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PUBLIC RELEASE DRAFT
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EPA Document# EPA-740-D-24-023
December 2024

United States	Office of Chemical Safety and

Environmental Protection Agency	Pollution Prevention

Draft Environmental Hazard Assessment for Dibutyl Phthalate

(DBP)

Technical Support Document for the Draft Risk Evaluation

CASRN 84-74-2

December 2024


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TABLE OF CONTENTS

ACKNOWLEGEMENTS	5

SUMMARY	6

1	INTRODUCTION	7

1.1 Approach and Methodology	7

2	ENVIRONMENTAL HAZARD	8

2.1	Aquatic Species	8

2.1.1	Acute Toxicity of DBP in Aquatic Vertebrates	8

2.1.2	Chronic Toxicity of DBP in Aquatic Vertebrates	10

2.1.3	Acute Toxicity of DBP in Aquatic Invertebrates	13

2.1.4	Chronic Toxicity of DBP in Aquatic Invertebrates	14

2.1.5	Acute Toxicity of DBP in Benthic Invertebrates	16

2.1.6	Chronic Toxicity of DBP in Benthic Invertebrates	16

2.1.7	Toxicity of DBP in Aquatic Plants and Algae	19

2.2	Terrestrial Species	20

2.2.1	Toxicity of DBP in Terrestrial Vertebrates	20

2.2.2	Toxicity of DBP in Soil Invertebrates	21

2.2.3	Toxicity of DBP in Terrestrial Plants	22

2.3	Hazard Thresholds	24

2.3.1	Acute Aquati c C oncentrati on of C oncern	25

2.3.2	Chronic Aquatic Vertebrate Concentration of Concern	26

2.3.3	Chronic Aquatic Invertebrate Concentration of Concern	26

2.3.4	Acute Benthic Concentration of Concern	26

2.3.5	Chronic Benthic Concentration of Concern	27

2.3.6	Aquatic Plant and Algae Concentration of Concern	27

2.3.7	Terrestrial Vertebrate Hazard Value	27

2.3.8	Soil Invertebrate Hazard Value	28

2.3.9	Terrestrial Plant Hazard Value	28

2.4	Weight of Scientific Evidence and Conclusions	29

2.4.1	Quality of the Database; Consistency; Strength (Effect Magnitude) and Precision; and
Biological Gradient (Dose-Response)	30

2.4.2	Relevance (Biological; Physical/Chemical; Environmental)	34

3	CONCLUSIONS	36

REFERENCES	38

APPENDICES	46

Appendix A RUBRIC FOR WEIGHT OF THE SCIENTIFIC EVIDENCE	46

A. 1 Confidence Levels	46

A.2 Types of Uncertainties	46

Appendix B SPECIES SENSITIVITY DISTRIBUTION FOR ACUTE AQUATIC HAZARD 50

Appendix C ENVIRONMENTAL HAZARD STUDIES	57

Appendix D SUPPLEMENTAL SUBMITTED DATA TO BE CONSIDERED FOR FINAL

RISK EVALUATION	64


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LIST OF TABLES	

Table ES-1. Environmental Hazard Thresholds for DBP	6

Table 2-1. Acute Toxicity of DBP in Aquatic Vertebrates	9

Table 2-2. Chronic Toxicity of DBP in Aquatic Vertebrates	12

Table 2-3. Acute Toxicity of DBP in Aquatic Invertebrates	14

Table 2-4. Chronic Toxicity of DBP in Aquatic Invertebrates	15

Table 2-5. Acute Toxicity of DBP in Aquatic Benthic Invertebrates	16

Table 2-6. Chronic Toxicity of DBP in Benthic Invertebrates	17

Table 2-7. Toxicity of DBP in Aquatic Plants and Algae	20

Table 2-8. Toxicity of DBP to Terrestrial Vertebrates	21

Table 2-9. Toxicity of DBP in Soil Invertebrates	21

Table 2-10. Toxicity of DBP in Terrestrial Plants	23

Table 2-11. Acute Aquatic COC and Multiomics PODs	26

Table 2-12. DBP Evidence Table Summarizing the Overall Confidence Derived from Hazard

Thresholds	29

LIST OF APPENDIX TABLES	

TableApx A-l. Considerations that Inform Evaluations of the Strength of the Evidence within an

Evidence Stream (i.e., Apical Endpoints, Mechanistic, or Field Studies)	48

Table Apx B-l. Species Sensitivity Distribution (SSD) Model Input for Acute Exposure Toxicity in

Aquatic Vertebrates and Invertebrates - Empirical Data	50

Table Apx B-2. SSD Model Predictions0 for Acute Exposure Toxicity to Aquatic Vertebrates (Fish).. 51
Table Apx B-3. Species Sensitivity Distribution (SSD) Model Input for Acute Exposure Toxicity in

Aquatic Vertebrates and Invertebrates - Web-ICE Data	51

Table_Apx C-l. Acute Aquatic Vertebrate Toxicity of DBP	57

Table_Apx C-2. Chronic Aquatic Vertebrate Toxicity of DBP	58

Table_Apx C-3. Acute Aquatic Invertebrate Toxicity of DBP	59

Table Apx C-4. Chronic Aquatic Invertebrate Toxicity of DBP	59

TableApx C-5. Chronic Benthic Invertebrate Toxicity of DBP	59

Table Apx C-6. Aquatic Plants and Algae Toxicity of DBP	60

Table_Apx C-l. Terrestrial Vertebrate Toxicity of DBP	61

Table_Apx C-8. Acute Soil Invertebrate Toxicity of DBP	62

Table Apx C-9. Chronic Soil Invertebrate Toxicity of DBP	62

Table_Apx C-10. Terrestrial Plant Toxicity of DBP	63

LIST OF APPENDIX FIGURES	

FigureApx B-l. AICc for the Six Distribution Options in the SSD Toolbox for Acute DBP Toxicity

to Aquatic Vertebrates and Invertebrates (Etterson, 2020)	 54

Figure Apx B-2. Q-Q Plots of Acute DBP Toxicity to Aquatic Vertebrates and Invertebrates with the

A) Gumbel, B) Weibull, C) Burr, and D) Logistic Distributions (Etterson, 2020)	 55

Figure Apx B-3. Species Sensitivity Distribution (SSD) for Acute DBP Toxicity to Aquatic

Vertebrates and Invertebrates (Etterson, 2020)	 56

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ABBREVIATIONS AND ACRONYMS

AF

Assessment factor

ChV

Chronic value

CI

Confidence interval

coc

Concentration(s) of concern

EC50

Effect concentration at which 50% of test organisms exhibit an effect

EPA

Environmental Protection Agency

HC05

Hazard concentration that is protective of 95% of the species in the SSD

HV

Hazard value

LC50

Lethal concentration at which 50 percent of test organisms die

LD50

Lethal dose at which 50 percent of test organisms die

LOAEC

Lowest-observed-adverse-effect-concentration

LOAEL

Lowest-observed-adverse-effect-level

LOEC

Lowest-observed-effect-concentration

LOEL

Lowest-observed-effect-level

MATC

Maximum acceptable toxicant concentration

NOAEC

No-observed-adverse-effect-concentration

NOAEL

No-observed-adverse-effect-level

NOEC

No-observed-effect-concentration

NOEL

No-observed-effect-level

OCSPP

Office of Chemical Safety and Pollution Prevention

OPPT

Office of Pollution Prevention and Toxics

POD

Point of departure

QSAR

Quantitative structure-activity relationship

SSD

Species sensitivity distribution

TOC

Total organic carbon

TRV

Toxicity reference value

TSCA

Toxic Substances Control Act

U.S.

United States

Web-ICE

Web-based Interspecies Correlation Estimation

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ACKNOWLEGEMENTS	

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

Acknowledgements

The Assessment Team gratefully acknowledges the participation, review, and input from EPA OPPT
and OSCPP senior managers and science advisors. The Agency is also grateful for assistance from the
following EPA contractors for the preparation of this draft technical support document: General
Dynamics Information Technology, Inc. (Contract No. HHSN316201200013W); ICF, Inc. (Contract No.
68HERC23D0007); SpecPro Professional Services, LLC (Contract No. 68HERC20D0021); and SRC,
Inc. (Contract No. 68HERH19D0022 and 68HERC23D0007).

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 Research and Development
(ORD).

Docket

Supporting information can be found in the public docket, Docket ID EPA-HQ-QPPT-2018-0503.
Disclaimer

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

Authors: Collin Beachum (Management Lead), Mark Myer (Assessment Lead, Environmental Hazard
Assessment Lead), Jennifer Brennan (Assessment Lead, Environmental Hazard Discipline Lead),
Christopher Green (Environmental Hazard Discipline Lead), Emily Griffin (Environmental Hazard
Assessor)

Contributors: Azah Abdallah Mohamed, Rony Arauz Melendez, Sarah Au, Maggie Clark, Jone
Corrales, Daniel DePasquale, Lauren Gates, Ryan Klein, Sydney Nguyen, Brianne Raccor, Maxwell
Sail, Andrew Sayer, Joe Valdez, Leora Vegosen.

Technical Support: Mark Gibson, Hillary Hollinger, S. Xiah Kragie

This draft technical support document was reviewed and cleared for release by OPPT and OCSPP
leadership.

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SUMMARY	

This technical document is in support of the Draft Risk Evaluation for Dibutyl Phthalate (DBP) (U.S.
EPA, 2024). DBP is a common name for the chemical substance 1,2-benzenedicarboxylic acid, 1,2-
dibutyl ester (CASRN 84-74-2). See the draft risk evaluation for a complete list of all the technical
support documents for DBP.

EPA considered all reasonably available information identified through the systematic review process
under the Toxic Substances Control Act (TSCA) to characterize environmental hazard endpoints for
DBP. After evaluating the reasonably available information, environmental hazard thresholds were
derived for aquatic vertebrates, aquatic invertebrates, benthic invertebrates, aquatic plants and algae,
terrestrial vertebrates, soil invertebrates, and terrestrial plants. These hazard thresholds are summarized
in Table ES-1. EPA's rationale for selecting these hazard thresholds, as well as the level of confidence
in each is based on the weight of scientific evidence, is described in Section 2.42.3 and 0.

Table ES-1. Environmental Hazard Thresholds for DBP

Receptor Group

Exposure Duration

Hazard Threshold (COC or HV)

Citation

Aquatic Vertebrates

(Including

Amphibians)

Acute (96 hours)

347.6 jig/L

SSD (see Section 2.3.1)

Chronic (112 days)

1.56 jig/L

(EAG Laboratories. 2018)

Aquatic
Invertebrates

Acute (96 hours)

347.6 jig/L

SSD (see Section 2.3.1)

Chronic (14 days)

12.23 jig/L

(Taeatz et al.. 1983)

Benthic
Invertebrates

Acute (96 hours)

347.6 jig/L

SSD (see Section 2.3.1)

Chronic (10 days)

114.3 mg DBP/kg dry sediment

(Call et al.. 2001a)

Aquatic Plants and
Algae

96 hours

31.6 jig/L

(Adachi et al.. 2006)

Terrestrial
Vertebrates

17 weeks

80 mg/kg-bw/day

(NTP. 1995)

Soil Invertebrates

21 days

14 mg DBP/kg dry soil

(Jensen et al.. 2001)

Terrestrial Plants

40 days

10 mg DBP/kg dry soil

(Gao et al.. 2019)

COC = concentration of concern; HV = hazard value

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1 INTRODUCTION	

Dibutyl phthalate is an organic substance primarily used as a plasticizer in a wide variety of consumer,
commercial and industrial products. DBP may be released during industrial activities and through
consumer use, with most releases occurring to air and water. EPA reviewed studies of the toxicity of
DBP to aquatic and terrestrial organisms and its potential environmental hazards.

1.1 Approach and Methodology	

EPA utilized studies with overall quality determinations of high and medium to characterize the
environmental hazards of DBP to surrogate species representing various receptor groups, including
aquatic vertebrates, aquatic invertebrates, benthic invertebrates, aquatic plants and algae, terrestrial
mammals, soil invertebrates, and terrestrial plants. Mechanistic (transcriptomic and metabolomic) and
behavioral points of departure (PODs) from an acute exposure of DBP to fathead minnows were
compared to the acute aquatic vertebrate hazard threshold. Hazard studies with mammalian wildlife
exposed to DBP were not available; therefore, EPA used ecologically relevant endpoints from the
laboratory rat and mouse—model organisms that are commonly used to evaluate human health
hazards—to establish a hazard threshold for terrestrial mammals. Although two studies with overall
quality determinations of high and medium containing avian hazard data were available for exposures to
DBP, no apical hazards were observed in those studies. Because no apical hazards were observed in any
avian studies, EPA was not able to establish a definitive hazard threshold for avian species.

TSCA requires that EPA use data and/or information in a manner consistent with the best available
science and that the Agency base decisions on the weight of scientific evidence. To meet the TSCA
science standards, EPA applies a systematic review process to identify data and information across
taxonomic groups for both aquatic and terrestrial organisms with a focus on apical endpoints (e.g., those
affecting survival, growth, or reproduction). The data collection, data evaluation, and data integration
stages of the systematic review process are used to develop the hazard assessment to support the
integrative risk characterization. EPA completed the review of environmental hazard data/information
sources during risk evaluation using the data quality review evaluation metrics and the rating criteria
described in the 2021 Draft Systematic Review Protocol Supporting TSCA Risk Evaluations for
Chemical Substances (U.S. EPA. 2021) and the Draft Systematic Review Protocol for Dibutyl Phthalate
(DBP) (U.S. EPA. 2024c). Studies identified and evaluated by the Agency were assigned an overall
quality determination of high, medium, low, or uninformative. Study quality was evaluated based on a
rubric that included consideration of the following seven overarching domains: test substance, test
design, exposure characterization, test organism, outcome assessment, confounding/variable control, and
data presentation/analysis. Several metrics within each of these domains were evaluated for each study,
and an overall study quality determination was assigned based on the overall evaluation. Because data
on toxicity of DBP are numerous, EPA systematically evaluated all data for this hazard characterization,
but relied only on high-quality and medium-quality studies for purposes of risk characterization.

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2 ENVIRONMENTAL HAZARD

2.1 Aquatic Species	

EPA reviewed 68 studies for DBP toxicity to aquatic organisms. Some studies may have included
multiple endpoints, species, and test durations. Of these 68 studies, those that received an overall quality
determination of low or uninformative were not considered for quantitative risk evaluation. For the 55
studies that received an overall quality determination of high and medium, those that demonstrated no
acute or chronic adverse effects at the highest concentration tested (unbounded no-observed-effect-
concentration [NOECs]), or where hazard values exceeded the limit of solubility for DBP in water as
determined by EPA at 11.2 mg/L (U.S. EPA. 2024a). are listed in Appendix C. Those studies were
excluded from consideration for development of hazard thresholds (see Section 2.3). Of the 68 studies,
55 were considered for the development of hazard thresholds as outlined below.

2.1.1 Acute Toxicity of DBP in Aquatic Vertebrates	

EPA reviewed 17 studies that received overall quality determination of high or medium for acute
toxicity in aquatic vertebrates (Table 2-1). Two studies received overall quality determinations of low or
unacceptable and were not considered. Of the 17 high and medium quality studies, 13 contained
acceptable acute endpoints that identified definitive hazard values below the DBP limit of water
solubility. Additionally, predicted hazard data for 53 species were generated using EPA's Web-Based
Interspecies Correlation Estimation (Web-ICE) tool, including predictions for 31 aquatic vertebrates, 5
aquatic invertebrates, 16 benthic invertebrates, and 1 amphibian species. For the fathead minnow
{Pimephales promelas), bluegill {Lepomis macrochirus), and rainbow trout (Oncorhynchus mykiss), the
96-hour mortality LC50s ranged from 0.48 to 2.02 mg/L DBP (Smithers Viscient. 2018; Adams et al..
1995; EnviroSvstem. 1991; Defoe et al.. 1990; McCarthy and Whitmore. 1985; EG&G Bionomics.
1983a. b; Buccafusco et al.. 1981). Additional endpoints were established in two fish species, including
a 144-hour mortality LC50 of 0.92 mg/L and 96-hour mortality NOEC/LOEC of 0.53/8.3 mg/Lin the
fathead minnow (Smithers Viscient. 2018; EG&G Bionomics. 1984a) and a 72-hour mortality LC50 of
0.63 mg/L in the zebrafish {Danio rerio) (Chen et al.. 2014). Hazard values for development and growth
were also identified in the African clawed frog (Xenopus laevis).

For these endpoints, the 96-hour EC50s ranged from 0.9 to 8.40 mg/L. DBP was found to have
significant effects on developmental malformations in tadpoles at 0.5 mg/L (0.1 mg/L NOEC) with a 96-
hour EC50 of 0.9 mg/L (Lee et al.. 2005) at 6.3 mg/L (lowest concentration tested) with a 96-hour EC50
of 7.5 mg/L (Xu and Gve. 2018). and in tadpole embryos at 8.3 mg/L (5.8 mg/L NOEC) with a 96-hour
EC5 of 8.4 mg/L (Gardner et al.. 2016). The bolded values in Table 2-1 describe data which were used
as inputs for generating Web-ICE predictions and within a species-sensitivity distribution analysis
(SSD) (Appendix B).

TSCA section 4(h)(1)(B) requires EPA to encourage and facilitate the use of scientifically valid test
methods and strategies that reduce or replace the use of vertebrate animals while providing information
of equivalent or better scientific quality and relevance that will support regulatory decisions. One avenue
of research for reducing the time needed for toxicity testing in vivo is the use of transcriptomic and
metabolomic points of departure, which allow for studies with much shorter durations that still provide
the necessary robust experimental data to characterize hazard and provide important evidence for
mechanisms of action and affected cellular and metabolic pathways. A multiomics study was conducted
by EPA in which fathead minnows {Pimephales promelas) were exposed for 24 hours to several
phthalates, including DBP (Bencic et al.. 2024). PODs were derived for transcriptomic change (tPOD),
metabolomic change (mPOD), and behavioral change (bPOD). Additionally, a 24-hour mortality

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NOEC/LOEC of 0.8/2.1 mg/L was identified. At 2.1 mg/L DBP, 100 percent mortality was observed.
The tPOD identifies the DBP concentration at which gene expression is significantly affected. RNA was
isolated from exposed minnows at each treatment level and analyzed for significant deviation from the
control fish, and the tPOD was defined as the median benchmark dose limit (BMDL) for the lowest
affected gene ontology. For DBP, the tPOD was 0.12 mg/L. The mPOD identifies the DBP
concentration at which the metabolome is significantly affected. The mPOD was defined as the 10th
percentile benchmark dose (BMD) for change in metabolomics vs. the control. For DBP, the mPOD was
0.11 mg/L. The bPOD identifies the DBP concentration at which startle behavior is significantly
affected. The bPOD was defined as the SD50, or the concentration which causes a 50% reduction in
startle response in the fish larvae. For DBP, the bPOD was 0.24 mg/L.

Table 2-1. Acute Toxicity of DBP in Aquatic Vertebrates

Test Organism
(Species)

Hazard Values

Endpoint

Effect

Citation
(Study Quality)



0.1/0.5 mg/L

96-hour
NOEC/LOEC

Development/
Growth

(Lee et al.. 2005)
(High)



0.9 mg/L

96-hour EC50





African clawed frog

(Xenopus laevis)









7.5 mg/L

96-hour EC50

Growth

(Xu and Gve. 2018)



6.3 mg/L

96-hour LOEC



(High)



8.40 mg/L

96-hour EC50

Growth

(Gardner et al.. 2016) (Medium)



5.8/8.3 mg/L

96-hour
NOEC/LOEC







1.54 mg/L

96-hour LC50

Mortality

(Adams et al.. 1995) (High)



0.8/2.1 mg/L

24-hour NOEC/
LOEC

Mortality

(Bencic et al.. 2024)

Fathead minnow

0.92 mg/L

144-hour
LC50

Mortality

(EG&G Bionomics. 1984a)
(High)

(Pimephales
promelas)

2.02 mg/L

96-hour LC50

Mortality

(McCarthy and Whitmore. 1985)
(Medium)



0.85 mg/L

96-hour LC50

Mortality

(Defoe et al.. 1990) (Hish)



1.1 mg/L









1.0 mg/L

96-hour LC50

Mortality

(Smithers Viscient. 2018)
(Medium)



0.53/1.4 mg/L

96-hour
NOEC/LOEC



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Test Organism
(Species)

Hazard Values

Endpoint

Effect

Citation
(Study Quality)



0.8/2.1 mg/L

24-hour
NOEC/LOEC

Mortality

(Bencic et al.. 2024)

0.12 mg/L

24-hour tPOD

Transcriptomic
change

0.11 mg/L

24-hour mPOD

Metabolomic
change

0.24 mg/L

24-hour bPOD

Behavior

Bluegill (Lepomis
mctcrochirus)

1.2 mg/L

96-hour LC50

Mortality

(Buccafusco et al.. 1981)
(Medium)

0.85 mg/L

96-hour LC50

Mortality

(EG&G Bionomics. 1983b)
(High)

0.48 mg/L

96-hour LC50

Mortality

(Adams et al.. 1995) (High)

Rainbow trout

(iOncorhynchus
mykiss)

1.60 mg/L

96-hour LC50

Mortality

(Adams et al.. 1995) (High)

1.60 mg/L

96-hour LC50

Mortality

(EG&G Bionomics. 1983a)
(High)

1.4 mg/L

96-hour LC50

Mortality

(EnviroSvstem. 1991) (High)

Zebrafish (Danio
rerio)

0.63 mg/L

72-hour LC50

Mortality

(Chen et al.. 2014) (Medium)

Bolded values indicate data used to derive acute aquatic COC using SSD

2.1.2 Chronic Toxicity of DBP in Aquatic Vertebrates

EPA reviewed 16 studies with overall quality determinations of high or medium for chronic toxicity in
aquatic vertebrates (Table 2-2). One study received an overall quality determination of unacceptable and
was not considered. Of the 16 high and medium quality studies, 11 contained acceptable chronic
endpoints that identified definitive hazard values below the DBP limit of water solubility for five fish
species and two amphibians.

In zebrafish, there was a significant effect on offspring mortality resulting from females exposed to 0.1
and 0.5 mg/L DBP for 15 days. In the same study, zebrafish embryos exposed to 0.025 and 0.1 mg/L
DBP experienced developmental malformations. Further, exposure to DBP incited liver peroxisome
proliferation and up-regulation of aromatases in zebrafish embryos and adult females (Ortiz-Zarragoitia
et al.. 2006). In rainbow trout, the 99-day growth NOEC/LOEC was 0.10/0.19 mg/L (0.14 mg/L
maximum acceptable toxicant concentration [MATC]), representing significant effects on fish length

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and weight (Rhodes et al.. 1995; EnviroSvstem. 1991). Additionally, a 13-day NOEC/LOEC of 0.52/1.0
mg/L (1.3 mg/L LC50) and a 99-day NOEC/LOEC of 0.19/0.40 mg/L (0.28 mg/L MATC) for mortality
was identified in the rainbow trout (EnviroSvstem. 1991).

In a Bagrid catfish (Pseudobagrusfitlvidraco) feeding study, which used DBP concentrations of 100,
500, and 1000 mg DBP/kg diet, there was an observed significant reduction in body weight in fish that
were fed 1000 mg/kg over 8-weeks resulting in an 8-week NOEC/LOEC of 500/1000 mg DBP/kg diet.
Additionally, significant effects of acetylcholinesterase activity were observed in the brain at
concentrations of 100 mg DBP/kg diet, in the liver, muscle, and kidney at 500 mg DBP/kg diet, in the
heart at 1000 mg DBP/kg diet, and in gill tissue at 1000 mg DBP/kg diet. The authors stated that feeding
was conducted at a rate of 3% body weight per day based on group biomass at Week 0 and Week 4.
Based on this rate, the three given doses in dietary concentration (100/500/1000 mg DBP/kg diet) can be
converted to a dose in terms of fish body weight as 3/15/30 mg DBP/kg-bw/day. No significant effects
were observed in fish mortality during the 8-week period (Jee et al.. 2009). In the fathead minnow, a 20-
day NOEC/LOEC of 0.53/0.97 mg/L and 0.97/1.74 mg/L were identified for hatching rate and larval
survival, respectively (McCarthy and Whitmore. 1985).

In a multi-generational Japanese medaka (Oryzias latipes) study, an LC50 of 0.82 mg/L was identified
in embryos exposed (in an aqueous solution) to 0, 0.67, 0.74, 0.80, 1.0, and 1.3 mg/L DBP. In the F0
generation exposed to DBP concentrations of 0, 12, 65, and 776 mg/kg-bw/day via diet, egg production
per female fish was significantly reduced at all test concentrations, however there were no significant
effects on survival, growth, or sexual development. In the F1 and F2 generations, there were no effects
on survival and growth, but there was an observed increase in hepatic vitellogenin levels in the F2 65
mg/kg-bw/day DBP group (12/65 mg/kg-bw/day NOEL/LOEL). Further, in the F1 and F2 generations,
there was no egg production at the highest DBP dose (776 mg/kg-bw/day). (Patyna. 1999). In another
multigenerational Japanese medaka study in which parental fish were aqueously exposed to DBP at
concentrations of 0.015, 0.038, 0.066, 0.103, and 0.305 mg/L for 218 days, significant effects were
observed in growth of both male and female F1 and F2 generations. In the male and female F1
generation (subadults), weight was significantly less when compared to controls at 70-days, resulting in
NOEC/LOECs of 0.103/0.305 and 0.0387/0.066 mg/L in males and females, respectively. Additionally,
in the female F2 generation (subadults), length was significantly less compared to controls at day 70,
resulting in a NOEC/LOEC of 0.0156/0.0387 mg/L. Similarly, in the male and female F2 generation
(adults), weight was significantly less compared to controls at 98-days, resulting in NOEC/LOECs of
0.103/0.305 and 0.0156/0.0387 mg/L in males and females, respectively. In this study, unbounded
effects (unbounded LOEC) were also observed for growth at the lowest concentration tested.
Specifically, male F1 adult weight at 112-days, male F2 adult weight and length at 70-days, and male F2
adult length at 98-days were significantly inhibited at 0.015 mg/L DBP (EAG Laboratories. 2018).

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345 Table 2-2. Chronic Toxicity of DBP in Aquatic Vertebrates

Test Organism
(Species)

Hazard Values

Endpoint

Effect

Citation
(Study Quality)

African clawed frog

(Xenopus laevis)

0.1 mg/L

30-week LOEC

Growth

(Lee and

Veeramachaneni. 2005)
(High)

2/10 mg/L

22-day NOEC/LOEC

Growth

(Shen et al.. 2011)
(High)

0.00476/0.0134
mg/L

21-day NOEC/LOEC

Growth

(Battelle. 2018) (Hiah)

Japanese wrinkled
frog (Glandirana
rugosa)

0.28/2.8 mg/L

21-day NOEC/LOEC

Growth

(Ohtani et al.. 2000)
Medium)

Zebrafish (Danio
rerio)

0.1 mg/L

5-week LOEC

Mortality

(Ortiz-Zarraaoitia et al..
2006) (Medium)

Rainbow trout

(iOncorhynchus
mvkiss)

0.1/0.19 mg/L

99-day NOEC/LOEC

Growth

(Rhodes et al.. 1995)
(High)

1.3 mg/L

13-day LC50

Mortality

(EnviroSvstem. 1991)

0.52/1.0 mg/L

13-day NOEC/LOEC

0.28 mg/L

99-day MATC

0.19/0.40 mg/L

99-day NOEC/LOEC

(High)

0.14 mg/L

99-day MATC

Growth

0.10/0.19 mg/L

99-day NOEC/LOEC

Bagrid catfish

(Pseudobagrus
fitlvidraco)

15/30 mg/kg-
bw/day (feeding
study)

8-week NOEC/
LOEC

Growth

(Jee et al.. 2009) (Hiah)

Fathead minnow

{Pimephales
promelas)

0.53/0.97 mg/L

20-day NOEC/
LOEC

Mortality - hatch rate

(McCarthy and
Whitmore. 1985)
(Medium)

0.97/1.74 mg/L

Mortality - larval
survival

Japanese medaka

0Oryzias latipes)

<12/12 mg/kg-
bw/day (Feeding
study)

180-day LOEC

Reproduction - F0
egg production per
female

(Patvna. 1999) (Hiah)

65/776 mg/kg-
bw/day (Feeding
study)

180-day NOEC/
LOEC

Reproduction - F1
egg production per
female

65/776 mg/kg-
bw/day (Feeding
study)

180-day NOEC/
LOEC

Growth - weight,
female F1

0.82 mg/L

17-day LC50

Mortality

0.103/0.305
mg/L

70-day NOEC/
LOEC

Growth - weight,
male F1 subadults

(EAG Laboratories.
2018)(High)

0.0156/0.0387

Growth - length,

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Test Organism
(Species)

Hazard Values

Endpoint

Effect

Citation
(Study Quality)

Japanese medaka

0Oryzias latipes)

mg/L



male F1 subadults



0.0387/0.066
mg/L

Growth - weight,
female F1 subadults

0.0156/0.0387
mg/L

Growth - length,
female F1 subadults

<0.0156 mg/L/
0.0156 mg/L

112-day LOEC

Growth - weight,
male F1 adults

0.0156/0.0387
mg/L

112-day
NOEC/LOEC

Growth - length,
male F1 adults

0.066/0.103
mg/L

Growth - weight and
length, female
F1 adults

<0.0156 mg/L/
0.0156 mg/L

70-day LOEC

Growth - weight and
length, male F2
subadults

0.0156/0.0387
mg/L

70-day NOEC
/LOEC

Growth - length and
weight, female F2
subadults

<0.0156 mg/L/
0.0156 mg/L

98-day LOEC

Growth - length,
male F2 adults

0.103/0.305
mg/L

98-day NOEC
/LOEC

Growth - weight,
male F2 adults

0.0156/0.0387
mg/L

Growth - length and
weight, female F2
adults

Bolded values indicate hazard value used in determining concentration of concern (COC).

2.1.3 Acute Toxicity of DBP in Aquatic Invertebrates	

EPA reviewed 11 studies that received overall quality determinations of high or medium for acute
toxicity in aquatic invertebrates (Table 2-3). Three studies received overall quality determinations of low
or unacceptable and were not considered. All 11 of the high and medium quality studies contained
acceptable chronic endpoints that identified definitive hazard values below the DBP limit of water
solubility for 9 aquatic invertebrate species. Additionally, predicted hazard data for 53 species were
generated using EPA's Web-ICE tool, including predictions for 31 aquatic vertebrates, 5 aquatic
invertebrates, 16 benthic invertebrates, and 1 amphibian species.

In the opposum shrimp, the mortality 96-hour LC50s ranged from 0.50 to 0.75 mg/L. Mortality was
assessed at 48- and 72-hours, resulting in a 0.87 and 0.77 mg/L LC50, respectively (EG&G Bionomics.
1984b). In the water flea {Daphnia magna), the 48-hour mortality LC50s ranged from 2.55 to 5.2 mg/L
(Wei et al.. 2018; McCarthy and Whitmore. 1985). In the water flea, additional endpoints of
immobilization were also identified, resulting in 24-hour LC 50 of 8.0 mg/L and 48-hour EC50 of 2.99
mg/L. In Taiwan abalone (Haliotis diversicolor), at DBP concentrations of 0, 0.5, 0.2, 2.0, 10, and 15
mg/L, one study identified abnormal growth of embryos exposed to 10 mg/L DBP, resulting in a 96-
hour NOEC/LOEC of 2.0/10 mg/L (Yang et al.. 2009). Another Taiwan abalone embryo study that

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utilized DBP concentrations of 0, 0.0017, 0.0207, 0.196, 1.984, 20.09, 9.22, and 39.47 mg/L
demonstrated significant effects on embryonic development resulting in a 9-hour EC50 of 8.37 mg/L.
Additionally, metamorphosis was found to be disrupted at 10 mg/L DBP resulting in a 96-hour
NOEC/LOEC of 2.0/10 mg/L. Lastly, there was a significant increase in population growth and a
negative effect on sexual reproduction in the rotifer {Brcichionus calyciflorus) with a resulting 0.5/1.0
mg/L 48-hour no-observed-adverse-effect-concentration (NOAEC)/lowest-observed-adverse-effect-
concentration (LOAEC) and 1.0/2.0 mg/L 96-hour NOAEC/LOAEC, respectively (Cruciani et al..
2015). The bolded values in Table 2-3 describe data which were used as inputs for generating Web-ICE
predictions and within an SSD (Appendix B).

Table 2-3. Acute Toxicity of DBP in Aquatic Invertebrates

Test Organism
(Species

Hazard Values

Endpoint

Effect

Citation
(Study Quality)

Opossum shrimp

(Americamysis
bcthia)

0.75 mg/L

96-hour LC50

Mortality

(EG&G Bionomics.
1984b)

0.77 mg/L

72-hour LC50

Mortality

0.87 mg/L

48-hour LC50

Mortality

0.50 mg/L

96-hour LC50

Mortality

(Adams et al.. 1995)
(High)

Water flea

(Daphnict magna)

2.99 mg/L

48-hour EC50

Immobilization

Taiwan abalone

(Haliotis
diversicolor)

2/10 mg/L

96-hour NOEC/
LOEC

Development/Growth

(Yane et al.. 2009)
(Medium)

Taiwan abalone

(Haliotis
diversicolor)

8.37 mg/L

9-hour EC50

Development/Growth

(Liu et al.. 2009)
(Medium)

0.0207/0.196 mg/L

96-hour NOEC/
LOEC

Development/Growth
- metamorphosis

Water flea

(Daphnia magna)

5.2 mg/L

48-hour LC50

Mortality

(McCarthy and
Whitmore. 1985)
(Medium)

2.55 mg/L

48-hour LC50

Mortality

(Wei et al.. 2018)
(High)

4.31 mg/L

Mortality

2.83 mg/L

Mortality

8.0 mg/L

24-hour LC50

Immobilization

(Huana et al.. 2016)
(High)

Rotifer (Brachionus
calyciflorus)

1.0/2.0 mg/L

96-hour NOAEC/
LOAEC

Reproduction

(Cruciani et al.. 2015)

0.5/1.0 mg/L

48-hour NOAEC/
LOAEC

Population

(Medium)

Bolded values indicate data used to derive acute aquatic COC using SSD.

2.1.4 Chronic Toxicity of DBP in Aquatic Invertebrates	

EPA reviewed 13 studies which received an overall quality determination of high or medium for chronic
toxicity in aquatic invertebrates (Table 2-4). One study received an overall quality determination of low
and was not considered. Of the 13 high and medium quality studies, 8 contained chronic endpoints that

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identified definitive hazard values below the DBP limit of water solubility for 10 aquatic invertebrate
species.

A 21-day mortality NOEC/LOEC of 0.96/2.5 mg/L and a 21-day mortality LC50 of 1.92 mg/L were
identified in the water flea (Daphnia magna). Reproduction, population, development, and growth
endpoints were also identified. For reproduction, there was an observed decrease of fecundity in three
studies resulting in a range of NOEC/LOECs of 0.07/0.23 (number of days between eggs laid) to
1.05/1.92 mg/L (1.64 mg/L, 21-day EC50). In the water flea, there was also an observed reduction in
population growth rate (total neonates) with a NOEC/LOEC of 0.42/0.48 mg/L and a reduction in
development/growth (length) with a NOAEC/LOAEC of 0.278/2.78 mg/L (Wei et al.. 2018; Defoe et
al.. 1990; Springborn Bionomics. 1984b). In the rotifer {Brachionus calyciflorus) at aqueous
concentrations of 0, 0.000005, 0.00005, 0.0005, 0.005, 0.05, 0.5, and 5.0 mg/L, significant effects on
mortality and reproductive rates were observed after 6 days, resulting in a NOEC/LOEC of 0.05/0.5
mg/L for both endpoints (Zhao et al.. 2009).

Table 2-4. Chronic Toxicity of DBP in Aquatic Inverl

ebrates

Test Organism
(Species)

Hazard Values

Endpoint

Effect

Citation
(Study Quality)

Water flea

(Daphnia magna)

0.07/0.23 mg/L

21-day NOAEC/
LOAEC

Reproduction - #
days between egg
laid

(Wei et al.. 2018) (Hieh)

<0.07/0.07 mg/L

Reproduction -
Fecundity

0.42/0.48 mg/L

21-day NOAEC/
LOAEC

Population

0.278/2.78 mg/L

14-day NOAEC/
LOAEC

Development/
Growth

(Sevoum and Pradhan.
2019) (Medium)

0.96/2.5 mg/L

21-day NOAEC/
LOAEC

Mortality

(Sprineborn Bionomics.
1984b) (Medium)

0.96/2.5 mg/L

21-day NOAEC/
LOAEC

Reproduction

1.92 mg/L

21-day LC50

Mortality

(Defoe et al.. 1990) (Hiah)

1.64 mg/L

21-day EC50

Reproduction

1.05/1.91 mg/L

21-day NOAEC/
LOAEC



Scud (Gammarus
piilex)

0.1/0.5 mg/L

20-day NOAEC/
LOAEC

Behavior

(Thuren and Woin. 1991)
(Medium)

Amphipod
crustacean
(Corophium

acherusicum)

0.044/0.34 mg/L

14-day NOAEC/
LOAEC

Population -
Abundance

(Taaatz et al.. 1983)
(Medium)

Rotifer (Brachionus
calyciflorus)

0.05/0.5 mg/L

6-day

NOAEC/LOAEC

Mortality

(Zhao et al.. 2009)
(Medium)

Rotifer (Brachionus
calyciflorus)

0.05/0.5 mg/L

6-day

NOAEC/LOAEC

Reproduction

Bolded values indicate hazard value used in determining COC.

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2.1.5 Acute Toxicity of DBP in Benthic Invertebrates

EPA reviewed four studies that received an overall quality determination of high or medium for acute
toxicity in aquatic benthic invertebrates (Table 2-5). All four studies contained acute endpoints that
identified definitive hazard values below the DBP limit of water solubility for three aquatic invertebrate
species. In the harpacticoid copepod (Nitocra spinipes) and the midge (Paratcmytarsus
parthenogeneticus), the 96-hour mortality LC50s ranged from 1.7 to 6.29 mg/L (Adams et al.. 1995;
Linden et al.. 1979). In the midges (Paratcmytarsus parthenogeneticus and Chironomus plumosus), the
48-hour mortality LC50s ranged from 4.0 to 5.8 mg/L (EG&G Bionomics. 1984c; Streufort 1978).

Table 2-5. Acute Toxicity of DBP in Aquatic Benthic Invertebrates

Test Organism

Hazard Values

Endpoint

Effect

Citation
(Study Quality)

Harpacticoid copepod

(Nitocra spinipes)

1.7 mg/L

96-hour LC50

Mortality

(Linden et al.. 1979) (Medium)

Midge (Paratanytarsus
parthenogeneticus)

5.8 mg/L

48-hour LC50

Mortality

(EG&G Bionomics. 1984c)
(High)

Midge (Paratanytarsus
parthenogeneticus)

6.29 mg/L

96-hour LC50

Mortality

(Adams et al.. 1995) (Hish)

Midge (Chironomus
plumosus)

4.0 mg/L

48-hour LC50

Mortality

(Streufort. 1978) (Medium)

Bolded values indicate data used to derive acute aquatic COC using SSD.

2.1.6 Chronic Toxicity of DBP in Benthic Invertebrates

EPA reviewed five studies which received an overall quality determination of high or medium for
chronic toxicity in benthic invertebrates (Table 2-6). All five studies contained acceptable chronic
endpoints that identified definitive hazard values below the DBP limit of water solubility for six benthic
invertebrate species. A study (Call et al.. 2001a) examining the effects of DBP in sediment pore water
and sediment for high, medium, and low TOC (total organic carbon) in Hyalella azteca resulted in 10-
day development/growth (decrease in weight compared to controls) NOEC/LOECs of 4.76/10.7
mg/L and 3,410/26,200 mg/kg, 4.20/12.9 mg/L and 748/3,340 mg/kg, and 0.70/4.59 mg/L and 41.6/360
mg/kg, respectively. In that study, there were no significant effects on H. azteca mortality. In the midge
(Chironomus tentans), effects on mortality and growth were observed in the high, medium, and low
TOC sediment groups. For high TOC, a 10-day NOEC/LOEC of 0.448/5.85 mg/L in sediment pore
water and 508/3550 mg/kg in sediment was observed for an increase in weight. For medium TOC, a 10-
day NOEC/LOEC of 3.85/16 mg/L in sediment pore water and 423/3090 mg/kg in sediment was
observed for an increase in weight relative to controls. For mortality, the 10-day NOEC/LOEC for
sediment pore water and sediment in high, medium, and low TOC was 0.448/5.85 mg/L and 508/3550
mg/kg, 3.85/16 mg/L and 423/3090 mg/kg, and 0.672/4.59 mg/L and 50.1/315 mg/kg, respectively (Call
et al.. 2001a). Another benthic invertebrate study examined the effects of DBP aqueous exposures and
observed significant effects in the midge and H. azteca. Specifically, in the midge, a 10-day growth and
development (weight) NOEC/LOEC of 1.78/4.52 mg/L (2.81 mg/LEC50) and a 10-day mortality LC50
of 2.64 mg/L was observed. In H. azteca, a 10-day mortality LC50 of 0.63 mg/L was identified (Call et
al.. 2001b).

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Lake Superior Research Institute (1997) also examined the effects of aqueous and sediment (high,
medium, and low TOC) DBP exposures in the midge and the scud. Ten-day LC50s were calculated via
multiple methods including Trimmed Spearman-Karber, probit analysis, and/or linear interpolation. In
the midge, the high, medium, and low TOC pore water 10-day mortality LC50s ranged from 4.22 to 6.21
mg/L, 10.3 mg/L (one value), and 6.86 to 6.95 mg/L, respectively. The high, medium, and low TOC
sediment 10-day mortality LC50s ranged from 4,730 to 5,213 mg/kg, 2,261 to 4,730 mg/kg, and 706 to
827 mg/kg, respectively. Most LC50s were unable to be calculated for the scud due to low mortality,
however there was a calculated 10-day mortality LC50 of 52,363 mg/kg in the medium sediment TOC
group. That study also conducted water only tests in which 10-day mortality LC50s for the midge
ranged from 2.64 to 3.08 mg/L and 0.59 to 0.63 mg/L for the scud (Lake Superior Research Institute.
1997).

In the mollusk (several species), segmented worm (several species), Actiniaria (unidentified species),
and sea squirt (Molgula manhcittensis), the 14-day population (abundance and diversity) NOEC/LOECs
were 0.34/3.7 mg/L. In the amphipod crustacean (Corophium acherusicum), the abundance
NOEC/LOEC was slightly more sensitive at 0.044/0.34 mg/L (Tagatz et al.. 1983). Two additional
endpoints were available in two studies for the worm (Lumbricuius variegatus) and the scud (Gammarus
pulex). In the worm, a 2.48 mg/L (in water) 10-day LC50 was identified for mortality (Call et al..
2001b). In the scud (Gammarus pulex), there was a significant effect on distance moved and changes in
direction resulting in a 20-day NOAEC/ LOAEC of 0.1/0.5 mg/L (in water) (Thuren and Woin. 1991).

Table 2-6. Chronic r

"oxicity of DBP in Bent

thic Invertebrates

Test Organism
(Species) and TOC

Hazard Values

Endpoint

Effect

Citation
(Study Quality)

Hvalella azteca high
TOC

4.76/10.7 mg/L

10-day NOAEC/
LOAEC

Development/
Growth

(Call et al.. 2001a)
(High)

3,410/26,200 mg/kg dry
sediment

Hyalella azteca
Medium TOC

4.20/12.9 17 mg/L

10-day NOAEC/
LOAEC

Development/
Growth

(Call et al.. 2001a)
(High)

748/ 3340 mg/kg dry
sediment

52,363 mg/kg bulk
sediment (Probit)

10-day LC50

Mortality

(Lake Superior
Research Institute.
1997)(Hiah)

Hvalella azteca low
TOC

0.70/4.59 mg/L

10-day NOAEC/
LOAEC

Development/
Growth

(Call et al.. 2001a)
(High)

41.6/360 mg/kg dry
sediment

Midge (Chironomns
ten tans) high TOC

6.12 mg/L (Probit)

10-day LC50

Mortality

(Lake Superior
Research Institute.
1997)(High)

6.21 mg/L (Linear
Interpolation)

5,213 mg/kg (Linear
Interpolation)

4.22 mg/L (Trimmed
Spearman-Karber)

4,730 mg/kg (Trimmed
Spearman-Karber)

0.448/5.85 mg/L

10-day NOAEC/
LOAEC

Development/
Growth

(Call et al.. 2001a)

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Test Organism
(Species) and TOC

Hazard Values

Endpoint

Effect

Citation
(Study Quality)

Midge (Chironomus
ten tans) high TOC

508/3,550 mg/kg dry
sediment

10-day NOAEC/
LOAEC



(High)

0.448/5.85 mg/L

10-day NOAEC/
LOAEC

Mortality

4.22 mg/L

10-day LC50

508/3,550 mg/kg dry
sediment

10-day NOAEC/
LOAEC

4,730 mg/kg diy
sediment

10-day LC50

Midge (Chironomus
tentcms) medium TOC

2,261 mg/kg dry
sediment (Linear
Interpolation)

10-day LC50

Mortality

(Lake Superior

10.3 mg/L (Trimmed
Speannan-Karber)

Research Institute.
1997)(High)

4,730 mg/kg (Trimmed
Speannan-Karber)

423/3,090 mg/kg diy
sediment

10-day NOAEC/
LOAEC

Development/
Growth

(Call et al.. 2001a)
(High)

3.85/16 mg/L

10-day NOAEC/
LOAEC

423/3,090 mg/kg diy
sediment

10-day NOAEC/
LOAEC

Mortality

1,664 mg/kg dw

10-day LC50

3.85/16 mg/L

10-day NOAEC/
LOAEC

10.3 mg/L

10-day LC50

Midge (Chironomus
tentcms) low TOC

6.95 mg/L (Trimmed
Speannan-Karber)

10-day LC50

Mortality

(Lake Superior
Research Institute.
1997)(Hiah)

827 mg/kg (Trimmed
Speannan-Karber)

6.88 mg/L (Probit)

820 mg/kg (Probit)

6.86 mg/L (Linear
Interpolation)



706 mg/kg dry sediment
(Linear Interpolation)

0.672/4.59 mg/L

10-day NOAEC/
LOAEC

(Call et al.. 2001a)
(High)

50.1/315 mg/kg diy
sediment

10-day NOAEC/
LOAEC

Midge (Chironomus
tentcms)

1.78/4.52 mg/L

10-day NOAEC/
LOAEC

Development/
Growth

(Call et al.. 2001b)
(High)

2.81 mg/L

10-day EC50

2.64 mg/L

10-day LC50

Mortality

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Test Organism
(Species) and TOC

Hazard Values

Endpoint

Effect

Citation
(Study Quality)

Hyalella azteca

0.63 mg/L

10-day LC50

Mortality

(Call etal.. 2001b)
(High)

Midge (Chironomus
tentcms) water only test

2.64 mg/L (Trimmed
Speannan-Karber)

10-day LC50

Mortality

(Lake Superior
Research Institute.
1997)(High)

3.08 mg/L (Linear
Interpolation)

Hyalella Azteca
water only test

0.63 mg/L (Trimmed
Speannan-Karber)

10-day LC50

Mortality

(Lake Superior

0.62 mg/L (Probit)

Research Institute.
1997)(High)

0.59 mg/L (Linear
Interpolation)

Mollusk (several
species)

0.34/3.7 mg/L

14-day NOAEC/
LOAEC

Population -
Abundance and
Diversity

(Taeatz et al.. 1983)
(Medium)

Segmented worm

(several species)

0.34/3.7 mg/L

14-day NOAEC/
LOAEC

Population -
Abundance and
Diversity

Amphipod crustacean
(Corophium

acheriisicum)

0.044/0.34 mg/L

14-day NOAEC/
LOAEC

Population -
Abundance

Actiniaria

(unidentified species)

0.34/3.7 mg/L

14-day NOAEC/
LOAEC

Population -
Diversity

Sea squirt (Molgula
manhattensis)

0.34/3.7 mg/L

14-day NOAEC/
LOAEC

Population -
Abundance and
Diversity

Worm (Lumbricuius
variegatus)

2.48 mg/L

10-day LC50

Mortality

(Call etal.. 2001b)
(High)

Scud (Gammarus
pulex)

0.1/0.5 mg/L

20-day NOAEC/
LOAEC

Behavior

(Thuren and Woin.
1991) (Medium)

TOC = total organic carbon

" Value slightly greater than DBP water solubility. Species included for mollusk are Diastema varium. Laevicardium
mortoni, Tellina sp.,Anomalocardia auberiana, Marginella apicina, Morula didyma, Anadara transversa, Mitrella lunata,
Crassostrea virginica, Eupleura sulcidentata, Mangelia quadrata, Thais haemastoma, Bursatella leachii pleii, Atrina
rigida. and Polinices duplicatus. Species included for the segmented worm include Haploscoloplos robustus, Tharvx
marioni, Loimia viridis, Scoloplos rubra, Mediomastus californiensis, Malacoceros vanderhorsti, Aricidea fragilis,

Arm an di a agilis,Axiothella mucosa, Nephtvs picta, Prionospio heterobranchia, Unidentified Sabcllidac.. Imphictene sp.,
Galathowenia sp., Glycera americana, Lumbrineris sp..Magelona rosea, Minuspio sp., Neanthes succinea, and Pectinaria
gouldii.

Bolded values indicate hazard value used in determining COC.

450	2.1.7 Toxicity of DBP in Aquatic Plants and Algae

451	EPA reviewed seven studies which received overall quality determinations of high or medium for

452	toxicity in aquatic plants and algae (Table 2-7). Three studies received overall quality determinations of

453	low or unacceptable and were not considered. Of the 7 high and medium quality studies, 3 contained

454	acceptable endpoints that identified definitive hazard values below the DBP limit of water solubility for

455	one species of green algae (Selenastrum capricormitum). A 10-day static toxicity test examined the

456	percent increase or decrease of chlorophyll a at DBP concentrations of 0.05, 0.08, 0.13, 0.39, 0.77, and

457	1.45 mg/L. Chlorophyll a was found to increase slightly at lower concentrations, then decreased at

458	higher concentrations with an observed 100 percent decrease in chlorophyll a at 1.45 mg/L DBP

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resulting in a 10-day EC50 of 0.75 mg/L. The study authors noted that there was considerable loss of
phthalate esters from the test solutions and thus the EC50 values were calculated based on
concentrations measured at the beginning of the study (Springborn Bionomics. 1984c). Two other
studies examined the effects of DBP on S. capricornutum abundance. Adams et al. (1995) identified a
96-hour EC50 of 0.40 in S. capricornutum with DBP concentrations ranging from 0.21 to 377 mg/L and
Adachi et al. (2006) identified a 96-hour NOEC/LOEC of 0.1/1.0 mg/L in S. capricornutum at
concentrations ranging from 0.1 to 10 mg/L.

Table 2-7. Toxicit

y of DBP in Aquatic Plants and Algae

Test Organism
(Species)

Hazard Values

Endpoint

Effect

Citation
(Study Quality)

Green algae

(Selenastrum
capricornutum)

0.75 mg/L

10-day EC50

Population
(Chlorophyll a
concentration)

(SDrinsborn
Bionomics. 1984c)
(High)

0.40 mg/L

96-hour EC50

Population
(Abundance)

(Adams et al.. 1995)
(High)

0.1/1 mg/L

96-hour NOEC/
LOEC

Population
(Abundance)

(Adachi et al.. 2006)
(Medium)

Bolded values indicate hazard value used in determining COC.

2.2 Terrestrial Species	

EPA reviewed 35 studies for DBP toxicity to terrestrial organisms. Some studies may have included
multiple endpoints, species, and test durations. Of these 35 studies, those that received an overall quality
determination of low or uninformative were not considered for quantitative risk evaluation. For the 30
studies that received an overall quality determination of high and medium, those that demonstrated no
acute or chronic adverse effects at the highest dose tested (unbounded NOAELs) are listed in Appendix
C and were excluded from consideration for development of hazard thresholds. In addition to the 30
high or medium quality terrestrial wildlife studies, EPA considered 13 terrestrial vertebrate studies for
toxicity to DBP in human health animal model rodent species that contained ecologically relevant
reproductive endpoints (TableApx C-7).

2.2.1 Toxicity of DBP in Terrestrial Vertebrates	

No reasonably available information was identified for exposures of DBP to mammalian wildlife. EPA
reviewed 13 studies for toxicity to DBP in human health animal model rodent studies that contained
ecologically relevant reproductive endpoints (Table Apx C-7). EPA's decision to focus on ecologically
relevant (population level) reproductive endpoints in the rat and mouse data set for DBP for
consideration of a hazard threshold in terrestrial mammals is due to the known sensitivity of these taxa
to DBP in eliciting phthalate syndrome (U.S. EPA. 2024b). Of the 13 rat and mouse studies containing
ecologically relevant reproductive endpoints, EPA selected the study with the most sensitive LOAEL for
deriving the hazard threshold for terrestrial mammals (Table 2-8). The most sensitive endpoint resulted
from a Sprague-Dawley rat (Rattus norvegicus) study in which a 17-week LOAEL for significant
reduction in number of live pups per litter was observed at 80 mg/kg-bw/day DBP intake in dams (NTP.
1995).

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Table 2-8. Toxicit

y of DBP to Terrestrial Vertebrates

Test Organism
(Species)

Hazard Values

Endpoint

Effect

Citation
(Study Quality)

Rat (Rattus
norvegicus)

80 mg/kg-bw/day

17-week LOAEL

Reproduction

(NTP. 1995) (Hieh)

2.2.2 Toxicity of DBP in Soil Invertebrates	

EPA reviewed 14 studies that received an overall quality determination of high or medium for acute
toxicity in soil invertebrates (Table 2-9). One study received an overall quality determination of low and
was not considered. Of the 14 high and medium quality studies, 12 contained acute endpoints that
identified definitive hazard values below the DBP limit of water solubility for five soil invertebrate
species.

In the European house dust mite {Dermatophagoides pteronyssinus), American house dust mite
(Dermatophagoides farina), and copra mite (Tyrophagusputrescentiae), the 24-hour mortality LC50s
with fabric contact to DBP were found to range from 0.017 to 0.03 mg/cm2 and 0.077 to 0.079 mg/cm2
(LD50s) via direct application of DBP (Wang et al.. 2011; Kim et al.. 2008. 2007; Kang et al.. 2006; Tak
et al.. 2006; Kim et al.. 2004). In the earthworm (Eiseniafetida), the 48-hour mortality LC50 via DBP
on filter paper ranged from 1.3 to 6.8 mg/cm2 (Du et al.. 2015; Neuhauser et al.. 1985). Because filter
paper contact is not considered a relevant exposure pathway for soil invertebrates due to the absorbed
amount of chemical via dermal contact being uncertain, EPA did not establish a hazard threshold from
the filter paper data set. In the nematode (Caenorhabditis e/egans), the 24-hour reproduction
NOEC/LOEC were 2.783/27.83 mg/L and 27.83/139.17 mg/L for hatching rate and brood size,
respectively. Specifically, nematodes exposed to DBP at concentrations of 0.0278, 2.78, 27.8, and 139
mg/L experienced an increase in embryonic lethality (reduced hatch rate) at 27.8 mg/L and a decrease in
mean number of eggs laid at 139 mg/L (Shin et al.. 2019).

In the springtail (Folsomiaflmetarid), the 21-day mortality LC10 and LC50 for juveniles was 11.3 and
19.4 mg/kg, respectively, and 33 and 305 mg/kg, respectively, for adults. Adult springtail reproduction
was also significantly affected with an observed 21-day EC10 and EC50 of 14 and 68 mg/kg (Jensen et
al.. 2001). A 14-day earthworm {Eisenia fetida) study identified a mortality LC50 of 2,364.8 mg/kg. In
this study, mechanistic endpoints were also observed; superoxide dismutase and catalase were found to
be significantly reduced at 100 mg/kg DBP on day 28; glutathione-S-transferase was increased after day
21 in the 10 to 50 mg/kg DBP group; glutathione was found to increase on days 7 to 28 in the 50 mg/kg
DBP group; and malondialdehyde was greater in all dosage groups and time frames compared to
controls (Du et al.. 2015).

Table 2-9. Toxicity of DBP in Soil Invertebrates

Test Organism
(Species)

Hazard Values

Endpoint

Effect

Citation
(Study Quality)

European house
dust mite

(Dermatophagoides
pteronyssinus)

0.07779 mg/cm2 (Direct
application)

24-hour LD50

Mortality

(Kans et al.. 2006)
(Medium)

0.02323 mg/cm2 (Fabric
contact)

24-hour LC50

(Wans et al.. 2011)
(Medium)

0.02851 mg/cm2 (Fabric
contact)

24-hour LC50

(Kim et al.. 2008)
(Medium)

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Test Organism
(Species)

Hazard Values

Endpoint

Effect

Citation
(Study Quality)



0.03159 mg/cm3 (Fabric
contact)

24-hour LC50



(Kim et al.. 2004)
(Medium)

0.01881 mg/cm2 (Fabric
contact)

24-hour LC50

(Kim et al.. 2007)
(Medium)

American house
dust mite

(Dermatophagoides
farina)

0.07954 mg/cm2 (Direct
application)

24-hour LD50

Mortality

(Kana et al.. 2006)
(Medium)

0.02189 mg/cm2 (Fabric
contact)

24-hour LC50

(Wans et al.. 2011)
(Medium)

0.0281 mg/cm2 (Fabric
contact)

24-hour LC50

(Kim et al.. 2008)
(Medium)

0.03392 mg/cm3 (Fabric
contact)

24-hour LC50

(Kim et al.. 2004)
(Medium)

0.01739 mg/cm2 (Fabric
contact)

24-hour LC50

(Kim et al.. 2007)
(Medium)

Copra mite

(Tyrophagns
putrescentiae)

0.02523 mg/cm2 (Fabric
contact)

24-hour LC50

Mortality

(Tak et al.. 2006)
(Medium)

Earthworm
(Eisenia fetida)

6.8 mg/cm2 (Filter
paper)

48-hour LC50

Mortality

(Du et al.. 2015) (Medium)

1.360 mg/cm2 (Filter
paper)

48-hour LC50

Mortality

(Neuhauser et al.. 1985)
(Medium)

Nematode

(Caenorhabditis
elegans)

2.783/27.83 mg/L in
solution

24-hour
NOEC/LOEC

Reproduction
(Hatch rate)

(Shin et al.. 2019) (Hieh)

27.83/139.17 mg/L in
solution

24-hour
NOEC/LOEC

Reproduction
(Brood size)

Springtail

(Folsomia
fimetaria) -
Juvenile

11.3 mg/kg dry soil

21-day LC10

Mortality

(Jensen et al.. 2001) (Hieh)

19.4 mg/kg dry soil

21-day LC50

Springtail

(Folsomia
fimetaria) - Adult

33 mg/kg dry soil

21-day LC10

305 mg/kg dry soil

21-day LC50

14 mg/kg dry soil

21-day EC10

Reproduction

68 mg/kg dry soil

21-day EC50

Earthworm

{Eisenia fetida)

2364.8 mg/kg dry soil

14-day LC50

Mortality

(Du et al.. 2015) (Medium)

Bolded values indicate hazard value used in determining a hazard value.

524	2.2.3 Toxicity of DBP in Terrestrial Plants

525	EPA reviewed 12 studies that received an overall quality determination of high or medium for hazard in

526	terrestrial plants (Table 2-10). Three studies received overall quality determinations of low or

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unacceptable and were not considered. Of the 12 high and medium quality studies, 6 contained
acceptable endpoints that identified definitive hazard values for 10 terrestrial plant species.

The main endpoint observed to be affected by exposure to DBP was growth. In the dutch clover
(Trifolium repens), turnip (Brassica rapa ssp. rapa), rippleseed plantain (Plantago major), and
velvetgrass {Holcus lanatus), there was an observed reduction in total biomass after DBP administration
via fumigation, resulting in a 62-day growth EClOs of 0.00033, 0.00077, 0.00239, 0.00879 mg/m3,
respectively. Similarly, in the common bean {Phaseolus vulgaris) that was harvested after 42 days, there
was an observed reduction in total biomass resulting in an EC 10 of 0.00232 mg/m3 (Dueck et al.. 2003).
Because fumigation is not considered a relevant exposure pathway for soil invertebrates due to the
exposure of the amount of chemical being uncertain, EPA did not establish a hazard threshold from the
fumigation data set. For plants exposed to DBP via soil, there was an observed reduction in biomass
resulting in a 72-hour EC50, 72-hour NOEC/LOEC, and 45-day NOEC/LOEC of 1,559 mg/kg, 5/20
mg/kg, and 10/100 mg/kg in mung bean (Vigna radiata), bread wheat (Triticum aestivum), and false bok
choy {Brassicaparachinensis), respectively (Zhao et al.. 2016; Ma et al.. 2015; Ma et al.. 2014).
Unbound LOAECs were also observed in which significant effects on growth were observed at the
lowest concentration tested. Specifically, in the common onion {Allium cepa), alfalfa {Medicago sativa),
radish (Raphamts sativus), cucumber (Cucumis sativus), and common oat (Avena sativa), growth was
significantly less compared to controls at 5 mg/kg soil (Ma et al.. 2015). In false bok choy there were
also observed mechanistic effects including a reduction in chlorophyll content, intercellular CO2
concentration, and catalase, as well as an increase in malondialdehyde—all of which resulted in a
NOEC/LOEC of 10/100 mg/kg (Zhao et al.. 2016).

In bread wheat exposed to DBP at concentrations of 0, 5, 10, 20, 30, and 50 mg/L, significant decreases
in the growth of roots and shoots up until germination were identified resulting in growth EClOs and
EC50s of 5.08 and 37.70 mg/L and 8.02 and 42.73 mg/L, respectively. Additionally, seed germination
was inhibited by DBP and was found to be 76.51 percent at 40 mg/L (Gao et al.. 2017). Similarly, a 40-
day LOEL of 10 mg/kg DBP (lowest concentration used in the study) for reduced weight in bread wheat
was observed (Gao et al.. 2019). In rapeseed {Brassica napus), a reduction in weight was also observed
at the lowest concentration used in the study resulting in an unbound LOEC of 50 mg/kg (Kong et al..
2018). Lastly, in the Chinese sprangletop {Leptochloa chinensis) and rice {Oryza sativa) exposed to DBP
concentrations of 1.2, 2.4, and 4.8 kg/ha via soil surface, there was an observed reduced seedling growth
(emergence) and weight in sprangletop resulting in a 14-day NOEC/LOEC of 1.2/2.4 kg/ha and reduced
root length, shoot height, and weight in rice resulting in a 14-day NOEC/LOEC 2.4/4.8 kg/ha (Chuahet
al.. 2014).

Table 2-10. Toxicity of DBP in Terrestrial Plants

Test Organism

Hazard Values

Endpoint

Effect

Citation

(Species)

(Study Quality)

Dutch clover

0.00033 mg/m3

62-day EC 10

Growth



{Trifolium repens)

(Fumigation)







Turnip {Brassica
rapa ssp. rapa)

0.00077 mg/m3
(Fumigation)

62-day EC 10

Growth



Rippleseed plantain

{Plantago major)

0.00239 mg/m3
(Fumigation)

62-day EC 10

Growth

(Dueck et al.. 2003)
(High)

Velvetgrass {Holcus
lanatus)

0.00879 mg/m3
(Fumigation)

62-day EC 10

Growth



Common bean

0.00232 mg/m3

42-day EC 10

Growth



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Test Organism
(Species)

Hazard Values

Endpoint

Effect

Citation
(Study Quality)

(Phctseolus
vulgaris)

(Fumigation)







Mung bean (Vigna
radiata)

1559 mg/kg dry soil

72-hour EC50

Growth

(Maetal.. 2014)
(Medium)

Common onion

(Allium cepct)

<5 mg/kg soil/5 mg/kg
soil

168-hour LOEC

Growth

(Maetal.. 2015)

Alfalfa (Medicctgo
sativa)

<5 mg/kg soil/5 mg/kg
soil

72-hour LOEC

Radish (Raphanus
sativus)

<5 mg/kg soil/5 mg/kg
soil

Cucumber (Cucumis
sativus)

<5 mg/kg soil/5 mg/kg
soil

(High)

Common oat (Avert a
sativa)

<5 mg/kg soil/5 mg/kg
soil

Bread wheat

(Triticum aestivum)

5/20 mg/kg soil

72-hour NOEC/
LOEC

5.08 mg/L

Until germination
EC10

Growth (Roots)

(Gao et al.. 2017)
(High)

37.70 mg/L

Until germination,
EC50

8.02 mg/L

Until germination
EC10

Growth
(Shoots)

42.73 mg/L

Until germination
EC50

30/40 mg/L

Until germination
NOEC/LOEC

Reproduction
(Germination)

<10 mg/kg dry soil/10
mg/kg dry soil

40-day LOEL

Growth

(Gao et al.. 2019)
(High)

False bok choy

(Brassica
parachinensis)

10/100 mg/kg dry soil

45-day NOAEC/
LOAEC

Growth

(Zhao et al.. 2016)
(Medium)

Chinese sprangletop

(Leptochloa
chinensis)

1.2/2.4 kg/ha

14-day NOEC/
LOEC

Growth

(Chuah et al.. 2014)
(Medium)

<500 mg/L

7-day LOEC

Reproduction
(Germination)

Rice (Oryza sativa)

2.4/4.8 kg/ha

14-day NOEC/
LOEC

Growth

Rapeseed (Brassica
napus)

<50 mg/kg dry soil/50
mg/kg dry soil

30-day LOEC

Growth

(Kona et al.. 2018)
(Medium)

Bolded values indicate hazard value used in determining a hazard value.

564	2.3 Hazard Thresholds	

565	EPA calculates hazard thresholds to identify potential concerns to aquatic and terrestrial species. After

566	weighing the scientific evidence, EPA selects the appropriate toxicity value from the integrated data to

567	use for hazard thresholds. See 0 for more details about how EPA weighed the scientific evidence and

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Section 2.4 for the weight of scientific evidence conclusions.

2.3.1 Acute Aquatic Concentration of Concern	

For aquatic species EPA uses probabilistic approaches (e.g., an SSD) when enough data are available,
and deterministic approaches (e.g., deriving a geometric mean of several comparable values) when more
limited data are available. An SSD is a type of probability distribution of toxicity values from multiple
species. It can be used to visualize which species are most sensitive to a toxic chemical exposure, and to
predict a concentration of a toxic chemical that is hazardous to a percentage of test species. This
hazardous concentration is represented as an HCP, where p is the percent of species below the threshold.
EPA used an HCos (a Hazardous Concentration threshold for 5% of species) to estimate a concentration
that would protect 95 percent of species. This HCos can then be used to derive a COC, which is the
estimated hazardous concentration of DBP in water for aquatic organisms. For the deterministic
approaches, COCs are calculated by dividing a hazard value by an assessment factor (AF) according to
EPA methods (U.S. EPA. 2016. 2014. 2012). However, for the probabilistic approach used for acute
aquatic hazard in this TSD, the lower bound of the 95 percent confidence interval (CI) of the HCos can
be used to account for uncertainty instead of dividing by an AF. EPA has more confidence in the
probabilistic approach when enough data are available because an HCos is representative of a larger
portion of species in the environment. Generally, EPA considers the probabilistic approach for aquatic
hazard (i.e., an SSD) appropriate when hazard values for at least eight species are represented in the data
set.

The aquatic acute COC for DBP was derived from an SSD that contained 96-hour LC50s for 9 species
identified in systematic review, bolstered by an additional 53 predicted LC50 values from the Web-ICE
v4.0 toxicity value estimation tool. Web-ICE (Web-based Interspecies Correlation Estimation) is a tool
developed by U.S. EPA's Office of Research and Development that estimates the acute toxicity of a
chemical to a species, genus, or family from the known toxicity of the chemical to a surrogate species. It
was used to obtain estimated acute toxicity values for DBP in species that were not represented in the
empirical data set. All empirical studies included in the SSD were rated high or medium quality. After
reviewing the possible statistical distributions for the SSD, the maximum likelihood method was chosen
with a Gumbel distribution. This choice was based on an examination of p-values for goodness of fit,
visual examination of Q-Q plots, and evaluation of the line of best fit near the low-end of the SSD. The
HC05 for this distribution is 414.9 |ig/L DBP. After taking the lower 5th percentile of this HC05 as an
alternative to the use of assessment factors, the acute aquatic COC for vertebrates and invertebrates is
347.6 |ig/L DBP.

See Appendix B for details of the SSD that was used to derive the acute aquatic COC for DBP.

The multiomics-based PODs derived by EPA in Bencic et al. (2024) suggest that Pimephalespromelas
(fathead minnow) larvae exhibited changes in gene expression, metabolite levels, and swimming
behavior at concentrations of DBP near the SSD-derived COC. EPA did not use the multiomics-based
PODs for hazard thresholds because it is uncertain if these sub-organismal and individual-level effects
(e.g., behavior) at short exposure durations scale up to ecologically relevant outcomes, such as survival
and reproduction, in wild fish populations. Notably, the PODs derived from the multiomics study are
similar to the SSD-derived acute aquatic COC (Table 2-11). This provides additional confidence in the
acute aquatic COC for DBP, as the multiomics approach resulted in a similar hazard value to that
derived from empirical and modeled data in the SSD.

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Table 2-11. Acute Aquatic COC and Multiomics

PODs

Acute Aquatic COC
(SSD-Derived)

Transcriptomic POD

Metabolomic POD

Behavioral POD

347.6 jig/L

120 jig/L

110 (ig/L

240 jig/L

2_.3.2 Chronic Aquatic Vertebrate Concentration of Concern	

EPA reviewed 17 studies on chronic toxicity in aquatic vertebrates. The most sensitive organism for
which a clear population-level fitness endpoint could be obtained was for Japanese medaka (O. latipes)
(EAG Laboratories. 2018). This study was rated high quality. In this multi-generational study, the
growth of the F1 and F2 generations of fish was significantly affected by exposure to DBP. In male F1
generation Japanese medaka, there was a significant inhibition of body weight at the lowest
concentration studied, with an unbounded LOEC value of 15.6 |ig/L DBP. The ChV (Chronic value, the
geometric mean of the NOEC and LOEC) for bodyweight inhibition in female F1 generation Japanese
medaka was 82.4 |ig/L DBP. In the F2 generation, the ChV for bodyweight inhibition in male fish was
177.2 |ig/L DBP, while the ChV for bodyweight inhibition in F2 female fish was 24.6 |ig/L DBP. The
most sensitive of these endpoints is the unbounded LOEC for inhibition of bodyweight in F1 males at
15.6 |ig/L DBP. At the lowest dose (15.6 |ig/L), bodyweight was inhibited by 13.4 percent relative to the
vehicle control, and there was a statistically significant trend toward greater bodyweight inhibition with
increasing dose, culminating at 34.0 percent inhibition at the highest dose (305 |ig/L). Based on the
presence of a statistically significant dose-response relationship and a population-level fitness endpoint,
the 112-day ChV for bodyweight inhibition in F1 male Japanese medaka was selected to derive the
chronic COC for aquatic vertebrates.

Because the most sensitive endpoint in this study was an unbounded LOEC, an AF of 10 was applied.
This is to account for the uncertainty in the actual threshold dose, which may have been lower than the
lowest dose studied. After applying an AF of 10, the chronic COC for aquatic vertebrates is 1.56 |ig/L
DBP.

^.3.3 Chronic Aquatic Invertebrate Concentration of Concern	

EPA reviewed 13 studies on chronic toxicity from DBP in aquatic invertebrates. The most sensitive
organism for which a clear population-level fitness endpoint could be obtained was for the marine
amphipod crustacean Monocorophium acherusicum (Tagatz et al.. 1983). with a 14-day ChV of 122.3
|ig/L DBP for reduction in population abundance. Populations were reduced by 91 percent at the LOEC,
which was 340 |ig/L DBP. Higher doses resulted in a complete loss of amphipods in the aquaria. This
study was rated medium quality. Based on the presence of a clear dose-response relationship and a
population-level fitness endpoint, the 14-day ChV for reduction in population abundance in the marine
amphipod crustacean was selected to derive the chronic COC for aquatic invertebrates. After applying
an AF of 10, the chronic COC for aquatic invertebrates is 12.23 |ig/L DBP.

2.3.4 Acute Benthic Concentration of Concern

Acute toxicity data from three empirical studies, representing LC50 estimates for three species of
benthic invertebrates, were included in the SSD for acute aquatic organisms. The acute aquatic COC
(see Section 2.3.1), because it was derived from an SSD that contained empirical LC50 data for benthic
invertebrates as well as WeblCE-derived predicted LC50s for additional benthic species including
worms (Lumbriculus variegatus), snails (Physella gyrina, Lymnaea stagricilis), and copepods (Tigriopus
jciponicus), is expected to encompass the level of concern for benthic invertebrates as well. The acute
benthic invertebrate COC is therefore 347.6 |ig/L DBP in water. There were no studies available to

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characterize the acute toxicity of DBP in sediment to benthic invertebrates; therefore, no COC was
derived for the sediment exposure pathway.

2.3.5	Chronic Benthic Concentration of Concern

EPA reviewed five studies on chronic toxicity from DBP in benthic invertebrates. Of these, the most
sensitive was the midge (Chironomus tentans) (Lake Superior Research Institute. 1997). with a 10-day
ChV for population loss of 1,143.3 mg DBP/kg dry sediment in medium-TOC sediments (4.80% TOC).
This study was rated high quality. This ChV was the middle of three for the midge; the experiment was
repeated with low, medium, and high TOC sediments and toxicity decreased with the increase in TOC,
as expected for a relatively hydrophobic compound like DBP based on equilibrium partitioning theory.
The chosen endpoint for deriving the COC, medium-TOC, was selected because it is the closest to the
assumed TOC level (4%) used in Point Source Calculator to estimate DBP exposure in benthic
organisms. Population was reduced by 76.7 percent at the LOEC, which was 3,090 mg DBP/kg dry
sediment. Higher doses resulted in a similar degree of population loss in the medium-TOC treatment;
however, all population losses were significantly different from controls. This endpoint was considered
acceptable to derive a COC because of population-level relevance and a demonstrated dose-response
relationship. After applying an AF of 10 to the ChV at 1,143.3 mg/kg, the chronic COC for benthic
invertebrates is 114.3 mg DBP/kg dry sediment.

2.3.6	Aquatic Plant and Algae Concentration of Concern	

EPA reviewed six studies on toxicity from DBP in aquatic plants and algae. Of these, the most sensitive
was green algae (Selenastrum capricornntam) (Adachi et al.. 2006) with a 96-hour ChV of 316 |ig/L
DBP for reduced population abundance. This study was rated medium quality. There was significant
reduction in the algal population at the LOEC, which was 1,000 |ig/L DBP, relative to an increase in the
algal population at the NOEC of 100 |ig/L DBP and in controls. The population reduction was increased
with a higher dose of DBP. Therefore, this endpoint was considered acceptable to derive a COC because
of population-level relevance and a demonstrated dose-response relationship. After applying an AF of
10, the COC for aquatic plants and algae is 31.6 |ig/L DBP.

2.3.7	Terrestrial Vertebrate Hazard Value	

EPA reviewed 15 studies on toxicity from DBP in terrestrial vertebrates. Of these, the most sensitive
among acceptable-quality studies was the Sprague-Dawley rat (Rattus norvegicus) (NTP. 1995). with a
17-week LOAEL for significant reduction in number of live pups per litter at 80 mg/kg-bw/day DBP
intake in dams. This study was assigned an overall quality determination of high.

The above referenced study also found a LOAEL for reduced bodyweight in F2 pups at the same dose
(80 mg/kg-bw/day). The lowest bounded NOAEL/LOAEL pair for which a ChV could be calculated
was significantly reduced bodyweight in F1 pups at a ChV of 115.4 mg/kg-bw/day, but this effect was
not as sensitive as reduced number of live pups per litter. Other effects of DBP exposure included
significantly decreased female body weight in dams, significantly reduced male sex ratio (percentage of
male pups), significantly decreased mating index and pregnancy index in the F1 generation, and
significantly reduced male pup weight gain.

Because the most sensitive endpoint in this study was an unbounded LOAEL, the actual threshold dose
may have been lower than the lowest dose studied. However, no information was available in the study
to adjust the value to account for this uncertainty. Other reproductive endpoints for which bounded
NOAEL/LOAEL pairs were observed in rats and mice (see Table Apx C-7) indicated ChV that were
higher than this unbounded LOAEL; therefore, it is not clear whether an adjustment for uncertainty is
necessary to adequately characterize the toxicity of DBP to terrestrial mammals. Based on reduction in

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live pups per litter, the results found in NTP (1995) indicated that toxicity in terrestrial vertebrates
occurs at 80 mg/kg-bw/day.

2.3.8	Soil Invertebrate Hazard Value	

EPA reviewed 10 studies on acute toxicity from DBP in terrestrial invertebrates; however, the majority
(8 of the 10 studies identified) focused on the use of DBP as a pesticide fumigant and the DBP dose that
was experienced by the invertebrates studied could not be determined from the available data. There
were two studies identified for which doses could be determined—for the fruit fly (Drosophila
melcmogaster) (Misra et al.. 2014) and the nematode (Caenorhabditis elegans) (Shin et al.. 2019). Both
studies were rated medium quality. For the fruit fly, the 72-hour LC50 value in feed (an agar-grape juice
solution) was 505,100 mg/L. This exposure was not considered ecologically relevant, as the dose would
need to be present in fruit at a concentration that is not possible based on the physicochemical properties
of DBP. Such a fruit would be nearly 33 percent DBP by mass. For the nematode, after 24-hours there
was no significant mortality observed at any dose examined up to the NOEL of 139.17 mg/L DBP in a
buffered water solution. However, this study did not observe any effect of DBP at any dose examined;
therefore, this exposure is not appropriate for use in calculating a hazard value.

The same study also examined hatch rate in the nematode (Caenorhabditis elegans) on agar plates and
had a 24-hour ChV of 8.8 mg/L DBP in agar. However, the magnitude of this effect was small even at
the highest DBP dose (an increase in embryonic mortality from approximately 3 to 8%), and it was
unclear whether a change of this magnitude has a population-level relevance. Therefore, this study was
not considered acceptable to derive a hazard threshold.

EPA reviewed two studies on chronic toxicity from DBP in soil invertebrates. Of these, the most
sensitive was the springtail (Folsomiafimetaria) (Jensen et al.. 2001) with a 21-day EC 10 of 14 mg
DBP/kg dry soil for reduced reproduction. This study was rated high quality. Reproduction was reduced
by approximately 60 percent at the lowest concentration tested, which was 100 mg DBP/kg dry soil,
with reproduction completely eliminated at higher doses. Therefore, this endpoint was considered
acceptable to derive a hazard value because of population-level relevance and a clear dose-response
relationship.

The hazard value for soil invertebrates is calculated as the geometric mean of ChV, EC20, and EC 10
values for mortality, reproduction, or growth endpoints from acceptable studies. Because the data set
contained one EC10 for reproduction of 14 mg DBP/kg dry soil, this value will be used as the hazard
value for soil invertebrates.

2.3.9	Terrestrial Plant Hazard Value

EPA reviewed 12 studies on toxicity from DBP in vascular plants. An unbounded LOEL for growth at
10 mg DBP/kg dry soil was obtained in a study rated high quality for a 40-day exposure in bread wheat
(Triticum aestivam) (Gao et al.. 2019). and at 50 mg DBP/kg dry soil for rapeseed (Brassica napus) in a
medium quality study (Kong et al.. 2018). The most sensitive endpoint was the LOEL for reduction in
leaf and root biomass in bread wheat seedlings observed in Gao et al. (2019). which was 10 mg/kg dry
soil. There was a clear dose-response observed, with biomass reduction increasing as the dose of DBP
increased. At the highest dose (40 mg/kg), root and leaf biomass were reduced by 29.93 and 32.10
percent, respectively. Because the most sensitive endpoint in this study was an unbounded LOAEL, the
actual threshold dose may have been lower than the lowest dose studied. However, no information was
available in the study to adjust the value to account for this uncertainty. The HV for terrestrial plants for
DBP derived from this study is 10 mg/kg dry soil.

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for a 42-day exposure in bok choy (Brassica rapa ssp. Chinensis) (Liao et al.. 2009) at 3.16 mg/L DBP
in hydroponic solution. This study was rated medium quality. Biomass was reduced by 27 percent at the
LOAEL (10 mg/L), with a clear dose-response at increasing doses up to 76 percent reduced biomass at
the highest dose (100 mg/L). However, this study was conducted in hydroponic solution rather than in
soil; therefore, it was not considered ecologically relevant for the purpose of deriving a hazard value.
Other ChVs included a 72-hour exposure in bread wheat (Triticum aestivum) (Ma et al.. 2015) at 100 mg
DBP/kg wet soil. This study was rated high quality. Unbounded LOELs for growth inhibition were also
obtained from (Ma et al.. 2015) for a 72-hour exposure in the common oat (Avena sativa), a 168-hour
exposure in the common onion (Allium cepa), a 72-hour exposure in alfalfa (Medicago sativa), and a 72-
hour exposure in the radish (Raphanus sativus). All of the aforementioned unbounded LOELs were at 5
mg/kg wet soil. However, because the study did not provide information on the water content of the soil,
this study was not considered acceptable to derive a hazard value. Furthermore, in this study a
comparator non-food crop plant (perennial ryegrass, Lolium perenne) had no observable effects on
growth even at the highest dose of 500 mg/kg wet soil.

Other studies investigated soil fumigation, application to fields (in kg/hectare), or direct application to
leaves (in |ig/cm2), and the dose to each plant could not be calculated from the information given.
Another study, rated medium quality, examined a 45-day exposure in false bok choy (Brassica
parachinensis) with a ChV of 31.62 mg DBP/kg dry soil (Zhao et al.. 2016); however, the lowest dose
(10 mg DBP/kg dry soil) resulted in statistically increased growth relative to controls.

2.4 Weight of Scientific Evidence and Conclusions	

After calculating the hazard thresholds that were carried forward to characterize risk, a table describing
the weight of the scientific evidence and uncertainties was completed to support EPA's decisions (Table
2-12). See 0 for more detail on how EPA weighed the scientific evidence.

Table 2-12. DBP Evidence Table Summarizing the Overall Confidence Derived from Hazard
Thresholds

Types of
Evidence

Quality of

the
Database

Consistency

Strength

and
Precision

Biological
Gradient/Dose-
Response

Relevance

Hazard
Confidence

Aquatic

Acute Aquatic
(SSD)

+++

+++

+++

+++

+++

Robust

Chronic Aquatic
Vertebrates

+++

++

+++

+++

+++

Robust

Chronic Aquatic
Invertebrates

+++

+++

+++

+++

+++

Robust

Chronic Benthic
Invertebrates

++

+++

+++

++

+++

Robust

Aquatic Plants &
Algae

++

+++

++

++

++

Moderate

Terrestrial

Terrestrial
Vertebrates

+++

++

++

+++

++

Moderate

Soil Invertebrates

++

++

+++

+++

+++

Robust

Terrestrial Plants

++

++

++

++

++

Moderate

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Types of
Evidence

Quality of

the
Database

Consistency

Strength

and
Precision

Biological
Gradient/Dose-
Response

Relevance

Hazard
Confidence

11 Relevance includes biological, physical/chemical, and environmental relevance.

+++ Robust confidence suggests thorough understanding of the scientific evidence and uncertainties. The supporting
weight of the scientific evidence outweighs the uncertainties to the point where it is unlikely that the uncertainties
could have a significant effect on the hazard estimate.

++ Moderate confidence suggests some understanding of the scientific evidence and uncertainties. The supporting
scientific evidence weighed against the uncertainties is reasonably adequate to characterize hazard estimates.
+ Slight confidence is assigned when the weight of the scientific evidence may not be adequate to characterize the
scenario, and when the assessor is making the best scientific assessment possible in the absence of complete
information. There are additional uncertainties that may need to be considered.	

2.4.1 Quality of the Database; Consistency; Strength (Effect Magnitude) and Precision;
	and Biological Gradient (Dose-Response)	

For the acute aquatic assessment, the database consisted of 28 studies with overall quality
determinations of high/medium with both aquatic invertebrates and vertebrates represented. Data from
nine of these studies were supplemented by using Web-ICE version 4.0 to obtain additional estimated
acute toxicity values and generate a subsequent SSD output; therefore, a robust confidence was assigned
to quality of the database. DBP had similar effects on the same species across multiple studies, well
within one order of magnitude. For example, 96-hour LC50 values in the fathead minnow {Pimephales
promelas) ranged from 0.85 mg/L to 2.02 mg/L across three independent studies, from 0.48 mg/L to 1.2
mg/L in the bluegill {Lepomis macrochirus) across three independent studies, and from 1.4 to 1.60 mg/L
in the rainbow trout (Oncorhynchus mykiss) across two independent studies. For the water flea {Daphnia
magna), 48-hour LC50s ranged from 2.55 mg/L to 5.2 mg/L across two independent studies. Because
LC50 values were comparable among independent studies conducted in well-characterized test
organisms, a robust confidence was assigned to consistency of the acute aquatic assessment. The effects
observed in the DBP empirical data set for acute aquatic assessment were mortality, with 48-, 72-, or 96-
hour LC50s represented empirically (depending on species) with additional predicted LC50 values
reported from Web-ICE. Because more than 50 species were represented in the acute data set with LC50
values, robust confidence was assigned to the strength and precision consideration. Dose-response is a
prerequisite of obtaining reliable LC50 values and was observed in the empirical studies that were used
in the SSD. Because dose-response was observed in the empirical studies, a robust confidence was
assigned to the dose-response consideration.

For the chronic aquatic vertebrate assessment, the database consisted of 16 studies with overall quality
determinations of high/medium. Of these studies, 11 contained acceptable chronic endpoints that
identified definitive hazard values below the DPB limit of water solubility for 5 fish species and 2
amphibians, resulting in robust confidence for quality of the database. DBP had chronic effects on
growth which spanned several orders of magnitude among aquatic vertebrate taxa, with effects on
growth in the African clawed frog (Xenopus laevis) ranging from NOEC/LOEC pairs of 0.00476/0.0134
mg/L to 2/10 mg/L in 21- and 22-day independent studies, respectively. Among fish, effects on growth
ranged from an unbounded LOEC at 0.0156 mg/L in Japanese medaka (Oryzias latipes) to 0.19/0.40
mg/L in rainbow trout (Oncorhynchus mykiss) in 112-day and 99-day studies, respectively. Among the
same species, in a three-generation reproductive study that received a high quality study evaluation,
(EAG Laboratories. 2018). effects on growth in Japanese medaka (Oryzias latipes) ranged from an
unbounded LOEC at 0.0156 mg/L in F1 male fish to a NOEC/LOEC pair at 0.103/0.305 mg/L in F2
male fish. Because chronic effects were seen at concentrations that spanned several orders of magnitude
among aquatic vertebrates, a moderate confidence was assigned to the consistency consideration. In the
study chosen to derive the COC, EAG Laboratories (2018). body weight was inhibited by 13.4 percent

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relative to the vehicle control, and there was a statistically significant trend toward greater bodyweight
inhibition with increasing dose, culminating at 34.0 percent inhibition at the highest dose (305 |ig/L).
Similarly strong dose-response effects were observed in other studies in the database. Because there was
a strong biologically relevant effect and dose-response effects were observed in the study chosen to
derive the COC and among other studies in the database, a robust confidence was assigned to the
strength and precision consideration and the dose-response consideration for the chronic aquatic
invertebrate assessment.

For the chronic aquatic invertebrate assessment, the database consisted of 13 studies with overall quality
determinations of high/medium. Of these studies, 8 contained acceptable chronic endpoints that
identified definitive hazard values below the DPB limit of water solubility for 10 aquatic invertebrate
species, resulting in robust confidence for quality of the database. DBP had similar effects on the same
species across multiple studies, within one order of magnitude. For example, in the water flea (Daphnia
magna), 21-day mortality studies resulted in paired NOEC/LOECs of 0.96/2.5 mg/L, and an LC50 of
1.92 mg/L, in independent studies. Paired 21-day NOEC/LOECs for reproductive effects on the number
of juveniles produced ranged from 0.42/0.48 mg/L to 0.96/2.5 mg/L across three independent studies. In
other species, effects on population, reproduction, and mortality were observed. Because effects were
similar across multiple studies and were seen at concentrations that were within an order of magnitude
within the same species, a robust confidence was assigned to the consistency consideration. In the study
chosen to derive the COC, (Tagatz et al.. 1983). populations of the marine amphipod Monocorophhm
acherusicum were reduced by 91 percent at the LOEC. Higher doses resulted in a complete loss of
amphipods in the aquaria. Similarly strong dose-response effects were observed in other studies in the
database. Because there was a strong biologically relevant effect and dose-response effects were
observed in the study chosen to derive the COC and among other studies in the database, a robust
confidence was assigned to the strength and precision consideration and dose-response consideration for
the chronic aquatic invertebrate assessment.

For the chronic benthic invertebrate assessment, the database consisted of three studies with overall
quality determinations of high. Reporting of these studies was extremely detailed and included multiple
species, endpoints, durations, and organic carbon contents, but only two species were represented.
Additionally, some of the results were repeated among the three studies and the author lists overlapped,
and it was unclear in some cases whether certain experiments were conducted independently among the
three studies. This lack of clarity about whether the studies were conducted independently resulted in a
moderate confidence assigned for the quality of the database consideration. In the studies examined, the
experiment was repeated with low, medium, and high TOC sediments and toxicity decreased with the
increase in TOC, as expected for a relatively hydrophobic compound like DBP based on equilibrium
partitioning theory. Among the same species, effects were generally within one order of magnitude for
repeated experiments in the same TOC. Because effects were seen at comparable concentrations within
species, a robust confidence was assigned to the consistency consideration. In the study chosen to derive
the COC, Lake Superior Research Institute (1997). population in the midge (Chironomus tentans) was
reduced by 76.7 percent at the LOEC, which was 3,090 mg DBP/kg dry sediment. Population reduction
in other treatments and TOC levels was generally as expected given equilibrium partitioning theory.
Because the effect size of DBP exposure was large, and other treatments resulted in effects that were as
expected based on equilibrium partitioning theory, a robust confidence was assigned to the strength and
precision consideration for the chronic benthic invertebrate assessment. Higher doses resulted in a
similar degree of population loss in the medium-TOC treatment; however, all population losses were
significantly different from controls. There was a clear dose-response effect observed in other studies in
the database, and among sub-studies using different TOC levels. Because dose-response was non-
monotonic in the medium-TOC treatment—but was as expected, with higher doses increasing the

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observed population loss, in other sub-studies involving different TOC levels within the same study—a
moderate confidence was assigned to the dose-response consideration for the chronic benthic
invertebrate assessment.

For the aquatic plants and algae assessment, the database consisted of seven high/medium quality
studies for toxicity in aquatic plants and algae. Of these studies, three contained acceptable endpoints
that identified definitive hazard values below the DBP limit of water solubility for one species of green
algae (Selencistrum capricornutum). Because only one species was identified, and several of the studies
in the database were not acceptable because exposure concentrations were above the limit of solubility
for DBP, confidence was decreased in the quality of the database. However, because three independent
studies were available in the species examined, a moderate confidence level was assigned for the quality
of the database. DBP had similar effects on population, measured as either chlorophyll a concentration
or cell abundance, in three independent studies. Effects were within an order of magnitude, ranging from
a 96-hour NOEC/LOEC pair at 0.1/1 mg/L to a 10-day EC50 at 0.75 mg/L. Because effects on the same
species were observed at DBP concentrations within one order of magnitude, a robust confidence was
assigned to the consistency consideration. In the study chosen to derive the COC, (Adachi et al.. 2006). a
significant reduction in the algal population at the LOEC, which was 1,000 |ig/L DBP, relative to an
increase in the algal population at the NOEC of 100 |ig/L DBP and in controls. The population reduction
was increased with a higher dose of DBP. Due to the increase in algal population at the NOEC relative
to controls, a moderate confidence was assigned to the strength and precision and dose-response
considerations for the aquatic plants and algae assessment.

For the terrestrial vertebrate assessment, the database consisted of 2 high/medium quality studies for
toxicity in environmentally relevant terrestrial vertebrates (chicken, Gallus gallus, and Japanese quail,
Coturmx japonica), supplemented by 13 high/medium quality studies for toxicity in human-relevant
terrestrial vertebrates (rat, Rattus norvegicus, and mouse, Mus musculus). Because 15 studies
representing four species were identified, a robust confidence was assigned to the quality of the
database. Among the two avian species, no effects were observed on growth at any DBP dose. Among
studies in rats, effects on reproduction were observed at NOEC/LOEC pairs ranging from 100/200
mg/kg-bw/day from gestational day 1 to 14 (Giribabu et al.. 2014). to 10,000/20,000 mg/kg-bw/day
from gestational day 0 to 20 (NTP. 1995). In mice, effects on reproduction were observed at
NOEC/LOEC pairs ranging from 50/300 mg/kg-bw/day from gestational day 7 to 9 (Xia et al.. 2011) to
10,000/20,000 mg/kg-bw/day from gestational day 0 to postnatal day 28 (NTP. 1995). Because effective
doses spanned two orders of magnitude among independent studies in the same species, but effective
doses for similar reproductive endpoints were much closer within each study, a moderate confidence
was assigned to the consistency consideration for terrestrial vertebrates. In the study chosen to derive the
HV, (NTP. 1995). 17-week LOAEL for significant reduction in number of live pups per litter was
identified at 80 mg/kg-bw/day DBP intake in dams. That study also found a LOAEL for reduced
bodyweight in F2 pups at the same dose (80 mg/kg-bw/day). The lowest bounded NOAEL/LOAEL pair
for which a ChV could be calculated was significantly reduced bodyweight in F1 pups at a ChV of 115.4
mg/kg-bw/day, but this effect was not as sensitive as reduced number of live pups per litter.

Other effects of DBP exposure included significantly decreased female body weight in dams,
significantly reduced male sex ratio (percentage of male pups), significantly decreased mating index and
pregnancy index in the F1 generation, and significantly reduced male pup weight gain. Because clear
dose-response relationships were found for many endpoints, robust confidence was assigned for the
dose-response consideration. However, the effect size for reduction in live pups per litter was relatively
small (a 7.8% reduction in litter size at the LOAEL, with a 17% reduction at the highest dose

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administered), leading to a moderate confidence for the strength and precision consideration for the
terrestrial vertebrate assessment.

For the soil invertebrate assessment, the database consisted of three high/medium quality studies, of
which two contained acceptable chronic endpoints that identified definitive hazard values below the
DPB limit of water solubility for two soil invertebrate species. Because only two high/medium quality
studies were identified that contained usable hazard values, and two species were represented, a
moderate confidence was assigned to the quality of the database. Among multiple endpoints and
lifestages, 21-day LC50 values in the springtail (Folsomiafimetaria) ranged from 19.4 mg/kg dry soil in
juveniles to 305 mg/kg dry soil in adults. No comparison to other studies was available for the EC10 and
EC50 values for reproduction in springtails, or for the 14-day LC50 value from a second study in the
earthworm (Eisenia fetida). Because comparisons among organisms within the same study or for the
same organisms among independent studies were not possible given the available data, but no
inconsistencies were observed among the studies examined (i.e., widely different toxicities among the
same organism), a moderate confidence evaluation was assigned to the consistency criterion.

In the study chosen to derive the HV, (Jensen et al.. 2001). reproduction was reduced by approximately
60 percent at the lowest concentration tested, which was 100 mg DBP/kg dry soil, with reproduction
completely eliminated at higher doses. Clear dose-response relationships were observed in other studies
in the data set for soil invertebrates. Because there was a strong biologically relevant effect and dose-
response effects were observed in the study chosen to derive the HV and among other studies in the
database, robust confidence was assigned to the strength and precision and dose-response criteria for the
soil invertebrate assessment.

For the terrestrial plant assessment, the database comprised 12 high/medium quality studies, of which 6
contained acceptable endpoints that identified definitive hazard values below the DBP limit of water
solubility for 10 terrestrial plant species. However, the majority of acceptable studies characterized
doses in a way that was unsuitable for a hazard determination (in mg/m3 soil fumigation, kg DBP/ha
agricultural application, or mg/kg wet soil). These dosing regimes made it impossible to characterize
dose in the unit EPA uses for exposure estimates to terrestrial plants, mg/kg dry soil. After filtering the
database to only those endpoints that characterized dose in mg/kg dry soil, four studies remained.
Because most of the studies characterized doses in a way that was not useful for developing a hazard
value, moderate confidence was assigned to the quality of the database. Effects on growth were
observed at a wide range of concentrations among terrestrial plants, ranging from unbounded 72- or 168-
hour LOECs at 5 mg/kg soil in agricultural crops including common oat (Avena sativa), alfalfa
(Medicago sativa), radish (Raphanus sativus), cucumber (Cucumis sativus), and common onion (Allium
cepa\ to an unbounded 72-hour NOEC at 500 mg/kg soil in perennial ryegrass (Loliumperetme) and a
72-hour EC50 at 1559 mg/kg dry soil in the mung bean (Vigna radiata).

Since consistent growth effects were seen in a variety of species, but the observed effects were
distributed over a wide range of concentrations, a moderate confidence was assigned to the consistency
consideration. In the study selected to derive the HV, (Gao et al.. 2019). the most sensitive endpoint was
the LOEL for reduction in leaf and root biomass in bread wheat seedlings at 10 mg/kg dry soil. There
was a clear dose-response observed, with biomass reduction increasing as the dose of DBP increased. At
the highest dose (40 mg/kg), root and leaf biomass were reduced by 29.93 and 32.10 percent,
respectively. However, for other studies in the data set, strong and precise effects of DBP on plant
growth were not observed, and dose-response was not observed in all studies. For example, in Zhao et
al. (2016). a 45-day exposure in false bok choy (Brassicaparachinensis) had a ChV of 31.62 mg
DBP/kg dry soil; however, the lowest dose (10 mg DBP/kg dry soil) resulted in statistically increased

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growth relative to controls. A strong biologically relevant effect was not observed among all studies in
the database, and dose-response effects were not observed among some studies in the database. Because
of the added uncertainty from some studies in similar plants showing a lack of strong biologically
relevant effects or clear dose-response, moderate confidence was assigned to the strength and precision
and dose-response considerations for the terrestrial plants assessment.

2.4.2 Relevance (Biological; Physical/Chemical; Environmental)

For the acute aquatic assessment, mortality was observed in the empirical data for 9 invertebrates and
fish, several of which are considered representative test species for aquatic assessments; mortality was
predicted in 53 additional species using Web-ICE. Although modeled approaches such as Web-ICE can
have more uncertainty than empirical data when determining the hazard or risk, the use of the
probabilistic approach within this risk evaluation increases confidence compared to a deterministic
approach. The use of the lower 95 percent CI of the HC05 in the SSD instead of a fixed AF also
increases confidence, as it is a more data-driven way of accounting for uncertainty. Because empirical
data was available for mortality for nine species, and predicted mortality data was available for 53 more
through Web-ICE, robust confidence was assigned to the relevance consideration for the acute aquatic
assessment.

For the chronic aquatic vertebrate assessment, ecologically relevant population level effects (growth and
mortality) were observed in seven different species, five of which are considered representative test
species for aquatic toxicity tests (African clawed frog, Xenopus laevis\ zebrafish, Danio rerio\ rainbow
trout, Oncorhynchus mykiss\ fathead minnow, Pimephales promelas; and Japanese medaka, Oryzicis
latipes). Because relevant population level effects were observed in several species, including
representative test species, robust confidence was assigned to the relevance consideration for the chronic
aquatic vertebrate assessment.

For the chronic aquatic invertebrate assessment, ecologically relevant population level effects (mortality
and reproduction) were observed in 10 species, 2 of which (water flea, Daphnia magna, and the worm
Lumbricuius variegatus) are considered representative test species for aquatic toxicity tests. Although
the COC was derived from a less-common species (the amphipod crustacean Monocorophium
acherusicum), effects on reproduction were seen at similar DBP doses in Daphnia magna, which
increases confidence in the biological relevance of effects that are expected to occur at the COC.
Because ecologically relevant effects were observed in 10 species, including 2 representative test
species, robust confidence was assigned to the relevance consideration for the chronic aquatic
invertebrate assessment.

For the chronic benthic invertebrate assessment, ecologically relevant population level effects (growth
and mortality) were observed in two different species (scud, Hyalella azteca, and midge, Chironomus
plumosus), both of which are considered representative test species for benthic toxicity tests. Because
ecologically relevant effects were observed in two representative test species, robust confidence was
assigned to the relevance consideration for the chronic benthic invertebrate assessment.

For the aquatic plant and algae assessment, an ecologically relevant population level effect (population
abundance, measured as either chlorophyll a concentration or cell count) was observed in one species of
green algae (Selenastrum capricornutum). This species is ubiquitous in the environment and is
considered a representative test species for algal toxicity tests. However, because only one species was
represented in the database, moderate confidence was assigned to the relevance consideration for the
aquatic plant and algae assessment.

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For the terrestrial vertebrate assessment, ecologically relevant population level effects were not observed
in ecologically relevant species. Data from human-relevant terrestrial vertebrates (rat, Rattus norvegicus,
and mouse, Mus musculus) were used to supplement the data set. A relevant population-level effect
(reproduction) was observed in both species. Because the study used to develop the COC was conducted
in rats, which are less ecologically relevant than other potential vertebrate species, moderate confidence
was assigned to the relevance consideration for the terrestrial vertebrate assessment.

For the soil invertebrate assessment, ecologically relevant endpoints (mortality and reproduction) were
observed for two ecologically relevant species (springtail, Folsomici fimetaria, and earthworm, Eisenici
fetida). Both species are considered representative test species for soil invertebrate toxicity testing.
Because ecologically relevant effects were observed in two representative test species, robust confidence
was assigned to the relevance consideration for the chronic benthic invertebrate assessment. Robust
confidence was also assigned to the relevance consideration for the soil invertebrate assessment.

For the terrestrial plant assessment, an ecologically relevant endpoint (growth) was observed for 10
plant species. However, of those species for which doses were measured in a way that was usable for
determining an HV (in mg/kg dry soil), only agricultural crops were represented. Additionally, for non-
food crop plants represented in the data set (Norway spruce, Picea abies, and perennial ryegrass, Lolium
perenne), no effects were observed at any tested DBP dose. This raises doubts whether ecologically
relevant effects of DBP exposure can be expected to occur in a non-agricultural context, so moderate
confidence was assigned to the relevance consideration for the terrestrial plant assessment.

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3 CONCLUSIONS	

EPA considered all reasonably available information identified through the systematic review process
under TSCA to characterize environmental hazard endpoints for DBP. The following bullets summarize
the hazard values and overall hazard confidence:

•	Aquatic species:

o LC50 values from nine exposures to DBP in fish and aquatic invertebrates were used
alongside quantitative structure-activity relationship (QSAR)-derived hazard estimates to
develop an SSD. The lower confidence interval of the HCos was used as the COC and
indicated that acute toxicity occurs at 347.6 |ig/L. EPA has robust confidence that this
hazard value represents the level of acute DBP exposure at which ecologically relevant
effects will occur in fish and aquatic invertebrates.

o A three-generation reproductive study in Japanese medaka (Oryzias latipes) found
significantly reduced bodyweight in F1 male fish after a 112-day exposure to DBP. The
COC based on this study indicated that chronic toxicity in aquatic vertebrates occurs at
1.56 |ig/L. EPA has robust confidence that this hazard value represents the level of
chronic DBP exposure at which ecologically relevant effects will occur in aquatic
vertebrates.

o A 14-day exposure to DBP in the marine amphipod crustacean Monocorophium

acherusicum found a significant reduction in population abundance. The COC based on
this study indicated that chronic toxicity in aquatic invertebrates occurs at 12.23 |ig/L.
EPA has robust confidence that this hazard value represents the level of chronic DBP
exposure at which ecologically relevant effects will occur in aquatic invertebrates.

o A 96-hour exposure to DBP in the green algae Selenastrum capricornutum found a
significant reduction in population growth. The COC based on this study indicated that
toxicity in aquatic plants and algae occurs at 31.6 |ig/L. EPA has moderate confidence
that this hazard value represents the level of DBP exposure at which ecologically relevant
effects will occur in algae, because hazard information for only one species was
identified in the database, and several of the studies in the database were not acceptable
since exposure concentrations were above the limit of solubility for DBP.

•	Benthic species:

o A 10-day exposure to DBP in the midge (Chironomus tentcms) in sediment found a
significant reduction in population abundance. The COC based on this study indicated
that chronic toxicity in benthic invertebrates occurs at 114.3 mg/kg dry sediment. EPA
has robust confidence that this hazard value represents the level of chronic DBP exposure
at which ecologically relevant effects will occur in benthic invertebrates.

•	Terrestrial species:

o A 17-week perinatal exposure to DBP in Sprague-Dawley rats (Rattus norvegicus) found
a significant reduction in number of live pups born per litter. The HV derived from this
study indicated that chronic toxicity in terrestrial vertebrates occurs at 80 mg/kg-bw/day.
EPA has moderate confidence that this hazard value represents the level of DBP exposure
at which ecologically relevant effects will occur in terrestrial vertebrates, because
effective doses for reproductive effects spanned two orders of magnitude among
independent studies in the same species, effect sizes were relatively small, and human-
toxicology model organisms were used instead of ecologically relevant species.

o A 21-day exposure to DBP in the springtail (Folsomia fimetarici) found a significant
reduction in reproduction. The HV derived from this study indicated that chronic toxicity
in soil invertebrates occurs at 14 mg/kg dry soil. EPA has robust confidence that this
hazard value represents the level of DBP exposure at which ecologically relevant effects

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will occur in soil invertebrates,
o A 40-day exposure to DBP in bread wheat (Triticum aestivum) found a significant

reduction in leaf and root biomass in seedlings. The HV derived from this study indicated
that toxicity in terrestrial plants occurs at 10 mg/kg dry soil. EPA has moderate
confidence that this hazard value represents the level of DBP exposure at which
ecologically relevant effects will occur in terrestrial plants, because most of the studies
characterized doses in a way that was not useful for developing a hazard value, and
because only agricultural crops were represented in the studies for which an adverse
effect of DBP exposure was observed.

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REFERENCES	

Abdul-Ghani. S; Yanai. J; Abdul-Ghani. R; Pinkas. A; Abdeen. Z. (2012). The teratogenicity and

behavioral teratogenicity of di(2-ethylhexyl) phthalate (DEHP) and di-butyl Phthalate (DBP) in a
chick model. Neurotoxicol Teratol 34: 56-62. http://dx.doi.Org/10.1016/i.ntt.2011.10.001
Adachi. A: Asa. K; Okano. T. (2006). Efficiency of rice bran for removal of di-n-butyl phthalate and its
effect on the growth inhibition of Selenastrum capricornutum by di-n-butyl phthalate. Bull
Environ Contam Toxicol 76: 877-882. http://dx.doi.org/10.1007/s00128-006-100Q-4
Adams. WJ: Biddinger. GR; Robillard. KA; Gorsuch. JW. (1995). A summary of the acute toxicity of 14
phthalate esters to representative aquatic organisms. Environ Toxicol Chem 14: 1569-1574.
http://dx.doi.org/10.1002/etc.562014Q916
Aoki. KA: Harris. CA; Katsiadaki. I; Sumpter. JP. (2011). Evidence suggesting that di-n-butyl phthalate
has antiandrogenic effects in fish. Environ Toxicol Chem 30: 1338-1345.
http: //dx. doi. or g/10.1002/etc. 5 02
BASF Aktiengesellschaft. (1989). Report on the study of the acute toxicity of dibutylphthalat on the

Golden Orfe (Leuciscus idus L., golden variety). (10F0449/895178). Ludwigshafen, Germany.
Battelle. (2018). 21-d amphibian metamorphosis assay (AMA) of dibutyl phthalate with African clawed
frog, xenopus laevis. (BATT01-00397). Washington, DC: U.S. Environmental Protection
Agency.

Bello. UM; Madekurozwa. M. -C: Groenewald. HB; Aire. TA; Arukwe. A. (2014). The effects on

steroidogenesis and histopathology of adult male Japanese quails (Coturnix coturnix japonica)
testis following pre-pubertal exposure to di(n-butyl) phthalate (DBP). Comp Biochem Physiol C
Toxicol Pharmacol 166: 24-33. http://dx.doi.Org/10.1016/i.cbpc.2014.06.005
Bencic. DC: Flick. RW: Bell. ME: Henderson. WM; Huang. W: Purucker. ST: Glinski. DA: Blackwell.
BR: Christen. CH; Stacy. EH: Biales. AD. (2024). A multiomics study following acute exposures
to phthalates in larval fathead minnows (Pimephales promelas) - The potential application of
omics data in risk evaluations under TSCA (internal use only). (EPA/600/X-24/098). Cincinnati,
OH: U.S. Environmental Protection Agency.

Bhatia. H; Kumar. A: Chapman. JC: McLaughlin. MJ. (2015). Long-term exposures to di-n-butyl

phthalate inhibit body growth and impair gonad development in juvenile Murray rainbowfish
(Melanotaenia fluviatilis). J Appl Toxicol 35: 806-816. http://dx.doi.org/10.1002/iat.3076
Bhatia. H; Kumar. A: Du. J: Chapman. J: McLaughlin. MJ. (2013). Di-n-butyl phthalate causes

antiestrogenic effects in female murray rainbowfish (Melanotaenia fluviatilis). Environ Toxicol
Chem 32: 2335-2344. http://dx.doi.org/10.1002/etc.2304
Bhatia. H; Kumar. A: Ogino. Y; Gregg. A: Chapman. J: McLaughlin. MJ: Iguchi. T. (2014). Di-n-butyl
phthalate causes estrogenic effects in adult male Murray rainbowfish (Melanotaenia fluviatilis).
Aquat Toxicol 149: 103-115. http://dx.doi.Org/10.1016/i.aquatox.2014.01.025
Buccafusco. RJ: Ells. SJ: Leblanc. GA. (1981). Acute toxicity of priority pollutants to bluegill (Lepomis
macrochirus). Bull Environ Contam Toxicol 26: 446-452. http://dx.doi.org/10.1007/BFQ1622118
Burnham. KP; Anderson. DR. (2002). Model selection and multimodel inference: a practical
information-theoretic approach (2nd ed.). New York: Springer.

http://www.springer.com/statistics/statistical+theory+and+methods/book/978-0-387-95364-9
Call. DJ: Cox. DA: Geiger. PL: Genisot. KI; Markee. TP: Brooke. LT; Polkinghorne. CN: Vandeventer.
FA: Gorsuch. JW: Robillard. KA: Parkerton. TF: Reilev. MC: Anklev. GT: Mount. DR. (2001a).
An assessment of the toxicity of phthalate esters to freshwater benthos. 2. Sediment exposures.
Environ Toxicol Chem 20: 1805-1815. http://dx.doi.org/10.1002/etc.562020Q826
Call. DJ: Markee. TP: Geiger. PL: Brooke. LT: Vandeventer. FA: Cox. DA: Genisot. KI; Robillard.

KA; Gorsuch. JW; Parkerton. TF; Reilev. MC; Anklev. GT; Mount. DR. (2001b). An assessment
of the toxicity of phthalate esters to freshwater benthos. 1. Aqueous exposures. Environ Toxicol
Chem 20: 1798-1804. http://dx.doi.org/10.1002/etc.562020Q825

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1177

1178

1179

1180

1181

1182

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PUBLIC RELEASE DRAFT
December 2024

Chen. P; Li. S; Liu. L; Xu. N. (2015). Long-term effects of binary mixtures of 17a-ethinyl estradiol and
dibutyl phthalate in a partial life-cycle test with zebrafish (Danio rerio). Environ Toxicol Chem
34: 518-526. http://dx.doi.org/10.1002/etc.2803
Chen. X: Xu. S: Tan. T: Lee. ST: Cheng. SH: Lee. FWF: Xu. SJL; Ho. KC. (2014). Toxicity and
estrogenic endocrine disrupting activity of phthalates and their mixtures. Int J Environ Res
Public Health 11: 3156-3168. http://dx.doi.org/10.3390/iierphl 10303156
Chuah. TS: Oh. HY; Habsah. M; Norhafizah. MZ; Ismail. BS. (2014). Potential of crude extract and

isolated compounds from golden beard grass (Chrysopogon serrulatus) for control of sprangletop
(Leptochloa chinensis) in aerobic rice systems. Crop and Pasture Science 65: 461-469.
http://dx.doi.org/10.1071/CP13339
Cruciani. V: Iovine. C: Thome. JP; Joaquim-Justo. C. (2015). Impact of three phthalate esters on the

sexual reproduction of the Monogonont rotifer, Brachionus calyciflorus. Ecotoxicology 25: 192-
200. http://dx.doi.org/10.1007/slQ646-015-1579-5
Defoe. PL: Holcombe. GW: Hammermeister. DE; Biesinger. KE. (1990). Solubility and toxicity of

eight phthalate esters to four aquatic organisms. Environ Toxicol Chem 9: 623-636.

Deng. J: Zhang. Y; Hu. J: Jiao. J: Hu. F; Li. H; Zhang. S. (2017). Autotoxicity of phthalate esters in

tobacco root exudates: Effects on seed germination and seedling growth. Pedosphere 27: 1073-
1082. http://dx.doi.org/10.1016/S1002-0160( 17)60374-6
Du. L; Li. G: Liu. M; Li. Y; Yin. S: Zhao. J. (2015). Biomarker responses in earthworms (Eisenia fetida)
to soils contaminated with di-n-butyl phthalates. Environ Sci Pollut Res Int 22: 4660-4669.
http://dx.doi.org/10.1007/sll356-014-3716-8
Dueck. TA; Van Dijk. CJ: David. F; Scholz. N: Vanwalleghem. F. (2003). Chronic effects of vapour
phase di-n-butyl phthalate (DBP) on six plant species. Chemosphere 53: 911-920.
http://dx.doi.org/10.1016/S0045-6535(03)00580-0
EAG Laboratories. (2018). Dibutyl phthalate: Medaka extended one generation reproduction test (final

report). (83260). Washington, DC: U.S. Environmental Protection Agency.

EG&G Bionomics. (1983a). Acute toxicity of fourteen phthalate esters to rainbow trout (Salmo
gairdneri) under flow-through conditions (final report) report no BW-83-3-1373 [TSCA
Submission], (Bionomics Report No. BW-83-3-1373. OTS0508403. 42005 B4-5. 40-8326144.
TSCATS/206776). Washington, DC: Chemical Manufacturers Association.
https://ntrl.ntis.gov/NTRL/dashboard/searchResults/titleDetail/OTS05084Q3.xhtml
EG&G Bionomics. (1983b). Exhibit III: Acute toxicity of thirteen phthalate esters to bluegill (Lepomis
macrochirus) [TSCA Submission], In Exhibit III: Acute toxicity of thirteen phthalate esters to
fathead minnow (Pimephales promelas) under flow-through conditions. (Bionomics report No.
BW-83-3-1368. OTS0508481. 42005 G5-2. 40-8326129. TSCATS/038115). Washington, DC:
Chemical Manufacturers Association.

https://ntrl.ntis.gov/NTRL/dashboard/searchResults/titleDetail/QTS0508481.xhtml
EG&G Bionomics. (1984a). Acute toxicity of thirteen phthalate esters to fathead minnows (Pimephales
promelas) under flow-through conditions [TSCA Submission], (BW-83-3-1374; EPA/OTS Doc
#FYI-AX-0184-0286). Washington, DC: Chemical Manufacturers Association.
https://ntrl.ntis.gov/NTRL/dashboard/searchResults/titleDetail/OTS0000286Q.xhtml
EG&G Bionomics. (1984b). Acute toxicity of twelve phthalate esters to mysid shrimp (Mysidopsis
bahia) [TSCA Submission], (EPA/OTS Doc #40-8426078). Washington, DC: Chemical
Manufacturers Association.

https://ntrl.ntis.gov/NTRL/dashboard/searchResults/titleDetail/OTS05084Q5.xhtml
EG&G Bionomics. (1984c). Acute toxicity of twelve phthalate esters to Paratanytarsus parthenogenica
(final report) report no BW-83-6-1424 [TSCA Submission], (EPA/OTS Doc #40-8426146).
Chemical Manufacturers Association.

https://ntrl.ntis.gov/NTRL/dashboard/searchResults/titleDetail/OTS05084Q4.xhtml

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PUBLIC RELEASE DRAFT
December 2024

Ema, M; Amano. H; Itami. T; Kawasaki. H. (1993). Teratogenic evaluation of di-n-butyl phthalate in

rats. Toxicol Lett 69: 197-203. http://dx.doi.org/10.1016/03 78-4274(93)90104-6
EnviroSvstem. (1991). Early life-stage toxicity of di-n-butyl phthalate (DnBP) to the rainbow trout
(Oncorhynchus mykiss) under flow-through conditions [TSCA Submission], (9102-CMA.
OTS0533141. 42005 L5-5. 40-9126399). Washington, DC: Chemical Manufacturers
Association. https://ntrl.ntis.gov/NTRL/dashboard/searchResults/titleDetail/OTSQ533141.xhtml
Erkmen. B; Benli. ACK; Agus. HH; Yildirim. Z; Mert. R; Erkoc. F. (2017). Impact of sublethal di-n-
butyl phthalate on the aquaculture fish species Nile tilapia (Oreochromis niloticus):
Histopathology and oxidative stress assessment. Aquaculture Research 48: 675-685.
http://dx.doi.Org/10.l 111/are. 12914
Etterson. M. (2020). Species Sensitivity Distribution (SSD) Toolbox. Duluth, MN: U.S. Environmental
Protection Agency. Retrieved from https://www.epa.gov/sciencematters/species-sensitivitv-
distribution-toolbox-new-tool-identifv-and-protect-vulnerable
Gao. M; Dong. Y; Zhang. Z; Song. W: Oi. Y. (2017). Growth and antioxidant defense responses of

wheat seedlings to di-n-butyl phthalate and di (2-ethylhexyl) phthalate stress. Chemosphere 172:
418-428. http://dx.doi.org/10.1016/i.chemosphere.2017.01.034
Gao. M; Guo. Z; Dong. Y; Song. Z. (2019). Effects of di-n-butyl phthalate on photosynthetic

performance and oxidative damage in different growth stages of wheat in cinnamon soils.
Environ Pollut 250: 357-365. http://dx.doi.org/10.1016/i.envpol.2019.04.022
Gardner. ST: Wood. AT: Lester. R; Onkst. PE; Burnham. N: Perygin. DH; Ravburn. J. (2016).

Assessing differences in toxicity and teratogenicity of three phthalates, Diethyl phthalate, Di-n-
propyl phthalate, and Di-n-butyl phthalate, using Xenopus laevis embryos. J Toxicol Environ
Health A 79: 71-82. http://dx.doi.org/10.1080/15287394.2Q15.1106994
Giribabu. N: Sainath. SB: Reddv. PS. (2014). Prenatal di-n-butyl phthalate exposure alters reproductive
functions at adulthood in male rats. Environ Toxicol 29: 534-544.
http: //dx. doi. or g/10.1002/tox. 217 79
Gray. LE; Ostbv. J: Sigmon. R; Ferrell. J: Rehnberg. G: Linder. R; Cooper. R; Goldman. J: Laskev. J.
(1988). The development of a protocol to assess reproductive effects of toxicants in the rat
[Review], Reprod Toxicol 2: 281-287. http://dx.doi.org/10.1016/0890-6238(88)90032-9
Gu. S: Zheng. H; Xu. O; Sun. C: Shi. M; Wang. Z; Li. F. (2017). Comparative toxicity of the plasticizer
dibutyl phthalate to two freshwater algae. Aquat Toxicol 191: 122-130.
http://dx.doi.Org/10.1016/i.aquatox.2017.08.007
Huang. B; Li. D; Yang. Y. (2016). Joint toxicity of two phthalates with waterborne copper to Daphnia
magna and Photobacterium phosphoreum. Bull Environ Contam Toxicol 97: 380-386.
http://dx.doi.org/10.1007/sQ0128-016-1879-3
Isogai. Y; Komoda. Y; Okamoto. T. (1972). Biological activities of n-butyl phthalate and its analogous
compounds on various bioassays of plant growth regulators. Scientific papers of the College of
General Education, University of Tokyo 22: 129-135.

Jee. JH; Koo. JG: Keum. YH; Park. KH; Choi. SH; Kang. JC. (2009). Effects of dibutyl phthalate and
di-ethylhexyl phthalate on acetylcholinesterase activity in bagrid catfish, Pseudobagrus
fulvidraco (Richardson). J Appl Ichthyol 25: 771-775. http://dx.doi.org/10.Ill 1/i. 1439-
0426.2009.01331.x

Jensen. J: van Langevelde. J: Pritzl. G: Krogh. PH. (2001). Effects of di(2-ethylhexyl) phthalate and

dibutyl phthalate on the collembolan Folsomia fimetaria. Environ Toxicol Chem 20: 1085-1091.
http://dx.doi.org/10.1002/etc.562020052Q
Kang. SW: Kim. HK; Lee. WJ: Ahn. YJ. (2006). Toxicity of bisabolangelone from Ostericum koreanum
roots to Dermatophagoides farinae and Dermatophagoides pteronyssinus (Acari :

Pyroglyphidae). J AgricFood Chem 54: 3547-3550. http://dx.doi.org/10.1021/if060140d
Khalil. S. R.; Abd Elhakim. Y; El-Murr. AE. (2016). Sublethal concentrations of di-n-butyl phthalate

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December 2024

promote biochemical changes and DNA damage in juvenile Nile tilapia (Oreochromis niloticus).
Jpn J Vet Res 64: 67-80.

Kim. HK; Tak. JH; Ahn. YJ. (2004). Acaricidal activity of Paeonia suffruticosa root bark-derived
compounds against Dermatophagoides farinae and Dermatophagoides pteronyssinus (Acari:
Pyroglyphidae). J AgricFood Chem 52: 7857-7861. http://dx.doi.org/10.1021/if048708a
Kim. HK: Yun. YK; Ahn. YJ. (2007). Toxicity of atractylon and atractylenolide III identified in

Atractylodes ovata rhizome to Dermatophagoides farinae and Dermatophagoides pteronyssinus.
J AgricFood Chem 55: 6027-6031. http://dx.doi.org/10.102l/if0708802
Kim. HK: Yun. YK: Ahn. YJ. (2008). Fumigant toxicity of cassia bark and cassia and cinnamon oil
compounds to Dermatophagoides farinae and Dermatophagoides pteronyssinus (Acari:
Pyroglyphidae). Exp Appl Acarol 44: 1-9. http://dx.doi.org/10.1007/slQ493-008-9129-v
Kong. X: Jin. D; Jin. S: Wang. Z; Yin. H; Xu. M; Deng. Y. (2018). Responses of bacterial community to
dibutyl phthalate pollution in a soil-vegetable ecosystem. J Hazard Mater 353: 142-150.
http://dx.doi.Org/10.1016/i.ihazmat.2018.04.015
Kuang. QJ: Zhao. WY; Cheng. SP. (2003). Toxicity of dibutyl phthalate to algae. Bull Environ Contam

Toxicol 71: 602-608. http://dx.doi.org/10.1007/sQ0128-003-8559-9
Lake Superior Research Institute. (1997). Sediment toxicity testing program for phthalate esters.

(Unpublished Report PE-88.0-SED-WIS). Arlington, VA: Chemical Manufacturers Association.
Lamb. J: Chapin. R; Teague. J: Lawton. A: Reel. J. (1987). Reproductive effects of four phthalic acid
esters in the mouse. Toxicol Appl Pharmacol 88: 255-269. http://dx.doi.org/10.1016/0Q41-
008X(87)90011-1

Laughlin Rb. JR; Neff. JM; Hrung. YC: Goodwin. TC: Giam. CS. (1978). The effects of three phthalate
esters on the larval development of the grass shrimp Palaemonetes pugio (Holthuis). Water Air
Soil Pollut 9: 323-336.

Lee. SK; Owens. GA; Veeramachaneni. DN. (2005). Exposure to low concentrations of di-n-butyl
phthalate during embryogenesis reduces survivability and impairs development of Xenopus
laevis frogs. J Toxicol Environ Health A 68: 763-772.
http://dx.doi.org/10.1080/1528739059093Q243
Lee. SK: Veeramachaneni. DNR. (2005). Subchronic exposure to low concentrations of di-n-butyl
phthalate disrupts spermatogenesis in Xenopus laevis frogs. Toxicol Sci 84: 394-407.
http://dx.doi.org/10.1093/toxsci/kfi087
Liao. CS: Yen. JH: Wang. YS. (2009). Growth inhibition in Chinese cabbage (Brassica rapa var.
chinensis) growth exposed to di-n-butyl phthalate. J Hazard Mater 163: 625-631.
http://dx.doi.Org/10.1016/i.ihazmat.2008.07.025
Linden. E; Bengtsson. BE: Svanberg. O; Sundstrom. G. (1979). The acute toxicity of 78 chemicals and
pesticide formulations against two brackish water organisms, the bleak (Alburnus alburnus) and
the harpacticoid Nitocra spinipes. Chemosphere 8: 843-851. http://dx.doi.org/10.1016/0045-
653 5(79)90015-8

Liu. Y; Guan. Y; Yang. Z: Cai. Z: Mizuno. T; Tsuno. H: Zhu. W: Zhang. X. (2009). Toxicity of seven
phthalate esters to embryonic development of the abalone Haliotis diversicolor supertexta.
Ecotoxicology 18: 293-303. http://dx.doi.org/10.1007/sl0646-008-Q283-0
Lokke. H: Rasmussen. L. (1983). Phytotoxicological effects of Di-(2-ethyl hexyl)-phthalate and Di-n-
butyl-phthalate on higher plants in laboratory and field experiments. Environ Pollut Ser A 32:
179-199. http://dx.doi.org/10.1016/0143-1471(83)90035-1
Ma. T; Teng. Y; Christie. P; Luo. Y. (2015). Phytotoxicity in seven higher plant species exposed to di-n-
butyl phthalate or bis (2-ethylhexyl) phthalate. Front Env Sci Eng 9: 259-268.
http://dx.doi.org/10.1007/sll783-014-Q652-2
Ma. TT: Christie. P; Luo. YM; Teng. Y. (2014). Physiological and antioxidant responses of germinating
mung bean seedlings to phthalate esters in soil. Pedosphere 24: 107-115.

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1328

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1330

1331

PUBLIC RELEASE DRAFT
December 2024

http://dx.doi.org/10.1016/S 1002-0160(13)60085-5
McCarthy. JF; Whitmore. DK. (1985). Chronic toxicity of di-n-butyl and di-n-octyl phthalate to
daphnia-magna and the fathead minnow. Environ Toxicol Chem 4: 167-179.
http://dx.doi.org/10.1002/etc.56200402Q6
Medlin. LK. (1980). Effects of di-n-butyl phthalate and salinity on the growth of the diatom
Skeletonema costatum. Bull Environ Contam Toxicol 25: 75-78.
http://dx.doi.org/10.1007/BF0198549Q
Melin. C: Egneus. H. (1983). Effects of di-n-butyl phthalate on growth and photosynthesis in algae and
on isolated organelles from higher plants. Physiol Plant 59: 461-466.
http://dx.doi.org/10.1111/i. 1399-3054.1983.tb04230.x
Misra. S: Singh. A: Ch. R; Sharm a. V: Mudiam. MKR; Ram. KR. (2014). Identification of Drosophila-
based endpoints for the assessment and understanding of xenobiotic-mediated male reproductive
adversities. Toxicol Sci 141: 278-291. http://dx.doi.org/10.1093/toxsci/kful25
Mylchreest. E; Cattlev. RC: Foster. PMD. (1998). Male reproductive tract malformations in rats
following gestational and lactational exposure to di(n-butyl) phthalate: An antiandrogenic
mechanism? Toxicol Sci 43: 47-60. http://dx.doi.org/10.1006/toxs.1998.2436
Neuhauser. EF; Loehr. RC: Malecki. MR: Milligan. PL: Durkin. PR. (1985). The toxicity of selected
organic chemicals to the earthworm Eisenia fetida. J Environ Qual 14: 383-388.
http://dx.doi.org/10.2134/ieal985.00472425001400030Q15x
Nikonorow. M; Mazur. H: Piekacz. H. (1973). Effect of orally administered plasticizers and polyvinyl
chloride stabilizers in the rat. Toxicol Appl Pharmacol 26: 253-259.
http://dx.doi.org/10.1016/0041-008X(73)90259-7
Ntp. (1984). Di(n-butyl) phthalate: Reproduction and fertility assessment in CD-I mice when

administered in the feed (pp. 1-197). (NTP-84-411). Research Triangle Park, NC: National
Toxicology Program, National Institute of Environmental Health Sciences.
http://ntp.niehs.nih.gov/testing/types/repro/abstracts/racb/index-14.html
NTP. (1995). NTP technical report on the toxicity studies of dibutyl phthalate (CAS No. 84-74-2)

administered in feed to F344/N rats and B6C3F1 mice (pp. 1-G5). (ISSN 1521-4621
Toxicity Report Series Number 30; NIH Publication 95-3353). Research Triangle Park, NC: National

Toxicology Program. https://ntp.niehs.nih.gov/publications/reports/tox/000s/toxQ30
Ohtani. H: Miura. I: Ichikawa. Y. (2000). Effects of dibutyl phthalate as an environmental endocrine

disruptor on gonadal sex differentiation of genetic males of the frog Rana rugosa. Environ Health
Perspect 108: 1189-1193. http://dx.doi.org/10.2307/3434832
Ortiz-Zarragoitia. M; Trant. JM; Caiaravillet. MP. (2006). Effects of dibutylphthalate and

ethynylestradiol on liver peroxisomes, reproduction, and development of zebrafish (Danio rerio).
Environ Toxicol Chem 25: 2394-2404. http://dx.doi.Org/10.1897/05-456R.l
Patyna. PJ. (1999) Reproductive effects of phthalate esters in Japanese medaka (Oryzias latipes).

(Doctoral Dissertation). Rutgers The State University of New Jersey - New Brunswick, New
Brunswick, NJ. Retrieved from https://primo.lib.umn.edu/primo-
explore/openurl?url ver=Z39.88-

2004&rft val fmt=info:ofi%2Ffmt:kev:mtx:dissertation&genre=dissertations%20%26%20these
s&sid=ProO:Dissertations%20%26%20Theses%20@%20CIC%20Institutions&atitle=&title=Re
productive%20effects%20of%20phthalate%20esters%20in%20japanese%20medaka%20(0ryzia
s%201atipes)&issn=&date= 1999-01 -

01&volume=&issue=&spage=&au=Patvna.%20Przemyslaw%20J.&isbn=0599385669&ititle=&
btitle=&rft id=info:eric%2F&rft id=info:doi%2F&vid=DULUTH&institution=DULUTH&url
ctx val=&url ctx fmt=null&isSerivcesPage=true
Rhodes. JE; Adams. WJ: Biddinger. GR; Robillard. KA; Gorsuch. JW. (1995). Chronic toxicity of 14
phthalate esters to Daphnia magna and rainbow trout (Oncorhynchus mykiss). Environ Toxicol

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1369

1370

1371

1372

1373

1374

1375

1376

1377

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1379

1380

PUBLIC RELEASE DRAFT
December 2024

Chem 14: 1967-1976. http://dx.doi.org/10.1002/etc.5620141119
Sevoum. A: Pradhan. A. (2019). Effect of phthalates on development, reproduction, fat metabolism and
lifespan in Daphnia magna. Sci Total Environ 654: 969-977.
http: //dx. doi. or g/10.1016/i. scitotenv .2018.11.158
Shen. O: Wu. W: Du. G: Liu. R: Yu. L: Sun. H: Han. X: Jiang. Y: Shi. W: Hu. W: Song. L: Xia. Y:

Wang. S: Wang. X. (2011). Thyroid disruption by Di-n-butyl phthalate (DBP) and mono-n-butyl
phthalate (MBP) in Xenopus laevis. PLoS ONE 6: el9159.
http://dx.doi.org/10.1371/iournal.pone.0019159
Shin. N: Cuenca. L; Karthikrai. R; Kannan. K; Colaiacovo. MP. (2019). Assessing effects of germline

exposure to environmental toxicants by high-throughput screening in C. elegans. PLoS Genet 15:
el007975. http://dx.doi.org/10.1371/iournal.pgen.1007975
Shiota. K; Chou. MJ; Nishimura. H. (1980). Embryotoxic effects of di-2-ethylhexyl phthalate (DEHP)
and di-n-butyl phthalate (DBP) in mice. Environ Res 22: 245-253.
http://dx.doi.org/10.1016/0013-9351(80)90136-X
Shiota. K; Nishimura. H. (1982). Teratogenicity of di(2-ethylhexyl) phthalate (DEHP) and di-n-butyl

phthalate (DBP) in mice. Environ Health Perspect 45: 65-70. http://dx.doi.org/10.2307/3429385
Smithers Viscient. (2018). Di-n-butyl phthalate - short-term reproduction assay with fathead minnow
(Pimephales promelas) following OPPTS 890.1350 and OECD 229 guidelines. (Smithers
Viscient Study No. 13784.6123). Washington, DC: U.S. Environmental Protection Agency.
Springborn Bionomics. (1984a). Acute toxicity of thirteen phthalate esters to the sheepshead minnow
(Cyprinodon variegatus) (final report) [TSCA Submission], (BP-84-2-14/10823.8000.
OTS0508409. 40-8426151. 42005 B4-11. TSCATS/206782). Washington, DC: Chemical
Manufacturers Association.

https://ntrl.ntis.gov/NTRL/dashboard/searchResults/titleDetail/OTS05084Q9.xhtml
Springborn Bionomics. (1984b). Chronic toxicity of fourteen phthalate esters to Daphnia magna with
cover letter dated 032585 [TSCA Submission] (pp. 95). (Report No. BW-84-5-1567.
OTS0000392-0. FYI-AX-0485-0392. TSCATS/032642). Wareham, MA: Chemical
Manufacturers Association.

https://ntrl.ntis.gov/NTRL/dashboard/searchResults/titleDetail/OTS0000392Q.xhtml
Springborn Bionomics. (1984c). FYI Submission: Toxicity of fourteen phthalate esters to the freshwater
green alga Selenastrum capricornutum [TSCA Submission], (EPA/OTS Doc #FYI-OTS-0485-
0392). Washington, DC: Chemical Manufacturers Association.
https://ntrl.ntis.gov/NTRL/dashboard/searchResults/titleDetail/OTS0000392Q.xhtml
Streufort. JM. (1978). Some effects of two phthalic acid esters on the life cycle of the midge
(Chironomus plumosus) [TSCA Submission], (OTS0000013-0. FYI-AX-1178-0013.
TSC ATS/029296). Washington, DC: Manufacturing Chemists Association.
https://ntrl.ntis.gov/NTRL/dashboard/searchResults/titleDetail/OTS0000013Q.xhtml
Tagatz. ME: Deans. CH; Moore. JC: Plaia. GR. (1983). Alterations in composition of field-developed
and laboratory-developed estuarine benthic communities exposed to di-normal-butyl phthalate.
Aquat Toxicol 3: 239-248. http ://dx. doi .org/10.1016/0166-445X(83 )90044-9
Tak. JH; Kim. HK; Lee. SH; Ahn. YJ. (2006). Acaricidal activities of paeonol and benzoic acid from
Paeonia suffruticosa root bark and monoterpenoids against Tyrophagus putrescentiae (Acari:
Acaridae). Pest Manag Sci 62: 551-557. http://dx.doi.org/10.1002/ps.1212
Thuren. A: Woin. P. (1991). Effects of phthalate esters on the locomotor activity of the freshwater
amphipod Gammarus pulex. Bull Environ Contam Toxicol 46: 159-166.
http://dx.doi.org/10.1007/BF0168827Q
U.S. EPA. (1998). Guidelines for ecological risk assessment [EPA Report], (EPA/630/R-95/002F).
Washington, DC: U.S. Environmental Protection Agency, Risk Assessment Forum.
https://www.epa.gov/risk/guidelines-ecological-risk-assessment

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1420

1421

1422

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December 2024

U.S. EPA. (2005). Guidelines for carcinogen risk assessment [EPA Report], (EPA630P03001F).
Washington, DC. https://www.epa.gov/sites/production/files/2013-
09/documents/cancer guidelines final 3-25-05.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. (2014). Framework for human health risk assessment to inform decision making. Final [EPA
Report], (EPA/100/R-14/001). Washington, DC: U.S. Environmental Protection, Risk
Assessment Forum, https://www.epa.gov/risk/framework-human-health-risk-assessment-inform-
decision-making

U.S. EPA. (2016). Weight of evidence in ecological assessment [EPA Report], (EPA/100/R-16/001).
Washington, DC: Office of the Science Advisor.
https://nepis.epa. gov/Exe/ZvPURL.cgi?Dockev=P100SFXR.txt
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. (2024a). Draft Environmental Exposure Assessment for Diisodecyl Phthalate (DIDP).
Washington, DC: Office of Pollution Prevention and Toxics.
https://www.regulations.gov/document/EPA-HQ-OPPT-2024-0073-0024
U.S. EPA. (2024b). Draft Non-cancer Human Health Hazard Assessment for Dibutyl Phthalate (DBP).

Washington, DC: Office of Pollution Prevention and Toxics.

U.S. EPA. (2024c). Draft Systematic Review Protocol for Dibutyl Phthalate (DBP). Washington, DC:

Office of Pollution Prevention and Toxics.

Wang. Z; Kim. HK; Tao. W: Wang. M; Ahn. YJ. (2011). Contact and fumigant toxicity of

cinnamaldehyde and cinnamic acid and related compounds to Dermatophagoides farinae and
Dermatophagoides pteronyssinus (Acari: Pyroglyphidae). J Med Entomol 48: 366-371.
http://dx.doi.org/10.1603/ME10127
Wei. J: Shen. O: Ban. Y; Wang. Y; Shen. C: Wang. T; Zhao. W: Xie. X. (2018). Characterization of

Acute and Chronic Toxicity of DBP to Daphnia magna. Bull Environ Contam Toxicol 101: 214-
221. http://dx.doi.org/10.1007/sQ0128-018-2391-8
Williams. MJ; Wiemerslage. L; Gohel. P; Kheder. S: Kothegala. LV; Schioth. HB. (2016). Dibutyl

Phthalate Exposure Disrupts Evolutionarily Conserved Insulin and Glucagon-Like Signaling in
Drosophila Males. Endocrinology 157: 2309-2321. http://dx.doi.org/10.1210/en.2015-20Q6
Wine. RN: Li. LH; Barnes. LH; Gulati. DK; Chapin. RE. (1997). Reproductive toxicity of di-n-
butylphthalate in a continuous breeding protocol in Sprague-Dawley rats. Environ Health
Perspect 105: 102-107. http://dx.doi.org/10.1289/ehp.97105102
Wolf. C: Lambright. C: Mann. P; Price. M; Cooper. RL; Ostbv. J: Gray. LE. Jr. (1999). Administration
of potentially antiandrogenic pesticides (procymidone, linuron, iprodione, chlozolinate, p,p'-
DDE, and ketoconazole) and toxic substances (dibutyl- and diethylhexyl phthalate, PCB 169,
and ethane dimethane sulphonate) during sexual differentiation produces diverse profiles of
reproductive malformations in the male rat. Toxicol Ind Health 15: 94-118.
http://dx.doi.org/10.1177/0748233799015001Q9
Xia. H: Chi. Y. i: Oi. X: Su. M: Cao. Y: Song. P: Li. X: Chen. T: Zhao. A: Zhang. Y: Cao. Y: Ma. X:
Jia. W. (2011). Metabolomic evaluation of di-n-butyl phthalate-induced teratogenesis in mice.
Metabolomics 7: 559-571. http://dx.doi.org/10.1007/sll306-011-0276-5
Xu. Y; Gve. MC. (2018). Developmental toxicity of dibutyl phthalate and citrate ester plasticizers in
Xenopus laevis embryos. Chemosphere 204: 523-534.

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http: //dx. doi. or g/ 10.1016/i. chemosphere .2018.04.077
Yang. ZH; Zhang. XJ; Cat ZH. (2009). Toxic effects of several phthalate esters on the embryos and
larvae of abalone Haliotis diversicolor supertexta. Chin J Oceanol Limnol 27: 395-399.
http://dx.doi.org/10.1007/s00343-009-91Q3-5
Zhao. HM: Du. H: Xiang. L: Li. YW: Li. H: Cai. OY: Mo. CH: Cao. G: Wong. MH. (2016).

Physiological differences in response to di-n-butyl phthalate (DBP) exposure between low- and
high-DBP accumulating cultivars of Chinese flowering cabbage (Brassica parachinensis L.).
Environ Pollut 208: 840-849. http://dx.doi.org/10.1016/i.envpol.2015.11.009
Zhao. LL; Xi. YL; Huang. L; Zha. CW. (2009). Effects of three phthalate esters on the life-table

demography of freshwater rotifer Brachionus calyciflorus Pallas. Aquatic Ecology 43: 395-402.
http://dx.doi.org/10.1007/slQ452-008-9179-6

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APPENDICES

Appendix A RUBRIC FOR WEIGHT OF THE SCIENTIFIC
	EVIDENCE	

The weight of the scientific evidence fundamentally means that the evidence is weighed (i.e., ranked)
and weighted (i.e., a piece or set of evidence or uncertainty may have more importance or influence in
the result than another). Based on the weight of the scientific evidence and uncertainties, a confidence
statement was developed that qualitatively ranks (i.e., robust, moderate, slight, or indeterminate) the
confidence in the hazard threshold. The qualitative confidence levels are described below.

The evidence considerations and criteria detailed within U.S. EPA (2021) guides the application of
strength-of-evidence judgments for environmental hazard effect within a given evidence stream and
were adapted from Table 7-10 of the 2021 Draft Systematic Review Protocol (U.S. EPA. 2021).

EPA used the strength-of-evidence and uncertainties from U.S. EPA (2021) for the hazard assessment to
qualitatively rank the overall confidence rating for environmental hazard (Table Apx A-l). Confidence
levels of robust (+ + +), moderate (+ +), slight (+), or indeterminant are assigned for each evidence
property that corresponds to the evidence considerations (U.S. EPA. 2021). The rank of the Quality of
the Database consideration is based on the systematic review overall quality determination (High,
Medium, or Low) for studies used to calculate the hazard threshold, and whether there are data gaps in
the toxicity data set. Another consideration in the Quality of the Database is the risk of bias (i.e., how
representative is the study to ecologically relevant endpoints). Additionally, because of the importance
of the studies used for deriving hazard thresholds, the Quality of the Database consideration may have
greater weight than the other individual considerations. The high, medium, and low systematic review
overall quality determination ranks correspond to the evidence table ranks of robust (+ + +), moderate (+
+), or slight (+), respectively. The evidence considerations are weighted based on professional judgment
to obtain the overall confidence for each hazard threshold. In other words, the weights of each evidence
property relative to the other properties are dependent on the specifics of the weight of the scientific
evidence and uncertainties that are described in the narrative and may or may not be equal. Therefore,
the overall score is not necessarily a mean or defaulted to the lowest score. The confidence levels and
uncertainty type examples are described below.

A.l Confidence Levels	

•	Robust (+ + +) confidence suggests thorough understanding of the scientific evidence and
uncertainties. The supporting weight of the scientific evidence outweighs the uncertainties to the
point where it is unlikely that the uncertainties could have a significant effect on the exposure or
hazard estimate.

•	Moderate (+ +) confidence suggests some understanding of the scientific evidence and
uncertainties. The supporting scientific evidence weighed against the uncertainties is reasonably
adequate to characterize exposure or hazard estimates.

•	Slight (+) confidence is assigned when the weight of the scientific evidence may not be adequate
to characterize the scenario, and when the assessor is making the best scientific assessment
possible in the absence of complete information. There are additional uncertainties that may need
to be considered.

A.2 Types of Uncertainties	

The following uncertainties may be relevant to one or more of the weight of scientific evidence

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1490

1491

1492

1493

1494

1495

1496

1497

1498

1499

1500

1501

1502

1503

1504

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considerations listed above and will be integrated into that property's rank in the evidence table:

•	Scenario Uncertainty: Uncertainty regarding missing or incomplete information needed to fully
define the exposure and dose.

o The sources of scenario uncertainty include descriptive errors, aggregation errors, errors
in professional judgment, and incomplete analysis.

•	Parameter Uncertainty: Uncertainty regarding some parameter.

o Sources of parameter uncertainty include measurement errors, sampling errors,
variability, and use of generic or surrogate data.

•	Model Uncertainty: Uncertainty regarding gaps in scientific theory required to make predictions
on the basis of causal inferences.

o Modeling assumptions may be simplified representations of reality.

Table 2-12 summarizes the weight of the scientific evidence and uncertainties, while increasing
transparency on how EPA arrived at the overall confidence level for each exposure hazard threshold.
Symbols are used to provide a visual overview of the confidence in the body of evidence, while de-
emphasizing an individual ranking that may give the impression that ranks are cumulative (e.g., ranks of
different categories may have different weights).

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1506	TableApx A-l. Considerations that Inform Evaluations of the Strength of the Evidence within an Evidence Stream {i.e., Apical

1507	Endpoints, Mechanistic, or Field Studies)		

Consideration

Increased Evidence Strength (of the Apical
Endpoints, Mechanistic, or Field Studies
Evidence)

Decreased Evidence Strength (of the Apical Endpoints, Mechanistic, or
Field Studies Evidence)

The evidence considerations and criteria laid out here guide the application of strength-of-evidence judgments for an outcome or environmental hazard effect
within a given evidence stream. Evidence integration or synthesis results that do not warrant an increase or decrease in evidence strength for a given
consideration are considered "neutral" and are not described in this table (and, in general, are captured in the assessment-specific evidence profile tables).

Quality of the database'1
(risk of bias)

•	A large evidence base of high- or medium-quality
studies increases strength.

•	Strength increases if relevant species are
represented in a database.

•	An evidence base of mostly /ow-quality studies decreases strength.

•	Strength also decreases if the database has data gaps for relevant species,
i.e., a trophic level that is not represented.

•	Decisions to increase strength for other considerations in this table should
generally not be made if there are serious concerns for risk of bias; in other
words, all the other considerations in this table are dependent upon the
quality of the database.

Consistency

Similarity of findings for a given outcome (e.g., of a
similar magnitude, direction) across independent
studies or experiments increases strength,
particularly when consistency is observed across
species, life stage, sex, wildlife populations, and
across or within aquatic and terrestrial exposure
pathways.

•	Unexplained inconsistency (i.e., conflicting evidence; see U.S. EPA
(2005) decreases strength.)

•	Strength should not be decreased if discrepant findings can be reasonably
explained by study confidence conclusions; variation in population or
species, sex, or life stage; frequency of exposure (e.g., intermittent or
continuous); exposure levels (low or high); or exposure duration.

Strength (effect magnitude)
and precision

•	Evidence of a large magnitude effect (considered
either within or across studies) can increase strength.

•	Effects of a concerning rarity or severity can also
increase strength, even if they are of a small
magnitude.

•	Precise results from individual studies or across the
set of studies increases strength, noting that
biological significance is prioritized over statistical
significance.

•	Use of probabilistic model (e.g., Web-ICE, SSD)
may increase strength.

Strength may be decreased if effect sizes that are small in magnitude are
concluded not to be biologically significant, or if there are only a few
studies with imprecise results.

Biological gradient/dose-
response

•	Evidence of dose-response increases strength.

•	Dose-response may be demonstrated across studies
or within studies and it can be dose- or duration-
dependent.

• A lack of dose-response when expected based on biological
understanding and having a wide range of doses/exposures evaluated in the
evidence base can decrease strength.

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Consideration

Increased Evidence Strength (of the Apical
Endpoints, Mechanistic, or Field Studies
Evidence)

Decreased Evidence Strength (of the Apical Endpoints, Mechanistic, or
Field Studies Evidence)



•	Dose response may not be a monotonic dose-
response (monotonicity should not necessarily be
expected, e.g., different outcomes may be expected
at low vs. high doses due to activation of different
mechanistic pathways or induction of systemic
toxicity at very high doses).

•	Decreases in a response after cessation of exposure
(e.g., return to baseline fecundity) also may increase
strength by increasing certainty in a relationship
between exposure and outcome (this particularly
applicable to field studies).

•	In experimental studies, strength may be decreased when effects resolve
under certain experimental conditions (e.g., rapid reversibility after
removal of exposure).

•	However, many reversible effects are of high concern. Deciding between
these situations is informed by factors such as the toxicokinetics of the
chemical and the conditions of exposure, see (U.S. EPA. 1998). cndooint
severity, judgments regarding the potential for delayed or secondary
effects, as well as the exposure context focus of the assessment (e.g.,
addressing intermittent or short-term exposures).

•	In rare cases, and typically only in toxicology studies, the magnitude of
effects at a given exposure level might decrease with longer exposures
(e.g., due to tolerance or acclimation).

•	Like the discussion of reversibility above, a decision about whether this
decreases evidence strength depends on the exposure context focus of the
assessment and other factors.

•	If the data are not adequate to evaluate a dose-response pattern, then
strength is neither increased nor decreased.

Biological relevance

Effects observed in different populations or
representative species suggesting that the effect is
likely relevant to the population or representative
species of interest (e.g., correspondence among the
taxa, life stages, and processes measured or observed
and the assessment endpoint).

An effect observed only in a specific population or species without a clear
analogy to the population or representative species of interest decreases
strength.

Physical/chemical relevance

Correspondence between the substance tested and
the substance constituting the stressor of concern.

The substance tested is an analog of the chemical of interest or a mixture of
chemicals which include other chemicals besides the chemical of interest.

Environmental relevance

Correspondence between test conditions and
conditions in the region of concern.

The test is conducted using conditions that would not occur in the
environment.

" Database refers to the entire data set of studies integrated in the environmental hazard assessment and used to inform the strength of the evidence. In this context,
database does not refer to a computer database that stores aggregations of data records such as the ECOTOX Knowledgebase.

1508

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1509

1510

1511

1512

1513

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

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Appendix B SPECIES SENSITIVITY DISTRIBUTION FOR ACUTE
	AQUATIC HAZARD	

The SSD Toolbox is a resource that can fit SSDs to environmental hazard data (Etterson. 2020). It runs
on Matlab 2018b (9.5) for Windows 64 bit. For this draft DBP risk evaluation, EPA created one SSD
with the SSD Toolbox Version 1.1 to evaluate acute aquatic vertebrate and invertebrate toxicity. The use
of this probabilistic approach increases confidence in the hazard threshold identification as it is a more
data-driven way of accounting for uncertainty. For the acute SSD, acute exposure hazard data for
aquatic vertebrates and invertebrates were curated to prioritize study quality and to assure comparability
between toxicity values. For example, the empirical data set included only LCsos for high and medium
quality acute duration assays that measured mortality for aquatic vertebrates and invertebrates.
TableApx B-l shows the empirical data that were used in the SSD. To further improve the fit and
representativeness of the SSD, Web-ICE acute toxicity predictions for 53 additional species were added
(TableApx B-2).

With this data set, the SSD Toolbox was used to apply a variety of algorithms to fit and visualize SSDs
with different distributions. An HCos is calculated for each (Table Apx B-2)

The SSD Toolbox's output contained several methods for choosing an appropriate distribution and
fitting method, including goodness-of-fit, standard error, and sample-size corrected Akaike Information
Criterion (AICc, (Burnham and Anderson. 2002)). Most P values for goodness-of-fit were above 0.05,
showing no evidence for lack of fit. The distribution and model with the lowest AICc value, and
therefore the best fit for the data was the Gumbel Model (Figure Apx B-l). Because numerical methods
may lack statistical power for small sample sizes, a visual inspection of the data were also used to assess
goodness-of-fit. For the Q-Q plot, the horizontal axis gives the empirical quantiles while the vertical axis
gives the predicted quantiles (from the fitted distribution). The Q-Q plot demonstrates a good model fit
with the data points in close proximity to the line across the data distribution. Q-Q plots were visually
used to assess the goodness-of-fit for the distributions (Figure Apx B-2) with the Gumbel distribution
demonstrating the best fit near the low end of the distribution, which is the region from which the HC05
is derived. The results for this model (Figure Apx B-3) predicted 5 percent of the species (HC05) to
have their LC50s exceeded at 415 |ig/L (348 to 517 |ig/L 95% CI). The HCso was estimated at 1,159
|ig/L (951 to 1,444 |ig/L 95% CI) and the HC95 was estimated at 7,213 |ig/L (4,376 to 11,443 |ig/L 95%
CI).

Table Apx B-l. Species Sensitivity Distribution (SSD) Model Input for Acute Exposure Toxicity
in Aquatic Vertebrates and Invertebrates - Empirical Data		

Species

Description

Acute Toxicity Value LC50
(Hg/L)

Citation(s)

Americctmysis bahia

Aquatic invertebrate

612

(Adams et al.. 1995; EG&G
Bionomics. 1984b)

Danio rerio

Aquatic invertebrate

630

(Chen et al.. 2014)

Lepomis macrochirns

Aquatic vertebrate

788

(Adams et al.. 1995; EG&G
Bionomics. 1983b; Buccafusco
et al.. 1981)

Pimephcdes promelas

Aquatic vertebrate

1,178

(Smithers Viscient. 2018;

Adams et al.. 1995; Defoe et al..
1990; McCarthy and Whitmore.
1985; EG&G Bionomics. 1984a)

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Species

Description

Acute Toxicity Value LC50
(M.g/L)

Citation(s)

Oncorhvnchus mykiss

Aquatic vertebrate

1,497

(Adams et al.. 1995;
EnviroSvstem. 1991; EG&G
Bionomics. 1983a)

Nitocra spinipes

Benthic invertebrate

1,700

(Linden et al.. 1979)

Daphnia magna

Aquatic invertebrate

3,443

(Wei et al.. 2018; Adams et al..
1995; McCarthy and Whitmore.
1985)

Chironomus plumosus

Benthic invertebrate

4,648

(Streufort. 1978)

Paratany tarsus
parthenogeneticus

Benthic invertebrate

5,800

(EG&G Bionomics. 1984c)

1544

1545

1546	TableApx B-2. SSD Model Predictions" for Acute Exposure Toxicity to Aquatic Vertebrates

(Fish)

Distribution6

HC05 (jig/L)

P value

Normal

381

0.0839

Logistic

348

0.0100

Triangular

364

0.4386

Gumbel

415

0.0559

Weibull

239

0.0280

Burr

400

0.0150

11 The SSD was generated using SSD Toolbox vl.l.

h The model with the lowest AICc value, and therefore the best model fit, is bolded in this table.

1548

1549

1550	Table Apx B-3. Species Sensitivity Distribution (SSD) Model Input for Acute Exposure Toxicity

1551	in Aquatic Vertebrates and Invertebrates - Web-ICE Data		

Species

Description

Acute Toxicity Value LC50
(jig/L)

Gammarus pseudolimnaeus

Benthic invertebrate

228

Menidia peninsulae

Aquatic vertebrate

327

Lctgodon rhomboides

Aquatic vertebrate

451

Catostomus commersonii

Aquatic vertebrate

501

Menidia menidia

Aquatic vertebrate

502

Caecidotea brevicauda

Benthic invertebrate

532

Percaflavescens

Aquatic vertebrate

535

Allorchestes compressa

Benthic invertebrate

545

Cyprinodon bovinus

Aquatic vertebrate

546

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Species

Description

Acute Toxicity Value LC50
(^g/L)

Jordanellct floridae

Aquatic vertebrate

547

Sander vitreus

Aquatic vertebrate

549

Crcissostreci virginica

Benthic invertebrate

595

Ptychocheilus lucius

Aquatic vertebrate

647

Oncorhvnchiis kisutch

Aquatic vertebrate

673

Oncorhvnchiis clarkii

Aquatic vertebrate

674

Salvelinus namciycush

Aquatic vertebrate

782

Salmo scdar

Aquatic vertebrate

796

Lumbriculus variegatus

Benthic invertebrate

818

Salvelinus fontinalis

Aquatic vertebrate

853

Oreochromis mossambicus

Aquatic vertebrate

872

Micropterus salmoides

Aquatic vertebrate

908

Oncorhvnchiis tshawytscha

Aquatic vertebrate

920

Simocephalus vetulus

Aquatic invertebrate

930

Amblema plicata

Benthic invertebrate

1,039

Cyprinus carpio

Aquatic vertebrate

1,342

Acipenser brevirostrum

Aquatic vertebrate

1,342

Cyprinodon variegatus

Aquatic vertebrate

1,463

Xyrauchen texanus

Aquatic vertebrate

1,505

Oncorhvnchiis gilae

Aquatic vertebrate

1,506

Lasmigona complanata

Benthic invertebrate

1,521

Salmo trutta

Aquatic vertebrate

1,528

Poecilia reticulata

Aquatic vertebrate

1,541

Menidia beryllina

Aquatic vertebrate

1,573

Ictalurus punctatus

Aquatic vertebrate

1,581

Megalonaias nervosa

Benthic invertebrate

1,751

Lepomis cyanellus

Aquatic vertebrate

1,823

Lithobates catesbeianus

Amphibian

1,938

Oryzias latipes

Aquatic vertebrate

2,097

Oncorhvnchiis nerka

Aquatic vertebrate

2,141

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Species

Description

Acute Toxicity Value LC50
(^g/L)

Utterbackict imbecillis

Benthic invertebrate

2,244

Cctrctssius auratus

Aquatic vertebrate

2,275

Ceriodctphnict dubia

Aquatic invertebrate

2,372

Thamnocephcdus platyurus

Aquatic invertebrate

2,855

Margctritifera fcdcata

Benthic invertebrate

2,858

Daphnia pulex

Aquatic invertebrate

2,892

Physa gyrina

Benthic invertebrate

3,052

Brcmchinecta lynchi

Aquatic invertebrate

3,142

Lampsilis siliquoidea

Benthic invertebrate

3,155

Notropis mekistocholas

Aquatic vertebrate

3,447

Gammctrus fasciatus

Benthic invertebrate

3,539

Tigriopus japonicus

Benthic invertebrate

3,642

Lvmnaea stagnalis

Benthic invertebrate

3,738

Paratanytarsus dissimilis

Benthic invertebrate

5,419

1552

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-Aj ModelSelection

Percentile of interest:
Model-averaqed HCp;

Model-averaqed SE of HCd:
CV of HCp:

AICc Table

5

367.6771
46.3176
0.12597

X



Distribution

AICc

delta AICc

Wt

HCp

SE HCp



1

triangular

1.0313e+03

0

0.6230

364.0017

30.3446

2

normal

1.0328e+03

1.4955

0.2950

380.7975

55.3481

3

logistic

1.0372e+03

5.8312

0.0338

348.2229

58.4528

4

burr

1.0385e+03

7.1684

0.0173

400.4899

64.1671

5

weibull

1.0386e+03

7.2289

0.0168

238.9533

58.6521

6

gumbel

1.0389e+03

7.5662

0.0142

414.8611

40.8076

1553

1554	FigureApx B-l. AICc for the Six Distribution Options in the SSD Toolbox for Acute DBP

1555	Toxicity to Aquatic Vertebrates and Invertebrates (Etterson. 2020)

1556

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0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8 0.9
Predicted Quantiles

0.2 0.3 0.4 0.5 0.6 0.7
Predicted Quantiles

0.1 0.2 0.3 0.4 0.5 0.6 0.7
Predicted Quantiles

0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8
Predicted Quantiles

1557	FigureApx B-2. Q-Q Plots of Acute DBP Toxicity to Aquatic Vertebrates and Invertebrates with

1558	the A) Gumbel, B) Weibull, C) Burr, and D) Logistic Distributions (Etterson. 2020)

1559

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1560

1561

1562

0.9

0.8

0.7

2?

I06

1
e

0.



ro

D

I 0.4

0.3

0.2

0.1

gumbel-ML
~ HC05

•••• 95% CL HC05

Notmpis mekisteyholas •
Bmochiipa.i	X	S

' si s '

2.5

3.5

Log 10 LC50 (pg/L)



Benthic
Invertebrate



Aquatic
Vertebrate



Aquatic
Invertebrate



Amphibian

4.5

Figure Apx B-3. Species Sensitivity Distribution (SSD) for Acute DBP Toxicity to Aquatic Vertebrates and Invertebrates (Etterson.

2020)

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1563	Appendix C ENVIRONMENTAL HAZARD STUDIES	

1564	This appendix summarizes the aquatic and terrestrial endpoints and studies not included in the DBP

1565	quantitative risk evaluation, due to hazard values above the limit of solubility, lack of observed toxic

1566	effects, or inconsistency in the reported dose-response relationship.

1567

1568	Table Apx C-l. Acute Aquatic Vertebrate Toxicity of DBP 	

Test Organism
(Species)

Hazard Values

Duration

Endpoint

Citation
(Study Quality)

African clawed
frog (Xenopus
laevis)

14.1/21.0 mg/L

96-hour
NOEC/LOEC

Mortality

(Xu and Gve. 2018) (High)

12.88 mg/L

96-hour LC50

Mortality

(Gardner et al.. 2016) (Medium)

11.7/14.7 mg/L

96-hour
NOEC/LOEC

Sheepshead
Minnow

0Cyprinodon
variegatus)

>0.6 mg/L

96-hour NOEC

Mortality

(Sprinsborn Bionomics. 1984a)
(High)

Nile tilapia

(iOreochromis
niloticus)

11.8 mg/L

96-hour LC50

Mortality

(Khalil et al.. 2016) (Medium)

>10 mg/L

96-hour NOEC

Mortality
Growth

(Erkmen et al.. 2017) (Hieh)

Ide (Leuciscus
idus)

>10 mg/L

96-hour NOEC

Mortality

(BASF Aktieneesellschaft. 1989)
(Medium)

1569

1570

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Table Apx C-2. C

ironic Aquatic Vertebrate Toxicity of DBP

Test Organism
(Species)

Hazard Values

Duration

Endpoint

Citation
(Study Quality)

Zebrafish (Danio
rerio)

>0.1 mg/L

5-week
NOEC

Mortality

(Ortiz-Zarraaoitia et al.. 2006)
(Medium)

>0.5 mg/L

95-day
NOEC

Mortality

Growth

Reproduction

(Chen et al.. 2015) (High)

Three-spined
stickleback

(Gasterosteus
aculeatus)

>0.0352 mg/L

22-day
NOEC

Growth

(Aoki et al.. 2011) (Medium)

Fathead minnow

(Pimephales
promelas)

>0.062 mg/L

21-day
NOEC

Growth

Mortality

Reproduction

(Smithers Viscient. 2018) (Medium)

Crimson-spotted
rainbowfish

(Melanotaenia
fliiviatilis)

>0.457 mg/L

7-day
NOEC

Growth
Mortality

(Bhatia et al.. 2013) (Hiah)

>113 mg/L

7-day
NOEC

Growth
Mortality

(Bhatia et al.. 2014) (Hiah)

>0.05 mg/L
(Nominal)

90-day
NOEC

Mortality

(Bhatia et al.. 2015) (Hiah)

>0.005 mg/L
(Nominal)

Growth

Japanese medaka

0Oryzias latipes)

>12 mg/kg
bw/d

540-day
NOEC

Growth
Reproduction

(Patvna. 1999) (Hiah)

1572

1573

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Table Apx C-3. Acute Aquatic Inveri

tebrate Toxicity of DBP

Test Organism
(Species)

Hazard Values

Duration

Endpoint

Citation
(Study Quality)

Opossum shrimp

(Americamysis
bcthia)

>1.3 mg/L

24-hour LC50

Mortality

(EG&G Bionomics.
1984c)(Hiah)

1575

1576

1577	Table Apx C-4. Chronic Aquatic Invertebrate Toxicity of DBP

Test Organism
(Species)

Hazard Values

Duration

Endpoint

Citation
(Study Quality)

Water flea (Daphnict
magna)

>2.08 mg/L

16-day NOAEC

Reproduction

(McCarthy and

Whitmore.

1985)(Medium)

Midge (Chironomiis
plamoms)

>0.695 mg/L

40-day NOAEC

Growth

(Streufort.
1978)(Medium)

Daggerblade grass shrimp
(Palaemonetes pugio)

>21.5 mg/L

38-day NOAEC

Development/
Growth

(Lauahlin Rb et al..
1978)(Medium)

1578

1579

Table Apx C-5. C

ironic Benthic Invertebrate Toxicity of DBP

Test Organism

Hazard Values

Duration

Endpoint

Citation
(Study Quality)

Scud (Hyalella
azteca) high TOC

>71,900 mg/kg dw

10-day LC50

Mortality

(Call et al..
2001a)(High)

>13.2 mg/L

10-day NOAEC

Scud (Hyalella
azteca) medium
TOC

>29,500 mg/kg dw

10-day LC50

Mortality

(Call et al..
2001a)(Hiah)

82.4 mg/L (Probit)

10-day LC50

Mortality

(Lake Superior
Research Institute.
1997)(Hiah)

Scud (Hvalella
azteca) low TOC

>62.9 mg/L

10-day NOAEC

Mortality

(Call et al..
2001a)(High)

>17,400 mg/kg dw

10-day LC50

Midge

{Chironomiis
tentans) medium
TOC

12.2 mg/L (Linear
Interpolation)

10-day LC50

Mortality

(Lake Superior
Research Institute.
1997)(Hiah)

3.85/16 mg/L

10-day NOAEC/
LOAEC

(Call et al..
2001a)(Hiah)

Midge

{Chironomiis
tentans) low TOC

>74.2 mg/L

10-day NOAEC/
LOAEC

Development/ Growth

(Call et al..
2001a)(High)

>17,000 mg/kg dry
sediment

10-day NOAEC

1581

1582

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Table Apx C-6. Aquatic Plants and Algae Toxicity of

DBP

Test Organism
(Species)

Hazard Values

Duration

Endpoint

Citation
(Study Quality)

Green algae

(Selenastrum
cctpri cornu turn)

2.78/27.8 mg/L

7-day

NOEC/LOEC

Population (Biomass)

(Melin and Eaneus.
1983) (Medium)

Green algae

0Scenedesmus
acutus var. acutus)

15.3 mg/L

96-hour EC50

Population
(Abundance)

(Guetal.. 2017)
(High)

Green algae

0Scenedesmus
acutus vctr. acutus)

30.2 mg/L

96-hour EC50

Population
(Abundance)

(Kuans et al.. 2003)
(Medium)

39.8 mg/L

Population (Population
growth rate)

44.7 mg/L

Population
(Chlorophyll a
concentration)

Diatom

(Skeletonema
costatum)

200/500 mg/L

4-day

NOEC/LOEC

Population (Population
growth rate)

(Medlin. 1980)
(Medium)

1584

1585

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1586 Table Apx C-7. Terrestrial Vertebrate Toxicity of DBP

Test Organism
(Species)

Hazard Values

Duration

Endpoint

Citation



1250/2,500 ppm

GD-0 PND 28



(NTP. 1995)



100/200 mg/kg-
bw/day

GD 1-14



(Giribabu et al.. 2014)



120/600 mg/kg-
bw/day

GD 0-20



(Nikonorow et al.. 1973)



250/500 mg/kg-
bw/day

PND 21-25



(Wolf et al.. 1999)



250/500 mg/kg-
bw/day

PND 21-25



(Grav et al.. 1988)



500/1,000 mg/kg-
bw/day

PND 21-25







256/509 mg/kg-
bw/day

17 weeks



(NTP. 1995) (Wine et al..
1997)



385/794 mg/kg-
bw/day

17 weeks







5,000/10,000 ppm

63 days





Rat (Rattus

500/630 mg/kg-
bw/day

GD 7-15



(Ema et al.. 1993)

norvegicus)

630/750 mg/kg-
bw/day

GD 7-15







500/1,000 mg/kg-
bw/day

GD 15-17

Reproduction





1,000/1,500 mg/kg-
bw/day

GD 12-14







500/750 mg/kg-
bw/day

GD 3-PND 20



(Mvlchreest et al.. 1998)



579/879 mg/kg-
bw/day

4 weeks post-
weaning



(NTP. 1995)



7,500/10,000 mg/kg-
bw/day

GD 0-PND 28







10,000/20,000
mg/kg-bw/day

GD 0-20







10,000/20,000
mg/kg-bw/day

GD 0-PND 28







10,000/30,000
mg/kg-bw/day

PND 1-22







50/300 mg/kg-
bw/day

GD 7-9



(Xia et al.. 2011)

Mice

370/660 mg/kg-
bw/day

GD 0-18



(Shiota and Nishimura. 1982)
(Shiota et al.. 1980)



660/2,100 mg/kg-
bw/day

Gd 0-18





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Test Organism
(Species)

Hazard Values

Duration

Endpoint

Citation



3,000/10,000 mg/kg-
bw/day

15 weeks



(NTP. 1995)

5,000/7,500 mg/kg-
bw/day

GD 0-PND 28

7,500/10,000 ppm

GD 0-PND 28

10,000/20,000 ppm

GD 0-PND 28

525/1,750 mg/kg-
bw/day

18 weeks

(Ntp. 1984) (Lamb et al..
1987)

525/1,750 mg/kg-
bw/day

18 weeks

(Nto. 1984)

Chicken (Gctllus
gallus)

>100 mg/kg egg

NR (until
hatching) NOEL

Mortality
Growth

(Abdul-Ghani et al.. 2012)
(High)

Japanese quail

(Coturnix jctponicct)

>400 mg/kg bw/d

30-day NOEL

Growth

(Bello et al.. 2014) (Medium)

1587

1588

Table Apx C-8. Acute Soil Invertebrate Toxicity of DI

tP

Test Organism
(Species)

Hazard Values

Duration

Endpoint

Citation
(Study Quality)

European house
dust mite

(Dermatophagoides
pteronvs sinus)

>0.152 mg/cm3
(Fumigation)

24-hour NOEC

Mortality

(Kane et al.. 2006)
(Medium)

American house
dust mite

(Dermatophagoides
farina)

>0.152 mg/cm3
(Fumigation)

24-hour NOEC

Mortality

(Kane et al.. 2006)
(Medium)

Fruit fly

{Drosophila
melanogaster)

505,100 mg/L feed

72-hour LC50

Mortality

(Misra et al.. 2014)
(Medium)

278.3/2783 mg/L
feed

72-hour
NOEC/LOEC
(Adult exposure)

Reproduction

27.83/139.17 mg/L
in solution

24-hour
NOEC/LOEC

Nematode

(Caenorhabditis
elegans)

>139.17 mg/L

24-hour NOEC

Mortality

(Shin et al.. 2019)
(High)

27.83/139.17 mg/L
in solution

24-hour
NOEC/LOEC

Reproduction (Brood
size)

Table Apx C-9. C

ironic Soil Inverte

irate Toxicity of

DBP

Test Organism
(Species)

Hazard Values

Duration

Endpoint

Citation
(Study Quality)

Fruit fly

(Drosophila
melanogaster)

>0.418 mg/L feed

NR (egg until 5 to
6 days post hatch)

Mortality

(Williams et al.. 2016)
(Medium)

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1592

Table Apx C-10.r

"errestrial Plant Toxicity of DBP

Test Organism

Hazard Values

Duration

Endpoint

Citation
(Study Quality)

Tobacco

(Nicotiana
tabacum)

>2783 mg/L

7-day NOEC

Growth

(Dene et al.. 2017)
(High)

139.17/278.34
mg/L

3-day

NOEC/LOEC

Reproduction
(Germination)

Norway spruce
(Piceci abies)

>0.010 mg/m3
(Fumigation)

76-day NOEC

Growth

(Dueck et al.. 2003)
(High)

Perennial ryegrass
(Lolium perenne)

>500 mg/kg soil

72-hour NOEC

Growth

(Ma et al.. 2015)
(High)

Rapeseed (Brassica
napus)

<2.4 jig/cnr leaf

15-day LOEL

Physiology (Injury -
Chlorosis)

(Lokke and Rasmussen.
1983) (Medium)

Common yarrow
(Achillea
millefolium)

>2.9 jig/cnr leaf

15-day NOEL

Physiology (Injury -
Chlorosis)

White mustard

(Sinapis alba)

<3.5 (ig/cm2 leaf

15-day LOEL

Physiology (Injury -
Chlorosis)

Rice (Oryza
sativa)

>100 mg/L

5-day NOEC

Growth

(Isoaai et al.. 1972)
(Medium)

1594

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1595	Appendix D SUPPLEMENTAL SUBMITTED DATA TO BE

1596		CONSIDERED FOR FINAL RISK EVALUATION

1597	On July 10, 2024, EPA received supplemental information from DBP Consortium member companies

1598	related to ecotoxicity data supporting the risk evaluation for DBP. The Agency was unable to

1599	incorporate this data into the draft DBP ecological hazard assessment due to its late submission in the

1600	draft risk evaluation development process. However, EPA has included these data in the DBP risk

1601	evaluation docket (Docket ID: EPA-HO-OPPT-2Q18-0503) and will be considering the submission in

1602	the development of the final risk evaluation for DBP.

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