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|EPA Document # EPA-740-D-24-011

A	JlVIay 2024

iBKf	United States	Office of Chemical Safety and

" * Environmental Protection Agency	Pollution Prevention

Draft Environmental Hazard Assessment for Diisononyl

Phthalate (DINP)

Technical Support Document for the Draft Risk Evaluation

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

(Representative Structure)

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27	TABLE OF CONTENTS

28	SUMMARY	5

29	1 INTRODUCTION	6

30	2 APPROACH AND METHODOLOGY	7

31	3 AQUATIC SPECIES HAZARD	8

32	3.1 Aquatic Organism Hazard Conclusions	15

33	4 TERRESTRIAL SPECIES HAZARD	16

34	4.1 Terrestrial Organism Hazard Conclusions	19

35	5 WEIGHT OF SCIENTIFIC EVIDENCE CONCLUSIONS FOR ENVIRONMENTAL

36	HAZARD	21

37	6 ENVIRONMENTAL HAZARD THRESHOLDS	23

38	REFERNCES	27

39	Appendix A ENVIRONMENTAL HAZARD DETAILS	31

40	A,1 Evidence Integration	31

41	A.2 Weight of Scientific Evidence	31

42	A.3 Strengths, Limitations, Assumptions, and Key Sources of Uncertainty for Environmental

43	Hazard	35

44

45	LIST of tables

46	Table 3-1. Aquatic Vertebrate Environmental Hazard Studies for DINP	11

47	Table 3-2. Aquatic Invertebrate Environmental Hazard Studies for DINP	14

48	Table 3-3. Aquatic Plant Environmental Hazard Studies for DINP	15

49	Table 4-1. Terrestrial Mammal Hazard Studies of DINP Used for TRV Derivation	17

50	Table 5-1. DINP Evidence Table Summarizing the Overall Confidence Derived from Hazard

51	Thresholds	22

52	Table 6-1. Environmental Hazard Threshold for Aquatic and Terrestrial (TRV) Environmental

53	Toxicity	24

54

55	LIST of figures

56	Figure 6-1. Terrestrial Mammal TRV Derivation for DINP in Mammal Diets	25

57	Figure 6-2. TRV Flow Chart	26

58

59	LIST of appendix tables

60	Table Apx A-l. Considerations that Inform Evaluations of the Strength of the Evidence within an

61	Evidence Stream {i.e., Apical Endpoints, Mechanistic, or Field Studies)	33

62

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ABBREVIATIONS AND ACRONYMS

AF

Assessment factor

bw

Body weight

coc

Concentration(s) of concern

dw

Dry weight

EC50

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

GD

Gestation day

HC05

Hazard concentration that is protective of 95 percent of the species in the sensitivity



distribution

LC50

Lethal concentration at which 50 percent of test organisms die

LD50

Lethal dose at which 50 percent of test organisms die

LOAEL

Lowest-ob served-adverse-effect level

LOEC

Lowest-observed-effect concentration

NITE

National Institute of Technology and Evaluation

NOAEL

No-observed-adverse-effect level

NOEC

No-observed-effect concentration

OCSPP

Office of Chemical Safety and Pollution Prevention

OPPT

Office of Pollution Prevention and Toxics

PND

Postnatal day

QSAR

Quantitative structure-activity relationship (model)

SSD

Species sensitivity distribution

TRV

Toxicity reference value

TSCA

Toxic Substances Control Act

U.S.

United States

Web-ICE

Web-based Interspecies Correlation Estimation

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ACKNOWLEDGMENTS

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

Acknowledgements

The Assessment Team gratefully acknowledges the participation, input, and review comments from
OPPT and OCSPP senior managers and science advisors and assistance from EPA contractors SRC, Inc.
(Contract No. 68HERH19D0022).

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

Docket

Supporting information can be found in the public docket, Docket ID (EPA-HQ-OPPT-2024-0073).
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: Jennifer Brennan (Discipline Lead), John Allran (Management Lead), Collin Beachum
(Branch Chief), Randall Bernot, and Christopher Green

Contributors: Emily Griffen, Mark Myer, Andrew Sayer, Kelley Stanfield

Technical Support: Mark Gibson, Hillary Hollinger

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

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SUMMARY

EPA evaluated the reasonably available information for environmental hazard endpoints associated with
diisononyl phthalate (DINP) exposure. EPA reviewed 46 references and determined that 32 references
had high or medium data quality. These references included acute and chronic exposures via water, soil,
sediment, and food in both aquatic and terrestrial habitats.

Experimental aquatic hazard data were available from studies of the effects from acute exposures of
DINP on five fish species, one amphibian species, five aquatic invertebrate species, and two algal
species. Three fish taxa were represented in chronic exposure DINP feeding studies. Results from
standard laboratory tests suggest that DINP has low hazard potential in aquatic species. Few adverse
effects on survival, growth, development, or reproduction were observed in acute and chronic exposure
duration tests at concentrations up to and exceeding the DINP solubility and saturation limits.

In terrestrial habitats, a Toxicity reference value (TRV) of 139 mg/kg-bw/d was derived for the chronic
exposure effects of DINP on a generalized terrestrial mammal. One study of earthworm survival and
reproduction found no hazards at the maximum experimental soil concentration of 1,000 mg/kg dw
DINP. Also, no toxicity studies on avian or terrestrial plant species were identified.

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136	1 INTRODUCTION

137	Diisononyl phthalate (DINP) is an organic substance primarily used as a plasticizer in a wide variety of

138	consumer, commercial and industrial products (	2021b). Like most phthalates, DINP would be

139	expected to cause adverse effects on aquatic organisms through a non-specific, narcotic mode of toxic

140	action (Parkerton and Konkel. 2000); however, previous assessments have found few to no effects of

141	DINP on organism survival and fitness (EC/HC. 2015; ECJRC. 2003). EPA reviewed studies of the

142	potential toxicity of DINP to aquatic and terrestrial organisms and its potential environmental hazards.

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2 APPROACH AND METHODOLOGY

EPA calculates hazard thresholds to identify potential concerns to aquatic and terrestrial species. For
aquatic species, the hazard threshold is called a concentration of concern (COC), and for terrestrial
species, the hazard threshold is called a hazard value or TRV. These terms (COC, TRV, and hazard
value) describe how the values are derived and can encompass multiple taxa or ecologically relevant
groups of taxa, as the environmental risk characterization serves populations of organisms within a wide
diversity of environments. After weighing the scientific evidence, EPA selects the appropriate toxicity
value from the integrated data to use for hazard thresholds. See Section 5 for more details about how
EPA weighed the scientific evidence.

For terrestrial species, EPA estimates hazard by calculating a TRV, in the case of terrestrial mammals
and birds, or by assigning the hazard value as the hazard threshold in the case of terrestrial plants and
soil invertebrates. When possible, EPA prefers to derive the TRV by calculating the geometric mean of
the no-observed-adverse-effect-level (NOAELs) across sensitive endpoints (growth and reproduction)
rather than using a single endpoint. The TRV method is preferred because the geometric mean of
NOAELs across studies, species, and endpoints provides greater representation of environmental hazard
to terrestrial mammals and/or birds. However, when the criteria for using the geometric mean of the
NOAELs as the TRV are not met (according to methodology described in EPA's Guidance for
Developing Ecological Screening levels (Eco-SSLs) (U.S. EPA. 2007). the TRVs for terrestrial
mammals and birds are derived using a single endpoint.

During the scoping process, EPA reviewed the potential environmental hazards associated with DINP
and identified 35 references (see Figure 2-9) from Final Scope of the Risk Evaluation for Di-isononyl
Phthalate (DINP) CASRN28553-12-0 and 68515-48-0 (U.S. EPA. 2021c). EPA reviewed the
environmental hazard data in these and additional referenced studies using the data quality evaluation
metrics and criteria described in the 2021 Draft Systematic Review Protocol (	21a). Studies

were assigned an overall quality determination of high, medium, low, or uninformative. High or medium
data quality determinations were assigned to 19 aquatic organism references, several of which contained
hazard data from multiple organisms and endpoints. EPA also considered 12 animal toxicity references
that contained data used to determine a TRV, and 1 terrestrial earthworm toxicity reference. Thus, 32
references contained environmental hazard data with high or medium data quality determinations and
were included in this assessment.

EPA assigned high or medium quality determinations to 19 aquatic toxicity references, and one
terrestrial earthworm reference. Five references indicated hazard values from feeding or water-based
exposure within fishes (Camevali et ai. 2019; Forner-Piquer et ai. 2019; Forner-Piquer et ai. 2018b;
Forner-Piquer et al. 2018a; Patyna et ai. 2006). All other studies did not result in estimates of
population-level effects (e.g., mortality, development, growth) up to the highest concentration tested.
The maximum test concentrations reported in these aquatic studies exceeded the estimates of the water
solubility limit for DINP which is approximately 6.1 xl0~4 mg/L (	2024). No studies on

terrestrial wildlife vertebrate species (birds and mammals) were identified. In lieu of terrestrial wildlife
studies, 12 references with controlled laboratory studies that used mice and rats as human health model
organisms were used to calculate a TRV that is expressed as doses in units of mg/kg-bw/day. Although
the TRV for DINP was derived from laboratory mice and rat studies, because body weight is
normalized, EPA used it as a screening surrogate for effects on ecologically relevant wildlife species to
evaluate chronic dietary exposure to DINP. An additional 12 studies of dietary DINP exposures to
laboratory rodents with high or medium data quality evaluations were used to derive a TRV.

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3 AQUATIC SPECIES HAZARD

EPA assigned an overall quality level of high or medium to 19 references. These references contained
relevant aquatic toxicity data for sheepshead minnow (Cyprinodon variegatus), rainbow trout
(Oncorhynchus mykiss), zebrafish (Danio rerio), fathead minnow (Pimephales promelas), bluegill
sunfish (Lepomis macrochirus), Japanese medaka (Oryzias latipes), gilthead sea bream (Sparus aurata),
moorfrog (Rana arvalis), waterflea (Daphnia magna), amphipod (Hyalella azteca), midge
(Paratany tarsusparthenogenetica & Chironomus tentans), mysid shrimp (Americamysis bahia), green
algae (Selanastrum capricornutum), and marine dinoflagellate (Karinia brevis). EPA summarized
aquatic toxicity studies for quantitative assessment of aquatic vertebrates (Table 3-1), invertebrates
(Table 3-2), and algae (Table 3-3).

Aquatic Vertebrates

Fish: EPA identified references with data from acute exposures and chronic exposures of DINP on
different fishes. Acute exposure studies found no effects of DINP at any of the tested concentrations
(Table 3-1). Chronic studies found no effects of DINP water exposure on fish and found inconsistent
effects of dietary exposure to fish.

Seven of the eight acute studies on aquatic vertebrates consisted of 96-hour toxicity tests conducted on
juvenile and adult fish species and were all assigned overall quality determinations of high. These acute
exposure studies tested up to the limit of solubility, were conducted without the use of solvents, and
were not able to establish LC50 or LOEC values due to lack of mortality. Also, the maximum test
concentrations reported in these studies exceeded EPA's estimate of the water solubility limit for DINP
which is approximately 6.1 xl0~4 mg/L (	2024). One study with an assigned overall quality

determination of medium used 0.1 percent methanol as a solvent to enhance solubility and reported an
LC50 of greater than 500 mg/L (the highest tested nominal concentration) from 72-hour exposures with
newly fertilized zebrafish embryos (4-128 cell stage) (Chen et al. ).

Of the studies of chronic dietary DINP exposure, a chronic duration study with Japanese medaka
(Oryzias latipes) with a high data quality determination found statistically significant but inconsistent
effects of DINP-amended diets on survival in second-generation fish, but not first- or third-generation
fish (Patvna et al.. 2006). This two-generation feeding study fed one elevated dose of 1 mg/kg-bw/day
DINP-amended dried food to juvenile and adult fish. Lower survival of embryos occurred in one assay
of Fo embryos, but not during a second assay in the Fo generation or in multiple assays in the Fi and F2
generations. Thus, fish embryos exhibited an inconsistent effect of parental dietary exposure to 1 mg/kg-
bw/day DINP with most assays finding no effects across three generations. The study also found a
transient effect of 16 percent lower survival among Fi adult fish fed 1 mg/g-bw/day DINP over 140 days
compared to control fish. This effect on survival did not occur in the Fo generation despite identical
dietary exposure over 140 days or in the F2 generation despite 40 more days of dietary exposure. Thus,
dietary DINP induced a transient 16 percent reduction in survival only in the second generation of
continuous feeding exposure, but not in the first or third generations. The authors measured several other
endpoints and found no DINP effects on reproduction and development except for an increase in
testosterone metabolites in males and a delay in red blood cell pigmentation in fish fed DINP daily. The
DINP-amended food dose was analytically verified as 21.9 ± 2.8 |ig/g fed at a rate of 5 percent body
weight per day with brine shrimp fed as a supplement three times per week for the Fo generation and
five times per week for the Fi generation resulting in an average lipid-based feeding rate of 1 mg/kg-
bw/day DINP per fish. Patvna et al. (2006) conducted this study with five replicates per treatment and
included untreated and solvent (acetone) controls.

Three 21-day feeding studies on gilthead sea bream (S. aurata) with overall quality determinations of

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medium, found non-apical effects of DINP on fish (Camevali et al. 2019 and Forner-Piquer.2019.
5534689 and Fomer-Pique:	59; Forner-Piquer et al.. 2018a). A 21-day study of gilthead sea

bream fed with 1.5 mg/kg-bw DINP-amended food resulted in increased presence of lipids and
triglycerides and decreased glycogen and phospholipids in the liver (Forner-Piquer et al.. 2018a). This
study also found DINP exposure upregulated genes associated with disrupted metabolic activity. No
statistically significant differences in body mass were observed among treatments, however. Similarly,
Camevali et al. ( found that gilthead sea bream exhibited decreased muscle protein and lipid
content after being fed 1.5 mg/kg-bw DINP, which was the highest nominal concentration of DINP
administered. This study also found that dietary DINP exposure resulted in upregulated catd mRNA
levels and more enzymes that break down proteins. Finally, Forner-Piquer et al. ( found reduced
levels of endocannabinoids and endocannabinoid-like mediators along with higher fatty acid amide
hydrolase activity in gilthead sea bream fed 1.5 mg/kg bw DINP per day compared to no-DINP controls
and low-DINP treatments of 15 |ig/kg-bw DINP per day. The authors documented fewer motile sperm
cells due to DINP despite overall sperm production being unaffected. The production of 11-
ketotestosterone, which is an active androgen in fish, was greater than 50 percent lower in males fed
diets of 1.5 mg/kg-bw DINP per day compared to no-DINP control fish after 21 days. EPA has slight
confidence in the hazard values from all three of these studies for several reasons that they all share.

First, all effects were non-apical in that they were not directly related to fish survival, growth, or
reproduction. Second, each study used experimental designs and analyses that resulted in a mismatch
between experimental unit replication and the statistical and biological inferences that were made. For
example, treatment diets were given to fish in duplicate {i.e., five fish in each of two aquaria, or n = 2),
but results were presented from analyses using individual fish as replicates {i.e., n = 10). Thus,
inferences about the results could be inferred about the tissues of individual fish but not a population of
fish. Third, the studies did not analytically verify DINP concentrations in the food and relied on nominal
concentrations across 21 days. Finally, the relatively short duration (21-day) of feeding exposure to adult
fish may be inadequate for detecting apical effects that are most likely to translate to effects on fish
populations.

In a chronic 21 -day DINP adult zebrafish study with a data quality determination of high, Santangeli et
al. (2017) reported 30 percent reductions in eggs per female and gonadosomatic index at a water
concentration of 0.42 |ig/L DINP compared to controls. However, the effects disappeared at higher
nominal DINP concentrations of 4.2, 42, 420, and 4,200 |ig/L. The nominal concentrations of 0.42, 4.2,
42, 420, and 4,200 |ig/L were not analytically verified and exceeded water solubility. Also, the study
was conducted with treatment and control groups in duplicate with all fish in each treatment
concentration housed in a single net-divided aquarium, resulting in limited statistical power. Although
this study received an initial high data quality determination, EPA has low confidence in the reported
effects due to the lack of dose-response effects, analytical DINP verification, and experimental unit
replication.

An additional chronic study, with an overall quality determination of medium, exposed zebrafish to
multiple water concentrations of DINP (Fomer-Piquer et al.. 20181). ler-Piquer et al. (2018b) found
>30 percent reductions in zebrafish fertilization rates, greater than 20 percent reductions in
gonadosomatic index, and statistically significant changes in a number of lipid-signaling endpoints at the
lowest exposure concentration of 0.42 |ig/L DINP over 21 days. However, these effects were not
observed at higher nominal DINP concentrations of 4.2 and 42 |ig/L. These concentrations were not
analytically verified, exceeded water solubility, and exposure was only replicated twice resulting in
limited statistical power.

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288	EPA identified one study on an amphibian, the moorfrog (R. arvalis), and assigned an overall quality

289	determination of high (TVL. 2001). Moorfrog embryos were exposed to two sediment types (fine and

290	coarse) spiked with nominal DINP concentrations of 0 (negative control), 0 (acetone solvent control),

291	100, 300, and 1,000 mg DINP/kg-dw to investigate hatchability and embryo survival with observations

292	at 9, 12, 16, and 21 days of exposure. Hatching success, median hatching time, mortality, and

293	deformities were not statistically different among DINP, control and solvent control treatments. Tadpole

294	growth, recorded as wet weight, was assessed after 26 days of exposure with results indicating no

295	difference among DINP, control, and solvent control treatments.

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296 Table 3-1. Aquatic Vertebrate Environmental Hazard Studies for DINP

Duration

Test Organism (Species)

Endpoint

Hazard Values

Effect

Citation
(Data Evaluation Rating)

Acute

Sheepshead minnow (Cyprinodon
variegatus)

96-hour LC50

>0.52 mg/L fl

Mortality

(Adams etaL 1995) (High)

Rainbow trout

(Oncorhynchus mykiss)

96-hour LC50

>0.16 mg/L a

Mortality

(Adams etaL, 1995) (High)

Fathead minnow

(Pimephales promelas)

96-hour LC50
(static)

>0.10 mg/L fl

Mortality

(Adams etaL. 1995) (High)

Fathead minnow

(Pimephales promelas)

96-hour LC50

>0.14 mg/L fl

Mortality

(I nomics, 1983a) (Hinh)

Fathead minnow

(Pimephales promelas)

96-hour LC50
(flow-through)

>0.19 mg/L fl

Mortality

(Adams etaL, 1995) (High)

Bluegill sunfish

(Lepomis macrochirus)

96-hour LC50

>0.14 mg/L fl

Mortality

(Adams etaL. 1995) (High)

Bluegill sunfish

(Lepomis macrochirus)

96-hour LC50

>0.17 mg/L fl

Mortality

(I nomics, 1983c) (Hinh)

Zebrafish

(Danio rerio)

72-hour LC50

>500 mg/L b

Mortality

(Chen et al.. 2014) (Medium)

Chronic

Japanese Medaka

(Oryzias latipes)

2nd generation
140-day LOEC

1 mg/kg bw/day a c

Posthatch survival

(Patvna et al.. 2006) (Hiszh)

Japanese Medaka

(Oryzias latipes)

2nd generation
140-day LOEC

1 mg/kg bw/day a d

Survival/Reproduction
/Growth

(Patvna et al., 2006) (Hinh)

Zebrafish

(Danio rerio)

21-day LOEC

>0.0004 mg/L fl

Egg production,
oocyte biochemical
composition

(Santanseli et al.. 2017) (Hiah)

Zebrafish

(Danio rerio)

21-day LOEC

0.0004 mg/L fl

Egg production, lipid
signaling system

(Fomer-Piouer et al.. 2018b) (Medium)

Gilthead sea bream

(Sparus aurata)

21-day LOEC

1.5 mg/kg bw be

Muscle molecular
composition

(Carnevali et al., 2019) (Medium)

Gilthead sea bream

(Sparus aurata)

21-day LOEC

1.5 mg/kg bw hf

Lipid signaling system

(Fomer-Piciuer et al., 2018a) (Medium)

Gilthead sea bream

(Sparus aurata)

21-day LOEC

1.5 mg/kg bw bg

Increase in

Gonadosomatic Index

(Fomer-Piouer et al.. 2019) (Medium)

Moorfrog
(Rana arvalis)

21-day LOEC

>1,000 mg/kg/ dry
weighta

Hatching success,
mortality, growth

(I LD (High)

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Duration

Test Organism (Species)

Endpoint

Hazard Values

Effect

Citation
(Data Evaluation Rating)

a indicates measured concentration.
b indicates nominal concentration.

c authors state that posthatch survival was lower in one assay of the Fo generation, but not in a second assay of the Fo generation and no posthatch survival
effects were observed in the Fi or F2 embryos. DINP diets delayed the pigmentation of red blood cells and increased testosterone hydroxylase activity.
d dietary DINP induced a transient 16% reduction in survival after 140 days of exposure to parental and then another 140 days of individual fish exposure. DINP
effects were not observed in Fo or F2 generations. The authors concluded that DINP in diet did not affect adult fish survival overall.

'' dietary DINP exposure resulted in decreased lipid and protein content in muscle tissue due to upregulated catd mRNA levels and more enzymes that break
down proteins.

^dietary DINP exposure resulted in more lipids and triglycerides in fish livers along with upregulated genes associated with disrupted metabolic activity.
g dietary DINP exposure resulted in lipid metabolism disruption, reduced androgen production, increase 17(3-estradiol production leading to a higher
gonadosomatic index, and fewer motile sperm cells in male fish.

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Aquatic Invertebrates

EPA identified 11 studies with aquatic invertebrate hazard data from DINP exposure: seven studies
representing acute DINP exposures and four studies representing chronic DINP exposures, all with
overall quality determinations of high (Table 3-2).

Acute studies conducted on aquatic invertebrates included results of three different 48-hour exposures of
DINP to D. magna, one 48-hour exposure and one 96-hour exposure of DINP to P. parthenogenetica,
and two different 96-hour exposures of DINP to A. bahia, (Brown etal	Adams et al. \ , i

& G Bionomics. 1984b; Sprinebom Bionomics. 1984a). Adverse acute effects were not observed at
DINP concentrations up to and beyond the limit of solubility (6.1 xl0~4, (	2024). For example,

Sprinebom Bionomics (1984a) studied the acute toxicity of 14 phthalate esters to I), magna under static
conditions. Based on the 0 and 48 hour mean measured concentrations, the DINP EC50 exceeded 0.086
mg/L. No visible film or apparent insoluble test material was observed in the test solution; however,
since there were entrapped daphnids on the test vessel's surface, the authors suggested that the test
material aggregated on the surface during tests. No mortality was reported even though more than 50
percent of the daphnids were caught on the surface of the test solution.

Chronic studies with aquatic invertebrates included two aquatic DINP exposures on I), magna (Brown et
al.. 1998; Rhodes et al.. 1995) and two 10-day studies with sediment DINP exposures conducted on H.
azteca and C. tentans (Call et al.. 2001). D. magna exposed to nominal concentrations of DINP for 21
days resulted in a reduced survival and reproduction LOEC of 0.089 mg/L and a NOEC of 0.034 mg/L,
for a chronic value (ChV) of 0.06 mg/L (Rhodes et al.. 1995). Although authors reported that no visible
film was observed, physical entrapment of D. magna with the water surface boundary was observed
within test vessels at the LOEC. The authors concluded that this physical entrapment contributed to their
observed animal mortality and reproduction effects. Thus, because of this uncertainly between physical
and chemical toxicity, EPA is not considering these as concentrations of concern. A similar 21-day
exposure study conducted by Brown et al. (1998) increased the solubility of DINP in solution with the
addition of a dispersant, castor oil 40 ethoxylate (10 mg/L) and found no differences in reproduction or
survival from a 1 mg/L exposure to DINP when compared to the control or dispersant control. Longer
duration studies with C. tentans and H. azteca were conducted with subchronic 10-day exposures of
sediment spiked with nominal concentrations of DINP (Call et al.. 2001). Adverse effects were not
observed for the highest DINP spiked sediment concentrations used in these studies at 2,900 mg/kg dw
DINP sediment and 2,680 mg/kg dw DINP sediment for H. azteca and C. tentans, respectively.

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Table 3-2. Aquatic Invertebrate Environmental Hazard Studies for DINP

Duration

Test Organism (Species)

Endpoint

Hazard Values"

Effect

Citation
(Data Evaluation Rating)

Acute

Waterflea
(Daphnia magna)

48-hour EC50

>0.06 mg/L

Immobilization

(Adams et al. 1995) (High)

Waterflea
(Daphnia magna)

48-hour EC50

>1.00 mg/L

Immobilization

(Brown et al„ 1998) (High)

Waterflea
(Daphnia magna)

48-hour EC50

>0.09 mg/L

Immobilization

(Soringborn Bionomics.
1984a) (Medium)

Midge

{Paratany tarsus
varthenogenetica)

48-hour LC50

>0.12 mg/L

Mortality

( 3nomics„
1984c) (High)

Midge

{Paratany tarsus
varthenogenetica)

96-hour LC50

>0.08 mg/L

Mortality

(Adams et al. 1995) (High)

Mysid shrimp
(Americamysis bahia)

96-hour LC50

>0.39 mg/L

Mortality

(Adams et al. 1995) (High)

Mysid shrimp
(.Americamysis bahia)

96-hour LC50

>0.77 mg/L

Mortality

( 3nomics„
1984b)(High)

Chronic

Waterflea
(Daphnia magna)

21-day LOEC

0.034 mg/L NOEC
0.089 mg/L LOEC for
all effects b

Mortality, Offspring per female

(Rhodes et al. 1995) (High)

Waterflea
(Daphnia magna)

21-day NOEC

>1.0 mg/L

Mortality, Reproduction, Growth

(Brown et al.. 1998) (High)

Amphipod
(Hyalella azteca)

10-day NOEC

>0.44 mg/L porewater;
>2900 mg/kg dw
sediment

Mortality

(Call et al.. 2001) (High)

Midge

(Chironomus tentans)

10-day NOEC

>0.869 mg/L
porewater, >2680
mg/kg dw sediment

Mortality

(Call et al. 2001) (High)

3 all hazarc
'' the autho

values represent measured concentrations.

rs concluded that D. magna physical entrapment with surface tension contributed to animal mortality and reproduction effects.

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Aquatic Plants

EPA identified two studies with an overall quality determination of high and one study with an overall
quality determination of low for aquatic plants exposed to DINP (Table 3-3).

Both studies with overall quality determinations of high were conducted on green algae, Salanastrum
capricornutum. Sprinebom Bionomics (1984c) determined that an EC50 based on cell numbers at 5
days of DINP exposure was greater than2.8 mg/L, well over EPA's estimated water solubility of
6.1 xl0~4 (	124). Specifically, chlorophyll a concentration was not different from the control

treatment after 5 days of DINP exposure but cell numbers within the single DINP concentration tested
(2.8 mg/L) were 34 percent less than the control treatment. Adams et; 5) did not observe adverse
effects at the highest tested concentration of DINP (1.8 mg/L) from 96-hour exposures of DINP.
Concentrations of DINP were verified analytically with gas-liquid chromatography and gas
chromatography for Sprinebom «11< momi.^ ^ I • ?N4c) and Adams et al. (1995). respectively, with neither
study using a solvent within treatment and control groups. A 96-hour exposure study conducted on the
marine dinoflagellate, K. brevis, resulted in no significant effect of DINP on algal cell number compared
to the controls up to the highest reported nominal concentration of DINP at 50 mg/L (Liu et al.. 2016).

Table 3-3. Aquatic Plant Environmental Hazard Studies for DINP

Test Organism (Species)

Endpoint

Hazard
Values

Effect

Citation
(Data Evaluation Rating)

Green algae

(Selanastrum capricornutum)

96-hour
EC50

>2.80 mg/L a

Cell numbers,
chlorophyll a

(Serin
(High)

Green algae

{Selanastrum capricornutum)

96-hour
EC50

>1.80 mg/L a

Cell numbers

(Adams et al.. 1995) (High)

Marine dinoflagellate
(Karinia brevis)

96-hour
EC50

>50 mg/L h

Cell numbers

(Liu et al.. 2016) (Low)

3 indicates measured concentration.
'' indicates nominal concentration.

3.1 Aquatic Organism Hazard Conclusions

Overall, EPA has robust confidence in the evidence that DINP has low hazard potential in aquatic
species (Table 5-1). No consistent effects of DINP on aquatic organism survival or reproduction were
observed in studies of aquatic organisms across taxonomic groups, habitats, exposure type, and exposure
duration. Studies of DINP exposure via water to fish, amphibians, invertebrates, and algae reported no
effects up to and well above the solubility limit in the water column and in the sediment pore water.
Studies of dietary exposure of DINP to two fish species indicate no consistent population-level DINP
effects and inconsistent effects of DINP on mechanistic endpoints such as gene expression and protein
synthesis. Thus, EPA has moderate confidence in the studies that describe the potential effects of
chronic dietary DINP exposure to fish populations.

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4 TERRESTRIAL SPECIES HAZARD

EPA identified 12 terrestrial animal toxicity references with overall quality determinations of high or
medium that use rat (Rattus norvegicus) or mouse (Mus musculus) species to study reproductive,
growth, or survival endpoints. These studies were used to derive a TRV of DINP for a representative
small mammal. EPA also identified one invertebrate toxicity study on chronic exposure of DINP to
earthworms (Eiseniafetida) in soil.

Terrestrial Vertebrates

No terrestrial vertebrate studies were reasonably available to assess the potential effects or hazards from
DINP exposure in bird or mammalian wildlife species. Therefore, EPA considered ecologically relevant
definitive hazard data from studies conducted on laboratory mammals (e.g., rats, mice, etc.) that are
routinely used to inform human health hazard. These data were then used in accordance with EPA's
Guidance for Developing Ecological Soil Screening Levels (Eco-SSLs) (	2007) to formulate a

TRV to represent terrestrial mammals (see Table 4-1 and Table 6-1).

Mammals

Multiple studies of DINP administered in rat diets found reductions in rat offspring body weight over the
course of 2 to 19 weeks (LOAEL range from 288 to 1,500 mg/kg-bw/d) (Gray. 2023; Clewell et at..
20.| 3; I-Sobers et at.. 2011; Masutomi et at.. 2003; NTP-CERHR. 2003; Waterman et at.. 20001
Masutomi et at. (2003) found a decrease in body weight of male pups at prepubertal necropsy (PND 27)
in 306.7 and 1,165.5 mg/kg/day groups. Exposure duration was 18 days (assuming GD 15-22, PND 1-
10) and ceased on PND 10. Dams were fed control diet for the remainder of lactation, and pups were fed
the control diet after weaning. Treatment exposures were 0, 30.7, 306.7 and 1,164.5 mg/kg/day).

Hettwie et at. (1997) found mean maternal body weights were lower in rats gavaged 1,000 mg/kg DINP
at days 13, 15, and 17 post gestation day after being administered DINP from day 6 to day 15 post
gestation day. Waterman et at. (1999) found reductions in maternal rat body weight gain in 1,000
mg/kg/day treatments after being gavaged from GD 6 to 15. In a one-generation study, Exxon
Biomedical (1996a) found lower body weight in parental female rats in 741 mg/kg/day and 1,087
mg/kg/day groups during GD 0 to 21. Rats were fed DINP in diet through gestation and post-partum.

In similar two-generation studies, Exxon Biomedical (1996b) found lower body weight of Fi male pups
and Fi female pups at birth (PND 0) in the 0.4 and 0.8 percent dietary concentrations groups and lower
body weight as GD 21 of PI adult females. Sobers et al. (2 found lower male pup weight at PND
13 in a 900 mg/kg bw/day DINP-fed treatment. Exposure duration was 33 days (GD 7-22, PND 1-17).
Finally, Clewell et t	found lower male pup weight on PND 14 at 247 mg/kg/day DINP

treatments. Adult rats were fed DINP diets through gestation and lactation.

Growth: Across a range of study durations, DINP fed to adult rats resulted in lighter body weights
compared to control adult rats and mice (LOAEL range 152 to 1,5 13 mg/kg/d) (Clewell et al.. 2013;
Masutomi et al.. 2003; NTP-CERHR. 2003; Waterman et al.. 2000; Covance Labs. 1998c; Lington et
al s ,	» ^nics. 1987). Bio/dynamics (1987) found lower body weight in high dose (672

mg/kg-bw/day) females during most timepoints from week 11 through 94. In a 104-week feeding study
with mice, Covance La 8c) found lower body weight in 741 mg/kg-bw/day DINP diet fed male
mice. This effect was consistent in weeks 29 through 105. The same study found a lower female mouse
body weights when fed 741 mg/kg-bw/day DINP. This effect was observed in weeks 29, 37, and weeks
45 through 105. In one- and two-generation studies with rats, Exxon Biomedical (1996a) found
reductions in parental male and female body weights in both generations at feeding doses as low as 301
mg/kg-bw/day in the first generation and 288 mg/kg-bw/day DINP in the second generation.

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410	Survival: In studies of adult rat survival, fewer rats survived while being fed DINP compared to control

411	rats (LOAEL range 184 to 733.2 mg/kg/d) (Covance Labs. 1998c; Lington et at.. 1997). DINP diets also

412	lowered the survival of adult mice compared to controls (LOAEL= 1560.2 mg/kg/d) (Covar

413	1998a; Lington et a 7). Using these studies and guidance from Eco-SSLs (	),

414

415	Avian

416	No avian hazard studies were reasonably available to assess potential hazards from DINP exposure.

417

Table 4-1. Terrestrial Mammal Hazard Studies of

)INP Used for TRY Derivation

Test
Organism

NOAEL/
LOAEL
(mg/kg-day)

Effect

Study Description
(Duration/Dose)

Citation
(Rating)

Sprague-
Dawley rat

(Rattus
norvegicus)

31/307

Reproduction: lower
male pup body weight
at prepubertal necropsy
(PND 27)

18-day diet exposure to maternal
females (GD 15-22, PND 1-10). Target
concentrations were 400, 4,000, and
20,000 ppm (0, 30.7, 306.7 and 1,164.5
mg/kg/day).

(Masutomi et

al. 2003)
(Medium)

Wistar rat

(Rattus
norvegicus)

200/1000

Reproduction: lower
maternal body weights
in the 1,000 mg/kg
group at GD 15

10-day gavage exposure to pregnant
females (GD 6-15). Target
concentrations were 0, 40, 200, 1,000
mg/kg/day.

(Hellwig et
)

(Medium)

Sprague-
Dawley rat

(Rattus
norvegicus)

500/1000

Reproduction: lower
maternal body weight
gain GD 6-9 and 6-15.

10-day gavage exposure to pregnant
females (GD 6-15). Target
concentrations were 0, 100, 500, 1,000
mg/kg/day.

(Waterman et
9)

(High)

Sprague-
Dawley rat

(Rattus
norvegicus)

377/741

Reproduction: lower
maternal body weight
at GD 21

One generation study diet exposure (10
weeks prior to mating, through mating,
gestation, and lactation). Target doses
correspond to dietary concentrations of
0, 0.5, 1, and 1.5% (0, 377, 741, 1,087
mg/kg/day).

(Exxon
nedical.
)

(Medium)

Sprague-
Dawley rat

(Rattus
norvegicus)

287/555

Reproduction: lower
male F1 body weight at
birth (PND 0)

Two-generation study diet exposure.
Target doses correspond to dietary
concentrations of 0, 0.2, 0.4, and 0.8%
(0, 146, 287, 555 mg/kg/day).

(Exxon
Biomedical.
5b)(High)

Sprague-
Dawley rat

(Rattus
norvegicus)

139/274

Reproduction: lower
male F2 offspring body
weight at PND 7 and
PND 21.

Two-generation study diet exposure.
Target doses correspond to dietary
concentrations of 0, 0.2, 0.4, and 0.8%
(0, 143, 288, 560 mg/kg/day).

(Exxon
Biomedical.
5b)(High)

Wistar rat

(Rattus
norvegicus)

750/900

Reproduction: lower
male pup weight at
PND 13

3 3-day gavage exposure to maternal
females (GD 7-22, PND 1-17). Target
concentrations were 0, 300, 600, 750,
900 mg/kg/day.

(Boberg et al..
2011)
(Medium)

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Test
Organism

NOAEL /
LOAEL
(mg/kg-day)

Effect

Study Description
(Duration/Dose)

Citation
(Rating)

Sprague-
Dawley rat

(Rattus
norvegicus)

56/288

Reproduction: lower
male pup body weight
at PND 14

25-day feeding exposure to gestational
and lactating females (GD 12 through
PND 14). Target doses correspond to
dietary concentrations of 0, 760, 3,800,
and 11,400 ppm (0, 56, 288, 720
mg/kg/day).

(Clewell et ah,

2013)

(Medium)

Sprague-
Dawley rat

(Rattus
norvegicus)

307/1165

Growth: lower
maternal body weight
(PND 2-10)

18-day dietary exposure to maternal
animals (GD 15 to PND 10). Target
concentrations were 0, 400, 4,000, and
20,000 ppm (30.7, 306.7 and 1,164.5
mg/kg/day).

(Masutomi et

al.. 2003)
(Medium)

Sprague-
Dawley rat

(Rattus
norvegicus)

331/672

Growth: lower
maternal body weight
(PND 2-10)

2-year chronic dietary. Doses
correspond to dietary concentrations of
0, 33, 331, and 672 mg/kg/day.

(Bio/dvnamics.
1*^7) (High)

Fischer 344
rat (Rattus
norvegicus)

88/359

Growth: lower male
body weight gain

Chronic (105-week) diet exposure.
Target dietary doses were 0, 500, 1,500,
6,000, and 12,000 ppm (0, 29.2, 88.3,
358.7, 733.2 mg/kg/day).

(Covance Labs.
t(^8c) (High)

Sprague-
Dawley rat

(Rattus
norvegicus)

301/622

Growth: lower male
body weight

One-generation reproduction study.
Target dietary concentrations were 0,
0.5, 1, and 1.5% (0, 301, 622, 966
mg/kg/day).

(Exxon

Biomedical
>a)

(Medium)

Sprague-
Dawley rat

(Rattus
norvegicus)

363/734

Growth: lower parental
female body weight

One-generation reproduction study.
Target dietary concentrations were 0,
0.5, 1, and 1.5% (0, 363, 734, 1,114
mg/kg/day).

(Exxon

Biomedical
>a)

(Medium)

Sprague-
Dawley rat

(Rattus
norvegicus)

347/673

Growth: lower PI adult
body weight

Two-generation reproduction study.
Target dietary concentrations were 0,
0.2, 0.4, and 0.8% (0, 146, 287, 555
mg/kg/day).

(Exxon
Biomedical
»b) (High)

Sprague-
Dawley rat
{Rattus
norvegicus)

348/718

Growth: lower P2 adult
female body weight

Two-generation reproduction study.
Target dietary concentrations were 0,
0.2, 0.4, and 0.8% (0, 143, 288, 560
mg/kg/day).

(Exxon
Biomedical

»b)
(Medium)

B6C3F1
Mouse (Mils

musculus)

276/742

Growth: lower adult
male body weight

104-week dietary exposure to adult
mice. Target dietary concentrations
were 0, 500, 1,500, 4,000, and 8,000
ppm (0, 90.3, 275.6, 741.8, 1,560.2
mg/kg/day).

(Covance Labs.
t()()8b) (High)

B6C3F1
Mouse (Mils

musculus)

336/910

Growth: lower adult
female body weight

104-week exposure to adult mice.
Target dietary concentrations were 0,
500, 1,500, 4,000, and 8,000 ppm (0,
112, 335.6, 910.3, 1,887.6 mg/kg/day).

(Covance Labs.
t()()8b) (High)

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Test
Organism

NOAEL /
LOAEL
(mg/kg-day)

Effect

Study Description
(Duration/Dose)

Citation
(Rating)

Fischer 344
rat (Rattus
norvegicus)

15/152

Growth: lower adult
male body weight

Chronic (2-year) dietary study in rats.
Target dietary concentrations were 0,
0.03, 0.3, and 0.6% (0, 15, 152, 307
mg/kg/day).

(Bio/dvnamics.
l«>Sii) (High)

Sprague-
Dawley rat

(Rattus
norvegicus)

555/1513

Growth: lower
maternal body weight
at PND 2 and PND 14

25-day exposure (GD 12 through PND
14) to maternal rats. Target dietary
concentrations were 0, 760, 3,800, and
11,400 ppm (0, 109,555, 1513
mg/kg/day).

(Clewell et ah,

2013)

(Medium)

Fischer 344
rat (Rattus
norvegicus)

1192/2289

Growth: lower male
body weight

21-day dietary exposure to male and
female rats. Target dietary
concentrations were 0, 0.6, 1.2, and
2.5% (0, 639, 1,192, 2,195 mg/kg/day).

( beret ah,

(<^7)

(Medium)

Fischer 344
rat (Rattus
norvegicus)

359/733

Survival: lower male
survival

105-week dietary exposure to male and
female rats. Target dietary
concentrations were 0, 500, 1,500,
6,000, and 12,000 ppm (0, 29.2, 88.3,
358.7, 733.2 mg/kg/day).

(Covance Labs,
t(^8c) (High)

B6C3F1
Mouse (Mils

musculus)

742/1560

Survival: lower male
survival

104-week dietary exposure to male and
female mice. Target dietary
concentrations were 0, 500, 1,500,
4,000, and 8,000 ppm (0, 90.3, 275.6,
741.8, 1,560.2 mg/kg/day).

(Covance Labs.
t()()8b) (High)

Fischer 344
rat (Rattus
norvegicus)

18/184

Survival: lower female
survival

2-year dietary exposure to male and
female rats. Target dietary
concentrations were 0, 0.03, 0.3, and
0.6% (0, 18, 184, 375 mg/kg/day).

(Bio/dvnatnics.
l«>Sii) (High)

Terrestrial Invertebrates

EPA identified one study of DINP chronic exposure to the earthworm Eisenia fetida in artificial soil
(ExxonMobil. 2010). This study found no difference in mortality between earthworms in control soil
and soil containing nominal concentrations of 1,000 mg/kg dw DINP. The soil concentrations were
analyzed by gas chromatography with flame ionization detection and ranged from 925.2 to 1052 mg/kg
on Day 0 and from 651.4 to 795.8 mg/kg on Day 28 and from 389.6 to477.1 mg/kg on Day 56.

However, the study found a difference between the number of juveniles found in 1,000 mg/kg dw DINP
soils (mean=90) versus a mean of 39 worms found in no-DINP control soils.

Terrestrial Plants

No terrestrial plants studies were available to assess potential hazards from DINP exposure.

4.1 Terrestrial Organism Hazard Conclusions

Overall, EPA has moderate confidence in the evidence that DINP poses low hazard to terrestrial
mammals via dietary exposure, but robust confidence that DINP poses no hazard to soil invertebrates
(see Table 5-1). No studies on DINP exposure to wild mammals, birds, or plants were available to assess
DINP hazard, indicating that no hazard has been observed in these groups under realistic exposure

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436	conditions. EPA reviewed studies of laboratory rodents to derive a TRV of 139 mg/kg-bw/day dietary

437	DINP exposure. This TRV represents the potential chronic exposure dose at which the dietary effects of

438	DINP may affect a general mammal. Thus, EPA has only moderate confidence that the TRV represents

439	realistic hazards to wild populations. Chronic DINP exposure to an earthworm species in soil did not

440	affect earthworm survival, indicating little to no hazard of DINP to soil dwelling invertebrates.

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5 WEIGHT OF SCIENTIFIC EVIDENCE CONCLUSIONS FOR
ENVIRONMENTAL HAZARD

Overall, EPA has determined that DINP poses low hazard potential in aquatic species and has robust
confidence in the evidence showing low acute aquatic hazard, low acute benthic hazard, low chronic
benthic hazard, and low aquatic plant hazard and moderate confidence in the evidence showing low
chronic aquatic hazard to fish (see Aquatic Organism Hazard Conclusions). Within the terrestrial
environment, EPA has moderate confidence in the evidence showing low chronic dietary hazards of
DINP to terrestrial mammals and robust confidence in the evidence for low soil invertebrate hazard (see
Terrestrial Organism Hazard Conclusions). Thus, the weight of scientific evidence leads EPA to having
robust confidence in the overall conclusion that DINP has little to no hazards to wild organism
populations. However, EPA has more uncertainty and less confidence in the size and quality of the
studies in the database, the strength and precision of more subtle and mechanistic effects found within a
few studies, and whether study design allowed for dose-response effects to be detected for mechanistic
endpoints. A more detailed explanation of the weight of the scientific evidence, uncertainties, and
overall confidence levels is presented in Appendix A.l. EPA uses several considerations when weighing
the scientific evidence to determine confidence in the environmental hazard data. These considerations
include the quality of the database, consistency, strength and precision, biological gradient/dose
response, and relevance (see Appendix A.2), and are consistent with the 2021 Draft Systematic Review
Protocol (	21a). Table 5-1 summarizes how these considerations were determined for each

environmental hazard.

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461 Table 5-1. DINP 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 assessment

+++

+++

+++

++

+++

Robust

Acute benthic assessment

++

+++

+++

++

+

Robust

Chronic aquatic assessment

++

+

+

+

+++

Moderate

Chronic benthic assessment

++

++

++

+

+++

Robust

Algal assessment

+

+++

++

++

+++

Robust

1 envslnal

Avian assessment

Indeterminate

Indeterminate

Indeterminate

Indeterminate

Indeterminate

Indeterminate

Chronic mammalian assessment

++

+++

+++

+++

+

Moderate

Soil invertebrate assessment

+

Not applicable

+

+

+++

Robust

Terrestrial plant assessment

Indeterminate

Indeterminate

Indeterminate

Indeterminate

Indeterminate

Indeterminate

a Relevance includes biological, physical/chemical (including use of analogues), and environmental relevance.

h Hazard Confidence reflects the overall confidence in the conclusions about the presence or absence of hazard thresholds and the weight of support and

uncertainties around all the available data and does not necessarily represent a summation of the individual evidence properties.

+++ 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.

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479

480

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482

483

484

485

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487

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6 ENVIRONMENTAL HAZARD THRESHOLDS

Aquatic Species Hazard Values

Acute Aquatic Threshold: No definitive hazard values or concentrations of concern were identified from
the studies of acute exposure of DINP on aquatic organisms that live in the water column. Thus, EPA
found no hazards from acute water exposure of DINP to aquatic organisms.

Acute Benthic Threshold: No definitive hazard values or concentrations of concern were identified from
the studies of acute exposure of DINP on benthic organisms. Thus, EPA found no hazards from acute
exposure of DINP to aquatic organisms living in benthic habitats.

Chronic Aquatic Threshold: No definitive hazard concentrations via water or dietary exposure were
identified from the studies of chronic exposure of DINP on aquatic organisms. Thus, EPA found no
survival or reproductive hazards of chronic DINP to aquatic organism populations.

Chronic Benthic Threshold: No definitive hazard values or concentrations of concern were identified
from the studies of chronic exposure of DINP on benthic organisms. Thus, EPA found no hazards from
chronic exposure of DINP to aquatic organisms living in benthic habitats.

Aquatic Plant Threshold: No definitive hazard values or concentrations of concern were identified from
the studies of DINP effects on algae. Thus, EPA found no hazards from acute or chronic exposure of
DINP to aquatic plants.

Terrestrial Species Hazard Values

Terrestrial Vertebrate Threshold: For terrestrial species exposed to DINP, EPA estimated hazard using a
deterministic approach to calculate a TRV expressed as doses in units of mg/kg-bw/day (for mammals)
(Figure 6-1). Although the TRV for DINP was derived from laboratory mice and rat studies, body
weight was standardized, therefore the TRV can be used with ecologically relevant wildlife species to
evaluate the potential toxicity of chronic dietary exposure to DINP. The following criteria and steps
(Figure 6-2) were used to select the data to calculate the TRV for DINP with NOAEL and/or LOAEL
data using (U.S. EPA. 2007). General step descriptions are in italics, while EPA's step by step decisions
for DINP are in regular text (Figure 6-2).

Step 1: The minimum data set required to derive either a mammalian or avian TRV consists of three
results (NOAEL or LOAEL values) for reproduction, growth, or mortality for at least two mammalian or
avian species.

EPA assessed 12 studies with 24 reported NOAELs and 24 reported LOAELs. The studies included
multiple strains of rat (R. norvegicus) including Sprague-Dawley, Wistar, and Fischer344, and one strain
of mouse (M musculus).

Because this condition was met, EPA proceeded to Step 2.

Step 2: Calculation of a geometric mean requires at least three NOAEL results from the reproduction
and growth effect groups.

Nine reproduction NOAEL results and 12 growth NOAEL results were reported from these studies.
Because this condition was met, EPA proceeded to Step 4.

Step 4: When the geometric mean of the NOAEL for reproduction and growth is higher than the lowest
bounded LOAEL for reproduction, growth, or mortality, then the TRV is equal to the highest bounded
NOAEL below the lowest bounded LOAEL.

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The geometric mean of NOAELs for reproduction and growth was 230 mg/kg-bw/day, which was
higher than the lowest bounded LOAEL of 152 mg/kg-bw/day DINP from a study of reduced male body
weight after 2 years of dietary exposure (Lington et ai. 1997). The highest bounded NOAEL less than
the lowest bounded LOAEL was 139 mg/kg-bw/day DINP (Waterman et ai. 2000) a concentration
corresponding to a reduction in second generation male rat body weight after 19 weeks of dietary
exposure. Therefore, the terrestrial mammal TRV was 139 mg/kg-bw/day DINP in the diet.

Summary of Environmental Hazard Thresholds

The effects of DINP on a generalized small mammal after consistent and prolonged ingestion of DINP
in their diets (Table 6-1).

Table 6-1. Environmental Hazard Threshold for Aquatic and Terrestrial (TRV) Environmental
Toxicity			

Environmental Terrestrial Toxicity

Assessment Medium

Hazard Value or TRV

Mammal (TRV)

Dietary

139 mg DINP/kg-bw/day

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10000

1000 =

— 100
>

-8
Sf

"a io

0)

i

o

0.1 ,

0.01

t

i 8 j S o 8 8 8 8 ? i

I i I i 1 I in

2 2

~ 5

ec

£ 2

oo oo

5 a

I

TRV = Highest bounded NOAEL lower than
the lowest bounded LOAEL for
Reproduction, Growth and Survival = 139

Geometric Mean of NOAELs for
Reproduction and Growth = 230

i

i

O

2

I

5

DC

Reproduction
(REP)

Growth
(GRO)

Survival
(SUR)

527

528

(closed circle) - No-Observed Adverse Effect Dose

(open circle) - Lowest-Observed Adverse Effect Dose

Result number —

1) 10 - Rat. MORT

X t \

Reference number
Test Species Kev:

Effect Measure

Effect Measure Kev:

BDWT - body weight changes
GREP - general reproduction
MORT - mortality
PRWT - progeny weight

o

"Lowest-Observed Adverse Effect Dose

	Paired values from same study when

jomed by lme
h— No-Observed Adverse Effect Dose

Test Species
Mou - Mouse
Rat - Rat
Wildlife TRV Derivation Process

1)	There are at least three results available for two test species within the growth, reproduction, and survival effect groups. There are enough data to derive a
TRV.

2)	There are at least three NOAEL results available in the growth and reproduction effect groups for calculation of a geometric mean.

3)	The geometric mean of the NOAEL values for growth and reproductive effects equals 230 mg Di-isononyl phthalate/kg BW'day. which is greater than the
lowest bounded LOAEL of 152 mg Di-isononyl phthalate/kg BW 'day for reproduction, growth or survival.

4)	The Mammalian wildlife TRV for Di-isononyl phthalate is equal to 139 mg Di-isononyl phthalate/kg BW 'day. which is the highest bounded NOAEL
below the lowest bounded LOAEL for reproduction, growth or survival.

Figure 6-1. Terrestrial Mammal TRV Derivation for DINP in Mammal Diets

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529

NO

Stq> 1:

YES

Step 2:

Ate there at least 3

Ate there 3 or mote
NQAELs in REP
and GRO?

tonicityvalues fbr 2
species REP. GRO or



MOR?



NO 1 YES

TRV = lowest
LOAEL 10

No TRV can
be derived

NO

St


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carcinogenic evaluation of diisononyl phthalate in rats. Fundam Appl Toxicol 36: 79-89.
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on Karenia brevis. Chemosphere 155: 498-508.
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Masutomi. N: Shibutani. M; Takagi. H; Unevama. C: Takahashi. N: Hirose. M. (2003). Impact of

dietary exposure to methoxychlor, genistein, or diisononyl phthalate during the perinatal period
on the development of the rat endocrine/reproductive systems in later life. Toxicology 192: 149-
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N	(2003). NTP-CERHR monograph on the potential human reproductive and

developmental effects of di-isononyl phthalate (DINP) (pp. i-III90). (NIH Publication No. 03-
4484). Research Triangle Park, NC: National Toxicology Program Center for the Evaluation of
Risks to Human Reproduction.

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Parkerton. TF; Konkel. WJ. (2000). Application of quantitative structure—activity relationships for
assessing the aquatic toxicity of phthalate esters. Ecotoxicol Environ Saf 45: 61-78.
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Patyn . I i Y xut JY 1 Uvi. RA; Letinsl i Y < liomas. PE; Cooper. KR; Parkerton. TF. (2006).
Hazard evaluation of diisononyl phthalate and diisodecyl phthalate in a Japanese medaka
multigenerational assay. Ecotoxicol Environ Saf 65: 36-47.
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Rhod	ims. 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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Santangeli. S: Maradonna. F; Zanardini. M; Notarstefano. V: Gioacch:	ler-Piqiier. I: Habibi. H;

Carneval (2017). Effects of diisononyl phthalate on Danio rerio reproduction. Environ Pollut
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report) [TSCA Submission], (Report No. BW-84-4-1567. OTS0508408. 42005 B4-10. 40-
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OTS0508409. 40-8426151. 42005 B4-11. TSCATS/206782). Washington, DC: Chemical
Manufacturers Association.

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Sprineborn 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.
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1	\ (1998). Guidelines for ecological risk assessment [EPA Report], (EPA/630/R-95/002F).

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U.S. EPA. (2005). Guidelines for carcinogen risk assessment [EPA Report], (EPA630P0300IF).
Washington, DC. https://www.epa.gov/sites/production/files/2Q13-
09/docum ents/can cer guidelines final 3-25-05.pdf
U.S. EPA. (2007). Attachment 4-3 Guidance for Developing Ecological Soil Screening Levels (Eco-
SSLs) Eco-SSL Standard Operating Procedure (SOP) #4: Wildlife Toxicity Reference Value
Literature Review, Data Extraction and Coding. (OSWER9285755F).
http://nepis.epa.eov/exe/ZvPURL.cei?Dockev=P 100CDHC.txt
U.S. EPA. (202 la). Draft systematic review protocol supporting TSC A 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.reeiilations.eov/dociiment/EPA-HQ-OPPT-2'

0005

U.S. EPA. (202 lb). Final scope of the risk evaluation for di-isononyl phthalate (DINP) (1,2-benzene-
dicarboxylic acid, 1,2-diisononyl ester, and 1,2-benzenedicarboxylic acid, di-C8-10-branched
alkyl esters, C9-rich); CASRNs 28553-12-0 and 68515-48-0 [EPA Report], (EPA-740-R-21-
002). Washington, DC: Office of Chemical Safety and Pollution Prevention.

https://www.epa.eov/sYStem/files/documents/2021-08/casrn-2l	di-isononyl-phthalate-

final-scope.pdf

U.S. EPA. (202 lc). Final use report for di-isononyl phthalate (DINP) - (1,2-benzene-dicarboxylic acid,
1,2-diisononyl ester, and 1,2-benzenedicarboxylic acid, di-C8-10-branched alkyl esters, C9-rich)
(CASRN 28553-12-0 and 68515-48-0). (EPA-HQ-OPPT-2018-0436-0035). Washington, DC:
U.S. Environmental Protection Agency. https://www.reeiilations.eov/dociiment/EPA-HQ-OPPT-
2018-0436-0035

U.S. EPA. (2024). Draft Physical Chemistry Assessment for Diisononyl Phthalate (DINP). Washington,
DC: Office of Pollution Prevention and Toxics.

Waterman. SJ: Ambroso. JL; Keller. LH; Trimmer. GW; Mkiforoi [arris. SB. (1999).

Developmental toxicity of di-isodecyl and di-isononyl phthalates in rats. Reprod Toxicol 13:
131-136. http://dx.doi.ore/i 0. i 0 i 6/80890-6238(99)0000:

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McKee. RH. (2000). Two-generation reproduction study in rats given di-isononyl phthalate in
the diet. Reprod Toxicol 14: 21-36. http://dx.doi.ore/10.1016/50890-6238(99)00067-2

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Appendix A ENVIRONMENTAL HAZARD DETAILS

A.I Evidence Integration

Data integration includes analysis, synthesis, and integration of information for the draft risk evaluation.
During data integration, EPA considers quality, consistency, relevancy, coherence, and biological
plausibility to make final conclusions regarding the weight of the scientific evidence. As stated in the
Draft Systematic Review Protocol Supporting TSCA Risk Evaluations for Chemical Substances (U.S.
E 2la), data integration involves transparently discussing the significant issues, strengths, and
limitations as well as the uncertainties of the reasonably available information and the major points of
interpretation.

The general analytical approaches for integrating evidence for environmental hazard is discussed in
Section 7.4 of the 2021 Draft Systematic Review Protocol (U.S. EPA. 2021a).

The organization and approach to integrating hazard evidence is determined by the reasonably available
evidence regarding routes of exposure, exposure media, duration of exposure, taxa, metabolism and
distribution, effects evaluated, the number of studies pertaining to each effect, as well as the results of
the data quality evaluation.

The environmental hazard integration is organized around effects to aquatic and terrestrial organisms as
well as the respective environmental compartments (e.g., pelagic, benthic, soil). Environmental hazard
assessment may be complex based on the considerations of the quantity, relevance, and quality of the
available evidence.

For DINP, environmental hazard data from toxicology studies identified during systematic review have
used evidence that characterizes apical endpoints; that is, endpoints that could have population-level
effects such as reproduction, growth, and/or mortality. Additionally, mechanistic data that can be linked
to apical endpoints will add to the weight of the scientific evidence supporting hazard thresholds.

A.2 Weight of Scientific Evidence

After calculating the hazard thresholds that were carried forward to characterize risk, a narrative
describing the weight of scientific evidence and uncertainties was completed to support EPA's
decisions. The weight of 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 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 (	321a) 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 (	)21a).

EPA used the strength-of-evidence and uncertainties from (	) for the hazard assessment

to qualitatively rank the overall confidence using evidence Table 5-1 for environmental hazard.
Confidence levels of robust (+ + +), moderate (+ +), slight (+), or indeterminant are assigned for each
evidence property that corresponds to the evidence considerations (U.S. EPA. 2021a). The rank of the
Quality of the Database consideration is based on the systematic review overall quality determination

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(High, Medium, or Low) for studies used to calculate the hazard threshold, and whether there are data
gaps in the toxicity dataset. 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 determinations rank 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 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.

Confidence Levels

•	Robust (+ + +) confidence suggests thorough understanding of the scientific evidence and
uncertainties. The supporting weight of 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 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.

•	Indeterminant (N/A) corresponds to entries in evidence tables where information is not available
within a specific evidence consideration.

Types of Uncertainties

The following uncertainties may be relevant to one or more of the weights of scientific evidence
considerations listed above and will be integrated into that property's rank in the evidence table (Table
5-1):

•	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 Apx A-l summarizes the weight of 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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813	TableApx A-l. Considerations that Inform Evaluations of the Strength of the Evidence within an Evidence Stream (Le., Apical

814	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"
(risk of bias)

•	A large evidence base of high- or medium-qaa[ity
studies increases strength.

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

•	An evidence base of mostly low-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 CU.S. EPA. 1998). enduoint
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 analogue 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 dataset 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.

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A.3 Strengths, Limitations, Assumptions, and Key Sources of Uncertainty
for Environmental Hazard

Quality of the Database; Consistency; Strength (Effect Magnitude); and Precision

The database for the acute aquatic assessment consisted of 14 studies representing five fishes and four
invertebrate species (Chen et al. 2014; Brown et ai. 1998; Adams et ai. 1995;	>nomics.

1984n. h; c.prineborn Bionomics. 198 I b, >1 *, r^onomics. 1983 a. b). Twelve of the 14 studies
received overall quality determinations of high, while the two other studies received overall quality
determinations of medium increasing the overall strength of evidence for database quality (Chen et al..
2014; Springborn Bionomics. 1984a). Five fish species were represented with acute duration studies,
and two aquatic invertebrate species were represented by three studies on D. magna and two studies on
M. bahia, resulting in robust confidence in the overall quality of the database. All studies within the pool
of reasonably available information resulted in similar findings of no acute adverse effects up to the
limit of water solubility across all species and between vertebrate (Table 3-1) and invertebrate taxa
(Table 3-2), resulting in robust confidence in the consistency of the database. Seven out of the eight
acute aquatic studies were conducted with analytical verification of concentrations of DINP and details
on precise results among treatment and control groups indicates robust confidence in the strength and
precision of the exposure-response relationship.

The database for the acute benthic assessment consisted of two studies, both with overall quality
determinations of high and representing P. yarthenogenetica (Adams et al \ , * *, * Y^nomics.
1984c). Moderate confidence in the overall quality of the database was determined, as the studies on
benthic and epibenthic aquatic invertebrates produced two independent results. These studies
demonstrated similar results within the same species tested (Table 3-2) leading to robust confidence
assigned to the consistency consideration. Both studies were conducted with analytical verification of
concentrations of DINP and provides precise detailed results of the data recorded, thereby providing
robust confidence in the strength and precision of the exposure concentrations and associated response.

The database for the chronic aquatic assessment consisted of ten studies representing three fish species

(Carnevali et al.. 2019; Fomer-Piquer et al.. 2019; Fomer-Piquer et al.. 2018b; Fomer-Piquer et al..
2018a; Santangeli et al. < s , I .tvna et al.. 2006) and two aquatic invertebrates (Carnevali et al.. 2019;
Fomer-Piquer et al.. 2019; Fomer-Piquer et al.. 2018b; Fomer-Piquer et al.. 2018a; Santangeli et al..
2017; Patyna et al.. 2006; Brown et al.. 1998; Lake Superior Research Institu 7; Rhodes et al..
1995) and two aquatic invertebrates (Call et al.. 1 ^M, Tm< *wn et al.. 1998; Lake Superior Research
Institute. 1997; Rhodes et al.. 1995). Four subchronic studies were conducted with 21 -day aquatic
exposures of DINP with two studies on zebrafish, two studies on D. magna, and two studies on the
epibenthic amphipod, H. azteca. The remaining four studies were on dietary exposures of DINP to
Japanese medaka (O. latipes) and gilthead sea bream (S. aurata). The dietary study conducted on 0.
latipes received an overall quality determination of high, while the remaining three dietary studies
conducted on S. aurata received medium overall quality determinations. Studies conducted with 21-day
aquatic exposures were of limited statistical power, observed inconsistent dose-response effects, were
not analytically verified, and exceeded solubility (Fomer-Piquer et al.. 2018b; Santangeli et al.. 2017).
The 21-day feeding studies conducted on aquatic vertebrates displayed limited replication and sample
sizes, relied on nominal concentration with no analytical verification of DINP within the feed, and did
not demonstrate impacts on apical endpoints (Carnevali et al.. 2019; Fomer-Piquer et al.. 2019; Forner-
Piquer et al.. 2018a). Moderate confidence was assigned to the overall quality of the chronic aquatic
assessment database due to the low number of studies with apical endpoints from relative few species
represented.

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Both chronic duration studies conducted on aquatic invertebrates resulted in similar observations of no
adverse effects from 21-day exposures of DINP, with one study observing presumed adverse effects
from surface entrapment at the highest concentration tested (Rhodes et al. 1995) and the other study
observing no adverse effects at an increased concentration of 1 mg/L DINP aided by the application of a
dispersant n et al. (1998). Two studies with aquatic exposures of DINP to J). rerio for 21-days
resulted in reproductive impacts (Forner-Piquer et al.. 2018b; Santangeli et al.. 2017). and two of the
four studies conducted with dietary exposures of DINP were consistent in demonstrating adverse apical
effects (Forner-Piquer et al.. 2019; Patvna et al.. 2006). indicating slight confidence regarding the
consistency of effects on aquatic species from chronic exposure. Effect size, replication, and analytical
verification of DINP within studies on chronic exposures to invertebrates and vertebrates was observed
within studies such as Patvna et al. (2006) and Rhodes et al. (1995); however, low sample sizes and lack
of analytical verification within other studies Santangeli et;	decreased evidence strength

resulting in slight confidence in the strength and precision of the exposure-response relationship.

The database for the chronic benthic assessment consisted of one study representing sediment exposures
of DINP to an amphibian species (R. arvalis) and two studies with an invertebrate species, C. tentans
(Call et al.. 2*- ^ m , 1 1 2001; Lake Superior Research Institute. 1997). All three studies received overall
quality determinations of high with studies on benthic invertebrates using subchronic 10-day exposures.
Slight confidence was assigned to the overall quality of the database due to the limited number of
studies, subchronic exposure duration, and the relevant species represented. No adverse effects were
observed for the amphibian study, R. arvalis, throughout the 26-day exposures of DINP spiked sediment
which was conducted from the embryo to tadpole stage (IVL. 2001). Moderate confidence was assigned
to consistency for the chronic benthic assessment. Decreased confidence strength for the invertebrate
chronic benthic assessment originates from the subchronic duration exposures to DINP spiked sediment
within the two invertebrates studies, predominately following the OCSPP test guideline detailed within
OCSPP 850.1735 Spiked Whole Sedimer> i,»;i\ l'<»\i.citv Test hyshwater Invertebrates (Call et al.
2001; Lake Superior Research Institute. 1997). All three studies were conducted with analytical
verification of concentrations of DINP, therefore moderate confidence was attributed to the strength and
precision.

The database for the aquatic plant assessment consisted of three studies of algae, with two studies
having overall quality determinations of high conducted on S. capricornutum (Adams et al.. 1995;
Springborn Bionomics. 1984c) and one study having an overall quality determination of medium
conducted on the marine dinoflagellate, K. brevis (Liu et al.. 2016). Slight confidence was assigned to
the overall quality of the database due to the relatively limited number of studies and species
represented. All studies were conducted with exposure durations of 96-hour and resulted in similar
findings of no acute adverse effects on cell number up to the limit of water solubility across both species
investigated (Table 3-3), providing robust confidence in the consistency in results of the algal
assessment. Both studies conducted on S. capricornutum included analytical verification of DINP
concentrations, while the study conducted on K. brevis reported nominal concentrations, indicating
moderate confidence in the strength and precision consideration for the algal assessment.

The database for terrestrial mammals and the TRV derivation consisted of 12 studies that documented
the DINP effects on laboratory rat and mouse reproduction, growth, and survival endpoints. EPA has
moderate confidence in this database because the studies used model mammals to inform human health
and not wildlife species. EPA has robust confidence in the consistency of the DINP effects on mammals
because the effects were consistently observed at concentrations within the same order of magnitude.
Similar strength and precision of the effects were observed across strains of rat and one mouse species,

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resulting in a TRV that can be interpreted across many studies. Thus, EPA has robust confidence in
these effects and the resultant TRV.

The database for terrestrial invertebrates consisted of one study (ExxonMobil. 2010) that found no
mortality effects of soil DINP on E. fetida. EPA has slight confidence in quality of the database,
consistency, and strength (effect magnitude) and precision because it is one study that represents one
unbounded hazard soil concentration.

Biological Gradient/Dose-Response: Several acute toxicity tests for aquatic and benthic organisms were
conducted with initial range finding tests followed by a definitive test with a single treatment
concentration near the limit of solubility. In general, this approach would be interpreted to decrease the
strength of the evidence for acute studies with aquatic and benthic organisms. However, given the fact
that there is consistency among acute tests in the demonstration of no adverse effects up to the limit of
solubility, EPA has moderate confidence in the biological gradient/dose-response for the acute toxicity
assessments for aquatic and benthic organisms.

Among the six chronic studies conducted on fishes, two aquatic exposure studies included three or more
treatment concentrations and had a medium overall quality determination (Forner-Piquer et ai. 2018b)
demonstrating evidence of concentration-response. A corresponding study from the same laboratory
reported non-linear adverse effects at all five treatment concentrations for the number of eggs per female
per day (Santangeli et al. 2017). The same study also reported adverse effects on gonadosomatic index
at the two lowest and highest concentrations among three out of five aquatic DINP treatments. None of
the chronic invertebrate studies with aquatic or benthic exposures reported any adverse effects resulting
from DINP exposure. Rhodes et al. (1995) reported adverse effects at the highest concentration tested
from 21-day DINP exposures to D. magna; however, as previously discussed impacts on mortality and
subsequent reproduction were attributed to entrapment at the water surface. Moderate confidence in the
biological gradient/dose-response consideration was assigned for the chronic toxicity assessments for
aquatic organisms. Slight confidence in the biological gradient/dose-response consideration was
assigned for the chronic assessments for benthic organisms due to a lack of DINP concentration
gradients in these studies (Call et al.. 2001; Lake Superior Research Institute. 1997).

Two of the three algal toxicity tests were conducted with initial range finding tests followed by a
definitive test with a single treatment concentration near the limit of solubility, limiting the assessment
of the biological gradient/dose-response consideration (Adams	5; Sprinebom Bionomics.

1984c). Liu et al. (2 used five concentrations and a control for their investigations of acute DINP
toxicity to the marine dinoflagellate, K. brevis, with no adverse effect on cell number at nominal
concentrations compared to controls. Moderate confidence in the biological gradient/dose-response
consideration was assigned for the algal assessment.

The database for terrestrial invertebrates consisted of one study (ExxonMobil. 2010) that found no
mortality effects of soil DINP on E. fetida. EPA has slight confidence in Biological Gradient/Dose-
response because only one test concentration was used. EPA has robust confidence in the dose-
responses in rodent studies used to derive the TRV because they used gradients of DINP concentration
in animal diets in their experimental designs.

Relevance (Biological; Physical/Chemical; Environmental): Acute aquatic studies similarly observed no
adverse impacts of mortality or immobilization from acute DINP exposures within five species of fish
and one invertebrate species. Test conditions for these species corresponded well with expected natural
environmental conditions. Seven of the eight acute aquatic studies were conducted without the use of a

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solvent and reported analytical verification of DINP treatment concentrations. Robust confidence in the
relevance considerations was assigned for the acute aquatic assessment.

Acute benthic studies were represented by 48- and 96-hour exposure studies on the midge, P.
parthenogenetica, (Adams et ai. 1995; EG & rnomics. 1984c) The consistency in results among
these independent studies on representative sediment-oriented species increases evidence strength for
this consideration. All acute benthic studies were conducted without the use of a solvent and reported
analytical verification of DINP treatment concentrations, providing moderate confidence in the
relevance consideration for the acute benthic assessment.

Chronic aquatic studies are represented by studies with both invertebrates and vertebrates. Test
concentrations were either not reported or not analytically verified for chronic aquatic studies with
zebrafish (Forner-Piquer et ai. 2018b; Santangeli et ai. 2017) and chronic feeding studies with gilthead
sea bream (Carnevali et ai. 2019; Forner-Piquer et ai. 2019; Forner-Piquer et ai. 2018a). Because of
this lack of analytical verification of concentrations, Moderate confidence in the relevance
considerations was assigned for the chronic aquatic assessment.

Chronic benthic studies were limited to subchronic duration exposures conducted with the amphipod, H.
azteca, the midge, C. tentans, and the moorfrog, R. arvalis, which are considered relevant study
organisms for sediment toxicity testing. Although no adverse effects on mortality or
development/growth were reported, these studies were conducted with 10-day exposures from DINP
spiked sediment. Both studies conducted analytical verification of DINP within sediment, and one study
(Lake Superior Research Institute. 1997) reported the corresponding concentration of DINP within
porewater. Slight confidence in the relevance considerations was assigned for the chronic benthic
assessment.

Algal toxicity studies are narrowly represented with the green algae, S. capricornutum, (Adams et ai.
1995; Sprineborn Bionomics. 1984c) and the marine dinoflagellate, K. brevis (Liu et ai. 2016). The two
studies on S. capricornutum were conducted with analytical verification of DINP concentrations, while
the remaining study on K. brevis did not perform analytical verification of the treatment concentrations
but reported the purity, source, and nominal concentration of DINP. Based on the limited landscape of
available studies for algal organisms and the duration of exposure, slight confidence in the relevance
consideration was assigned for the algal assessment.

The database for terrestrial invertebrates consisted of one study (ExxonMobil. 2010) that found no
mortality effects of soil DINP on earthworms. EPA has moderate confidence in its relevance (biological;
physical/chemical; environmental) because soil concentrations were analytically verified, and
earthworms are a relevant representative species. However, only one test concentration was used.

EPA has slight confidence in the relevance of the rodent studies and resultant TRV because they were
conducted on non-wildlife species in highly controlled laboratory experiments, and they mainly found
DINP effects after long term dietary exposures that may be unlikely in ecosystems. Additional
uncertainties associated with laboratory to field variation in exposures to DINP are likely to have some
effect on the hazard threshold; that is, formulated diets vs. natural forage diet for mammals (rats and
mice).

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