six months (beta [95% confidence interval] for the change in hand digit ratio
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per ln-unit increase in MBP: 0.0122 [0.0018, 0.0227]). Another medium confidence study, Bloom et al.
(2019) considered MIBP to be a metabolite of DBP rather than DIBP. The study reported a significant
positive association between maternal urinary MiBP at gestational weeks 24-32 and odds of small for
gestational age (OR [95% CI] per ln-unit increase maternal urinary DBP = 2.82 [1.21, 6.56]). Results for
MBP were not statistically significant. No significant findings were found for other female reproductive
outcomes such as anthropometric measures of female reproductive organs, fecundity/increased time to
pregnancy, female reproductive hormones and uterine fibroids.
Other Developmental Outcomes. A low confidence study by Amin et al. (2018) reported significant
positive associations between urinary MBP and BMI z-score (beta = 0.22; p-value < 0.001) and waist
circumference (beta = 0.29; p-value < 0.001) in Iranian children and adolescents. Another low
confidence study by Durmaz et al. (2018) reported significant positive correlations between urinary
MBP and weight (Spearman correlation coefficient = 0.550; p-value < 0.01) and BMI (Spearman
correlation coefficient = 0.611; p-value < 0.01) in 4- to 8-year-old Turkish girls. A medium confidence
study by Boss et al. (2018) reported a significant positive association between maternal urinary MBP
throughout pregnancy and gestational age at delivery (HR [95% CI] for change in gestational age per
IQR increase in urinary MBP = 1.17 (1.05, 1.29)]. No significant associations were observed for risk of
preterm birth in this study. No significant findings for birth measures (placental). No significant findings
were found for fetal loss.
Reproductive Outcomes for Males. Another medium confidence study by Tian et al. (2018) reported a
significant positive association between urinary MBP and urinary androstenedione levels among healthy
reproductive-age men in Xiamen, China (beta [95% confidence interval] for the change in ln-
androstenedione per ln-unit increase in MBP: 0.35 [0.11, 0.60]). No significant findings were found for
other male reproductive outcomes such as sperm quality parameters and biomarkers of prostate health.
EPA concurs with the conclusions of Health Canada (2018b) systematic review articles published by
Radke et al. (2018) and NASEM (2017) that there is some evidence of association but not enough to
conclude a causal relationship between DBP exposure and developmental and reproductive outcomes.
Moreover, new studies identified by EPA from 2018 to 2019 do not alter the previous conclusions from
Health Canada (2018b). NASEM (2017). and systematic review articles published by Radke et al.
(2018). Although there is moderate level of confidence in the association between DBP and health
outcomes such as AGD and time to pregnancy, discussed above, causality was not established.
Therefore, EPA preliminarily concludes that the existing epidemiological studies do not support
quantitative exposure-response assessment due to uncertainty associated exposure characterization of
individual phthalates, including source or exposure and timing of exposure as well as co-exposure
confounding with other phthalates, discussed in 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 considered 52 studies across 39 publications that examined effects on the developing male
reproductive system following oral exposure to DBP, including prenatal and perinatal exposure studies,
and multi-generational studies of reproduction (Table 3-3; Table 3-4). All 52 of these studies were
identified because they were key studies considered in dose-response analyses in previous assessments
and the endpoints are consistent with phthalate syndrome. While EPA identified additional studies
through systematic review, none were included for further analysis due to study limitations (see
discussion in Sections 1.2.3 and 3.1.3). No studies evaluating effects on the developing male
reproductive system following exposure to DBP are available for the dermal or inhalation exposure
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routes. Studies that have evaluated male reproductive outcomes following developmental exposure to
DBP are discussed in Section 3.1.2.1. Other developmental and reproductive outcomes, such as changes
in fetal body weight or reproductive organ weight, post-implantation loss, resorptions, or skeletal
variations, are discussed in Section 3.1.2.2. Data from chronic studies of DBP are limited in sensitivity
compared to the database of developmental exposure studies. Data from chronic studies of DBP are
limited to one well-conducted NTP technical report (NTP. 2021) with 2-year studies in mice and rats
that provide far less sensitive LOAELs (i.e., above 500 mg/kg-day) than the developmental studies
compared to the database of developmental exposure studies. Indeed, there is one study available: the
technical report by the NTP (2021) that evaluated the toxicity of DBP in mice and rats exposed for up to
2 years. In rats exposed to 510 mg/kg-day, NTP reported increased gross findings (cryptorchidism,
agenesis, small testis), increased microscopic findings in the testes (e.g., seminiferous tubule dysgenesis,
Ley dig cell hyperplasia, and hypospermia), increased incidence of hepatocyte alteration in the liver of
males and females, and increased incidence of hypertrophy in the pars distalis in males.
3.1.2.1 Developing Male Reproductive System
As part of the Draft Proposed Approach for Cumulative Risk Assessment of High-Priority and a
Manufacturer-Requested Phthalate under the Toxic Substances Control Act, EPA has previously
considered the weight of evidence and concluded that oral exposure to DBP 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 Science Advisory Committee on Chemicals (SACC) (U.S. EPA. 2023b). A summary
of the MOA for phthalate syndrome and data available for DBP supporting this MOA is provided in
3.1.3. Readers are also 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 DBP'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.
Three studies evaluated effects on the developing male reproductive system following prepubertal or
pubertal exposures to DBP (Moody et al.. 2013; Xiao-Feng et al.. 2009; Srivastava et al.. 1990). Of
these, only Moody et al. (2013) was considered for dose-response analysis in Section 4 because of
methodological limitations in the latter two (i.e., qualitative histopathological assessment of testes).
Additionally, Srivastava et al. (1990) received an uninformative study evaluation rating.
There is a robust database showing adverse effects on the male reproductive system following
developmental exposure to DBP in rats. Adverse effects include decreased fetal testis testosterone,
histopathological alterations in the testis, decreased anogenital distance, increased male nipple retention
gross malformations of the male reproductive tract (e.g., undescended testes, hypospadias, etc.), and
sperm parameters. EPA identified 40 publications of oral exposure studies that have evaluated at least
one of these effects, 36 of which are oral exposure studies following in utero exposure to DBP (34
studies of rats; one in mice; one in marmosets; Table 3-3), and four of which follow pubertal exposures
(2 in rats, 1 in mice; Table 3-4). One publication in rabbits evaluated in utero and prepubertal exposure
to DBP in two separate experiments. Studies considered by EPA are summarized in Table 3-3 and Table
3-4, including study findings and limitations. Effects on the developing male reproductive system in the
context of the mode of action for phthalate syndrome is further discussed in 3.1.3.
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3.1.2.2 Other Developmental and Reproductive Outcomes
In addition to effects on the developing male reproductive system, developmental exposure to DBP has
been associated with other developmental and reproductive effects in experimental animals. These
include decreases in litter size, changes in sex ratio, increases in pup mortality, decreases in fetal
weight, resorptions, post-implantation loss, and increase in skeletal variations (Table 3-3; Table 3-4).
These effects generally, but not exclusively, occur at higher doses than those that elicit effects on the
developing male reproductive system. Indeed, the majority of studies reviewed by EPA that observed
developmental effects other than those on the male reproductive system observed them at doses
ranging from 500 to 712 mg/kg-day or higher (Giribabu et al.. 2014; Kim et al.. 2010; Drake et al..
2009; Li et al.. 2009; Jiang et al.. 2007; Lee et al.. 2004; Ema et al.. 1998; Mylchreest et al.. 1998).
Nevertheless, there are two studies that reported decreased pup body weights at lowest doses of around
250 to 400 mg/kg-day. A multigeneration study by the NTP (reported by (Wine et al.. 1997)) exposed
pregnant rats to dietary concentrations of DBP for equivalent to 52, 256, 509 mg/kg-day [males] or 80,
385, or 794 mg/kg-day [females]. The bodyweights of F1 pups (both absolute and adjusted for litter
size) from exposed females were decreased in the mid and high dose groups (LOAEL = 385 mg/kg-
day). The body weights of female F1 pups from the high dose group were decreased (10 to 15 percent)
on PND0, 14, or 21. F2 pup body weights were significantly decreased at birth in all exposure groups,
and 6 percent decreased from controls at the low dose (equivalent to 80 mg/kg-day). Zhang et al.
(2004) also reported decreased pup body weights. In that study, pregnant SD rats were exposed to 0,
50, 250, or 500 mg/kg-day DBP via gavage from GDI to PND21. Pup body weight at birth was
decreased in the 250 mg/kg-day dose group, which coincided with decreased male AGD on PND4 as
well as reductions in sperm motility and absolute epididymis weight in PND70 adults. No changes in
PND70 body weight were observed, indicating that the decrease in pup body weight at birth was not
permanent. Other unaffected outcomes included sex ratio and pup survival to weaning.
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1219 Table 3-3. Summary of Studies Evaluating Effects on the Developing Male Reproductive System Following In Utero Exposures to
1220 DBP
Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
(Lee et al..
2004)
Pregnant rats (6-8
dams/group) were
exposed to 0, 20, 200,
2000, or 10,000 ppm
DBP via diet from GDI5
-PND21 (equivalent to
0, 1.5-3, 14-29, 148-
291, 712- 1372 mg/kg-
day). Male and female F1
offspring were evaluated
at PND 2, PND14,
PND21, and PNW 8-11
and PNW20.
ND/3
I spermatocyte
development on PND 21
and t vacuolar
degeneration of alveolar
cells and alveolar atrophy
of mammary gland in PNW
II males
Maternal Effects
-1 BW gain on GDs 15-20 (712 mg/kg-day)
Other Developmental Effects
-1 male:female ratio (712 mg/kg-day)
-1 absolute AGD (males) on PND2 & | male NR on PND 14 (712
mg/kg-day)
-1 relative liver (both sexes) & J. testes weight on PND21
(712 mg/kg-day)
- Testicular pathology on PND 21 (aggregated foci of Leydig cells and
decreased epididymal duct cross section at >148 mg/kg-day);
Testicular pathology on PNW 11 (loss of germ cell development at
>148 mg/kg-day)
Unaffected Outcomes
- Dam BW gain on PND 2 - PND 21; food consumption; # live
offspring; offspring BW on PND 2; F1 relative kidney, adrenal,
epididymis, ovary, uterus weight on PND 21; PPS; vaginal opening;
estrous cyclicity; testicular pathology on PNW 20
Limitations:
-Individual animal was the statistical unit, not the litter; Small sample
size; Insufficient methodological details provided regarding
histopathology; outcome measure timing concerns (i.e., male rats just
beginning to develop spermatocytes around PND21)
(Boekelheide et
al. 2009)
Pregnant SD rats (4-10
litters/group) gavaged
with 0 0.1, 1, 10,30,50,
100, 500 mg/kg-day DBP
on GD 12-21.
10/30
t testicular pathology (J.
testicular cell number;
disorganized seminiferous
tubules)
Developmental Effects
-1 number of tubular cross sections (50, 100, 500 mg/kg-day)
-1 cell proliferation on GD20 & GD21 (500 mg/kg-day)
-1 number of MNGs (>100 mg/kg-day)
Limitations:
- Qualitative histopathology (no incidence data provided)
(Mahood et al..
Pregnant Wistar rats (4-6
litters/group) gavaged
with 0, 4, 20, 100, 500
mg/kg-day DBP on GD
13.5-20.5 (fetal tissue,
for endpoints of testicular
testosterone, MNGs, LC
20/100
I fetal testicular
testosterone content, t
MNGs, t Leydig cell
aggregation
Developmental Effects
2007)
-1 testicular testosterone content on GD 21.5 (>100 mg/kg-day)
-1 MNGs on GD 21.5 (>100 mg/kg-day)
- Changes in Leydig cell distribution (i.e., [ # of total Leydig cell
clusters, t occurrence of medium (100 mg/kg-day) and large (500
mg/kg-day) Leydig cell clusters)- increased dysgenic areas (not
statistically significant)
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Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
distribution) or GD 13.5—
21.5 (postnatal tissue for
endpoints of infertility,
cryptorchidism, testis
weights).
Pregnant Wistar rats
gavaged with 0, 4, 20,
100, 500 mg/kg-day DBP
on GD 13.5-21.5. "
100/500
t infertility,
cryptorchidism, [ testis
weight
Developmental Effects
-1 incidence of infertility (i.e., male produce offspring with untreated
females) and cryptorchidism on PND 90 (500 mg/kg-day)
-1 incidence of Sertoli cell only tubules (SCO) in cryptorchid testes
(>100 mg/kg-day; 11/11 animals at 500 mg/kg-day) and increased
incidence of SCO tubules in scrotal testes (>20 mg/kg-day; flat dose-
response)
-1 absolute testis weight on GD 21.5 and PND 90 (500 mg/kg-day)
(Furr et al..
2014)c
Pregnant Harlan SD rats
(3-4/dose) gavaged with
0, 1, 10, 100 mg/kg-day
DBP on GDs 14-18.
Dams sacrificed on GD
18. (Block 22)
10/100
I ex vivo fetal testicular
testosterone production
(36%)
Unaffected Outcomes
- Dam weight gain; fetal viability
Pregnant Harlan SD rats
(2-3/dose) gavaged with
0, 33,50, 100,300
mg/kg-day DBP on GDs
14-18. Dams sacrificed
on GD 18. (Block 18)
50/100
I ex vivo fetal testicular
testosterone production
(35%)
Unaffected Outcomes
- Dam weight gain; fetal viability
Pregnant Harlan SD rats
(3-4/dose) gavaged with
0, 1, 10, 100 mg/kg-day
DBP on GDs 14-18.
Dams sacrificed on GD
18. (Block 26)
100/ND
NA, no LOAEL identified
Unaffected Outcomes
- Dam weight gain; fetal viability; ex vivo fetal testicular testosterone
production
Pregnant Harlan SD rats
(3-4/dose) gavaged with
0, 750 mg/kg-day DBP
on GDs 14-18. Dams
ND / 750
I ex vivo fetal testicular
testosterone production
(89%)
Unaffected Outcomes
- Dam weight gain; fetal viability
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Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
sacrificed on GD 18.
(Block 34)
(Lehmann et al..
2004)
Pregnant SD rats (5-
7/dose) gavaged with 0,
0.1, 1, 10,30,50, 100,
500 mg/kg-day DBP on
GD 12-19.
30/50
I fetal testicular
testosterone content
Maternal Effects
- Not reported
Developmental Effects
-1 testicular mRNA & protein expression of genes involved in
steroidogenesis (e.g., StAR, P450scc, CYP17) (>50 mg/kg-day) and
testis descent (Insl3) (>500 mg/kg-day)
-1 fetal testicular testosterone content on GD 19 (>50 mg/kg-day)
Limitations/Uncertainties
^Authors state that the study was repeated, and a 30-mg/kg/day dose
group was included for the testosterone radioimmunoassay (RIA). For
other endpoints in this study, the 30 mg/kg-day dose group was not
included.
(Mvlchreest et
al.2000)
Pregnant SD rats (19-20
or 11 (high-dose) per
dose) gavaged with 0,
0.5,5,50, 100,500
mg/kg-day DBP on GDs
12-21.
50/100
t males with nipples and/or
areolae on PND 14
Maternal Effects
- None
Developmental Effects
-1 absolute AGD on PND 1 (500 mg/kg-day)
-1 absolute epididymal, dorsal prostate, LABC weight on PND 110
(500 mg/kg-day)
-1 hypospadias, absent or partial epididymis, vas deferens, SV and
prostate on PND 110 (500 mg/kg-day)
-1 Seminiferous tubule degeneration, interstitial cell hyperplasia,
interstitial cell adenoma (500 mg/kg-day)
Unaffected Outcomes
- Dam BW and food consumption; live pups per litter; sex ratio; birth
weight; survival to weaning; absolute liver, kidney, adrenal, testis, vas
deferens, SV, ventral prostate weight on PND 110; PPS; age at
vaginal opening
(MacLeod et al..
2010)
Pregnant Wistar rats (at
least 3 dams/dose)
gavaged with 0 or 500
mg/kg-day DBP on GD
13.5-16.5 and sacrificed
on GD17.5.
ND / 500
I fetal testicular
testosterone concentration
Developmental Effects
-1 fetal testicular testosterone concentration on GDI7.5 (500 mg/kg-
day)
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Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
Pregnant Wistar rats (at
least 3 dams/dose)
gavaged with 0 or 500
mg/kg-day DBP on GD
13.5-20.5 and sacrificed
on GD21.5.
ND/500
I fetal testicular
testosterone and J. AGD
Developmental Effects
-1 fetal testicular testosterone concentration on GD21.5 (500 mg/kg-
day)
-1 absolute AGD (male) on GD21.5 (500 mg/kg-day)
Pregnant Wistar rats (at
least 3 dams/dose)
gavaged with 0, 100, or
500 mg/kg-day DBP on
GD 13.5-21.5"and
sacrificed on PND 25.
ND/100
I ventral prostate weight
on PND25
Developmental Effects
-1 absolute SV and testis (500 mg/kg-day) and ventral prostate weight
(>100 mg/kg-day) on PND 25
- i penis length and absolute AGD on PND 25 (500 mg/kg-day)
Pregnant Wistar rats (at
least 3 dams/dose)
gavaged with 0 and 500
mg/kg-day DBP on GD
13.5-PND 15 and
sacrificed on PND 25.
ND / 500
I male AGD and penis
length
Developmental Effects
-1 male absolute AGD and penis length on PND 25 (500 mg/kg-day)
(Zhang et al..
Pregnant SD rats
(20/group) gavaged with
0, 50, 250, 500 mg/kg-
day DBP on GD 1-PND
21
50/250
I pup birth weight (12%
[males]; 9.8% [females]; J.
male AGD on PND 4
(absolute and BW
normalized), J. absolute
epididymis weight on PND
70; J. sperm motility and
total sperm heads per testis
on PND 70
Maternal Effects
2004)
- None
Developmental Effects
-1 live pups per litter (500 mg/kg-day)
-1 sperm number on PND 70 (500 mg/kg-day)
- Testicular pathology (i.e., small diameter tubules, degeneration or
exfoliation of the germinal epithelium of the seminiferous tubules
[250 mg/kg-day]; degeneration of seminiferous tubules, depletion of
germ cells [500 mg/kg-day])
Unaffected Outcomes
- Dam BW during pregnancy or lactation; gestation length; sex ratio;
pup survival to weaning; F1 male BW on PND 70; absolute testis,
prostate, pituitary weight on PND 70
Limitations
Qualitative histopathology (i.e., no incidence data)
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Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
(Giribabu et al..
2014)fe
Pregnant Albino Wistar
rats (6/group) were
gavaged with 0, 100, or
500 mg/kg DBP on GD 1,
7, and 14 (equivalent to 0,
21, or 107 mg/kg-day for
3 doses). PND100 F1
males (8/group) were
mated with unexposed
females to evaluate
reproductive
performance.
ND/100
I sperm count; sperm
motility; | percent
abnormal sperm; | serum
FSH & LH in F1 at
PND100; I serum
testosterone in F1 at
PND100; | levels of 17-
hydroxysteroid
dehydrogenase & 3-
hydroxysteroid
dehydrogenase in F1 at
PND 100; Abnormal testis
histopathology (i.e.,
disorganized seminiferous
tubules & t interstitial
spaces and ruptured
epithelium)
I No. of pups
I relative weight of the
seminal vesicle at PND 100
in F1
Maternal Effects
-1 number of pups delivered, mean number of live F2 fetuses (>100
mg/kg)
-1 number of resorptions in F2 on GD6 (>100 mg/kg)
Developmental Effects
-1 No. of pups (500 mg/kg)
-1 relative weight of the seminal vesicle at PND 100 in F1 500 mg/kg)
-1 sperm viability (500 mg/kg)
Unaffected Outcomes
- Fertility index; number of corpora lutea in F2 on GD6;
Developmental landmarks (e.g., pinna unfolding, eye opening) of Fl;
survival rate of Fl on PND4 and PND21; body weights & most organ
weights in Fl; skeletal system & external anomalies in F2 on GDI8
Limitations:
-Qualitative histopathology (incidence data not provided)
-Did not use litter as the statistical unit
(Therlmmunc
Research
Corporation.
Continuous breeding
protocol. Pregnant VAF
Crl:CD BR outbred
Sprague-Dawley albino
rats (20/sex/group; 40/sex
for controls) exposed to
0, 0.1, 0.5, or 1% DBP
via diet starting 10 weeks
prior to mating and
throughout gestation and
lactation periods
continuously for 2
generations (equivalent to
52, 256, 509 mg/kg-day
[males]; 80, 385, or 794
mg/kg-day [females]).
ND/80
- F2: I live pup weight (all
doses; not dose-
dependent);
- F1:1 live pups per litter
(dose-dependent)
Maternal Effects:
-1 body weight gain (11%) in PI females at week 17 (794 mg/kg-day)
Developmental Effects:
2002; Wine et
al. 1997: NTP.
1995)
- Fl: I number of live pups per litter (>385 mg/kg-day)
- Fl: I live pup weight (high dose [794 mg/kg-day] female x
unexposed male)
- Fl: Testicular pathology (i.e., degeneration of seminiferous tubules
[385 mg/kg-day]; interstitial cell hyperplasia and underdeveloped
epididymis [794 mg/kg-day]; sperm content reduction [794 mg/kg-
day],
- F2: I mating index, fertility index, pregnancy index (794 mg/kg-day
only)
Other Outcomes:
Fl: I terminal BW (13%) in females only (794 mg/kg-day)
F2: J. seminal vesicle weight (794 mg/kg-day only); | relative liver
weight (males; 794 mg/kg-day); t relative kidney (males;
>385 mg/kg-day)
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Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
F2: I terminal BW (males [7.9%] and females [13%]; 794 mg/kg-day)
Unaffected Outcomes:
-F1 fertility & average number of litters per pair; Fl: live pup weight
(high dose male x unexposed female)
(Li et al. 2009)
Pregnant Wistar rats (9-
10/dose) fed diets from
GD 6 - PND 28; diets
contained 0, 0.037, 0.111,
0.333, 1% DBP (31,94,
291, 797 mg/kg-day on
GD 6-21; 55; 165,486,
1,484 mg/kg-day on PND
0-15; 47, 140,433, 1,283
mg/kg-day on PND 16-
28).
94/291
I male absolute AGD on
PND1
Developmental Effects
-1 gestation length (797 mg/kg-day)
-1 male and female BW on PND 0, 7, 14, 21, 28 (797 mg/kg-day)
-1 relative liver weight (both sexes) and J. relative testes weight (797
mg/kg-day)
Unaffected Outcomes
- Live pups per litter; dam BW on GD 6-20; sex ratio; pinna
detachment; incisor eruption; eye opening
(Mvlchreest et
Pregnant SD rats
(10/dose) gavaged with 0,
100, 250, 500 mg/kg-day
DBP on GD 12-21.
100/250
| AGD on PND 1: | NR on
PND 14; epididymal
dysgenesis/ agenesis,
cryptorchidism, and
degeneration of
seminiferous epithelium in
F1 males on PND 100-105
Maternal Effects
al. 1999)
- None
Developmental Effects
-1 age at PPS (at 100 and 500, but not 250 mg/kg-day)
-1 hypospadias and prostate agenesis (>500 mg/kg-day)
-1 Interstitial cell hyperplasia or adenoma (>500 mg/kg-day)
-1 absolute kidney, testis, epididymis, SV weight in Fl offspring on
PND 100-105 (>500 mg/kg-day)
Unaffected Outcomes
- BW gain GD 0-21; BW during dosing; litter size; live pups per
litter; sex ratio; live pup weight on PND 1; offspring BW, absolute
liver, adrenal, vas deferens, prostate weight on PND 100-105
(Howdeshell et
al.2008)
Pregnant SD rats (3-
4/dose) gavaged with 0,
33, 50, 100, 300, 600
mg/kg-day DBP on GDs
14-18. Dams sacrificed
on GD 18.
100/300
I ex vivo fetal testicular
testosterone production
Maternal Effects
- None
Unaffected Outcomes
- # of dams with whole litter loss; maternal body weight gain; # of
implantations; # of live/dead fetuses; resorptions; fetal mortality
(Grav et al..
Pregnant Sprague-
Dawley rats (3-4 litters
group) exposed GD 14-
ND/300
I ex vivo fetal testicular
testosterone production
Other effects:
2021)c
- Dose-dependent reduction in genes involved in cholesterol
absorption (CYP461aj, cholesterol homeostasis (Ldlr), cholesterol
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Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
18 via gavage to 0, 300,
600, or 900 mg/kg-day
DBP (based on block 70
and 71 experiments)
(62% [block 70; 47%
[block 71])
biosynthesis (Cyp51, Dhcr24, Dhcr7, Ebp, Hmgcr, Hmgcsl, Idil,
Mvd, Nsdhl, RGD1564999, & Tm7sf2), or other functions in
cholesterol metabolism (Cypllal, Insigl) (>300 mg/kg-day)
Notes
- Testicular testosterone data for additional blocks of animals are
presented in Furr et al. (2014).
(Li etal.2015)
Pregnant Wistar rats (2-
5/dose) gavaged with 0,
100, 300, 900 mg/kg-day
DBP on GD 12.5-20.5
100/300
I testicular testosterone
concentration, Leydig cell
aggregation, J. AGD,
hypospadias, J. testis
weight
Developmental Effects
-1 testicular testosterone on GD17.5 (>300 mg/kg-day), GD19.5 (900
mg/kg-day), GD21.5 (900 mg/kg-day)
-1 Leydig cell aggregation on GD19.5 and GD20.5 (>300 mg/kg-day)
-1 absolute AGD (males) on PND2, PND21, PND63 (>300 mg/kg-
day)
-1 hypospadias (>300 mg/kg-day) and cryptorchidism (900) on PND
63
-1 absolute testis weight on GD17.5 (>300), GD19.5 (900 mg/kg-
day), GD21.5 (900 mg/kg-day)
(Martino-
Pregnant Wistar rats (7-
8/dose) gavaged with 0,
100, 500 mg/kg-day DBP
on GDs 13-21. Dams
terminated on GD21
(fetal study)
ND/100
I male AGD
Developmental Effects
Andrade et al..
2008)
-1 fetal testicular testosterone (63%) on GD 21 (500 mg/kg-day)
-1 MNGs, seminiferous cord diameter, Leydig cell aggregates on GD
21 (500 mg/kg-day)
-1 absolute AGD (males) (500 mg/kg-day) and AGD normalized to
cube root of BW (>100 mg/kg-day) on GD 21
Unaffected Outcomes
- Dam BW gain GD 12-21; implantation sites, post-implantation loss
Pregnant Wistar rats (4-
7/dose) gavaged with 0,
100, 500 mg/kg-day DBP
on GDs 13-21. Dams
allowed to deliver, and
offspring examined up to
PND90.
100/500
t male offspring NR on
PND13
Developmental Effects
-1 male pup NR on PND 13 (500 mg/kg-day)
- Unaffected Outcomes
- Dam BW gain GD 12-21; male F1 BW on PND 90; absolute testis,
epididymis, prostate, LABC, SV weight on PND 90; # of spermatids
per testis on PND 90; reproductive tract malformations; PPS
(Kuhl et al..
Pregnant SD rats
(10/dose) gavaged with 0,
100, 500 mg/kg DBP on
GD 18 and sacrificed 24-
hours later on GD 19.
100/500
I fetal testicular
testosterone concentration
(67%)
Developmental Effects
2007)
-1 fetal testicular mRNA levels oi' St A R, SR-B1, Cypllal, CYP17
(>100 mg/kg-day)
Not considered adverse at 100 mg/kg-day in absence of decreases in
fetal testosterone at this dose.
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Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
(Drake et al..
2009)
Pregnant Wistar rats (13-
15 dams/dose) gavaged
with 0, 100, 500 nig/kg -
day DBP from GD 13.5—
21.5 and reproductive
outcomes evaluated in
offspring at birth and
throughout adulthood.
100/500
I AGD during adulthood; J.
penis length, | hypospadias
& cryptorchidism, [
absolute testis and ventral
prostate weight
Maternal Effects
- Maternal effects not evaluated
Developmental Effects
-1 birth weight (8%; 500 mg/kg-day)
Pregnant Wistar rats (8-
17 litters/dose) gavaged
with 0, 500 mg/kg-day
DBP from GD 13.5-16.5
& sacrificed on GD 17.5.
ND / 500
I fetal intratesticular
testosterone, J. testicular
Star and Cypllal mRNA
Maternal Effects
Not reported
(Barlow et al..
Pregnant SD rats (10-
11/dose) gavaged with 0,
100, 500 mg/kg-day DBP
on GDs 12-21.
ND /100
| F1 males with NR on
PND13
Developmental Effects
2004)
-1 absolute AGD (male) on PND 1 and PND 180 (500 mg/kg-day)
-1 males with areolae on PND 13 (>100 mg/kg-day) and nipples on
PND 180 (500 mg/kg-day)
-1 incidence of gross lesions in testes (atrophied, enlarged, or absent),
epididymides (agenesis), vas deferens (absent), SVs ( mall or absent),
prostate (small or absent), penis (hypospadias) on PND 180, PND
370, PND 540 (500 mg/kg-day)
-1 testicular pathology (e.g., unilateral and/or bilateral testicular
dysgenesis and germ cell degeneration) on PND 180, PND 370, PND
540 (500 mg/kg-day)
(Scarano et al..
Pregnant Wistar rats
(5/group) gavaged with 0
or 100 mg/kg-day DBP
from GD 12 - PND21.
ND /100
Histopathological
abnormalities of fetal testis
(e.g., Leydig-cell clusters,
presence of MNGs, t
interstitial tissue area
relative to tubular area)
Maternal Effects
2010)
- Maternal effects not evaluated
Developmental Effects
-1 male AGD on PND4 (not statistically significant; 3.5 ± 0.2 mm vs.
3.1± 0.4 mm)
Unaffected Outcomes
- Serum testosterone levels in PND90 adults & in vitro testicular
testosterone from PND90 animals; male F1 body weight at PND1 and
PND90; sperm morphology and motility
(Struve et al..
Pregnant SD rats (7-
9/dose) fed diets
containing 0, 100, 500
ND / 112
I fetal testicular
testosterone (71%)
concentration on GD 20; j
Maternal Effects
2009)
- None
Developmental Effects
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Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
ppm (equivalent to 112,
582 mg/kg-day) DBP on
GDs 12-19. Dams
sacrificed on GD 19 or 20
(4- or 24-hours post-DBP
exposure).
Leydig cell aggregates and
seminiferous cord diameter
on GD 19 and 20
-1 absolute AGD (males) on GD 19 and 20 (500 mg/kg-day)
-1 fetal testicular testosterone concentration on GD 19 (500 mg/kg-
day) and GD 20 (>100 mg/kg-day)
-1 fetal testis mRNA levels for Star, Scarbl, Cypl 7al, P450scc!
Cyplla on GD 19 (>100) and GD 20 (500 mg/kg-day)
-1 MNGS (500 mg/kg-day)
Unaffected Outcomes
- Dam BW; litter size; sex ratio; fetal survival; fetal weights
(Ema et al..
1998)
Pregnant Wistar rats
(11/dose) fed diets
containing 0, 0.5, 1.0
2.0% (equivalent to 331,
555, 661 mg/kg-day)
DBP on GDs 11-21.
331/555
I AGD (absolute and BW
normalized) of male fetuses
on GD21; f incidence of
undescended testes
Maternal Effects
-1 BW gain and food consumption on GDs 11-21 (>555 mg/kg-day)
Developmental Effects
-1 fetal weight (661 mg/kg-day)
-1 incidence of cleft palate, fusion of stemebrae, fusion of ribs (661
mg/kg-day)
Unaffected Outcomes
- Resorptions; post-implantation loss; # live fetuses per litter; sex
ratio;
(Gaido et al..
2007)
Pregnant C57BL/6 mice
gavaged with 250 mg/kg
DBP on GD 16, 17, and
18.
ND/250
I seminiferous cord
formation and | MNGs
(quantitative
histopathology); t
seminiferous cord
diameter, MNGs per cord,
& nuclei/MNG
Maternal Effects
- Not evaluated for 250 mg/kg-day experiment; authors report
evidence of maternal toxicity at 1500 mg/kg-day in a preliminary
experiment.
Developmental Effects
- Changes in gene expression (j Btg2, Ctgf, Fos, Ier3, Nr4a, Pawr,
Tnfrsfl2a & \ Hsdllb2, Tkl at 4- & 8-hour timepoints (500 mg/kg
DBP on GDI8);
Unaffected Outcomes
- fetal testicular testosterone concentration
Limitations:
Qualitative histopathology for some endpoints
Insufficient information available to determine maternal toxicity
(Mvlchreest et
al. 1998)
Pregnant SD rats
(10/dose) gavaged with 0,
250, 500, 750 mg/kg-day
DBP on GD 3 - PND 20
ND/250
Reproductive tract
malformations
(hypospadias, non-scrotal
testes, epididymal
dysgenesis/ agenesis on
PND 100)
Maternal Effects
- None
Developmental Effects
- i male pup absolute AGD on PND 1 (>500 mg/kg-day)
- SV dysgenesis on PND 100 (>500 mg/kg-day)
-1 absolute testis and SV weight (>500) and epididymis and prostate
weight (750 mg/kg-day)
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Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
-1 live pups per litter and pup survival to weaning (750 mg/kg-day)
Unaffected Outcomes
- Dam BW and food consumption during pregnancy and lactation;
Dam absolute liver, kidney, adrenal, ovary, uterus weight on PND 21;
pup sex ratio; offspring BW on PND 1,21, 100 (both sexes); age at
vaginal opening; age at first estrus; length of estrous cycle
(Jiang et al..
Pregnant SD rats (10
dams/dose) gavaged with
0, 250, 500, 750, 1000
mg/kg-day DBP on GD
14-18. Dams allowed to
deliver pups naturally.
ND/250
t cryptorchidism
Maternal Effects
2007)
-1 maternal BW gain on GDs 14-18 and 18-20 (>750 mg/kg-day)
Developmental Effects
-1 live pups (>750 mg/kg-day)
-1 BW normalized AGD (males) on PND 1 (>500 mg/kg-day)
-1 hypospadias (>500) and cryptorchidism (>250 mg/kg-day) on PND
70
-1 BW, I relative liver, kidney, prostate, testis, epididymis, adrenal,
pituitary weight on PND 70 (>500 mg/kg-day)
Unaffected Outcomes
- Maternal mortality; relative heart and spleen weight on PND 70
(Kim et al..
2010)
Pregnant SD rats
(minimum of 3
dams/dose) gavaged with
0, 250, 500, 700 mg/kg-
day DBP on GD 10-19
and allowed to deliver
naturally.
ND/250
Delayed PPS
Developmental Effects
-1 BW and absolute testes, epididymis, ventral prostate, SV,
Cow pcr's gland, glans penis weight on PND 31 (700 mg/kg-day); J.
absolute LABC weight on PND 31 (>500 mg/kg-day)
-1 incidence of cryptorchidism and hypospadias on PND 11 (700
mg/kg-day)
-1 incidence of degeneration of seminiferous epithelium (700 mg/kg-
day)
-1 BW normalized AGD on PND 11 and t F1 male NR (>500 mg/kg-
day)
- i Serum DHT and total testosterone on PND 31 (700 mg/kg-day)
(Mylchreest et
Pregnant SD rats gavaged
with 0 and 500 mg/kg-
day DBP on GD 12-21.
ND / 500
I fetal testis testosterone on
GD 18 and GD 21,
testicular pathology
(Leydig cell hyperplasia,
testis atrophy, MNGs)
Developmental Effects
al. 2002)
- Leydig cell hyperplasia on GDs 16, 18, 21; testis atrophy on GD 18
and 21; MNGs on GD 21
(Howdeshell et
Pregnant SD rats (6/dose)
gavaged with 0 or 500
mg/kg-day DBP on GDs
ND / 500
I AGD, I LABC weight, J.
ex vivo fetal testicular
testosterone production
Developmental Effects
al.2007)c
-1 absolute AGD (males) on PND 3
-1 absolute LABC weight at 7-11 months of age
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Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
14-18 and allowed to
deliver pups naturally.
(34%), testicular
degeneration
- Low incidence of testicular malformations (not statistically
significant)
-1 ex vivo testicular testosterone production and mRNA for StAR on
GD 18
Unaffected Outcomes
- Maternal BW gain on GDs 14-18; litter size; fetal and neonatal
mortality; F1 BW on PND 3 (both sexes); # areolae per PND 14 male;
# nipples per adult male
(Ferrara et al..
2006)
Pregnant Wistar rats
gavaged with 0 or 500
mg/kg-day DBP on GD
15.5-21.5.
ND / 500
t MNGs and effects on
germ cell numbers
Developmental Effects
-1 incidence of MNGs in seminiferous cords on el9.5, e21.5, and
PND 4
-1 incidence of apoptotic gonocytes on el5.5, 17.5
-1 germ cell # per testis on e21.5, PND 4, PND8, PND15, PND25
- i germ cell proliferation index on PND 6 and PND 25
(Johnson et al..
2011)
Pregnant SD rats
(4/group) were gavaged
with 0 or 500 mg/kg DBP
from GD-12 - GD20 and
evaluated at GD 20.
ND / 500
I testicular testosterone
(34%);
i absolute male AGD; |
percentage of seminiferous
cords with one or more
MNGs
Maternal Effects
- None
Unaffected Outcomes
- Maternal body weights (qualitative statement in text)
(Johnson et al..
2007)
Pregnant SD rats gavaged
with a single dose of 1,
10, 100, or 500 mg/kg-
day DBP on GDI9 and
evaluated 1 hour after
dosing.
ND / 500
I intratesticular
testosterone (62%) on GD
19, 1 hour after dosing
Maternal Effects
- Not reported
Developmental Effects
- Altered gene expression of Ergl, Fos, Thbsl, CxcllO, Nr4al, Stcl,
Ednl, Tnfrsfl2ci, and ler3 from interstitial cells, Sertoli cells, and/or
peritubular myoid cells.
(Hiauchi et al..
Pregnant rabbits (8
litters/group) were
exposed to 0 or 400
mg/kg-day DBP from
GDI5-29 and male
offspring were evaluated
at PNW 6, 12, and 25.
ND / 400
I ejaculated sperm (43%);
t abnormal sperm; [
weight of testes (23%) and
accessory sex glands (36%)
on PNW12;
I serum testosterone at
PNW6 (32%);
Histopathological
alterations in the
Maternal effects:
2003)
- None
Other effects:
-Hypospadias, hypoplastic prostate, and cryptorchid testes with
carcinoma in situ-like cells in one male.
Unaffected Outcomes:
- Weight of epididymides (PNW12 and PNW25); weight of thyroid,
liver, or testes (PNW25); hypothalamic content of GnRH (PNW12 or
PNW25)
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Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
seminiferous tubule
epithelium and interstitium
of the testes (e.g.,
desquamated premature
germ cells)
(van den
Driesche et al..
2012)
Pregnant Wistar rats (3-7
litters/group) were
gavaged with 0, 500, or
750 mg/kg-day DBP from
GD 13.5-20.5 and AGD
and intratesticular
testosterone was
evaluated at GD21.5.
ND / 500
I intratesticular
testosterone on GD 21.5; [
absolute AGD (males) on
PND8
Maternal Effects
- Not reported
Developmental Effects
-1 Intratesticular testosterone on GD 21.5 (750 mg/kg-day)
-1 Focal testicular dysgenesis (t percentage of large & small Leydig
cell aggregates at PND8)
Pregnant Wistar rats (3-7
litters/group) were
gavaged with 0, 500, or
750 mg/kg-day DBP from
GD 19.5 - 20.5 and AGD
and intratesticular
testosterone was
evaluated at GD21.5.
ND / 500
I intratesticular
testosterone on GD 21.5; |
germ cell aggregation on
GD21.5
Maternal Effects
- Not evaluated
Developmental Effects
-1 intratesticular testosterone on GD 21.5 ( 750 mg/kg-day)
- focal testicular dysgenesis (t percentage of large & small Leydig cell
aggregates at PND8)
Unaffected Outcomes
- Male AGD (absolute) on PND8; focal testicular dysgenesis
(percentage of large & small Leydig cell aggregates at PND8)
(McKinnell et
al. 2009)
Pregnant marmoset
monkeys were exposed
from gestational week 7-
15 with 500 mg/kg-day
MBP, and male offspring
(11 offspring from 9
mothers) were evaluated
at birth (n=6) or later in
adulthood (n=5).
ND / 500
- Histopathological
alterations in the testes
(unusual clusters of
undifferentiated germ cells)
Unaffected Outcomes
- Gross testicular morphology; reproductive tract development;
testosterone levels at birth; germ cell number and proliferation, Sertoli
cell number, germ:Sertoli cell ratio
(Spade et al..
2018)
Pregnant SD rats (3-
6/dose) gavaged with 0
and 750 mg/kg-day DBP
on GDs 17-21
ND / 750
I ex vivo fetal testicular
testosterone production, |
incidence of MNGs
Unaffected Outcomes
- Litter size; resorptions; fetal loss; terminal maternal BW
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Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
(Wilson et al..
2004)
Pregnant SD rats (3/dose)
gavaged with 0 and 1,000
mg/kg-day DBP on GDs
14-18. Dams sacrificed
on GD 18.
ND/1,000
I testes testosterone
production and testicular
Ins/3 mRNA
Unaffected Outcomes
- Testis progesterone production
(Ema et al..
2000)
Pregnant Wistar rats (10-
13/dose) gavaged with 0,
1000, 1500 mg/kg DBP
on GDs 12-14.
ND/1000
I absolute AGD in male
pups on GD21; [ fetal body
weight (both sexes); J.
maternal body weight gain
and food consumption
I fetal body weight (both
sexes) (>1000 mg/kg-day)
I absolute AGD (males)
(>1000 mg/kg-day)
Maternal Effects
-1 maternal body weight gain and food consumption (>1000 mg/kg-
day)
Developmental Effects
-1 total litter resorptions (1500 mg/kg-day)
-I# live fetuses per litter (1500 mg/kg-day)
-1 fetuses with undescended testes (1500 mg/kg-day)
Unaffected Outcomes
- Sex ratio; AGD (females)
Considerations:
- Decreased fetal body weights may be attributed to decreased
maternal body weight gain and decreased food consumption.
Pregnant Wistar rats
(10/dose) gavaged with 0,
1000, 1500 mg/kg DBP
on GDs 18-20.
ND / 1000
| fetal BW and j AGD
(males) on GD21; [
maternal BW gain
Maternal Effects
-1 maternal BW gain (>1000 mg/kg-day) and food consumption
(1500 mg/kg-day)
Developmental Effects
-1 fetal weight (both sexes) (>1000 mg/kg-day)
-1 absolute AGD (males) (>1000 mg/kg-day)
-1 fetuses with undescended testes (1500 mg/kg-day)
Unaffected Outcomes
- Sex ratio; total litter resorptions; # of dead and live fetuses; # of
fetuses with undescended testes; AGD in females
Pregnant Wistar rats
(10/dose) gavaged with 0,
500, 1000, 1500 mg/kg
DBP on GDs 15-17.
ND / 500
t fetuses with undescended
testes and J. AGD on GD21
(absolute and BW
normalized)
Maternal Effects
-1 maternal BW gain and food consumption (>1,000 mg/kg-day)
Developmental Effects
-1 # of resorptions per litter (1500) mg/kg-day
-I# live fetuses per litter (1,500 mg/kg-day)
-1 fetal weight (both sexes) (1,500 mg/kg-day)
Unaffected Outcomes
- Sex ratio; total litter resorptions; AGD (females)
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Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
Abbreviations: [ = statistically significant decrease; t = statistically significant increase; ND = NOAEL or LOAEL not established; NOAEL = No observed adverse effect
level; LOAEL = lowest observed adverse effect level; GD = gestation day; PND = postnatal day; PNW = postnatal week; F1 = first-generation offspring; F2 = second-
generation offspring; AGD = anogenital distance; BW = body weight; LABC = levator ani plus bulbocavernosus muscles; MNGs = multinucleated gonocytes; LC = Leydig
cell; NR = nipple retention; PPS = preputial separation; SV = seminal vesicle; DHT = dihydrotestosterone; FSH = follicle stimulating hormone; LH = luteinizing hormone;
IHC = immunohistochemistry; StAR = steroidogenic acute regulatory protein; P450scc/ Cvpllal = cytochrome P450 family 11, subfamily a, polypeptide 1; CYP17 =
cytochrome P450 family 17; InsB = insulin-like hormone 3; SR-Bl/Scarbl = scavenger receptor class B member 1.
b Time-weighted doses calculated for 3 doses spanning 14 days (e.g., GDI, GD7, and GD14 = 3 doses; GD1-GD14 = 14 days of dosing; 100 mg/kg x 3 doses = 300 mg/kg/14
days = 21.4 mg/kg-day). Effects considered after a single dose for acute POD listed for the LOAEL.
"These studies were conducted by EPA's Office of Research and Development (ORD).
1221
1222
1223 Table 3-4. Summary of Studies Evaluating Effects on the Developing Male Reproductive System following Prepubertal and Pubertal
1224 Exposure to DBP
Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
(Xiao-Feng et
al.. 2009)
Male SD rats (8/group)
gavaged with 0, 250, 500,
1,000, or 2,000 mg/kg-day
DBP from PND35-PND65.
An additional recovery group
was maintained for 15
additional days after
cessation of DBP exposure.
ND / 250 (LOEL)
i Leydig cell number
(not considered
adverse)
Other Effects
- i serum testosterone (>500 mg/kg-day)
-1 serum glucocorticoid hormone (>1000 mg/kg-day)
- Histopathological changes in the testes (>500 mg/kg-day)
-1 gene expression of 11/3-HSD1 & Glucocorticoid Receptor: { StAR (>1000 mg/kg-
day)
- i relative weight of testes (<28%; >500 mg/kg-day) & epididymis weight (absolute
weight not reported)
Limitations:
- Qualitative histopathology (no incidence data provided)
Unaffected outcomes
- Body weight; relative adrenal weight
(Moodv et al..
2013)
Male and female C57BL/6
mice (< 6/group) gavaged
withO, 1, 10, 50, 100, 250,
or 500 mg/kg-day DBP from
PND 4-14
ND/ 1
Defective
spermatogenesis (t
incidence of partial
spermatogenesis); j
AGD relative to body
weight at adulthood; j
AGD relative to trunk
length at PND 14 &
adulthood
Other Effects
- Delayed spermatogenesis (J, cords containing pachytene spermatocytes
[ >10 mg/kg-day])
- i Serum testosterone (PND 14; 500 mg/kg-day); t serum inhibin alpha subunit
(PND 14; 500 mg/kg-day)
- Immature Sertoli cell and disorganization (100 mg/kg-day)
- i AGD relative to BW at PND 14 (500 mg/kg-day)
-1 relative heart weight (PND3; 500 mg/kg-day)
Limitations:
-No dose-response in AGD (absolute or normalized to BW on PND 14)
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Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
(Srivastava et
al. 1990)
Male Wistar albino rats
gavaged with 0, 250, 500, or
1,000 mg/kg-day DBP from
PNW5 - 7 (15 days).
ND / 250
t testes
histopathology (i.e.,
defective
spermatogenesis;
shrunken tubules in
testes); t activity of
enzymes in testes,
lactate
dehydrogenase, acid
phosphatase, and
glucose-6-phosphate
dehydrogenase; t
activity of enzymes in
testes, including
lactate dehydrogenase
Other effects:
-1 activity of enzymes in testes, including lactate dehydrogenase (dose-responsive,
beginning at 250 mg/kg-day), gamma-glutamyl transpeptidase (500, 1,000 mg/kg-
day), acid phosphatase, and glucose-6-phosphate dehydrogenase (dose-responsive,
beginning at 250 mg/kg-day)
- i absolute and relative testes weight (500, 1,000 mg/kg-day)
- i terminal BW (500 [19% change], 1,000 mg/kg-day [36% change])
Limitations:
- Qualitative histopathology
(Hisuchi et
al.. 2003)
Pregnant rabbits (8
litters/group) were exposed
to 0 or 400 mg/kg-day DBP
from PNW 4-12 and male
offspring were evaluated at
PNW 6, 12, and 25.
ND / 400
-Hypothalamic
content of GnRH
(PNW12 orPNW25)
- weight of accessory
sex organs at PNW 12
-1 abnormal sperm
Other effects:
- Hypospadias, hypoplastic prostate, and cryptorchid testes with carcinoma in situ-
like cells in one male
Unaffected Outcomes:
- Absolute organ weights including liver, kidney, thyroid, testes, and epididymides at
PNW12 orPNW25.
Abbreviations: I = statistically significant decrease; t = statistically significant increase; ND = NOAEL orLOAEL not established; NOAEL = No observed adverse effect level;
LOAEL = lowest observed adverse effect level; LOEL = lowest observed effect level; ND = no data; GD = gestation day; PND = postnatal day; PNW = postnatal week; AGD =
anogenital distance; BW = body weight; FSH = follicle stimulating hormone; 1 l/3-HSDl=l 1 [3-Hydroxysteroid dehydrogenase type 1; StAR =Steroidogenic acute regulatory
protein; GnRH = gonadotropin releasing hormone.
1225
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3.1.3 Mode of Action for Phthalate Syndrome
EPA previously developed a weight of scientific evidence analysis and concluded that oral exposure to
DBP can induce effects on the developing male reproductive system consistent with a disruption of
androgen action. The proposed MOA for phthalate syndrome is shown in Figure 3-1, which explains the
link between gestational and/or perinatal exposure to DBP and effects on the male reproductive system
in rats. The MOA has been described in greater detail in 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 ( J.S. EPA. 2023a) and is described briefly below. The MOA
underlying phthalate syndrome has not been fully established; however, key events at the cellular-,
organ-, and organism-level are generally understood (Figure 3-1).
Chemical Structure
and Properties
Phthalate
exposure during
critical window of
development
Cellular
Responses
Fetal Male Tissue
•J, AR dependent
mRNA/protein I
synthesis
^ x
Adverse Organism
Outcomes
Metabolism to
monoester &
transport to fetal
testes
Unknown MIE
(rot believed to be
AR or PPARa
mediated)
Key genes involved in the AOP \
for phthalate syndrome
Scarbl Cher7 Mvd Cla3b
StAP Ebp Nsdhl Insl3
Cypllal Fdps RGD1S64999 Lhcqr
Cypllbl Hmqcr Tm7sf2 Inha
Cypllb2 Hmgcsl Cyp46ol NrObl
Cypl7al HsdJb Ldlr RhoxlO
Cyp51 Fldil Insigl Wnt7a
s.
4/ Testosterone
synthesis
. V .
4/ Gene
expression
(INSL3, lipid
? metabolism,
cholesterol and
androgen synthesis
and transport)
3E
4> INSL3 synthesis
Fetal Leydig cell
V
Abnormal cell
apoptosis/
proliferation
(Nipple/areolae
retention, 4 AGD,
Disrupted testis
tubules, Leydig cell
clusters, MNGs,
agenesis of
reproductive tissues)
Suppressed
gubernacular cord
development
(inguinoscrotal phase)
4- Androgen-
dependent tissue
weights, testicular
pathology [e.g.,
seminiferous tubule
atrophy),
malformations (e.g.,
hypospadias), 4/
sperm production
£
(
Impaired
V-
fertility
Suppressed
gubernacular cord
development
(transabdominal
phase) J
Undescended
testes
Figure 3-1. Hypothesized Phthalate Syndrome Mode of Action Following Gestational Exposure
Figure taken directly from (U.S. EPA. 2023a) and adapted from (Conlev et al.. 2021; Gray et at.. 2021; Schwartz ejaL 202 J:
Howdeshell et al.. 2017).
AR = androgen receptor; INSL3 = insulin-like growth factor 3; MNG = multinucleated gonocyte; PPARa = peroxisome
proliferator-activated receptor alpha.
Molecular Initiating Event
The molecular events {i.e., the molecular initiating event) preceding cellular changes remain unknown.
Several studies have provided evidence against the involvement of androgen receptor antagonism and
peroxisome proliferator-activated receptor alpha (PPARa) activation (Gray et al.. 2021; Foster, 2005;
Foster et al.. 2001; Parks et al.. 2000). Other studies have suggested depletion of zinc concentration in
rodents (Gray et al.. 1982; Foster et al.. 1980). which could perturb the function of zinc-containing
proteins (e.g., zinc-finger transcription factors or as an enzyme cofactor). Of note, SF-1, a transcription
factor that regulates the INSL3 promoter, contains two zinc-finger motifs that are required for DNA
binding. However, it is unclear if depletion is a consequence or a cause of decreased fetal testosterone
synthesis and subsequent steps in the MOA shown in Figure 3-1.
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Cellular Responses
Cellular responses are more well understood. There is abundant evidence that DBP disrupts the
production of fetal testicular testosterone in rodents. Disruption of testicular testosterone production
during the masculinization programming window (i.e., GDs 15.5 to 18.5 for rats; GDs 14 to 16 for mice;
gestational weeks 8 to 14 for humans) can lead to antiandrogenic effects on the developing male
reproductive system (MacLeod et al.. 2010; Welsh et al.. 2008; Carruthers and Foster. 2005). Consistent
with the MOA outlined in Figure 3-1, many studies of DBP identified by EPA have demonstrated that
oral exposure to DBP during the masculinization programming window can reduce testosterone
synthesis in the fetal male Ley dig cell and/or reduce expression (mRNA and/or protein) of insulin-like
growth factor 3 (INSL3), as well as genes involved in steroidogenesis in the fetal testes of rats.
Testosterone production drives extratesticular male reproductive tract development and, together with
INSL3, drives adverse organism-level outcomes, such as testicular descent. The vast majority of studies
identified have found decreased fetal testicular testosterone (ranging from 34 to 85 percent) following
exposures of pregnant rats to 500 mg/kg-day or higher (Table 3-3; Table 3-4). However, reductions in
fetal testicular testosterone have also been observed at lower doses ranging from 50 to 112 mg/kg-day
(Gray et al.. 2021; Furr et al.. 2014; Struve et al.. 2009; Mahood et al.. 2007; Lehmann et al.. 2004). Furr
et al. (2014) carried out several experiments in "blocks" conducted over 2 to 3 years, and observed
decreased ex vivo fetal testicular testosterone production in male rats from multiple blocks at doses as
low as 100 mg/kg-day, reflecting a 35 percent (Block 18) or 36 percent (Block 22) decrease in
testosterone. The data set from Mahood et al. (2007). demonstrates that a 14 percent decrease in
testicular testosterone content coincides with other male reproductive effects including increased Ley dig
cell aggregation and increased incidence of MNGs, and are therefore biologically significant. In parallel
with their observations of decreased fetal testicular testosterone, Lehmann et al. (2004) reported
reductions testicular mRNA and protein expression of genes involved in steroidogenesis (e.g., StAR,
P450scc, CYP17) at doses of 50 mg/kg-day and up, and testis descent (Insl3) at 500 mg/kg-day and up.
Additionally, significant decreases in gene expression of SR-B1, 3/3-HSD, and c-Kit were observed at
lower doses (0.1 or 1.0 mg/kg-day). Other studies of rats have also reported decreased fetal testicular
testosterone production or content coinciding with decreased expression of genes involved in cholesterol
transport and steroidogenesis (e.g., see (Gray et al.. 2021; Struve et al.. 2009)).
Moreover, several studies in rats have demonstrated that even a single exposure on a single day during
the critical window (i.e., GD 14 to 18) could elicit decreases in testicular testosterone and steroidogenic
gene expression (Johnson et al.. 2012; Johnson et al.. 2011; Johnson et al.. 2007; Kuhl et al.. 2007).
Kuhl et al. (2007) reported that fetal testicular mRNA levels of St A R, SR-B1 Scarbl, P450scc Cypl lal,
and CYP17 were decreased in GDI9 fetuses of pregnant rats exposed to doses as low as 100 mg/kg-day
DBP on GD18. Fetal testicular testosterone concentration was decreased at 500 mg/kg-day. Another
single exposure study reported decreased intratesticular testosterone on GDI9 one hour after dosing with
500 mg/kg-day (Johnson et al.. 2007). In later publications by the same authors (Johnson et al.. 2012;
Johnson et al.. 2011). reductions in steroidogenic gene expression was observed in the fetal testes 3
hours (Cypl7al) to 6 hours (P450scc Cypl lal, StAR) post-exposure in pregnant SD rats gavaged with
a single dose of 500 mg/kg DBP on GD 19. Fetal testicular testosterone was reduced starting at 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 (P450scc Cypl lal,
Cypl7al, Scarbl) hours post-exposure, and protein levels of these genes were reduced 6 to 12 hours
post-exposure. Altogether, these data support a mode of action where key changes in genes involved in
steroidogenesis or testosterone transport precede cellular responses and subsequent organ-level
responses.
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Organ-level Responses
Organ-level responses in the reproductive system include Leydig cell aggregation or altered distribution
of Ley dig cells, reduced AGD, and increased nipple retention. Perturbations in Leydig cell morphology
are indicative of disrupted androgen action. Leydig cells of the testes produce testosterone, INSL3, and
dihydrotestosterone (DHT), which forms from its precursor, testosterone. Reduced AGD (which is an
externally visible marker) stems from reduced production of testosterone by the Leydig cell during the
masculinization programming window, as DHT functions to lengthen the perineum (i.e., skin between
the genitals and anus) of males. AGD is therefore a sensitive indicator of prenatal androgen exposure.
Increased nipple retention also stems from reduced testosterone production, as DHT in peripheral tissues
is necessary for apoptosis and regression of nipples in male rats. Each of these responses have been well
documented in rodents exposed to DBP following gestational exposure. Indeed, three studies have
reported increased incidences of Leydig cell aggregates at doses ranging from 100 mg/kg-day (Scarano
et al.. 2010; Struve et al.. 2009; Mahood et al.. 2007) to 300 mg/kg-day (Li et al.. 2015). Aside from
these studies, the majority of studies report histopathological alterations at doses above 100 mg/kg-day.
At higher levels of exposure (i.e., 250 to 750 mg/kg-day), Leydig cell hyperplasia has been observed
(Mylchreest et al.. 2002).
Many studies have demonstrated that oral exposure of rats to DBP during the masculinization
programming window can reduce male pup AGD measured earlier in the postnatal window (i.e., on
PND1 through PND4) (Li et al.. 2015; Jiang et al.. 2007; Barlow et al.. 2004; Lee et al.. 2004; Zhang et
al.. 2004; Mylchreest et al.. 2000; Mylchreest et al.. 1999; Mylchreest et al.. 1998). after lactation on
PND25 (MacLeod et al.. 2010). or during adulthood (Drake et al.. 2009; Barlow et al.. 2004). Similarly,
increased male pup nipple retention (NR) around PND13 or PND14 has been consistently reported (Kim
et al.. 2010; Martino-Andrade et al.. 2008; Barlow et al.. 2004; Lee et al.. 2004; Mylchreest et al.. 2000;
Mylchreest et al.. 1999). 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, as well
as result in permanent NR, and increase the frequency of reproductive tract malformations and testicular
pathology in adult rats that received two doses of DBP during the critical window (i.e., GD 14 to 18).
Reduced AGD has been reported in rats following exposures during the masculinization programming
window at doses as low as 100 mg/kg-day (Martino-Andrade et al.. 2008). but most commonly between
300 and 500 mg/kg-day (Table 3-3). Similarly, two studies have reported increased nipple retention at
doses as low as 100 mg/kg-day (Barlow et al.. 2004; Mylchreest et al.. 2000). but this effect is more
commonly observed at doses of 250 mg/kg-day and higher (Table 3-3). Consistent with the animal
literature, there is epidemiological evidence that supports an inverse association between in utero
exposure to DBP and anogenital distance (Radke et al.. 2018). which may reflect exposure or
responsiveness to testosterone during fetal development.
Phthalates can also affect Sertoli cell function and development. Formation of lesions such as
multinucleated gonocytes (MNGs) is one indication of perturbed Sertoli cell function and development.
Increases in MNGs (Spade et al.. 2018; Boekelheide et al.. 2009; Ferrara et al.. 2006) have been
observed at higher levels of exposure to DBP (i.e., 250 to 750 mg/kg-day). While MNGs are also
observed in mice exposed to DBP during the critical window, decreased expression of genes involved in
steroidogenesis and cholesterol homeostasis that are observed in the testicular tissues of rats are not also
found in mice, suggesting that altered formation of MNGs is not mechanistically related to decreased
testosterone in mice as it is in rats (Gaido et al.. 2007).
Additionally, 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
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Substances Control Act (U.S. EPA. 2023 a). several explant (Lambrot et al.. 2009; Hallmark et al.. 2007)
and xenograft studies (van Den Driesche et al.. 2015; Heger et al.. 2012; Spade et al.. 2014; Mitchell et
al.. 2012) using human donor fetal testis tissue have been conducted to investigate the antiandrogenicity
of DBP and its monoester metabolite, MBP, as well as mono-2-ethylhexyl phthalate (MEHP; a
monoester metabolite of DEHP) in a human model. Generally, results from human explant and
xenograft studies (i.e., host serum testosterone production, host serum testosterone concentration, and
MNG formation) suggest that human fetal testes are less sensitive to the antiandrogenic effects of
phthalates, however, increased incidence of MNGs have been observed in two human xenograft studies
of DBP (van Den Driesche et al.. 2015;; Heger 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.
Organism-level Responses
Adverse outcomes at the organism-level have been observed following exposure to DBP during the
masculinization programming window, including effects on androgen-dependent organ weights (e.g.,
testes weight), testicular histopathology, seminiferous tubule atrophy, malformations (e.g., hypospadias),
cryptorchidism, or impaired fertility (Table 3-3).
Androgen-dependent testicular histopathology has been reported across a number of studies including
degeneration of the seminiferous tissue (Mylchreest et al.. 1999) or of the testicular tissues more
generally (Howdeshell et al.. 2007). or other perturbations (van den Driesche et al.. 2012; Johnson et al..
2011; McKinnell et al.. 2009; Zhang et al.. 2004; Higuchi et al.. 2003). Hypospadias and/or
cryptorchidism following gestational exposure to DBP during the critical window has been reported in
several rodent studies, some of which demonstrate lasting effects in adults that had been exposed in
utero (LOAELs range 250 to 700 mg/kg-day), demonstrating the permanence of these effects (Li et al..
2015; Kim et al.. 2010; Jiang et al.. 2007; Mahood et al.. 2007; Mylchreest et al.. 2000; Mylchreest et
al.. 1998). Reproductive tract malformations (Mylchreest et al.. 1998) or delayed male puberty (i.e.,
preputial separation) (Kim et al.. 2010) have also been reported at doses of 250 mg/kg-day. Similarly,
seminiferous tubule atrophy has been observed in adult rats that had been exposed to doses of DBP
ranging from 250 to 500 mg/kg-day during the critical window of fetal development (e.g., (Barlow et al..
2004; Mylchreest et al.. 1999; Wine et al.. 1997; NTP. 1995). and others in Table 3-3). Epidemiological
evidence is consistent with the findings of rodent studies. Indeed, the Radke et al. (2018) study
determined that the level of evidence was slight for the association between in utero exposure to DBP
and hypospadias and/or cryptorchidism (Radke et al.. 2018).
Gestational exposure to DBP has also been associated with reductions in reproductive performance
measures. In a multigenerational study with a continuous breeding protocol, decreased indices of
mating, pregnancy, and fertility were observed in Fl, but not F0 (Wine et al.. 1997; NTP. 1995)
generation rats, indicating the heightened sensitivity of the Fl generation due to the gestational
exposure. Mahood et al. (2007) reported increased incidence of infertility (approximately 75 percent of
infertile/fertile animals per litter and overall) in adult rats exposed to 500 mg/kg-day DBP in utero (GD
13.5 to GD12.5). Although increased incidences were observed at the lower doses (i.e., 4, 20, or 100
mg/kg-day), changes were not statistically significant and there was no dose-response (i.e., the incidence
of infertility across the 0, 4, 20, and 100 mg/kg-day groups was 1, 22, 14, and 33 percent, respectively).
Increased incidence of cryptorchidism was observed in parallel with the increased incidence of infertility
at 500 mg/kg-day, although cryptorchidism was observed in 1 of 19 animals in the 100 mg/kg-day
group. Dose-responsive increases in the percent of seminiferous cords with MNGs (LOAEL =100
mg/kg-day) and decreases in testis testosterone (LOAEL =100 mg/kg-day) were also observed.
Impaired fertility, reflected by reduced sperm count, reduced sperm motility, or increased percentages of
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abnormal sperm have also been reported in two studies following gestational exposures during the
critical window (Giribabu et al.. 2014; Zhang et al.. 2004). One study in rabbits also observed changes in
post-puberty sperm parameters following gestational exposure to 400 mg/kg-day DBP, providing further
evidence that the effects of DBP on fertility extend across species (Higuchi et al.. 2003). This study also
included a postnatal exposure, where fewer effects on fertility were observed compared to the
gestational exposure. However, some studies that evaluated male fertility following DBP exposures
during the critical window of up to 500 mg/kg-day did not observe any changes (Scarano et al.. 2010;
Martino-Andrade et al.. 2008). Further details on these studies are provided in Table 3-3 and Table 3-4.
An important limitation of the majority of these studies is that histopathological evaluations were
qualitative, which impacts the ability to interpret the results. Nevertheless, the few studies that provide
quantitative histopathological data (e.g., Mahood et al. (2007) and Mylchreest et al. (2000)) report
similar findings to the qualitative findings (e.g., Mylchreest et al. (1999)). and when considered together
support that seminiferous tubule atrophy, MNG formation, and changes in Ley dig cell morphology
occur following exposure to DBP. In support of the animal data, there is epidemiological evidence that
supports the association between exposure to DBP and indicators of fertility including semen parameters
(e.g., semen concentration, motility, and/or morphology), and time to pregnancy measured in adults.
3.2 New Literature Considered for Non-Cancer Hazard Identification
EPA identified 63 new animal toxicology studies that provide data on PECO-relevant health effects
following exposure to DBP. Of these, 12 studies provided LOAELs for PECO-relevant outcomes within
an order of magnitude of the most sensitive PODs identified from prior assessments (i.e., 20 mg/kg-day
or lower). These studies evaluated reproductive and developmental outcomes (seven studies),
neurological outcomes (three studies), nutritional/metabolic outcomes (three studies), cardiovascular
outcomes (one study), and the immune adjuvant capacity of DBP (two studies). Limitations in most of
these 12 studies impacted the interpretation of the results, and there was substantial resulting uncertainty
in the new data. Therefore, EPA ultimately did not consider these new animal toxicology studies further
in Section 4. EPA did not conduct a full evidence integration for health outcomes other than those of the
male reproductive system following developmental exposure (Section 3.1.2.1). Details and summaries
of EPAs consideration of new literature for Non-Cancer Hazard Identification are provided in Appendix
B. Summarized study information on the remining 51 studies is available in a supplemental file is (U.S.
EPA. 2024n).
3.3 Summary
Collectively, reasonably available studies consistently demonstrate that oral exposure to DBP during the
masculinization programming window can disrupt androgen action, leading to a spectrum of effects on
the developing male reproductive system consistent with phthalate syndrome. Evidence from
epidemiological studies indicates a moderate level of confidence in the association between DBP and
health effects on the male reproductive system, such as AGD. Evidence from animal studies, including
the robust database of studies in rats, demonstrates adverse effects on the male reproductive system
following developmental exposure to DBP. EPA's MOA analysis concluded that available studies
consistently demonstrate that oral exposure to DBP during the masculinization programming window
can disrupt androgen action, leading to a spectrum of effects on the developing male reproductive
system consistent with 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 DraftProposed 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 DBP's effects on the developing male reproductive system and EPA's MOA analysis.
EPA is considering effects on the developing male reproductive system for dose-response analysis and
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1449 for use in estimating risk to human health. The observed developmental effects are assumed to be
1450 relevant for extrapolating human risk. EPA further considered effects on the developing male
1451 reproductive system in Section 4.
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4 DOSE-REPONSE ASSESSMENT
EPA is considering non-cancer hazard endpoints related to effects on the developing male reproductive
system for dose-response analysis as described in the following sections. These hazard endpoints were
selected for dose-response analysis because EPA has the highest confidence in these hazard endpoints
for estimating non-cancer risk to human health and effects on the developing male reproductive system
are the most sensitive based on available data. Other non-cancer hazard endpoints were therefore not
considered for dose-response analysis or for estimating risk to human health.
For most hazard endpoints, EPA used a NOAEL/LOAEL approach for the dose-response analysis based
on a subset of critical studies. EPA considered NOAEL and LOAEL values from oral toxicity studies in
experimental animal models. For one hazard endpoint (i.e., 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), Diisobutyl Phthalate (DIBP), Dicyclohexyl Phthalate (DCHP),
andDiisononylPhthalate (U.S. EPA. 2024g). Acute, intermediate, and chronic non-cancer NOAEL/
LOAEL values identified by EPA are discussed further below in Section 4.2. 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. 2011b). Differences in dermal and oral absorption are corrected for
as part of the dermal exposure assessment. In the absence of inhalation studies, EPA performed route-to-
route extrapolation to convert oral HEDs to inhalation human equivalent concentrations (HECs)
(Appendix D).
4.1 Selection of Studies and Endpoints for Non-cancer Health Effects
EPA considered a 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 identified 39 studies that evaluated effects on the developing male reproductive system DBP
exposure (Table 3-3;Table 3-4). In order to focus its dose-response assessment, EPA further considered
the most sensitive studies of DBP supporting a LOAEL of 100 mg/kg-day or less in Section 4.2. Studies
supporting a LOAEL of greater than 100 mg/kg-day are discussed in Section 3.1.2 as part of the non-
cancer hazard identification and characterization. EPA identified 11 studies investigating effects on the
developing male reproductive system consistent with phthalate syndrome that support a LOAEL of 100
mg/kg-day or less and these studies are discussed further in Section 4.2 (Furr et al.. 2014; Moody et al..
2013; Boekelheide et al.. 2009; Clewell et al.. 2009; Martino-Andrade et al.. 2008; Mahood et al.. 2007;
Barlow et al.. 2004; Lee et al.. 2004; Lehmann et al.. 2004; Mylchreest et al.. 2000; Wine et al.. 1997).
EPA considered the following factors during study and endpoint selection for POD determination from
11 studies with relevant non-cancer health effects based on the following considerations:
• Exposure duration;
• Dose range;
• Relevance (i.e., considerations of species, direct vs. indirect effects, suitability of the endpoint as
a biomarker or indicator of the toxicological outcome,);
• Uncertainties not captured by the overall quality determination;
• Endpoint/POD sensitivity; and
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• 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 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
EPA considered effects on the developing male reproductive system across 11 studies of rats with
endpoints considered relevant to acute exposure duration (U.S. EPA. 1996. 1991). in addition to being
relevant for intermediate and chronic durations (Furr et al.. 2014; Moody et al.. 2013; Boekelheide et al..
2009; Clewell et al.. 2009; Martino-Andrade et al.. 2008; Mahood et al.. 2007; Barlow et al.. 2004; Lee
et al.. 2004; Lehmann et al.. 2004; Mylchreest et al.. 2000; Wine et al.. 1997). There is evidence that
effects on the developing male reproductive system consistent with a disruption of androgen action can
result from a single exposure during the critical window of development (i.e., GD 14 to 18) (Appendix
C). 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. 2024a V
These studies were previously discussed in Section 3.1.2.1 and are summarized in Table 4-1. The
majority of studies in Table 4-1 entailed exposure durations that exceeded a single day but evaluated
endpoints consistent with disruption of androgen action and included at least one dose during the critical
window. Effects observed across these studies included testicular histopathology consistent with
decreased spermatocyte development, decreased fetal testicular testosterone, male mammary gland
histopathology, decreased steroidogenic gene expression in the fetal testes, decreased male pup body
weights, effects on fetal Ley dig cells, increased incidence of MNGs, decreased anogenital distance, and
increased nipple retention.
Studies in Table 4-1 were subjected to dose-response analysis to select the study and endpoint most
appropriate to derive the POD for acute, intermediate, and chronic hazards. Candidate PODs range from
1 to 100 mg/kg-day based on antiandrogenic effects. Eight of these studies provided more sensitive
candidate PODs of 50 mg/kg-day of less for effects on the developing male reproductive system,
including decreased fetal testicular testosterone. EPA considers decreased fetal testicular testosterone
(reported as ex vivo fetal testicular testosterone production or fetal testicular testosterone content) to be
adverse and relevant to human health (U.S. EPA. 2023a. b).
As part of the dose response analysis, EPA also reviewed a meta-regression analysis and benchmark
dose (BMD) modeling analysis of decreased fetal testicular testosterone data published by The National
Academies of Sciences, Engineering, and Medicine (NASEM) (2017). Based on results from 12 studies
of rats (Li et al.. 2015; Furr et al.. 2014; van den Driesche et al.. 2012; Johnson et al.. 2011; Clewell et
al.. 2009; Struve et al.. 2009; Howdeshell et al.. 2008; Martino-Andrade et al.. 2008; Johnson et al..
2007; Kuhl et al.. 2007; Mahood et al.. 2007; Lehmann et al.. 2004). NASEM found high confidence in
the body of evidence and a high level of evidence that fetal exposure to DBP is associated with a
reduction in fetal testosterone in rats. NASEM further conducted a meta-regression analysis and BMD
modeling analysis on decreased fetal testicular testosterone production data from 7 studies of rats (Furr
et al.. 2014; Johnson et al.. 2011; Struve et al.. 2009; Howdeshell et al.. 2008; Martino-Andrade et al..
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2008; Johnson et al.. 2007; Kuhl et al.. 2007). Five studies were excluded from this meta-analysis
analysis due to deficiencies in data reporting (i.e., sample sizes were not reported for each dose group)
(Li et al.. 2015; van den Driesche et al.. 2012; Clewell et al.. 2009; Mahood et al.. 2007; Lehmann et al..
2004). Some of these studies were also reviewed for dose-response assessment by EPA (Table 4-1).
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 DBP. The linear-
quadratic model provided the best fit (based on lowest AIC) (Table 4-4). BMD estimates from the
linear-quadratic model were 12 mg/kg-day [95% confidence interval: 8, 22] for a 5 percent change
(BMR = 5%) and 125 mg/kg-day [85, 205] for a 40 percent change (BMR = 40%) (Table 4-4).
Since EPA identified new fetal testicular testosterone data (Gray et al.. 2021) for DBP, 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 E provides justification for the evaluated BMRs of 5, 10, and 40 percent. F Fetal rat
testosterone data from eight studies was included in the updated analysis, including new data from Gray
et al. (2021) and data from the same 7 studies included in the 2017 NASEM analysis. 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 4-3). There was substantial, statistically
significant heterogeneity in all cases (I2>90%). 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 4-4). BMD estimates from the linear-quadratic model were 14 mg/kg-day [95% confidence
interval: 9, 27] for a 5 percent change (BMR = 5%), 29 mg/kg-day [20, 54] for a 10 percent change
(BMR = 10%), and 149 mg/kg-day [101, 247] for a 40 percent change (BMR = 40%) (Table 4-4).
Notably, BMDs and BMD40 estimates calculated by NASEM and as part of EPA's updated analysis are
nearly identical (i.e., BMDs values of 12 and 14 mg/kg-day; BMD40 values of 125 and 140 mg/kg-day).
Further 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), Diisobutyl Phthalate (DIBP),
DicyclohexylPhthalate (DCHP), andDiisononylPhthalate (U.S. EPA. 2024g). EPA considered the
BMDLs of 9 mg/kg-day further as a candidate POD.
Eleven studies were considered by EPA that provided relatively sensitive candidate PODs based on
antiandrogenic effects to the developing male reproductive system. Of these, four studies support
LOAELs of 50 to 100 mg/kg-day based on decreases in fetal testicular testosterone (Clewell et al..
2009). decreases in live pup weight and number of live pups per litter in second generation offspring
(Wine et al.. 1997). decreased male AGD (mm/cube root BW) on GD21 (Martino-Andrade et al.. 2008).
and increased nipple retention in males on PND13 (Barlow et al.. 2004). However, these studies are
limited by poor dose-selection and didn't test sufficiently low doses to establish a NOAEL. The Clewell
et al. (2009) study is further limited by the fact that it only one dose (i.e., the LOAEL of 50 mg/kg-day)
in addition to control and had a low sample size of 3 to 4 animals per group. Given the limitations, EPA
did not select these studies and endpoints for an acute/intermediate/chronic POD. Other studies tested
lower doses that allowed for the identification of a NOAEL.
Five studies identified NOAELs ranging from 20 to 50 based on increased nipple retention in males in
males on PND 14 (Mylchreest et al.. 2000). increased incidence in testicular pathology (Boekelheide et
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al.. 2009). or decreased fetal testicular testosterone (Furr et al.. 2014; Mahood et al.. 2007; Lehmann et
al.. 2004). The NOAEL of 20 mg/kg-day from Mahood et al, (2007) was based on increases in MNGs
and Ley dig cell aggregation at 100 mg/kg-day, in addition to decreased fetal testicular testosterone.
However, as described further below each study contained limitations or areas of uncertainty that
impacted the ability of EPA to interpret the results and ultimately EPA did not select any of these
candidate PODs.
Mylchreest et al. (2000) provided a POD of 11.8 mg/kg-day (HED) based on a NOAEL of 50 mg/kg-
day. The POD was based on increased nipple retention in 31 percent of the males from the 100 mg/kg-
day group on PND 14. Although the study used large sample sizes (i.e., 20 litters/dose, per OECD TG
414 guidelines), and the exposure period encompassed the critical window, there were considerable
factors that decreased confidence in the study, primarily the lack of other adverse reproductive effects at
the same dose. Indeed, decreases in AGD at birth (PND1), decreased reproductive organ weights, and
histopathological lesions (i.e., interstitial cell hyperplasia in the seminiferous epithelium) were only
observed in the high dose group (i.e., 500 mg/kg-day).
Although Mahood et al. (2007) and Lehman et al. (2004) identified sensitive candidate PODs of 4.7
mg/kg-day and 7.1 mg/kg-day (based on NOAELs of 20 and 30 mg/kg-day), respectively, neither was
considered further due to reporting deficiencies (e.g., sample sizes unclear or did not reflect the litter as
the statistical unit). For similar reasons, NASEM (2017) had not included these two studies in their
original meta-analysis and BMD modeling. Additionally, other studies provided more sensitive
candidate PODs.
Furr et al. (2014) and Boekelheide et al. (2009) support candidate PODs based on NOAELs of 10 or 50
mg/kg-day. Furr et al. (2014) identified two candidate PODs based on NOAELs of 10 and 50 mg/kg-day
(HED = 2.4 and 11.8 mg/kg-day, respectively) based on decreased fetal testicular testosterone at the next
highest dose group (LOAEL =100 mg/kg-day, for both studies). However, given the same LOAEL and
large dose-spacing of the Block 22 data, the NOAEL of 10 mg/kg-day from Block 22 is likely an artifact
of dose-selection, which decreases EPAs confidence in its utility as a POD. Additionally, there was no
clear dose-response in Block 18 (i.e., NOAEL = 50 mg/kg-day), and both blocks had low sample sizes.
Low sample size as also a limitation of Boekelheide et al, (2009). This study observed increased
incidences of testicular pathology (i.e., decreased testicular cell number and disorganized seminiferous
tubules) in rats exposed to 30 mg/kg-day during gestation (NOAEL = 10 mg/kg-day; HED = 2.4 mg/kg-
day). However, the adversity of this outcome is uncertain, which reduces EPAs confidence in this
outcome as a POD.
Two additional studies (Moody et al.. 2013; Lee et al.. 2004). identified the lowest candidate PODs
reviewed by EPA (Table 4-1). Of note, these were below the BMDLs of 9 mg/kg-day (HED= 2.1 mg/kg-
day) identified from the updated meta-analysis and BMD modeling conducted by EPA. Moody et al.
(2013) and Lee et al. (2004) provided candidate PODs of 0.13 mg/kg-day (Moody et al.. 2013) and
0.71 mg/kg-day (Lee et al.. 2004). based on effects on the developing male reproductive system.
Although the studies offer sensitive PODs and provide data from two different species (mice and rats)
consistent with decreased spermatocyte development following gestational and/or postnatal exposure to
DBP, limitations related to insufficient methodological detail, study design and exposure timing, and
evidence of maternal toxicity reduced EPAs confidence in the results.
Moody et al. (2013) offered a sensitive POD of 0.13 mg/kg-day (HED) based on delayed
spermatogenesis in mice (LOAEL = 1 mg/kg-day). These data are inconsistent with the abundant
literature indicating that the rat is more sensitive than the mouse to the antiandrogenic effects of
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phthalates. Nevertheless, the study represents an expansion of the data set to include additional sensitive
lifestages and examine prepubertal exposure at the beginning of the first wave of spermatogenesis, as
mice were exposed to DBP from PND4 to PND14. However, given this exposure window {i.e., PND4 to
PND14), which is outside the masculinization programming window in mice, the permanence of these
effects are therefore less well known than effects that stem from exposures during the critical window.
In adult animals exposed to 1 mg/kg-day, semi-quantitative histopathological observations suggest
defective spermatogenesis in addition to decreased AGD relative to body weight. However,
methodological limitations hinder the interpretation of these results, as data are presented as individual
values rather than litter means, so it is unclear of the quantitative data are statistically analyzed correctly.
In addition to this limitation, there is no clear dose-response in either outcome, which increases
uncertainty in the data set. Given the uncertainty regarding the permanence of effects and data
presentation, EPA did not consider this study further.
Lee et al. (2004) offered a POD of 0.71 mg/kg-day (HED; LOAEL = 3 mg/kg-day) based on increased
incidence of reduced spermatocyte development in PND21 rats exposed to DBP from GDI 5 to PND21.
However, several limitations increased the uncertainty in this endpoint, including lack of a linear dose-
response, and the fact that the severity score was minimal to mild for the lowest two doses and did not
linearly increase in severity with increasing dose. Additional sources of uncertainty for this study
include the age of outcome assessment being close to the beginning of spermatocyte development
(which begins around PND21), which impacts the interpretation of the severity scores. Interpreting the
histopathological data is further limited by insufficient methodological detail required to understand how
the outcomes were assessed. Additionally, maternal weight gain during pregnancy was significantly
decreased in the low dose group, which potentially confounds the observed effects on spermatocyte
development.
Both Moody et al.(2013) and Lee et al. (2004) point to sensitive effects following exposure to DBP
during a sensitive lifestage that is observed in both mice and rats. It is likely that species differences in
sensitivity of these pubertal effects across the two studies is a function of study design, as in both cases,
no NOAEL was identified (i.e., lowest dose tested has an effect). However, the aforementioned
limitations in each study impact the interpretation of the results and contribute uncertainty and as a result
EPA did not select either study for the POD for acute and/or intermediate exposures.
Data on chronic studies of DBP did not offer a more sensitive POD than the database of developmental
exposure studies (Table 4-1). Moreover, NTP (2021) identified a LOAEL of 510 mg/kg-day
(HED = 120.6 mg/kg-day) based on increased gross findings in male rats (cryptorchidism, agenesis,
small testis), increased microscopic findings in the testes (e.g., seminiferous tubule dysgenesis, Ley dig
cell hyperplasia) and hypospermia), increased incidence of hepatocyte alteration in the liver of males
and females, and increased incidence of hypertrophy in the pars distalis male rats. Because the scarce
data that exist on chronic exposure durations of DBP (NTP. 2021) do not offer more sensitive PODs
than those considered relevant for acute exposure durations (Table 4-1), EPA is considering acute
duration PODs for intermediate and chronic durations as well.
Ultimately, EPA is proposing the BMDLs of 9 mg/kg-day based on the decreased fetal testicular
testosterone as the POD for assessing risks from acute, intermediate, and chronic durations of exposure.
Numerous factors increase EPA's confidence in using the HED of 2.1 mg/kg-day based on the decreased
fetal testicular testosterone. Notably, the BMDLs of 9 mg/kg-day falls within the narrow range of the
NOAEL or LOAELs (i.e., 1 to 10 mg/kg-day) identified in additional studies that evaluated effects on
the developing male reproductive svstem(Moodv et al.. 2013; Boekelheide et al.. 2009; Mahood et al..
2007; Lee et al.. 2004; Lehmann et al.. 2004). which provides support and confidence in both the effect
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1691 and the dose at which it occurs. Additionally, the BMDLs is not constrained to one of the experimental
1692 doses within a given study, as a NOAEL or LOAEL would be, which may better define the POD (U.S.
1693 EPA. 2012). Using allometric body weight scaling to the three-quarters power, EPA extrapolated an
1694 HED of 2.1 mg/kg-day. A total uncertainty factor of 30 was selected for use as the benchmark margin of
1695 exposure (based on an interspecies uncertainty factor (UFa) of 3 (see Appendix D for further discussion)
1696 and an intraspecies uncertainty factor (UFh) of 10).
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1697 Table 4-1. Studies Being Considered for POD Selection
Study Details
(Species, duration, exposure route/
method, doses [mg/kg-day])
Study POD/ Type
(mg/kg-day)
Effect
HED
(mg/kg-day)
HEC
(mg/m3)
[ppm]
Uncertainty
Factors abc
Reference d
Male C57BL/6J mice (n = 5-10/group)
were fed corn oil with 0, 1, 10, 50, or
500 mg/kg-day DBP from PND4-PND14
LOAEL = 1
Delayed
spermatogenesis,
reduced abs. AGD
(rel. to BW at higher
dose) in mice
(PND 4-14)
0.13
0.7 [0.06]
UFa = 3
UFh=10
UFl = 10
Total UF =300
(Moodv et al..
2013)
Pregnant rats (6-8 dams/group) were
exposed to 0, 20, 200, 2000, or 10,000
ppm DBP via diet from GDI5 - PND21
(equivalent to 0, 1.5-3, 14-29, 148-291,
712 - 1372 mg/kg-day)e
LOAEL = 3
i spermatocyte
development
(PND 21), t vacuolar
degeneration of
alveolar cells, alveolar
atrophy of mammary
gland (PNW 11 males)
0.71
3.9 [0.34]
UFa = 3
UFh=10
UFl = 10
Total UF =300
(Lee et al..
2004)
Meta-regression and BMD modeling of
fetal testicular testosterone in rats across
seven studies of rats exposed to 1-600
mg/kg-day DBP at various times during
gestation
BMDL5 — 8
i Fetal testicular
testosterone
1.89
10.3 [0.90]
UFa = 3
UFH=10
Total UF =30
(NASEM.
2017)h
Pregnant SD rats (4-10 litters/group)
gavagedwithO 0.1, 1, 10, 30, 50, 100,
500 mg/kg-day DBP on GD 12-21
NOAEL = 10
t testicular pathology
(J, testicular cell
number; disorganized
seminiferous tubules)
2.36
12.9 (1.13)
UFa = 3
UFH=10
Total UF =30
(Boekelheide
et al.. 2009)
Pregnant Harlan SD rats (3-4/dose)
gavagedwithO, 1, 10, 100 mg/kg-day
DBP on GDs 14-18 (Block 22)'
NOAEL = 10
I ex vivo fetal
testicular testosterone
production
2.36
12.9 (1.13)
UFa = 3
UFH=10
Total UF =30
(Furr et al..
2014)8
Pregnant Harlan SD rats (2-3/dose)
gavaged with 0, 33, 50, 100, 300 mg/kg-
day DBP on GDs 14-18 (Block 18) '
NOAEL = 50
I ex vivo fetal
testicular testosterone
production
11.82
64.3 (5.65)
UFa = 3
UFH=10
Total UF =30
Pregnant Wistar rats gavaged with 0, 4,
20, 100, 500 mg/kg-day DBP on GD
13.5-20.5
NOAEL = 20
i fetal testicular
testosterone content, t
MNGs, t Leydig cell
aggregation
4.73
25.7 (2.26)
UFa = 3
UFH=10
Total UF =30
(Maliood et
al.. 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-day)
HEC
(mg/m3)
[ppm]
Uncertainty
Factors abc
Reference d
Pregnant SD rats (3-4 separate rat fetuses
from 1-4 dams/group) gavaged with 0,
0.1. 1, 10, 30, 50, 100, 500 mg/kg-day
DBP on GD 12-19 '
NOAEL = 30
i fetal testis
testosterone on GD 19
7.09
38.6 (3.39)
UFa = 3
UFh=10
Total UF =30
(Lehmann et
al.. 2004)
Pregnant SD rats (19-20 or 11 (high-
dose) per dose) gavaged with 0, 0.5, 5,
50, 100, 500 mg/kg-day DBP on GDs
12-21
NOAEL = 50
t males with nipples
and/or areolae on
PND 14
11.82
64.3 (5.65)
UFa = 3
UFh=10
Total UF =30
(Mylchreest
et al.. 2000)
Pregnant SD rats (4 litters/dose) exposed
to 0 or 50 mg/kg-day DBP from GD 12-
19 via gavage, 12 hours after final dose
LOAEL = 50
mg/kg-day
i fetal testicular
testosterone
concentration
11.82
64.3 (5.65)
UFa = 3
UFh=10
UFl = 10
Total UF =300
(Clewell et
al.. 2009)
Continuous breeding protocol. Pregnant
VAF Crl:CD BR outbred Sprague-
Dawley albino rats (20/sex/group; 40/sex
for controls) exposed to 0, 0.1, 0.5, or 1%
DBP via diet starting 10 weeks prior to
mating and throughout gestation and
lactation periods continuously for 2
generations (equivalent to 52, 256, 509
mg/kg-day [males]; 80, 385, or 794
mg/kg-day [females])
LOAEL = 80
F2: I live pup weight;
F1: j live pups per
litter
18.91
102.9 (9.04)
UFa = 3
UFh=10
UFl = 10
Total UF =300
(Wine et al..
1997: NTP.
1995)
Pregnant Wistar rats (7-8/dose) gavaged
with 0, 100, 500 mg/kg-day DBP on GDs
13-21 (fetal study)
LOAEL = 100
| male AGD (GD21)
23.64
128.7 (11.30)
UFa = 3
UFH=10
UFl = 10
Total UF =300
(Martino-
Andrade et
al.. 2008F
Pregnant SD rats (10/11/dose) gavaged
with 0, 100, 500 mg/kg-day DBP on GDs
12-21
LOAEL = 100
t F1 males withNR
(PND 13)
23.64
128.7 (11.30)
UFa = 3
UFH=10
UFl = 10
Total UF =300
(Barlow et al..
2004)
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Study Details
(Species, duration, exposure route/
method, doses [mg/kg-day])
Study POD/ Type
(mg/kg-day)
Effect
HED
(mg/kg-day)
Reference d
EPA identified the above listed studies supporting derivation of candidate acute, intermediate, and chronic PODs; the selected POD from NASEM (2017) is in
bold.
PND = postnatal day; GD = gestation day ; LOAEL = lowest observed adverse effect level; NOAEL = No-observed-adverse-effect level; POD = point of
departure; HED = human equivalent dose; HEC = human equivalent concentration; MNG= Multinucleated gonocytes; UF = uncertainty factor; UFA= interspecies
uncertainty factor; UFH = intraspecies uncertainty factor; UFL = LOAEL-to-NOAEL uncertainty factor.
" EPA used allometric body weight scaling to the three-quarters power to derive the HED. Consistent with EPA Guidance (U.S. EPA. 2011b). the interspecies
uncertainty factor (UFA), was reduced from 10 to 3 to account remaining uncertainty associated with interspecies differences in toxicodynamics.
h 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 Overall data quality determinations were not made for these studies because the acute POD was more sensitive than the acute/intermediate, or chronic candidate
PODs, and these studies are not used quantitatively in the draft DBP risk evaluation.
e Equivalent doses provided by (NICNAS. 2008).
^Multiple time blocks in this experiment, which was carried out over 2-3 years, with each block consisting of 15 pregnant dams divided into 4-5 exposure groups.
g Considered in the metanalysis of the effect of DBP on fetal testosterone by NASEM (2017).
h R code supporting NASEM's meta-regression and BMD analysis of DBP is publicly available through GitHub (https://github.com/wachiuphd/NASEM-2017-
Endocrine-Low-Dose)."
' Authors state that the study was repeated, and a 30-mg/kg/day dose group was included for the testosterone radioimmunoassay (RIA). All other endpoints in this
study do not have a 30 mg/kg-day group.
' Data from block 70 and 71 rats in Gray et al. (2021).
1698
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1699 Table 4-2. Summary of Effects of Gestational Exposure to DBP on Testicular Testosterone Across Select Studies
Study Details
(Species, Duration, Exposure Route/
Method, Endpoint, Measurement timing.
Reference; TSCA Study Quality Rating)
% of Control Testosterone Response by Dose (mg/kg-day)"
0
1
10
33
50
100
112
300
500
581
600
900
SD Rats (Block 18); GD 14-18;
Oral/gavage; ex vivo fetal testicular
testosterone Droduction: GD 18 (Furr et al..
2014)*
High confidence
100%
(ii = 3)
-
32%
(ii=3)
OO
£
65%*
(n=3)
-
23%*
(n=3)
-
-
-
-
SD Rats (Block 22); GD 14-18;
Oral/gavage; ex vivo fetal testicular
testosterone Droduction: GD 18 (Furr et al..
2014)*
High confidence
100%
(11 = 3)
88%
(ii=3)
80%
(ii=4)
-
-
64%*
(n=4)
-
-
-
-
-
-
SD Rats; GD 19; Oral/gavage; testicular
testosterone; GDI9 (1 lirpost exposure)
(Johnson et al.. 2007)*
Medium confidence
100%
(11 = 5)
-
109%
(ii=5)
67%
(ii=5)
-
84%*
(ii=5)
-
-
-
-
-
-
SD Rat; GD 8-18; Oral/gavage; ex vivo
testicular testosterone production; GD 18 (2
lir incubation) (Howdeshell et al.. 2008)*
High confidence
100%
(11 = 3)
-
-
94%
(ii=4)
78%
(ii=4)
84%
(ii=4)
-
66%*
(n=4)
-
-
33%*
(n=4)
-
WistarRat; GD 13-21; Oral/gavage;
testicular testosterone: GD21 (Martino-
Andrade et al.. 2008)*
Medium confidence
100%
(11 = 7)
-
-
-
-
71%
(11=8)
-
-
37%*
(n=7)
-
-
-
SD Rat; GD18; Oral/gavage; testicular
testosterone: GDI9 (Kuhl et al.. 2007)b
Low confidence
100%
(11=10)
-
-
-
-
71%
(n=10)
-
-
33%*
(n=10)
-
-
-
SD Rat; GD12-19; Oral/diet; testicular
testosterone; GDI9 (4 hr post exposure)
(Strove et al.. 2009)*
Medium confidence
100%
(11 = 9)
-
-
-
-
-
56%
(ii=7)
-
-
3.7%*
(n=7)
-
-
SD Rat; GD12-19; Oral/diet; testicular
testosterone; GD20 (24 hr post exposure)
(Strove et al.. 2009)*
Medium confidence
100%
(11 = 9)
-
-
-
-
-
29%*
(n=7)
-
-
7.1%*
(n=7)
-
-
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Study Details
(Species, Duration, Exposure Route/
Method, Endpoint, Measurement timing.
Reference; TSCA Study Quality Rating)
% of Control Testosterone Response by Dose (mg/kg-day)"
0
1
10
33
50
100
112
300
500
581
600
900
SD Rat; GD12-20; Oral/gavage; testicular
testosterone: GD20 (Johnson et al.. 2011)b
Medium confidence
100%
n = 6)
-
-
-
-
-
-
-
15%*
(n=5)
-
-
-
SD Rats (Block 70); GD 14-18;
Oral/gavage; ex vivo fetal testicular
testosterone Droduction: GD 18 (Grav et al..
2021)c
High confidence
100%
(ii = 3)
-
-
-
-
-
-
62%
(n=4)
-
-
25%
(n=4)
16%
(n=4)
SD Rats (Block 71); GD 14-18;
Oral/gavage; ex vivo fetal testicular
testosterone Droduction: GD 18 (Grav et al..
2021)c
High confidence
100%
(11 = 4)
-
-
-
-
-
-
47%
(n=3)
-
-
22%
(n=4)
13%
(n=4)
SD = Sprague-Dawley; GD = Gestation Day; lir = hour
The following studies reported fetal testicular testosterone data but are not represented in this table because the sample sizes were not reported for each dose
aroiiD: (Mahood et al.. 2007); (Lehmann et al.. 2004); (Clewell et al.. 2009); (Li et al.. 2015); (van den Driesche et al.. 2012).
"Effect on fetal testicular testosterone production reported as percent of control. Asterisks indicate statistically significant pairwise comparison to control, as
reported by study authors.
h Data used in meta-analysis and BMD modeling analysis of fetal testosterone.
"Data from Block 70 and 71 rats reported in supplemental information file associated with Gray et al. (2021). Ex vivo testosterone production data from Block 70
and 71 rats was not subjected to statistical analysis.
d No data; dose not evaluated in this study.
1700
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1701 Table 4-3. Overall Analyses and Sensitivity Analyses of Rat Studies of DBP and Fetal Testosterone (Updated Analysis Conducted by
1702 EPA) ^
Analysis
Estimate
Beta
CI, Lower
Bound
CI, Upper
Bound
P value
Tau
I2
P value for
Heterogeneity
AICs
Primary Analysis
Overall
intrcpt
-71.85
-95.76
-47.95
3.82E-09
67.01
95.60
2.74E-152
383.39
Trend in loglO(dose)
loglO(dose)
-62.44
-81.70
-43.19
2.08E-10
41.61
88.70
4.43E-50
349.26
Linear in dose 100
doselOO
-25.69
-31.55
-19.83
8.64E-18
57.78
94.26
3.38E-119
354.71
LinearQuadratic in doselOO
doselOO
-36.78
-54.53
-19.03
4.89E-05
54.79
93.26
1.72E-117
343.82*
LinearQuadratic in doselOO
I(dosel00A2)
1.70
-0.86
4.26
1.94E-01
54.79
93.26
1.72E-117
343.82
Sensitivity Analysis
Overall minus Furr et al. 2014
intrcpt
-88.38
-117.31
-59.45
2.14E-09
67.21
93.19
2.16E-55
270.22
Overall minus Johnson et al. 2007
intrcpt
-76.78
-102.25
-51.31
3.47E-09
68.66
96.10
3.84E-153
350.04
Overall minus Howdeshell et al. 2008
intrcpt
-78.30
-105.70
-50.91
2.11E-08
70.83
95.72
3.63E-139
329.10
Overall minus Johnson et al. 2011
intrcpt
-69.59
-93.70
-45.48
1.53E-08
65.39
95.51
3.39E-148
359.45
Overall minus Kuhl et al. 2007
intrcpt
-72.06
-97.37
-46.75
2.39E-08
68.92
95.94
3.87E-152
362.13
Overall minus Martino-Andrade et al. 2009
intrcpt
-72.43
-97.80
-47.06
2.19E-08
69.11
95.94
1.74E-152
362.26
Overall minus Struve et al. 2009
intrcpt
-63.19
-86.77
-39.61
1.50E-07
62.87
95.50
2.53E-148
329.62
Overall minus Gray et al. 2021
intrcpt
-56.97
-80.64
-33.31
2.37E-06
59.25
94.78
3.05E-115
311.44
* Indicates lowest AIC.
1703
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Table 4-4. Benchmark I
)ose Estimates for DBP and Fetal Testosterone in Rats
Analysis
BMR
BMD
CI, Lower
Bound
CI, Upper Bound
2017 NASEM Analysis for all strains of rats using Metafor Version 2.0.0
(as reported in Table C6-8 of NASEM, 2017)
Linear in doselOO
5%
17
14
22
Linear in doselOO
40%
174
143
222
LinearQuadratic in
doselOO*
5%
12
8
22
LinearQuadratic in
doselOO*
40%
125
85
205
Updated Analysis using Metafor Version 4.6.0
Linear in doselOO
5%
20
16
26
Linear in doselOO
10%
41
33
53
Linear in doselOO
40%
199
162
258
LinearQuadratic in
doselOO*
5%
14
9
27
LinearQuadratic in
doselOO*
10%
29
20
54
LinearQuadratic in
doselOO*
40%
149
101
247
* Indicates model with lowest AIC.
1705
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4.3 Weight of Scientific Evidence: POD for Acute, Intermediate, and
Chronic Durations
EPA has reached the preliminary conclusion that the HED of 2.1 mg/kg-day (BMDLs of 9 mg/kg-day) is
appropriate for calculation of risk from acute, intermediate, and chronic exposures to DBP. This POD is
based on a meta-analysis and BMD modeling of decreased fetal testicular testosterone in eight studies of
rats exposed to DBP during gestation. 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. 2002. 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 D). EPA has robust overall confidence
in the proposed POD for acute, intermediate, and chronic durations based on the following weight
of the scientific evidence:
• EPA has previously considered the weight of evidence and updated here and concluded that oral
exposure to DBP 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).
• DBP exposure resulted in effects on the developing male reproductive system consistent with a
disruption of androgen action during the critical window of development in over 20 studies of
rats (Section 3.1.2.1), 11 of which reported LOAELs at or below 100 mg/kg-day (Table 3-3).
Observed effects in rats perinatally exposed to DBP included: disruption of testicular
testosterone production; reductions testicular mRNA and protein expression of genes involved in
steroidogenesis (e.g., StAR, P450scc, CYP17) and testis descent (Insl3); decreased AGD;
increased NR; disrupted testis tubules; Leydig cell clusters; increased incidence of MNGs;
changes in androgen-dependent organ weights (e.g., testes weight); testicular histopathology;
and/or malformations (e.g., hypospadias).
• Alignment across epidemiological, animal toxicology, and mechanistic streams of evidence
(Section 3.3).
• The proposed POD is based on meta-regression analysis of fetal testicular testosterone data from
eight studies of rats (Gray et al.. 2021; Furr et al.. 2014; Johnson et al.. 2011; Struve et al.. 2009;
Howdeshell et al.. 2008; Martino-Andrade et al.. 2008; Johnson et al.. 2007; Kuhl et al.. 2007).
• Chronic studies do not offer a more sensitive chronic POD. The NTP (2021) identified a POD of
510 mg/kg-day (based on LOAEL in rats; HED = 130 mg/kg-day).
• The BMDLs of 9 mg/kg-day (HED 2.1 mg/kg-day) is within the range of PODs (i.e., 1 to 10
mg/kg-day) identified from other studies based on antiandrogenic effects on the developing male
reproductive system (Furr et al.. 2014; Boekelheide et al.. 2009). These studies support the
selection of the BMDLs of 9 mg/kg-day for the acute, intermediate, and chronic duration PODs.
• Three developmental toxicity studies (Furr et al.. 2014; Mahood et al.. 2007; Lehmann et al..
2004) provide NOAEL values ranging from 10 to 30 mg/kg-day based on decreased fetal
testicular testosterone.
• 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 duration exposures, based on
studies of DBP which have demonstrated that a single exposure during the critical window of
development can disrupt expression of steroidogenic genes and decrease fetal testes testosterone.
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There are no studies conducted via the dermal and inhalation route relevant for extrapolating human
health risk, which remains a limitation. DBP undergoes hydrolysis by skin esterases to the bioactive
metabolite, MBP, which permeates the skinCSugino et al.. 2017). However, beyond this study, there are
insufficient data to support quantitative adjustment that accounts for this. Therefore, EPA is using the
oral HED of 2.1 mg/kg-day 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 Dibutyl phthalate (U.S. EPA. 2024f).
EPA is also using the oral HED of 2.1 mg/kg-day to extrapolate to the inhalation route. EPA assumes
similar absorption for the oral and inhalation routes, and no adjustment was made when extrapolating to
the inhalation route. For the inhalation route, EPA extrapolated the daily oral HEDs to inhalation HECs
using a human body weight and breathing rate relevant to a continuous exposure of an individual at rest.
Appendix D 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 proposed PODs in
Section 6.
5.2 PESS Based on Greater Susceptibility
EPA addressed subpopulations expected to be more susceptible to DBP 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 DBP.
EPA did not identify direct evidence of differences in susceptibility among human populations. EPA
identified indirect evidence for differences among human populations in ADME properties that may
impact lifestage susceptibility to DBP. For instance, the activity of glucuronosyltransferase differs
between adults and infants; adult activity is achieved at 6 to 18 months of age (Leeder and Kearns.
1997). Also, preexisting chronic liver or kidney disease may enhance susceptibility to DBP as a
consequence of impaired metabolism and clearance (i.e., altered functionality of phase I and phase II
metabolic enzymes); impaired activity of UGTs can reduce metabolism of chemicals that rely on UGT
conjugation to be excreted (Sugatani. 2013). including DBP (Section 2.1). Additional indirect evidence
of differences among human populations that confer enhanced susceptibility to DBP, including other
preexisting diseases, lifestyle factors, sociodemographic factors, genetic factors, and chemical co-
exposures are presented in Table 5-1. Animal studies provide direct evidence of several factors that
enhance susceptibility to DBP, including that gestation is a particularly sensitive lifestage for effects on
male reproductive development to manifest. These, and other lines of evidence are summarized in Table
5-1. EPA is quantifying risks based on developmental toxicity in the draft DBP risk evaluation.
As summarized in Table 5-1, EPA identified a range of factors that may have the potential to increase
biological susceptibility to DBP, including lifestage, chronic liver or kidney disease, pre-existing
diseases, physical activity, diet, stress, and co-exposures to other environmental stressors that contribute
to related health outcomes. The effect of these factors on susceptibility to health effects of DBP is not
known. Therefore, EPA is uncertain about the magnitude of any possible increased risk from effects
associated with DBP 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 DBP. The Risk Assessment Forum, in
A Review of the Reference Dose and Reference Concentration Processes (U.S. EPA. 2002). 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).
However, U.S. EPA (2002) did not discuss all the factors presented in Table 5-1. Although U.S. EPA
(2002) did not discuss all the factors presented in Table 5-1, EPA considers the POD proposed for use in
characterizing risk from exposure to DBP to be protective of effects on the developing male
reproductive system consistent with phthalate syndrome in humans. Thus, uncertainty remains whether
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1812 additional susceptibility factors would be covered by the default UFh value of 10 chosen for use in the
1813 draft DBP risk evaluation.
1814
1815 As discussed in U.S. EPA (2023a). exposure to DBP and other toxicologically similar phthalates {i.e.,
1816 DEHP, DIBP, BBP, DCHP, DINP) that disrupt androgen action during the development of the male
1817 reproductive system cause dose additive effects. Cumulative effects from exposure to DBP and other
1818 toxicologically similar phthalates will be evaluated as part of U.S. EPA's cumulative risk assessment of
1819 phthalates.
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1820 Table 5-1. PESS Evidence Crosswalk for Biological Susceptibility Considerations
Susceptibility
Category
Examples of
Specific Factors
Direct Evidence this Factor
Modifies Susceptibility to DBP
Indirect Evidence of Interaction with Target Organs
or Biological Pathways Relevant to DBP
Susceptibility Addressed in Risk
Evaluation?
Description of Interaction
Key Citations
Description of Interaction
Key Citation(s)
Lifestage
Embryos/
fetuses/infants
Direct quantitative animal evidence
for developmental toxicity including
multigenerational effects (e.g.,
increased skeletal and visceral
variations, decreased live births,
decreased offspring body weight gain,
and decreased offspring survival with
increased severity in the second
generation).
There is direct quantitative animal
evidence for effects on the developing
male reproductive system consistent
with a disruption of androgen action.
(Wine et al..
1997)
(U.S. EPA.
2023a)
(U.S. EPA.
2023b)
(Lee et al.. 2004)
(Boekelheide et
al.. 2009)
(Furretal.,2014)
(Mvlchreest et
al.. 2000)
POD proposed for assessing risks
from acute, intermediate, and
chronic exposures to DBP is based
on developmental toxicity (i.e.,
reduced fetal testicular
testosterone production) and is
protective of effects on the fetus
and offspring.
Pregnancy/
lactating status
Rodent dams less susceptible that
developing fetus during pregnancy
and lactation during a continuous
breeding multigenerational
experiment. Dams reduction in body
weight (14%) occurred at doses
higher than those that caused
developmental toxicity and pup
weight changes observed in the
absence of changes in maternal
weight for other doses.
(Wine et al..
1997)
POD proposed for assessing risks
from acute, intermediate, and
chronic exposures to DBP is based
on developmental toxicity (i.e.,
reduced fetal testicular
testosterone production) and is
protective of effects in dams.
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Susceptibility
Category
Examples of
Specific Factors
Direct Evidence this Factor
Modifies Susceptibility to DBP
Indirect Evidence of Interaction with Target Organs
or Biological Pathways Relevant to DBP
Susceptibility Addressed in Risk
Evaluation?
Description of Interaction
Key Citations
Description of Interaction
Key Citation(s)
Lifestage
Males of
reproductive age
and adolescence
Crossover mating trial study by Wine
et al. (1997) demonstrates that effects
on F1 offspring are attributable to
female (dam) exposure rather than
male (sire) exposure. Nevertheless,
three other studies of DBP exposure
to males of reproductive age or
adolescence suggest adverse
metabolic effects (e.g., increased BW
gain, BMI, serum glucose, and serum
total cholesterol) (Maieed et al..
2017). adverse outcomes of the
cardiovascular svstem (Xie et al..
2019). and neurobehavioral effects
(Farzanehfar et al.. 2016) in male rats
and mice. Effects observed at doses
ranging from 1 to 12.5 mg/kg-day
DBP. See Section 3.1.3 for individual
study details and Radke et al.
summary of human evidence on adult
male semen parameters.
(Wine et al..
1997)
(Xie et al.. 2019)
(Maieed et al..
2017)
(Farzanehfar et
al.. 2016)
(Moodv et al..
2013)
(Lee et al.. 2004)
POD proposed for assessing risks
from acute, intermediate, and
chronic exposures to DBP based
on developmental toxicity (i.e.,
reduced fetal testicular
testosterone production) is
protective of adult male
reproductive effects.
Use of default lOx UFH
Children
Reduced F1 and F2 rodent offspring
bodyweight (live pup weight) was
observed in a continuous breeding
experiment. Decreased F2 live pup
weight observed at lower dose.
(Wine et al..
1997)
POD proposed for assessing risks
from acute, intermediate, and
chronic exposures to DBP is based
on developmental toxicity (i.e.,
reduced fetal testicular
testosterone production) and is
protective of effects of offspring
bodyweight gain.
Use of default lOx UFH
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Susceptibility
Category
Examples of
Specific Factors
Direct Evidence this Factor
Modifies Susceptibility to DBP
Indirect Evidence of Interaction with Target Organs
or Biological Pathways Relevant to DBP
Susceptibility Addressed in Risk
Evaluation?
Description of Interaction
Key Citations
Description of Interaction
Key Citation(s)
Lifestage
Elderly
Two cross sectional studies suggest
associations with obesity in elderly
populations and combined MBP and
MiBP (80% is MBP) metabolites in
serum and associations of DBP
metabolites with adverse cognitive
functioning in the elderly.
(Wens et al..
2022)
(Li et al.. 2020)
Use of default lOx UFH
Toxicokinetics
The activity of enzymes involved in
metabolism of DBP differ between
adults and infants (e.g.,
glucuronosyltransferases, lipases,
CYPs) and may result in abnormal
toxicity.
(Leeder and
Kearns. 1997)
Use of default lOx UFH
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 (2023d)
Use of default lOx UFH
Pre-existing
disease or
disorder
Toxicokinetics
No direct evidence identified
Chronic liver and kidney disease are
associated with impaired metabolism
and clearance (altered expression of
phase 1 and phase 2 enzymes,
impaired clearance), which may
enhance exposure duration and
concentration of DBP.
(Sueatani. 2013)
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 (2023e)
Qualitative discussion in Section
5.2 and this table
Alcohol
consumption
No direct evidence identified
Alcohol use during pregnancy can
cause developmental outcomes (e.g.,
fetal alcohol spectrum disorders).
CDC (2023c)
Qualitative discussion in Section
5.2 and this table
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Susceptibility
Category
Examples of
Specific Factors
Direct Evidence this Factor
Modifies Susceptibility to DBP
Indirect Evidence of Interaction with Target Organs
or Biological Pathways Relevant to DBP
Susceptibility Addressed in Risk
Evaluation?
Description of Interaction
Key Citations
Description of Interaction
Key Citation(s)
Lifestyle
activities
Physical activity
No direct evidence identified
Insufficient activity may increase
susceptibility to multiple health
outcomes.
Overly strenuous activity may also
increase susceptibility.
CDC (2022)
Qualitative discussion in Section
5.2 and this table
Sociodemo-
Race/ethnicity
No direct evidence identified (e.g., no
information on polymorphisms in
DBP metabolic pathways or diseases
associated race/ethnicity that would
lead to increased susceptibility to
effects of DBP by any individual
group).
Qualitative discussion in Section
5.2 and this table
graphic status
Socioeconomic
status
No direct evidence identified
Individuals with lower incomes may
have worse health outcomes due to
social needs that are not met,
environmental concerns, and barriers
to health care access.
ODPHP (2023b)
Sex/gender
The key effect is male reproductive
development.
(U.S. EPA.
2023a)
Use of default lOx UFH
Nutrition
Diet
No direct evidence identified
Poor diets can lead to chronic
illnesses such as heart disease, type 2
diabetes, and obesity, which may
contribute to adverse developmental
outcomes. Additionally, diet can be a
risk factor for fatty liver, which could
be a pre-existing condition that
impairs liver enzyme metabolism of
DBP, thereby enhancing
susceptibility to DBP toxicity.
CDC (2023d)
CDC (2023a)
Qualitative discussion in Section
5.2 and this table
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Susceptibility
Category
Examples of
Specific Factors
Direct Evidence this Factor
Modifies Susceptibility to DBP
Indirect Evidence of Interaction with Target Organs
or Biological Pathways Relevant to DBP
Susceptibility Addressed in Risk
Evaluation?
Description of Interaction
Key Citations
Description of Interaction
Key Citation(s)
Nutrition
Malnutrition
No direct evidence identified
Micronutrient malnutrition can lead
to multiple conditions that include
birth defects, maternal and infant
deaths, preterm birth, low birth
weight, poor fetal growth, childhood
blindness, undeveloped cognitive
ability.
Thus, malnutrition may increase
susceptibility to some developmental
outcomes associated with DBP.
CDC (2021)
CDC (2023a)
Qualitative discussion in Section
5.2 and this table
Genetics/
epigenetics
Target organs
No direct evidence identified
Polymorphisms in genes may
increase susceptibility to
developmental toxicity, metabolic
outcomes, or neurological effects.
(Cassina et al..
2012)
(Inselman-
Sundbers. 2004)
Use of default lOx UFH
Toxicokinetics
No direct evidence identified
Polymorphisms in genes encoding
phase 1 or phase 2 metabolic
enzymes (e.g., UGTs, CYPs) or other
enzymes (e.g., lipases, esterases)
involved in metabolism of DBP may
influence metabolism and excretion
of DBP
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)
Qualitative discussion in Section
5.2 and this table
Social
enviromnent
No direct evidence identified
Social isolation and other social
determinants (e.g., decreased social
capital, stress) can lead to negative
health outcomes.
CDC (2023b)
ODPHP (2023c)
Qualitative discussion in Section
5.2 and this table
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Susceptibility
Category
Examples of
Specific Factors
Direct Evidence this Factor
Modifies Susceptibility to DBP
Indirect Evidence of Interaction with Target Organs
or Biological Pathways Relevant to DBP
Susceptibility Addressed in Risk
Evaluation?
Description of Interaction
Key Citations
Description of Interaction
Key Citation(s)
Other
chemical and
nonchemical
stressors
Chemical co-
exposures
Studies have demonstrated that co-
exposure to DBP and other
toxicologically similar phthalates
(e.g., DIBP, DEHP, DINP, BBP) and
other classes of antiandrogenic
chemicals (e.g., certain pesticides and
pharmaceuticals - discussed more in
(U.S. EPA. 2023a)) can induce
effects on the developing male
reproductive system in a dose-
additive manner.
See (U.S. EPA.
2023a) and (U.S.
EPA 2023b)
Qualitative discussion in Section
5.2 and this table and will be
quantitatively addressed as part of
the phthalate cumulative risk
assessment.
1821
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6 POINTS OF DEPARTURE USED TO ESTIMATE RISKS FROM
DBP EXPOSURE, CONCLUSIONS, AND NEXT STEPS
EPA considered the identified hazards, dose-response evaluation, and weight of the scientific evidence
of POD candidates, and ultimately chose one non-cancer endpoint for use in determining the risk from
acute, intermediate, and chronic exposure scenarios (Table 6-1). 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 9 mg/kg-day
(HED = 2.1 mg/kg-day) for acute, intermediate, and chronic durations. EPA has robust overall
confidence in the proposed 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 12 mg/m3 (1.0 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.
Table 6-1. Non-cancer HECs and HEDs Used to Estimate Risks for Acute, Intermediate, and
Chronic Exposure Scenarios
Target Organ
System
Species
Duration
POD
(mg/kg-day)
Effect
HEP"
(mg/kg-day)
HEC
(mg/m3)
[ppm]
Benchmark
MOE
Reference
Development
/Reproductive
Rat
5 to 14 days
throughout
gestation
BMDL5 = 9
| fetal
testicular
testosterone
2.1
12
[1.0]
UFa= 3
UFh=10
Total UF=30
_b
Abbreviations: POD = Point of Departure; HEC = human equivalent concentration; HED = human equivalent dose; MOE =
margin of exposure; UF = uncertainty factor BMDL5 = Benchmark dose (lower confidence limit) associated with a 5%
response level.
"EPA used allometric body weight scaling to the three-quarters power to derive the HED. Consistent with EPA Guidance
(U.S. EPA. 201 lb), the interspecies uncertainty factor (UFa). was reduced from 10 to 3 to account remaining uncertainty
associated with interspecies differences in toxicodynamics. EPA used a default intraspecies (UFH) of 10 to account for
variation in sensitivity within human populations.
b The BMDLs was derived through meta-regression and BMD modeling of fetal testicular testosterone data from eight studies
of DBP with rats (Gray et al.. 2021; Furr et al.. 2014; Johnson et al.. 2011; Starve et al.. 2009; Howdeshell et al.. 2008;
Martino-Andrade et al.. 2008; Johnson et al.. 2007; Kuhl et al.. 2007).
The POD of 9 mg/kg-day (HED = 2.1 mg/kg-day) will be used in the Draft Risk Evaluation for Dibutyl
Phthalate (U.S. EPA. 2024m) to estimate acute, intermediate, and chronic non-cancer risk. EPA
summarizes the cancer hazards of DBP in a separate technical support document, Draft Cancer Human
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.
2024a).
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
proposed POD for use in risk characterization of DBP.
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(DEHP), Dibutyl Phthalate (DBP), Butyl Benzyl Phthalate (BBP), Diisobutyl Phthalate (DIBP),
and Dicyclohexyl Phthalate (DCHP). Washington, DC: Office of Pollution Prevention and
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(DBP). Washington, DC: Office of Pollution Prevention and Toxics.
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Hazard Animal Toxicology and Epidemiology for Dibutyl Phthalate (DBP). Washington, DC:
Office of Pollution Prevention and Toxics.
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Toxicology for Dibutyl Phthalate (DBP). Washington, DC: Office of Pollution Prevention and
Toxics.
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Testosterone for Di(2-ethylhexyl) Phthalate (DEHP), Dibutyl Phthalate (DBP), Butyl Benzyl
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U.S. EPA. (20240. Draft Non-cancer Human Health Hazard Assessment for Dibutyl Phthalate (DBP).
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(DCHP). Washington, DC: Office of Pollution Prevention and Toxics.
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(DEHP). Washington, DC: Office of Pollution Prevention and Toxics.
U.S. EPA. (20241). Draft Non-cancer Human Health Hazard Assessment for Diisobutyl phthalate
(DIBP). Washington, DC: Office of Pollution Prevention and Toxics.
U.S. EPA. (2024m). Draft Risk Evaluation for Dibutyl Phthalate (DBP). Washington, DC: Office of
Pollution Prevention and Toxics.
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Phthalate (DBP) - Literature Published from 2014 to 2019. Washington, DC: Office of Pollution
Prevention and Toxics.
U.S. EPA. (2024o). Draft Systematic Review Protocol for Dibutyl Phthalate (DBP). Washington, DC:
Office of Pollution Prevention and Toxics.
U.S. EPA. (2024p). Non-Cancer Human Health Hazard Assessment for Diisononyl Phthalate (DINP)
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2622 in China. Sci Total Environ 613-614: 1573-1578.
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2625 induced Oxidative Stress and Depression-like Behavior in Mice with or without Ovalbumin
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2627
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2628 APPENDICES
2629
2630 Appendix A Existing Assessments from Other Regulatory Agencies of DBP
2631 The available existing assessments of DBP are summarized in TableApx A-l, which includes details regarding external peer review, public
2632 consultation, and systematic review protocols that were used.
2633
2634 Table Apx A-l. Summary of Peer-review, Public Comments, and Systematic Review for Existing Assessments of DBP
Agency
Assessment(s) (Reference)
External
Peer-
Review?
Public
Consultation?
Systematic
Review
Protocol
Employed?
Remarks
U.S. EPA (IRIS
Program)
Phthalate exposure and male
reproductive outcomes: A systematic
review of the human epidemiological
evidence fRadke et al.. 2018)
Phthalate exposure and female
reproductive and developmental
outcomes: A systematic review of the
human epidemiological evidence fRadke
etal.. 2019b)
Phthalate exposure and metabolic
effects: A systematic review of the
human epidemiological evidence fRadke
et al.. 2019a)
Phthalate exposure and
neurodevelopment: A systematic review
and meta-analysis of hum an
epidemiological evidence fRadke et al..
2020a).
No
No
Yes
- Publications were subjected to peer review prior
to being published in a special issue of
Environment International
- Publications employed a systematic review
process that included literature search and
screening, study evaluation, data extraction, and
evidence synthesis. The full systematic review
protocol is available as a supplemental file
associated with each publication.
ATSDR
Toxicological profile for di-b-phthalate
(ATSDR. 2001)
Yes
Yes
No
- Draft reviewed by peer review panel of four
experts (see v. xi of (ATSDR. 2001) for more
details).
U.S. CPSC
Toxicitv review of di-n-butvlphthalate
(DBPf(U.S. CPSC. 2010)"
Yes
Yes
No
- Peer-reviewed by panel of four experts. Peer-
review report available at:
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Agency
Assessment(s) (Reference)
External
Peer-
Review?
Public
Consultation?
Systematic
Review
Protocol
Employed?
Remarks
Chronic Hazard Achnsory Panel on
Phthalates andPhthalate Alternatives
(U.S. CPSC. 2014)
https ://www.cpsc. aov/s3fs-public/Peer-Review-
RcDort-Commcnts.Ddf
-Public comments available at:
httos ://www.cpsc. gov/chap
- No formal systematic review protocol
employed.
- Details regarding CPSC's strategy for
identifying new information and literature are
provided on paae 12 of (U.S. CPSC. 2014)
NASEM
Application of systematic review
methods in an overall strategy for
evaluating low-dose toxicity from
endocrine active chemicals fNASEM.
2017;
Yes
No
Yes
- Draft report was reviewed by individuals
chosen for their diverse perspectives and
technical expertise in accordances with the
National Academies peer review process. See
Acknowledgements section of (NASEM. 2017)
for more details.
- Employed NTP's Office of Heath Assessment
and Translation (OHAT) systematic review
method
Health Canada
State of the science report: Phthalate
substance grouping: Medium-chain
phthalate esters: Chemical Abstracts
Sen'ice 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. 2015)
Supporting documentation: Evaluation
of epidemiologic studies on phthalate
compounds and their metabolites for
hormonal effects, growth and
development and reproductive
parameters (Health Canada. 2018b)
Yes
Yes
No (Animal
studies)
Yes
(Epidemiologic
studies)
- Ecological and human health portions of the
screening assessment reoort (ECCC/HC. 2020)
were subject to external review and/or
consultation. See oaae 2 of (ECCC/HC. 2020) for
additional details.
- State of the science reoort (EC/HC. 2015) and
draft screening assessment report for the
phthalate substance group subjected to 60-day
public comment periods. Summaries of received
public comments available at:
https://www.canada.ca/en/health-
canada/services/chemical-substances/substance-
aroumnas-initiativeA>hthalate.html#al
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Agency
Assessment(s) (Reference)
External
Peer-
Review?
Public
Consultation?
Systematic
Review
Protocol
Employed?
Remarks
Supporting documentation: Evaluation
of epidemiologic studies on phthalate
compounds and their metabolites for
effects on behaviour and
neurodevelopment, allergies,
cardiovascular function, oxidative stress,
breast cancer, obesity, and metabolic
disorders (Health Canada. 2018a)
Screening Assessment - Phthalate
Substance Grouvine (ECCC/HC. 2020)
- 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 is
orovidcd in Section 1 of (EC/HC. 2015) and
(ECCC/HC. 2020)
- Human epidemiologic studies evaluated using
Downs and Black Method (Health Canada.
2018a, b)
NICNAS
Priority existing chemical assessment
report no. 36: Dibutylphthalate
(NICNAS. 2013)
No
Yes
No
- NICNAS (2013) states "The reoort has been
subjected to internal peer review by NICNAS
during all stages of preparation." However, a
formal external peer review was not conducted.
- NICNAS (2013) states "Arolicants for
assessment are given a draft copy of the report
and 28 days to advise the Director of any errors.
Following the correction of any errors, the
Director provides applicants and other interested
parties with a copy of the draft assessment report
for consideration. This is a period of public
comment lasting for 28 days during which
requests for variation of the report may be made."
See Preface of (NICNAS. 2013) for more details.
- No formal systematic review protocol
employed.
- Details regarding NICNAS's strategy for
identifying new information and literature is
orovidcd in Section 1.3 of (NICNAS. 2013)
ECHA
Opinion on an Annex XV dossier
proposing restrictions on four phthalates
(DEHP, BBP, DBP, DIBP) (ECHA.
2017b)
Yes
Yes
No
- Peer-reviewed by ECHA" s Committee for Risk
Assessment (RAC)
- Subject to public consultation
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Agency
Assessment(s) (Reference)
External
Peer-
Review?
Public
Consultation?
Systematic
Review
Protocol
Employed?
Remarks
Annex to the Background document to
the Opinion on the Annex XV dossier
proposing restrictions on four ph thai ate s
(DEHP. BBP. DBP. DIBP) (ECHA.
2017a)
- No formal systematic review protocol
employed.
EFSA
Update of the Risk Assessment ofDi-
butylphthalate (DBP), Butyl-benzyl-
phthalate (BBP), Bis(2-
ethylhexyl)phthalate (DEHP'), Di-
isononylphthalate (DINP) and Di-
isodecylphthalate (DIDP) for Use in
Food Contact Materials (EFSA. 2019)
No
Yes
No
- Draft report subject to public consultation.
Public comments and EFSA's response to
comments are available at:
httos://doi.ora/10.2903/saefsa.2019.EN-1747
- No formal systematic review protocol
employed.
- Details regarding EFSA's strategy for
identifying new information and literature are
orovidcd on oaae 18 and Aoocndix B of (EFSA.
2019)
NTP-CERHR
NTP-CERHR Monograph on the
Potential Human Reproductive and
Developmental Effects of Di-n-Butyl
Phthalate (DBP) (NTP-CERHR. 2003b)
No
Yes
No
- Report prepared by NTP-CERHHR Phthalates
Expert Panel and was reviewed by CERHR Core
Committee (made up of representatives of NTP-
participating agencies, CERHR staff scientists,
member of phthalates expert panel)
- Public comments summarized in Appendix III
of CNTP-CERHR. 2003 a)
- No formal systematic review protocol
employed.
2635
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2636
2637
2638
2639
2640
2641
2642
2643
2644
2645
2646
2647
2648
2649
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Appendix B New Literature Considered for Non-Cancer Hazards
B.l Reproductive and Developmental Effects
EPA evaluated seven new studies that provide data on reproductive and developmental outcomes in
rodents following oral exposure to DBP. The data set included 3 intermediate duration studies (Zhang et
al.. 2018a; Ahmad et al.. 2015; Sen et al.. 2015). 1 subchronic study (Xie et al.. 20191 and 3 one-
generation studies (Xie et al.. 2016; de Jesus et al.. 2015; Ahmad et al.. 2014). These studies provided
data on the effect of DBP exposure on reproductive hormone levels, the estrus cycle, reproductive organ
weights, histopathological alterations of the uterus or ovary, and fertility, including evaluations of
sperm. Developmental endpoints included measures of pup body weight. The effects that were most
sensitive to DBP exposure (i.e., the lowest LOELs) included decreases in the levels of 17-P-estradiol
(E2) at doses ranging from 0.01 to 1 mg/kg-day (Xie et al.. 2019; Zhang et al.. 2018a; Sen et al.. 2015)
and decreased pup body weight (Ahmad et al.. 2014). However, each individual study had limitations
that contributed uncertainty that impacted interpretation of the results and therefore none were
considered further for dose response in Section 4. Detailed information on study designs is provided in
TableApx B-l.
In two rodent studies, (Zhang et al.. 2018a; Sen et al.. 2015) decreased E2 was observed at doses ranging
from 0.01 to 1 mg/kg-day, but these decreases did not consistently correspond with other reproductive
health effects (e.g., changes in histopathology or changes in estrus cyclicity). Zhang et al. (2018a)
exposed adolescent (PND21) Sprague-Dawley rats to 0, 1, 10, or 500 mg/kg-day DBP via gavage from
PND21 to 33 and observed increases in progesterone and relative uterus weight (non-dose-related) at 1
mg/kg-day. However, there were no measures of estrous cyclicity or attainment of puberty in this study.
Vaginal opening, a marker of puberty, was not assessed. Female rats (SD) start cycling on average at
PND 32, which is the first day of estrus. Without knowing if the females had begun to cycle, the data on
uterine weight and hormones are difficult to interpret. Uterine weights are increased due to the effects of
estradiol from growing follicles on the uterine epithelium. Uterine weight is highest on the day of
proestrus when estrogen levels reach their peak in the estrus cycle (Goldman et al.. 2007). The
uncertainty of these measures without cycle day data limits any interpretation of any of the results as the
variation in hormone levels and associated tissue changes are not aligned with puberty or cycle day.
Similar results were reported in Sen et al. (2015). where adolescent (PND35) female CD-I mice were
orally exposed to 0.01, 0.1, or 1,000 mg/kg/day DBP for 10 days. Decreased E2 was observed in the
0.01 mg/kg-day dose group, but no other effects were reported at that dose. At the next highest dose
group, 0.1 mg/kg-day, E2 was decreased with corresponding increases in serum FSH and LH, as well as
decreased number of antral ovarian follicles. It is plausible that DBP acts on the ovary to elicit these
effects, as E2 is produced by developing follicles. However, decreased E2 was also observed in the
high-dose group (1,000 mg/kg-day DBP) but the average number of antral follicles was increased, albeit
not at a statistically significant level. Moreover, similar to Zhang et al. (2018a). there is some
uncertainty in the data set from Sen et al. (2015). including the hormone analysis, uterine weights and
ovarian measures. Sen et al. (2015) did not measure the endpoints on the same day of diestrus (Dil or
Di2), which is problematic because E2 increases from Dil to Di2, as the follicles grow and secrete more
E2. In addition, the study was conducted immediately following the onset of puberty when cyclicity
inconsistent and no evaluation of normal cyclicity can be determined or compared between dose groups.
In addition to the aforementioned limitations, there were several factors that further increased
uncertainty in the data set of new studies of reproductive effects following DBP exposure including low
sample size and the lack of an appropriate dose range (very low or very high, 0.01, 0.1 or 1000 mg/kg).
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A study designed to evaluate cardiovascular outcomes (More information provided in Appendix B.4)
also provided data on serum E2 levels following a 6-week gavage exposure to 0.1, 1, or 10 mg/kg-day
DBP in adult male mice. A non-monotonic, "U" shaped dose-response was observed in E2, with
decreased E2 at the lowest dose tested (0.1 mg/kg-day), but increased E2 at higher dose levels (1 and 10
mg/kg-day) (Xie et al.. 2019).
A one-generation study by Xie et al. (2016) also provided data on reproductive hormone levels
following developmental exposures to DBP. Increased serum E2 was observed during specific phases of
the estrus cycle in adult F1 offspring following in utero and lactational exposure (GDI2 to PND21) to
doses as low as 10 mg/kg-day. Specifically, increased serum E2 was observed during proestrus, diestrus,
and metestrus in F1 females on PND63. These effects coincided with decreases in serum progesterone
during proestrus, estrus, diestrus, and metestrus in F1 females PND63. There was no dose-response, and
exposure to the mid and high doses (100 and 600 mg/kg-day) did not lead to significant increases in
these hormones across multiple phases of the estrus cycle as was observed in the low-dose group.
Furthermore, the changes in hormone levels at 10 mg/kg-day did not coincide with any functional
changes such as those in estrus cyclicity, onset of vaginal opening, uterus weights or ovarian weights
with doses tested up to 600 mg/kg-day. More data are needed to understand the impact of gestational
and/or lactational DBP exposure on ovarian development and function in adult F1 offspring following
maternal exposure.
Two additional studies provide data on reproductive and developmental effects in offspring following
maternal exposure to DBP (de Jesus et al.. 2015; Ahmad et al.. 2014). A one-generation reproductive
study by de Jesus et al. (2015) reported reproductive effects in Mongolian gerbils at 5 mg/kg-day based
on histopathological effects in the prostate of F1 offspring (i.e., decreased epithelium height and
decreased SMC thickness) and increased weight of the prostatic complex (seminal vesicle, coagulating
gland, & dorsolateral, ventral & dorsal lobes) (de Jesus et al.. 2015). In that study, pregnant gerbils were
exposed to DBP from GD 0 to PND 28, then F1 (12 litters, 6 to 8 pups/litter) continued the same
exposure through study termination at PNW14. This study contained several issues that limit the
interpretation of results, including those related to chemical characterization (e.g., drinking water
exposure for a non-water-soluble phthalate), insufficient information on measures to reduce bias from
the litter effect, dose-range issues, and only one dose other than the control.
In a gestational exposure study by Ahmad et al. (2014). pregnant albino rats were gavaged with 0, 2, 10,
or 50 mg/kg-day DBP from GD14 to parturition, and endpoints were evaluated in F1 from PND1 to
PND75. Decreased pup body weight was observed at doses as low as 2 mg/kg-day in PND21 males
exposed to DBP from GDI4 to parturition. The reduction in body weight was dose-dependent (other
doses included 10 and 50 mg/kg-day). However, by adulthood (PND75), the effect was no longer dose-
responsive and was significant at the high dose only. Of note, reduced pup body weight in this study was
observed at a dose much lower dose than those which have been observed in the majority of other
studies cited in existing assessments (Section 3.1.2.2). Indeed, the aforementioned study by Lee et al.
(2004) (see Section 3.1.2.2) did not observe any change in pup body weight on PND21 following
exposure to doses as low as 2 mg/kg (equivalent to 1.5 to 3 mg/kg-day) for a longer duration than
Ahmad et al. (2014). Lee et al. included most of the critical window (i.e., the critical window is GD 14
to 19 and the exposure range in Lee et al., was GD15 to PND21). Changes in pup body weight are not
considered exposure-related.
A study by the same authors (Ahmad et al.. 2015) evaluated the estrogenic effects of DBP in a 3-day
uterotrophic assay and a 20-day pubertal assay, though several methodological limitations impact the
ability to interpret results and draw conclusions from these studies. In the uterotrophic assay, PND20
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female rats were exposed to 0, 10, or 100 mg/kg-day DBP once per day for 3 consecutive days via
gavage. Decreased uterine wet weight was observed one day after exposure ended in the 100 mg/kg-day
group, but the effect is difficult to interpret as there was also an increase in body weight (over 10
percent) in this dose-group. In the pubertal assay, PND21 female rats were exposed to 0, 10, or 100
mg/kg-day DBP for 20 days via gavage, and animals were examined daily for body weights, vaginal
opening (VO). The pubertal data are not conclusive; neither control nor DBP-exposed animals attained
puberty (i.e., first day of VO and the first day of estrus), although rats typically attain VO by PND32.
The authors also reported significant decreases in uterine and ovarian wet weights. However, the
females exposed to DBP had not yet begun to cycle, making it difficult to interpret the significance of
the observed decreases in uterine and ovarian weights in DBP-exposed animals, as well as the lack of
reporting of relative weights since there was a decrease in body weight. Altogether, these data do not
suggest that DBP is an estrogen agonist, as it would have increased the uterine weight in the three-day
uterotrophic assay.
New studies have provided data on developmental and reproductive health outcomes other than the male
reproductive system following exposure to DBP. However, the data set still contains several limitations
that increase uncertainty. Moreover, while decreased E2 (Xie et al.. 2019; Zhang et al.. 2018a; Sen et al..
2015) or decreased pup body weight (Ahmad et al.. 2014) were observed at doses lower than some of
the most sensitive PODs (i.e., 2 mg/kg (equivalent to 1.5 to 3 mg/kg-day) in Lee et al. (2004)) identified
in existing assessments (e.g., (ECHA. 2017a; OEHHA. 2007 EFSA. 2019. 6548141)). the data set is not
sufficiently robust given the amount of uncertainty resulting from the limitations in each individual
study. Additionally, the increased E2 and progesterone observed in adult offspring following in utero
and lactational exposure to 5 mg/kg-day do not coincide with other functional reproductive endpoints
which makes it difficult to interpret the biological relevance of the changes in hormone levels (Xie et al..
2016). Concerns with other studies included evidence of a transient effect on body weight (Ahmad et al..
2014) and study design limitations (de Jesus et al.. 2015). Therefore, EPA did not further consider these
six new studies on reproductive effects for POD selection (Section 4).
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2758 TableApx B-l. Summary of New Animal Toxicology Studies Evaluating Effects on the Developmental and Reproductive System
2759 Following Exposure to DBP
Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-
day)
Effect at LOAEL
Remarks
(Sen et al.. 2015)
Adolescent (PND35) female CD-I
(8/dose) were mice exposed to DBP
at concentrations of 0.01, 0.1, or
1,000 mg/kg/day for 10 days via
administered orally by placing a
pipette tip containing the dosing
solution into the mouth past the
incisors and into the cheek pouch.
ND/0.01
(LOEL)
I serum E2
Effects at 0.1 mg/kg-dav
-t FSH; t LH; j E2
- j No. of antral ovarian follicles
-[ relative liver weight
Effects at 1.000 mg/kg-dav
-t FSH; | E2
-Changes in estrus cycle (J. time in proestrus/estrus and
t time in metestrus/diestrus)
- j No. of corpora lutea
Limitations
- Poorly designed study without adequate estrous cycle
assessments
- Large dose spacing; many effects non-monotonic or
displayed flat D-R; small sample size (n =8)
(Xie et al.. 2019)
Male C57BL/6 mice (9/group) were
exposed via gavage to 0.1, 1, or 10
mg/kg-day DBP for 6 weeks.
ND/0.1
(LOEL)
I serum E2;
Effects at 1 mg/kg-dav
-| serum E2
ffects at 10 mg/kg-dav
-| serum E2
imitations
-Study only included males
(Zhang et al.. 2018a)
Adolescent (PND21) Sprague-
Dawley rats (10/group) were
exposed to 0, 1, 10, or 500 mg/kg-
day DBP via gavage from PND21-
33.
ND/1
t serum progesterone J.
serum E2; changes in
ovarian histopathology;
t relative weight of
uterus
Effects at 10 mg/kg-dav
-1 serum E2; | progesterone
Effects at 500 mg/kg-dav
-[ serum E2; [ serum progesterone
Limitations
-Only evaluated females; Large dose spacing;
Qualitative histopathology
(Ahmad et al.. 2014)
Pregnant albino rats (6-9/group)
were gavaged with 0, 2, 10, or 50
mg/kg-day DBP from GDI4 -
ND/2
(LOEL)
I pup body weight on
PND21 (males)
Maternal Effects
-1 maternal BW gain
Developmental Effects
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Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-
day)
Effect at LOAEL
Remarks
parturition Endpoints evaluated in
F1 from PND1-PND75.
-1 pup BW on PND1 (10 & 50 mg/kg-day) & PND21
(2, 10 and 50 mg/kg-day)
-1 BW in Fl adults on PND75 (50 mg/kg-day)
Reproductive Effects in adult Fl
-1 absolute weight of epididymis, testis, prostate, &
seminal vesicle in Fl adults on PND75 (50 mg/kg-day)
-1 sperm count, J. percent motile sperm, |percent
abnormal sperm (50 mg/kg-day)
Other effects:
-1 absolute weight of adrenal gland, liver & kidney in
Fl adults on PND75 (50 mg/kg-day)
Unaffected Outcomes
- Serum testosterone in Fl adults (PND75); Litter size,
live/dead pups, sex ratio (PND1); Anogenital distance
(PND5 & PND25);Viability index (PND4); Weaning
index (PND21)
Limitations:
- No statistical method to account for litter effects (i.e.,
statistics on offspring presented as means of individual
animals rather than litter means)
(de Jesus et al.. 2015)
Pregnant Mongolian gerbils (12
dams /group) exposed via drinking
water to 5 mg/kg-day DBP at
concentrations of GD 0 to PND 28,
then F1 (12 litters, 6-8 pups/litter)
continued exposure through
PNW14.
ND/5
Prostate histopathology
in Fl; | wet weight of
the prostatic complex
(seminal vesicle,
coagulating gland, &
dorsolateral, ventral &
dorsal lobes)
Limitations
- DBP administered in drinking water (solubility
concerns)
- Insufficient information on measures to reduce bias
from the litter effect
- Unconventional experimental animal used (gerbils)
- Study only used one dose other than control
(Xie et al.. 2016)
Pregnant Sprague-Dawley rats
(8/group) were exposed to 0, 10,
1000, or 600 mg/kg-day DBP via
gavage from GDI2 - PND21, and
F1 female evaluated on PND63.
ND/10
t serum E2 during
proestrus, diestrus, &
metestrus in Fl females
(PND63); I serum
progesterone during
proestrus, estrus,
Effects at 100 ma/ka-dav
- 1 serum progesterone during proestrus in Fl females
Effects at 600 ma/ka-dav
- 1 serum progesterone during proestrus in Fl females
Limitations:
- Large dose spacing
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Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-
day)
Effect at LOAEL
Remarks
diestrus, and metestrus
in F1 females (PND63)
- Changes in hormone levels did not coincide with
changes in other reproductive outcomes
Abbreviations: I = statistically significant decrease; t = statistically significant increase; NOAEL = no observed adverse effect level; LOAEL = Lowest-observed-
adverse-effect level; LOEL = Lowest observed effect level; GD = Gestation Day; PND = Postnatal Day; PNW = Postnatal Week; F1 = First generation offspring; E2 =
(^-estradiol: FSH = follicle stimulating hormone; LH = Luteinizing Hormone; BW = body weight; ND = No data
2760
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B.2 Neurotoxicity
Three studies in male mice identified alterations in neurological health outcomes following DBP
exposure (Farzanehfar et al.. 2016; Yan et al.. 2016; Zuo et al.. 2014). These studies provided data on
neurobehavioral effects, sometimes paired with brain histopathology. Behavioral alterations were
observed at doses as low as 0.45 mg/kg-day (Zuo et al.. 2014). albeit in a study with several limitations.
Ultimately, no studies were considered further for dose response in Section 4. Detailed information on
the study designs are provided in TableApx B-2.
Zuo et al. (2014) observed reduced performance in the tail suspension test (TST; increased time spent
immobile) in male BALB/c mice that had been exposed to 0.45 mg/kg-day DBP via gavage for 32 days.
The TST is considered a proxy for depression-like behavior in mice. The test is conducted by
suspending a mouse by its tail and recording the duration of time spent immobile or hanging passively.
A normal mouse will be more mobile and try to free itself, but a mouse that exhibits depression-like
behavior with not. The authors observed increase time spent immobile in the TST at the next highest
dose level or 45 mg/kg-day as well. Effects at the high dose, but not the low dose, coincided with other
neurobehavioral effects, such as reduced performance in the forced swim test (FST), specifically
increased time spend time spent immobile. Performance in the FST is another proxy for depression-like
behavior in rodents. In the FST, normal mice will struggle to free themselves from the water to escape,
but a mouse with depression-like behavior spends more time floating passively in the water without
struggling. An important consideration of both the TST and FST is that they both involve a motor
component, and correct interpretation relies on paring these tests with specific tests that evaluate motor
function in the animals. Other limitations of this study include: subjective outcome measures for
behavioral examinations; failure to state measures to reduce observer bias (i.e., blinding); insufficient
detail on the order in which neurobehavior tests were conducted; and restricting the experiment to male
animals without justification. EPA is not considering neurological endpoints in Zuo et al. (2014) further
for POD selection based on reporting deficiencies that compromise the ability to interpret results of the
study.
Similar limitations were noted in Yan et al. (2016). Yan et al. reported reduced performance in the
elevated plus maze (EPM; decreased time spent in the open arms) at in male Kunming mice exposed to
5 mg/kg-day DBP via gavage for 28 days. The EPM can be considered a proxy for anxiety-like behavior
in rodents and is a type of maze that has sections that are open (with no top, just walls) and closed/dark
(walls and a closed top). Mice that exhibit an anxiety-like behavior will spend more time in the closed
arms than the open arms. Other dose tested included 25 and 125 mg/kg-day and a dose-responsive
decrease in time spend in the open arms was observed. The majority of adverse neurobehavioral and
functional effects were observed at 25 and 125 mg/kg-day, which are summarized in Table Apx B-2.
Similar to the limitations in Zuo et al. (2014). Yan et al. (2016) only used male animals without
providing an explanation, did not present information on animal body weight, provided qualitative
histopathology data for the high dose only, and didn't report measures used to account of observer bias
in their tests.
Neurobehavioral effects were also observed in a study of male NMRI mice exposed to 0, 6.25, 12.5, 25,
50, 100, or 200 mg/kg-day DBP via gavage for 14 days (Farzanehfar et al.. 2016). Reduced exploratory
behavior in the open field test (OFT) was observed in mice exposed to 12.5 mg/kg-day DBP, reflected in
decreased total distance traveled and decreased percent time spent central to the peripheral zone. These
effects were also observed at the next highest dose level, in addition to decreased performance in the
EPM, where the mice exposed to 25 mg/kg-day DBP or higher spent more time in the closed arms. A
linear dose-dependent decrease in avoidance latency time in the passive avoidance test was observed,
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beginning at 25 mg/kg-day. Avoidance latency is one outcome measured in the passive avoidance test,
which is considered a proxy for long term memory in rodents. The test involves training mice to learn
that one of two compartments will deliver an electric shock, which a mouse will normally learn to avoid.
However, a mouse with a memory impairment may not avoid the room where they previously received
the electric shock or may venture into that room after some time has elapsed. Neurobehavioral deficits
observed at 25 mg/kg-day corresponded with histopathological changes in the granular cells of the
dentate gyrus (decreased nuclei area and condensation). While the authors observed neurobehavioral
effects at the 12.5 mg/kg-day dose (i.e., decreased total distance movement), they do not present
histopathological data for animals at this dose or the low dose of 6.25 mg/kg-day. No changes in rotarod
performance or forelimb grip strength were observed at any dose level, suggesting that the reductions in
performance in the passive avoidance test and OFT were not likely to be explained by deficits in motor
function. Moreover, the increase in avoidance latency, paired with the locomotor data (i.e., decreased
exploratory behavior in OFT), yet exposed animals reenter dark compartment of the EPM), suggest that
the memory of the negative stimulus delivered in the passive avoidance test is the result of impaired
learning and memory. This study was more well designed than those of Zuo et al. (2014). Yan et al.
(2016) and the methods were sufficiently detailed for neurobehavioral examinations, but several other
limitations of this study exist, including only using males without an explanation, and the lack of
histopathology data that correspond with the LOAEL. These data provide a LOAEL of 12.5 mg/kg-day
based neurobehavioral effects following al4-day exposure in adults. However, the LOAEL's identified
for reproductive and developmental effects are more well supported by a robust database and are
sometimes more sensitive. Although there is some evidence of neurotoxicity following exposure to DBP
in experimental animals, EPA is not further considering these effects for dose-response assessment or
for use in extrapolating human risk in Section 4. The database of experimental animal studies is not as
robust as that of developmental and reproductive health outcomes, which remains the most sensitive and
robust outcome from which to derive a POD.
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2834 Table Apx B-2. Summary of New Animal Toxicology Studies Evaluating Effects on the Nervous System Following Exposure to DBP
Reference
Brief Study
Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
(Zuo et al.. 2014)
Male BALB/c mice
(8/group) were
exposed to 0, 0.45,
or 45 mg/kg-day
DBP via gavage for
32 days.
Neurobehavioral
examinations
conducted days 36,
37, and 39 prior to
study termination
on day 40.
ND/0.45
| immobile time in
TST on day 37
Effects at 45 mg/kg-dav
- j immobile time in FST & TST
Unaffected Outcomes:
- OFT endpoints (defecation numbers; distance in outer ring) on day 36; relative
brain weight on day 40
Limitations:
- Only male animals were evaluated; Not guideline; Quantitative data for FST,
authors do not state use of measures to reduce observer bias (i.e., blinding);
Insufficient detail on training period for behavior tests;
(Tan et al.. 2016)
Male Kunming
mice (9/group)
exposed to 0, 5, 25,
or 125 mg/kg-day
DBP via gavage for
28 days.
ND/5
Neurobehavioral
changes: J. percent
time spent in open
arms (EPM)
Effects at 25 mg/kg-dav
- 1 number of open arm entries (EPM); [ percent time spent in open arms
(EPM); I total distance traveled (OFT); | percent distance in outer ring (OFT);
t defecations (OFT)
Effects at 125 mg/kg-dav
- 1 number of open arm entries (EPM); [ percent time spent in open arms
(EPM); I total distance traveled (OFT); | percent distance in outer ring (OFT);
t defecations (OFT)
- t histopathological observations (damaged cells, hippocampal CA1 region)
- 1 relative brain weight (brain coefficient)
Limitations:
- Only male animals were evaluated; No data provided on animal body weight;
Qualitative histopathology results provided only for high dose group; Authors
do not state use of measures to reduce observer bias (i.e., blinding)
(Farzanehfar et
Male NMRI mice
(10/group) were
exposed to 0, 6.26,
12.5,25,50, 100, or
200 mg/kg-day
DBP for 14 days
via gavage.
6.25/12.5
Neurobehavioral
changes:! total
distance (OFT); J.
percent time spent
central to peripheral
zone (OFT)
Effects at 25 mg/kg-dav
al.. 2016)
-[ total distance (OFT); [ percent time spent central to peripheral zone (OFT); [
percent time spent in open arm (EPM); [ avoidance latency time
- Histopathological findings in granular cells of dentate gyrus (J. nuclei area and
condensation)
Effects at 50 mg/kg-dav or higher
- 1 total distance (OFT); [ percent time spent central to peripheral zone (OFT);
I percent time spent in open arm (EPM); [ avoidance latency time
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Reference
Brief Study
Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
- Histopathological findings in granular cells of dentate gyrus (J. nuclei area and
condensation)
Limitations:
- Histopathology data were only provided for 25 and 100 mg/kg-day DBP
groups
Abbreviations: I = statistically significant decrease; t = statistically significant increase; NOAEL = no observed adverse effect level; LOAEL = Lowest-observed-
adverse-effect level; LOEL = Lowest observed effect level; EPM = Elevated Plus Maze; TST = Tail suspension test; FST = Forced Swim test; OFT = open field test
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B.3 Metabolic/Nutritional
Three new studies were identified that provided data on nutritional or metabolic effects following
exposure to DBP. The data set included one study (Ahmad et al.. 2015) in female rats exposed to DBP
for 20 days beginning on PND21, one 13-week study (Majeed et al.. 2017). and one one-generation
study (de Jesus et al.. 2015). These studies reported effects at low doses ranging from 5 to 10 mg/kg-
day. Detailed information on the study designs are provided in Table Apx B-3.
de Jesus et al. (2015) exposed pregnant Mongolian gerbils to 5 mg/kg-day DBP in drinking water from
GD 0 to PND28. After weaning the F1 offspring continued the same exposure as their mothers until
PNW14. Increased terminal body weight (approximately 8 percent) was observed in PNW14 offspring,
which coincided with increased adiposity index (approximately 35 percent) as well as decreased total
cholesterol, decreased serum LDL levels, and increased triglycerides. However, the results are difficult
to interpret because the study contained serious flaws that limit its use for deriving a robust POD,
including concerns regarding chemical administration in drinking water; DBP is not soluble in water.
Other limitations include insufficient information on measures to reduce observer bias or control for
intra litter correlations, and the study only used one dose other than control.
EPA identified a LOEL of 10 mg/kg-day in both Ahmad et al. (2015) and Majeed et al. (2017). Ahmad
et al. (2015) exposed adolescent female rats to 0, 10, or 100 mg/kg-day DBP via gavage from PND21 to
42. Decreased body weight gain was reported at PND27 (7.29 percent), PND33 (10.1 percent), and
PND43 (9.39 percent). Limitations of this study include the short exposure duration, low sample size
(6/group) and large dose spacing. Majeed et al. (2017) exposed male and female albino rats to 0, 10, or
50 mg/kg-day DBP for 13 weeks via feed and reported increased body weight gain, increased AC/TC
ratio, and decreased energy intake. This study was adequately designed (e.g., reported feed consumption
data, evaluated males and females, evaluated endpoints at several timepoints and in both sexes).
However, it is difficult to reconcile the biological plausibility of increased body weight gain and
increased body size (AC/TC ratio) given other known effects of DBP, namely decreased testosterone,
which would more likely coincide with a decrease in body weight. Although it is possible that DBP acts
through a different mechanism to elicit these effects. Nevertheless, even though these studies provide
some evidence of metabolic effects following exposure to DBP in experimental animals, EPA is not
further considering these effects for dose-response assessment or for use in extrapolating human risk.
The database of experimental animal studies is not as robust as that of developmental and reproductive
health outcomes, which remains the most sensitive and robust outcome from which to derive a POD.
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Table Apx B-3. Summary of New Animal Toxicology Studies Evaluating Effects on IV
etabolism Following Exposure to DBP
Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
(de Jesus et al..
Pregnant Mongolian gerbils (12
dams /group) exposed via
drinking water to 5 mg/kg-day
DBP at concentrations of GD 0
to PND 28, then F1 (12 litters, 6-
8 pups/litter) continued the same
exposure through PNW14 (study
tenn ination).
ND/5
I Total cholesterol; J. serum
low density lipoprotein (LDL)
levels; | serum triglycerides
t terminal (PNW14) body
weight (~8%)
t adiposity index (-35%)
Limitations
2015)
- DBP administered in drinking water (solubility
concerns)
- Insufficient information on measures to reduce bias
from the litter effect
- Unconventional experimental animal used (gerbils)
- Study only used one dose other than control
(Maieed et al..
Male and female albino rats
(24/sex/dose) were exposed to 0,
10, or 50 mg/kg-day DBP via
diet for 13 weeks.
Anthropometric measures
recorded after 0, 45, or 90 (i.e.,
study termination) days of
exposure.
ND/10
t BW gain (males) &
| AC7TC ratio at study
termination; [ energy intake
(females)
Effects at 50 ma/ka-dav:
2017)
t BW gain (males only); | BMI (males only)
t energy intake (males); J. energy intake (females)
t Glucose (10 mg/kg-day)
t total serum cholesterol
Other Effects
- Change in relative liver weight
- | ALP (females); t ALT; | albumin
Limitations:
-Organ weight data and most serum chemistry
presented as pooled values for both sexes, with result
of analysis for sex by treatment effect provided &
authors provide insufficient information to discern
directionality and magnitude of the effect specific to
each sex.
(Ahmad et al..
Female rats (6/group) were
exposed to 0, 10, or 100 mg/kg-
day DBP via gavage from
PND21 -42.
ND/10
I BW at multiple timepoints
(PND27, 33, & 42)
Effects at 100 ma/ka-dav
2015)
- 1 BW at multiple timepoints (PND27, 33, & 42)
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Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
Abbreviations: I = statistically significant decrease; t = statistically significant increase; NOAEL = no observed adverse effect level; LOAEL = Lowest-observed-
adverse-effect level; LOEL = Lowest observed effect level; AGD = anogenital distance; GD = gestation day; PND = postnatal day; AC/TC = abdominal
circumference to thoracic circumference ratio; BW = body weight; ALP = Alkaline phosphatase; ALT = alanine aminotransferase; BMI = body mass index
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B.4 Cardiovascular Health Effects
EPA identified one subchronic study that was designed to evaluate cardiovascular outcomes (Xie et al..
2019). The study provided data on histopathological alterations in the heart and aorta of male C57BL/6
mice exposed to 0.1, 1, or 10 mg/kg-day DBP via gavage for 6 weeks (Table_Apx B-4). At 0.1 mg/kg-
day, the authors observed increased vascular wall thickness of the aortic vessels and increased ACE
staining density in the thoracic aorta based on quantitative histopathology. Increased vascular wall
thickness of aortic vessels was also observed at 10 mg/kg-day, but not 1 mg/kg-day. There were several
limitations of the study including only including male animals and inconsistent reporting of results for
mean blood pressure (in the figure, it is noted as a significant increase at 10 mg/kg-day, while the
running text indicates no effect at this dose). The inconsistency reduces confidence in the study
reporting overall. Despite the sensitive LOAEL for cardiovascular outcomes, the limitations of the
single study available introduce enough uncertainty that EPA is not selecting a POD based on
cardiovascular effects.
TableApx B-4. Summary of New Animal Toxicology Study Evaluating Effects on the
Cardiovascular Sysi
tem Following Exposure to DBP
Reference
Brief Study
Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at
LOAEL
Remarks
(Xie et al.. 2019)
Male C57BL/6 mice
were exposed via
gavage to 0.1, 1, or
10 mg/kg-day DBP
for 6 weeks.
ND/0.1 (LOEL)
t vascular wall
thickness of
aortic vessels; t
ACE staining
density in the
thoracic aorta;
Effects at 1 ma/ka-dav
t serum E2
Effects at 10 ma/ka-dav
t serum E2; [ ACE
staining density in the
thoracic aorta
Limitations
-Study only included
males
-Inconsistent reporting of
results for mean BP
reported in text vs
Figure2B
Abbreviations:J, = statistically significant decrease; t = statistically significant increase; NOAEL = no observed adverse
effect level; LOAEL = Lowest-observed-adverse-effect level; LOEL = Lowest observed effect level; E2 = 17(3- estradiol;
ACE = Angiotensin Converting Enzyme; BP = blood pressure
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B.5 Immune adjuvant effects
EPA identified two new studies that provide data on the immune adjuvant properties of DBP following
intermediate duration exposure to DBP. LOELs based on immune adjuvant effects in male BALB/c
mice in two studies (Li et al.. 2014; Zuo et al.. 2014). Details on the study designs for each study are not
provided in the text for brevity and are instead summarized in TableApx B-5. The aforementioned
study by Zuo et al. (2014) was designed to evaluate the relationship between atopic allergy and
neurobehavioral effects in male mice exposed to DBP, and therefore included a second set of animals
that were sensitized with ovalbumin (OVA) antigen in addition to exposure to DBP and challenged via
OVA aerosol leading up to neurobehavioral testing. While exposure to DBP alone did not affect
performance in the open field test, exposure to DBP and OVA (both 0.45 mg/kg-day and 45 mg/kg-day
doses) resulted in changes in one parameter measured in the open field test, an increase in distance in the
outer ring. This result implies OVA exposure exacerbated the reduction in performance in one parameter
measured in the open field test. OVA exacerbated the reduction in performance in the TST and FST at
the high-dose only. Serum IgE and IL-4 were increased in all groups exposed to OVA relative to their
saline-controls (i.e., groups that received no antigen). IgE was increased, and IL-4 was decreased in
animals exposed to 45 mg/kg-day DBP (no OVA) relative to untreated controls (no OVA). A second
study by Li et al. (2014) exposed mice for 40 days to DBP via dermal application in addition to
sensitization with FITC via dermal application to their backs. Mice were challenged with FITC
application to their right ear prior to behavioral testing. Dermal sensitization and immunological effects
were observed in mice exposed to 4 mg/kg-day DBP + FITC relative to the comparator group (0 mg/kg-
day DBP + FITC). Specifically, mice from the 4 mg/kg-day DBP + FITC group had increased ear
swelling, increased bilateral ear weight, and histopathological changes in the ear such as an increased
number of infiltrating inflammatory cells and degranulating mast cells. Other effects included an
increase in cytokines and other molecules associated with inflammation in ear tissues (Table Apx B-5).
Although these studies provide some evidence for immune adjuvant effects of DBP in sensitized
animals, EPA is not further considering these effects for dose-response assessment or for use in
extrapolating human risk. Several sources of uncertainty reduce EPA's confidence in this outcome. First,
the database of new experimental animal studies that provide data on immune effects of DBP is limited
to two studies, each in male mice of the same strain, so it is difficult to understand effects in other sexes,
strains, or species. Second, available studies evaluate the adjuvant properties of DBP in experimental
rodent models pre-sensitized by exposure to other compounds (i.e., FITC, OVA). This co-exposure to
DBP and other compounds is another source of uncertainty that further reduced EPA's confidence in
this outcome. EPA is not further considering immune adjuvant effects for dose-response analysis or for
use in estimating risk to human health.
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2926 Table Apx B-5. Summary of New DBP Studies Evaluating Effects on the Immune System
Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
(Zuo et al.. 2014)
Male BALB/c mice
(8/group) exposed to 0,
0.45, 45 mg/kg-day DBP
via gavage, 32 days with
OVA sensitization via s.c.
injection on days 7, 21, and
28. Mice challenged with
aerosolized OVA for 30
mins each day on days 33-
39. Neurobehavioral
examinations conducted
days 36, 37, 39 prior to
study term, on day 40.
Groups include: OVA + 0
mg/kg-day DBP
(comparator), OVA + 0.45
mg/kg-day DBP, or OVA +
45 mg/kg-day DBP.
ND/0.45
| immobile time in FST (not
dose-dependent) & TST
(dose-dependent); t distance
in outer ring on day 36
(OFT); t serum IgE & jIL-4
(neither are dose-dependent)
Effects at 45 ma/ka-dav
Neurologic
| immobile time in FST & TST
t distance in outer ring on day 36 (OFT)
Immune
I relative spleen weight
Limitations:
Only male animals were evaluated
Not guideline; insufficient detail on recording
equipment - data collection presumed to not have
been automated.
Quantitative data for FST, TST, and OFT based on
highly subjective, qualitative observations & and
authors do not state use of measures to reduce
observer bias (i.e., blinding) (observer bias away
from null)
Insufficient detail on training period for behavior
tests
(Li et al.. 2014)
Male BALB/c mice
(8/group) exposed to 0, 0.4,
4, 40 mg/kg-day DBP, 40
days via dermal application
to their shaven backs. Mice
were sensitized with FITC
via dermal application to
backs; on day 41 and 42
(i.e., after the exposure
period); challenged with
FITC (ear) on day 47.
Groups: 0 mg/kg-day DBP
+ FITC (comparator), 0
mg/kg-day DBP + saline,
0.4 mg/kg-day DBP +
FITC or saline, 4 mg/kg-
day DBP + FITC or saline,
0.45/4
Dermal sensitization and
immunological effects: t ear
swelling; | bilateral ear
weight; quantitative
histopathological changes (t
no. infiltrating inflammatory
cells; t degranulating mast
cells in the ear); | ECP &
TSLP in ear tissues; t
cytokines IL-4, IL-5, IL-13,
& IL-17A in ear tissue
Effects at 40 ma/ka-dav:
t serum IgE 24 hours after FITC challenge
t ear swelling; | bilateral ear weight
t quantitative histopathological changes in the ear
t cytokines IL-4, IL-5, IL-13, & IL-17A in ear tissue
t ECP & TSLP in ear tissue
Limitations
Only male animals were evaluated
Study did not evaluate T cell subpopulations in
primary or secondary immune organs (i.e., spleen,
thymus, lymph nodes)
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Reference
Brief Study Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
40 mg/kg-day DBP + FITC
or saline.
Abbreviations:J, = statistically significant decrease; t = statistically significant increase; NOAEL = no observed adverse effect level; LOAEL = Lowest-observed-
adverse-effect level; LOEL = Lowest observed effect level; OVA = ovalbumin; s.c. = subcutaneous; IL = interleukin; IgE = immunoglobulin E; OFT = open field test;
FST = Forced Swim test; TST = tail suspension test; TSLP = thymic stromal lymphopoietin; ECP = eosinophil cationic protein; FITC = Fluorescein isothiocyanate
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Appendix C Fetal Testicular Testosterone as an Acute Effect
Several studies of experimental animal models are available that investigate the antiandrogenic effects
of DBP following single dose, acute exposures. Available studies indicate a single acute exposure during
the critical window of development (i.e., GD 14 to 18) 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
(P450scc 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 (P450scc Cypllal,
Cypl7al, Scarbl) hours post-exposure, and protein levels of these genes were reduced 6 to 12 hours
post-exposure. Additionally, studies by Carruthers et al. (2005) further demonstrate that exposure to as
few as two oral doses of 500 mg/kg DBP on successive days between GDs 15 to 20 can reduce male pup
AGD, cause permanent nipple retention, and increase the frequency of reproductive tract malformations
and testicular pathology in adult rats that received two doses of DBP during the critical window.
In summary, studies of DBP provide evidence to support use of effects on fetal testosterone as an acute
effect. However, the database is limited to just a few studies of DBP that test relatively high (500 mg/kg)
single doses of DBP, which contributes additionally uncertainty.
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Appendix D Calculating Daily Oral Human Equivalent Doses and
Human Equivalent Concentrations
For DBP, all data considered for PODs are obtained from oral animal toxicity studies in rats or mice.
Because toxicity values for DBP are from oral animal studies, EPA must use an extrapolation method to
estimate human equivalent doses (HEDs). The preferred method would be to use chemical-specific
information for such an extrapolation. EPA identified one study reporting a diffusion4imited, pH
trapping PBPK model for DBP and MBP (Keys et al.. 2000). However, the model was not fit for
purpose {i.e., the model was developed to predict blood concentrations of DBP and MBP following oral
exposure in the rat, not to extrapolate HEDs between species). EPA relied on the guidance from U.S.
EPA (2011b). 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 D-l.
EquationApx D-l. Dosimetric Adjustment Factor
/BWa\1/4
DAF=W)
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 (2011b). 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 DBP risk evaluation (U.S. EPA. 2011a). EPA used the body weights of 0.025 kg
for mice and 0.25 kg for rats, as presented in U.S. EPA (2011b). 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 uncertainty factor (UFa) used to set the benchmark MOE; the
default value of 10 can be decreased to 3, which accounts for any toxicodynamic differences that are not
covered by use of BW3 4. Using the appropriate DAF from Equation Apx D-l, EPA adjusts the POD to
obtain the HED using Equation Apx D-2:
Equation Apx D-2. Daily Oral Human Equivalent Dose
HEDDaiiy = PODDaiiy X DAF
Where:
HEDDaiiy = Human equivalent dose assuming daily doses (mg/kg-day)
PODDaiiy = Oral POD assuming daily doses (mg/kg-day)
DAF = 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
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extrapolated the daily oral HEDs to inhalation HECs using a human body weight and breathing rate
relevant to a continuous exposure of an individual at rest, as follows:
EquationApx D-3. Extrapolating from Oral HED to Inhalation HEC
iirn r BW a
HECDaily,continuous ~ HEDDaily X ( )
lA^ * C U Q
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. Equation Apx D-4 presents the conversion of
the HEC from mg/m3 to ppm.
Equation Apx D-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 DBP = 278.35 g/mol)
D.l DBP 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 9 mg/kg-day,
and the critical effect is decreased fetal testicular testosterone. The BMDLs was derived by a meta-
regression and BMD modeling of fetal testicular testosterone data from eight studies of DBP with rats
(Gray et al.. 2021: Furr et al.. 2014: Johnson et al.. 2011: Struve et al.. 2009: Howdeshell et al.. 2008:
Martino-Andrade et al.. 2008: Johnson et al.. 2007: Kuhl et al.. 2007). R code supporting NASEM's
original meta-regression and BMD analysis of DBP (NASEM. 2017) is publicly available on GitHub
(https://github.com/wachiuphd/NASEM-2017-Endocrine-Low-Dose). This non-cancer POD is
considered protective of effects observed following all duration exposures to DBP.
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
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(DBP), Butyl Benzyl Phthalate (BBP), Diisobutyl Phthalate (DIBP), Dicyclohexyl Phthalate (DCHP),
andDiisononylPhthalate (U.S. EPA. 2024g).
EPA used EquationApx D-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 D-2:
mq mq
2.1 — = 9- — X 0.236
kg — day kg — day
EPA then calculated the continuous HEC for an individual at rest using Equation Apx D-3:
mq mq 80 kq
12 —j = 2.1- x( )
m kg day 0.6125* 24 hr
hr
Equation Apx D-4 was used to convert the HEC from mg/m3 to ppm:
mg 24.45
1.0 ppm = 13 —=- x ————
m3 278.35
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Appendix E Considerations for Benchmark Response (BMR) Selection
for Reduced Fetal Testicular Testosterone
E.l Purpose
EPA has conducted an updated meta-analysis and benchmark dose modeling (BMD) analysis of
decreased fetal rat testicular testosterone (U.S. EPA. 2024g). 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. 2024q). 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.)
E.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. 2024g).
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 data sets 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-
ethylhexyl) Phthalate (DEHP), Dibutyl Phthalate (DBP), Butyl Benzyl Phthalate (BBP), Diisobutyl
Phthalate (DIBP), andDicyclohexylPhthalate (DCHP) (U.S. EPA. 2024g).
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E.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 E-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 E-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 E-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 E-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
E-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
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).
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E.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 reflect 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 E-l). 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 health protective.
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TableApx E-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. 2024k)
DBP
BMDL5 = 9
(J, fetal testicular testosterone)
30
(t Testicular Pathology)
14 [9, 27]
29 [20, 54]
149 [101,247]
(U.S. EPA. 2024i)
DIBP
BMDL5 = 24
(J, fetal testicular testosterone)
125
(t Testicular Pathology)
_b
55 [NA, 266f
279 [136, 517]
(U.S. EPA. 20241)
BBP
NOAEL = 50
(phthalate syndrome-related effects)
100
(IAGD)
_b
_b
284 [150, 481]
(U.S. EPA. 2024h)
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. 2024i)
DINP
BMDL5 = 49
(J, fetal testicular testosterone)
600
(J, Sperm motility)
74 [47, 158]
152 [97, 278]
699 [539, 858]
(U.S. EPA. 2024d)
Abbreviations: AGD = anogenital distance; BMD = benchmark dose; BMDL = lower 95% confidence limit on BMD; CI = 95% confidence interval; LOAEL = lowest
observed-adverse-effect level; NOAEL = no observed-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.
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