PUBLIC RELEASE DRAFT
DECEMBER 2024
EPA Document# EPA-740-D-24-029
December 2024
Office of Chemical Safety and
Pollution Prevention
xvEPA
United States
Environmental Protection Agency
Draft Non-cancer Human Health Hazard Assessment for
Butyl Benzyl Phthalate (BBP)
Technical Support Document for the Draft Risk Evaluation
CASRN 85-68-7
December 2024
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
27 TABLE OF CONTENTS
28 ACKNOWLEGEMENTS 7
29 SUMMARY 8
30 1 INTRODUCTION 11
31 1.1 Human Epidemiologic Data: Approach and Preliminary Conclusions 11
32 1.2 Laboratory Animal Findings: Summary of Existing Assessments, Approach, and
33 Methodology 13
34 1.2.1 Existing Assessments of BBP 13
35 1.2.2 Approach to Identifying and Integrating Laboratory Animal Data 16
36 1.2.3 New Literature Identified and Hazards of Focus for BBP 18
37 2 TOXICOKINETICS 20
38 2.1 Oral Route 20
39 2.2 Inhalation Route 23
40 2.3 Dermal Route 23
41 2.4 Summary 25
42 3 NON-CANCER HAZARD IDENTIFICATION 27
43 3.1 Effects on the Developing Male Reproductive System 27
44 3.1.1 Summary of Available Epidemiological Studies 27
45 3.1.1.1 Previous epidemiology assessment (conducted in 2019 or earlier) 27
46 3.1.1.1.1 Health Canada (2018b) 28
47 3.1.1.1.2 Radkeetal. (2019b; 2018) 29
48 3.1.1.1.3 NASEM report (2017) 31
49 3.1.1.1.4 Summary of the existing assessments of male reproductive effects 31
50 3.1.1.2 EPA summary of new studies (2018-2019) 32
51 3.1.2 Summary of Laboratory Animal Studies 34
52 3.1.2.1 Developing Male Reproductive System 42
53 3.1.2.2 New Literature Considered for Non-Cancer Hazard Identification 46
54 3.1.2.3 Other Developmental and Reproductive Outcomes 48
55 4 DOSE RESPONSE ASSESSMENT 49
56 4.1 Selection of Studies and Endpoints for Non-Cancer Health Effects 50
57 4.2 Non-cancer Oral Points of Departure for Acute, Intermediate, and Chronic Exposures 50
58 4.2.1 Studies with Lack of Dose-Response Sensitivity and Increased Uncertainty 54
59 4.2.2 Meta-analysis and BMD Modeling of Fetal Testicular Testosterone Data 55
60 4.2.3 Co-critical Studies Supporting a Consensus LOAEL of 100 mg/kg-day and NOAEL of 50
61 mg/kg-day 59
62 4.3 Weight of the Scientific Evidence 61
63 5 CONSIDERATION OF PESS AND AGGEGRATE EXPOSURE 64
64 5.1 Hazard Considerations for Aggregate Exposure 64
65 5.2 PESS Based on Greater Susceptibility 64
66 6 POINTS OF DEPARTURE USED TO ESTIMATE RISKS FROM BBP EXPOSURE,
67 CONCLUSIONS, AND NEXT STEPS 70
68 REFERENCES 72
Page 2 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
69 APPENDICES 83
70 Appendix A EXISTING ASSESSMENTS FROM OTHER REGULATORY AGENCIES OF
71 BBP 83
72 Appendix B NEW LITERATURE CONSIDERED FOR NON-CANCER HAZARDS 88
73 B. 1 Reproductive and Developmental Effects 88
74 B.2 Neurotoxicity 94
75 B.3 Immune adjuvant effects 96
76 B.4 Renal 96
77 B.5 Hepatic 97
78 Appendix C FETAL TESTICULAR TESTOSTERONE AS AN ACUTE EFFECT 98
79 Appendix D CALCULATING DAILY ORAL HUMAN EQUIVALENT DOSES AND
80 HUMAN EQUIVALENT CONCENTRATIONS 99
81 D. 1 BBP Non-cancer HED and HEC Calculations for Acute, Intermediate, and Chronic Duration
82 Exposures 100
83 Appendix E CONSIDERATIONS FOR BENCHMARK RESPONSE (BMR) SELECTION
84 FOR REDUCED FETAL TESTICULAR TESTOSTERONE 102
85 E.l Purpose 102
86 E.2 Methods 102
87 E.3 Results 103
88 E.4 Weight of Scientific Evidence Conclusion 104
89 Appendix F BENCHMARK DOSE MODELING OF FETAL TESTICULAR
90 TESTOSTERONE 106
91 F. 1 BMD Model Results of Howdeshell et al. (2008) 107
92 F.2 BMD Model Results of Gray et al. (2021) 110
93 F.3 BMD Model Results of Furr et al. (2014) (Block 36 Rats) 114
94 F.4 BMD Model Results of Furr et al. (2014) (Block 37 Rats) 120
95
96 LIST OF TABLES
97 Table 1-1. Summary of BBP Non-cancer PODs Selected for Use by Other Regulatory Organizations.. 14
98 Table 2-1. Metabolites of BBP Identified in Urine from Rats and Humans after Oral Administration... 21
99 Table 3-1. Summary of Scope and Methods Used in Previous Assessments to Evaluate the Association
100 Between BBP Exposure and Male Reproductive Outcomes 28
101 Table 3-2. Summary of Epidemiologic Evidence of Male Reproductive Effects Associated with
102 Exposure to BBP° 30
103 Table 3-3. Summary of BBP Oral Exposure Studies Evaluating Effects on the Developing Male
104 Reproductive System 36
105 Table 4-1. Dose-Response Analysis of Selected Studies Considered for Acute, Intermediate, and
106 Chronic Exposure Scenarios 51
107 Table 4-2. Effect of BBP Exposure on Fetal Testicular Testosterone Production0 56
108 Table 4-3. Summary of NASEM (2017) Meta-Analysis and BMD Modeling for Effects of BBP on Fetal
109 Testosterone0, b-c 57
110 Table 4-4. Overall Analyses of Rat Studies of BBP and Fetal Testosterone (Updated Analysis
111 Conducted by EPA using Metafor Version 4.6.0) 58
112 Table 4-5. Benchmark Dose Estimates for BBP and Fetal Testosterone in Rats 58
Page 3 of 122
-------
113
114
115
116
117
118
119
120
121
122
123
124
125
126
127
128
129
130
131
132
133
134
135
136
137
138
139
140
141
142
143
144
145
146
147
148
149
150
151
152
153
154
155
156
157
PUBLIC RELEASE DRAFT
DECEMBER 2024
Table 5-1. PESS Evidence Crosswalk for Biological Susceptibility Considerations 65
Table 6-1. Non-cancer HECs and HEDs Used to Estimate Risks 70
LIST OF FIGURES
Figure 1-1. Overview of BBP Human Health Hazard Assessment Approach 17
Figure 2-1. Proposed Metabolic Pathway of BBP Following Oral Exposure (Figure from Health Canada
(EC/HC, 2015b)) 23
Figure 3-1. Hypothesized Phthalate Syndrome Mode of Action Following Gestational Exposure 42
LIST OF APPENDIX TABLES
TableApx A-l. SUMMARY OF PEER-REVIEW, PUBLIC COMMENTS, AND SYSTEMATIC
REVIEW FOR EXISTING ASSESSMENTS OF BBP 83
TableApx B-l. Summary of New Animal Toxicology Studies Evaluating Additional Effects on the
Developmental and Reproductive System Following Exposure to BBP 90
Table_Apx B-2. Summary of New Animal Toxicology Study Evaluating Effects on the Nervous System
Following Exposure to BBP 95
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 .105
Table Apx F-l. Summary of BMD Model Results for Decreased Ex Vivo Fetal Testicular Testosterone
107
Table Apx F-2. Ex Vivo Fetal Rat Testicular Testosterone Data (Howdeshell et al., 2008) 107
Table Apx F-3. BMD Model Results Ex Vivo Fetal Testicular Testosterone (Howdeshell et al., 2008)
108
Table Apx F-4. Ex Vivo Fetal Rat Testicular Testosterone Data (Gray et al., 2021) 110
Table Apx F-5. BMD Model Results Ex Vivo Fetal Testicular Testosterone (All Dose Groups) (Gray et
al., 2021) 112
Table Apx F-6. Ex Vivo Fetal Rat Testicular Testosterone Data (Furr et al., 2014) (Block 36 Rats)... 115
Table Apx F-7. BMD Model Results Ex Vivo Fetal Testicular Testosterone (Block 36 - All Dose
Groups) (Furr et al., 2014) 116
Table Apx F-8. Ex Vivo Fetal Rat Testicular Testosterone Data (Furr et al., 2014) (Block 37 Rats)... 120
Table Apx F-9. BMD Model Results Ex Vivo Fetal Testicular Testosterone (Block 37 rats - All Dose
Groups) (Furr et al., 2014) 121
LIST OF APPENDIX FIGURES
FigureApx F-l. Frequentist Exponential Degree 3 Model of Howdeshell et al. (2008) data 109
FigureApx F-2. User Input of Frequentist Exponential Degree 3 Model of Howdeshell et al. (2008)
Data 109
FigureApx F-3. Model Results of Frequentist Exponential Degree 3 Model of Howdeshell et al. (2008)
Data 110
Figure Apx F-4. Frequentist Exponential Degree 3 Model of Gray et al. (2021) Data 113
FigureApx F-5. User Input of Frequentist Exponential Degree 3 Model of Gray et al. (2021) Data... 113
Figure Apx F-6. Model Results of Frequentist Exponential Degree 3 Model of Gray et al. (2021) Data
114
Page 4 of 122
-------
158
159
160
161
162
163
164
165
166
167
168
169
170
171
172
173
174
175
176
177
178
179
180
181
182
183
184
185
186
187
188
189
190
191
192
193
194
195
196
197
198
199
200
201
202
203
204
205
PUBLIC RELEASE DRAFT
DECEMBER 2024
KEY ABBREVIATIONS AND ACRONYMS
ADME
Absorption, distribution, metabolism, and excretion
AGD
Anogenital distance
BBP
Butyl benzyl phthalate
BMD
Benchmark dose
BMDL
Benchmark dose, lower confidence limit
BMR
Benchmark response
BW
Body weight
CASRN
Chemical abstracts service registry number
CD
Charles River Sprague-Dawley
CPSC
Consumer Product Safety Commission (U.S.)
DBP
Dibutyl phthalate
DC HP
Dicyclohexyl phthalate
DEHP
Di(2-ethylhexyl) phthalate
DIBP
Diisobutyl phthalate
DIDP
Diisodecyl phthalate
DINP
Diisononyl phthalate
ECB
European Chemicals Bureau
ECCCHC
Environment and Climate Change Canada Health Canada
ECHA
European Chemicals Agency
ECHC
Environment Canada Health Canada
EFSA
European Food Safety Authority
EPA
Environmental Protection Agency (U.S.)
F344
Fischer 344 rat
GD
Gestation Day
HEC
Human equivalent concentration
HED
Human equivalent dose
hCG
Human chorionic gonadotropin
INSL3
Insulin-like factor 3
LOAEL
Lowest-observed-adverse-effect level
LOEL
Lowest-observed-effect level
MBP
Monobutyl phthalate
MBP co-ox
Monobutyl phthalate co-ox
MBzP
Monobenzyl phthalate
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
NTP-CERHR
National Toxicology Program Center for the Evaluation of Risks to Human Reproduction
OCSPP
Office of Chemical Safety and Pollution Prevention
OEHHA
Office of Environmental Health Hazard Assessment
OHAT
Office of Health Assessment and Translation
OPPT
Office of Pollution Prevention and Toxics
PBPK
Physiologically based pharmacokinetic
PECO
Population, exposure, comparator, and outcome
PESS
Potentially exposed or susceptible subpopulations
Page 5 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
206
PND
Postnatal Day
207
POD
Point of departure
208
SACC
Science Advisory Committee on Chemicals
209
SD
Sprague-Dawley
210
TSCA
Toxic Substances Control Act
211
UF
Uncertainty factor
212
U.S.
United States
213
WOE
Weight of evidence
Page 6 of 122
-------
214
215
216
217
218
219
220
221
222
223
224
225
226
227
228
229
230
231
232
233
234
235
236
237
238
239
240
241
242
243
244
245
246
247
248
249
250
251
PUBLIC RELEASE DRAFT
DECEMBER 2024
ACKNOWLEGEMENTS
This report was developed by the United States Environmental Protection Agency (U.S. EPA or the
Agency), Office of Chemical Safety and Pollution Prevention (OCSPP), Office of Pollution Prevention
and Toxics (OPPT).
Acknowledgements
The Assessment Team gratefully acknowledges the participation, review, and input from EPA OPPT
and OSCPP senior managers and science advisors. The Agency is also grateful for assistance from the
following EPA contractors for the preparation of this draft technical support document: ICF (Contract
No. 68HERC23D0007); and SRC, Inc. (Contract No. 68HERH19D0022). Special acknowledgement is
given for the contributions of technical experts from EPA's Office of Research and Development (ORD)
including Justin Conley, Earl Gray, and Tammy Stoker.
As part of an intra-agency review, this technical support document was provided to multiple EPA
Program Offices for review. Comments were submitted by EPA's Office of General Counsel (OGC) and
ORD.
Docket
Supporting information can be found in the public docket, Docket ID EPA-HQ-QPPT-2018-0501.
Disclaimer
Reference herein to any specific commercial products, process, or service by trade name, trademark,
manufacturer, or otherwise does not constitute or imply its endorsement, recommendation, or favoring
by the United States Government.
Authors: Collin Beachum (Management Lead), Brandall Ingle-Carlson (Assessment Lead), Devin
Alewel (Human Health Hazard Assessment Lead), Anthony Luz (Human Health Hazard Discipline
Lead), Christelene Horton, Ashley Peppriell, (Human Health Hazard Assessors)
Contributors: Azah Abdallah Mohamed, 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,
Susanna Wegner
Technical Support: Kelley Stanfield, Hillary Hollinger, S. Xiah Kragie
This draft technical support document was reviewed and cleared for release by OPPT and OCSPP
leadership.
Page 7 of 122
-------
252
253
254
255
256
257
258
259
260
261
262
263
264
265
266
267
268
269
270
271
272
273
274
275
276
277
278
279
280
281
282
283
284
285
286
287
288
289
290
291
292
293
294
295
296
297
298
299
PUBLIC RELEASE DRAFT
DECEMBER 2024
SUMMARY
This technical support document is in support of the Toxic Substances Control Act (TSCA) Draft Risk
Evaluation for Butyl Benzyl Phthalate (BBP) (U.S. EPA. 2025f). This document describes the use of
reasonably available information to identify the non-cancer hazards associated with exposure to BBP
and the points of departure (PODs) to be used to estimate risks from BBP exposures in the draft risk
evaluation of BBP. Environmental Protection Agency (EPA, or the Agency) summarizes the cancer and
genotoxicity hazards associated with exposure to BBP in the Draft Cancer Raman Health Hazard
Assessment for Di(2-ethylhexyl) Phthalate (DEHP), Dibutyl Phthalate (DBP), Diisobiityl Phthalate
(DIBP), Butyl Benzyl Phthalate (BBP) and Dicyclohexyl Phthalate (DCHP) (U.S. EPA. 2025a).
EPA identified effects on the developing male reproductive system as the most sensitive and robust non-
cancer hazard associated with oral exposure to BBP 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 BBP by existing assessments of the U.S. EPA (2002a).
U.S. Consumer Product Safety Commission (CPSC) (2014. 2010). Health Canada (ECCC/HC. 2020;
EC/HC. 2015a; 2015; Environment Canada. 2000). European Chemical Agency (ECHA) (2017a. b,
2014. 2010. 2008). the Australian National Industrial Chemicals Notification and Assessment Scheme
(NICNAS) (2015). European Chemicals Bureau (ECB) (2007). European Food Safety Authority (EFSA)
(2019); California Office of Environmental Health Hazard Assessment (OEHHA) (1986). and the
National Academies of Sciences, Engineering, and Medicine (NASEM) (2017). EPA also considered
epidemiologic evidence qualitatively as part of hazard identification and characterization. However,
epidemiologic evidence for BBP was not considered further for dose response analysis due to limitations
and uncertainties in exposure characterization (discussed further in Sections 1.1 and 3.1.1). Use of
epidemiologic evidence qualitatively is consistent with phthalates assessment by Health Canada, U.S.
CPSC, NICNAS, ECHA, and NASEM.
As discussed further in Section 3.1.2, EPA identified 14 oral exposure studies (all of rats) that have
investigated developing male reproductive system effects of BBP following gestational and/or perinatal
exposure (Gray et al.. 2021; Spade et al.. 2018; Debartolo et al.. 2016; Schmitt et al.. 2016; Ahmad et
al.. 2014; Furr et al.. 2014; Howdeshell et al.. 2008; Aso et al.. 2005; Tyl et al.. 2004; Wilson et al..
2004; Ema et al.. 2003; Ema and Miyawaki. 2002; Gray et al.. 2000; Nagao et al.. 2000). Three studies
were multi-generation reproduction studies of BBP oral exposure (Aso et al.. 2005; Tyl et al.. 2004;
Nagao et al.. 2000). Across available studies, the most sensitive 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 POD of 50 mg/kg-day (human equivalent dose
[HED] of 12 mg/kg-day) based on phthalate syndrome-related effects on the developing male
reproductive system (organ-level outcomes such as decreased anogenital distance (AGD); decreased
fetal testicular testosterone; testicular histopathology) to estimate non-cancer risks from oral exposure to
BBP for acute, intermediate, and chronic durations of exposure in the draft risk evaluation of BBP. The
proposed POD was derived from 4 co-critical prenatal exposure studies of BBP that that support a no-
observed-adverse-effect level (NOAEL) of 50 mg/kg-day (Tyl et al.. 2004) and consensus lowest-
observable-adverse-effect level (LOAEL) of 100 mg/kg-day (Ahmad et al.. 2014; Furr et al.. 2014; Aso
et al.. 2005). Across the co-critical studies, 1 multi-generational study identified a NOAEL of 50 mg/kg-
day based on decreased anogenital distance (AGD) (Tyl et al.. 2004). and the other 3 studies support a
LOAEL of 100 mg/kg-day based upon decreased AGD, reduced ex vivo fetal testicular testosterone
production, and slight increases in testicular pathology (i.e., decreased epididymal and prostate weight,
decreased sperm count and motility, decreased epididymal germ cells, and testes softening) (Ahmad et
al.. 2014; Furr et al.. 2014; Aso et al.. 2005). The latter studies support the identified NOAEL.
Page 8 of 122
-------
300
301
302
303
304
305
306
307
308
309
310
311
312
313
314
315
316
317
318
319
320
321
322
323
324
325
326
327
328
329
330
PUBLIC RELEASE DRAFT
DECEMBER 2024
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 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 AGP and related phthalate syndrome effects
for use in characterizing risk from exposure to BBP for acute, intermediate, and chronic exposure
scenarios. The applicability and relevance of this POD for all exposure durations (acute, intermediate,
and chronic) is described in the introduction to Section 4 and additionally in 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 to assess risk for other age groups (e.g.,
older children, adult males, and the elderly) is considered to be conservative and appropriate for a
screening level assessment for these other age groups.
No data are available for the dermal or inhalation routes that are suitable for deriving route-specific
PODs. Therefore, EPA is using the acute/intermediate/chronic oral PODs to evaluate risks from dermal
and inhalation exposure to BBP. For the dermal route, differences in absorption are being accounted for
in dermal exposure estimates in the draft risk evaluation for BBP. For the inhalation route, EPA is
extrapolating the oral HED to an inhalation human equivalent concentration (HEC) per EPA's Methods
for derivation of inhalation reference concentrations and application of inhalation dosimetry (U.S.
EPA. 1994) using the updated human body weight and breathing rate relevant to continuous exposure of
an individual at rest provided in EPA's Exposure factors handbook: 2011 edition (U.S. EPA. 201 lb).
Table ES-1 and Section 6 summarize EPA's selection of the oral HED and inhalation HEC values used
to estimate non-cancer risk from acute/intermediate/chronic exposure to BBP in the draft risk evaluation
of BBP.
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 non-cancer risk characterization of BBP.
Page 9 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Table ES-1
. Non-cancer HED and HEC Used to
Estimate Risks
Exposure
Scenario
Target
Organ
System
Species
Duration
POD
(mg/kg-
day)
Effect
HEC"
(mg/m3
[ppm]
HED"
(mg/
kg-
day)
Benchmark
MOE*
References'
Acute,
Intermediate,
Chronic
Developing
male
reproductive
system
Rat
Multi-
generational
or 5-8 days
during
gestation
NOAEL
= 50
Phthalate
syndrome-
related effects
(e.g., J,AGD; j
fetal testicular
testosterone; j
reproductive
organ weights;
Leydig cell
effects; j
mRNA and/or
protein
expression of
steroidogenic
genes;
4INSL3)
64.2
[5.03]
12
UFa= 3
UFh=10
Total UF=30
(Alunad et
al.. 2014;
Furr et al..
2014; Aso
et al.. 2005;
Tvl et al..
2004)
Abbreviations: AGD = Anogenital distance; HEC = Human equivalent concentration; HED = Human equivalent dose;
INSL3: Insulin-like 3; MOE = Margin of exposure; NOAEL = No-observed-adverse-effect level; POD = Point of departure;
UF = Uncertainty factor
a HED and HEC values were calculated based on the most sensitive NOAEL of 50 mg/kg-day.
b EPA used allometric body weight scaling to the three-quarters power to derive the HED. Consistent with EPA Guidance
(U.S. EPA. 201 lc). the interspecies uncertainty factor (UF-,). 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.
c Tvl et al. (2004) suDDort a NOAEL of 50 me/ke-dav based on decreased AGD and decreased reproductive orsan weiehts in
a multi-generational study at 250 mg/kg-day (LOAEL); the remaining effects listed reached statistical significance at higher
doses (most of which are not considered adverse in isolation). Ahmad et al. (2014). Furr et al. (2014). and Aso et al. (2005)
reflect supporting phthalate syndrome-related effects (e.g., reduced ex vivo testicular testosterone production or testicular
histopathological changes) at LOAEL = 100 mg/kg-day.
332
Page 10 of 122
-------
333
334
335
336
337
338
339
340
341
342
343
344
345
346
347
348
349
350
351
352
353
354
355
356
357
358
359
360
361
362
363
364
365
366
367
368
369
370
371
372
373
374
375
376
377
378
379
PUBLIC RELEASE DRAFT
DECEMBER 2024
1 INTRODUCTION
In December 2019, EPA designated butyl benzyl phthalate (BBP) (CASRN 85-68-7) 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 part 702) (U.S. EPA.
2019). Following publication of the draft and final scope documents for BBP in 2020 (U.S. EPA. 2020a.
b), one of the next steps in the TSCA risk evaluation process is to identify and characterize the human
health hazards of BBP, conduct a dose-response assessment, and to determine toxicity values, such as a
point of departure (POD), to be used to estimate risks from BBP exposures. This technical support
document for BBP summarizes the non-cancer hazards associated with exposure to BBP and proposes
non-cancer toxicity values to be used to estimate risks from BBP exposures. Cancer human health
hazards associated with exposure to BBP are summarized in EPA's Draft Cancer Raman Health Hazard
Assessment for Di(l-ethylhexyl) Phthalate (DEHP), Dibutyl Phthalate (DBP), Diisobiityl Phthalate
(DIBP), Butyl Benzyl Phthalate (BBP) and Dicyclohexyl Phthalate (DCHP) (U.S. EPA. 2025a).
Over the past several decades, the human health effects of BBP have been reviewed by several
regulatory and authoritative agencies, including the: U.S. Consumer Product Safety Commission (U.S.
CPSC); 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); Australian National Industrial Chemicals Notification and Assessment
Scheme (NICNAS); and the California Office of Environmental Health Hazard Assessment (OEHHA).
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 BBP. Additionally, EPA
considered new literature published since the most recent existing assessments of BBP to determine if
additional data might support the identification of new human health hazards or lower PODs for use in
estimating human health risk. EPA's process for considering and incorporating new BBP literature is
described in the Draft Systematic Review Protocol for Butyl Benzyl Phthalate (BBP) (also referred to as
the Draft BBP Systematic Review Protocol) (U.S. EPA. 2025g). EPA's approach and methodology for
identifying and using human epidemiologic data and experimental laboratory animal data is described in
Sections 1.1 and 1.2.
1.1 Human Epidemiologic Data: Approach and Preliminary Conclusions
To identify and integrate human epidemiologic data into the draft BBP Risk Evaluation, EPA first
reviewed existing assessments of BBP conducted by regulatory and authoritative agencies, as well as
systematic reviews of epidemiological studies published by Radke et al. (2020b). Although the authors
(i.e., Radke et al.) are affiliated with the U.S. EPA's Center for Public Health and Environmental
Assessment, the reviews do not reflect EPA policy. Existing assessments of BBP identified by EPA are
listed below. As described further here and in Appendix A, most of these assessments have been
subjected to peer-review and/or public comment periods and 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 neurodevelopment, allergies, cardiovascular
function, oxidative stress, breast cancer, obesity, and metabolic disorders (Health Canada.
2018a);
Page 11 of 122
-------
380
381
382
383
384
385
386
387
388
389
390
391
392
393
394
395
396
397
398
399
400
401
402
403
404
405
406
407
408
409
410
411
412
413
414
415
416
417
418
419
420
421
422
423
424
425
426
427
PUBLIC RELEASE DRAFT
DECEMBER 2024
Application of systematic review methods in an overall strategy for evaluating low-dose toxicity
fi'om endocrine active chemicals (NASEM. 2017);
Phthalate exposure and male reproductive outcomes: A systematic review of the human
epidemiological evidence (Radke et al.. 2018);
Phthalate exposure andfemale reproductive and developmental outcomes: A systematic review
of the human epidemiological evidence (Radke et al.. 2019b);
Phthalate exposure and metabolic effects: A systematic review of the human epidemiological
evidence (Radke et al.. 2019a);
Phthalate exposure and neurodevelopment: A systematic review and meta-analysis of human
epidemiological evidence (Radke et al.. 2020a); and
Application of US EPA IRIS systematic review methods to the health effects of phthalates:
Lessons learned and path forward (Radke et al.. 2020b).
EPA relied on conclusions from Health Canada (2018a. b) and systematic review publications in the
open literature from authors affiliated with EPA's Center for Public Health and Environmental
Assessment ((Radke et al.. 2020b; Radke et al.. 2020a; Radke et al.. 2019b; Radke et al.. 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 BBP. PECO-relevant literature published since the most recent existing
assessment(s) of BBP was identified by applying a literature inclusion cutoff date from existing
assessments of BBP. For BBP, 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 BBP. 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 BBP metabolites and health outcomes. New PECO-relevant literature published between 2018 to
2019 was identified through the literature search conducted by EPA in 2019, as well as references
published between 2018 to 2023 that were submitted with public comments to the BBP docket (EPA-
HQ-OPPT-2018-0501). and these studies were evaluated for data quality and extracted consistent with
EPA's Draft Systematic Review Protocol Supporting TSCA Risk Evaluations for Chemical Substances
(U.S. EPA. 2021). Data quality evaluations for new studies reviewed by EPA are provided in the Draft
Data Quality Evaluation Information for Raman Health Hazard Epidemiology for Butyl Benzyl
Phthalate (BBP) (U.S. EPA. 2025d).
As described further in the Draft Systematic Review Protocol for Butyl Benzyl Phthalate (BBP) (U.S.
EPA. 2025g). EPA considers phthalate metabolite concentrations in urine to be an appropriate proxy of
exposure from all sourcesincluding exposure through ingestion, dermal absorption, and inhalation. As
described in the Application of US EPA IRIS systematic review methods to the health effects of
phthalates: Lessons learned and path forward (Radke et al.. 2020b). the "problem with measuring
phthalate metabolites in blood and other tissues is the potential for contamination from outside sources
(Calafat et al.. 2015). Phthalate diesters present from exogenous contamination can be metabolized to
the monoester metabolites by enzymes present in blood and other tissues, but not urine." Therefore, EPA
has focused its epidemiologic evaluation on urinary biomonitoring data; new epidemiologic studies that
Page 12 of 122
-------
428
429
430
431
432
433
434
435
436
437
438
439
440
441
442
443
444
445
446
447
448
449
450
451
452
453
454
455
456
PUBLIC RELEASE DRAFT
DECEMBER 2024
examined BBP metabolites in matrices other than urine were considered supplemental and not evaluated
for data quality.
The Agency used epidemiologic studies of BBP qualitatively. This approach is consistent with Health
Canada, U.S. CPSC, ECHANICNAS, and other agencies. EPA did not use epidemiology studies
quantitatively for dose-response assessment, primarily due to uncertainty associated with exposure
characterization. Primary sources of uncertainty include the source(s) of exposure; timing of exposure
assessment that may not be reflective of exposure during outcome measurements; and use of spot-urine
samples, which due to rapid elimination kinetics may not be representative of average urinary
concentrations that are collected over a longer term or calculated using pooled samples. The majority of
epidemiological studies introduced additional uncertainty by not considering BBP in isolation and
failing to account for confounding effects from co-exposure to mixtures of multiple phthalates (Shin et
al.. 2019; Aylward et al.. 2016). Conclusions from Health Canada (2018a. b) and systematic review
articles (Radke et al.. 2020a; Radke et al.. 2019b; Radke et al.. 2019a; Radke et al.. 2018) regarding the
level of evidence for association between urinary BBP metabolites and each health outcome were
reviewed by EPA and used as a starting point for its human health hazard assessment. The Agency also
evaluated and summarized new epidemiologic studies identified by EPA's systematic review process to
use qualitatively during evidence integration to inform hazard identification and the weight of scientific
evidence (Shin et al.. 2019; Aylward et al.. 2016) (Section 3.1.1).
1.2 Laboratory Animal Findings: Summary of Existing Assessments,
Approach, and Methodology
1.2.1 Existing Assessments of BBP
The human health hazards of BBP have been evaluated in existing assessments by U.S. EPA (2002a).
U.S. CPSC (2014. 2010). Health Canada (ECCC/HC. 2020; EC/HC. 2015a: 2015; Environment Canada.
2000). ECHA (2017a. b, 2014. 2010. 2008). ECB (2007). EFSA (2019); OEHHA (1986). NTP (2003).
NICNAS (2016. 2015). and NASEM (2017). These assessments have consistently identified toxicity to
the developing male reproductive system as the most sensitive outcomes for use in estimating human
risk from exposure to BBP. The PODs from these assessments are shown in Table 1-1.
Page 13 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Table 1-1. Summary of BBP Non-cancer POPs Selected for Use by Other Regulatory Organizations
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Critical Effect (s)
O
6 dams/group) were
exposed to 0, 4, 20, or 100 mg/kg-day BBP via
oral gavage from GD 14-21. Dams were allowed
to give birth naturally, and male offspring were
sacrificed on PND 5, 25, or 75. (Alunad et al
2014)
20/100
I Serum testosterone (PND 75),
i Epididymal and prostate
weights (PND 75), and j Sperm
count/motility (PND 75)
y/c
Page 14 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Critical Effect (s)
o
-------
459
460
461
462
463
464
465
466
467
468
469
470
471
472
473
474
475
476
477
478
479
480
481
482
483
484
485
486
487
488
489
490
491
492
493
494
495
496
497
498
499
500
501
502
503
PUBLIC RELEASE DRAFT
DECEMBER 2024
1.2.2 Approach to Identifying and Integrating Laboratory Animal Data
Table 1-1 provides an overview of EPA's approach to identifying and integrating laboratory animal data
into the draft BBP Risk Evaluation. EPA first reviewed existing assessments of BBP conducted by
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 BBP, and identify key studies used to establish PODs for estimating human risk. Existing
assessments reviewed by EPA are listed below. As described further in Appendix A, most of these
assessments have been subjected to external peer-review and/or public comment periods.
Provisional Peer Reviewed Toxicity Values for butyl benzyl phthalate (U.S. EPA. 2002a);
Chronic Hazard Advisory Panel on phthalate s and phthalate alternatives (U.S. CPSC. 2014);
Toxicity review for benzyl-n-butyl phthalate (U.S. CPSC. 2010);
Supporting documentation: Carcinogenicity ofphthalate s - mode of action and human relevance
(Health Canada. 2015);
Canadian environmental protection act priority substances list assessment report:
Butylbenzylphthalate (Environment Canada. 2000);
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. 2015a);
Screening assessment - Phthalate substance grouping (ECCC/HC. 2020);
Substance name: Benzyl butyl phthalate, EC number: 201-622-7, CAS number: 85-68-7:
Member state committee support documentation for identification of benzyl butyl phthalate
(BBP) as a substance of very high concern (ECHA. 2008);
Evaluation of new scientific evidence concerning the restriction contained in Annex XVII to
regulation (EC) no. 1907 2006 (REACH): Review of new available information for benzyl butyl
phthalate (BBP) CAS no. 85-68-7 Einecs no. 201-622-7 (ECHA. 2010);
Support document to the opinion of the member state committee for identification of benzyl butyl
phthalate (bbp) as a substance of very high concern because of its endocrine disrupting
properties which cause probable serious effects to human health and the environment which give
rise to an equivalent level of concern to those of cmrl and phi vpvb2 substances (ECHA. 2014);
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 on an Annex XV dossier proposing restrictions on four phthalates (DEHP, BBP, DBP,
DIBP) (ECHA. 2017b);
European union risk assessment report: Benzyl butyl phthalate (BBP) (ECB. 2007);
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);
Safe Drinking Water and Toxic Enforcement Act of 1986 Proposition 65. Initial Statement of
Reasons. Title 27, California Code of Regulations. Proposed amendment to Section 25805(b),
Specific Regulatory Levels: Chemicals Causing Reproductive Toxicity. Butyl benzyl phthalate
(oral exposure) (OEHHA. 1986);
NTP-CERHR monogi'aph on the potential human reproductive and developmental effects of
butyl benzyl phthalate (BBP) (NTP-CERHR. 2003);
Priority existing chemical assessment report no. 40: Butyl benzyl phthalate (NICNAS. 2015);
C4-6 side chain transitionalphthalates: Human health tier II assessment (NICNAS. 2016); and
Page 16 of 122
-------
504
505
506
507
508
509
510
511
512
513
514
515
516
517
518
519
520
521
522
523
524
525
526
527
528
529
530
531
532
533
534
535
536
537
538
539
540
541
PUBLIC RELEASE DRAFT
DECEMBER 2024
Application of systematic review methods in an overall strategy for evaluating low-dose toxicity
fi'om endocrine active chemicals (NASEM. 2017).
Figure 1-1. Overview of BBP Human Health Hazard Assessment Approach
Abbreviations: HED = Human equivalent dose; LOAEL = Lowest-observed-adverse-effect level; LOEL =
Lowest-observed-effect level; PECO = Population, exposure, comparator, and outcome; POD = Point of
departure.
11 Any study that was considered for dose-response assessment, not necessarily limited to the study used for POD
selection.
b 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. 2015a)
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 Systematic Review Protocol for Butyl Benzyl Phthalate (U.S. EPA. 2025g). 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, subchronic, chronic) and routes of exposure
(i.e., oral, dermal, inhalation), as shown in Table 1-1 (literature evaluated from 2014-2019 is described
below). EPA first screened titles and abstracts and then full texts for relevancy using PECO screening
criteria described in the Draft Systematic Review Protocol for Butyl Benzyl Phthalate (U.S. EPA.
2025g).
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 Systematic Review
Protocol for Butyl Benzyl Phthalate (U.S. EPA. 2025 g). 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) (Table 1-1).
New information for BBP 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
Page 17 of 122
-------
542
543
544
545
546
547
548
549
550
551
552
553
554
555
556
557
558
559
560
561
562
563
564
565
566
567
568
569
570
571
572
573
574
575
576
577
578
579
580
581
582
583
584
585
586
587
PUBLIC RELEASE DRAFT
DECEMBER 2024
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. 2011c). 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) 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 BBP.
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 integrated into
the hazard identification process.
Additionally, because effects on the developing male reproductive system are a focus of EPA's BBP
hazard assessment, 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). EPA identified one additional more recent PECO-relevant
study that provided information pertaining to one primary hazard outcomes (i.e.,
reproductive/developmental toxicity) (Gray et al.. 2021).
Data quality evaluations for BBP animal toxicity studies reviewed by EPA are provided in the Draft
Data Quality Evaluation Information for Human Health Hazard Animal Toxicology for Butyl Benzyl
Phthalate (BBP) (U.S. EPA. 2025c).
1.2.3 New Literature Identified and Hazards of Focus for BBP
As described in Section 1.2.2, and as described further in the Draft Systematic Review Protocol for Butyl
Benzyl Phthalate (U.S. EPA. 2025 g). 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
Systematic Review Protocol for Butyl Benzyl Phthalate (U.S. EPA. 2025g). EPA identified 10 new
PECO-relevant animal toxicology studies that provided information pertaining to various primary
hazard outcomes, including: reproductive/developmental, neurotoxicity, immune adjuvant effects, renal,
and hepatic outcomes. Further details regarding EPA's handling of new information provided in these
10 studies are provided below.
Reproductive/Developmental. EPA identified 7 new studies evaluating reproductive/
developmental outcomes (Gray et al.. 2021; Integrated Laboratory Systems. 2017; Debartolo et
al.. 2016; Schmitt et al.. 2016; Ahmad et al.. 2015; Alam and Kurohmaru. 2015; Ahmad et al..
2014). These new studies of BBP are discussed further in Section 3.1. Of these, only 4 studies
(Gray et al.. 2021; Debartolo et al.. 2016; Schmitt et al.. 2016; Ahmad et al.. 2014) evaluated
endpoints relevant to phthalate syndrome outcomes from developing male exposure (i.e.,
histopathology and/or organ weights of the male reproductive system, anogenital distance). The
other 3 studies evaluated a range of endpoints including changes in the estrus cycle or serum
estradiol, progesterone, follicle stimulating hormone, luteinizing hormone, number of ovarian
follicles, reproductive organ weights (i.e., ovary and/or uterus), pup body weights, or used a non-
Page 18 of 122
-------
588
589
590
591
592
593
594
595
596
597
598
599
600
601
602
603
604
605
606
607
608
609
610
611
612
PUBLIC RELEASE DRAFT
DECEMBER 2024
developmental exposure design (Integrated Laboratory Systems. 2017; Ahmad et al.. 2015; Alam
and Kurohmaru. 2015).
Neurotoxicity. EPA identified 3 new studies evaluating neurological effects following BBP
exposure (Debartolo et al.. 2016; Schmitt et al.. 2016; Min et al.. 2014).
Immune adjuvant effects. EPA identified 1 new study evaluating immunological effects
following BBP exposure (Jahreis et al.. 2018).
Renal EPA identified 2 new studies evaluating renal effects following BBP exposure
(Integrated Laboratory Systems. 2017; Nakagomi et al.. 2017).
Hepatic. EPA identified 1 new study evaluating hepatic effects following BBP exposure
(Nakagomi et al.. 2017).
The most sensitive and robust PODs selected from existing hazard assessments of BBP have been based
on effects on the developing male reproductive system (ECCC/HC. 2020; EFSA. 2019; ECHA. 2017a;
NICNAS. 2015; U.S. CPSC. 2014). Existing assessments have consistently shown that effects on other
health outcomes (i.e., female reproduction, neurological, hepatic/renal, immune, and 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 NOAELs/LOAELs were identified for male 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). New literature relevant to developing male reproductive toxicity presenting phthalate
syndrome-related effects were considered in non-cancer hazard assessment and are further discussed in
Section 3.1.2. All other new studies, as well as brief justification for their exclusion from further
evaluation of dose-response and derivation of a POD for use in human health risk assessment, are
discussed in Appendix B.
Page 19 of 122
-------
613
614
615
616
617
618
619
620
621
622
623
624
625
626
627
628
629
630
631
632
633
634
635
636
637
638
639
640
641
642
643
644
645
646
647
648
649
650
651
652
653
654
655
656
657
658
659
PUBLIC RELEASE DRAFT
DECEMBER 2024
2 TOXICOKINETICS
2.1 Oral Route
EPA identified two animal studies available on the metabolism of BBP following oral exposure
(Nativelle et al.. 1999; Eigenberg et al.. 1986). as well as one human oral exposure study (Anderson et
al.. 2001) and three human biomonitoring assessments identifying BBP metabolites in urine and feces
(Apel et al.. 2020; Frederiksen et al.. 2011; Stahlhut et al.. 2007). Based upon few experimental
absorption, distribution, metabolism, and excretion (ADME) toxicokinetic assessments, orally
administered BBP is readily absorbed through the gastrointestinal tract and mainly processed via first
pass metabolism by intestinal and hepatic esterases (Anderson et al.. 2001; Nativelle et al.. 1999;
Eigenberg et al.. 1986). Following oral absorption in rats and humans, BBP is hydrolyzed into the
diester monobutyl phthalate (MBP), followed by an appreciable amount of monobenzyl phthalate
(MBzP) metabolite formation (Anderson et al.. 2001; Nativelle et al.. 1999; Eigenberg et al.. 1986).
However, identification of BBP metabolites in female Wistar rat urine has shown hippuric acid,
generated from further hydrolyzation of MBP and MBzP to increase solubility for excretion, as the main
recovered metabolite (Nativelle et al.. 1999). Additional oxidized metabolites detected in urine,
including phthalic acid, which has been observed in rats and humans (Anderson et al.. 2001; Nativelle et
al.. 1999). as well as benzoic acid and monobutyl phthalate co-ox (MBP co-ox), are also observed in
small quantities (Nativelle et al.. 1999). A summary of BBP metabolite formation pathway of 6 different
metabolites identified in rat and human samples (primarily urine and feces) after oral administration of
BBP is presented in Table 2-1.
In an oral metabolism study, female Wistar rats were gavaged with 150, 475, 780, and 1500 mg/kg-day
BBP for 3 consecutive days, followed by molecular characterization of urinary metabolites at 24, 48,
and 72 hours post exposure (Nativelle et al.. 1999). Here, Nativelle et al. (1999) identified 6 metabolites
in urine collections, where the parent compound was not recovered. Hippuric acid represented the major
recovered metabolite (51-56%), followed by MBP (29-34%), MBzP (7-12%), and small percentages
(less than 3%) of MBP co-ox and terminal acid hydrolysis products. It should also be noted that the study
by Nativelle et al. (1999) observed a dose-dependent impact on metabolite excretion rate. Metabolite
recovery over the three days at 150 to 1500 mg/kg-day showed a dose-dependent effect on metabolite
quantity. Daily elimination analysis following administration of 475 mg/kg-day BBP resulted in a
steady-state excretion rate over 72 hours post exposure, but 1500 mg/kg-day exposure resulted in
increased relative MBP, MBzP, and hippuric acid metabolite levels at 72 hours. This toxicokinetic
response indicates a time-dependent dose effect, where gastrointestinal absorption may be saturated and
may shift to fecal elimination at excessive levels (Eigenberg et al.. 1986). In sum, the proposed BBP
degradation pathway based upon oral exposure data obtained from the Nativelle et al. (1999) study is
shown in Figure 2-1.
Eigenberg et al. (1986) performed a single oral gavage exposure using radiolabeled BBP (14C-BBP) at 2,
20, 200, and 2000 mg/kg in male F344 rats and made examinations at 24 and 96 hours post exposure.
Here, 75 to 86 percent of the total dose was excreted within 24 hours in urine and feces, and 92 percent
was recovered by the 96 hours post collection. For groups receiving 2 to 200 mg/kg, 75 percent of the
dose was eliminated in urine vs. 20 percent in feces. However, in the high dose group of 2000 mg/kg,
the predominant excretion route was feces (72%) and not urine (22%). In this same study, investigators
also administered intravenous infusion of 14C-BBP (20 mg/kg) through the tail vein to assess tissue
distribution and toxicokinetic properties. In this kinetics assessment, blood BBP monoester metabolite
levels peaked within 5 minutes of BBP administration. To determine the extent of biliary excretion, the
bile ducts of rats were cannulated, and bile was collected over the course of 4 hours at regular time
Page 20 of 122
-------
660
661
662
663
664
665
666
667
668
669
670
671
672
673
674
675
676
677
678
679
680
681
682
683
684
685
686
687
688
689
690
691
692
693
694
695
696
697
698
699
PUBLIC RELEASE DRAFT
DECEMBER 2024
intervals after dosing. After 4 hours, 55 percent of the dose was excreted in bile, whereas 34 percent was
excreted in the urine, and larger quantities of BBP metabolites were found in the bile compared to the
urine. Altogether, these data demonstrate that biliary excretion is the predominant route of excretion,
and that reabsorption occurs (via enterohepatic circulation). In addition to biliary and enterohepatic
recirculation for excretion, BBP metabolites rapidly distributed into multiple tissues, including brain,
lung, liver, kidney, spleen, testes, small intestine, muscle (thigh), skin (abdominal), and adipose. Aside
from urinary and fecal excretion levels, peak distribution levels (measured at 30 minutes) were observed
in the small intestine, muscle, and skin. Half-lives of parent compound and monoester metabolites was
approximately 6 hours across all tissues examined, with 84 percent of the total dose cleared within 24
hours of administration.
One controlled human BBP oral exposure study was identified (Anderson et al.. 2001). In this study,
participants (n = 13 volunteers, age and sex not specified) were orally exposed (ingestion) to the
deuterated form of BBP (d4-BBP) at a single low (253 jag) and high (506 jag) dose, followed by urinary
metabolite measures 24 hours after dosing and at 2 and 6 days post exposure. MBzP was the
predominant urinary metabolite (67% at low dose and 78% of the excretion fraction at high dose) in
humans following oral d4-BBP exposure when measured in urine 24 hours after dosing. MBP was
identified as a minor urinary metabolite in humans, accounting for 6 percent of the excretion fraction at
the high dose but was undetectable in the low dose group. No labeled phthalate monoester levels were
found in urine when measured at 2 or 6 days following exposure, suggesting rapid uptake and excretion
occurring within the first 24 hours. It should be noted participant sex was not reported in the Anderson
et al. (2001) assessment, which may impact toxicokinetic assumptions, albeit variability of inter-
individual excretion fractions was determined to be acceptable. Nevertheless, MBzP as such a major
excretion fraction suggests this metabolite as a dominant biomarker of human exposure. Human
biomonitoring assessments of multiple phthalates, including BBP, have consistently identified MBzP as
the predominant metabolite (along with lesser amounts of MBP) in urine collections of multiple human
sampling collections (Apel et al.. 2020; Frederiksen et al.. 2011; Stahlhut et al.. 2007).
The available rodent studies (Nativelle et al.. 1999; Eigenberg et al.. 1986) established elimination
profiles shifting from urine to feces at high doses of BBP (1500 to 2000 mg/kg-day), which indicates
that oral absorption of BBP is saturable in rodents at high doses. However, elimination rates in urine
were fairly constant across doses less than 200 mg/kg-day in rodents and in a low single dose human
study (Anderson et al.. 2001). which is within the range considered for dose response in this assessment.
Given that approximately 75 percent BBP is excreted at lower doses in both rats (Eigenberg et al.. 1986)
and humans (Anderson et al.. 2001) at 24 hours, no adjustment is needed to account for oral absorption
between species. Therefore, based on the available data, and in accordance with prior agency
assessments, EPA will assume an oral absorption 100 percent for the draft risk evaluation of BBP.
Table 2-1. Metabolites of BBP Identified in Urine from Rats and Humans after Oral
Administration
Urinary Metabolite
Abbreviation
Rat
Human"
Reference(s) (Species)
Monobutyl phthalate
MBP
V
V
(Eieenbere et al.. 1986) (rat)
(Nativelle et al.. 1999) (rat)
(Anderson et al.. 2001) (human)
(Frederiksen et al.. 2011)* (human)
(Ariel et al.. 2020) (human)
(Stahlhut et al.. 2007) (human)
Page 21 of 122
-------
PUBLIC RELEASE DRAFT
DECE]
VIBER 2024
Urinary Metabolite
Abbreviation
Rat
Human"
Reference(s) (Species)
Monobenzyl
phthalate
MBzP
(Eisenbers et al.. 1986) (rat)
(Nativelle et al.. 1999) (rat)
(Anderson et al.. 2001) (human)
(Frederiksen et al.. 2011) (human)
(Ariel et al.. 2020) (human)
(Stahlhut et al.. 2007) (human)
Monobutyl phthalate
ffl-OX
MBP ©-ox
ND
(Nativelle et al.. 1999) (rat)
Hippuric acid
-
ND
(Nativelle et al.. 1999) (rat)
Phthalic acid
-
ND
(Nativelle et al.. 1999) (rat)
Benzoic acid
-
ND
(Nativelle et al.. 1999) (rat)
Abbreviations: ND = no data available
"Metabolites detected as Dart of human controlled experimental (Anderson et al.. 2001) or biomonitorina/DODiilation data
assessment studies (Ariel et al.. 2020; Frederiksen et al.. 2011; Stahlhut et al.. 2007). Althoushbiomonitorins studies do
not distinguish between routes or pathways of exposure, urinary metabolites are shown for comparison to urinary
metabolites detected in rodent models.
b Urinary MBP detection was reported as the sum of MBP and mono-iso-butyl phthalate isofonns due to chromatographic
characterization limitations.
700
Page 22 of 122
-------
701
702
703
704
705
706
707
708
709
710
711
712
713
714
715
716
717
718
719
a:
MBuP w-ox (n°5)
o
li 4
C - O- CH,- CH,- CH,-
C-OH
II
o
f
I
PUBLIC RELEASE DRAFT
DECEMBER 2024
BBP
o
II
C-O- at;CHjCH;CH,
c-o aiji
o I
ijr
1
MBuP (n°3)
MBeP (n°6)
o
O CH; CHj CHj cn3
OH
Of°" ^
o '
~
HO CEI2 CI
1 J
[j
CH3 OH
I
phthalic acid (n°2)
benzoic acid (n°l)
~
ii
C" OH
C~ OK
II
o
phthalic acid (n°2)
CHj CH2 CH3 Cs
o
II
- C-OH
hippuric acid (n°4)
a
O
o
II *
C-NHCH2 Cs
O
OH
Figure 2-1. Proposed Metabolic Pathway of BBP Following Oral Exposure (Figure from Health
Canada (EC/HC. 2015b))
Notes: Metabolic pathway is based upon data collected from oral administration of BBP in female Wistar rats.
Original pathway is taken from Native lie et al. (1999) and is also found in Health Canada (EC/HC. 2015b)
report (Figure H-3). MBuP = Monobutyl phthalate (MBP); MBeP = Monobenzyl phthalate (MBzP); MBuP co-ox
= Oxidized monobutyl phthalate.
2.2 Inhalation Route
No controlled human exposure studies or in vivo animal studies are available that evaluate the ADME
properties of BBP for the inhalation route. As discussed further in Sections 3 and 6, no data from
experimental animal models are available for the inhalation route that are suitable for deriving a route-
specific PODs. Therefore, EPA extrapolated the inhalation POD from the oral POD. For this draft risk
evaluation, EPA assumed similar absorption for the oral and inhalation routes (100% absorption),
as done in previous assessments (NICNAS. 2015). and no adjustment was made when extrapolating to
the inhalation route of exposure.
2.3 Dermal Route
EPA identified two in vivo/ex vivo rodent studies (Sugino et al.. 2017; Elsisi et al.. 1989) and three ex
vivo human studies evaluating ADME properties following dermal application of BBP (Sugino et al..
2017; DuPont. 2006a. b).
Page 23 of 122
-------
720
721
722
723
724
725
726
727
728
729
730
731
732
733
734
735
736
737
738
739
740
741
742
743
744
745
746
747
748
749
750
751
752
753
754
755
756
757
758
759
760
761
762
763
764
765
766
767
768
PUBLIC RELEASE DRAFT
DECEMBER 2024
In the report by Elsisi et al. (1989). ADME properties of eight phthalates, including BBP, were analyzed
through dermal application of radiolabeled parent compounds in male F344 rats. 14C-BBP (5 to 8
mg/cm2) was applied to shaved dorsal area skin (1.3 cm diameter application area) and covered with a
plastic cap, and urine and feces were collected every 24 hours for the seven days. After 7 days, animals
were sacrificed, and levels of 14C-BBP were determined in organs. After 7 days, 86 percent of the
applied dose was recovered, including approximately 30 percent of 14C-BBP in urine and feces, 44.9
percent in skin at the site of application, 6.3 percent in the plastic cap, 4.6 percent in muscle, 0.17
percent in adipose tissue, 0.08 percent in skin, and less than 0.5 percent in other tissues (i.e., brain, lung,
liver, spleen, small intestine, kidney, testis, spinal cord and blood). These results indicate that
approximately 35 percent of the applied dose was absorbed over 7 days. However, combined urinary
and fecal excretion was linear over 7 days, indicating approximately 5 percent adsorption per day for 7
days. Relative to other phthalates tested, BBP had a linear and intermediate excretion rate, with slower
absorption and excretion likely being due to its higher molecular weight, as other medium-chain
phthalates with a low molecular weight, such as dibutyl phthalate, showed rapid excretion. Elsisi et al.
(1989) also observed low levels of distribution in muscle (4.6%), adipose (0.17%), and small amounts
across brain, lung, liver, spleen, small intestine, kidney, testis, spinal cord, and blood (summation of less
than 0.5%). Thus, in addition to biliary excretion and enterohepatic recirculation, BBP metabolites may
distribute into multiple non-circulatory or -hepatic compartments following dermal exposure. Oral
exposure studies have noted relatively short BBP metabolite half-lives in both rats and humans
(Anderson et al.. 2001; Nativelle et al.. 1999; Eigenberg et al.. 1986). Because of the short metabolic
half-life, it is assumed that BBP metabolites do not accumulate in tissues.
Sugino et al. (2017) used ex vivo epidermal membranes (0.95 cm2) prepared from abdominal excisions
of male hairless rats (WBN/Ila-Ht) and human females to assess skin permeation properties of multiple
phthalates, including BBP. Application of BBP to skin showed species-specific metabolite permeation
outcomes, but no diffusion of the parent compound. In sections prepared from hairless rats, only the
monoester metabolites MBP and MBzP diffused across dermal membranes, with more MBP metabolites
relative to MBzP. Conversely, MBzP was the dominant metabolite recovered in human skin, and the
permeability coefficient was markedly lower in human skin relative to the rat. An additional important
finding of this ex vivo assessment is that there also appeared to be metabolism-dependent processes
impacting dermal uptake. Sugino et al. (2017) applied diisopropyl fluorophosphates (DFP), as serine-
esterase inhibitor, to additional rat skin treatment groups, and noted both a shift toward MBzP
metabolite production and impermeability of BBP metabolites following DFP application.
Lastly, Dupont et al. (2006a. b) conducted two independent assessments using ex vivo human female
cadaver abdominal skin sections (n = 3, 2 replicates from each donor). The first experiment utilized skin
collections (468 to 487 |im thick) and exposed 0.64 cm2 size sections to an infinite dermal load of 100
|iL/cm2 BBP for 8 hours, which was spiked with 14C-BBP (for recovery estimate) into a non-
radiolabeled formulation that was uniformly mixed. Recovery of the non-absorbed applied dose at the
end of 8 hours was 96.4 percent, with a total estimated absorbed dose of 0.197 percent (165 |ig BBP)
(DuPont. 2006b). The second experiment by Dupont et al. (2006a) exposed 0.64 cm2 abdominal skin
sections (248 to 470 |im thick) to a BBP film matrix containing 14C-BBP occluded parafilm directly
placed on the skin for 8 hours. In this case, the total estimated exposure was 5958 |ig BBP, and the total
absorbed dose at the end of 8 hours was less than .01 percent (0.57 |ig BBP).
Although human evidence is limited, multiple regulatory agencies assume that BBP dermal absorption is
low, and that dermal migration is reportedly lower in human compared to rat skin for phthalates,
including BBP (Sugino et al.. 2017; Scott et al.. 1987). This assumption, along with the lack of data and
Page 24 of 122
-------
769
770
111
772
773
774
775
776
777
778
779
780
781
782
783
784
785
786
787
788
789
790
791
792
793
794
795
796
797
798
799
800
801
802
803
804
805
806
807
808
809
810
811
812
813
814
815
PUBLIC RELEASE DRAFT
DECEMBER 2024
uncertainty in available studies, has led several agency assessments to adopt a worst-case dermal
bioavailability of 5 percent in humans (NICNAS. 2015; ECB. 2007).
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 Draft
Environmental Release and Occupational Exposure Assessment for Butyl benzyl phthalate (BBP) (U.S.
EPA. 2025e) and Draft Consumer and Indoor Dust Exposure Assessment for Butyl benzyl phthalate
(BBP) (U.S. EPA. 2025b). Briefly, EPA is proposing to use BBP dermal absorption data from the
Dupont (2006b) study to estimate flux-limited dermal absorptions from liquid formulations of BBP, and
Dupont (2006a) to estimate the flux-limited dermal absorption of BBP in solids. Using the Dupont
(2006b) study estimate of 0.165 mg on a 0.64 cm2 area of BBP (0.258 mg/cm2) over an 8 hour period,
the steady-state flux of neat BBP is estimated as 3.22 x 10 2 mg/cm2/hr. Using the Dupont (2006a) study
estimate of 0.00057 mg over a 0.64 cm2 area off BBP (0.0008906 mg/cm2 of BBP) over an 8 hour
period, the steady-state flux of neat BBP is estimated as 1.113 x 10-4 mg/cm2/hr. EPA estimated the
steady-state flux and assumed is equal to the average flux.
2.4 Summary
The majority of ADME information on BBP are obtained from rodent oral exposure studies. Following
oral exposure, BBP rapidly undergoes esterase hydrolysis into MBP and MBzP, and then subsequent
metabolism to the predominant urinary metabolite of hippuric acid (in rats), however, other minor
urinary metabolites have been detected, including glucuronidated MBP and/or MBzP metabolites.
Whereas hippuric acid is the major BBP urinary metabolite in rats, MBzP is the predominant metabolite
detected in human urine following oral exposure. BBP is rapidly absorbed by the gastrointestinal tract
and generally undergo hepatic metabolism and biliary excretion, along with some distribution
throughout many bodily compartments. Reasonably available data suggest most of the administered
dose of BBP is excreted through urine within 24 hours, albeit at excessively high doses, there is major
fecal elimination potentially associated with saturated oral absorption.
ADME data on non-oral routes of exposure remains limited, with no quantitative inhalation studies
currently available. The few in vivo/ex vivo studies for dermal BBP exposure suggest a much lower
dermal absorption rate compared to ingestion. However, there remain uncertainties in the data available
on the toxicokinetics of BBP, particularly pertaining to inter-species and inter-individual factors and
lack of comprehensive experimental data. In the Sugino et al. (2017) study using ex vivo epidermal
membranes from rats and humans, dermal absorption and metabolite formation was reportedly impacted
by serine esterase inhibitors, suggesting dermal rate-limiting enzymatic activity may be a consideration
in inter-species differences or a source of uncertainty.
Although the exposure routes or conditions cannot be specified, human biomonitoring assessments have
noted, in addition to urine metabolite detection, the presence of phthalate metabolites (including MBP
and MBzP) in fetal serum, breast milk, and semen (Main et al.. 2006; Lashlev et al.. 2004; Rozati et al..
2002). Given these findings, along with the observation of BBP metabolites in multiple tissues following
dermal exposure (Elsisi et al.. 1989) and intravenous infusion (Eigenberg et al.. 1986). and comparative
analysis of phthalate kinetics in prior assessments (NICNAS. 2008). BBP metabolites are assumed to
widely distribute after exposure, even being able to cross the placental barrier. However, metabolites
appear to have short half-lives and there is no evidence for tissue accumulation.
Given the toxicokinetic information available for BBP, EPA will assume an oral absorption of 100
percent and an inhalation absorption of 100 percent for the draft risk evaluation. For dermal absorption,
Page 25 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
816 EPA is using a flux-limited absorption rate as described further in the Draft Consumer and Indoor Dust
817 Exposure Assessment for Butyl benzyl phthalate (BBP) (U.S. EPA. 2025b) and Draft Environmental
818 Release and Occupational Exposure Assessment for Butyl benzyl phthalate (BBP) (U.S. EPA. 2025e).
Page 26 of 122
-------
819
820
821
822
823
824
825
826
827
828
829
830
831
832
833
834
835
836
837
838
839
840
841
842
843
844
845
846
847
848
849
850
PUBLIC RELEASE DRAFT
DECEMBER 2024
3 NON-CANCER HAZARD IDENTIFICATION
As discussed in Section 1.2, the effects on the developing male reproductive system has consistently
been identified in existing assessments of BBP as the most sensitive effects associated with oral
exposure to BBP in experimental animal models (ECCC/HC. 2020; EFSA. 2019; ECHA. 2017a. b;
NASEM. 2017; NICNAS. 2016; EC/HC. 2015a; Health Canada. 2015; NICNAS. 2015; U.S. CPSC.
2014; ECHA. 2010; U.S. CPSC. 2010; ECHA. 2008; ECB. 2007; NTP-CERHR. 2003; U.S. EPA.
2002a; Environment Canada. 2000; OEHHA. 1986). EPA identified no new information through
systematic review that would change this conclusion. Therefore, EPA focused its non-cancer hazard
characterization on developing male reproductive toxicity, which is discussed in the sections below.
New literature on non-cancer hazards identified by EPA in studies published between 2014 to 2019, but
not used for POD derivation, are briefly presented in Section 3.1.2.2 and Appendix B.
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 BBP metabolites and male and female 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 BBP by Health Canada
(2018b). Radke et al. (2019b; 2018). and NASEM (2017). respectively, including conclusions related to
exposure to BBP and health outcomes. Additionally, EPA also evaluated epidemiologic studies
published after the Health Canada (2018b) assessment as part of its literature search (i.e., published
between 2018 and 2019) to determine if newer epidemiologic studies would change the conclusions of
existing epidemiologic assessments or provide useful information for evaluating exposure-response
relationship (Section 3.1.1.2). Overall, EPA considered there to be limitations in the epidemiological
evidence for association between urinary metabolites of BBP and the developing male reproductive
system. This stems from uncertainty associated with exposure characterization of individual phthalates,
including source or exposure and timing of exposure as well as co-exposure confounding with other
phthalates. Therefore, EPA considered epidemiologic studies of BBP qualitatively.
Page 27 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
851 Table 3-1. Summary of Scope and Methods Used in Previous Assessments to Evaluate the
852 Association Between BBP Exposure and Male Reproductive Outcomes
Previous Assessment
Outcomes Evaluated
Method Used for Study Quality
Evaluation
Health Canada (2018b)
Hormonal effects:
Sex hormone levels (e.g.. testosterone)
Growth & Development:
AGD
Birth measures
Male infant genitalia (e.g..
hypospadias/cryptorchidism)
Placental development and gene expression
Preterm birth and gestational age
Postnatal growth
DNA methylation
Reproductive:
Altered male puberty
Gynecomastia
Changes in semen parameters
Sexual dysfunction (males)
Sex ratio
Downs and Black (Downs and
Black. 1998)
Radke et al. (2018)
AGD
Hypospadias/cryptorchidism
Pubertal development
Semen parameters
Time to pregnancy (male exposure)
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
ROB INS-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; GRADE = Grading of Recommendations, Assessment, Development and
Evaluation; OHAT = National Toxicology Program's Office of Health Assessment and Translation; ROBINS-I= Risk
of Bias in Non-randomized Studies of Interventions.
853 3.1.1.1.1 Health Canada (2018b)
854 Health Canada (2018b) considered 83 studies that evaluated the association between BBP and its
855 metabolite (MBzP) and reproductive outcomes such as altered male puberty, altered female puberty,
856 gynecomastia (i.e., the increase of male breast glands in pubescent boys), changes in semen parameters,
857 pregnancy complications and loss, altered fertility and time to pregnancy, endometriosis and
858 adenomyosis, uterine leiomyoma, sexual dysfunction in males, sexual dysfunction in females, polycystic
859 ovary syndrome, age at menopause, as well as sex ratio.
Page 28 of 122
-------
860
861
862
863
864
865
866
867
868
869
870
871
872
873
874
875
876
877
878
879
880
881
882
883
884
885
886
887
888
889
890
891
892
893
894
895
896
897
898
899
900
901
PUBLIC RELEASE DRAFT
DECEMBER 2024
Data quality evaluation criteria and methodology used by Health Canada considered 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 BBP and its metabolites and decreased odds of
polycystic ovary syndrome. There was also limited evidence for the association with infant sex ratio at
birth (i.e., male excess associated with maternal exposure to MBzP and/or MBP). There was inadequate
evidence for the association between BBP and its metabolites and sexual dysfunction in males and
females, changes in semen parameters and time to pregnancy. The level of evidence could not be
established for the association between BBP and its metabolites and altered fertility. There was no
evidence for the association between exposure to BBP and its metabolites and altered male puberty,
gynecomastia, pregnancy loss, endometriosis and adenomyosis, and uterine leiomyoma. All other
reproductive outcomes (i.e., altered male or female puberty, gynecomastia, pregnancy complication and
loss) did not have reported evidence of association with BBP and/or its metabolites.
Sixty-five studies were assessed by Health Canada (2018b) to evaluate the association between exposure
to BBP and growth and developmental outcomes (outcomes listed in Table 3-1). There was limited
evidence of association for BBP and its metabolites and postnatal growth in infants/children (with some
variations regarding the direction of the associations) and altered placental gene expression. There was
inadequate evidence of association for BBP and its metabolites and the following outcomes: birth
measures, placental development, preterm birth and gestational age, postnatal DNA methylation, and
sperm DNA damage/apoptosis. There was no evidence of association for BBP and its metabolites and
AGD, as well as male infant genitalia (e.g., hypospadias and cryptorchidism).
The relationship between BBP 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 MBzP with thyroid-related hormones and 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 MBzP and 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
between BBP or its metabolite (MBzP) and male reproductive outcomes, including AGD and
hypospadias/cryptorchidism following in utero exposures; pubertal development following in utero or
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."
Page 29 of 122
-------
902
903
904
905
906
907
908
909
910
911
912
913
914
915
916
917
918
919
920
921
922
923
924
925
926
927
928
929
930
931
932
933
934
PUBLIC RELEASE DRAFT
DECEMBER 2024
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., National Toxicology Program's Office of Health Assessment and
Translation (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 (BBP, DEP, DBP,
DIBP) 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, which contribute to 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. (2018) found that although it is difficult to determine whether phthalates cause male
reproductive toxicity due to inconsistency across studies, the most consistent studies were those looking
at semen parameters. Several medium quality studies contributed to the moderate level of evidence for
the association between BBP exposure and a decline in the motility and overall quality of sperm. There
is also moderate level of evidence from a single high confidence study that reported statistically
significant associations between increased exposure to BBP and either longer time to pregnancy or
reduced fecundability. Evidence for BBP exposure and testosterone, as well as pubertal development,
was deemed indeterminate due to inconsistency in available studies. In five studies, three of which were
medium confidence and reported a non-statistically significant inverse association, and two low
confidence studies which reported no association, Radke et al. (2018) determined that there was slight
evidence for an association between exposure to BBP and AGD which may be due to data availability
and low exposure levels in the studies. Evidence for Hypospadias/cryptorchidism was considered to be
slight and found in one low confidence study.
Table 3-2. Summary of Epidemiologic Evidence of Male Reproductive Effects Associated with
Exposure to BBP"
Timing of Exposure
Outcome
Level of Confidence in Association
In utero
Anogenital distance
Slight
Hypospadias/cryptorchidism
Slight
In utero or childhood
Pubertal development
Indeterminate
Semen parameters
Moderate
Adult
Time to pregnancy
Moderate
Testosterone
Indeterminate
Male Reproductive Outcomes Overall
Moderate
"Table fromFieure 3 in Radke et al. (2018).
Page 30 of 122
-------
935
936
937
938
939
940
941
942
943
944
945
946
947
948
949
950
951
952
953
954
955
956
957
958
959
960
961
962
963
964
965
966
967
968
969
970
971
972
973
974
PUBLIC RELEASE DRAFT
DECEMBER 2024
Radke et al. (2019b) evaluated the associations between BBP or its metabolite (MBzP) and female
reproductive outcomes, including pubertal development (5 studies), time to pregnancy (3 studies),
spontaneous abortion (5 studies) and preterm birth (6 studies). Radke et al. (2019b) determined the
evidence for whether there is a relationship between BBP exposure and pubertal development is
indeterminate2 because the investigations reported conflicting results regarding the onset of puberty,
pubic hair development, and breast development. The evidence for association between fecundity and
spontaneous abortion and BBP exposure was also indeterminate. Finally, the authors determined that
there was slight evidence of association between preterm birth and BBP exposure.
3.1.1.1.3 NASEM report (2017)
NASEM (2017) evaluated the association between BBP exposure and the following outcomes:
hypospadias, testosterone and AGD. 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 BBP, 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
OHAT to assign confidence ratings and determine the certainty of the evidence to ultimately draw
hazard conclusions (NTP. 2015).
NASEM concluded that there was inadequate evidence to establish an association between prenatal
exposure to BBP 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 BBP is associated with a decrease in fetal testosterone in males,
given the various matrices used to measure testosterone (amniotic fluid, maternal serum, or cord blood),
the differences in timing of exposure (during pregnancy or at delivery), and the limited number of
studies. NASEM concluded that the available studies (meta-analysis included three prospective cohort
studies) do not support an association between BBP exposure and decreased AGD. However, NASEM
found moderate confidence in the evidence of association between BBP (MBzP) and AGD. This finding
is inconsistent with the conclusions of Radke et al. (2018). who found slight evidence of an association
between exposure to BBP and AGD. The AGD effect estimates in NASEM (2017) for BBP (% change
[95% CI] = -1.43 [-3.47, 0.61] [p = 0.17]) 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 BBP exposure and male reproductive outcomes. Radke et al.
(2018) concluded that there was a slight level of confidence in the association between exposure to BBP
and AGD, while Health Canada (2018b) and NASEM (2017) did not. Radke et al. (2018). also found a
slight level of confidence in the association between exposure to BBP and cryptorchidism/hypospadias,
but this association was not consistent with the findings of Health Canada (2018b) or NASEM (2017).
The scope and purpose of the assessments by Health Canada (2018b). systematic review articles by
2 Radke et al. (2019b; 2018) define moderate evidence descriptors as "evidence that supports a hazard, differentiated by
the quantity and quality of information available to rule out alternative explanations for the results." Slight and
indeterminant evidence descriptors are defined as "evidence that could support a hazard or could support the absence of a
hazard. These categories are generally limited in terms of quantity or confidence level of studies and serve to encourage
additional research across the exposure range experienced by humans."
Page 31 of 122
-------
975
976
977
978
979
980
981
982
983
984
985
986
987
988
989
990
991
992
993
994
995
996
997
998
999
1000
1001
1002
1003
1004
1005
1006
1007
1008
1009
1010
1011
1012
1013
1014
1015
1016
1017
1018
1019
1020
1021
PUBLIC RELEASE DRAFT
DECEMBER 2024
Radke et al. (2018). and the report by NASEM (2017) differ from that of Health Canada and may be
related to differences in confidence conclusions drawn for AGD. Health Canada (2018b) was the most
comprehensive review, and considered 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 Section 4). The results of the animal and
epidemiological systematic reviews were considered together by NASEM (2017) to draw hazard
conclusions. Each of the existing assessments covered above considered a different number of
epidemiological outcomes and used different data quality evaluation methods for risk of bias. Despite
these differences and limitations of the epidemiological data, each assessment provides qualitative
support as part of the weight of scientific evidence.
3.1.1.2 EPA summary of new studies (2018- 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 24 new developmental
and 16 new reproductive epidemiology studies published between 2018 to 2019. Eleven of those studies
covered female reproductive outcomes (1 high confidence, 9 medium confidence, and 1 uninformative),
and 5 medium confidence studies investigated male reproductive outcomes. Three medium confidence
studies (Lee et al.. 2020; Chin et al.. 2019; Arbuckle et al.. 2018) found significant associations with
exposure to BBP and female reproductive outcomes, including associations with a slower rise in hCG
(Chin et al.. 2019). increased AGD at birth (Arbuckle et al.. 2018). and increased uterine fibroids (Lee et
al.. 2020). However, there were no significant findings for male reproductive outcomes, aside from male
pubertal outcomes. On the other hand, of the 24 male developmental studies, there were six studies
(Burns et al.. 2022; Bloom et al.. 2019; Amin et al.. 2018; Berger et al.. 2018; Boss et al.. 2018; Huang
et al.. 2018) with significant outcomes (1 high confidence, 2 medium confidence, 5 low confidence).
Studies reporting an association are discussed further below.
In text below, EPA discussed the evaluation of the new studies by outcome with significant results that
contribute to the weight of scientific evidence. 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 Butyl
BenzylPhthalate (BBP) (U.S. EPA. 2024b);
Draft Data Extraction Information for Environmental Hazard and Human Health Hazard
Animal Toxicology and Epidemiology for Butyl Benzyl Phthalate (BBP) (U.S. EPA. 2024a).
Developmental Outcomes for Males.
Twenty studies were evaluated for the association between BBP and developmental outcomes including
birth measures, size trajectory, fetal loss, pubertal development, and gestational duration. Of those
studied, two were high confidence, 12 were of medium confidence, and six were of low confidence. Two
medium confidence studies looked at the associations between BBP and birth measures without
significant results. In a low confidence study (Huang et al.. 2018). mother-infant pairs from Wuhan,
China reported a significant positive association between late pregnancy maternal urinary MBzP and
birth length in boys [Beta (95% CI) per ln-ug/L increase in MBzP = 0.15 (0.01, 0.28)]. This study also
reported significant associations with BBP metabolites and gestational age [beta (95% CI) per ln-ug/L
increase in MBzP overall = 0.16 (0.03, 0.29); and in boys = 0.22 (0.04, 0.41)]. A second low confidence
study reported associations between BBP and gestational duration (Boss et al.. 2018) of mother-infant
pairs from Boston, Massachusetts reported significant positive associations between prenatal urinary
MBzP and gestational age [HR (95% CI) per interquartile range increase in MBzP averaged over three
Page 32 of 122
-------
1022
1023
1024
1025
1026
1027
1028
1029
1030
1031
1032
1033
1034
1035
1036
1037
1038
1039
1040
1041
1042
1043
1044
1045
1046
1047
1048
1049
1050
1051
1052
1053
1054
1055
1056
1057
1058
1059
1060
1061
1062
1063
1064
1065
1066
1067
1068
1069
1070
PUBLIC RELEASE DRAFT
DECEMBER 2024
samples collected between 4.7 and 29.3 weeks gestation = 1.15 (1.03, 1.27); and for repeated measures
of urinary MBzP collected at up to 38.3 weeks gestation = 1.13 (1.05, 1.22)].
Developmental Outcomes for Females.
One high confidence study of mother-infant pairs from Charleston, South Carolina reported significant
inverse associations across all tertiles of prenatal (18 to 22 weeks gestation) urinary MBzP and small for
gestational age [OR (95% CI) = 0.30 (0.10, 0.85) for T2 vs. T1 and 0.29 (0.10, 0.81) for T3 vs. Tl]
(Bloom et al.. 2019). One medium confidence study looked at the association between BBP and female
reproductive hormones. This study (Arbuckle et al.. 2018) reported a significant inverse association
between prenatal first trimester urinary MBzP and anoclitoris distance in infant girls [Beta (95% CI) per
ln-ug/L increase in MBzP = -1.2401 (-1.9080, -0.5723); p-value = 0.0004], No significant results were
reported for other anthropometric measurements in females.
Other Developmental Outcomes.
Two low confidence studies reported associations between BBP and its metabolite and size trajectory.
The first low confidence study (Amin et al.. 2018) reported significant positive associations between
MBzP exposure and body mass index z-score (Beta = 0.18, p-value = 0.002) and waist circumference
(Beta = 0.22, p-value < 0.001) in Iranian children and adolescents aged 6 tol8 years. The other low
confidence study (Durmaz et al.. 2018) reported a significant positive association between MBzP and
body mass index in Turkish girls with premature thelarche (Spearman correlation coefficient = 0.375, p-
value = 0.041). No significant results were found for fetal loss, anthropometric measures of female
reproductive organs, polycystic ovary syndrome, or male reproductive outcome measures such as
anthropometric measures of male reproductive organs, sperm parameters, prostate, and male
reproductive hormones.
Reproductive Outcomes for Males.
Two medium confidence studies (Burns et al.. 2022; Berger et al.. 2018) reported associations between
BBP and pubertal development. The first medium confidence study (Burns et al.. 2022) reported
significant positive associations between prepubertal BBP exposure (measured between 8 and 13 years
of age) and pubertal onset outcomes in Russian boys. These included testicular volume > 3 mL [mean
shift in months (95% CI) = 5.6 (0.3, 11.0) for Q3 vs. Q1 ; 5.6, (0.6, 10.7) for Q4 vs. Ql; p-value for
trend =0.006], Tanner Genitalia Stage > 2 [Mean shift in months (95% CI) = 7.5, (1.1,13.8) for Q4 vs.
Ql; p-value for trend =0.02], and Tanner Pubarche stage > 2 [Mean shift in months (95% CI) =15.1
(8.0, 22.2) for Q3 vs. Ql and 14.2 (7.4, 21.0) for Q4 vs. Ql; p-value for trend < 0.001], The other
medium confidence study (Berger et al.. 2018) reported significant positive associations were between
prenatal (mean 14.0 and 26.9 weeks' gestation) MBzP exposure and age at onset of thelarche in girls
[Mean shift in months (95% CI) =1.9 (0.2, 3.6); for overweight/obese girls = 3.9 (1.2, 6.7)] and with
pubarche onset in normal weight boys [Mean shift in months (95% CI) = 3.5, (0.4, 6.5)]. Significant
inverse associations were observed for boys for onset of gonadarche [Mean shift in months (95% CI) = -
3.1(-5.2, - 0.9); for overweight/obese boys = -4.3 (-6.8, -1.8)] and for pubarche onset in
overweight/obese boys [Mean shift in months (95% CI) = -3.6, (-5.7, -1.4)].
Reproductive Outcomes for Females.
One medium confidence study looked at the association between BBP exposure and fecundity/increased
time to pregnancy. This medium confidence study (Chin et al.. 2019) of North Carolina women without
known fertility issues reported a significantly altered pattern of human chorionic gonadotropin (hCG)
rise during the first 6 days after implantation among women with urinary MBzP levels above vs. below
the median [p-value for the association between MBzP concentration above median and rate of hCG rise
= 0.04], No significant results were reported for fecundity outcomes (type of corpus luteum rescue, time
Page 33 of 122
-------
1071
1072
1073
1074
1075
1076
1077
1078
1079
1080
1081
1082
1083
1084
1085
1086
1087
1088
1089
1090
1091
1092
1093
1094
1095
1096
1097
1098
1099
1100
1101
1102
1103
1104
1105
1106
1107
1108
1109
1110
1111
1112
1113
1114
1115
PUBLIC RELEASE DRAFT
DECEMBER 2024
from ovulation to implantation). One medium confidence study (Lee et al.. 2020) looked at the
association between BBP and fibroids in adult premenopausal women in Korea. Significantly increased
odds of uterine fibroids in quartile 2 (Q2) compared to quartile 1 (Ql) of urinary MBzP [OR (95% CI)
for Q2 vs. Ql = 4.82 (1.09-21.27)] were reported. Associations for quartiles 3 or 4 were positive but not
statistically significant.
Conclusion.
EPA considered the conclusions of Health Canada (2018b) and NASEM (2017) as well as Radke et al.
(2018) and agrees that while there may be evidence of an association between BBP and male
development and reproductive outcomes including sperm quality and AGD, it is not enough to conclude
a causal relationship. Moreover, new studies identified by EPA from 2018 to 2019 do not alter the
previous conclusions from Health Canada (2018b) and NASEM (2017). and systematic review articles
published by Radke et al. (2018) regarding developmental and reproductive outcomes. Although there is
moderate evidence of association between BBP and AGD health outcomes 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 with exposure characterization
of individual phthalates, including source or exposure and timing of exposure as well as co-exposure
confounding with other phthalates, discussed in Section 1.1. The epidemiological studies provide
qualitative support as part of the weight of scientific evidence.
3.1.2 Summary of Laboratory Animal Studies
EPA identified 14 oral exposure studies (all of rats) that have investigated the effects of BBP on the
developing male reproductive system (Gray et al.. 2021; Spade et al.. 2018; Debartolo et al.. 2016;
Schmitt et al.. 2016; Ahmad et al.. 2014; Furr et al.. 2014; Howdeshell et al.. 2008; Aso et al.. 2005; Tyl
et al.. 2004; Wilson et al.. 2004; Ema et al.. 2003; Ema and Miyawaki. 2002; Gray et al.. 2000; Nagaoet
al.. 2000). EPA identified ten of these through review of prior assessments as described in Section 1.2.1,
and 4 of these were identified through systematic review of literature as described in Sections 1.2.2 and
1.2.3. No studies evaluating the developmental and/or reproductive toxicity of BBP are available for
inhalation or dermal exposure routes.
There are numerous sources and assessments creating a robust data set demonstrating adverse male
reproductive system effects following developmental exposure to BBP, which are summarized in Table
3-3, and include 3 multi-generational exposure assessments. These assessments include a variety of
endpoints to be used for evidence integration, including phenotypic changes, organ-level changes, and
mechanistic outcomes. Importantly, all animal studies conducted exposures during the developmental
masculinization programming window (i.e., GD 15.5 to 18.5 for rats; GD 14 to 16 for mice; gestational
weeks 8 to 14 for humans), which may disrupt cellular responses (e.g., testicular testosterone
production) and lead to antiandrogenic effects on the developing male reproductive system (MacLeod et
al.. 2010; Welsh et al.. 2008; Carruthers and Foster. 2005). Available oral exposure studies of BBP
evaluating developmental and reproductive outcomes are summarized in Table 3-3. Most of the
available studies evaluate effects on the developing male reproductive system consistent with a
disruption of androgen action following gestational, perinatal, pre-pubertal, or multi-generational oral
exposures to BBP. However, several studies are available that evaluate other developmental outcomes
(e.g., post-implantation loss, resorptions, fetal body weight, female developmental effects, etc.). Effects
Page 34 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
1116 on the developing male reproductive system (Sections 3.1.2.1 and 3.1.2.2) and other developmental and
1117 reproductive outcomes (Section 3.1.2.3) are discussed below.
Page 35 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Table 3-3. Summary of BBP Oral Exposure Studies Evaluating Effects on
the Developing Male Reproductive System
Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
(Ema et al
2003)
Pregnant Wistar rats (16
dams/group) were exposed to
0, 167, 250, or 375 mg/kg-
day MBP via oral gavage
from GD 15-17. Dams were
sacrificed, and fetal tissue
collected on GD 21.
167/250
I AGD, Cryptorchidism
Developmental Outcomes
- | AGD
- Cryptorchidism
(Howdeshel
1 et al.,
2008)
Pregnant SD rats (4-9
dams/group) were exposed to
0, 100, 300, 600, or 900
mg/kg-day BBP via oral
gavage from GD 8-18. Dams
were sacrificed, and fetal
tissue collected on GD 18.
100/300
I ex vivo fetal testicular
testosterone production
Developmental Outcomes
- 1 ex vivo fetal testes testosterone production
(Furr et al.,
2014)
Pregnant Harlan SD rats
were exposed to 0, 100, 300,
600, or 900 mg/kg-day BBP
(Block 36) via oral gavage
from GD 14-18. Dams were
sacrificed, and fetal tissue
collected on GD 18. (Block
36)
None/100
I ex vivo fetal testicular
testosterone production
Developmental Outcomes
- 1 ex vivo fetal testicular testosterone production
Pregnant Harlan SD rats
were exposed to 0, 11, 33, or
100 mg/kg-day BBP (Block
37) via oral gavage from GD
14-18. Dams were
sacrificed, and fetal tissue
collected on GD 18. (Block
37)
100/None
None
Unaffected Outcomes
- No effect on ex vivo fetal testicular testosterone production
(Grav et al.,
2000)
Pregnant SD rats (5-10
dams/group) were exposed to
0 or 750 mg/kg/-day BBP via
oral gavage from GD 14-
None/750
| AGD (PND2), t male NR
(PND13), I reproductive
Developmental Outcomes
- | AGD (PND 2)
- t NR (PND 13)
Page 36 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
PND 3. Dams were allowed
to give birth naturally and
outcomes were evaluated in
male offspring on PND 2,
PND 3, PND 13, and at
maturity (3-7 months of
age).
organ weights, reproductive
organ malformations
- Permanent nipples (3-7 months)
- i absolute testes, LABC, SV, ventral prostate, glans penis, epididymis, cauda
epididymis, caput-corpus epididymis weight (3-7 months)
- Incomplete PPS due to genital malformations
- Reproductive tract malformations (cleft phallus, hypospadias, vaginal pouch, SV
and epididymal agenesis, fluid filled testis, small testis, testis absent, abnormal
gubernaculum) (3-7 months)
- Undescended testes (3-7 months)
Unaffected outcomes
- Mean age at PPS
- Serum testosterone (3-7 months)
(Soade et
al.. 2018)
Pregnant SD rats (3-6
dams/group) were exposed to
0 or750 mg/kg-day BBP via
oral gavage from GD 17-21.
Dams were sacrificed, and
fetal tissue collected on GD
21.
None/750
I ex vivo fetal testicular
testosterone production, f
MNG incidence
Developmental Outcomes
- 1 ex vivo fetal testicular testosterone production
- t Incidence of MNGs
(Wilson et
al.. 2004)
Pregnant SD rats (3
dams/group) were exposed to
0 or 1000 mg/kg-day BBP
via oral gavage from GD 14-
18. Dams were sacrificed,
and fetal tissue collected on
GD 18.
None/1000
I ex vivo fetal testicular
testosterone production, J,
testicular InsB mRNA
expression
Developmental Outcomes
- 1 ex vivo fetal testicular testosterone production
Mechanistic Outcomes
- i Testicular Ins!3 mRNA
Unaffected Outcomes
- Testicular progesterone production
(Ema and
Mivawaki,
2002)
Pregnant Wistar rats (16
dams/group) were exposed to
0, 250, 500, or 1000 mg/kg-
day BBP via gastric
intubation from GD 15-17.
Dams were sacrificed and
fetal tissue collected on GD
21.
250/500
| AGD, | AGI,
Cryptorchidism, f
Transabdominal testicular
ascent
Developmental Outcomes
- | AGD
- | AGI
- Cryptorchidism
- t Transabdominal testicular ascent
Unaffected outcomes
Fertility index, gestation indices
Page 37 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
(Nasao et
al.. 2000)
Two-generation Study of
Reproduction (Guideline not
stated): Male and female SD
rats (20-24/group) were
exposed continuously via
oral gavage to 0, 20, 100,
500 mg/kg-day from 8-10
weeks of age for two-
generations.
100/500
I AGD, I serum
testosterone, j reproductive
organ weights, testicular
pathological changes,
delayed PPS
Developmental Outcomes
- | AGD (F1 PND 0)
- i serum testosterone (F0 & F1 adults)
- i absolute testes & epididymis weight (F1 PND 22)
- i absolute testes, epididymis, ventral prostate weight (F1 adults)
- Testicular pathology (j spermatocytes in seminiferous tubules (F1 PND 22);
atrophy of seminiferous tubules (F1 adults); j germ cells in seminiferous tubule
(F1 adults); testicular edema (F1 adults); decreased sperm in epididymis, with
cell debris (F1 adults)
- Delayed PPS (Fl)
Unaffected Outcomes
- Mating, fertility, delivery indices (F0, Fl); gestation length (F0, Fl); absolute
reproductive organ weight (testes, epididymides, ventral prostate, SV; F0 adults);
absolute SV weight (Fl adults); testicular pathology (F0); sperm motility and
concentration (F0, Fl adults); serum testosterone (Fl PND 22); hypospadias
(Fl), cryptorchidism (Fl)
(Aso et al..
2005)
Two-generation Study of
Reproduction (Guideline not
stated): Cij:CD(SD)IGS rats
(24/dose) were exposed via
oral/gavage to 0, 100, 200,
400 mg/kg-day continuously
for two generations.
None/100
I AGD, softening of testes j
spermatozoa in epididymis, j
germ cells in epididymal
lumen
Developmental Outcomes
- | AGD (100-400/F2 PND 4)
- Low rate for completed PPS (400/F1 males)
- i absolute epididymis weight (400/F0 adults; 200/F1 adults) & SV (400/F1
adults)
- t incidence of small testes (400/F1 adult), softening of testes (100/F1 adult); t
incidence of small or hypoplastic epididymides (400/F1 adult)
- Testicular pathology (e.g., Leydig cell hyperplasia (400/F0 & 400/F1 adults),
diffuse atrophy of testicular seminiferous tubules (400/F1 adults); j spermatozoa
in epididymides (400/F0; 100/F1 adults), j germ cells in epididymal lumen (Fl
adults at 100), bilateral or unilateral partial aplasia or unilateral aplasia of
epididymides (400/F1 adults)
Unaffected outcomes
- Mating index, days required for mating, gestation length, # implantations,
fertility index, delivery index, gestation index, # of pups delivered, # of sperm in
testis, epididymal sperm motility or morphology (F0 and Fl parents); serum
hormones (FSH, LH, testosterone, estradiol (F0 and Fl parents); absolute testis
and ventral prostate weight (Fl adults); AGD (Fl pups)
Page 38 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 202^
Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
(Tvl et al
2004)
Two-generation Study of
Reproduction (GLP-
compliant and adhered to
OPPTS 870.3800 [August
1998]): CD rats (~20/dose)
were exposed via oral/diet to
0, 750, 3750, 11,250 ppm
BBP (eq. 0, 50, 250, 750
mg/kg-day) continuously for
two generations.
50/250
I AGD, I absolute testes
weight
Developmental Outcomes
- | AGD (F1 and F2 at PND 0)
- i absolute testes weight (F1 weanlings, PND 21)
- i Mating and fertility indices (750/F1)
- i epididymal sperm concentration & motility (750/F1 adults)
- i absolute testes, epididymis, prostate, SV weight (750/F1 adult)
- i absolute testes weight (750/F2 weanlings, PND 21) and j epididymis weight
(750/F1 weanlings, PND 21)
- NR (750/F1 and F2, PND 11-13)
- Delayed PPS (750/F1)
- Undescended testes (750/F1 pups, PND 4)
- Gross malformations (missing epididymis (whole or part), epididymis reduced in
size, missing testes, testes reduced in size, and undescended testis(es) (750/F1
weanlings, PND 21)
- Gross malformations (hypospadias, missing reproductive organ or portion(s) of
organs and/or abnormal organ size and/or shape) (750/F1 adults)
- Gross malformations (missing SVs, missing epididymides) (750/F2 pups, PND
4)
- Testicular pathology (epididymal aspennia, testis dilation, seminiferous tubule
degeneration & atrophy) (750/F1 adult)
Unaffected Outcomes
- Mating, fertility, gestation, pregnancy indices (F0); gestational and pregnancy
indices (Fl); absolute testes, epididymis, prostate, SV weight (F0); epididymal
sperm concentration and motility (F0 adults)
New Literature, as Identified in Section 1.2.3
(Ahmad et
al.. 2014)
Pregnant Albino rats (>6
dams/group) were exposed to
0, 4, 20, or 100 mg/kg-day
BBP via oral gavage from
GD 14-21. Dams were
allowed to give birth
naturally, and male offspring
were sacrificed on PND 5,
20/100
i serum testosterone, j
absolute weight of
epididymis and prostate, j
sperm count, j percent
motile sperm, f percent
abnormal sperm
Developmental Outcomes
- i serum testosterone (Fl adults, PND 75)
- i absolute epididymis and prostate weight (Fl adults, PND 75)
- i sperm count, j percent motile sperm, f percent abnormal sperm (Fl adults,
PND 75)
- i pup body weight (4-100, Fl, PND 1)
Page 39 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
25, or 75. Endpoints
evaluated in F1 from PND 1-
PND 75.
- i body weight (20 and 100, PND 75)
Unaffected Outcomes
- Litter size, live/dead pups, sex ratio (PND1); Anogenital distance (PND5 &
PND25); testis descent; Viability index (PND4); Weaning index (PND21);
testicular 17(3-HSD activity (PND 75)
(Grav et al..
2021)
Pregnant Harlan SD rats (3-4
dams/group) were exposed to
0, 11,33, 100, 300, 600, or
900 mg/kg-day BBP via oral
gavage from GD 14-18.
Dams were sacrificed, and
fetal tissue collected on GD
18.
11/33
i fetal testicular mRNA
expression of steroidogenic
genes, including his/3
Developmental Outcomes
- 1 ex vivo fetal testes testosterone production (300)
Mechanistic Outcomes
- i fetal testicular expression of lnsl3, as well as steroidogenic genes (Star (100),
CypllaL Cypllbl, Cvpl7al (300), Dhcr7 (11), Cypllbl (11), Hsd3b (100),
and Scarbl)
Additional Remarks
Data are an expansion of previous dose response studies (Furr et al., 2014;
Howdeshell et al., 2008)
Pregnant Charles River SD
rats (3-4 dams/group) were
exposed to 0, 100, 300, 600,
or 900 mg/kg-day BBP via
oral gavage from GD 14-18.
Dams were sacrificed and
fetal tissue collected on GD
18. (Block 78)
100/300
I ex vivo fetal testicular
testosterone production
Developmental Outcomes
- 1 ex vivo fetal testes testosterone production
Mechanistic Outcomes
i fetal testicular expression of Insl3 (600) and steroidogenic genes (Star (600),
Cypllal (600), Cypl7al (600), Dhcr7 (900), Cypllbl (600), Hsd3b (900),
Scarbl (600))
(Schmitt et
al.. 2016)
Female C57B1/6J mice were
gavaged with 0 or 500
mg/kg-day BBP on GD 9-16.
Dams were allowed to give
birth naturally, and male
pups were sacrificed at 4, 10,
or 20 weeks of age.
None/500
I AGD, I Serum testosterone
Developmental Outcomes
- i AGD (10 and 20 weeks)
- i Serum testosterone (10 and 20 weeks)
(Debartolo
et al.. 2016)
Pregnant SD rats were
exposed to 0 or 10 ng/mL-
_Cl
I AGD, I Relative body
weight
Developmental Outcomes
- | AGD
Page 40 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
day BBP via spiked food
pellet (solution pipetted onto
pellet) during GD 5-7
through weaning on PND 23.
Pups were necropsied on
PND 23.
- i Relative body weight
Abbreviations: j = Statistically significant decrease; t = Statistically significant increase; AGD = Anogenital distance; AGI = Anogenital index; BW = Body weight; CD =
Charles River Sprague-Dawley; GD = Gestation day; LABC = Levator ani/bulbocavernosus muscles; LOAEL = Lowest-observed-adverse-effect level; MNGs = Multinucleated
gonocytes; NOAEL = No-observed-adverse-effect level NR = Nipple retention; PND = Postnatal day; PPS = Preputial separation; SD = Sprague-Dawley; SV = Seminal vesicle.
" Achieved dose, including NOAEL/LOAEL dose, cannot be calculated in mg/kg-dav, because dam bodv weight and food consumption were not reported (Debartolo et al..
2016).
1119
Page 41 of 122
-------
1120
1121
1122
1123
1124
1125
1126
1127
1128
1129
1130
1131
1132
1133
1134
1135
1136
1137
1138
1139
1140
1141
1142
1143
1144
1145
1146
PUBLIC RELEASE DRAFT
DECEMBER 2024
3.1.2.1 Developing Male Reproductive System
EPA previously developed a weight of scientific evidence analysis and concluded that oral exposure to
BBP can induce effects on the developing male reproductive system consistent with a disruption of
androgen action (see EP A's Draft Proposed Approach for Cumulative Risk Assessment of High-Priority
and a Manufacturer-Requested Phthalate under the Toxic Substances Control Act (U.S. EPA. 2023 a)).
Notably, EPA's conclusion was supported by the Science Advisory Committee on Chemicals (SACC)
(U.S. EPA. 2023b). A brief summary of the MOA for phthalate syndrome and data available for BBP
supporting this MOA is provided below. Readers are directed to EPA's Draft Proposed Approach for
Cumulative Risk Assessment of High-Priority and a Manufacturer-Re quested Phthalate under the Toxic
Substances Control Act (U.S. EPA. 2023a) for a more thorough discussion of BBP'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.
Mode of Action for Phthalate Syndrome
A MOA for phthalate syndrome is shown in Figure 3-1, which explains the link between gestational
and/or perinatal exposure to BBP and effects on the male reproductive system in rats. The MOA has
been described in greater detail in EPA's Draft Proposed Approach for Cumulative Risk Assessment of
High-Priority Phthalates and a Manufacturer-Requested Phthalate under the Toxic Substances Control
Act (J.S. EPA, 2023a) and is described briefly below.
Chemical Structure
and Properties
Adverse Organism
Outcomes
Phthalate
exposure during
critical window of
development
Fetal Male Tissue
AR dependent
mRNA/protein
synthesis
^ ^ #
¦=:>
ZS
V
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 Chcr7 Mvd Ela3b
StAfi Ebp Nsdhl Insl3
Cypllal Fdps RGD1B64999 Lhcgr
Cypllbl Hmgcr Tm7sf2 Inha
Cypllb2 Hmgcsl Cyp46al NrQbl
Cypl7al HSdSb Ldlr RhaxlO
Cyp51 Fldil InsigJ Wnt7a
4, Testosterone
synthesis
I*
l* Gene
expression
(INSL3, lipid
j metabolism,
cholesterol and
androgen synthesis
and transport)
' ^ '
K
-i- INSL3 synthesis
Fetal Leydig cell
Abnormal cell
apoptosis/
proliferation
(Nipple/areolae
retention, 4, AGD,
Disrupted testis
tubules, Leydig celi
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
<
<0
l=D>
Suppressed
gubernacular cord
development
(transabdominal
Phase)
Impaired
s
fertility
1}
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 al.. 2021; Schwartz
et al.. 202.1; Howdcshcll et al.. 2017).
Abbreviations: AR = Androgen receptor; INSL3 = Insulin-like growth factor 3; MNG = Multinucleated gonocyte;
PPARa = Peroxisome proliferator-activated receptor alpha.
Page 42 of 122
-------
1147
1148
1149
1150
1151
1152
1153
1154
1155
1156
1157
1158
1159
1160
1161
1162
1163
1164
1165
1166
1167
1168
1169
1170
1171
1172
1173
1174
1175
1176
1177
1178
1179
1180
1181
1182
1183
1184
1185
1186
1187
1188
1189
1190
1191
1192
1193
1194
1195
PUBLIC RELEASE DRAFT
DECEMBER 2024
Although the MOA underlying phthalate syndrome has not been fully established, key events at the
cellular-, organ-, and organism-level are generally understood (Figure 3-1). In general, molecular events
(i.e., the molecular initiating event) of disrupted steroid and lipid metabolism cellular responses are
critical in the phthalate syndrome MOA. Several studies have provided evidence against the direct
impact of phthalates on androgen receptor and peroxisome proliferator-activated receptor alpha binding
for transcriptional activity modulation (Gray et al.. 2021; Foster. 2005; Foster et al.. 2001; Parks et al..
2000). Other studies have suggested depletion of elemental zinc, which is essential in testicular function,
could perturb function of zinc-containing proteins (e.g., zinc-finger transcription factors or as an enzyme
cofactor), conceivably resulting in adverse organ-level reactions (Gray et al.. 1982; Foster et al.. 1980).
Of note, SF-1, a transcription factor that regulates the INSL3 (insulin-like factor 3) promoter, contains
two zinc-finger motifs that are required for DNA binding. INSL3 is a small peptide hormone critical in
Leydig cell steroidogenic machinery for cellular differentiation and testosterone production (Ivell et al..
2013). However, it is unclear if zinc depletion is a consequence or an upstream event preceding
decreased fetal testosterone synthesis and subsequent steps in the MOA shown in Figure 3-1.
Exposure to BBP during the masculinization programming window (i.e., GDs 15.5 to 18.5 for rats; GDs
14 to 16 for mice; gestational weeks 8 to 14 for humans), in which androgen action drives development
of the male reproductive system, can lead to antiandrogenic effects on the male reproductive system
(MacLeod et al.. 2010; Welsh et al.. 2008; Carruthers and Foster. 2005). Consistent with the MOA
outlined in Figure 3-1, there is experimental evidence of disrupted expression of key genes involved in
lipid metabolism and cholesterol and androgen synthesis. In an early assessment by Wilson et al. (2004).
pregnant SD rats were orally gavaged with 0 or 1000 mg/kg-day BBP during GD 14 to 18, and on GD
18, offspring were assessed for testicular effects. In this assessment, decreased testicular Insl3
expression was noted at 1000 mg/kg-day, coinciding with decreased fetal testicular testosterone
production. However, a clear limitation of this study was large dose spacing without dose-response
assessment. More recently, Gray et al. (2021) investigated the effects of in utero exposure to various
phthalates, including BBP. In this multi-cohort study, pregnant Harlan SD rats were orally gavaged with
11,33, 100, 300, 600, or 900 mg/kg-day BBP from GD 14 to 18, followed by assessment of relevant
anti androgenic genomic and hormonal biomarkers. In offspring, decreased fetal testicular expression of
Insl3 was noted at levels as low as 33 mg/kg-day, along with decreased expression of multiple
steroidogenic genes at all levels of exposure, including steroidogenic acute regulatory protein, Star, and
various cytochromes. This same assessment also included a group pregnant Charles River SD rats
gavaged with 100 to 900 mg/kg-day BBP on GD 14 to 18. Analysis in these offspring found decreased
testicular expression of Insl3 occurring at 600 mg/kg-day, with additional steroidogenic genes
expression disruption occurring at the same level or higher. Overall, available studies provide consistent
evidence that gestational exposure to BBP disrupts mRNA expression of steroidogenic genes and Ins/3
in the fetal testes. Altogether, these data support a mode of action where changes in key genes involved
in steroidogenesis or testosterone transport precede cellular responses and subsequent organ-level
responses consistent with phthalate syndrome.
Within the MOA of phthalate syndrome, disrupted expression of key genes is suggested to impact fetal
Leydig cell production of testosterone, which may contribute to organ-level adverse outcomes. Multiple
studies indicate that gestational BBP exposure during the critical developmental window disrupts
offspring testicular testosterone production (Gray et al.. 2021; Spade et al.. 2018; Furr et al.. 2014;
Howdeshell et al.. 2008; Wilson et al.. 2004). In addition to previously mentioned mechanistic
outcomes, Wilson et al. (2004) noted decreased fetal testicular testosterone production in rats exposed to
1000 mg/kg-day BBP. Likewise, although there appeared to be slight strain-specific effects in BBP
sensitivity, both SD cohorts in Gray et al. (2021) displayed decreased ex vivo testicular testosterone
production, with Harlan SD showing 27 percent decrease in testosterone production at 100 mg/kg-day
Page 43 of 122
-------
1196
1197
1198
1199
1200
1201
1202
1203
1204
1205
1206
1207
1208
1209
1210
1211
1212
1213
1214
1215
1216
1217
1218
1219
1220
1221
1222
1223
1224
1225
1226
1227
1228
1229
1230
1231
1232
1233
1234
1235
1236
1237
1238
1239
1240
1241
1242
1243
1244
PUBLIC RELEASE DRAFT
DECEMBER 2024
and Charles River SD showing 38 percent decrease in testosterone production at 300 mg/kg-day.
Howdeshell et al. (2008) orally-gavaged SD dams to doses ranging from 100 to 900 mg/kg-day BBP
during GD 8 to 18 and assessed ex vivo testicular testosterone production. Here, significant effects were
noted in a dose-response fashion, with decreased testosterone levels noted at doses of 300 mg/kg-day
and above (22% decrease in testosterone production at 300 mg/kg-day). Similarly, Spade et al. (2018)
exposed SD dams to 750 mg/kg-day BBP during GD 17 to 21 and accordingly noted a higher dose effect
of 69 percent decrease in testosterone production in BBP-exposed rats. Lastly, Furr et al. (2014)
conducted a series of studies with BBP. In those studies, Harlan SD rats were orally gavaged with BBP
at levels of 100 to 900 mg/kg-day (Block 36) or 11 to 100 mg/kg-day (Block 37) during GD 14 to 18,
followed by ex vivo testosterone production assessment on GD 18. Interestingly, low-dose assessment in
Block 37 did not indicate any significant BBP effects, even at 100 mg/kg-day. However, a dose-
dependent decrease in ex vivo testicular testosterone production was observed at all levels of BBP
exposure in Block 36 (100, 300, 600, and 900 mg/kg-day), with 100 mg/kg-day resulting in a 53 percent
decrease and 900 mg/kg-day resulting in an 85 percent decrease in ex vivo fetal testicular testosterone
production (Furr et al.. 2014). Although a limitation of the Furr et al. (2014) assessment is the
inconsistent findings of Blocks 36 and 37 at 100 mg/kg-day BBP, likely due to the variability within
limited sample size of a small data set available for BBP, findings are generally consistent with effect-
levels noted in other studies assessing BBP effects on testicular testosterone production (Gray et al..
2021; Howdeshell et al.. 2008). Collectively, these studies demonstrate the ability of gestational BBP
exposure to decrease testicular testosterone production.
In addition to BBP exposure-related decrements in testicular testosterone production and gene
expression of steroidogenic genes, corroborating organ-level responses are also noted across multiple
studies. As with effects commonly noted from other phthalates, sensitive organ-level responses include
testicular histopathological changes, reproductive organ weight changes, and antiandrogenic-related
abnormal growth and development effects, such as decreased anogenital distance (AGD) and nipple
retention in male offspring. AGD, typically corrected for the cube root of body weight ratio, is regarded
a sensitive hallmark of early-life reproductive and developmental androgen disruption from phthalate
exposure (Schwartz et al.. 2019).
Multiple studies have evaluated organ responses to perinatal BBP exposure. Ema et al. (2003) exposed
pregnant Wistar rats to 0, 250, 500, or 1000 mg/kg-day BBP via gastric intubation during GD 15 to 17
and examined fetal offspring on GD 21. In this study, a significant increase in fetuses with
cryptorchidism (i.e., undescended testes) and male fetuses with decreased AGD was noted in groups that
received 500 and 1000 mg/kg-day BBP. Ema et al. (2003) orally gavaged pregnant Wistar rats with
either 0, 167, 250, or 375 mg/kg-day MBP, the major BBP metabolite. Dams were dosed from GD 15 to
17, and dams were sacrificed on GD 21 for fetal collection. In this study, effects of cryptorchidism
incidence and significantly decreased AGD were noted at levels of 250 mg/kg-day and above. In other
studies, many histopathological findings coincide with antiandrogenic effects, such as Spade et al.
(2018) finding, in addition to decreased testosterone production at 750 mg/kg-day BBP, an increased
incidence of multinuclear gonocytes (MNGs). In a study that tested a single dose level, Gray et al.
(2000) orally gavaged pregnant SD rats with 0 or 750 mg/kg-day BBP from GD 14 to PND 3. Dams
were allowed to give birth, and male offspring were sacrificed from PND 2 to mature adults, with
evaluations occurring at PND 2, PND 13, and 3 to 7 months. Decreased AGD was noted at PND 2, and
increased nipple retention was noted at PND 13 in BBP-exposed rats. Additionally, these early-life
effects were accompanied by incomplete preputial separation, increased incidence of undescended testis,
and decreased testes and accessory sex organ weights (e.g., absolute testes, seminal vesicles, LABC,
ventral prostate, glans penis, paired epididymides, cauda epididymides, and caput-corpus epididymides)
at 3 to 7 months of age. Reproductive tract malformations at 750 mg/kg-day BBP were also noted,
Page 44 of 122
-------
1245
1246
1247
1248
1249
1250
1251
1252
1253
1254
1255
1256
1257
1258
1259
1260
1261
1262
1263
1264
1265
1266
1267
1268
1269
1270
1271
1272
1273
1274
1275
1276
1277
1278
1279
1280
1281
1282
1283
1284
1285
1286
1287
1288
1289
1290
1291
1292
1293
PUBLIC RELEASE DRAFT
DECEMBER 2024
including cleft phallus, hypospadias, epididymal agenesis, fluid filled or missing testis, and abnormal
gubernaculum (Gray et al.. 2000).
Three multi-generational studies of male reproductive and developmental outcomes were identified (Aso
et al.. 2005; Tyl et al.. 2004; Nagao et al.. 2000). Although Tyl et al. (2004) was the only study to
explicitly report adherence to multi-generation testing guidelines by U.S. EPA Office of Prevention,
Pesticides, and Toxic Substances Guidelines, the other two-generation studies identified adhered to
similar practices suggested by the Organisation for Economic Co-operation and Development two-
generation reproduction toxicity testing guidelines (OECD. 2018).
Nagao et al. (2000) conducted a two-generation reproductive study in male and female SD rats using
oral exposure doses of 0, 20, 100, and 500 mg/kg-day BBP from 8 to 10 weeks of age in the F0
generation, and female exposure continued during gestation and lactation until postpartum day 21.
Dosing continued into F2 generation, and F1 rat observations were made on PND 0 and PND 21 at
necropsy and in F2 rats on PND 0. In F1 offspring, multiple developmental outcomes were noted
pertaining to hormonal, histopathological, and organ-level changes only at the highest level of exposure.
In F1 adults, decreased circulating testosterone was noted in rats exposed to 500 mg/kg-day, along with
decreases absolute testicular, epididymal, and ventral prostate weights. Testicular pathological incidence
significantly increased at this same dose level and included decreased seminiferous tubule
spermatocytes, seminiferous tubule atrophy, decreased seminiferous tubule germ cell, testicular edema,
and decreased epididymal sperm. Importantly, 500 mg/kg-day BBP also resulted in delayed preputial
separation, an androgen-sensitive sign of puberty, and decreased AGD in F1 and F2 offspring (Nagao et
al.. 2000).
Aso et al. (2005) conducted a similar multi-generational reproductive study in Charles River SD rats, in
which rats were orally gavaged with 0, 100, 200, or 400 mg/kg-day BBP continuously for two
generations, with exposure starting in the F0 parental generation at 5 weeks of age and at 3 weeks of age
(i.e., at weaning) for the F1 parental generation. In F1 offspring, the majority of BBP exposure-related
effects occurred at 400 mg/kg-day, with a few effects also occurring at 100 mg/kg-day. Absolute organ
weights were significantly decreased in F1 adults, including absolute epididymis weight (at 200 and 400
mg/kg-day) and seminal vesicle weight (at 400 mg/kg-day). In F1 adults, in addition to decreased germ
cells in the epididymal lumen at 100 mg/kg-day and above, there was also an increased incidence of
Leydig cell hyperplasia, diffuse atrophy of seminiferous tubules, decreased epididymal spermatozoa,
and epididymal aplasia at 400 mg/kg-day. As expected, these histopathological changes in the F1
generation at 400 mg/kg-day were accompanied by the developmental outcome of lower rate of
completed preputial separation. In F2 pups, reproductive and developmental assessment was largely
limited to physical development analysis. Although AGD was not impacted by exposure in F1 male
pups, F2 male pups showed significantly decreased AGD at BBP exposure levels of 100 mg/kg-day and
above (Aso et al.. 2005). These results by Aso et al. (2005) were also briefly summarized in the
summary paper by Yamasaki et al. (2005).
Tyl et al. (2004) conducted a multi-generational study treating CD rats with BBP in the diet at 0, 750,
3750, and 11,250 ppm (equivalent to 0, 50, 250, and 750 mg/kg-day) continuously for two generations,
with observations made in F1 weanlings (PND 21) and adults and F2 pups. In F1 pups, decreased
absolute testicular and epididymal weight at levels 250 and 750 mg/kg-day were noted, which occurred
in F1 adults as well at 750 mg/kg-day. Also, in F1 adults, decreased epididymal sperm concentration and
motility, as well as decreased absolute prostate and seminal vesicle weight, were noted at 750 mg/kg-
day. In F1 weanling and adult examinations, a myriad of gross malformations (e.g., reduced epididymal
size, missing epididymis, hypospadias, and cryptorchidism) were noted at the highest BBP exposure
Page 45 of 122
-------
1294
1295
1296
1297
1298
1299
1300
1301
1302
1303
1304
1305
1306
1307
1308
1309
1310
1311
1312
1313
1314
1315
1316
1317
1318
1319
1320
1321
1322
1323
1324
1325
1326
1327
1328
1329
1330
1331
1332
1333
1334
1335
1336
1337
1338
1339
1340
1341
PUBLIC RELEASE DRAFT
DECEMBER 2024
level. Further, adverse developmental effects were noted in both F1 and F2 pups. At 750 mg/kg-day,
delayed preputial separation was noted in F1 pups, and increased nipple retention was noted in F1 and
F2 pups. However, in both F1 and F2 pups, the lower level of BBP exposure (at 250 mg/kg-day and
above) resulted in significantly reduced AGD (Tyl et al.. 2004).
In sum, these studies provide consistent evidence that oral BBP exposure in rats, particularly during the
critical window of organogenesis and masculinization of the male reproductive system, can disrupt
androgen action, leading to a cluster of anti-androgenic mechanistic-, cellular, and organ-level outcomes
that are consistent with the MOA for phthalate syndrome (Figure 3-1). As previously noted, this
conclusion was supported by the Science Advisory Committee on Chemicals (SACC) (U.S. EPA.
2023b). presented in 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. 2023a).
which provides a thorough discussion of BBP exposure effects on the developing male reproductive
system and EPA's MOA analysis. Consistent with animal toxicology literature, epidemiological
evaluations across several studies, as reported by Radke et al. (2018). provide some evidence that BBP
exposure is associated with male reproductive toxicity. For example, Radke et al. (2018) found moderate
evidence of an association between BBP and effects on semen parameters and increased time to
pregnancy. These observed developmental effects are assumed to be relevant for extrapolating human
risk, and thus EPA is considering developmental toxicity for dose-response analysis and for use in
estimating risk to human health. EPA's further consideration of developmental toxicity is discussed in
Section 4.
3.1.2.2 New Literature Considered for Non-Cancer Hazard Identification
EPA identified 10 new animal toxicology studies that provide data on PECO-relevant health effects
following exposure to BBP, as discussed in Section 1.2.3. Of these, only 4 studies provided outcomes
evaluating gestational exposure and relevant male reproductive and developmental effects (Table 3-3).
For two of these new studies, EPA identified several limitations or lack of sensitivity, including that of
exposure dose uncertainty, lack of monotonic dose-response, or studies that only included relatively
high doses) (Debartolo et al.. 2016; Schmitt et al.. 2016). Two new studies identified potentially relevant
NOAELs or LOAELs for adverse outcomes on developing male reproductive system (Gray et al.. 2021;
Ahmad et al.. 2014). which are further considered in Section 4 for dose-response assessment. 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 EPA's
consideration of new literature for non-cancer hazard identification are provided in Appendix B.
In a behavioral assessment by Schmitt et al. (2016). pregnant mice were exposed to 0 or 500 mg/kg-day
BBP via gavage from GD 9 to 16, and pups were weaned and analyzed for behavioral physical activity
levels and markers of endocrine disruption activity from postnatal week (PNW) 8 to PNW 20.
Regarding relevant effects on developing male reproductive outcomes identified in this study,
significantly decreased AGD and serum testosterone levels were noted at PNW 10 and 20. However,
notable limitations of this study include that it only included a single high dose group, which precludes
its use in understanding dose-response, and this study does not offer a more sensitive POD than those
provided by earlier literature for effects of the developing male reproductive system; therefore, it was
not considered further. Lastly, it was considered a limitation of the experimental design that BBP
exposure ended on GD 16, which did not include the entire susceptible gestational exposure window of
rats for BBP-related anti androgenic effects (GD 15.5 to 18.5), as was done in other studies. This study is
shown in Table 3-3, and is included in Appendix B. 1 discussion of new literature presenting
reproductive/developmental outcomes and Appendix B.2 discussion of new literature depicting
neurotoxicity outcomes.
Page 46 of 122
-------
1342
1343
1344
1345
1346
1347
1348
1349
1350
1351
1352
1353
1354
1355
1356
1357
1358
1359
1360
1361
1362
1363
1364
1365
1366
1367
1368
1369
1370
1371
1372
1373
1374
1375
1376
1377
PUBLIC RELEASE DRAFT
DECEMBER 2024
A study by DeBartolo et al. (2016) provided data on neurobehavioral effects {i.e., fear conditioning) and
endocrine-disrupting outcomes following developmental exposure to BBP. Pregnant SD rats were
exposed to 0 or 10 [j,g/mL BBP pipetted onto food pellets from GD 14 to PND 23. Regarding relevant
effects on developing male reproductive outcomes identified in this study, significantly decreased
relative body weight and AGD were noted in BBP-exposed male offspring when measured at PND 23.
However, EPA identified substantial limitations of this study which impact the interpretation of the
results and contribute to uncertainty in the data set. The largest limitation of this study includes
substantial uncertainty regarding the achieved dose. BBP stock solution (10 (ag/m L) was pipetted onto
sweetened food pellets and fed to pregnant dams; however, the resulting concentration of the test diets
were not determined, and maternal body weights and feed consumption were not reported, therefore
achieved dose cannot be calculated. Also, treatment for the control and BBP-exposed groups was also
stated to occur -5-7 days during gestation, meaning not all exposed animals may have had the same
exposure time duration. This study is shown in Table 3-3, and is presented in Appendix B.l discussion
of new literature on reproductive/developmental outcomes and Appendix B.2 discussion of new
literature on neurotoxicity outcomes.
Two new studies considered relevant to BBP dose-response assessment and POD derivation were
identified (Gray et al.. 2021; Ahmad et al.. 2014). In the study by Ahmad et al. (2014). albino rats were
gavaged with 0, 4, 20, or 100 mg/kg-day BBP from GD 14 to 21, and male offspring were sacrificed for
analysis on PND, 5, 25, or 75. In this study, authors reported a LOAEL of 100 mg/kg-day (NOAEL of
20 mg/kg-day) based upon decreased serum testosterone, decreased epididymal and prostate weights,
and sperm quality effects in F1 adults measured at PND 75. In Gray et al. (2021). two experiments were
performed in SD rats. In the first experiment, pregnant Harlan SD rats were gavaged with 0, 11,33, 100,
300, 600, or 900 mg/kg-day BBP from GD 14-18, with ex vivo fetal testicular testosterone and
steroidogenic gene expression measurement on GD 18. This study yielded a NOAEL of 11 mg/kg-day
BBP based on decreased Insl3 expression at 33 mg/kg-day; significantly reduced ex vivo fetal testicular
testosterone occurred at 300 mg/kg-day BBP levels and higher. The second experiment was conducted
with pregnant Charles River SD rats gavaged with 0, 100, 300, 600, or 900 mg/kg-day from GD 14 to
18, with ex vivo fetal testicular testosterone and steroidogenic gene expression measured at GD 18. This
study yielded a NOAEL of 100 mg/kg-day BBP based on decreased ex vivo fetal testicular testosterone
occurring at levels of 300 mg/kg-day BBP and higher. These studies are shown in Table 3-4 and are
presented in Appendix B.l discussion of new literature on reproductive/developmental outcomes.
Further, these studies were among new literature considered for dose-response assessment and POD
derivation based upon potentially identified more sensitive effect levels for BBP-related developing
male reproductive toxicity effects, and thus are further discussed in Section 4.
Page 47 of 122
-------
1378
1379
1380
1381
1382
1383
1384
1385
1386
1387
1388
1389
1390
1391
1392
1393
1394
1395
1396
1397
1398
1399
1400
1401
1402
1403
1404
1405
1406
1407
1408
1409
1410
1411
1412
1413
1414
1415
1416
1417
1418
1419
1420
PUBLIC RELEASE DRAFT
DECEMBER 2024
3.1.2.3 Other Developmental and Reproductive Outcomes
In addition to effects on the developing male reproductive system, other developmental effects (e.g.,
decreased fetal weight, resorptions, decreased mating and fertility index, and effects on female
reproductive outcomes) have been observed in experimental animal models following oral exposure to
BBP. However, these effects generally occur at equal or higher doses than those that result in effects on
the developing male reproductive system and frequently coincide with maternal toxicity (Table 3-3).
Data supporting other developmental effects of BBP are discussed below.
In the two-generation reproduction oral exposure studies, Aso et al. (2005) noted developmental toxicity
outcomes in addition to effects on the developing male reproductive system. In SD rats orally gavaged
with 0, 100, 200, or 400 mg/kg-day BBP, Aso et al. (2005) noted decreased AGD in F1 female offspring
at 100 mg/kg-day and above and a reduced fertility index in the F1 generation at 400 mg/kg-day.
In another two-generation study, Nagao et al. (2000) gavaged SD rats to BBP doses of 0, 20, 100, and
500 mg/kg-day from 8 to 10 weeks of age in the F0 generation, and female exposure continued during
gestation and lactation until postpartum day 21. F1 rat observations were made on PND 0 and PND 21 at
necropsy. Here, Nagao et al. (2000) reported decreased male and female F1 pup weight at 100 and 500
mg/kg-day. Further developmental female toxicity was also noted in F1 offspring, where decreased
AGD was reported in F1 pups at 500 mg/kg-day.
In the two-generation study by Tyl et al. (2004). CD rats were exposed to 0, 50, 250, and 750 mg/kg-day
via the diet and numerous developmental and reproductive outcomes in F1 and F2 male and female
offspring were examined. Authors reported numerous effects at the highest dose tested of 750 mg/kg-
day across both sexes, including decreased uterine and ovarian weights in F1 and F2 offspring,
decreased mating and fertility index in F1 generation, decreased implantations in F2 generation,
decreased number of F2 live pups, and decreased fetal body weight in F1 male and female offspring.
Lastly, Howdeshell et al. (2008) gavaged SD rats with 0, 100, 300, 600, or 900 mg/kg-day BBP from
GD 8 to 18 and examined fetal outcomes on GD 18. Authors reported developmental toxicity occurring
at 600 mg/kg-day and above through decreased live fetuses, increased resorption, and increased fetal
mortality. Further, in a single dose-level study conducted by Gray et al. (2000). SD rats were gavaged
with 0 or 750 mg/kg-day BBP from GD 14 to PND 3. In this study, authors reported decreased mean
pup weight at birth when assessed on PND 3 (Gray et al.. 2000).
Collectively, available studies provide consistent evidence that gestational exposure to BBP can result in
a spectrum of developmental effects in addition to those of the developing male reproductive system.
However, effects on the developing male reproductive system (Section 3.1.2.1) occur at much lower
doses than the aforementioned other developmental effects. Specifically, the lowest LOAELs for effects
on the developing male reproductive system occur around 100 mg/kg-day, while the lowest LOAELs for
other developmental outcomes discussed here range from 100 to 750 mg/kg-day, with most effects
occurring at or above 200 mg/kg-day (Table 3-3). Therefore, effects on the developing male
reproductive system are as sensitive and often robust than other endpoints to BBP exposure and are
consistent with a disruption of androgen action and phthalate syndrome.
Page 48 of 122
-------
1421
1422
1423
1424
1425
1426
1427
1428
1429
1430
1431
1432
1433
1434
1435
1436
1437
1438
1439
1440
1441
1442
1443
1444
1445
1446
1447
1448
1449
1450
1451
1452
1453
1454
1455
1456
1457
1458
1459
1460
1461
1462
1463
1464
1465
1466
1467
PUBLIC RELEASE DRAFT
DECEMBER 2024
4 DOSE RESPONSE ASSESSMENT
EPA is focusing its dose-response analysis on developmental and reproductive toxicity, particularly
effects relevant to phthalate syndrome in male rats. These effects are consistently observed across
different strains of rat, varying exposure durations including single and multi-generations, and occur in a
dose-related manner.
EPA identified evidence of other non-cancer hazard endpoints (i.e., liver and kidney toxicity), but did
not perform dose-response analysis of these endpoints because endpoints associated with developing
male reproductive effects are supported by the most robust data set and available information that
indicates male reproductive effects are at least as or more sensitive as other reported effects, increasing
EPA's confidence in using these endpoints for estimating risk to human health. According to previous
assessments by U.S. CPSC (2010) and NICNAS (2015). there is evidence for systemic toxicity
following BBP exposure, including liver and kidney weight effects. However, Health Canada (2015a)
concluded BBP has lower systemic toxicity effects that occur only at much higher exposure levels than
developmental reproductive effects based upon multiple repeated oral dose toxicity studies. For acute
developmental oral exposure studies, LOAEL systemic toxicity effects (such as changes in liver weight)
occurred at levels >200 mg/kg-day, where some systemic effects are observed in females at lower doses
but are not accompanied by changes in clinical chemistry markers or histopathological effects (EC/HC.
2015a). Further, for intermediate exposure duration, the lowest LOAEL for repeated intermediate oral
exposure was determined to be 313 mg/kg-day based on increased liver and kidney weights,
accompanied by histopathological changes (EC/HC. 2015a). No new studies indicating more extensive
hepatic or renal effects, or a more sensitive POD, were identified through the TSCA systematic review
process, and thus EPA considered the conclusions of previous assessments of male developmental
reproductive health effects as most sensitive toxicity indicators as valid.
For hazard endpoints, EPA used a NOAEL/LOAEL approach for the BBP dose-response analysis based
on a subset of critical studies. EPA considered NOAEL and LOAEL values from oral toxicity studies in
experimental animal models (all available studies conducted in rats). The use of a NOAEL/LOAEL
approach is supported by consistency across several studies that have evaluated effects on the
developing male reproductive system consistent with phthalate syndrome that are similar and cluster
around a single human equivalent dose (HED) NOAEL value, which supports identification of a
consensus NOAEL. For one hazard endpoint (i.e., reduced fetal testicular testosterone in rats), EPA
conducted benchmark dose (BMD) modeling as well as an updated 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), DicyclohexylPhthalate (DCHP), andDiisononylPhthalate (U.S. EPA. 2024c). No
dermal or inhalation studies were available that could be used for dose-response assessment. Acute,
intermediate, and chronic non-cancer NOAEL/ LOAEL values identified by EPA are discussed further
in Section 4.2. As discussed further in Section 4.2, EPA considers effects on the developing male
reproductive system consistent with a disruption of androgen action relevant for setting a POD for acute
exposure durations. However, because these acute effects are the most sensitive effects following
exposure to BBP, they are also considered protective of intermediate and chronic duration exposures. As
described in Appendix D, EPA converted oral PODs derived from animal studies to HEDs using
allometric body weight scaling to the three-quarters power (U.S. EPA. 2011c). Differences in dermal
and oral absorption are corrected for in the dermal exposure assessment, allowing the same HEDs to be
used for both oral and dermal routes. In the absence of inhalation studies, EPA performed route-to-route
Page 49 of 122
-------
1468
1469
1470
1471
1472
1473
1474
1475
1476
1477
1478
1479
1480
1481
1482
1483
1484
1485
1486
1487
1488
1489
1490
1491
1492
1493
1494
1495
1496
1497
1498
1499
1500
1501
1502
1503
1504
1505
1506
1507
1508
1509
1510
1511
PUBLIC RELEASE DRAFT
DECEMBER 2024
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 the suite of oral animal toxicity studies primarily demonstrating effects on the
developing male reproductive system consistent with phthalate syndrome when considering non-cancer
PODs for estimating risks for acute, intermediate, and chronic exposure scenarios, as described in
Section 4.2. EPA considered the following factors during study and endpoint selection for POD
determination from relevant non-cancer health effects:
Exposure duration;
Dose range;
Relevance (e.g., considerations of species, whether the study directly assesses the effect, whether
the endpoint the best marker for the toxicological outcome, etc.);
Uncertainties not captured by the overall quality determination;
Endpoint/POD sensitivity; and
Total uncertainty factors (UFs). EPA considers the overall uncertainty with a preference for
selecting studies that provide a lower uncertainty (e.g., lower benchmark MOE) because
provides higher confidence (e.g., use of a NOAEL 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 14 developmental and reproductive toxicity studies across 12 publications (all rat
studies) with endpoints relevant to acute, intermediate, and chronic exposure durations (U.S. EPA. 1996.
1991). These studies were previously discussed in Section 3.1.2 and are summarized in Table 4-1.
Primary endpoints considered relevant to all exposure durations include effects on the developing male
reproductive system consistent with a disruption of androgen action during the critical window of male
reproductive development in rats and other developmental effects, such as resorptions and decreased
body weight. Although single dose studies evaluating the effects of BBP on the developing male
reproductive system are not available, Alam et al. (2015) conducted a single dose gavage study in male
adolescent SD rats (briefly discussed in Appendix C). In this study, one high-dose oral exposure to 500
mg/kg BBP resulted in a spectrum of antiandrogenic outcomes, including increased seminiferous tubule
spermatocyte cell apoptosis and decreased absolute testis weight. Regarding acute developmental
exposures, studies of the toxicologically similar phthalate dibutyl phthalate (DBP) have demonstrated
that a single exposure during the critical window of development can disrupt expression of steroidogenic
genes and decrease fetal testicular testosterone (Johnson et al.. 2012; Johnson et al.. 2011; Thompson et
al.. 2005). Therefore, EPA considers effects on the developing male reproductive system consistent with
a disruption of androgen action to be relevant for setting a non-cancer POD for acute, intermediate, and
chronic exposure durations (see Appendix C for further discussion). Notably, the 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 diisodecyl phthalate (DIDP) and diisononyl phthalate (DINP) draft human
health hazard assessment (U.S. EPA. 2024i).
Page 50 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
1512 Table 4-1. Dose-Response Analysis of Selected Studies Considered for Acute, Intermediate, and Chronic Exposure Scenarios
Brief Study Description
(Reference)
Study POD/ Type
(mg/kg-day)
Effect
HEP
(mg/kg)
Uncertainty
Factors"b
Pregnant Harlan SD rats (n = 3-4 dams/group)
were exposed to 0, 11,33, 100, 300, 600, or 900
mg/kg-day BBP via oral gavage from GD 14-18.
Dams were sacrificed and fetal tissue collected on
GD 18 (Grav et al.. 2021).
NOEL = IF
i Fetal testicular mRNA expression of
steroidogenic genes (including Insl3)
2.6
UFa= 3
UFh=10
Total UF=30
Pregnant Albino rats (>6 dams/group) were
exposed to 0, 4, 20, or 100 mg/kg-day BBP via
oral gavage from GD 14-21. Dams were allowed
to give birth naturally and male offspring were
sacrificed on PND 5. 25. or 75 (Ahmad et al..
2014).
NOAEL = 20
i serum testosterone, j absolute weight of
epididymis and prostate, j sperm count, j
percent motile sperm, f percent abnormal sperm
4.72
UFa= 3
UFh=10
Total UF=30
CD rats (n ~20/group) exposed via oral/diet to 0,
750, 3750, 11,250 ppm (eq. 0, 50, 250, 750 mg/kg-
dav) continuously for 2 generations (Tvl et al..
2004)
NOAEL = 50
I AGD (F1 and F2), cryptorchidism (F1 and F2)
11.8
UFa= 3
UFh=10
Total UF=30
SD rats were exposed via oral gavage from GD 8-
18 to 0, 100, 300, 600, 900 mg/kg-day BBP. Dams
were sacrificed and fetal tissue collected on GD 18
(Howdeshell et al.. 2008).
BMDL5 = 81
I ex vivo fetal testicular testosterone production
19.2
UFa= 3
UFh=10
Total UF=30
Pregnant Charles River SD rats (n = 3-4
dams/group) were exposed to 0, 100, 300, 600, or
900 mg/kg-day BBP via oral gavage from GD 14-
18. Dams were sacrificed and fetal tissue collected
on GD 18 (Grav et al.. 2021).
NOAEL = 100
I ex vivo fetal testicular testosterone production
23.6
UFa= 3
UFh=10
Total UF=30
Pregnant Harlan SD rats (n = 2-4/group) were
exposed to 0, 11, 33, or 100 mg/kg-day BBP via
oral gavage from GD 14-18. Dams were sacrificed
and fetal tissue collected on GD 18 (Block 37)
(Furret al.. 2014).
NOAEL = 100
I ex vivo fetal testicular testosterone production
23.6
UFa= 3
UFh=10
Total UF=30
Male and female SD rats (n = 20-24/group) dosed
(gavage) from 8-10 weeks of age with 0, 20, 100,
and 500 mg/kg-day BBP continuously for 2-
senerations (Nasao et al.. 2000).
NOAEL = 100
I AGD (Fl), I serum testosterone (Fl), J,
reproductive organ weights (Fl), testicular
pathological changes (Fl; e.g., 1 spermatocytes
in seminiferous tubules, atrophy of seminiferous
23.6
UFa= 3
UFh=10
Total UF=30
Page 51 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Brief Study Description
(Reference)
Study POD/ Type
(mg/kg-day)
Effect
HEP
(mg/kg)
Uncertainty
Factors"b
tubules, i germ cells in seminiferous tubule,
testicular edema, decreased sperm in epididymis)
Pregnant Harlan SD rats (n = 2-4/group) were
exposed to 0, 100, 300, 600, or 900 mg/kg-day
BBP via oral gavage from GD 14-18. Dams were
sacrificed and fetal tissue collected on GD 18
(Block 36) (Furr et al.. 2014).
LOAEL = 100
I ex vivo fetal testicular testosterone production
23.6
UFa= 3
UFh=10
ufl=io
Total UF=300
Cij:CD(SD)IGS rats (n = 24/dose) oral/gavage
with 0, 100, 200, 400 mg/kg-day BBP
continuously for 2 generations (Aso et al.. 2005).
LOAEL = 100
I AGD (F2); softening of testes; j spermatozoa
in epididymis; j germ cells in epididymal lumen
23.6
UFa= 3
UFh=10
ufl=io
Total UF=300
Pregnant Wistar rats (16 dams/group) were orally
gavaged with 0, 167, 250, 375 mg/kg-day MBP on
GD 15-17. Dams sacrificed and fetal tissue
collected on GD 21 (Ema et al.. 2003).
NOAEL = 167
I AGD, cryptorchidism
39.4
UFa= 3
UFh=10
Total UF=30
Pregnant Wistar rats (16 dams/group) were
exposed to 0, 250, 500, or 1000 mg/kg-day BBP
via gastric intubation from GD 15-17. Dams were
sacrificed and fetal tissue collected on GD 21
(Ema and Mivawaki. 2002).
NOAEL = 250
I AGD, I AGI, cryptorchidism, f
transabdominal testicular ascent
59
UFa= 3
UFh=10
Total UF=30
Pregnant SD rats (5-10 dams/group) were exposed
to 0 or 750 mg/kg-day BBP via oral gavage from
GD 14-PND 3. Dams were allowed to give birth
naturally and male offspring were sacrificed
between PND 2-mature adults (3-7 months of
ase) (Grav et al.. 2000).
LOAEL = 750
I AGD, | male NR, J, reproductive organ
weights, reproductive organ malformations
111
UFa= 3
UFh=10
ufl=io
Total UF=300
Pregnant SD rats (3-6 dams/group) were exposed
to 0 or750 mg/kg-day BBP via oral gavage from
GD 17-21. Dams were sacrificed and fetal tissue
collected on GD 21 (Soadc et al.. 2018).
LOAEL = 750
I ex vivo fetal testicular testosterone production,
| MNG incidence
111
UFa= 3
UFh=10
ufl=io
Total UF=300
Pregnant SD rats (3 dams/group) were exposed to
0 or 1000 mg/kg-day BBP via oral gavage from
GD 14-18. Dams were sacrificed and fetal tissue
collected on GD 18 (Wilson et al.. 2004).
LOAEL = 1000
I ex vivo fetal testicular testosterone production,
i Testicular Ins!3 mRNA expression
236
UFa= 3
UFh=10
ufl=io
Page 52 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Brief Study Description
(Reference)
Study POD/ Type
(mg/kg-day)
Effect
HEP
(mg/kg)
Uncertainty
Factors"b
Total UF=300
Abbreviations: j = Statistically significant decrease; t = Statistically significant increase; AGD = Anogenital distance; AGI = Anogenital index; BMDL5 =
Lower 95% confidence limit on benchmark dose; CD = Charles River Sprague-Dawley; GD = Gestation day; LOAEL = Lowest-observed-adverse-effect
level; MNGs = Multinucleated gonocytes; NOAEL = No-observed-adverse-effect level; NOEL = No-observed-effect level; NR = Nipple retention; PND =
Postnatal day; SD = Sprague-Dawley.
" EPA used allometric body weight scaling to the three-quarters power to derive the HED. Consistent with EPA Guidance (U.S. EPA. 2011c). 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 DCHP. 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.
c NOAEL of 11 mg/kg-day is limited to decreased fetal testicular expression of genes involved in steroidogenesis, including his/3 (which are effects not
considered adverse in isolation). Statistically significant adverse effects, particularly decreased ex vivo fetal testicular testosterone production reached
statistical significance at higher doses, resulting in a LOAEL = 300 mg/kg-day based upon ex vivo fetal testicular testosterone production.
1513
Page 53 of 122
-------
1514
1515
1516
1517
1518
1519
1520
1521
1522
1523
1524
1525
1526
1527
1528
1529
1530
1531
1532
1533
1534
1535
1536
1537
1538
1539
1540
1541
1542
1543
1544
1545
1546
1547
1548
1549
1550
1551
1552
1553
1554
1555
1556
1557
1558
1559
1560
1561
PUBLIC RELEASE DRAFT
DECEMBER 2024
4.2.1 Studies with Lack of Dose-Response Sensitivity and Increased Uncertainty
Of the 14 studies considered and presented in Table 4-1, six were not further evaluated for quantitative
dose response due to the reasons detailed below, including limitations of dose selection and effect level
sensitivity that was determined to increase uncertainties (Spade et al.. 2018; Wilson et al.. 2004; Ema et
al.. 2003; Ema and Miyawaki. 2002; Gray et al.. 2000; Nagao et al.. 2000).
Three studies (Spade et al.. 2018; Wilson et al.. 2004; Gray et al.. 2000) that only tested one relatively
high-dose-level of BBP effects on male reproductive outcomes, including fetal testicular testosterone
production, were not further considered for dose-response analysis. Gray et al. (2000) gavaged SD rats
with 0 or 750 mg/kg-day BBP from GD 14 to PND 3 and assessed offspring effects from PND 2 into
adulthood. In this assessment, authors noted decreased AGD, increased nipple retention, and various
reproductive organ malformations at 750 mg/kg-day BBP (Gray et al.. 2000). Spade et al. (2018)
gavaged SD rats with 750 mg/kg-day BBP and noted 69 percent decrease in ex vivo fetal testicular
testosterone production relative to control, as well as an increase in testicular MNG incidence. In SD rats
gavaged with 1000 mg/kg-day BBP, Wilson et al. (2004) reported 12 percent ex vivo fetal testicular
testosterone relative to control, which coincided with decrease testicular mRNA expression of InsI3.
However, experiments in each of these studies only tested one high dose level in addition to vehicle
controls, support LOAELs ranging from 750 to 1000 mg/kg-day BBP, which are therefore not sensitive.
Furthermore, these studies do not allow for the identification of a NOAEL, which increases the
uncertainty in the data sets use for POD derivation. Ultimately, these studies were not selected in dose-
response assessment because other developmental studies of BBP are available that test more than one
dose level and support identification of more sensitive NOAELs.
Two studies (Ema et al.. 2003; Ema and Miyawaki. 2002) were similarly not considered further for
dose-response analysis because other studies provide more sensitive NOAELs. Ema et al. (2003)
gavaged Wistar rats with 0, 167, 250, or 375 mg/kg-day MBP (a major BBP metabolite) from GD 15 to
17 and made observations on GD 21. Here, Ema et al. (2003) noted a NOAEL of 167 mg/kg-day MBP
based on occurrence of cryptorchidism and decreased AGD. Ema et al. (2002) exposed Wistar rats with
0, 250, 500, or 1000 mg/kg-day BBP via gastric intubation from GD 15 to GD 17 and also made
observations on GD 21. Ema et al. (2002) reported a NOAEL of 250 mg/kg-day BBP based upon
decreased AGD (and decreased anogenital index, a standardized AGD value corrected for body weight),
cryptorchidism, and increased transabdominal testicular ascent. However, the doses at which
developmental effects were observed in these studies (LOAELs of 250 mg/kg-day and 500 mg/kg-day
supporting NOAELs of 167 and 250 mg/kg-day, respectively) were higher than doses at which similar
outcomes and sensitive effects of androgen insufficiency (e.g., decreased fetal testicular testosterone)
were observed in other studies (Table 4-1). Therefore, EPA did not select these studies because they do
not provide the most sensitive robust endpoint for a POD relevant to these durations.
Nagao et al. (2000) gavaged SD rats from 8-10 weeks of age and continuously through two generations
with either 0, 20, 100, or 500 mg/kg-day BBP. Authors noted significantly decreased AGD in F1
offspring at 500 mg/kg-day, along with decreased reproductive organ weights and increased testicular
pathological changes (e.g., decreased spermatocytes in seminiferous tubules, atrophy of seminiferous
tubules, decreased seminiferous germ cells, testicular edema, and decreased sperm in epididymis) also
occurring at 500 mg/kg-day, resulting in a study LOAEL of 500 mg/kg-day and NOAEL of 100 mg/kg-
day (Nagao et al.. 2000). However, as discussed later (Section 4.2.3), multiple studies, including Tyl et
al. (2004). support a considerably lower consensus LOAEL and NOAEL (i.e., 100 mg/kg-day and 50
mg/kg-day respectively), providing more sensitive suggested effect levels of BBP exposure-related
effects on developing male reproductive outcomes. Additionally, the large dose range between the
Page 54 of 122
-------
1562
1563
1564
1565
1566
1567
1568
1569
1570
1571
1572
1573
1574
1575
1576
1577
1578
1579
1580
1581
1582
1583
1584
1585
1586
1587
1588
1589
1590
1591
1592
1593
1594
1595
1596
1597
PUBLIC RELEASE DRAFT
DECEMBER 2024
NOAEL of 100 mg/kg-day and LOAEL of 500 mg/kg-day represents an additional source of effect level
uncertainty. Considering these factors, Nagao et al. (2000) was not further selected for dose-response
assessment, as other studies identified more sensitive LOAELs and NOAELs.
4.2.2 Meta-analysis and BMP Modeling of Fetal Testicular Testosterone Data
Of the 14 studies considered and presented in Table 4-1,4 studies across 3 publications (Gray et al..
2021; Furr et al.. 2014; Howdeshell et al.. 2008) were explored for their dose-related relationship to ex
vivo fetal testicular testosterone production in accordance with prior NASEM (2017) analysis that used
BMD modeling for sensitive BBP POD derivation.
Four studies across 3 publications provided consistent evidence of dose-related reductions in ex vivo
fetal testicular testosterone production (Gray et al.. 2021; Furr et al.. 2014; Howdeshell et al.. 2008).
Study details, including BBP dose-response effects on percent decreased ex vivo fetal testicular
testosterone production, are shown in Table 4-2. It is notable that the magnitude of effect on ex vivo fetal
testicular testosterone production was consistent at similar doses across all 4 studies (Table 4-2). In the
lower dose range at 0 to 33 mg/kg-day BBP, Furr et al. (2014) identified no BBP exposure-related
effect. However, a dose-dependent decrease in ex vivo testicular testosterone production was noted at
BBP doses of 100 to 900 mg/kg-day across studies (Gray et al.. 2021; Furr et al.. 2014; Howdeshell et
al.. 2008). In highest tested dose of 900 mg/kg-day in these studies, there was a range of 85 to 93 percent
decrease in ex vivo testicular testosterone production relative to control groups for each study (Table
4-2). However, only one study identified a statistically significant effect on ex vivo fetal testicular
testosterone production at 100 mg/kg-day, with Furr et al. (2014) reporting 53 percent decreased ex vivo
fetal testicular testosterone production in Block 36. This contrasts with the other 3 dose-response studies
indicating no BBP-related effect on this endpoint at 100 mg/kg-day (Gray et al.. 2021; Furr et al.. 2014;
Howdeshell et al.. 2008). suggesting some uncertainty in the identified LOAEL of 100 mg/kg-day by
Furr et al. (2014) (Block 36) based upon fetal testicular testosterone effects. It should also be noted these
individual studies are generally limited by small sample sizes of only 2 to 3 dams per dose group for
most of these studies, except for Howdeshell et al. (2008). Additionally, Gray et al. (2021) reported
testicular mRNA expression changes in pertinent steroidogenic genes (as discussed in Section 3.1.2.1).
These mRNA changes suggested a no-observed-effect-level of 11 mg/kg-day (Table 4-1); however,
these gene effects are not considered adverse in isolation, where additional study effects of diminished
ex vivo fetal testicular testosterone production did not occur until a considerably higher level of 300
mg/kg-day (Table 4-2). Given that effects on ex vivo fetal testicular testosterone production are
considered as a sensitive indicator of BBP exposure and critical in phthalate-syndrome related MO A
(Figure 3-1), EPA reviewed previous benchmark dose (BMD) modeling of fetal testosterone production
effects to inform current POD selection (NASEM. 2017). discussed below.
Page 55 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
1598 Table 4-2. Effect of BBP Exposure on Fetal Testicular Testosterone Production"
Reference
Study Details
(Species, Duration, Exposure Route/
Method, Doses [mg/kg-day])
0 mg/kg-
day
11 mg/kg-
day
33 mg/kg-
day
100
mg/kg-
day
300
mg/kg-
day
600
mg/kg-
day
900
mg/kg-
day
(Furr et al.,
2014)
Harlan SD rats; GD 14-18; oral/gavage; 0, 100,
300, 600, 900 (Block 36)
100%
(n=3)
b
b
47%*
(n=3)
33%*
(n=3)
24%*
(n=3)
15%*
(n=3)
Harlan SD rats; GD 14-18; oral/gavage; 0, 11,
33, 100 (Block 37)
100%
(n=3)
111%
(n=3)
91%
(n=3)
88%
(n=3)
b
b
b
(Howdeshell
et al.. 2008)
SD rats; GD 8-18; oral gavage; 0, 100, 300,
600, 900
100%
(n=9)
b
b
106%
(n=4)
78%*
(n=5)
34%*
(n=5)
10%*
(n=5)
(Grav et al.,
202 l)c
Charles River SD rats; GD 14-18; oral/gavage;
0, 100, 300, 600, 900 (Block 78)
100%
(n=3)
b
b
107%
(n=3)
62%*
(n=3)
37%*
(n=2)
7%*
(n=2)
Abbreviations: *= Statistical significance; GD = Gestation day; SD = Sprague-Dawley.
"Effect on ex vivo fetal testicular testosterone production reported as percent of control. Asterisks indicate statistically significant pairwise comparison to control, as reported
by study authors.
bExposure level was not tested in study (or block) and thus is not shown for given study.
c Data from Block 78 rats reported in siiDDlcmcntal information file associated with Grav et al. (2021).
1599
Page 56 of 122
-------
1600
1601
1602
1603
1604
1605
1606
1607
1608
1609
1610
1611
1612
1613
1614
1615
1616
1617
1618
1619
1620
1621
1622
1623
1624
1625
1626
1627
1628
1629
1630
1631
1632
1633
PUBLIC RELEASE DRAFT
DECEMBER 2024
In the NASEM (2017) analysis, experimental animal model evidence for BBP in atero exposure
effects on fetal testicular testosterone was assessed using the systematic review methodology
developed by the National Toxicology Program's (NTP) Office of Health Assessment and
Translation (OHAT). NASEM concluded a high rating in the confidence in the body of evidence and
evidence of outcome that exposure to BBP during the gestational window of susceptibility decreased
fetal testicular testosterone production in rats. At the time, NASEM used the two available prenatal
BBP exposure rat studies (Furr et al.. 2014; Howdeshell et al.. 2008) to conduct a meta-regression
analysis and BMD modeling analysis on fetal testicular testosterone. NASEM found a statistically
significant overall effect in logio(dose) and dose, with an effect magnitude >50 percent (significant
heterogeneity in all cases, I2> 85%) (Table 4-3). The best-fit linear quadratic model estimated 23
mg/kg-day [95% CI: 13, 74] for a 5 percent change (BMDs) (benchmark response, BMR= -5.1) and
228 mg/kg-day [95% CI: 140, 389] for a 40 percent change (BMD40) (BMR =51) (Table 4-3).
Table 4-3. Summary of NASEM (2017) Meta-Analysis and BMD Modeling for Effects of BBP on
Fetal Testosterone"'b'c
Database
Supporting
Outcome
Confidence
in Evidence
Evidence of
Outcome
Heterogeneity
in Overall
Effect
Model with
Lowest AIC
BMD? mg/kg-
day (95% CI)
BMD40 mg/kg-
day (95% CI)
2 rat studies1'
High
High
I2 > 85%
Linear
Quadratic
23 (13, 74)
228 (140, 389)
Abbreviations: AIC = Akaike Information Criterion; BMD5 = Benchmark dose associated with a 5% change; BMD40 =
Benchmark dose associated with a 40% change; CI = Confidence interval.
" R code supporting NASEM's meta-regression and BMD analysis of BBP is publicly available through GitHub
(https ://eithub. com/wachiuohd/NASEM-2017 -Endocrine-Low-Dose).
b NASEM (2017) calculated a BMD40 for this endpoint because "previous studies have shown that reproductive-tract
malformations were seen in male rats when fetal testosterone production was reduced by about 40%."
"Taken from Table C6-4 of NASEM (2017).
^Studies in assessment are Furr et al. (2014) and Howdeshell et al. (2008).
Since EPA identified new fetal testicular testosterone dose-response data (Gray et al.. 2021) for BBP, 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. Fetal rat
testosterone data from three studies was included in the analysis (Gray et al.. 2021: Furr et al.. 2014:
Howdeshell et al.. 2008).Overall, the meta-analysis found a statistically significant overall effect and
linear trends in logio(dose) and dose, with an overall effect that is large in magnitude (>50% change)
(Table 4-4). There was substantial, statistically significant heterogeneity in all cases (I2>90%). The
statistical significance of these effects was robust to leaving out individual studies (Table 4-4). The
linear-quadratic model provided the best fit (based on lowest AIC) (Table 4-4). The BMD for a 40
percent change (BMD40) under the best-fit linear quadratic model was 284 mg/kg-day [95% CI: 150,
481] (Table 4-4). BMD estimates could not be generated for a 5 or 10 percent change (BMR = 5% or
10%) (Table 4-5). 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-
Page 57 of 122
-------
1634
1635
1636
1637
1638
1639
1640
1641
1642
1643
1644
1645
1646
1647
1648
1649
1650
1651
1652
1653
1654
1655
1656
1657
1658
1659
1660
1661
PUBLIC RELEASE DRAFT
DECEMBER 2024
ethylhexyl) Phthalate (DEHP), Dibutyl Phthalate (DBP), Butyl Benzyl Phthalate (BBP), Diisobutyl
Phthalate (DIBP), DicyclohexylPhthalate (DCHP), andDiisononylPhthalate (U.S. EPA. 2024c).
Although the meta- and BMD modeling analyses conducted by NASEM and EPA provided similar
results for BMD40 and benchmark dose (lower confidence limit) associated with a 40 percent response
level (BMDL40) estimates, analysis including updated available data did not allow for the derivation of a
BMDs for effects of BBP exposure on fetal testosterone production. EPA did not further consider the
BMDL40 estimate as a candidate for deriving a POD because, as described in Appendix E, the 40 percent
response level is not considered health protective, and other available studies of BBP provide more
sensitive PODs.
Because no benchmark dose (lower confidence limit) associated with a 5 percent response level
(BMDLs) could be derived via the updated meta-analysis and BMD analysis, EPA modelled individual
fetal testicular testosterone data from the three publications included in the updated meta-analysis (Gray
et al.. 2021; Furr et al.. 2014; Howdeshell et al.. 2008) using EPA's BMD Software (BMDS version
3.3.2). For this analysis, BMRs of 1 control standard deviation, and 5, 10, and 40 percent were modelled
using the full suite of standard continuous models provided in BMDS (Exponential, Hill, Polynomial,
Power, Linear). Further methodological details and BMD modeling results are presented in Appendix F.
Of the modelled data sets, adequate BMD model fits were only obtained for fetal testicular testosterone
data from Howdeshell et al. (2008). The best fitting exponential 3 model supports BMDs and BMDLs
values of 138 and 81 mg/kg-day, respectively, as shown in Table 4-1 (outputs are shown in Appendix
F.l). However, as discussed below in Section 4.2.3, the BMDLs = 81 mg/kg-day was found to be
slightly less sensitive than other identified NOAEL and LOAEL co-critical effect level studies.
Table 4-4. Overall Analyses of Rat Studies of BBP and Fetal Testosterone (Updated Analysis
CI,
CI,
P value for
Heterogeneity
Analysis
Estimate
Beta
Lower
Bound
Upper
Bound
P value
Tau
I2
AICs
Primary Analysis
Overall
intrcpt
-83.62
-127.17
-40.06
1.68E-04
83.98
98.20
4.78E-151
169.89
Trend in loglO(dose)
loglO(dose)
-120.36
-169.45
-71.28
1.54E-06
49.93
94.66
3.34E-36
149.12
Linear in doselOO
doselOO
-22.98
-30.32
-15.63
8.69E-10
69.12
97.13
7.81E-82
153.33
LinearQuadratic in doselOO
doselOO
-15.00
-36.40
6.40
1.70E-01
50.89
93.85
8.24E-53
140.94*
LinearQuadratic in doselOO
I(dosel00A2)
-1.04
-3.78
1.69
4.54E-01
50.89
93.85
8.24E-53
140.94
Sensitivity Analysis
Overall minus
(Furret al.. 2014)
intrcpt
-90.83
-160.08
-21.59
1.01E-02
97.63
97.87
2.72E-33
91.46
Overall minus
(Gray et al.. 2021)
intrcpt
-78.47
-125.70
-31.24
1.13E-03
77.72
98.17
5.38E-125
122.09
Overall minus
(Howdeshell et al.. 2008)
intrcpt
-84.05
-134.86
-33.24
1.19E-03
84.27
98.27
8.30E-102
123.25
Abbreviations: AIC = Akaike Information Criterion; CI = Confidence interval.
* Indicates lowest AIC.
Table 4-5. Benchmark Dose Estimates for BBP and Fetal Testosterone in Rats
Analysis
BMR
BMD
CI, Lower Bound
CI, Upper Bound
2017 NASEM Analysis using Metafor Version 2.0.0
Page 58 of 122
-------
1662
1663
1664
1665
1666
1667
1668
1669
1670
1671
1672
1673
1674
1675
1676
1677
1678
1679
1680
1681
1682
1683
1684
1685
1686
1687
1688
1689
1690
1691
1692
PUBLIC RELEASE DRAFT
DECEMBER 2024
Analysis
BMR
BMD
CI, Lower Bound
CI, Upper Bound
fas reported in Table C6-4 of NASEM (2017))
Linear in doselOO
5%
23
19
29
Linear in doselOO
40%
231
192
290
LinearQuadratic in doselOO*
5%
23
13
74
LinearQuadratic in doselOO*
40%
228
140
389
Updated Analysis using Metafor Version 4.6.0
Linear in doselOO
5%
22
17
33
Linear in doselOO
10%
46
35
67
Linear in doselOO
40%
222
168
327
LinearQuadratic in doselOO*
5%
NA "
NA "
236
LinearQuadratic in doselOO*
10%
NA "
NA "
280
LinearQuadratic in doselOO*
40%
284
150
481
Abbreviations: BMD = Benchmark dose; BMR = Benchmark response; CI = Confidence interval.
* Indicates model with lowest AIC.
" BMD and BMDL estimates could not be derived.
4.2.3 Co-critical Studies Supporting a Consensus LOAEL of 100 mg/kg-day and NOAEL
of 50 mg/kg-day
Of the 14 studies considered and presented in Table 4-1,4 studies were considered co-critical in
dose-response assessment of BBP, supporting either a consensus LOAEL of 100 mg/kg-day (Ahmad
et al.. 2014; Furr et al.. 2014; Aso et al.. 2005) or NOAEL of 50 mg/kg-day (Tyl et al.. 2004).
In the study by Ahmad et al. (2014) albino rats were gavaged with 0, 4, 20, or 100 mg/kg-day BBP from
GD 14 to 21, and male offspring were evaluated on PND 1, 5, 25, or 75. A suggested LOAEL of 4
mg/kg-day was identified by authors based on significantly decreased offspring body weight on PND 1
and PND 21. However, EPA considers the results in the study to support a LOAEL of 100 mg/kg-day
and NOAEL of 20 mg/kg-day considering that the decreases in body weight at PND 1, though
statistically significant, were minor (3-4% decrease) in addition to not exhibiting a strong dose-response
{i.e., not likely to be biologically significant). The body weight decreases at PND 21 were greater in
magnitude (13-22% decrease compared to controls) but were also not dose-dependent. Furthermore, the
effect of statistically decreased body weights at 4, 20, or 100 mg/kg-day at PND 1 was not accompanied
by effects on other examined developmental landmarks (e.g., pinnae unfolding, eye opening, fur
development, or testes descent), and offspring body weight changes showed substantial recovery when
measured in adulthood at PND 75. Finally, only at 100 mg/kg-day were consistent effects on other
endpoints observed, including decreased prostate and epididymis weights, decreased serum testosterone
levels, decreased sperm count and motility, and increased sperm abnormalities. It was also noted that no
statistical methods were used to account for litter effects of this study (i.e., statistics on offspring were
presented as means of individual animal rather than litter means). Regarding consistency of results from
Ahmad et al. (2014). decreased offspring body weight gain is an effect that does not occur in other
studies until higher dose levels, such as at 500 mg/kg-day in Nagao et al. (2000) and 750 mg/kg-day in
Tyl et al. (2004) (Table 4-1). Overall, the effects on body weight in the study by Ahmad et al. (2014)
had too much uncertainty in results at low doses tested to be considered for POD derivation using the
identified NOAEL of 20 mg/kg-day, including the fact that body weight changes supporting the NOAEL
in this study were transient. However, results at the LOAEL of 100 mg/kg-day had more confidence and
were consistent with other consensus studies showing developing male reproductive effects at the same
level (Furr et al.. 2014; Aso et al.. 2005).
Page 59 of 122
-------
1693
1694
1695
1696
1697
1698
1699
1700
1701
1702
1703
1704
1705
1706
1707
1708
1709
1710
1711
1712
1713
1714
1715
1716
1717
1718
1719
1720
1721
1722
1723
1724
1725
1726
1727
1728
1729
1730
1731
1732
1733
1734
1735
1736
1737
1738
1739
1740
1741
PUBLIC RELEASE DRAFT
DECEMBER 2024
Furr et al. (2014) (Block 36), gavaged Harlan SD rats with 0, 100, 300, 600, 900 mg/kg-day BBP from
GD 14 to 18 and evaluated ex vivo testicular testosterone production on GD 18. Significantly decreased
ex vivo testicular testosterone production was found in all treatment groups, which decreased in a dose-
dependent fashion. Rats exposed to 100 mg/kg-day BBP showed a 53 percent decrease in ex vivo
testicular testosterone production, and at the highest tested dose of 900 mg/kg-day was further decreased
to an 85 percent reduction. No NOAEL was identified in this study.
Two multi-generational studies were determined to provide sensitive effect levels, including a
LOAEL of 100 mg/kg-day in the study by Aso et al. (2005) and a NOAEL of 50 mg/kg-day in the
study by Tyl et al. (2004) (Table 4-1). Both studies were two-generation studies that assessed BBP
gestational exposure effects on AGD and associated male reproductive organ abnormalities and
histopathological outcomes. As discussed in Section 3.1.2.1, decreased AGD is highly correlated with
in utero anti-androgenic activity and is considered one of the most sensitive biomarkers for phthalate
effects on the developing male reproductive system (Schwartz et al.. 2019).
Aso et al. (2005) gavaged SD rats with 0, 100, 200, or 400 mg/kg-day BBP continuously for two
generations. Although most developmental male reproductive effects occurred at levels above 100
mg/kg-day (e,g., incomplete preputial separation, decreased epididymal weight, small testes, and
Ley dig cell hyperplasia at 400 mg/kg-day), a study-wide LOAEL of 100 mg/kg-day was determined
based upon decreased AGD noted in F2 male offspring and histopathology findings in the testes (e.g.,
softening of testes; decreased spermatozoa in epididymis, and decreased germ cells in epididymal
lumen). However, 100 mg/kg-day BBP was the lowest dose tested, and thus the study by Aso et al.
(2005) did not derive a NOAEL, increasing uncertainty associated with the use of this study alone for
a POD based on the LOAEL with no NOAEL established.
Tyl et al. (2004) conducted a two-generation diet exposure study in CD rats exposed to 0, 50, 250, or
750 mg/kg-day BBP. Tyl et al. (2004) identified a LOAEL of 250 mg/kg-day based on incidences of
cryptorchidism and decreased AGD in both F1 and F2 offspring, which provides a more sensitive
NOAEL (50 mg/kg-day) than the NOAEL of 100 mg/kg-day observed in the two-generation study by
Nagao et al. (2000). It should be noted additional and more extensive anti-androgenic effects
(increased nipple retention, gross testicular/epididymal histopathology changes, and cryptorchidism)
occurred at the higher dose of 750 mg/kg-day BBP. In sum, Tyl et al. (2004) identified the most
sensitive NOAEL of 50 mg/kg-day BBP.
Using a NOAEL/LOAEL approach in the studies of Table 4-1 considered for BBP POD derivation,
EPA selected a NOAEL of 50 mg/kg-day for the BBP POD identified in Tyl et al. (2004). with
supporting co-critical studies suggesting a consensus LOAEL of 100 mg/kg-day based upon
associated cluster of anti androgenic outcomes, including reduced ex vivo testicular testosterone
production and testicular histopathological changes (Ahmad et al.. 2014; Furr et al.. 2014; Aso et al..
2005). Although BBP effects on ex vivo fetal testicular testosterone production has been used in prior
assessments for BMD modeling and effect level estimates (NASEM. 2017). updated meta-analysis
modeling data by the EPA could not derive a BMDs (Table 4-5). Of individually modelled ex vivo
fetal testicular testosterone production data sets, an adequate BMD model fit was obtained for
Howdeshell et al. (2008) at a BMDLs value of 81 mg/kg-day (Appendix F.l), which was less
sensitive than the NOAEL of 50 mg/kg-day identified in Tyl et al. (2004). While there were
inconsistencies in the dose at which fetal testosterone production was decreased across the studies,
the lowest LOAEL for decreased fetal testosterone was 100 mg/kg-day, identified in Block 36 in the
study by Furr et al. (2014) (Table 4-2). Aso et al. (2005) identified a LOAEL of 100 mg/kg-day based
upon decreased AGD and slight reproductive organ histopathological effects. Ahmad et al. (2014)
Page 60 of 122
-------
1742
1743
1744
1745
1746
1747
1748
1749
1750
1751
1752
1753
1754
1755
1756
1757
1758
1759
1760
1761
1762
1763
1764
1765
1766
1767
1768
1769
1770
1771
1772
1773
1774
1775
1776
1777
1778
1779
1780
1781
1782
1783
1784
1785
1786
1787
PUBLIC RELEASE DRAFT
DECEMBER 2024
also identified LOAEL of 100 mg/kg-day and was found to have substantial limitations and
uncertainty at the low doses tested (4 and 20 mg/kg-day). Further, using the lowest LOAEL of 100
mg/kg-day instead of a NOAEL would create a considerably larger UF and lower confidence in risk
characterization due to 100 mg/kg-day being the lowest study dose tested across Furr et al. (2014).
Aso et al. (2005). and Ahmad et al. (2014). This would require a LOAEL-to-NOAEL (UFl) of 10,
which would make the benchmark MOE 300 as opposed to a benchmark MOE of 30 by using a
NOAEL of 50 mg/kg-day from Tyl et al. (2004). In summary, EPA considers these 4 co-critical
studies (Ahmad et al.. 2014; Furr et al.. 2014; Aso et al.. 2005; Tyl et al.. 2004) to support a NOAEL
of 50 mg/kg-day and a LOAEL of 100 mg/kg-day based upon decreased ex vivo fetal testicular
testosterone production, decreased AGD, and slight testicular histopathology. EPA selected a
NOAEL of 50 mg/kg-day (HED = 12 mg/kg-day) as the POD for assessing risks from acute,
intermediate, and chronic durations of exposure. A total uncertainty factor of 30 was selected for use
as the benchmark MOE (based on an interspecies uncertainty factor [UFa] of 3 and an intraspecies
uncertainty factor [UFh] of 10). EPA's POD selection of a NOAEL of 50 mg/kg-day BBP is in
accordance with the POD selection of multiple existing assessments conducted by other regulatory
authorities, including U.S. CPSC (2014), Health Canada (ECCC/HC. 2020), ECHA (2017a).
NICNAS (2015). and EFSA (2019). where the majority of these assessments also explicitly used Tyl
et al. (2004) and Aso et al. (2005) as critical dose-response studies (see Table 1-1).
4.3 Weight of the Scientific Evidence
EPA considers 4 co-critical studies that support a consensus LOAEL of 100 mg/kg-day (Ahmad et al..
2014; Furr et al.. 2014; Aso et al.. 2005) or NOAEL of 50 mg/kg-day (Tyl et al.. 2004). based on effects
on the developing male reproductive system consistent with a disruption of androgen action and
phthalate syndrome. EPA has preliminarily concluded that the HED of 12 mg/kg-day (NOAEL of 50
mg/kg-day) based on decreased AGD and associated anti-androgenic effects (e.g., decreased fetal
testicular testosterone production) observed in gestational BBP exposure studies in rats. This is assumed
appropriate for calculation of risk from acute, intermediate, and chronic durations. A total UF of 30 was
selected for use as the benchmark MOE based upon an interspecies UF (UFa) of 3, intraspecies UF
(UFh) of 10, and LOAEL-to-NOAEL UF (UFl) of 1. It should be noted that due to the strength of
existing studies and amount of data available for other phthalate non-cancer hazards, EPA did not deem
it necessary to apply a database UF (UFd). This is because for BBP, there exists multiple dose-response
developmental and multi-generational studies. Further for toxicologically similar phthalates (i.e., DEHP,
DBP, DCHP), there exists larger databases of animal toxicology studies including numerous well-
conducted subchronic and chronic toxicity studies identifying effects on the developing male
reproductive system consistent with a disruption of androgen action, which has been identified by EPA
as the most sensitive and well-characterized hazard in experimental animal models. Because a robust
database of studies that have evaluated BBP for effects on the developing male reproductive system are
available, EPA is confident that the selected POD for BBP is health protective and that a UFd is not
warranted. Consistent with EPA guidance (2022. 2002b. 1993). EPA reduced the UFa from a value of
10 to 3 because allometric body weight scaling to the three-quarter power was used to adjust the POD to
obtain a HED (Appendix D). EPA has robust overall confidence in the proposed POD based on the
following weight of the scientific evidence:
EPA has previously considered the weight of scientific evidence and concluded that oral
exposure to BBP 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
Page 61 of 122
-------
1788
1789
1790
1791
1792
1793
1794
1795
1796
1797
1798
1799
1800
1801
1802
1803
1804
1805
1806
1807
1808
1809
1810
1811
1812
1813
1814
1815
1816
1817
1818
1819
1820
1821
1822
1823
1824
1825
1826
1827
1828
1829
1830
1831
1832
PUBLIC RELEASE DRAFT
DECEMBER 2024
Substances Control Act (U.S. EPA. 2023 a)). Notably, EPA's conclusion was supported by the
SACC (U.S. EPA. 2023b).
BBP exposure resulted in treatment-related effects on the developing male reproductive system
consistent with a disruption of androgen action and phthalate syndrome-related outcomes during
the critical window of development in 14 studies in rats (Section 3.1.2). Observed effects in male
offspring included: altered testicular mRNA expression of lipid metabolism and steriodogenic
synthesis genes; reduced fetal testicular testosterone content and/or testosterone production;
reduced AGD; nipple retention; reproductive tract malformations (e.g., hypospadias,
cryptorchidism, decreased reproductive tissue weights); delayed preputial separation; testicular
pathology (e.g., degeneration of seminiferous tubules, testes softening, Ley dig cell hyperplasia,
edema, multinucleated gonocytes); decreases epididymal and seminiferous tubule germ cells;
decreased sperm concentration and motility.
In human epidemiological studies (discussed in 3.1.1), authors such as Radke et al. (2018) found
that there was a slight level of confidence in the association between exposure to BBP and AGD,
as well as cryptorchidism/hypospadias; however, this association was not consistent with the
findings from Health Canada (2018b) or NASEM (2017). leading the EPA to conclude causality
was not established preliminarily due to uncertainty associated with exposure characterization of
individual phthalates. including source or exposure and timing of exposure as well as co-
exposure confounding with other phthalates.
There was some alignment across epidemiological, animal toxicology, and mechanistic streams
of evidence. The epidemiological evidence provides qualitative support for the association
between BBP exposure and male reproductive outcomes, including AGD.
EPA's proposed POD of 50 mg/kg-day (HED of 12 mg/kg-day) is consistent with other
regulatory and authoritative bodies that have also concluded developmental male reproductive
toxicity and anti-androgenic effects (i.e., decreased AGD, reduced ex vivo testicular testosterone
production, testicular pathology) is a sensitive indicator of BBP exposure and relevant for
estimating human risk (ECCC/HC. 2020; EFSA. 2019; ECHA. 2017a; NICNAS. 2015; U.S.
CPSC. 2014). These assessments have used a similar POD of NOAEL = 50 mg/kg-day to
quantify risk from BBP exposure Table 1-1.
The multi-generational study critical in POD derivation (Tyl et al.. 2004) was reported to adhere
to good laboratory practice guidelines of reproduction toxicity studies, specifically Health effects
test guidelines OPPTS 870.3800 reproduction andfertility effects (U.S. EPA. 1998).
EPA considered effects on the developing male reproductive systemic consistent with the
disruption of androgen action to be relevant for POD selection for acute, intermediate, and
chronic exposure durations, where BBP exposure during gestationally susceptible windows in
rats can elicit a spectrum of anti-androgenic effects related to the phthalate syndrome MOA.
EPA did not identify any studies conducted via the dermal route relevant for extrapolating human health
risk. Therefore, EPA is using the oral HED of 12 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 Butyl benzyl phthalate (BBP)
(U.S. EPA. 2025e).
EPA did not identify any inhalation studies of BBP. Therefore, EPA is also using the oral HED of 12
mg/kg-day to extrapolate to the inhalation route. EPA assumes similar absorption for the oral and
inhalation routes, and no adjustment was made when extrapolating to the inhalation route. For the
Page 62 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
1833 inhalation route, EPA extrapolated the daily oral HEDs to inhalation HECs using a human body weight
1834 and breathing rate relevant to a continuous exposure of an individual at rest. Appendix D provides
1835 further information on extrapolation of inhalation HECs from oral HEDs.
1836
Page 63 of 122
-------
1837
1838
1839
1840
1841
1842
1843
1844
1845
1846
1847
1848
1849
1850
1851
1852
1853
1854
1855
1856
1857
1858
1859
1860
1861
1862
1863
1864
1865
1866
1867
1868
1869
1870
1871
1872
1873
1874
1875
1876
1877
1878
1879
1880
PUBLIC RELEASE DRAFT
DECEMBER 2024
5 CONSIDERATION OF PESS AND AGGEGRATE EXPOSURE
5.1 Hazard Considerations for Aggregate Exposure
For use in the risk evaluation and assessing risks from other exposure routes, EPA conducted route-to-
route extrapolation of the toxicity values from the oral studies for use in the dermal and inhalation
exposure routes and scenarios. Health outcomes that serve as the basis for acute, intermediate, and
chronic hazard values are systemic and assumed to be consistent across routes of exposure. EPA
therefore concludes that for consideration of aggregate exposures, it is reasonable to assume that
exposures and risks across oral, dermal, and inhalation routes may be additive for the selected PODs in
Section 6.
5.2 PESS Based on Greater Susceptibility
In this section, EPA addresses subpopulations likely to be more susceptible to BBP 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 BBP.
As summarized in Table 5-1, EPA identified a range of factors that may have the potential to increase
biological susceptibility to BBP, 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 BBP is not
known; therefore, EPA is uncertain about the direction and magnitude of any possible increased risk
from effects associated with BBP exposure for relevant subpopulations.
Animal studies demonstrating effects on male reproductive development (as discussed in Section
3.1.2.1) and other developmental outcomes (as discussed in Section 3.1.2.3) provide direct evidence that
gestation is a particularly sensitive lifestage. Evidence from animal studies also demonstrates that the
liver and kidneys may be sensitive target organs although the liver and kidney effects observed across
reasonably available studies are generally not indicative of an adverse response. EPA is quantifying
risks, including those for PESS, based on developmental toxicity in the draft BBP risk evaluation.
As discussed throughout Section 3.1.1, EPA concluded epidemiological studies qualitatively contribute
to the weight of scientific evidence of demonstrating BBP toxicity effects on male reproductive
development and other developmental outcomes. Although there is uncertainty in the exposure
characterizations of epidemiological evidence such that it cannot be used as quantitative or direct
evidence (i.e., source of exposure, timing of exposure, co-exposures), EPA is acknowledging
epidemiological evidence that provide support to animal studies, including PESS considerations.
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 BBP. The Risk Assessment Forum, in
A Review of the Reference Dose and Reference Concentration Processes (U.S. EPA. 2002b). discusses
some of the evidence for choosing the default factor of 10 when data are lacking and describe the types
of populations that may be more susceptible, including different lifestages (e.g., of children and elderly).
Although U.S. EPA (2002b) did not discuss all the factors presented in Table , EPA considers the POD
proposed for use in characterizing risk from exposure to BBP to be protective of effects on the
developing male reproductive system consistent with phthalate syndrome in humans.
Page 64 of 122
-------
1881
PUBLIC RELEASE DRAFT
DECEMBER 2024
Table 5-1. PESS Evidence Crosswalk for Biological Susceptibility Considerations
Susceptibility
Category
Examples of
Specific Factors
Direct Evidence this Factor
Modifies Susceptibility to BBP
Description of Interaction
Key Citations
Indirect Evidence of Interaction with Target Organs
or Biological Pathways Relevant to BBP
Description of Interaction
Key Citation(s)
Susceptibility Addressed in Risk
Evaluation?
Embryos/
fetuses/infants
Direct quantitative animal evidence
for developmental toxicity (e.g.,
decreased live births, decreased
offspring body weight gain, and
decreased offspring survival with
increased severity in the second
generation) for BBP.
There is direct quantitative animal
evidence for effects on the developing
male reproductive system consistent
with a disruption of androgen action
(e.g., decreased ex vivo fetal testicular
testosterone production, decreased
anogenital distance, testicular
histopathology).
(U.S. EPA.
2023a):
(U.S. EPA.
2023b)
There is epidemiological evidence for
in utero exposure effects on
developing male reproductive system
(e.g., reduced anogenital distance,
hypospadias/cryptorchidism).
(Radke et al..
2018)
POD proposed for assessing risks
from acute, intermediate, and
chronic exposures to BBP is based
on developmental toxicity (i.e.,
phthalate syndrome-related
effects) and is protective of effects
on the fetus and offspring.
Lifestage
Pregnancy/
lactating status
Rodent dams not particularly
susceptible during pregnancy and
lactation, except for effects related to
reduced maternal weight gain, food
consumption, decreased ovarian and
uterine weights, and increased kidney
weight, which occurred at doses
higher than those that caused
developmental toxicity.
(Nagao et al..
2000): (Aso et
al.. 2005): (Tvl et
al.. 2004)
POD proposed for assessing risks
from acute, intermediate, and
chronic exposures to BBP is based
on developmental toxicity (i.e.,
phthalate syndrome-related
effects) and is protective of effects
in dams.
Males of
reproductive age
Reduced body weight gain, increased
liver and kidney and decreased lung
and left epididymal weights, and
decreased serum testosterone. Effects
observed at higher doses than those
that caused developmental toxicity.
(Nagao et al..
2000): (Aso et
al.. 2005): (Tvl et
al.. 2004)
There is epidemiological evidence for
adult male reproductive effects,
including reduced male fecundability
(e.g., impacted semen parameters,
time to pregnancy)
(Radke et al..
2018)
POD proposed for assessing risks
from acute, intermediate, and
chronic exposures to BBP is based
on developmental toxicity (i.e.,
phthalate syndrome-related
effects) and is protective of adult
male reproductive effects.
Use of default lOx UFH
Page 65 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Susceptibility
Category
Examples of
Specific Factors
Direct Evidence this Factor
Modifies Susceptibility to BBP
Indirect Evidence of Interaction with Target Organs
or Biological Pathways Relevant to BBP
Susceptibility Addressed in Risk
Evaluation?
Description of Interaction
Key Citations
Description of Interaction
Key Citation(s)
Children
Reduced rodent offspring bodyweight
gain between PNDs 1 to 21 was
observed in one and two-generation
studies of reproduction.
(Naeao et al..
2000); (Aso et
al.. 2005): (Tvl et
al.. 2004)
POD proposed for assessing risks
from acute, intermediate, and
chronic exposures to BBP is based
on developmental toxicity (i.e.,
phthalate syndrome-related
effects) and is protective of effects
of offspring bodyweight gain
Use of default lOx UFH
Elderly
No direct evidence identified
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. 2023e):
(CDC. 2023 e)
Use of default lOx UFH
Pre-existing
disease or
disorder
Individuals with chronic liver and
kidney disease may be more
susceptible to effects on these target
organs.
Viruses such as viral hepatitis can
cause liver damage.
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 BBP.
Use of default lOx UFH
Lifestyle
activities
Smoking
No direct evidence identified
Smoking during pregnancy may
increase susceptibility for
developmental outcomes (e.g., early
delivery and stillbirths).
(CDC. 2023f)
Qualitative discussion in Section
5.2 and this table
Page 66 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Susceptibility
Category
Examples of
Specific Factors
Direct Evidence this Factor
Modifies Susceptibility to BBP
Indirect Evidence of Interaction with Target Organs
or Biological Pathways Relevant to BBP
Susceptibility Addressed in Risk
Evaluation?
Description of Interaction
Key Citations
Description of Interaction
Key Citation(s)
Alcohol
consumption
No direct evidence identified
Alcohol use during pregnancy can
cause developmental outcomes (e.g.,
fetal alcohol spectrum disorders).
Heavy alcohol use may affect
susceptibility to liver disease.
(CDC. 2023d):
(CDC. 2023a)
Qualitative discussion in Section
5.2 and this table
Physical activity
No direct evidence identified
Insufficient activity may increase
susceptibility to multiple health
outcomes.
Overly strenuous activity may also
increase susceptibility.
(CDC. 2022)
Qualitative discussion in Section
5.2 and this table
Sociodemo-
Race/ethnicity
No direct evidence identified (e.g., no
information on polymorphisms in
BBP metabolic pathways or diseases
associated race/ethnicity that would
lead to increased susceptibility to
effects of BBP by any individual
group).
Qualitative discussion in Section
5.2 and this table
graphic status
Socioeconomic
status
No direct evidence identified
Individuals with lower incomes may
have worse health outcomes due to
social needs that are not met,
environmental concerns, and barriers
to health care access.
(ODPHP.
2023b)
Sex/gender
No direct evidence identified
Use of default lOx UFH
Nutrition
Diet
No direct evidence identified
Poor diets can lead to chronic
illnesses such as heart disease, type 2
diabetes, and obesity, which may
contribute to adverse developmental
outcomes. Additionally, diet can be a
risk factor for fatty liver, which could
be a pre-existing condition to enhance
susceptibility to BBP-induced liver
toxicity.
(CDC. 2023e):
(CDC. 2023b)
Qualitative discussion in Section
5.2 and this table
Page 67 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Susceptibility
Category
Examples of
Specific Factors
Direct Evidence this Factor
Modifies Susceptibility to BBP
Indirect Evidence of Interaction with Target Organs
or Biological Pathways Relevant to BBP
Susceptibility Addressed in Risk
Evaluation?
Description of Interaction
Key Citations
Description of Interaction
Key Citation(s)
Malnutrition
No direct evidence identified
Micronutrient malnutrition can lead
to multiple conditions that include
birth defects, maternal and infant
deaths, preterm birth, low birth
weight, poor fetal growth, childhood
blindness, undeveloped cognitive
ability.
Thus, malnutrition may increase
susceptibility to some developmental
outcomes associated with BBP.
(CDC. 2021);
(CDC. 2023b)
Qualitative discussion in Section
5.2 and this table
Genetics/
epigenetics
Target organs
No direct evidence identified
Polymorphisms in genes may
increase susceptibility to liver,
kidney, or developmental toxicity.
Use of default lOx UFH
Toxicokinetics
No direct evidence identified
Polymorphisms in genes encoding
enzymes (e.g., esterases) involved in
metabolism of BBP may influence
metabolism and excretion of BBP.
Use of default lOx UFH
Other
chemical and
nonchemical
stressors
Built
environment
No direct evidence identified
Poor-quality housing is associated
with a variety of negative health
outcomes.
(ODPHP.
2023a)
Qualitative discussion in Section
5.2 and this table
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. 2023c):
(ODPHP.
2023c)
Qualitative discussion in Section
5.2 and this table
Page 68 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Susceptibility
Category
Examples of
Specific Factors
Direct Evidence this Factor
Modifies Susceptibility to BBP
Indirect Evidence of Interaction with Target Organs
or Biological Pathways Relevant to BBP
Susceptibility Addressed in Risk
Evaluation?
Description of Interaction
Key Citations
Description of Interaction
Key Citation(s)
Chemical co-
exposures
Studies have demonstrated that co-
exposure to BBP and other
toxicologically similar phthalates
(e.g., DEHP, DBP, DINP) and other
classes of antiandrogenic chemicals
(e.g., certain pesticides and
pharmaceuticals - discussed more in
(U.S. EPA. 2023a)) can induce
effects on the developing male
reproductive system in a dose-
additive manner.
(U.S. EPA.
2023a): (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.
1882
Page 69 of 122
-------
1883
1884
1885
1886
1887
1888
1889
1890
1891
1892
1893
1894
1895
1896
1897
1898
1899
1900
1901
1902
PUBLIC RELEASE DRAFT
DECEMBER 2024
6 POINTS OF DEPARTURE USED TO ESTIMATE RISKS FROM
BBP EXPOSURE, CONCLUSIONS, AND NEXT STEPS
EPA evaluated the non-cancer hazards of BBP and identified effects on the developing male
reproductive system as the most sensitive health effects indicator of BBP exposure. Anti-androgenic
effects from exposure occurring during the critical window of male reproductive system development
(e.g., gestational exposure) leads to a spectrum of adverse developing male reproductive system
outcomes, consistent with proposed phthalate syndrome pathways. After considering hazard
identification and evidence integration, dose-response evaluation, and weight of the scientific evidence
of POD candidates, EPA chose one non-cancer endpoints for the risk evaluation in acute, intermediate,
and chronic exposure scenarios (Table 6-1). EPA considers the proposed non-cancer POD (NOAEL of
50 mg/kg-day; HED = 12 mg/kg-day) protective of non-cancer developmental toxicity effects. 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 64.2 mg/m3(5.03 ppm). EPA is also using the
oral HED to extrapolate to the dermal route.
Table 6-1. Non-cancer HECs and HEDs Used to Estimate Risks
Exposure
Scenario
Target
Organ
Species
Duration
POD
(mg/kg-
Effect
HEC"
(mg/m3
HED"
(mg/
Benchmar
kMOE*
References'
System
day)
[ppm]
kg-day)
Acute,
Developing
Rat
Multi-
NOAEL
Phthalate
64.2
12
II
<
IJ-H
&
(Ahmad et
Intermediate,
male
generational
= 50
syndrome-
[5.03]
UFH=10
al.. 2014;
Chronic
reproductive
toxicity
or 5-8 days
during
gestation
related effects
(e.g., |AGD; |
fetal testicular
testosterone; |
reproductive
organ weights;
Leydig cell
effects; {
mRNA and/or
protein
expression of
steroidogenic
genes; J.INSL3)
Total UF=
30
Furr et al..
2014; Aso
et al.. 2005;
Tvl et al..
2004)
Abbreviations: AGD = Anogenital distance; HEC = Human equivalent concentration; HED = Human equivalent dose; INSL3:
Insulin-like 3; MOE = Margin of exposure; NOAEL = No-observed-adverse-effect level; POD = Point of departure; UF =
Uncertainty factor
a HED and HEC values were calculated based on the most sensitive NOAEL of 50 mg/kg-day.
h EPA used allometric body weight scaling to the three-quarters power to derive the HED. Consistent with EPA Guidance
(U.S. EPA. 2011c). 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.
c Tyl et al. (2004) support a NOAEL of 50 mg/kg-day based on decreased AGD and decreased reproductive organ weights in
a multi-generational study at 250 mg/kg-day (LOAEL); the remaining effects listed reached statistical significance at higher
doses (most of which are not considered adverse in isolation). Ahmad et al. (2014). Furr et al. (2014). and Aso et al. (2005)
reflect supporting phthalate syndrome-related effects (e.g., reduced ex vivo testicular testosterone production or testicular
histopathological changes) at LOAEL = 100 mg/kg-day.
The proposed POD of 50 mg/kg-day (HED = 12 mg/kg-day) will be used in the Draft Risk Evaluation
for Butyl Benzyl Phthalate (BBP) (U.S. EPA. 2025f) to estimate acute, intermediate, and chronic non-
Page 70 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
1903 cancer risk. EPA summarizes the cancer hazards of BBP in a separate technical support document, Draft
1904 Cancer Raman Health Hazard Assessment for Di(2-ethylhexyl) Phthalate (DEHP), Dibutyl Phthalate
1905 (DBP), Diisobutyl Phthalate (DIBP), Butyl Benzyl Phthalate (BBP) andDicyclohexyl Phthalate
1906 (I)CHI1) (U.S. EPA. 2025a).
1907
1908 EPA is soliciting comments from the SACC and the public on the non-cancer hazard identification,
1909 dose-response and weight of evidence analyses, and the proposed POD for use in non-cancer risk
1910 characterization of BBP.
Page 71 of 122
-------
1911
1912
1913
1914
1915
1916
1917
1918
1919
1920
1921
1922
1923
1924
1925
1926
1927
1928
1929
1930
1931
1932
1933
1934
1935
1936
1937
1938
1939
1940
1941
1942
1943
1944
1945
1946
1947
1948
1949
1950
1951
1952
1953
1954
1955
1956
1957
PUBLIC RELEASE DRAFT
DECEMBER 2024
REFERENCES
Ahmad. R; Gautam. AK; Venn a. Y; Sedha. S; Kumar. S. (2014). Effects of in utero di-butyl phthalate
and butyl benzyl phthalate exposure on offspring development and male reproduction of rat.
Environ Sci Pollut Res Int 21: 3156-3165. http://dx.doi.org/10.1007/sl 1356-013-2281-x
Ahmad. R; Verma. Y; Gautam. A: Kumar. S. (2015). Assessment of estrogenic potential of di-n-butyl
phthalate and butyl benzyl phthalate in vivo. Toxicol Ind Health 31: 1296-1303.
http://dx.doi.org/10.1177/0748233713491803
Alam. MS: Kurohmaru. M. (2015). Butylbenzyl phthalate induces spermatogenic cell apoptosis in
prepubertal rats. Tissue Cell 48: 35-42. http://dx.doi.Org/10.1016/i.tice.2015.12.001
Allen. BC: Kavlock. RJ; Kimmel. CA; Faustman. EM. (1994a). Dose-response assessment for
developmental toxicity II: Comparison of generic benchmark dose estimates with no observed
adverse effect levels. Fundam Appl Toxicol 23: 487-495.
http://dx.doi.org/10.1006/faat.1994.1133
Allen. BC: Kavlock. RJ: Kimmel. CA: Faustman. EM. (1994b). Dose-response assessment for
developmental toxicity III: statistical models. Fundam Appl Toxicol 23: 496-509.
http://dx.doi.org/10.1006/faat.1994.1134
Amin. MM: Parastar. S: Ebrahimpour. K; Shoshtari-Yeganeh. B; Hashemi. M; Mansourian. M;
Kelishadi. R. (2018). Association of urinary phthalate metabolites concentrations with body mass
index and waist circumference. Environ Sci Pollut Res Int 25: 11143-11151.
http://dx.doi.org/10.1007/sll356-018-1413-8
Anderson. WAC: Castle. L; Scotter. MJ; Massev. RC: Springall. C. (2001). A biomarker approach to
measuring human dietary exposure to certain phthalate diesters. Food Addit Contam 18: 1068-
1074. http://dx.doi.org/10.1080/0265203011005Q113
Apel. P; Kortenkamp. A: Koch. HM; Vogel. N: Rtither. M; Kasper-Sonnenberg. M; Conrad. A: Briining.
T; Kolossa-Gehring. M. (2020). Time course of phthalate cumulative risks to male
developmental health over a 27-year period: Biomonitoring samples of the German
Environmental Specimen Bank. 137: 105467.
https://heronet.epa.gov/heronet/index.cfm/reference/download/reference id/6957503
Arbuckle. TE; Agarwal. A: Macpherson. SH; Fraser. WD: Sathyanaravana. S: Ramsay. T; Dodds. L;
Muckle. G: Fisher. M; Foster. W: Walker. M; Monnier. P. (2018). Prenatal exposure to
phthalates and phenols and infant endocrine-sensitive outcomes: The MIREC study. Environ Int
120: 572-583. http://dx.doi.Org/10.1016/i.envint.2018.08.034
Aso. S: Ehara. H; Miyata. K; Hosvuvama. S: Shiraishi. K; Umano. T; Minobe. Y. (2005). A two-
generation reproductive toxicity study of butyl benzyl phthalate in rats. J Toxicol Sci 30: S39-
S58. http://dx.doi.org/10.2131/its.30.S39
Aylward. LL; Hays. SM; Zidek. A. (2016). Variation in urinary spot sample, 24 h samples, and longer-
term average urinary concentrations of short-lived environmental chemicals: implications for
exposure assessment and reverse dosimetry. J Expo Sci Environ Epidemiol 27: 582-590.
http://dx.doi.org/10.1038/ies.2016.54
Berger. K; Eskenazi. B; Kogut. K; Parra. K; Lustig. RH; Greenspan. LC: Holland. N: Calafat. AM: Ye.
X: Harlev. KG. (2018). Association of Prenatal Urinary Concentrations of Phthalates and
Bisphenol A and Pubertal Timing in Boys and Girls. Environ Health Perspect 126: 97004.
http://dx.doi.org/10.1289/EHP3424
Bloom. MS: Wenzel. AG: Brock. JW: Kucklick. JR; Wineland. RJ: Cruze. L; Unal. ER; Yucel. RM;
Jivessova. A: Newman. RB. (2019). Racial disparity in maternal phthalates exposure;
Association with racial disparity in fetal growth and birth outcomes. Environ Int 127: 473-486.
https://heronet.epa.gov/heronet/index.cfm/reference/download/reference id/5494469
Page 72 of 122
-------
1958
1959
1960
1961
1962
1963
1964
1965
1966
1967
1968
1969
1970
1971
1972
1973
1974
1975
1976
1977
1978
1979
1980
1981
1982
1983
1984
1985
1986
1987
1988
1989
1990
1991
1992
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
PUBLIC RELEASE DRAFT
DECEMBER 2024
Boss. J; Zhai. J; Aung. MT; Ferguson. KK; Johns. LE; Mc El rath. TF; Meeker. JD; Mukheriee. B.
(2018). Associations between mixtures of urinary phthalate metabolites with gestational age at
delivery: a time to event analysis using summative phthalate risk scores. Environ Health 17: 56.
http://dx.doi.org/10.1186/sl2940-018-040Q-3
Burns. JS: Sergevev. O: Lee. MM: Williams. PL: Minguez-Alarcon. L; Plaku-Alakbarova. B; Sokolov.
S: Kovalev. S: Koch. HM; Lebedev. AT: Hauser. R; Korrick. SA; Russian Children's. S. (2022).
Associations of prepubertal urinary phthalate metabolite concentrations with pubertal onset
among a longitudinal cohort of boys. Environ Res 212: 113218.
http://dx.doi. org/10.1016/i. envres.2022.113218
Calafat. AM: Longnecker. MP: Koch. HM: Swan. SH: Hauser. R; Goldman. LR; Lanphear. BP: Rudel.
RA: Engel. SM: Teitelbaum. SL: Whyatt. RM: Wolff. MS. (2015). Optimal exposure biomarkers
for nonpersistent chemicals in environmental epidemiology. Environ Health Perspect 123: A166-
A168. http://dx.doi.org/10.1289/ehp.1510041
Carruthers. CM: Foster. PMD. (2005). Critical window of male reproductive tract development in rats
following gestational exposure to di-n-butyl phthalate. Birth Defects Res B Dev Reprod Toxicol
74: 277-285. http://dx.doi.org/10.1002/bdrb.2005Q
CDC. (2021). CDC Health Topics A-Z: Micronutrients [Website],
http s: //www, cdc. gov/nutriti on/mi cronutri ent-
malnutrition/index.html?CDC AA refVal=https%3A%2F%2Fwww.cdc.gov%2Fimmpact%2Fin
dex.html
CDC. (2022). CDC Health Topics A-Z: Physical activity [Website],
http s: //www, cdc. gov/phy si cal activity/index. html
CDC. (2023a). Alcohol and Public Health: Alcohol use and your health [Website],
https://www.cdc.gov/alcohol/fact-sheets/alcohol-use.htm
CDC. (2023b). CDC Health Topics A-Z: Nutrition [Website], https://www.cdc.gov/nutrition/index.html
CDC. (2023c). CDC Health Topics A-Z: Stress at work [Website],
https://www.cdc.gov/niosh/topics/stress/
CDC. (2023d). Fetal Alcohol Spectrum Disorders (FASDs): Alcohol use during pregnancy [Website],
https://www.cdc.gov/ncbddd/fasd/alcohol-use.html
CDC. (2023e). Pregnancy: During pregnancy [Website], https://www.cdc.gov/pregnancv/during.html
CDC. (2023f). Smoking & Tobacco Use: Smoking during pregnancy - Health effects of smoking and
secondhand smoke on pregnancies [Website],
https://www.cdc.gov/tobacco/basic information/health effects/pregnancv/index.htm
CDC. (2023g). Viral Hepatitis: What is viral hepatitis? [Website],
https://www.cdc.gov/hepatitis/abc/index.htm
Chin. HB: Jukic. AM: Wilcox. AJ: Weinberg. CR: Ferguson. KK: Calafat. AM: McConnaughev. PR:
Baird. DP. (2019). Association of urinary concentrations of phthalate metabolites and bisphenol
A with early pregnancy endpoints. Environ Res 168: 254-260.
http://dx.doi.Org/10.1016/i.envres.2018.09.037
Conlev. JM: Lambright. CS: Evans. N: Cardon. M: Medlock-Kakalev. E: Wilson. VS: Gray. LE. (2021).
A mixture of 15 phthalates and pesticides below individual chemical no observed adverse effect
levels (NOAELs) produces reproductive tract malformations in the male rat. Environ Int 156:
106615. http://dx.doi.org/10.1016/i.envint.2021.106615
Pebartolo. P: Javatilaka. S: Yan Siu. N: Rose. M: Ramos. RL: Betz. AJ. (2016). Perinatal exposure to
benzyl butyl phthalate induces alterations in neuronal development/maturation protein
expression, estrogen responses, and fear conditioning in rodents. Behav Pharmacol 27: 77-82.
http://dx.doi.org/10.1097/FBP.000000000000019Q
Powns. SH: Black. N. (1998). The feasibility of creating a checklist for the assessment of the
methodological quality both of randomised and non-randomised studies of health care
Page 73 of 122
-------
2007
2008
2009
2010
2011
2012
2013
2014
2015
2016
2017
2018
2019
2020
2021
2022
2023
2024
2025
2026
2027
2028
2029
2030
2031
2032
2033
2034
2035
2036
2037
2038
2039
2040
2041
2042
2043
2044
2045
2046
2047
2048
2049
2050
2051
2052
2053
2054
PUBLIC RELEASE DRAFT
DECEMBER 2024
interventions. J Epidemiol Community Health 52: 377-384.
http://dx.doi.Org/10.1136/iech.52.6.377
DuPont. (2006a). Polyvinyl chloride film plasticized with butyl benzyl phthalate: In vitro dermal
absorption rate testing. (DuPont-17805). Independence, OH: Ferro Corporation, Inc.
DuPont. (2006b). [Sanitized] Butyl benzyl phthalate: In vitro dermal absorption rate testing. (DuPont-
17755). Independence, OH: Ferro Corporation, Inc.
Dunnaz. E; Erkekoglu. P; Asci. A: Akcurin. S: Bircan. I; Kocer-Gumusel. B. (2018). Urinary phthalate
metabolite concentrations in girls with premature thelarche. Environ Toxicol Pharmacol 59: 172-
181. http://dx.doi.Org/10.1016/i.etap.2018.03.010
EC/HC. (2015a). State of the science report: Phthalate substance grouping: Medium-chain phthalate
esters: Chemical Abstracts Service Registry Numbers: 84-61-7; 84-64-0; 84-69-5; 523-31-9;
5334-09-8; 16883-83-3; 27215-22-1; 27987-25-3; 68515-40-2; 71888-89-6. Gatineau, Quebec:
Environment Canada, Health Canada. https://www.ec.gc.ca/ese-ees/4D845198-761D-428B-
A519-7548 !B25B3E5/SoS Phthalates%20%28Medium-chain%29 EN.pdf
EC/HC. (2015b). State of the Science Report: Phthalates Substance Grouping: Long-chain Phthalate
Esters. 1,2-Benzenedicarboxylic acid, diisodecyl ester (diisodecyl phthalate; DIDP) and 1,2-
Benzenedicarboxylic acid, diundecyl ester (diundecyl phthalate; DUP). Chemical Abstracts
Service Registry Numbers: 26761-40-0, 68515-49-1; 3648-20-2. Gatineau, Quebec:
Environment Canada, Health Canada, https://www.ec.gc.ca/ese-
ees/default.asp?lang=En&n=D3FB0F30-l
ECB. (2007). European Union Risk Assessment Report: Benzyl butyl phthalate (CAS No: 85-68-7,
EINECS: 201-622-7). (EUR 22773 EN). Luxembourg: European Commission.
https://echa.europa.eu/documents/10162/bad5c928-93a5-4592-a4f6-e02c5e89c299
ECCC/HC. (2020). Screening assessment - Phthalate substance grouping. (Enl4-393/2019E-PDF).
Environment and Climate Change Canada, Health Canada.
https://www.canada.ca/en/environment-climate-change/services/evaluating-existing-
substances/screening-assessment-phthalate-substance-grouping.html
ECHA. (2008). Substance name: Benzyl butyl phthalate, EC number: 201-622-7, CAS number: 85-68-7:
Member state committee support document for identification of benzyl butyl phthalate (BBP) as
a substance of very high concern. Helsinki, Finland.
https://echa.europa.eu/documents/10162/b4024445-00ce-4849-9ab0-69aed88c51f2
ECHA. (2010). 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
benzyl butyl phthalate (BBP) CAS No. 85-68-7 Einecs No. 201-622-7 (pp. 15).
ECHA. (2014). Support document to the opinion of the member state committee for identification of
benzyl butyl phthalate (bbp) as a substance of very high concern because of its endocrine
disrupting properties which cause probable serious effects to human health and the environment
which give rise to an equivalent level of concern to those of cmrl and pbt/vpvb2 substances.
Available online at
https://echa.europa.eu/documents/10162/21833221/svhc msc opinion support document bbp
20141211 en.pdf/c5e7a581 -db 15-4e09-8be9-37d42c5409d8. Agency, EC.
ECHA. (2017a). Annex to the Background document to the Opinion on the Annex XV dossier
proposing restrictions on four phthalates (DEHP, BBP, DBP, DIBP). (ECHA/RAC/RES-O-
0000001412-86-140/F; ECHA/SEAC/RES-O-OOOOOO1412-86-154/F).
https://heronet.epa.gov/heronet/index.cfm/reference/download/reference id/10328892
ECHA. (2017b). Opinion on an Annex XV dossier proposing restrictions on four phthalates (DEHP,
BBP, DBP, DIBP). (ECHA/RAC/RES-0-0000001412-86-140/F). Helsinki, Finland.
https://echa.europa.eu/documents/10162/e39983ad-lbf6-f402-7992-8a032b5b82aa
Page 74 of 122
-------
2055
2056
2057
2058
2059
2060
2061
2062
2063
2064
2065
2066
2067
2068
2069
2070
2071
2072
2073
2074
2075
2076
2077
2078
2079
2080
2081
2082
2083
2084
2085
2086
2087
2088
2089
2090
2091
2092
2093
2094
2095
2096
2097
2098
2099
2100
2101
2102
PUBLIC RELEASE DRAFT
DECEMBER 2024
EFSA. (2019). Update of the risk assessment of di-butylphthalate (DBP), butyl-benzyl-phthal ate (BBP),
bis(2-ethylhexyl)phthalate (DEHP), di-isononylphthalate (DINP) and di-isodecylphthalate
(DIDP) for use in food contact materials. EFSA J 17: ee05838.
http://dx.doi.Org/10.2903/i.efsa.2019.5838
Eigenberg. DA: Bozigian. HP: Carter. DE; Sipes. IG. (1986). Distribution, excretion, and metabolism of
butylbenzyl phthalate in the rat. J Toxicol Environ Health 17: 445-456.
http://dx.doi.org/10.1080/1528739860953Q839
Elsisi. AE; Carter. DE: Sipes. IG. (1989). Dermal absorption of phthalate diesters in rats. Fundam Appl
Toxicol 12: 70-77. http://dx.doi.org/10.1016/0272-0590(89)90063-8
Ema. M; Miyawaki. E. (2002). Effects on development of the reproductive system in male offspring of
rats given butyl benzyl phthalate during late pregnancy. Reprod Toxicol 16: 71-76.
http ://dx.doi .org/10.1016/S0890-623 8(01 )00200-3
Ema. M; Miyawaki. E; Hirose. A: Kamata. E. (2003). Decreased anogenital distance and increased
incidence of undescended testes in fetuses of rats given monobenzyl phthalate, a major
metabolite of butyl benzyl phthalate. Reprod Toxicol 17: 407-412.
http ://dx.doi .org/10.1016/S0890-623 8(03 )0003 7-6
Environment Canada. (2000). Canadian environmental protection act priority substances list assessment
report: Butylbenzylphthalate. Ottawa, Ontario: Environment Canada, Health Canada.
https://www.canada.ca/content/dam/hc-sc/migration/hc-sc/ewh-semt/alt formats/hecs-
sesc/pdf/pubs/contaminants/psl2-lsp2/butvlbenzvlphthalate/butvlbenzylphthalate-eng.pdf
Faustman. EM: Allen. BC: Kavlock. RJ; Kimmel. CA. (1994). Dose-response assessment for
developmental toxicity: I characterization of data base and determination of no observed adverse
effect levels. Fundam Appl Toxicol 23: 478-486. http://dx.doi.Org/10.1006/faat.1994.l 132
Foster. PMD. (2005). Mode of action: Impaired fetal Leydig cell function - Effects on male reproductive
development produced by certain phthalate esters [Review], Crit Rev Toxicol 35: 713-719.
http://dx.doi.org/10.1080/104084405910Q7395
Foster. PMD: Mylchreest. E; Gaido. KW: Sar. M. (2001). Effects of phthalate esters on the developing
reproductive tract of male rats [Review], Hum Reprod Update 7: 231-235.
http://dx.doi.Org/10.1093/humupd/7.3.231
Foster. PMD: Thomas. LV; Cook. MW: Gangolli. SD. (1980). Study of the testicular effects and
changes in zinc excretion produced by some n-alkyl phthalates in the rat. Toxicol Appl
Pharmacol 54: 392-398. http://dx.doi.org/10.1016/0041-008X(80)90165-9
Frederiksen. H; Aksglaede. L; Sorensen. K; Skakkebaek. NE; Juul. A: Andersson. AM. (2011). Urinary
excretion of phthalate metabolites in 129 healthy Danish children and adolescents: Estimation of
daily phthalate intake. Environ Res 111: 656-663. http://dx.doi.Org/10.1016/i.envres.2011.03.005
Furr. JR; Lambright. CS: Wilson. VS: Foster. PM; Gray. LE. Jr. (2014). A short-term in vivo screen
using fetal testosterone production, a key event in the phthalate adverse outcome pathway, to
predict disruption of sexual differentiation. Toxicol Sci 140: 403-424.
http://dx.doi.org/10.1093/toxsci/kfu081
Gray. LE: Furr. J: Tatum-Gibbs. KR; Lambright. C: Sampson. H; Hannas. BR: Wilson. VS: Hotchkiss.
A: Foster. PM. (2016). Establishing the Biological Relevance of Dipentyl Phthalate Reductions
in Fetal Rat Testosterone Production and Plasma and Testis Testosterone Levels. Toxicol Sci
149: 178-191. http://dx.doi.org/10.1093/toxsci/kfv224
Gray. LE: Lambright. CS: Conlev. JM: Evans. N: Furr. JR: Hannas. BR: Wilson. VS: Sampson. H:
Foster. PMD. (2021). Genomic and Hormonal Biomarkers of Phthalate-Induced Male Rat
Reproductive Developmental Toxicity Part II: A Targeted RT-qPCR Array Approach That
Defines a Unique Adverse Outcome Pathway. Toxicol Sci 182: 195-214.
http://dx.doi.org/10.1093/toxsci/kfab053
Page 75 of 122
-------
2103
2104
2105
2106
2107
2108
2109
2110
2111
2112
2113
2114
2115
2116
2117
2118
2119
2120
2121
2122
2123
2124
2125
2126
2127
2128
2129
2130
2131
2132
2133
2134
2135
2136
2137
2138
2139
2140
2141
2142
2143
2144
2145
2146
2147
2148
2149
2150
PUBLIC RELEASE DRAFT
DECEMBER 2024
Gray. LE; Ostbv. J; Furr. J; Price. M; Veeramachaneni. DNR; Parks. L. (2000). Perinatal exposure to the
phthalates DEHP, BBP, and DNIP, but not DEP, DMP, or DOTP, alters sexual differentiation of
the male rat. Toxicol Sci 58: 350-365. http://dx.doi.Org/10.1093/toxsci/58.2.350
Gray. TJB; Rowland. IR; Foster. PMD; Gangolli. SD. (1982). Species differences in the testicular
toxicity of phthalate esters. Toxicol Lett 11: 141-147. http://dx.doi.org/10.1016/Q378-
4274(82)90119-9
Hannas. BR: Lambright. CS: Furr. J: Howdeshell. KL; Wilson. VS: Gray. LE. (2011). Dose-response
assessment of fetal testosterone production and gene expression levels in rat testes following in
utero exposure to diethylhexyl phthalate, diisobutyl phthalate, diisoheptyl phthalate, and
diisononyl phthalate. Toxicol Sci 123: 206-216. http://dx.doi.org/10.1093/toxsci/kfrl46
Health Canada. (2015). Supporting documentation: Carcinogenicity of phthalates - mode of action and
human relevance. In Supporting documentation for Phthalate Substance Grouping. Ottawa, ON.
Health Canada. (2018a). 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. Ottawa,
ON.
Health Canada. (2018b). Supporting documentation: Evaluation of epidemiologic studies on phthalate
compounds and their metabolites for hormonal effects, growth and development and
reproductive parameters. Ottawa, ON.
Howdeshell. KL: Hotchkiss. AK; Gray. LE. (2017). Cumulative effects of antiandrogenic chemical
mixtures and their relevance to human health risk assessment [Review], Int J Hyg Environ
Health 220: 179-188. http://dx.doi.Org/10.1016/i.iiheh.2016.ll.007
Howdeshell. KL: Rider. CV: Wilson. VS: Furr. JR; Lambright. CR; Gray. LE. (2015). Dose addition
models based on biologically relevant reductions in fetal testosterone accurately predict postnatal
reproductive tract alterations by a phthalate mixture in rats. Toxicol Sci 148: 488-502.
http://dx.doi.org/10.1093/toxsci/kfvl96
Howdeshell. KL: Wilson. VS: Furr. J: Lambright. CR: Rider. CV: Blystone. CR: Hotchkiss. AK: Gray.
LE. Jr. (2008). A mixture of five phthalate esters inhibits fetal testicular testosterone production
in the Sprague-Dawley rat in a cumulative, dose-additive manner. Toxicol Sci 105: 153-165.
http://dx.doi.org/10.1093/toxsci/kfn077
Huang. LL: Zhou. B: Ai. SH: Yang. P: Chen. YJ: Liu. C: Deng. YL: Lu. O: Miao. XP: Lu. WO: Wang.
YX: Zeng. O. (2018). Prenatal phthalate exposure, birth outcomes and DNA methylation of Alu
and LINE-1 repetitive elements: A pilot study in China. Chemosphere 206: 759-765.
http://dx.doi. org/10.1016/i. chemosphere. 2018.05.030
Integrated Laboratory Systems. (2017). Pubertal development and thyroid function in intact
juvenile/peripubertal female rats; benzyl butyl phthalate. (10005.0302). Research Triangle Park,
NC: RTI International.
Ivell. R; Wade. JD; Anand-Ivell. R. (2013). INSL3 as aBiomarker of Leydig Cell Functionality
[Review], Biol Reprod 88: 147-147. http://dx.doi.Org/10.1095/biolreprod.l 13.108969
Jahreis. S: Trump. S: Bauer. M; Bauer. T; Thtirmann. L; Feltens. R; Wang. Q: Gu. L; Griitzmann. K;
Roder. S: Averbeck. M; Weichenhan. D; Plass. C: Sack. U; Borte. M; Dubourg. V: Schiiurmann.
G: Simon. JC: Martin Von. B; Hackermtiller. J; Eils. R; Lehmann. I; Polte. T. (2018). Maternal
phthalate exposure promotes allergic airway inflammation over 2 generations through epigenetic
modifications. J Allergy Clin Immunol 141: 741-753.
http://dx.doi.Org/10.1016/i.iaci.2017.03.017
Johnson. KJ; Heger. NE; Boekelheide. K. (2012). Of mice and men (and rats): phthalate-induced fetal
testis endocrine disruption is species-dependent [Review], Toxicol Sci 129: 235-248.
http://dx.doi.org/10.1093/toxsci/kfs206
Page 76 of 122
-------
2151
2152
2153
2154
2155
2156
2157
2158
2159
2160
2161
2162
2163
2164
2165
2166
2167
2168
2169
2170
2171
2172
2173
2174
2175
2176
2177
2178
2179
2180
2181
2182
2183
2184
2185
2186
2187
2188
2189
2190
2191
2192
2193
2194
2195
2196
2197
PUBLIC RELEASE DRAFT
DECEMBER 2024
Johnson. KJ; McDowell EN; Viereck. MP; Xia. JO. (2011). Species-specific dibutyl phthalate fetal
testis endocrine disruption correlates with inhibition of SREBP2-dependent gene expression
pathways. Toxicol Sci 120: 460-474. http://dx.doi.org/10.1093/toxsci/kfr020
Lash lev. S; Calafat A; Barr. D; Ledoux. T; Hore. P; Lake. M; Robson. M; Smulian. J. (2004).
Endocrine Disruptors In The Maternal And Fetal Compartments [Abstract], Am J Obstet
Gynecol 191: S140. http://dx.doi.Org/10.1016/i.aiog.2004.10.390
Lee. G; Kim. S; Bastiaensen. M; Malarvannan. G; Poma. G; Caballero Casero. N; Gvs. C; Covaci. A;
Lee. S; Lim. JE; Mok. S; Moon. HB; Choi. G; Choi. K. (2020). Exposure to organophosphate
esters, phthalates, and alternative plasticizers in association with uterine fibroids. 189: 109874.
https://heronet.epa.gov/heronet/index.cfm/reference/download/reference id/7274600
MacLeod. DJ; Sharpe. RM; Welsh. M; Fisken. M; Scott. HM; Hutchison. GR; Drake. AJ; van Den
Driesche. S. (2010). Androgen action in the masculinization programming window and
development of male reproductive organs. Int J Androl 33: 279-287.
http://dx.doi.org/10.1111/i. 1365-2605.2009.01005.X
Main. KM; Mortensen. GK; Kaleva. MM; Boisen. KA; Damgaard. IN; Chellakootv. M; Schmidt. IM;
Suomi. AM; Virtanen. HE; Petersen. JH; Andersson. AM; Toppari. J; Skakkebaek. NE. (2006).
Human breast milk contamination with phthalates and alterations of endogenous reproductive
hormones in infants three months of age. Environ Health Perspect 114: 270-276.
http://dx.doi.org/10.1289/ehp.8075
Min. A; Liu. F; Yang. X; Chen. M. (2014). Benzyl butyl phthalate exposure impairs learning and
memory and attenuates neurotransmission and CREB phosphorylation in mice. Food Chem
Toxicol 71: 81-89. http://dx.doi.Org/10.1016/i.fct.2014.05.021
Nagao. T; Ohta. R; Marumo. H; Shindo. T; Yoshimura. S; Ono. H. (2000). Effect of butyl benzyl
phthalate in Sprague-Dawley rats after gavage administration: A two-generation reproductive
study. Reprod Toxicol 14: 513-532. http://dx.doi.org/10.1016/S0890-6238(W)00105-2
Nakagomi. M; Suzuki. E; Saito. Y; Nagao. T. (2017). Endocrine disrupting chemicals, 4-nonylphenol,
bisphenol A and butyl benzyl phthalate, impair metabolism of estradiol in male and female rats
as assessed by levels of 15a-hydroxyestrogens and catechol estrogens in urine. J Appl Toxicol
38: 688-695. http://dx.doi.org/10.1002/iat.3574
NASEM. (2017). Application of systematic review methods in an overall strategy for evaluating low-
dose toxicity from endocrine active chemicals. In Consensus Study Report. Washington, D.C.:
The National Academies Press, http://dx.doi.org/10.17226/24758
Nativelle. C; Picard. K; Valentin. I; Lhuguenot. JC; Chagnon. MC. (1999). Metabolism of n-butyl
benzyl phthalate in the female Wistar rat. Identification of new metabolites. Food Chem Toxicol
37: 905-917. http://dx.doi.org/10.1016/S0278-6915(W)00071-X
NICNAS. (2008). Phthalates hazard compendium: A summary of physicochemical and human health
hazard data for 24 ortho-phthalate chemicals. Sydney, Australia: Australian Department of
Health and Ageing, National Industrial Chemicals Notification and Assessment Scheme.
https://www.regulations.gov/document/EPA-HQ-OPPT-2010-0573-00Q8
NICNAS. (2015). Priority existing chemical assessment report no. 40: Butyl benzyl phthalate. (PEC40).
Sydney, Australia: Australian Government Department of Health and Ageing.
https://heronet.epa.gov/heronet/index.cfm/reference/download/reference id/3664467
NICNAS. (2016). C4-6 side chain transitional phthalates: Human health tier II assessment. Sydney,
Australia: Australian Department of Health, National Industrial Chemicals Notification and
Assessment Scheme. https://www.industrialchemicals.gov.au/sites/default/files/C4-
6%20side%20chain%20transitional%20phthalates Human%20health%20tier%20II%20assessm
ent.pdf
Page 77 of 122
-------
2198
2199
2200
2201
2202
2203
2204
2205
2206
2207
2208
2209
2210
2211
2212
2213
2214
2215
2216
2217
2218
2219
2220
2221
2222
2223
2224
2225
2226
2227
2228
2229
2230
2231
2232
2233
2234
2235
2236
2237
2238
2239
2240
2241
2242
2243
2244
2245
PUBLIC RELEASE DRAFT
DECEMBER 2024
NTP-CERHR. (2003). NTP-CERHR monograph on the potential human reproductive and
developmental effects of butyl benzyl phthalate (BBP). (NIH Publication No. 03-4487).
https://ntp.niehs.nih.gov/ntp/ohat/phthalates/bb-phthalate/bbp monograph final.pdf
NTP. (2015). Handbook for conducting a literature-based health assessment using OHAT approach for
systematic review and evidence integration. Research Triangle Park, NC: U.S. Deptartment of
Health and Human Services, National Toxicology Program, Office of Health Assessment and
Translation, https://ntp.niehs.nih.gov/ntp/ohat/pubs/handbookian2015 508.pdf
ODPHP. (2023a). Healthy People 2030 - Social determinants of health literature summaries:
Neighborhood and built environment [Website], https://health.gov/healthypeople/prioritv-
areas/social-determinants-health/literature-summaries#neighborhood
ODPHP. (2023b). Healthy People 2030 - Social determinants of health literature summaries: Poverty
[Website], https://health.gov/healthvpeople/prioritv-areas/social-determinants-health/literature-
summaries/poverty
ODPHP. (2023c). Healthy People 2030 - Social determinants of health literature summaries: Social and
community context [Website], https://health.gov/healthvpeople/prioritv-areas/social-
determinants-health/literature-summaries#social
OECD. (2018). Revised Guidance Document 150 on Standardised Test Guidelines for Evaluating
Chemicals for Endocrine Disruption Two-generation reproduction toxicity study (OECD TG
416) (pp. 617 - 630). Paris, France: OECD Publishing.
http://dx.doi.org/10.1787/9789264304741-33-en
OEHHA. (1986). Safe Drinking Water and Toxic Enforcement Act of 1986 Proposition 65. Initial
Statement of Reasons. Title 27, California Code of Regulations. Proposed amendment to Section
25805(b), Specific Regulatory Levels: Chemicals Causing Reproductive Toxicity. Butyl benzyl
phthalate (oral exposure). California: California Environmental Protection Agency, Office of
Environmental Health Hazard Assessment, https://oehha.ca.gov/media/downloads/proposition-
65/chemicals/060112bbpisor.pdf
Parks. LG: Ostbv. JS: Lambright CR; Abbott. BP: Klinefelter. GR; Barlow. NJ: Gray. LE. Jr. (2000).
The plasticizer diethylhexyl phthalate induces malformations by decreasing fetal testosterone
synthesis during sexual differentiation in the male rat. Toxicol Sci 58: 339-349.
http://dx.doi.Org/10.1093/toxsci/58.2.339
Radke. EG: Braun. JM; Meeker. JD; Cooper. GS. (2018). Phthalate exposure and male reproductive
outcomes: A systematic review of the human epidemiological evidence [Review], Environ Int
121: 764-793. http://dx.doi.Org/10.1016/i.envint.2018.07.029
Radke. EG: Braun. JM: Nachman. RM; Cooper. GS. (2020a). Phthalate exposure and
neurodevelopment: A systematic review and meta-analysis of human epidemiological evidence
[Review], Environ Int 137: 105408. http://dx.doi.Org/10.1016/i.envint.2019.105408
Radke. EG: Galizia. A: Thayer. KA; Cooper. GS. (2019a). Phthalate exposure and metabolic effects: A
systematic review of the human epidemiological evidence [Review], Environ Int 132: 104768.
http: //dx. doi. or g/10.1016/i. envint .2019.04.040
Radke. EG: Glenn. BS: Braun. JM: Cooper. GS. (2019b). Phthalate exposure and female reproductive
and developmental outcomes: A systematic review of the human epidemiological evidence
[Review], Environ Int 130: 104580. http://dx.doi.Org/10.1016/i.envint.2019.02.003
Radke. EG: Yost. EE: Roth. N: Sathyanaravana. S: Whalev. P. (2020b). Application of US EPA IRIS
systematic review methods to the health effects of phthalates: Lessons learned and path forward
[Editorial], Environ Int 145: 105820. http://dx.doi.org/10.1016/i.envint.2020.105820
Rozati. R; Reddv. PP; Reddanna. P; Muitaba. R. (2002). Role of environmental estrogens in the
deterioration of male factor fertility. Fertil Steril 78: 1187-1194.
http://dx.doi. org/10.1016/S0015-0282(02)043 89-3
Page 78 of 122
-------
2246
2247
2248
2249
2250
2251
2252
2253
2254
2255
2256
2257
2258
2259
2260
2261
2262
2263
2264
2265
2266
2267
2268
2269
2270
2271
2272
2273
2274
2275
2276
2277
2278
2279
2280
2281
2282
2283
2284
2285
2286
2287
2288
2289
2290
2291
2292
PUBLIC RELEASE DRAFT
DECEMBER 2024
Schmitt EE; Vellers. HL; Porter. WW; Lightfoot. JT. (2016). Environmental endocrine disruptor affects
voluntary physical activity in mice. Med Sci Sports Exerc 48: 1251-1258.
http://dx.doi.org/10.1249/MSS.00000000000009Q8
Schwartz. CL; Christiansen. S; Hass. U; Ramhai. L; Axelstad. M; Lobl. NM; Svingen. T. (2021). On the
use and interpretation of areola/nipple retention as a biomarker for anti-androgenic effects in rat
toxicity studies [Review], Front Toxicol 3: 730752.
https://heronet.epa.gov/heronet/index.cfm/reference/download/reference id/10492323
Schwartz. CL; Christiansen. S; Vinggaard. AM; Axelstad. M; Hass. U; Svingen. T. (2019). Anogenital
distance as a toxicological or clinical marker for fetal androgen action and risk for reproductive
disorders [Review], Arch Toxicol 93: 253-272. http://dx.doi.org/10.1007/s00204-018-235Q-5
Scott. RC; Dugard. PH; Ramsey. JD; Rhodes. C. (1987). In vitro absorption of some o-phthalate diesters
through human and rat skin. Environ Health Perspect 74: 223-227.
http://dx.doi.org/10.2307/3430452
Shin. HM; Bennett. DH; Barkoski. J; Ye. X; Calafat. AM; Tancredi. D; Hertz-Picciotto. I. (2019).
Variability of urinary concentrations of phthalate metabolites during pregnancy in first morning
voids and pooled samples. Environ Int 122: 222-230.
http://dx.doi.Org/10.1016/i.envint.2018.l 1.012
Spade. DJ; Bai. CY; Lambright. C; Conlev. JM; Boekelheide. K; Gray. LE. (2018). Validation of an
automated counting procedure for phthalate-induced testicular multinucleated germ cells.
Toxicol Lett 290: 55-61. http://dx.doi.Org/10.1016/i.toxlet.2018.03.018
Stahlhut. RW; van Wijngaarden. E; Dye. TP; Cook. S; Swan. SH. (2007). Concentrations of urinary
phthalate metabolites are associated with increased waist circumference and insulin resistance in
adult U.S. males. Environ Health Perspect 115: 876-882. http://dx.doi.org/10.1289/ehp.9882
Sterne. JAC; Hernan. MA; Reeves. BC; Savovic. J; Berkman. ND; Viswanathan. M; Henry. D; Altman.
DG; Ansari. MT; Boutron. I; Carpenter. JR; Chan. AW; Churchill. R; Peeks. JJ; Hrobiartsson.
A; Kirkham. J; Juni. P; Loke. YK; Pigott. TP; Ramsay. CR; Regidor. P; Rothstein, HR; Sandhu,
L; Santaguida, PL; Schiinemann, HJ; Shea, B; Shrier, I; Tugwell, P; Turner, L; Valentine, JC;
Waddington, H; Waters, E; Wells, GA; Whiting, PF; Higgins, JPT. (2016). ROBINS-I: A tool
for assessing risk of bias in non-randomised studies of interventions. BMJ 355: i4919.
http://dx.doi. org/10.113 6/bmi ,i4919
Sugino. M; Hatanaka. T; Todo. H; Mashimo. Y; Suzuki. T; Kobavashi. M; Hosova. O; Jinno. H; Juni.
K; Sugibavashi. K. (2017). Safety evaluation of dermal exposure to phthalates: Metabolism-
dependent percutaneous absorption. Toxicol Appl Pharmacol 328: 10-17.
http://dx.doi.Org/10.1016/i.taap.2017.05.009
Thompson. CJ; Ross. SM; Henslev. J; Liu. K; Heinze. SC; Young. SS; Gaido. KW. (2005). Pifferential
steroidogenic gene expression in the fetal adrenal gland versus the testis and rapid and dynamic
response of the fetal testis to di(n-butyl) phthalate. Biol Reprod 73: 908-917.
http://dx.doi.org/10.1095/biolreprod.105.042382
Tvl. RW; Myers. CB; Marr. MC; Fail. PA; Seelv. JC; Brine. PR: Barter. RA; Butala. JH. (2004).
Reproductive toxicity evaluation of dietary butyl benzyl phthalate (BBP) in rats. Reprod Toxicol
18: 241-264. http://dx.doi.Org/10.1016/i.reprotox.2003.10.006
U.S. CPSC. (2010). Toxicity review of benzyl-n-butyl phthalate. Bethesda, MP: U.S. Consumer Product
Safety Commission, Pirectorate for Hazard Identification and Reduction.
https://www.cpsc.gov/s3fs-public/ToxicitvReviewOfBBP.pdf
U.S. CPSC. (2014). Chronic Hazard Advisory Panel on Phthalates and Phthalate Alternatives (with
appendices). Bethesda, MP: U.S. Consumer Product Safety Commission, Pirectorate for Health
Sciences. https://www.cpsc.gov/s3fs-public/CHAP-REPORT-With-Appendices.pdf
Page 79 of 122
-------
2293
2294
2295
2296
2297
2298
2299
2300
2301
2302
2303
2304
2305
2306
2307
2308
2309
2310
2311
2312
2313
2314
2315
2316
2317
2318
2319
2320
2321
2322
2323
2324
2325
2326
2327
2328
2329
2330
2331
2332
2333
2334
2335
2336
2337
2338
2339
2340
PUBLIC RELEASE DRAFT
DECEMBER 2024
U.S. EPA. (1989). IRIS Assessment of: Butyl benzyl phthalate (CASRN 85-68-7). Washington, DC:
U.S. Environmental Protection Agency, Intergrated Risk Information System.
https://cfpub.epa.gov/ncea/iris2/chemicalLanding.cfm7substance nmbr=293
U.S. EPA. (1991). Guidelines for developmental toxicity risk assessment. Fed Reg 56: 63798-63826.
U.S. EPA. (1993). Reference Dose (RfD): description and use in health risk assessments background
document 1A, March 15, 1993. Washington, DC: U.S. Environmental Protection Agency,
Integrated Risk Information System, https://www.epa.gov/iris/reference-dose-rfd-description-
and-use-health-risk-assessments
U.S. EPA. (1994). Methods for derivation of inhalation reference concentrations and application of
inhalation dosimetry [EPA Report], (EPA600890066F). Research Triangle Park, NC.
https://cfpub.epa.gov/ncea/risk/recordisplav.cfm?deid=71993&CFID=51174829&CFTOKEN=2
5006317
U.S. EPA. (1996). Guidelines for reproductive toxicity risk assessment [EPA Report], (EPA/630/R-
96/009). Washington, DC: U.S. Environmental Protection Agency, Risk Assessment Forum.
https://nepis.epa. gov/Exe/ZvPURL.cgi?Dockev=3 0004YQB.txt
U.S. EPA. (1998). Health effects test guidelines OPPTS 870.3800 reproduction and fertility effects
[EPA Report], (EPA 712-C-98-208). Washington D.C.: U.S. Environmental Protection Agency,
Office of Prevention, Pesticides and Toxic Substances.
U.S. EPA. (2002a). Provisional Peer Reviewed Toxicity Values for butyl benzyl phthalate (CASRN 85-
68-7): Derivation of a carcinogenicity assessment [EPA Report], Cincinnati, OH.
U.S. EPA. (2002b). A review of the reference dose and reference concentration processes.
(EPA630P02002F). Washington, DC. https://www.epa.gov/sites/production/files/2014-
12/documents/rfd-final.pdf
U.S. EPA. (201 la). Exposure factors handbook: 2011 edition [EPA Report], (EPA/600/R-090/052F).
Washington, DC: U.S. Environmental Protection Agency, Office of Research and Development,
National Center for Environmental Assessment.
https://nepis.epa. gov/Exe/ZvPURL.cgi?Dockev=P100F2QS.txt
U.S. EPA. (201 lb). Exposure factors handbook: 2011 edition (final) (EPA/600/R-090/052F).
Washington, DC. http://cfpub.epa.gov/ncea/cfm/recordisplav.cfm?deid=236252
U.S. EPA. (2011c). Recommended use of body weight 3/4 as the default method in derivation of the oral
reference dose. (EPA100R110001). Washington, DC.
https://www.epa.gov/sites/production/files/2013-09/documents/recommended-use-of-bw34.pdf
U.S. EPA. (2012). Benchmark dose technical guidance [EPA Report], (EPA100R12001). Washington,
DC: U.S. Environmental Protection Agency, Risk Assessment Forum.
https://www.epa.gov/risk/benchmark-dose-technical-guidance
U.S. EPA. (2019). Proposed designation of Dibutyl Phthalate (CASRN 84-74-2) as a high-priority
substance for risk evaluation. U.S. Environmental Protection Agency, Office of Chemical Safety
and Pollution Prevention, https://www.epa.gov/sites/production/files/2019-
08/documents/dibutvlphthalate 84-74-2 high-priority proposeddesignation 082319.pdf
U.S. EPA. (2020a). Draft Scope of the Risk Evaluation for Butyl Benzyl Phthalate (1,2-
Benzenedicarboxylic acid, 1-butyl 2-(phenylmethyl) ester) CASRN 85-68-7 [EPA Report],
(EPA-740-D-20-015). https://www.epa.gov/sites/production/files/2020-04/documents/casrn-84-
74-2 butyl benzyl phthalate draft scope 4-15-2020.pdf
U.S. EPA. (2020b). Final scope of the risk evaluation for butyl benzyl phthalate (1,2-
benzenedicarboxylic acid, 1-butyl 2-(phenylmethyl) ester); CASRN 85-68-7 [EPA Report],
(EPA-740-R-20-015). Washington, DC: Office of Chemical Safety and Pollution Prevention.
https://www.epa.gov/sites/default/files/2020-09/documents/casrn 85-68-
7 butyl benzyl phthalate finalscope.pdf
Page 80 of 122
-------
2341
2342
2343
2344
2345
2346
2347
2348
2349
2350
2351
2352
2353
2354
2355
2356
2357
2358
2359
2360
2361
2362
2363
2364
2365
2366
2367
2368
2369
2370
2371
2372
2373
2374
2375
2376
2377
2378
2379
2380
2381
2382
2383
2384
2385
2386
2387
2388
PUBLIC RELEASE DRAFT
DECEMBER 2024
U.S. EPA. (2021). Draft systematic review protocol supporting TSCA risk evaluations for chemical
substances, Version 1.0: A generic TSCA systematic review protocol with chemical-specific
methodologies. (EPA Document #EPA-D-20-031). Washington, DC: Office of Chemical Safety
and Pollution Prevention. https://www.regulations.gov/document/EPA-HQ-OPPT-2021-0414-
0005
U.S. EPA. (2022). ORD staff handbook for developing IRIS assessments [EPA Report], (EPA 600/R-
22/268). Washington, DC: U.S. Environmental Protection Agency, Office of Research and
Development, Center for Public Health and Environmental Assessment.
https://cfpub.epa.gov/ncea/iris drafts/recordisplav.cfm?deid=356370
U.S. EPA. (2023a). Draft Proposed Approach for Cumulative Risk Assessment of High-Priority
Phthalates and a Manufacturer-Requested Phthalate under the Toxic Substances Control Act.
(EPA-740-P-23-002). Washington, DC: U.S. Environmental Protection Agency, Office of
Chemical Safety and Pollution Prevention. https://www.regulations.gov/document/EPA-HQ-
QPPT-2022-0918-0009
U.S. EPA. (2023b). Science Advisory Committee on Chemicals meeting minutes and final report, No.
2023-01 - A set of scientific issues being considered by the Environmental Protection Agency
regarding: Draft Proposed Principles of Cumulative Risk Assessment (CRA) under the Toxic
Substances Control Act and a Draft Proposed Approach for CRA of High-Priority Phthalates and
a Manufacturer-Requested Phthalate. (EPA-HQ-OPPT-2022-0918). Washington, DC: U.S.
Environmental Protection Agency, Office of Chemical Safety and Pollution Prevention.
https://www.regulations.gov/document/EPA-HQ-OPPT-2022-0918-0Q67
U.S. EPA. (2024a). Draft Data Extraction Information for Environmental Hazard and Human Health
Hazard Animal Toxicology and Epidemiology for Butyl Benzyl Phthalate (BBP). Washington,
DC: Office of Pollution Prevention and Toxics.
U.S. EPA. (2024b). Draft Data Quality Evaluation Information for Human Health Hazard Epidemiology
for Butyl Benzyl Phthalate (BBP). Washington, DC: Office of Pollution Prevention and Toxics.
U.S. EPA. (2024c). 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), and Dicyclohexyl Phthalate (DCHP).
Washington, DC: Office of Pollution Prevention and Toxics.
U.S. EPA. (2024d). Draft Non-cancer Human Health Hazard Assessment for Butyl benzyl phthalate
(BBP). Washington, DC: Office of Pollution Prevention and Toxics.
U.S. EPA. (2024e). Draft Non-cancer Human Health Hazard Assessment for Dibutyl Phthalate (DBP).
Washington, DC: Office of Pollution Prevention and Toxics.
U.S. EPA. (2024f). Draft Non-Cancer Human Health Hazard Assessment for Dicyclohexyl Phthalate
(DCHP). Washington, DC: Office of Pollution Prevention and Toxics.
U.S. EPA. (2024g). Draft Non-cancer Human Health Hazard Assessment for Diethylhexyl Phthalate
(DEHP). Washington, DC: Office of Pollution Prevention and Toxics.
U.S. EPA. (2024h). Draft Non-cancer Human Health Hazard Assessment for Diisobutyl phthalate
(DIBP). Washington, DC: Office of Pollution Prevention and Toxics.
U.S. EPA. (2024i). Science Advisory Committee on Chemicals Meeting Minutes and Final Report No.
2024-2, Docket ID: EPA-HQ-OPPT-2024-0073: For the Draft Risk Evaluation for Di-isodecyl
Phthalate (DIDP) and Draft Hazard Assessments for Di-isononyl Phthalate (DINP). Washington,
DC: U.S. Environmental Protection Agency, Science Advisory Committee on Chemicals.
U.S. EPA. (2025a). Draft Cancer Human Health Hazard Assessment for Di(2-ethylhexyl) Phthalate
(DEHP), Dibutyl Phthalate (DBP), Butyl Benzyl Phthalate (BBP), Diisobutyl Phthalate (DIBP),
and Dicyclohexyl Phthalate (DCHP). Washington, DC: Office of Pollution Prevention and
Toxics.
Page 81 of 122
-------
2389
2390
2391
2392
2393
2394
2395
2396
2397
2398
2399
2400
2401
2402
2403
2404
2405
2406
2407
2408
2409
2410
2411
2412
2413
2414
PUBLIC RELEASE DRAFT
DECEMBER 2024
U.S. EPA. (2025b). Draft Consumer and Indoor Dust Exposure Assessment for Butyl benzyl phthalate
(BBP). Washington, DC: Office of Pollution Prevention and Toxics.
U.S. EPA. (2025c). Draft Data Quality Evaluation Information for Human Health Hazard Animal
Toxicology for Butyl Benzyl Phthalate (BBP). Washington, DC: Office of Pollution Prevention
and Toxics.
U.S. EPA. (2025d). Draft Data Quality Evaluation Information for Human Health Hazard Epidemiology
for Butyl Benzyl Phthalate (BBP). Washington, DC: Office of Pollution Prevention and Toxics.
U.S. EPA. (2025e). Draft Environmental Release and Occupational Exposure Assessment for Butyl
benzyl phthalate (BBP). Washington, DC: Office of Pollution Prevention and Toxics.
U.S. EPA. (2025f). Draft Risk Evaluation for Butyl Benzyl Phthalate (BBP). Washington, DC: Office of
Pollution Prevention and Toxics.
U.S. EPA. (2025g). Draft Systematic Protocol for Butyl Benzyl Phthalate (BBP). Washington, DC:
Office of Pollution Prevention and Toxics.
U.S. EPA. (2025h). Non-Cancer Human Health Hazard Assessment for Diisononyl Phthalate (DINP)
Washington, DC: Office of Pollution Prevention and Toxics.
Welsh. M; Saunders. PTK; Fisken. M; Scott. HM; Hutchison. GR; Smith. LB: Sharpe. RM. (2008).
Identification in rats of a programming window for reproductive tract masculinization, disruption
of which leads to hypospadias and cryptorchidism. J Clin Invest 118: 1479-1490.
http: //dx. doi. or g/10.1172/i ci34241
Wilson. VS: Lambright. C: Furr. J: Ostbv. J: Wood. C: Held. G: Gray. LE. Jr. (2004). Phthalate ester-
induced gubernacular lesions are associated with reduced insl3 gene expression in the fetal rat
testis. Toxicol Lett 146: 207-215. http://dx.doi.Org/10.1016/i.toxlet.2003.09.012
Yamasaki. K; Takahashi. M; Yasuda. M. (2005). Two-generation reproductive toxicity studies in rats
with extra parameters for detecting endocrine disrupting activity: Introductory overview of
results for nine chemicals. J Toxicol Sci 30: 1-4. http://dx.doi.org/10.2131/its.30.Sl
Page 82 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
2415 APPENDICES
2416 Appendix A EXISTING ASSESSMENTS FROM OTHER REGULATORY AGENCIES OF BBP
2417 The available existing assessments of BBP are summarized TableApx A-l, which includes details regarding external peer-review, public
2418 consultation, and systematic review protocols that were used.
2419
2420 Table Apx A-l. SUMMARY OF PEER-REVIEW, PUBLIC COMMENTS, AND SYSTEMATIC REVIEW FOR EXISTING
2421 ASSESSMENTS OF BBP
Agency
Assessment(s) (Reference)
External
Peer-
Review?
Public
Consultation?
Systematic
Review Protocol
Employed?
Remarks
U.S. EPA (IRIS
Program)
IRIS Assessment of: Butyl benzyl
vhthalate (CASRN85-68-7 (U.S. EPA.
1989),
Phthalate exposure and male
reproductive outcomes: A systematic
review of the human epidemiological
evidence (Radke et al.. 2018).
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 lull systematic review
protocol is available as a supplemental file
associated with each publication.
U.S. CPSC
Toxicity review for benzyl-n-butvl
vhthalate (U.S. CPSC. 2010).
Chronic Hazard Advisory Panel on
phthalates and phthalate alternatives
(U.S. CPSC. 2014).
Yes
Yes
No
- Peer-reviewed by panel of four experts. Peer-
review report available at:
httt>s://www.ct>sc.eov/s3fs-t>ublic/Peer-Review-
RcDort-Commcnts.Ddf
-Public comments available at:
httos ://www. cose. eov/chao
- No formal systematic review protocol employed.
- Details regarding CPSC's strategy for identifying
new information and literature are provided on page
12 of (U.S. CPSC. 2014).
NASEM
Application of systematic review methods
in an overall strategy for evaluating low-
dose toxicity from endocrine active
chemicals (NASEM. 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.
Page 83 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Agency
Assessment(s) (Reference)
External
Peer-
Review?
Public
Consultation?
Systematic
Review Protocol
Employed?
Remarks
- Employed NTP's Office of Heath Assessment and
Translation (OHAT) systematic review method.
California OEHHA
Safe Drinking Water and Toxic
Enforcement Act of1986 Proposition 65.
Initial Statement of Reasons. Title 27,
California Code of Regulations.
Proposed amendment to Section
25805(b), Specific Regulatory Levels:
Chemicals Causing Reproductive
Toxicity. Butyl benzyl phthalate (oral
exposure) (OEHHA. 1986).
No
No
No
- In a statement of reasons for this evaluation.
OEHHA (1986) states "BBP was added to the
Proposition 65 list, based on formal identification
as causing reproductive toxicity (developmental
endpoint) by the National Toxicology Program
(NTP) and in a report by its Center for the
Evaluation of Risks to Human Reproduction
(CERHR)" (page 2).
- No formal systematic review protocol employed,
although human and animal study selection is
explained
Health Canada
State of the science report: Phthalate
substance grouping: Medium-chain
phthalate esters: Chemical Abstracts
Ser\'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.
2015a),
Canadian environmental protection act
priority substances list assessment
report: Butvlbenzylphthalate
(Environment Canada. 2000).
Supporting Documentation:
Carcinogenicity of Phthalates - Mode of
Action and Human Relevance (Health
Canada. 2015).
Supporting documentation: Evaluation
of epidemiologic studies on phthalate
compounds and their metabolites for
hormonal effects, growth and
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 oase 2 of (ECCC/HC. 2020) for additional
details.
- EC/HC (2000) provides a summary of information
critical to assessment (page 7) and search strategies
employed for identification of relevant data (page
57).
- State of the science reoort (EC/HC. 2015a) and
draft screening assessment report for the phthalate
substance group subjected to 60-day public
comment periods. Summaries of received public
comments available at:
httos ://www. Canada, ca/en/health-
canada/services/chemical-substances/substance-
eroumnes-initiative/r>hthalate.html#al
- No formal systematic review protocol employed
to identify or evaluate experimental animal
toxicology studies.
- Details regarding Health Canada's strategy for
identifying new information and literature are
Page 84 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Agency
Assessment(s) (Reference)
External
Peer-
Review?
Public
Consultation?
Systematic
Review Protocol
Employed?
Remarks
development and reproductive
parameters (Health Canada. 2018b).
Supporting documentation: Evaluation
of epidemiologic studies on phthalate
compounds and their metabolites for
effects on behaviour and
neurodevelopment, allergies,
cardiovascular function, oxidative stress,
breast cancer, obesity, and metabolic
disorders (Health Canada. 2018a).
Screening Assessment - Phthalate
Substance Grouvine (ECCC/HC. 2020).
orovidcd in Section 1 of (EC/HC. 2015a) and
(ECCC/HC. 2020)
- Human epidemiologic studies evaluated using
Downs and Black Method (Health Canada. 2018a.
b).
NICNAS
Priority existing chemical assessment
report no. 40: Butvl benzvl phthalate
(NICNAS. 2015).'
C4-6 side chain transitional phthalates:
Human health tier II assessment
(NICNAS. 2016).
No
Yes
No
- NICNAS (2015) states "On comoletine a PEC
assessment, the Director of NICNAS, in accordance
with the Act, causes a draft report of the assessment
to be prepared and makes it available to the
applicants for factual correction and to the pubic
(including applicants and other interested parties)
for comments." See Preface of for more details.
- No formal systematic review protocol employed.
- Details regarding NICNAS's strategy for
identifying new information and literature are
orovidcd in Section 1.3 of (NICNAS. 2015).
ECHA
Substance name: Benzyl butyl phthalate,
EC number: 201-622-7, CAS number:
85-68-7: Member state committee
support documentation for identification
of benzyl butyl phthalate (BBP) as a
substance ofverv hish concern ('ECHA.
2008),
Evaluation of new scientific evidence
concerning the restriction contained in
Annex AT 'II to regulation (EC) no.
1907/2006 (REACH): Review of new
No
Yes
No
- ECHA (2017b) states "This document dresents
opinions adopted by RAC and SEAC. The
Background Document, as a supportive document
to both RAC and SEAC opinions and their
justifications, gives the details of the Dossier
Submitter's proposal, amended for further
information obtained during the public consultation
and other relevant information resulting from the
opinion making process." See document for more
details.
- No formal systematic review protocol employed.
Page 85 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Agency
Assessment(s) (Reference)
External
Peer-
Review?
Public
Consultation?
Systematic
Review Protocol
Employed?
Remarks
available information for benzyl butyl
ph thai ate (BBP) CAS no. 85-68-7 Einecs
no. 201-622-7 (ECHA. 2010).
Support document to the opinion of the
member state committee for
identification of benzyl butyl phthalate
(bbp) as a substance of very high
concern because of its endocrine
disrupting properties which cause
probable serious effects to human health
and the environment which give rise to
an equivalent level of concern to those of
cmrl and vbt/vvvb2 substances (ECHA.
2014),
Opinion on an Annex AT " dossier
proposing restrictions on four phthalates
(DEHP, BBP, DBP, DIBP) (ECHA.
2017b).
Annex to the Background document to
the Opinion on the Annex AT " dossier
proposing restrictions on four phthalates
(DEHP. BBP. DBP. DIBP) (ECHA.
2017a).
ECB
European union risk assessment report:
Benzvl butvl vhthalate (BBP) (ECB.
2007).
Nov
No
No
- ECB (2007) states "The Risk Assessment Rcoort
is then peer-reviewed by the Scientific Committee
on Health and Enviromnental Risks (SCHER)
which gives its opinion on the European
Commission on the quality of the risk assessment."
See Forward for more details.
- No formal systematic review protocol employed.
EFSA
Update of the Risk Assessment of Di-
bit tvlph thai ate (DBP), Butyl-benzvl-
phthalate (BBP), Bis(2-
ethvlhexvUphthalate (DEHP), Di-
No
Yes
No
- Draft report subject to public consultation. Public
comments and EFSA's response to comments are
available at:
httt>s://doi.ore/10.2903/saefsa.2019.EN-1747
Page 86 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Agency
Assessment(s) (Reference)
External
Peer-
Review?
Public
Consultation?
Systematic
Review Protocol
Employed?
Remarks
isononylphthalate (D1NP) and Di-
isodecylphthalate (DIDP) for Use in
Food Contact Materials (EFSA. 2019).
- No formal systematic review protocol employed.
- Details regarding EFSA's strategy for identifying
new information and literature are provided on page
18 and AoDcndix B of (EFSA. 2019).
NTP-CERHR
NTP-CERHR Monograph on the
Potential Human Reproductive and
Developmental Effects of Butyl Benzyl
Phthalate (BBP) (NTP-CERHR. 2003).
No
Yes
No
- Report prepared by NTP-CERHHR Phthalates
Expert Panel and was reviewed by CERHR Core
Committee (made up of representatives of NTP-
participating agencies, CERHR staff scientists,
member of phthalates expert panel)
- Public comments summarized in Appendix III of
(NTP-CERHR. 2003)
- No formal systematic review protocol employed.
2422
Page 87 of 122
-------
2423
2424
2425
2426
2427
2428
2429
2430
2431
2432
2433
2434
2435
2436
2437
2438
2439
2440
2441
2442
2443
2444
2445
2446
2447
2448
2449
2450
2451
2452
2453
2454
2455
2456
2457
2458
2459
2460
2461
2462
2463
2464
2465
2466
2467
2468
PUBLIC RELEASE DRAFT
DECEMBER 2024
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
animals following exposure to BBP. Of these, four studies provided relevant information specifically on
developmental male reproductive toxicity following oral BBP exposure (Gray et al.. 2021; Debartolo et
al.. 2016; Schmitt et al.. 2016; Ahmad et al.. 2014). These studies are discussed above in Section 3.1.2.3
and summarized in Table 3-3 or are discussed below. The remaining three studies provide data on other
reproductive and developmental outcomes, including changes in the estrus cycle or serum estradiol,
progesterone, follicle stimulating hormone, luteinizing hormone, number of ovarian follicles,
reproductive organ weights (i.e., ovary and/or uterus), pup body weights, or non-developmental
exposure design (Integrated Laboratory Systems. 2017; Ahmad et al.. 2015; Alam and Kurohmaru.
2015). LOAELs provided by the new data set ranged from 20 to 1000 mg/kg-day. The lowest LOAEL
identified from the new literature for reproductive and developmental effects was 20 mg/kg-day based
on effects on body weight gain and reproductive organ weights (Ahmad et al.. 2015). The other
LOAELs from the remaining studies were at least an order of magnitude higher and therefore did not
offer more sensitive PODs than those discussed in Section 3.1.2.1. These studies are summarized in
Table Apx B-l, but only Ahmad et al. (2015). Integrated Laboratory Systems (2017). and Alam et al.
(2015) are discussed in the text below. The remaining studies (Gray et al.. 2021; Debartolo et al.. 2016;
Schmitt et al.. 2016; Ahmad et al.. 2014) considered for new data are discussed in Section 3.1.2.2. For
most of these new studies, EPA identified several limitations, including that of exposure dose
uncertainty, lack of linear dose-response, or other exposure design deficiencies (e.g., single-dose studies
or studies that only included relatively high doses).
Ahmad et al. (2015) evaluated the estrogenic effects of BBP 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 pubertal assay, PND 21 female rats were exposed daily to 0, 20, or
200 mg/kg-day BBP for 20 days via oral gavage, and animals were examined daily for body weights and
vaginal opening. The pubertal data are not conclusive; neither control nor BBP-exposed animals attained
puberty (i.e., first day of vaginal opening and the first day of estrus), although rats typically attain
vaginal opening by PND32. Nevertheless, a LOAEL of 20 mg/kg-day was identified based on
significantly decreased body weight gain at PND 27, 33, and 42, compared to the control group in the
pubertal assay. There was also a significant decrease in uterus weight at 20 mg/kg-day compared to the
control. Ovary weight was slightly, but not significantly, decreased at 20 mg/kg-day compared to the
control group. In the uterotrophic assay, PND20 female rats were exposed to 0, 20, or 200 mg/kg-day
BBP once per day for 3 consecutive days via gavage. Decreased uterine wet weight was observed one
day after exposure ended in the 200 mg/kg-day group. The data do not support that BBP is an estrogen
agonist. Although this is a relatively sensitive LOAEL, the study was limited by a large dose spacing,
small sample size, and the study design, which was a non-guideline female pubertal assay that did not
justify the selected exposure duration or window (i.e., PND 21 to 42), and was therefore not considered
further.
In a Good Laboratory Practice (40 CFR part 160) study by Integrated Laboratory Systems (2017).
female SD rats were exposed to 0, 250, 500, 750, or 1000 mg/kg-day BBP via oral gavage from PND 22
through 42 or 43. Beginning on PND 22, animals were examined weekly for body weights,0 and daily
for vaginal opening, as well as for estrous cyclicity. Clinical and histopathological observations were
Page 88 of 122
-------
2469
2470
2471
2472
2473
2474
2475
2476
2477
2478
2479
2480
2481
2482
2483
2484
2485
2486
2487
2488
2489
2490
2491
2492
2493
2494
2495
2496
2497
2498
PUBLIC RELEASE DRAFT
DECEMBER 2024
made in female rats at the end of exposure on either PND 42 or 43. The majority of significant BBP-
related effects occurred at levels of 750 mg/kg-day or higher, with decreased ovarian weight occurring at
750 and 100 mg/kg-day and increased medium ovarian follicles observed at 1000 mg/kg-day. The most
sensitive effect level (LOAEL) was observed at 250 mg/kg-day through a significantly decreased
number of corpora lutea. Although these endpoints speak to possible BBP-related adverse reproductive
effects in female rats, this exposure design and endpoint evaluation was not deemed specifically relevant
to the current non-cancer hazard assessment of developmental and reproductive toxicity effects
associated with phthalate syndrome outcomes. Moreover, BBP-related effects in this study were
associated with exposure during postnatal and rodent adolescence, which are assumed to not produce
effective as sensitive as the observations made during critical gestational windows of susceptibility, as
noted in other studies discussed throughout.
In a multi-cohort study by Alam et al. (2015). three week old male SD rats in experiment one were
orally gavaged with 0 or 500 mg/kg BBP and necropsied for reproductive effects assessment at 3, 12, or
24 hours post exposure. In experiment two, male SD rats were also exposed to 0 or 500 mg/kg BBP, and
assessed at 2, 4, 6, 9, and 12 days after treatment. Briefly in experiment one, adverse histopathological
changes in seminiferous tubules (i.e., reduction and/or disappearance of tubular lumens) was noted at 3
hours post exposure and thinning seminiferous epithelia and wide tubular lumina were noted by 24
hours post exposure. Increased seminiferous tubule spermatocyte cell apoptosis was also increased at 3,
12, and 24 hours post dosing. In experiment two, authors reported decreased absolute testis weight at
days 6, 9, and 12 post exposure. With regard to seminiferous tubule spermatocyte apoptosis, increased
apoptotic spermatogenic cells were observed at 2, 4, and 6 days post exposure, which was an effect that
dissipated by the ninth day observation. Although these results by Alam et al. (2015) provide limited
evidence that a single oral exposure to a high dose of BBP (500 mg/kg), this study was not conducted in
developmental model, and this did not provide data on BBP-related effects on the developing male
reproductive system, which as discussed earlier is determined the most sensitive indicator of BBP-
related effects. Further, this was a single dose-response study, and limitations were identified, including
histopathology of the testes were not quantified (i.e., incidence of pathological changes observed) and
reporting deficiencies/selection bias was noted (i.e., day 12 histopathology results not presented) (Alam
and Kurohmaru. 2015).
Page 89 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
2499 TableApx B-l. Summary of New Animal Toxicology Studies Evaluating Additional Effects on the Developmental and Reproductive
2500 System Following Exposure to BBP
Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
(Ahmad et al.. 2014)
Pregnant Albino rats (>6 dams/group)
were exposed to 0, 4, 20, or 100
mg/kg-day BBP via oral gavage from
GD 14-21. Dams were allowed to
give birth naturally, and male
offspring were sacrificed on PND 5,
25, or 75. Endpoints evaluated in F1
from PND 1-PND 75.
20/100
i serum testosterone, j
absolute weight of
epididymis and prostate,
i sperm count, j percent
motile sperm, f percent
abnormal sperm
Developmental Outcomes
- i serum testosterone (F1 adults, PND 75)
- i absolute epididymis and prostate weight (F1 adults,
PND 75)
- i sperm count, j percent motile sperm, f percent
abnormal sperm (F1 adults, PND 75)
- i pup body weight (4-100, Fl, PND 1)
- i body weight (20 and 100, PND 75)
Unaffected Outcomes
- Litter size, live/dead pups, sex ratio (PND1); Anogenital
distance (PND5 & PND25); testis descent; Viability index
(PND4); Weaning index (PND21); testicular 17(3-HSD
activity (PND 75)
(Ahmad et al.. 2015)
Female rats (6/group) were exposed
to 20 or 200 mg/kg-day BBP via
oral/gavage from PND21 - 42 (20-
day pubertal onset assay).
None/20
i BW at multiple
timepoints (PND 27, 33,
& 42), I uterus weight
Effects at 200 me/ke-dav
-j BW at multiple timepoints (PND27, 33, & 42)
-j uterus wet weight; j ovary wet weight
Considerations:
-1 BW at multiple timepoints (PND27, 33, & 42) at 20
mg/kg-day
- No changes in day of VO
Limitations
- Large dose spacing; organ weight decreases displayed flat
D-R; organ weight decreases likely a reflection of BW
changes; small sample size (n =6)
- reporting deficiencies (i.e., rat strain not reported, results
not reported for all measured outcomes, including estrus
cyclicity)
Page 90 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
Immature female rats (6/group; 20
days old) were exposed to 20 or 200
mg/kg-day BBP via oral/gavage and
sacrificed on day 4 (3-day
uterotrophic assay).
20/200
i uterus weight
Other effects:
- slight decrease in ovary wet weight (did not attain
statistical significance)
Limitations
- Large dose spacing; organ weight decreases displayed flat
D-R; organ weight decreases likely a reflection of BW
changes; small sample size (n =6)
(Gray et al.. 2021)
Pregnant Harlan SD rats (3-4
dams/group) were exposed to 0, 11,
33, 100, 300, 600, or 900 mg/kg-day
BBP via oral gavage from GD 14-18.
Dams were sacrificed and fetal tissue
collected on GD 18.
11/33
i fetal testicular mRNA
expression of
steroidogenic genes,
including InsB
Developmental Outcomes
- 1 ex vivo fetal testes testosterone production (300)
Mechanistic Outcomes
- 1 ex vivo fetal testes testosterone production (300)
- i fetal testicular expression of InsB (33) and
steroidogenic genes (Star (100), Cvpllal (33), Cypllbl
(33), Cypl 7a1 (300), Dhcr7 (11), Cypllbl (11), Hsd3b
(100), Scarbl (33))
Additional Remarks
Data are an expansion of previous dose response studies
(Furr et al.. 2014; Howdeshell et al.. 2008)
Pregnant Charles River SD rats (3-4
dams/group) were exposed to 0, 100,
300, 600, or 900 mg/kg-day BBP via
oral gavage from GD 14-18. Dams
were sacrificed and fetal tissue
collected on GD 18. (Block 78)
100/300
I ex vivo fetal testicular
testosterone production
Developmental Outcomes
- 1 ex vivo fetal testes testosterone production (300)
Mechanistic Outcomes
- i fetal testicular expression of InsB (600) and
steroidogenic genes (Star (600), Cvpllal (600),
Cvpl 7al (600), Dhcr7 (900), Cvpllbl (600), Hsd3b
(900), Scarbl (600))
(Integrated Laboratory Sv
SD rats (16/group) were exposed to 0,
250, 500, 750, or 1000 mg/kg-day
BBP via oral/gavage from PND22 to
PND42 orPND43.
None/250
i number of corpora
lutea
Effects at 250 me/ke-dav
stems. 2017)
- i corpora lutea
Effects at 500 me/ke-dav
- none
Effects at 750 or 1000 me/ke-dav
Page 91 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
- i ovary weights (11% at 750 mg/kg-day; 17% at 1,000
mg/kg-day)
-1 number of medium ovarian follicles (1,000 mg/kg-day)
Considerations:
-GLP (40 CFR part 160)
- Non-linear dose-response for number of corpora lutea (no
change at other tested doses)
- No exposure-related effects on age at VO, onset of first
estrus, mean estrus cycle length, estrus cycle regularity,
gross pathology, microscopic histology of the left ovary,
uterus, or mammary gland.
- No effect on serum TSH or T4
(Schmitt et al.. 2016)
Pregnant C57BL/6J inbred mice
(6/group) were exposed via
oral/gavage to 0 or 500 mg/kg-day
BBP from GD9-16. Serum
testosterone and estradiol evaluated at
4, 10, and 20 weeks.
None/500
Males:
| AGD at 10 and 20
weeks inFl; J, serum
testosterone
concentration in F1 at 10
and 20 weeks;
Females:
t serum testosterone & j
estradiol concentrations
in F1 at 20 weeks; t days
to VO.
Limitations:
- Large dose spacing; small sample size (n =6); single dose
study
-Exposure did not encompass the full critical window (i.e.,
GD14-19) for male antiandrogenic effects
(Alain and Kurohmaru.
2015)
Experiment 1: Male SD rats (8/arouD)
were exposed to a single dose of 0 or
500 mg/kg-day BBP via oral/gavage
and outcomes evaluated 3, 12, or 24
hours after exposure.
None/500
Histopathology of
seminiferous tubules
(reduction and/or
disappearance of tubular
lumen by 3 hours, thin
seminiferous epithelia
and wide tubular lumina
by 24 hours), t
spermatocyte cells
Limitations:
- Single dose study
-Histopathology of testes not quantified (i.e., incidence)
-Reporting deficiencies identified (e.g., day 12
histopathology not presented)
Page 92 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Reference
Brief Study Description
NOAEL/
LOAEL
(mg/kg-day)
Effect at LOAEL
Remarks
apoptosis in seminiferous
tubules at 3, 12, and 24
hours pose-dosing
Experiment 2: Male SD rats (4/arouD)
were exposed to a single dose of 0 or
500 mg/kg-day BBP via oral/gavage
and outcomes evaluated 2, 4, 6, 9, or
12 days after exposure.
None/500
t spermatocyte cell
apoptosis in seminiferous
tubules at 2, 4, and 6
days pose-dosing, J,
absolute weight of testis
at 6, 9, and 12 days post-
dosing
(Debartolo et al.. 2016)
Pregnant SD rats (22 control dams, 29
BBP dams) exposed to 10 (ig/ml BBP
from GD14 - PND23.
_Cl
i mean body weight at
PND 23, | AGD at PND
23 in both males and
females
Considerations:
- Dam body weights not provided; dose (mg/kg-day)
cannot be calculated.
- No histopathology observed via Nissl staining in cortex,
hippocampus, or cerebellum
Limitations:
-Substantial limitations in study design
- Single dose study
-Inadequate exposure characterization
Abbreviations: j = statistically significant decrease; t = statistically significant increase; AGD = Anogenital distance; BW = Body weight; E2 = (^-estradiol: F1 = First
generation offspring; FSH = follicle stimulating hormone; GD = Gestational Day; LH = Luteinizing Hormone; LOAEL = Lowest-observed-adverse-effect level;
NOAEL = no observed adverse effect level; PND = Postnatal Day; PNW = Postnatal Week; SD = Sprague-Dawley; T4 = Thyroxine; TSH = Thyroid stimulating
hormone; VO = Vaginal Opening.
" Achieved dose, including NOAEL/LOAEL dose, cannot be calculated in me/ke-dav. because dam bodv weieht and food consumption were not rcDortcd (Debartolo et
al.. 2016).
2501
2502
Page 93 of 122
-------
2503
2504
2505
2506
2507
2508
2509
2510
2511
2512
2513
2514
2515
2516
2517
2518
2519
2520
2521
2522
2523
2524
2525
2526
2527
2528
2529
2530
2531
2532
2533
2534
2535
2536
2537
2538
2539
2540
2541
2542
2543
2544
2545
2546
2547
2548
2549
PUBLIC RELEASE DRAFT
DECEMBER 2024
B.2 Neurotoxicity
EPA identified three new studies (Debartolo et al.. 2016; Schmitt et al.. 2016; Min et al.. 2014) that
provided data on neurological outcomes following exposures to BBP. Each study did not offer a more
sensitive POD than those discussed in Section 4 and/or had limitations that impacted the interpretation
of the results and were therefore not considered further. Detailed information on the study designs is
provided in Table Apx B-2.
A study by DeBartolo et al. (2016) provided data on neurobehavioral effects {i.e., fear conditioning) and
brain histopathology following developmental exposure to BBP. Pregnant SD rats were exposed to 0 or
10 |ig/mL BBP pipetted onto food pellets from GD 14 to PND 23. The authors qualitatively report no
evidence of an effect of BBP on the neocortex, hippocampus, and cerebellum in experiment 1,
visualized by Nissl staining. The authors report decreased duration of freezing after the conditional
stimulus {i.e., an audible tone that precedes and electric shock) in both sexes, which they attribute to
learning and memory impairments due to BBP. However, substantial limitations of this study impact the
interpretation of the results and contribute to uncertainty in the data set. Reporting deficiencies were
identified for the histopathology data (i.e., histopathological changes only reported qualitatively and
insufficient details in methods). For the fear conditioning experiments, limitations in the experimental
design further impact the interpretation of the results. The largest limitation of this study includes
substantial uncertainty regarding internal dose exposure through lack of adequate exposure
characterization, which is due to BBP stock solution (10 jag/mL) being pipetted onto sweetened food
pellets fed to pregnant dams; however, it was not reported that exposure was according to dam body
weight, and there was no measure of food consumption noted in methods. Treatment for the control and
BBP-exposed groups was also stated to occur ~5 to 7 days during gestation, meaning not all exposed
animals may have had the same exposure time duration.
In a behavioral assessment by Schmitt et al. (2016). pregnant mice were exposed to 0 or 500 mg/kg-day
BBP via gavage from GD 9 to 16 and F1 offspring were subjected to running wheel testing from
postnatal week 8 to 20. Authors reported reductions in voluntary physical activity (i.e., running wheel
activity) in a pre-and perinatal exposure study in mice, which may reflect locomotor deficits. Male and
female mice that had been exposed to BBP ran significantly less distance by postnatal week 20
compared to controls. However, limitations of this study include that it was a single dose study, which
precludes its use in understanding dose-response relationships. Additionally, the study does not offer a
more sensitive POD than those provided for effects of the developing male reproductive system and was
not considered further.
Min et al. (2014) reported evidence of alterations in neurological health outcomes in male mice
administered 0, 50, 250, or 1250 mg/kg-day BBP via gavage for 14 days (age at exposure not specified).
Following exposure, the mice underwent swim trials in the Morris Water Maze, as well as trials in the
forced swim test and tail suspension test. Brain tissue was collected to measure neurotransmitter (5-HT)
levels as well as measurements of oxidative stress and a histopathologic evaluation of the hippocampal
region of the brain. Reactive oxygen species were also measured via the DCF-DA assay, along with
glutathione content and phosphorylated CREB to provide indices of oxidative stress. In the Morris
Water Maze, increased escape latency was observed in mice exposed to BBP at dosages of 250 mg/kg-
day and higher. Decreased time spent swimming in the target quadrant were only observed at the highest
dose (1250 mg/kg-day), suggesting impaired memory in mice from the highest exposure group.
Increased time spent immobile in the tail suspension and forced swim tests was observed in mice from
the highest exposure group, while mice from the 250 mg/kg-day group exhibited increased time spent
Page 94 of 122
-------
2550
2551
2552
2553
2554
2555
2556
2557
2558
2559
2560
2561
2562
2563
2564
PUBLIC RELEASE DRAFT
DECEMBER 2024
immobile in the tail suspension test, implying affected motor function. A dose-dependent decrease
hippocampal levels of 5-HT and phosphorylated CREB was observed, including significant decreases in
the endpoints at all tested BBP doses. Other findings were consistent with increased oxidative stress,
including increased reactive oxygen species in the brain in parallel with decreased glutathione content in
the brain in mice exposed to BBP at doses of 250 mg/kg-day BBP or higher. Altogether, data from Min
et al. (2014) provide LOAEL values of 50 mg/kg-day based on neurological effects following short-term
exposure to BBP in one sex of one strain of mice.
In sum, the current database supporting neurological effects of BBP is too limited, especially compared
to that of more sensitive male reproductive and developmental effects. Additionally, given the
aforementioned limitations of this study, EPA did not consider these effects or studies for dose-response
assessment or for use in extrapolating human risk in Section 4.
TableApx B-2. Summary of New Animal Toxicology Study Evaluating Effects on the Nervous
System Following Exposure to BBP
Reference
Brief Study
Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at
LOAEL
Remarks
(Min etal.. 2014)
Male SPF Kunming
mice (6/group) were
exposed to 0, 50, 250,
or 1250 mg/kg-day
BBP via gavage for
14 days. Mice were
trained for MWM
from days 1-11 and
trials conducted on
day 14. FST and TST
conducted on day 14.
Mice sacrificed one
day after exposure
and brains collected
for histopathological
evaluation and
measurements of
oxidative stress.
None/50
J,5-HT, IpCREB
Effects at 250 me/ke-dav
-1 Average escape latency
for 11 days (MWM)
-1 Immobile time in TST
- i Brain GSH content
- |ROS (brain)
Effects at 1250 me/ke-dav
- i Swimming time in
target quadrant (MWM)
concentration
| Immobile time in FST &
TST
-1 Average escape latency
for 11 days (MWM)
- i Brain GSH content
- |ROS (brain)
Limitations
-Qualitative histopathology
- GSH level interpretation
is difficult without
GSH:GSSH ratio
(Schmitt et al.. 2016)
Pregnant C57BL/6J
inbred mice (6/group)
were exposed via
oral/gavage to 0 or
500 mg/kg-day BBP
from GD9-16.
Running Wheel
activity monitored
from PNW8 -20.
None/500
I Performance
on voluntary
wheel running
(J,distance and
duration of
exercise) in male
and female mice
onPNW20
Unaffected Outcomes:
-Running wheel speed
Limitations:
-Single dose study
Page 95 of 122
-------
2565
2566
2567
2568
2569
2570
2571
2572
2573
2574
2575
2576
2577
2578
2579
2580
2581
2582
2583
2584
2585
2586
PUBLIC RELEASE DRAFT
DECEMBER 2024
Reference
Brief Study
Description
NOAEL/ LOAEL
(mg/kg-day)
Effect at
LOAEL
Remarks
(Debartolo et al..
2016)
Pregnant SD rats (11
control dams, 12 BBP
dams) exposed to 10
ug/ml BBP from
GD14 - PND23.
_a
i Duration of
freezing in
males and
females on
PND65 after
conditional
stimulus
(audible tone)
Considerations:
- Dam body weights not
provided so dose (mg/kg-
day) cannot be calculated.
- No histopathology
observed via Nissl staining
in cortex, hippocampus, or
cerebellum
Limitations:
-Substantial limitations in
study design
- Single dose study
-Inadequate exposure
characterization
Abbreviations: j = statistically significant decrease; t = statistically significant increase; 5-HT = 5-hydroxytryptamine;
FST = Forced Swim test; GD = Gestational Day; GSH = Glutathione; LOAEL = Lowest-observed-adverse-effect level;
MWM = Morris Water Maze; NOAEL = No-observed-adverse-effect level; pCREB = Phosphorylated cyclic adenosine
monophosphate (cAMP) response element binding protein (CREB); PND = Postnatal day; PNW = Postnatal week; ROS =
Reactive oxygen species; SD = Sprague-Dawley; TST = Tail suspension test.
" Achieved dose, including NOAEL/LOAEL dose, cannot be calculated in mg/kg-day, because dam body weight and food
consumption were not reported (Debartolo et al.. 2016).
B.3 Immune adjuvant effects
EPA identified one new study that provided data on immune adjuvant effects following exposure of
adult female BALB/cByJ mice (4 to 6/group) to 0 or 3 |ig/ml BBP via drinking water (Jahreis et al..
2018). Exposure began one week prior to mating and lasted through either delivery or until weaning of
the F1 (PND 21). The Floffpsring were immunized with ovalbumin and adjuvant (alum and MgOH) via
i.p. injection on days 1 and 14 prior to receiving intranasal administration of ovalbumin on days 14-16
and 21-23. Female F1 were mated to un-exposed males, and the immunization paradigm was repeated in
the F2. Outcomes evaluated included measurement of cell numbers in BAL, lung histology, IgE levels,
cytokine levels (e.g., IFN-y, IL-17, IL10, IL-4), differentially methylated regions of DNA, and a
regulatory T cell suppression assay. However, there were substantial limitations that impact the
interpretation of the results of this study, particularly dose characterization. Indeed, the authors
estimated the dose ranged from 0.48 to 0.6 mg BBP/kg/body weight per day (assuming 4-5 mL/d
drinking water intake containing 3 ug/mL BBP) but did not report water intake for this drinking water
study. Additionally, this was a single dose study that used a small number of animals. Ultimately, the
study was not considered further.
B.4 Renal
EPA identified two studies that provided more data for renal outcomes following exposure to BBP
(Integrated Laboratory Systems. 2017; Nakagomi et al.. 2017). In this study, male and female SD rats
were exposed to 0 or 500 mg/kg-day BBP for 14 days via gavage. Necropsies were performed on male
(6 to 9/group) and female (5 to 9/group) rats at the time of sacrifice and kidneys were collected for
histopathological evaluation. The data are limited to a qualitative description by authors. The authors
reported, "light histopathologic changes in kidneys and marked changes in the hormone status of rats
Page 96 of 122
-------
2587
2588
2589
2590
2591
2592
2593
2594
2595
2596
2597
2598
2599
2600
2601
2602
2603
2604
2605
2606
PUBLIC RELEASE DRAFT
DECEMBER 2024
exposed to BBP at 500 mg kg-day". Kidney weights were not reported for BBP-exposed animals.
Endpoints from this study were not considered further due to limitations that impact the interpretation of
the results, namely qualitative reporting of histopathology data, small sample size, and use of a single,
high dose of BBP. Additionally, the study by Integrated Laboratory Systems (2017) (described in
Section B. 1) reported decreased blood urea nitrogen at 250 mg/kg-day BBP and higher in SD rats
(16/group) exposed to 0, 250, 500, 750, or 1000 mg/kg-day BBP via oral/gavage from PND 22 to PND
42. However, there was no dose-response above 500 mg/kg-day performed in the Nakagomi et al.
(2017) assessment, and other clinical chemistry markers of renal dysfunction noted by Integrated
Laboratory Systems (2017) were only observed at doses of 750 mg/kg-day and higher (e.g., increased
serum phosphorus).
B.5 Hepatic
EPA identified one study that provided more data for hepatic outcomes (liver weight and liver
histopathology data) following exposure to BBP (Nakagomi et al.. 2017). which was described above in
Section B.4. Liver weights were significantly decreased in male rats from the BBP group compared to
control (approximately 14% increase). No significant change was reported for female rats. No
histopathological effects of the liver were observed for rats exposed to BBP, as reported qualitatively in
the text with one representative micrograph showing a section of the liver from a control and exposed rat
(sex not specified). Endpoints from this study were not considered further due to limitations that impact
the interpretation of the results, namely qualitative reporting of histopathology data, small sample size,
and use of a single, high dose of BBP.
Page 97 of 122
-------
2607
2608
2609
2610
2611
2612
2613
2614
2615
2616
2617
2618
2619
2620
2621
2622
2623
2624
2625
2626
2627
2628
2629
2630
2631
2632
2633
2634
2635
2636
2637
2638
2639
2640
2641
PUBLIC RELEASE DRAFT
DECEMBER 2024
Appendix C FETAL TESTICULAR TESTOSTERONE AS AN
ACUTE EFFECT
One experimental animal model study is available that investigates the antiandrogenic effects of BBP
following single dose, acute exposure. In a multi-cohort study by Alam et al. (2015), three week old
male SD rats were orally gavaged with 0 or 500 mg/kg BBP and necropsied for reproductive effects
assessment at various timepoints, including 3, 12, or 24 hours post exposure, and 2, 4, 6, 9, or 12 days
after treatment. Briefly, adverse histopathological changes in seminiferous tubules (i.e., reduction and/or
disappearance of tubular lumens) was noted at 3 hours post exposure and thinning seminiferous epithelia
and wide tubular lumina were noted by 24 hours post exposure. Increased seminiferous tubule
spermatocyte cell apoptosis was also increased at 3, 12, and 24 hours post dosing. In experiment two,
authors reported decreased absolute testis weight at days 6, 9, and 12 post exposure. With regard to
seminiferous tubule spermatocyte apoptosis, increased apoptotic spermatogenic cells were observed at 2,
4, and 6 days post exposure, which was an effect that dissipated by the ninth day observation (Alam and
Kurohmaru. 2015). Moreover, there are studies of dibutyl phthalate (DBP) available (toxicologically
similar to BBP) that indicate a single acute exposure during the critical window of development (i.e.,
GD 14 to 19) can reduce fetal testicular testosterone production and disrupt testicular steroidogenic gene
expression. Two studies were identified that demonstrate single doses of 500 mg/kg DBP can reduce
fetal testicular testosterone and steroidogenic gene expression. Johnson et al. (2012; 2011) gavaged
pregnant SD rats with a single dose of 500 mg/kg DBP on GD 19 and observed reductions in
steroidogenic gene expression in the fetal testes three (Cypl7al) to six (Cypllal, Star) hours post-
exposure, while fetal testicular testosterone was reduced starting 18 hours post-exposure. Similarly,
Thompson et al. (2005) reported a 50 percent reduction in fetal testicular testosterone 1-hour after
pregnant SD rats were gavaged with a single dose of 500 mg/kg DBP on GD 19, while changes in
steroidogenic gene expression occurred 3 (Star) to 6 (Cypllal, Cypl7al, Scarbl) hours post-exposure,
and protein levels of these genes were reduced 6 to 12 hours post-exposure. Additionally, studies by
Carruthers et al. (2005) further demonstrate that exposure to as few as two oral doses of 500 mg/kg DBP
on successive days between GDs 15 to 20 can reduce male pup AGD, cause permanent nipple retention,
and increase the frequency of reproductive tract malformations and testicular pathology in adult rats that
received two doses of DBP during the critical window.
In summary, single dose acute studies of BBP (Alam and Kurohmaru. 2015) and DBP (Johnson et al..
2012; Johnson et al.. 2011) provide evidence to support use of effects on the male reproductive system,
specifically testicular histopathological changes and reduced fetal testosterone, as an acute effect.
However, the database is limited to just a few studies that test relatively high (500 mg/kg) single doses
of BBP or DBP.
Page 98 of 122
-------
2642
2643
2644
2645
2646
2647
2648
2649
2650
2651
2652
2653
2654
2655
2656
2657
2658
2659
2660
2661
2662
2663
2664
2665
2666
2667
2668
2669
2670
2671
2672
2673
2674
2675
2676
2677
2678
2679
2680
2681
2682
2683
2684
2685
2686
2687
PUBLIC RELEASE DRAFT
DECEMBER 2024
CALCULATING DAILY ORAL HUMAN
EQUIVALENT DOSES AND HUMAN EQUIVALENT
CONCENTRATIONS
For BBP, all data considered for PODs are obtained from oral animal toxicity studies in rats. Because
toxicity values for BBP 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. However, EPA did not identify existing PBPK models for BBP or other BBP
information that would be useful in conducting chemical-specific quantitative dose extrapolation. In the
absence of such data, EPA relied on the guidance from U.S. EPA (2011 c), 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 Equation C-l.
Equation C-l. Dosimetric Adjustment Factor
/BWa\1/4
DAF = fenr)
Where:
DAF = Dosimetric adjustment factor (unitless)
BWa = Body weight of species used in toxicity study (kg)
BWh = Body weight of adult human (kg)
U.S. EPA (2011c). presents DAFs for extrapolation to humans from several species. However, because
those DAFs used a human body weight of 70 kg, EPA has updated the DAFs using a human body
weight of 80 kg for the DINP risk evaluation (U.S. EPA. 2011a). EPA used the body weights of 0.25 kg
for rats, as presented in U.S. EPA (2011c). The resulting DAFs for rats is 0.236.
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 C-l, EPA adjusts the POD to obtain
the HED using Equation C-2:
Equation C-2. Daily Oral Human Equivalent Dose
HEDDaUy = 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
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 using Equation C-3 as follows:
Appendix D
Page 99 of 122
-------
2688
2689
2690
2691
2692
2693
2694
2695
2696
2697
2698
2699
2700
2701
2702
2703
2704
2705
2706
2707
2708
2709
2710
2711
2712
2713
2714
2715
2716
2717
2718
2719
2720
2721
2722
2723
2724
2725
2726
2727
2728
2729
PUBLIC RELEASE DRAFT
DECEMBER 2024
Equation C-3. Extrapolating from Oral HED to Inhalation HEC
iirn r BW a
HEC Daily, continuous ~ HED Daily ^ (f D P n )
JA^ * Lj Lf £
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 C-4 presents the conversion of the
HEC from mg/m3 to ppm.
Equation C-4. Converting Units for HECs (mg/m3 to ppm)
mg 24.45
X ppm = Y 7 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 BBP = 312.37 g/mol)
D.l BBP 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 NOAEL = 50 mg/kg/day,
and the critical effect is developmental toxicity (i.e., decreased AGD) in gestationally exposed CD rats
(Tyl et al.. 2004). EPA used Equation C-l to determine a DAF specific to rats (0.236), which was in turn
used in the following calculation of the daily HED using Equation C-2:
mq mq
11.8 = 50- x 0.236
kg day kg day
EPA then calculated the continuous HEC for an individual at rest using Equation C-3:
mq mq 80 kq
64-2^ = n-Qi^rd^x( )
g y 0.6125?-* 24 hr
hr
Equation C-4 was used to convert the HEC from mg/m3 to ppm:
Page 100 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
2730
ma 24.45
2731 5.03 ppm = 64.2 | x
2732
m3 312.37
Page 101 of 122
-------
2733
2734
2735
2736
2737
2738
2739
2740
2741
2742
2743
2744
2745
2746
2747
2748
2749
2750
2751
2752
2753
2754
2755
2756
2757
2758
2759
2760
2761
2762
2763
2764
2765
2766
2767
2768
2769
2770
2771
2772
2773
2774
2775
2776
2777
PUBLIC RELEASE DRAFT
DECEMBER 2024
Appendix E CONSIDERATIONS FOR BENCHMARK RESPONSE
(BMR) SELECTION FOR REDUCED FETAL
TESTICULAR TESTOSTERONE
E.l Purpose
EPA has conducted an updated meta-analysis and BMD analysis of decreased fetal rat testicular
testosterone (U.S. EPA. 2024c). During the July 2024 SACC peer-review meeting of the draft risk DIDP
and draft human health hazard assessments for DINP, the SACC recommended that EPA should clearly
state its rational for selection of BMR levels evaluated for decreases in fetal testicular testosterone
relevant to the single chemical assessments (U.S. EPA. 20240. 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. 2024c).
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. 2024c).
Page 102 of 122
-------
2778
2779
2780
2781
2782
2783
2784
2785
2786
2787
2788
2789
2790
2791
2792
2793
2794
2795
2796
2797
2798
2799
2800
2801
2802
2803
2804
2805
2806
2807
2808
2809
2810
2811
2812
2813
2814
2815
2816
2817
2818
2819
2820
2821
2822
PUBLIC RELEASE DRAFT
DECEMBER 2024
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 MOA for phthalate syndrome, which is described elsewhere (U.S. EPA. 2023a) and in Section
3.1.2.1 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 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 percent 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).
Page 103 of 122
-------
2823
2824
2825
2826
2827
2828
2829
2830
2831
2832
2833
2834
2835
2836
2837
2838
2839
2840
2841
2842
2843
2844
2845
2846
2847
2848
2849
2850
2851
2852
2853
2854
2855
2856
2857
2858
2859
2860
2861
2862
2863
2864
2865
2866
PUBLIC RELEASE DRAFT
DECEMBER 2024
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 be a reflection of the larger
database of studies and wider range of endpoints evaluated for DEHP, compared to DBP and
DCHP.
NASEM (2017) modeled a BMR of 40 percent using the following justification: "previous
studies hcn'e shown that reproductive-tract malformations were seen in male rats when fetal
testosterone production was reduced by about 40% ^Grav et al.. 2016; Howdeshell et al.. 2015)."
However, publications supporting a 40 percent response level are relatively narrow in scope and
assessed the link between reduced fetal testicular testosterone in SD rats on GD 18 and later life
reproductive tract malformations in F1 males. More specifically, Howdeshell et al. (2015) found
reproductive tract malformations in 17 to 100 percent of F1 males when fetal testosterone on GD
18 was reduced by approximately 25 to 72 percent, while Gray et al. (2016) found dose-related
reproductive alterations in F1 males treated with dipentyl phthalate (a phthalate not currently
being evaluated under TSCA) when fetal testosterone was reduced by about 45 percent on GD
18. Although NASEM modeled a BMR of 40 percent based on biological considerations, there is
no scientific consensus on the biologically significant response level and no other authoritative
or regulatory agencies have endorsed the 40 percent response level as biologically significant for
reductions in fetal testosterone.
BMDL40 values for DEHP, DBP, DIBP, BBP, DCHP, and DINP are above the PODs selected for
use in risk characterization for each phthalate by 3X to 25.4X (Table Apx 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.
Page 104 of 122
-------
2867
2868
PUBLIC RELEASE DRAFT
DECEMBER 2024
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. 2024e)
DBP
BMDL5 = 9
(J, fetal testicular testosterone)
30
(t Testicular Pathology)
14 [9, 27]
29 [20, 54]
149 [101, 247]
(U.S. EPA. 2024e)
DIBP
BMDL5 = 24
(J, fetal testicular testosterone)
125
(t Testicular Pathology)
_b
55 [NA, 266]*
279 [136, 517]
(U.S. EPA. 2024h)
BBP
NOAEL = 50
(phthalate syndrome-related effects)
100
(4 AGD)
_b
_b
284 [150, 481]
(U.S. EPA. 2024d)
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. 2024f)
DINP
BMDL5 = 49
(J, fetal testicular testosterone)
600
(J, sperm motility)
74 [47, 158]
152 [97, 278]
699 [539, 858]
(U.S. EPA. 2025h)
Abbreviations: AGD = Anogenital distance; BMD = Benchmark dose; BMDL5 = Lower 95% confidence limit on BMD; CI = 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.
2869
Page 105 of 122
-------
2870
2871
2872
2873
2874
2875
2876
2877
2878
2879
2880
2881
2882
2883
2884
2885
2886
2887
2888
2889
2890
2891
2892
2893
2894
2895
2896
2897
2898
2899
2900
2901
2902
2903
2904
2905
2906
2907
2908
2909
2910
PUBLIC RELEASE DRAFT
DECEMBER 2024
Appendix F BENCHMARK DOSE MODELING OF FETAL
TESTICULAR TESTOSTERONE
EPA conducted BMD modeling of ex vivo fetal testicular testosterone data from four gestational
exposure studies of BBP reported in three publications (Gray et al.. 2021; Furr et al.. 2014; Howdeshell
et al.. 2008).
The BMD modeling for continuous data was conducted with the EPA's BMD software (BMDS 3.3.2).
All standard BMDS 3.3.2 continuous models that use maximum likelihood optimization and profile
likelihood-based confidence intervals were used in this analysis. Standard forms of these models
(defined below) were run so that auto-generated model selection recommendations accurately reflect
current EPA model selection procedures EPA's benchmark Dose Technical Guidance (U.S. EPA. 2012).
BMDS 3.3.2 models that use Bayesian fitting procedures and Bayesian model averaging were not
applied in this work.
Standard BMDS 3.3.2 Models Applied to Continuous Endpoints:
Exponential 3-restricted (exp3-r)
Exponential 5-restricted (exp5-r)
Hill-restricted (hil-r)
Polynomial Degree 3-restricted (ply3-r
Polynomial Degree 2-restricted (ply2-r)
Power-restricted (pow-r)
Linear-unrestricted (lin-ur)
EPA evaluated BMR levels of 1 control standard deviation (1 SD) and 5, 10, and 40 percent relative
deviation. Model fit was judged consistent with EPA's benchmark Dose Technical Guidance (U.S. EPA.
2012). An adequate fit was judged based on the %2 goodness-of-fit p-value (p > 0.1), magnitude of the
scaled residuals in the vicinity of the BMR, and visual inspection of the model fit. In addition to these
three criteria forjudging adequacy of model fit, a determination was made as to whether the variance
across dose groups was constant. If a constant variance model was deemed appropriate based on the
statistical test provided in BMDS (i.e., Test 2; p-value > 0.05 [note: this is a change from previous
versions of BMDS, which required variance p-value > 0.10 for adequate fit]), the final BMD results
were estimated from a constant variance model. If the test for homogeneity of variance was rejected
(i.e., p-value < 0.05), the model was run again while modeling the variance as a power function of the
mean to account for this nonconstant variance. If this nonconstant variance model did not adequately fit
the data (i.e., Test 3; p-value < 0.05), the data set was considered unsuitable for BMD modeling. Among
all models providing adequate fit, the lowest BMDL was selected if the BMDLs estimated from
different adequately fitting models varied >3-fold; otherwise, the BMDL from the model with the lowest
AIC was selected.
Table Apx F-l summarizes BMD modeling results for reduced ex vivo fetal testicular testosterone data,
while more detailed BMD model results are provided in Appendices F.l through F.4.
Page 106 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
2911 TableApx F-l. Summary of BMD Model Results for Decreased Ex Vivo Fetal Testicular
2912 Testosterone
Data Set
BMR
Best-Fit Model
(Variance)
BMD
(mg/kg-
day)
BMDL
(mg/kg-
day)
Notes
Appendix
Containing
Results
(Gray et al.. 2021;
Howdeshell et al.. 2008)
5%
Exponential 3
(constant)
138
81
F.l
(Gray et al.. 2021)
(Howdeshell et al.. 2008)
5%
a
a
a
No models adequately
fit the data set"
F.2
(Furret al.. 2014)
(Block 36 rats)
-
-
-
-
No models adequately
fit the data set
F.3
(Furret al.. 2014)
(Block 37 rats)
-
-
-
-
No models adequately
fit the data set
F.4
BMD = benchmark dose; BMDL = benchmark dose lower limit; BMR = benchmark response
" Although the polynomial degree 2 model (non-constant variance) provided an adequate statistical fit and supported BMD5
and BMDL5 values of 48 and 47 mg/kg-day, respectively, the model provided a poor visual fit, particularly in the low end
range of the dose-response curve. Therefore, EPA did not further consider the derived BMD and BMDL values.
2913
2914
2915
2916
F.l BMD Model Results of Howdeshell et al. (2008)
Dose
(mg/kg-day)
N
(# of litters)
Mean
Standard
Deviation
Notes
0
9
3.45
0.4500
Data from Table 6 in (Howdeshell et al.. 2008)
100
4
3.66
0.5200
300
5
2.68
0.6037
600
2
1.18
0.3394
900
2
0.34
0.1980
2917
Page 107 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
2918
Models"
Restrictionb
Variance
BMR = 5%
BMR = 10%
BMR = 40%
BMR = 1 SD
P Value
AIC
BMDS
Recommends
BMDS
Recommendation
Notes
BMD
BMDL
BMD
BMDL
BMD
BMDL
BMD
BMDL
Exponential
3
Restricted
Constant
137.878
80.51109
194.8991
129.4477
416.3536
350.4237
218.8987
149.5615
0.4883665
34.60071824
Viable -
Recommended
Lowest AIC
Exponential
5
Restricted
Constant
146.2356
81.66467
201.768
130.8391
411.2116
345.5512
224.6711
150.9235
0.2511005
36.48451024
Viable -
Alternate
Hill
Restricted
Constant
159.9532
74.84564
210.1501
126.9003
401.1389
335.9305
230.161
149.4785
0.2959987
36.25947488
Viable -
Alternate
Polynomial
Degree 3
Restricted
Constant
66.44438
44.68631
129.0148
89.37277
449.0893
357.795
169.5433
106.3084
0.0935472
37.90591815
Questionable
Goodness of fit p-
value < 0.1
Polynomial
Degree 2
Restricted
Constant
64.35011
44.69158
125.49
89.38335
444.2024
357.4874
164.8979
106.3203
0.0935793
37.90523307
Questionable
Goodness of fit p-
value < 0.1
Power
Restricted
Constant
85.29039
45.81249
148.472
91.65708
449.9177
366.6237
185.5577
111.5766
0.1375952
37.13421803
Viable -
Alternate
Linear
Unrestricted
Constant
49.94472
43.80531
99.88943
87.6111
399.5577
350.4423
133.5974
102.2819
0.1370596
36.69388381
Viable -
Alternate
Abbreviations: AIC = Akaike information criterion; BMD = Benchmark dose; BMDL = Benchmark dose lower limit; BMDS = Benchmark dose software; BMR = Benchmark response; SD = Standard deviation.
" Selected Model (bolded and shaded gray).
b Restrictions defined in the BMDS 3.3 User Guide.
2919
Page 108 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
2920
2921
2922
5
4.5
4
3.5
E
3
o
2.5
+-<
LO
3
(J
2
1.5
1
0.5
0
Frequentist Exponential Degree 3 Model with BMR of 0.05 Added
Risk for the BMD and 0.95 Lower Confidence Limit for the BMDL
C
)
Estimated Probability
Response at BMD
O Data
BMD
BMDL
100
200
300
400 500
mg/kg-day
600
700
800
900
FigureApx F-l. Frequentist Exponential Degree 3 Model of Howdeshell et al. (2008) data
User Input
Info
Model
frequentist Enponential degree 3
Model Restriction
Restricted
~ataset Name
Howdeshell (2008) - BBR Testosterone
User notes
[Add user notes here]
Dose-Response Modi
M[dose] = a " enp(±1" (b " dosefd)
Variance Model
Var[i] = alpha
Model Options
BMR Type
Rel. Dev.
BMRF
0.05
Tail Probability
-
Confidence Level
0.95
Distribution Type
Normal
Variance Type
Constant
Model Data
Dependent Variable
mg^kg-day
Independent Variable
[Custom]
Total # of Observation
5
Adverse Direction
Automatic
2924 Figure Apx F-2. User Input of Frequentist Exponential Degree
2925 3 Model of Howdeshell et al. (2008) Data
2926
Page 109 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Model Results
Benchmark Dose
BMD
137.8779883
BMDL
80.51103103
BMDU
214.4853617
AIC
34.60071824
Test 4 P-value
0.48838643
D.O.F.
2
Model Parameters
# of Parameters
4
Variable
Estimate
Std Error
Lower Conl
Upper Conl
a
3.528232323
0.12365373
3.274116
3.782343
b
0.00173887
1.60E-04
0.001425
0.002053
d
2.073743645
4.13E-01
1.260538
2.638361
log-alpha
-1.628753523
3.02E-01
-2.2137
-1.03781
Goodness of Fit
Dose
Size
Estimated
Calo'd
Observed
Estimated
Calo'd
Observe
Scaled
Median
Median
Mean
3D
3D
dSD
Residual
0
3
3.52823233
3.45
3.45
0.442315
0.45
0.45
-0.523831
100
4
3.43665147
3.66
3.66
0.442315
0.52
0.52
1.0085384
300
5
2.7248325
2.68
2.68
0.442315
0.6037
0.60374
-0.226641
600
2
1.18363214
1.18
1.18
0.442315
0.3334
0.33341
-0.011537
300
2
0.27675032
0.34
0.34
0.442315
0.138
0.13733
0.1355662
2927
2928
2929
2930
2931
2932
Likelihoods of Interest
Model
Log Likelihood"
#of
Parameters
AIC
A1
-12.58366337
6
37.16734
A2
-10.44153567
10
40.88313
A3
-12.58366337
6
37.16734
fitted
-13.30035912
4
34.60072
R
-33.95385731
2
71.30771
' Includes additive constant of -20.21665. This constant w as
Tests of Interest
Test
2"Log(Likeliho
od Ratio)
Test df
p-value
1
47.02452329
8
<0.0001
2
4.284148604
4
0.368314
3
4.284148604
4
0.368314
4
1.433378302
2
0.488366
not included in the LL derivation prior to BMD5 3.0.
FigureApx F-3. Model Results of Frequentist Exponential Degree 3 Model of Howdeshell et al.
(2008) Data
F.2 BMP Model Results of Gray et al. (2021)
Table Apx F-4. Ex Vivo Fetal Rat Testicular Testosterone Data (Gray et al., 2021)
Dose
(mg/kg-day)
N
(# of litters)
Mean
Standard
Deviation
Notes
0
3
7.848667
1.202189
100
3
8.409444
1.32299
Page 110 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Dose
(mg/kg-day)
N
(# of litters)
Mean
Standard
Deviation
Notes
300
3
4.878667
0.634855
Data for Block 78 rats reported in
SiiDDlcmcntarv Data file associated with (Gray et
al.. 2021)
600
3
2.921333
1.205674
900
3
0.603111
0.223092
2933
Page 111 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
2934
al.. 2021)
Models "
Restriction 6
Variance
BMR = 5%
BMR = 10%
BMR = 40%
BMR = 1 SD
P Value
AIC
BMDS
Recommends
BMDS Recommendation
Notes
BMD
BMDL
BMD
BMDL
BMD
BMDL
BMD
BMDL
Exponential 3
Restricted
Constant
79.44
36.49647
125.8696
68.52385
345.3638
264.8549
140.8441
77.63085
0.0822269
49.78570739
Questionable
Goodness of fit p-value <0.1
Exponential 5
Restricted
Constant
79.44
36.49897
125.8696
68.52385
345.3638
264.8549
140.8441
77.63085
0.0253979
51.78570739
Questionable
Goodness of fit p-value <0.1
Hill
Restricted
Constant
88.91208
29.98603
133.0772
61.37096
338.0458
259.4567
145.5466
69.51906
0.0344126
51.26324531
Questionable
Goodness of fit p-value <0.1
BMDL 3x lower than lowest
non-zero dose
Polynomial
Degree 3
Restricted
Constant
47.34491
42.7451
94.68978
85.49016
378.7592
341.9607
117.7362
87.2974
0.0896355
49.28984585
Questionable
Goodness of fit p-value <0.1
Polynomial
Degree 2
Restricted
Constant
47.33616
42.74652
94.67232
85.49304
378.6893
341.9722
117.7338
87.29664
0.0896369
49.28980913
Questionable
Goodness of fit p-value <0.1
Power
Restricted
Constant
47.32532
42.74833
94.65064
85.49666
378.6026
341.9866
117.6541
87.29526
0.0896377
49.28978915
Questionable
Goodness of fit p-value <0.1
Linear
LTnrestricted
Constant
47.32533
42.74833
94.65064
85.49666
378.6025
341.9866
117.6541
87.29417
0.0896377
49.28978915
Questionable
Goodness of fit p-value <0.1
Exponential 3
Restricted
Non-
Constant
132.0526
48.65299
188.3712
84.90216
410.5209
284.9006
262.7082
110.1942
0.0426644
48.1908878
Questionable
Goodness of fit p-value <0.1
Exponential 5
Restricted
Non-
Constant
132.0526
48.65299
188.3711
84.90215
410.5208
284.9006
262.7082
110.1942
0.0120141
50.19088786
Questionable
Goodness of fit p-value <0.1
Hill
Restricted
Non-
Constant
61.41628
27.89238
108.6075
55.80617
365.3485
263.4583
172.7294
84.31022
0.0447518
47.91007314
Questionable
Goodness of fit p-value <0.1
BMDL 3x lower than lowest
non-zero dose
Polynomial
Degree 3
Restricted
Non-
Constant
48.38865
46.50351
96.77729
93.00702
387.1092
372.0281
170.7388
117.151
0.2048224
44.46722414
Viable - Alternate
Polynomial
Degree 2
Restricted
Non-
Constant
48.4699
46.5102
96.93983
93.02042
387.7593
372.0812
177.4433
117.4939
0.2086964
44.42271558
Viable -
Recommended
Lowest AIC
Power
Restricted
Non-
Constant
49.9447
48.7917
99.20589
92.64517
391.4114
372.6303
189.1184
116.9643
0.0998621
46.49003582
Questionable
Goodness of fit p-value <0.1
Linear
LTnrestricted
Non-
Constant
48.4699
46.71352
96.93983
93.42708
387.7593
373.7082
177.4433
117.4921
0.2086964
44.42271558
Viable - Alternate
Abbreviations: AIC = Akaike information criterion; BMD = Benchmark dose; BMDL = Benchmark dose lower limit; BMDS = Benchmark dose software; BMR = Benchmark response; SD = Standard deviation.
" Selected Model (bolded and shaded gray).
b Restrictions defined in the BMDS 3.3 LTser Guide.
2935
Page 112 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
2936
2937
2938
2939
£
o
+-<
LO
3
u
12
10
8
6
4
2
0
Frequentist Polynomial Degree 2 Model with BMR of 0.05 Added
Risk for the BMD and 0.95 Lower Confidence Limit for the BMDL
(
)
o
100
200
300
400 500
mg/kg-day
600
700
800
900
Estimated Probability
Response at BMD
O Data
BMD
BMDL
FigureApx F-4. Frequentist Exponential Degree 3 Model of Gray et al. (2021) Data
User Input
Info
Model
frequentist Polynomial degree 2
Model Restriction
Restricted
Dataset Name
5ray(2G21) - BBP Testosterone (Block 73 rats)
User notes
[Add user notes here]
~ose-Response Modi
M[dose] = g + bTdose + b2"dose''2 +...
Variance Model
Var[i] = alpha" mean[i]A rho
Model Options
BMR Type
Rel. Dev.
BMRF
0.05
Tail Probability
-
Confidence Level
0.35
Distribution Type
Normal
Variance Type
Non-Constant
Model Data
Dependent Variable
mg/kg-day
Independent Variable
[Custom]
Total # of Observation
5
Adverse Direction
Automatic
2941 Figure Apx F-5. User Input of Frequentist
2942 Exponential Degree 3 Model of Gray et al. (2021) Data
2943
Page 113 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Model Results
Benchmark Dose
BMD
48.46990407
BMDL
46.51019855
BMDU
77.34071876
AIC
44.42271558
Test 4 P-value
0.208696414
D.O.F.
3
Model Parameters
# of Parameters
5
Variable
Estimate
Std Error
Lower Uonl
Upper Uonl
9
8.079313138
0.49034105
7.118262
9.040364
beta
-0.008334359
5.80E-04
-0.00947
-0.0072
beta2
Bounded
NA
NA
NA
rho
1.460822961
4.76E-01
0.527479
2.394166
alpha
0.103359443
7.75E-03
0.088165
0.118554
Goodness of Fit
Dose
Size
Estimated
Calo'd
Observed
Estimated
Calc'd
Observe
Scaled
Median
Median
Mean
3D
3D
d 3D
Residual
0
3
8.07931314
7.848667
7.848667
1.478876
1.2022
1.20219
-0.270132
100
3
7.24587722
8.409444
8.409444
1.365826
1.323
1.32299
1.4755594
300
3
5.57900537
4.878667
4.878667
1.128415
0.6349
0.63485
-1.074979
600
3
3.0786976
2.921333
2.921333
0.730944
1.2057
1.20567
-0.372892
900
3
0.57838984
0.603111
0.603111
0.215525
0.2231
0.22309
0.1986704
2944
2945
2946
2947
Likelihoods of Interest
Model
Log Likelihood"
#of
Parameters
AIC
A1
-18.39458048
6
48.78916
A2
-14.33273648
10
48.66547
A3
-15.94105261
7
45.88211
fitted
-18.21135779
4
44.42272
R
-38.05514411
2
80.11029
' Includes additive constant of -13.78408. This constant was
Tests of Interest
Test
2'Log(Likeliho
od Ratio)
Test df
p-value
1
47.44481524
8
<0.0001
2
8.123687994
4
0.087151
3
3.216632255
3
0.359415
4
4.54061036
3
0.208696
not included in the LL derivation prior to BMDS 3.0.
FigureApx F-6. Model Results of Frequentist Exponential Degree 3 Model of Gray
et al. (2021) Data
2948
2949
F.3 BMD Model Results of Furr et al. (2014) (Block 36 Rats)
Page 114 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
2950 Table Apx F-6. Ex Vivo Fetal
Rat Testicular Testosterone Data (Furr et al., 2014) (Block 36 Rats)
Dose
(mg/kg-day)
N
(# of litters)
Mean
Standard
Deviation
Notes
0
3
11.63
0.225167
Data from Table 2 in (Furr et al.. 2014)
100
2
5.43
0.820244
300
2
3.81
0.296985
600
3
2.77
1.143154
900
3
1.73
0.127279
2951
Page 115 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Table Apx F-7. B]V
D Model Results Ex
Vivo Fetal Tesl
icular Testosterone (Block 36 - All
)ose Groups) (Furr et al., 2014)
Models "
Restriction b
Variance
BMR = 5%
BMR = 10%
BMR = 40%
BMR = 1 SD
P Value
AIC
BMDS
Recommends
BMDS Recommendation
Notes
BMD
BMDL
BMD
BMDL
BMD
BMDL
BMD
BMDL
Exponential 3
Restricted
Constant
17.3487
11.35096
35.6356
23.31577
172.7742
113.0432
47.89558
28.74612
<0.0001
51.45915786
Questionable
Constant variance test failed
(Test 2 p-value <0.05)
Goodness of fit p-value <0.1
BMD 3x lower than lowest
non-zero dose
BMDL 3x lower than lowest
non-zero dose
Modeled control response
std. dev. >1.5 actual
response std. dev.
Exponential 5
Restricted
Constant
6.53742
4.690565
13.53295
9.716502
70.61377
51.01258
8.614115
5.656828
0.0087808
37.33552192
Questionable
Constant variance test failed
(Test 2 p-value <0.05)
Goodness of fit p-value <0.1
BMD 3x lower than lowest
non-zero dose
BMDL 3x lower than lowest
non-zero dose
BMD lOx lower than lowest
non-zero dose
BMDL lOx lower than
lowest non-zero dose
Modeled control response
std. dev. >1.5 actual
response std. dev.
Hill
Restricted
Constant
4.406264
3.14372
9.375719
6.362706
60.82678
42.65383
4.613014
2.955083
0.1345529
31.87673639
Questionable
Constant variance test failed
(Test 2 p-value <0.05)
BMD 3x lower than lowest
non-zero dose
BMDL 3x lower than lowest
non-zero dose
BMD lOx lower than lowest
non-zero dose
BMDL lOx lower than
lowest non-zero dose
Modeled control response
std. dev. >1.5 actual
response std. dev.
Polynomial
Degree 3
Restricted
Constant
48.39746
40.18862
96.79489
80.3777
387.1796
321.5089
223.096
154.1383
<0.0001
61.53466853
Questionable
Constant variance test failed
(Test 2 p-value <0.05)
Goodness of fit p-value <0.1
Residual for Dose Group
Near BMD > 2
Residual at control > 2
Modeled control response
std. dev. >1.5 actual
response std. dev.
Page 116 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Models "
Restriction 6
Variance
BMR = 5%
BMR = 10%
BMR = 40%
BMR = 1 SD
P Value
AIC
BMDS
Recommends
BMDS Recommendation
Notes
BMD
BMDL
BMD
BMDL
BMD
BMDL
BMD
BMDL
Polynomial
Degree 2
Restricted
Constant
48.44536
40.18586
96.89071
80.37172
387.5628
321.4869
223.4798
154.1353
<0.0001
61.53482916
Questionable
Constant variance test failed
(Test 2 p-value < 0.05)
Goodness of fit p-value <0.1
Residual for Dose Group
Near BMD > 2
Residual at control > 2
Modeled control response
std. dev. >1.5 actual
response std. dev.
Power
Restricted
Constant
49.1679
40.13384
98.3358
80.26768
393.3432
321.0707
222.1772
153.6257
<0.0001
61.59215522
Questionable
Constant variance test failed
(Test 2 p-value < 0.05)
Goodness of fit p-value <0.1
Residual for Dose Group
Near BMD > 2
Residual at control > 2
Modeled control response
std. dev. >1.5 actual
response std. dev.
Linear
Unrestricted
Constant
48.37868
40.18983
96.75735
80.37966
387.0294
321.5167
222.9468
154.1385
<0.0001
61.53465443
Questionable
Constant variance test failed
(Test 2 p-value < 0.05)
Goodness of fit p-value <0.1
Residual for Dose Group
Near BMD > 2
Residual at control > 2
Modeled control response
std. dev. >1.5 actual
response std. dev.
Exponential 3
Restricted
Non-
Constant
28.63226
22.79934
58.81294
46.83166
285.1463
227.0567
269.7963
100.0567
0.0001714
45.85655362
Questionable
Non-constant variance test
failed (Test 3 p-value < 0.05)
Goodness of fit p-value <0.1
BMD 3x lower than lowest
non-zero dose
BMDL 3x lower than lowest
non-zero dose
Modeled control response
std. dev. >1.5 actual
response std. dev.
Exponential 5
Restricted
Non-
Constant
6.584692
4.690942
13.63099
9.717238
71.13498
51.02266
8.597824
5.658575
0.0095036
37.33137285
Questionable
Non-constant variance test
failed (Test 3 p-value < 0.05)
Goodness of fit p-value <0.1
BMD 3x lower than lowest
non-zero dose
BMDL 3x lower than lowest
non-zero dose
BMD lOx lower than lowest
non-zero dose
BMDL lOx lower than
Page 117 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Models "
Restriction 6
Variance
BMR = 5%
BMR = 10%
BMR = 40%
BMR = 1 SD
P Value
AIC
BMDS
Recommends
BMDS Recommendation
Notes
BMD
BMDL
BMD
BMDL
BMD
BMDL
BMD
BMDL
lowest non-zero dose
Modeled control response
std. dev. >1.5 actual
response std. dev.
Hill
Restricted
Non-
Constant
3.889547
2.960594
8.292334
6.225717
54.90352
43.90247
1.436185
0.693566
0.3308113
30.0887978
Questionable
Non-constant variance test
failed (Test 3 p-value < 0.05)
BMD 3x lower than lowest
non-zero dose
BMDL 3x lower than lowest
non-zero dose
BMD lOx lower than lowest
non-zero dose
BMDL lOx lower than
lowest non-zero dose
Polynomial
Degree 3
Restricted
Non-
Constant
63.27256
56.91067
126.5451
113.8213
506.1804
455.2852
957.3279
467.3266
<0.0001
47.54806445
Questionable
Non-constant variance test
failed (Test 3 p-value < 0.05)
Goodness of fit p-value <0.1
Residual at control > 2
Modeled control response
std. dev. >1.5 actual
response std. dev.
Polynomial
Degree 2
Restricted
Non-
Constant
63.26384
56.91213
126.5277
113.8243
506.1108
455.2988
956.4118
467.3261
<0.0001
47.54806964
Questionable
Non-constant variance test
failed (Test 3 p-value < 0.05)
Goodness of fit p-value <0.1
Residual at control > 2
Modeled control response
std. dev. >1.5 actual
response std. dev.
Power
Restricted
Non-
Constant
71.50585
56.69262
139.027
113.4702
525.5507
453.3178
1023.701
454.3581
<0.0001
49.78518364
Questionable
Non-constant variance test
failed (Test 3 p-value < 0.05)
Goodness of fit p-value <0.1
Residual at control > 2
Modeled control response
std. dev. >1.5 actual
response std. dev.
Linear
Unrestricted
Non-
Constant
63.27256
60.30696
126.5451
120.6139
506.1804
455.2853
957.328
467.3295
<0.0001
47.54806445
Questionable
Non-constant variance test
failed (Test 3 p-value < 0.05)
Goodness of fit p-value <0.1
Residual at control > 2
Modeled control response
std. dev. >1.5 actual
response std. dev.
Abbreviations: AIC = Akaike information criterion; BMD = Benchmark dose; BMDL = Benchmark dose lower limit; BMDS = Benchmark dose software; BMR = Benchmark response; SD = Standard deviation.
" Selected Model (bolded and shaded gray).
b Restrictions defined in the BMDS 3.3 User Guide.
Page 118 of 122
-------
2953
PUBLIC RELEASE DRAFT
DECEMBER 2024
Page 119 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
2954
2955
2956
F.4 BMP Model Results of Furr et al. (2014) (Block 37 Rats)
Dose
(mg/kg-day)
N
(# of litters)
Mean
Standard
Deviation
Notes
0
4
10.94
3.24
Data from Table 2 in (Furr et al.. 2014)
11
3
12.17
0.536936
33
4
10
3.3
100
4
9.63
2.16
2957
Page 120 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
2958 Table Apx F-9. BMP Mode
Results Ex Vivo Fetal Testicular Testosterone (Block 37 rats - All Dose Groups) (Furr
et al., 2014)
Models "
Restriction 6
Variance
BMR = 5%
BMR = 10%
BMR = 40%
BMR = 1 SD
P Value
AIC
BMDS
Recommends
BMDS Recommendation
Notes
BMD
BMDL
BMD
BMDL
BMD
BMDL
BMD
BMDL
Exponential 3
Restricted
Constant
30.10844
11.26244
61.84514
23.13395
299.8475
101.5943
139.0458
48.51259
0.5816642
74.52267227
Questionable
Constant variance test
failed (Test 2 p-value <
0.05)
Exponential 5
Restricted
Constant
30.43784
1.996375
32.10822
5.958151
0.485658
75.92507984
Questionable
Constant variance test
failed (Test 2 p-value <
0.05)
BMD/BMDL ratio > 3
BMDL 3x lower than
lowest non-zero dose
Hill
Restricted
Constant
29.25549
0.12955
31.44254
0.157793
0.485658
75.92507985
Questionable
Constant variance test
failed (Test 2 p-value <
0.05)
BMD/BMDL ratio > 20
BMD/BMDL ratio > 3
BMDL 3x lower than
lowest non-zero dose
BMDL lOx lower than
lowest non-zero dose
Polynomial
Degree 3
Restricted
Constant
33.205
13.77656
65.87241
27.54961
250.9207
107.5624
136.5096
54.65605
0.2939821
76.54024317
Questionable
Constant variance test
failed (Test 2 p-value <
0.05)
Polynomial
Degree 2
Restricted
Constant
55.52211
13.56606
83.38946
27.12893
179.4662
107.9355
128.3908
53.79601
0.2641501
76.685824
Questionable
Constant variance test
failed (Test 2 p-value <
0.05)
BMD/BMDL ratio > 3
Power
Restricted
Constant
32.33752
13.78305
64.67505
27.56611
258.7002
101.476
136.8744
54.68127
0.5778416
74.53585932
Questionable
Constant variance test
failed (Test 2 p-value <
0.05)
Linear
Unrestricted
Constant
32.32747
13.78304
64.65496
27.56616
258.6198
110.2644
136.6855
54.68155
0.5778512
74.53582612
Questionable
Constant variance test
failed (Test 2 p-value <
0.05)
Exponential 3
Restricted
Non-
Constant
-
-
-
-
-
-
-
-
-
-
Unusable
BMD computation failed
Exponential 5
Restricted
Non-
Constant
30.43784
1.471662
32.10822
5.238397
0.0055921
77.92507984
Questionable
Goodness of fit p-value <
0.1
BMD/BMDL ratio > 20
BMD/BMDL ratio > 3
BMDL 3x lower than
lowest non-zero dose
Hill
Restricted
Non-
Constant
28.22122
8.46407
30.45435
9.456441
NA
79.91452333
Questionable
BMD/BMDL ratio > 3
d.f.=0, saturated model
(Goodness of fit test
cannot be calculated)
Page 121 of 122
-------
PUBLIC RELEASE DRAFT
DECEMBER 2024
Models "
Restriction 6
Variance
BMR = 5%
BMR = 10%
BMR = 40%
BMR = 1 SD
P Value
AIC
BMDS
Recommends
BMDS Recommendation
Notes
BMD
BMDL
BMD
BMDL
BMD
BMDL
BMD
BMDL
Polynomial
Degree 3
Restricted
Non-
Constant
33.20414
14.27532
66.40826
28.55064
265.6328
109.4742
154.2012
56.69838
0.0185327
76.22425247
Questionable
Goodness of fit p-value <
0.1
Polynomial
Degree 2
Restricted
Non-
Constant
33.30109
14.27474
66.60218
28.54949
266.4087
111.7167
154.7987
56.70352
0.018533
76.22422372
Questionable
Goodness of fit p-value <
0.1
Power
Restricted
Non-
Constant
33.3204
14.27492
66.64081
28.54985
266.5632
101.5496
154.8914
56.70377
0.018533
76.22422594
Questionable
Goodness of fit p-value <
0.1
Linear
Unrestricted
Non-
Constant
33.30108
14.27472
66.60217
28.54948
266.4087
114.1977
154.7986
56.7035
0.018533
76.22422372
Questionable
Goodness of fit p-value <
0.1
Abbreviations: AIC = Akaike information criterion; BMD = Benchmark dose; BMDL = Benchmark dose lower limit; BMDS = Benchmark dose software; BMR = Benchmark response; NA = Not Applicable;
SD = Standard deviation.
" Selected Model (bolded and shaded gray). No models were selected.
b Restrictions defined in the BMDS 3.3 User Guide.
2959
Page 122 of 122
------- |