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EPA Document# EPA-740-D-24-012
May 2024
United States Office of Chemical Safety and
Environmental Protection Agency Pollution Prevention
Draft Fate 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 6
29 1 INTRODUCTION 7
30 2 APPROACH AND METHODOLOGY 8
31 2.1 EPI Suite™ Model Inputs and Settings 9
32 3 TRANSFORMATION PROCESSES 11
33 3.1 Biodegradation 11
34 3.2 Hydrolysis 12
35 3.3 Photolysis 13
36 4 PARTITIONING 14
37 5 MEDIA ASSESSMENTS 15
38 5.1 Air and Atmosphere 15
39 5.1.1 Indoor Air and Dust 15
40 5.2 Aquatic Environments 16
41 5.2.1 Surface Water 16
42 5.2.2 Sediments 16
43 5.3 Terrestrial Environments 17
44 5.3.1 Soil 17
45 5.3.2 Biosolids 18
46 5.3.3 Landfills 18
47 5.3.4 Groundwater 19
48 6 PERSISTENCE POTENTIAL OF DINP 20
49 6.1 Destruction and Removal Efficiency 20
50 6.2 Removal in Wastewater Treatment 20
51 6.3 Removal in Drinking Water Treatment 21
52 7 BIO AC CUMULATION POTENTIAL OF DINP 22
53 8 OVERALL FATE AND TRANSPORT OF DINP 25
54 9 WEIGHT OF SCIENTIFIC EVIDENCE CONCLUSIONS FOR FATE AND
55 TRANSPORT 26
56 9.1 Strengths, Limitations, Assumptions, and Key Sources of Uncertainty for the Fate and
57 Transport Assessment 26
58 REFERENCES 27
59
60 LIST OF TABLES
61 Table 2-1. Summary of DINP's Environmental Fate Information 8
62 Table 3-1. Summary of DINP's Biodegradation Information 12
63 Table 7-1. Summary of DINP's Bioaccumulation Information 23
64
65
66
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67 LIST OF FIGURES
68 Figure 4-1. EPI Suite™ Level III Fugacity Modeling Graphical Result for DINP 14
69
70 ABBREVIATIONS AND ACRONYMS
BAF Bioaccumulation factor
BCF Bioconcentration factor
BMF Biomagnification factor
BSAF Biota-sediment accumulation factor
CASRN Chemical Abstract Service registry number
DOC Dissolved organic carbon
dw Dry weight
DW Drinking water
DWTP Drinking water treatment plant
EPA Environmental Protection Agency
ESI Electrospray Ionization
FID Flame ionization detector
FPD Flame photometric detector
GC Gas chromatography
HLC Henry's Law constant
HPLC High-performance liquid chromatography
ISO International Organization for Standardization
Koa Octanol-air partition coefficient
Koc Organic carbon-water partition coefficient
Kow Octanol-water partition coefficient
L/d Liters per day
LOD Limit of detection
LOQ Limit of quantification
lw Lipid weight
M Molarity (mol/L = moles per Liter)
mL/min milliliters per minute
mM millimolar
MDL Method Detection Limit
MRL Method Reporting Limit
MS Mass spectrometry
n Sample size
N/A Not applicable
ND Non-detection
nm nanometers
NR. Not reported
OECD Organisation for Economic Co-operation and Development
• OH Hydroxyl radical
OPE Organophosphate ester
pg/L picograms per liter
ppm parts per million
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QSAR
Quantitative structure activity relationship
RSD
Relative standard deviation
SI
Supplemental information
SIM
Selected ion monitoring
SPE
Solid phase extraction
STP
Sewage treatment plant
TMF
Trophic magnification factor
TOC
Total organic carbon
TOF
Time of flight
UPLC
Ultra-performance liquid chromatography
U.S.
United States
UV (UV-Vis)
Ultra-violet (visible) light
WW
Wet weight
WWTP
Wastewater treatment plant
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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: Juan Bezares-Cruz, Ryan Sullivan
Contributors: Andrew Middleton, Marcella Card
Technical Support: Mark Gibson, Hillary Hollinger
This report was reviewed and cleared by OPPT and OCSPP leadership.
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SUMMARY
This technical document is in support of the Draft Risk Evaluation for Diisononyl Phthalate (DINP)
(I ). DINP is a common chemical name for the category of chemical substances that
includes the following substances: 1,2-benzenedicarboxylic acid, 1,2-isononyl ester (CASRN 28553-12-
0) and 1,2-benzenedicarboxylic acid, di-C9-l 1-branched alkyl esters, C9-rich (CASRN 68515-48-0).
Both CASRNs contain mainly C9 dialkyl phthalate esters. See the draft risk evaluation for a complete
list of all the technical support documents for DINP.
In this document, EPA evaluated the reasonably available information to characterize the environmental
fate and transport of, the key points are summarized below. Given the consistent results from numerous
high-quality studies, there is robust evidence that DINP
• Is expected to undergo significant direct photolysis and will rapidly degrade in the atmosphere
(ti/2 = 8.5 hours) (Section 3.3).
• Is expected to degrade rapidly via direct and indirect photolysis (Section 3.3).
• Is not expected to appreciably hydrolyze under environmental conditions (Section 3.2).
• Is expected to have environmental biodegradation half-life in aerobic environments on the order
of days to weeks (Section 3.1).
• Is not expected to be subject to long range transport.
• Is expected to transform in the environment via biotic and abiotic processes to form
monoisononyl phthalate, isononanol, and phthalic acid (Section 3).
• Is expected to show strong affinity and sorption potential for organic carbon in soil and sediment
(Sections 5.2.2, 5.3.2).
• Will be removed at rates greater than 93 percent in conventional wastewater treatment systems
(Section 6.2).
• When released to air, will not likely exist in gaseous phase, but will show strong affinity for
adsorption to particulate matter (Sections 4 and 5).
• Is likely to be found in, and accumulate in, indoor dust (Section 5).
As a result of limited studies identified, there is moderate evidence that DINP
• Is not expected to biodegrade under anoxic conditions and may have high persistence in
anaerobic soils and sediments (Sections 3.1, 5.2.2, 5.3.2).
• Is not bioaccumulative in fish in the water column (Section 7).
• May be bioaccumulative in benthic organisms exposed to sediment with elevated concentrations
of DINP proximal to continual sources of release (Section 7).
• Is expected to be removed in conventional water treatment systems both in the treatment process,
and via reduction by chlorination and chlorination byproducts in post treatment storage and
drinking water conveyance (Section 6.3).
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139 1 INTRODUCTION
140 DINP is considered ubiquitous in various environmental media to due to its presence in both point and
141 non-point source discharges from industrial and conventional wastewater treatment effluents, biosolids
142 and sewage sludge, storm water runoff, and landfill leachate (Net et at.. 2015). As an isomeric mixture,
143 the fate and transport properties of DINP can be difficult to classify. However, the following sections of
144 the fate and transport analysis of DINP are present the general fate and transport characteristics of
145 DINP.
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2 APPROACH AND METHODOLOGY
Reasonably available environmental fate data—including biotic and abiotic biodegradation rates,
removal during wastewater treatment, volatilization from lakes and rivers, and organic carbon:water
partition coefficient (log Koc)—are the parameters used in the current draft risk evaluation. In assessing
the environmental fate and transport of DINP, EPA considered the full range of results from data
sources that were rated high-quality. Information on the full extracted dataset is available in the file
Draft Risk Evaluation for Diisononyl Phthalate (DINP) - Systematic Review Supplemental File: Data
Quality Evaluation and Data Extraction Information for Environmental Fate and Transport (U.S. EPA.
2024c). Other fate estimates were based on modeling results from EPI Suite™ (U.S. EPA. 2012). a
predictive tool for physical and chemical properties and environmental fate estimation.
Table 2-1 provides a summary of the selected environmental fate data that EPA considered while
assessing the fate of DINP and were updated after publication of Final Scope of the Risk Evaluation for
Di-isononylPhthalate (DINP) CASRN28553-12-0 68515—(8-0 (U.S. EPA. 2021) with additional
information identified through the systematic review process.
Table 2-1. Summary of DINP's Environmental Fate Information
Parameter
Value
Source(s)
Octanol: Water (Log Kow)
8.8
ECHA (2016)
Organic Carbon:Water
(Log Koc)
5.5 (estimated; MCI method); 5.7 (estimated; Kow
method
U.S. EPA (2017b)
Adsorption Coefficient
(Log Kd)
2.97 (suspended particulate matter/water)
Li et al. (2017a)
3.27 (sediment/water)
Li et al. (2017a)
Octanol:Air (Log Koa)
11.9 (estimated)
U.S. EPA (2017b)
AirWater (Log Kaw)
-2.20 (estimated)
Lu (2009)
-2.43 (estimated)
Cousins and Mackav (2000)
Aerobic primary
biodegradation in water
32 to 67.8% in 24 hours
>90% in 5 days
>99% in 28 days
(O'Gradv et al.. 1985; SRC.
1983; Monsanto. 1978)
Aerobic ready
biodegradation in water
57 to 81% in 28 days
(ECJRC. 2003b)
Aerobic ultimate
biodegradation in water
57 to 84% in 28 days
(HSDB. 2015; Monsanto.
1983)
Aerobic biodegradation in
sediment
0.54% in 14 days
1.11% in 28 days
(Johnson et al.. 1983)
Anaerobic biodegradation
in sediment
0% in 100 days
(Eilertsson et al.. 1996)
Aerobic biodegradation in
soil
No significant change in concentration after 2 years
(ECJRC. 2003b)
Hydrolysis
152 days at pH 8 and 25 °C, and 4.2 years at pH 7 and
25 °C
(U.S. EPA. 2017a)
Photolysis
ti/2 (air) = 5.36 to 8.5 hours
ti/2 (waterPH=7) = 140 days
(U.S. EPA. 2017a;
Lertsirisopon et al.. 2009;
Peterson and Staples. 2003)
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Parameter
Value
Source(s)
Environmental degradation
half-lives (selected values
for modeling)
5.36 hours (air)
10 days (water)
20 days (soil)
90 days (sediment)
(U.S. EPA. 2017a)
WWTP Removal
>93%
(U.S. EPA. 2017a)
Aquatic Bioconcentration
(BCF)
<3 L/kg wet weight (rainbow trout; Oncorhvnchiis
my kiss)
5.2 L/kg wet weight (upper trophic Arnot-Gobas
estimation)
(U.S. EPA. 2017a; EC/HC.
2015a)
Aquatic Bioaccumulation
(BAF)
68 (75 ug/kg wet weight in mussel from field study in
Seine estuary, France)
21 L/kg wet weight (upper trophic Arnot-Gobas
estimation)
(U.S. EPA. 2017a; ECJRC.
2003b)
Aquatic Food web
Magnification Factor
(FWMF)
0.46
(Experimental; 18 marine species)
(Mackintosh et al.. 2004)
Terr. Bioconcentration
(BCF)
0.01 -0.02
Experimental; earthworms (Eisenici fetida)
(ECJRC. 2003b)
Terr. Biota-sediment
accumulation factor
(BSAF)
0.018
OECD Test Guideline 207 (Eisenici fetida)
(EC/HC. 2015a)
2.1 EPI Suite™ Model Inputs and Settings
The approach described by (Mackav et al.. 1996) using the Level III Fugacity model in EPI Suite™
(LEV3EPI™) was used for this Tier II analysis. LEV3EPI is described as a steady-state, non-equilibrium
model that uses a chemical's physical and chemical properties and degradation rates to predict
partitioning of the chemical between environmental compartments and its persistence in a model
environment (U.S. EPA. 2017a). A Tier II analysis involves reviewing environmental release
information for DINP to determine whether further assessment is warranted for each environmental
medium. Environmental release data for DINP was not available from the Toxics Release Inventory
(TRI) or Discharge Monitoring Reports (DMRs); however, between 250 million and 550 million lb of
CASRN 28553-12-0 and between 100 million and 1,000 million lb of CASRN 68515-48-0 were
produced annually from 2016 to 2019 for use in commercial products, chemical substances or mixtures
sold to consumers, or at industrial sites according to production data from the Chemical Data Reporting
(CDR) 2020 reporting period. DINP is used as a plasticizer in polyvinyl chloride (PVC) and non-PVC
products (U.S. EPA. 2020; EC/HC. 2015a). DINP may be released to the environment during
production, distribution, processing in PVC and non-PVC polymers, use of products such as paints and
sealants, disposal or recycling, wastewater treatment, and disposal of solid and liquid waste (ECJRC.
2003b).
Environmental release information is also useful for fugacity modeling because the emission rates will
predict a real-time percent mass distribution for each environmental medium. Environmental
degradation half-lives were taken from high and medium quality studies that were identified through
systematic review to reduce levels of uncertainties. Based on DINP's observed and calculated
environmental half-lives, partitioning characteristics, and the results of Level III Fugacity modeling (see
Figure 4-1 below), DINP is expected to partition primarily to soil and sediment—regardless of the
compartment of the environmental release. The LEV3EPI™ results were consistent with environmental
monitoring data. Further discussion of DINP partitioning can be found in Section 4.
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189 The following inputs parameters were used for the Level III Fugacity model in EPI Suite™:
190
•
Melting point = -48.00 °C
191
•
Vapor pressure = 5.40><10~7 mm Hg
192
•
Water solubility = 6.10x 10~4 mg/mL
193
•
Log Kow= 8.8
194
•
SMILES: CCCCCCC(C)COC(=0)clccccclC(=0)OCCCCC(C)C(C)C (representative
195
structure)
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3 TRANSFORMATION PROCESSES
DINP released to the environment will transform to the monoester form (monoisononyl phthalate) via
abiotic processes such as photolysis (direct and indirect) and hydrolysis of the carboxylic acid ester
group 023). Biodegradation pathways for the phthalates consist of primary biodegradation
from phthalate diesters to phthalate monoesters, then to phthalic acid, and ultimately biodegradation of
phthalic acid to form CO: and/or CH4 (Huang et al.. 2013). The monoisononyl phthalate is both more
soluble and more bioavailable than DINP. It is also expected to undergo biodegradation more rapidly
than the diester form. EPA considered DINP transformation products and degradants qualitatively but
due to their lack of persistence we do not expect them to substantially contribute to risk; thus, EPA is not
considering them further in this draft risk evaluation. Both biotic and abiotic routes of degradation for
DINP are described in the sections below.
3.1 Biodegradation
DINP can be considered readily biodegradable under most aquatic and terrestrial environments. The
EPA extracted and evaluated fourteen data sources containing DINP biodegradation information in
water, soil, and sediments under aerobic and anaerobic conditions (Table 3-1). Eight of the sources were
classified as overall high-quality, five as overall medium-quality and one as overall low-quality data
sources. DINP is considered an isomeric mixture, and certain components of DINP might biodegrade
more readily than others (ECJRC. 2003b). DINP's aerobic primary biodegradation in water has reported
to be 32 to 67.8 percent in 24 hours (O'Gradv et al.. 1985; Monsanto. 1978). greater than 90 percent in 5
days (O'Gradv et al.. 1985). and greater than 99 percent in 28 days (1 c. ! V \ JO I'") with half-life of
1.5 to 5.3 1 days under acclimated conditions ( 4, 1983) and 7 to 40 days under unacclimated
conditions (EC/HC. 2015a). Several studies evaluating the readily biodegradability of phthalate esters in
water have reported DINP's half-life of 10.3 days (ExxonMobil. 2010) and 57 to 81 percent DINP
removal in 28 days by CO: evolution (11SDB. 20< , •! v Wa \>) The required 60 percent
degradation during the 10-day pass window was met only in two of the four available studies (ECJRC.
2003b). However, DINP in water has been reported to completely biodegrade into its basic elements by
57 to 84 percent after 28 days (based on the available ultimate biodegradation information) (EC/HC.
2015a; H.SDB. 2015; Monsanto. 1983; SRC. 1983). In contrast to the rapid biodegradation of DINP in
aerobic environments, available information suggests that DINP is expected to have very low
biodegradation potential under low oxygen conditions (Ejlertsson et al.. 1996) and could remain longer
in subsurface sediments and soils (Kickham et al k u4msonetai. 1984. 1983).
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228 Table 3-1. Summary of DINP's Biodegradation Information
Environmental
Conditions
Degradation Value
Half-Life
(days)
Reference
Overall Data
Quality Ranking
Aerobic primary
biodegradation in
water
32% in 24 hours
ND
(Monsanto. 1978)
High
94 to 96% in 9 days
1.5 days (average;
1-1.9 days)
(SRC. 1984)
High
67.8% in 24 hours
greater than 90% in 5 days
ND
(O'Gradv et al.. 1985)
High
greater than 99% in 28
days
5.31 days
(SRC. 1983)
High
91 to 100% in 7 days
7 to 40 days
(EC/HC. 2015a)
Medium
greater than 95% in 12
days
ND
(HSDB. 2015)
Medium
90 to 100% in 5 to 28 days
68% in 1 day
ND
(U.S. EPA. 2019)
Medium
Aerobic ready
biodegradation in
water
70.5% in 28 days
ND
(ECJRC. 2003b)
Medium
57% in 28 days
ND
81% in 28 days
ND
74% in 28 days
ND
(HSDB. 2015)
Medium
ND
10.3 days
(ExxonMobil, 2010)
Low
Aerobic ultimate
biodegradation in
water
61.5% in 28 days
ND
(SRC. 1983)
High
84% in 28 days
ND
(Monsanto. 1983)
Medium
57 to 71% in 28 days
ND
(HSDB. 2015)
Medium
56.6 % in 29 days
ND
(EC/HC. 2015a)
Medium
67.5% in 28 days
ND
74% in 28 days
ND
Aerobic
biodegradation in
sediment
0.54% in 14 days
1.11% in 28 days
ND
(Johnson et al.. 1983)
High
ND
12,000 days
(Kickham et al..
2012)
High
0.7% at 12 °C
1.2% at 22 °C
2.2% at 28 °C
ND
(Johnson et al.. 1984)
High
Anaerobic
biodegradation in
sediment
0% in 100 days
ND
(Eilertsson et al..
1996)
High
Aerobic
biodegradation in
soil
No significant change in
concentration after 2 years
ND
(ECJRC. 2003b)
Medium
229
230 3.2 Hydrolysis
231 Traditionally accepted methods of testing for abiotic hydrolysis (OECD Guideline Test 111) are not
232 viable for DINP due to the low aqueous solubility (ECJRC. 2003a). Therefore, hydrolysis rates of DINP
233 are difficult to accurate measure experimentally (ECJRC. 2003a). EPI Suite™ was utilized to estimate
234 the hydrolysis half-lives of DINP at 152 days at pH 8 and 25 °C, and 4.2 years at pH 7 and 25 °C (U.S.
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) indicating that hydrolysis is a possible degradation pathway of DINP under more caustic
conditions. Lertsirisopon (2009) reported the hydrolysis half-lives of 720 days (pH = 5), 1,200 days (pH
= 6), negligible (pH = 7), 1,000 days (pH = 8), and 460 days (pH = 9), at average temperature of 10.8
°C. However, this study received a low data quality ranking in the systematic review process due to
poorly documented and variable test conditions.
When compared to other degradation pathways, it is not expected that hydrolysis is a significant source
of degradation under typical environmental conditions. However, the higher temperatures, variations
from typical environmental pH, and chemical catalysts present in the deeper anoxic zones of landfills
may be favorable to the degradation of DINP via hydrolysis (Huang et ai. 2013). This is discussed
further in Section 5.3.3.
3.3 Photolysis
DINP contains chromophores that absorb light at greater than 290 nm wavelength ( . ),
therefore, direct photodegradation is a relevant degradation pathway for DINP released to air. Modelled
indirect photodegradation half-lives indicated a slightly more rapid rate of degradation, estimating a
half-life of 0.22 days (5.36 hours) ( OH rate constant of 2.39xl0~u cm3 /molecule-second and a 12-
hour day with 1.5x 106 OH/cm3) (U.S. EPA. ^ ). Similarly, Peterson (2003) reported a calculated
DINP photodegradation half-life of 0.35 days (8.5 hours) ( OH rate constant of 2,35/10 " cm3
/molecule-second and 1 x 106 OH/cm3). DINP photodegradation in water is expected to be slower than
air, due to the typical light attenuation in natural surface water. The aquatic direct photodegradation
half-lives of 32, 52, 140, 61 and 36 days were observed at pH 5, 6, 7, 8, and 9, respectively, when
exposed to natural sunlight in artificial river water at 0.4 to 27.4 °C (average temperature of 10.8 °C)
(Lertsirisopon et ai. 2009).
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4 PARTITIONING
Environmental release data for DINP was not available from the Toxics Release Inventory (TRI) or
Discharge Monitoring Reports (DMRs), therefore DINP releases to the environment could not be
estimated. The approach described by (Mackav et al.. 1996) using the Level III Fugacity model in EPI
Suite™ (LEV3EPI™) was used for this Tier II analysis. LEV3EPI is described as a steady-state, non-
equilibrium model that uses a chemical's physical and chemical properties and degradation rates to
predict partitioning of the chemical between environmental compartments and its persistence in a model
environment (U.S. EPA. 2017a). DINP's physical and chemical properties were taken directly from
Section 2.1 of Draft Physical Chemistry Assessment for Diisononyl Phthalate (DINP) (U.S. EPA.
2024a). Environmental release information is useful for fugacity modeling because the emission rates
will predict a real-time percent distribution for each medium. Environmental degradation half-live in
water of 10 days was selected in this draft risk evaluation to represent the range of identified primary
biodegradation half-life values (Section 3.1) from high and medium quality studies to reduce levels of
uncertainties. EPA used environmental degradation half-lives of 5.36 hours in air (based on
AEROWIN™ predicted values, an atmospheric fate prediction model within EPI Suite™), 20 days in
soil (double the half-life in water), and 90 days in sediment (9 times the half-life in water) as
recommended for EPIWIN estimations (U.S. EPA. 2017a). Based on DINP's environmental half-lives,
partitioning characteristics, and the results of Level III Fugacity modeling, DINP is expected to be found
predominantly in water, soil, and sediment (Figure 4-1). The LEV3EPI™ results were consistent with
environmental monitoring data. Further discussion of DINP partitioning can be found in Sections 5.1,
5.2, and 5.3.
100
90
BO
70
-t—' 60
1.
<
S 40
ra
30
20
10
0
100% Soi
_
1
¦
.
Release 100% Air Release 100% Water Release Equal Releases
Air BWaer BSoil ¦Sediment
Figure 4-1. EPI Suite™ Level III Fugacity Modeling Graphical Result for DINP
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5 MEDIA ASSESSMENTS
DINP has been reported to be present in the atmosphere, aquatic environments, and terrestrial
environments. Once in the air, DINP will be most predominant in the organic matter present in airborne
particles and expected to have a short half-life in the atmosphere. Based on the physical and chemical
properties, DINP is likely to partition to house dust and airborne particles and is expected to have a
longer half-life compared to ambient (outdoor) air. DINP present in surface water is expected to mostly
partition to aquatic sediments. DINP is expected to have an aerobic biodegradation half-life between 14
and 28 days. In terrestrial environments DINP has the potential to be present in soils and groundwater
but is likely to be immobile in both media types. In soils, DINP is expected to be deposited via air
deposition and land application of biosolids. DINP in soils is expected to have a half-life on the order of
days to weeks, have low bioaccumulation potential and biomagnification potential in terrestrial
organisms. DINP is released to groundwater via wastewater effluent and landfill leachates, expected to
have a half-life of 14 to 56 days, and not likely to be persistent in most groundwater/subsurface
environments.
5.1 Air and Atmosphere
DINP is a liquid at environmental temperatures with a melting point of-48 °C (Havnes. 2014; O'Neit.
2013) and a vapor pressure of 5.40 10 7 nimHg at 25 °C CNLM. 2015). Based on its physical and
chemical properties and short half-life in the atmosphere, ti/2 = 5.36 hours ( a), DINP was
assumed to not be persistent in the air. The AEROWIN™ module in EPI Suite™ estimated that a large
fraction of DINP could be sorbed to airborne particles and these particulates may be resistant to
atmospheric oxidation. DINP has not been detected in ambient air; however, studies have detected DINP
in settled house dust, indoor air samples and in indoor particulate phase air samples (NCBI. 2020;
Kubwabo et at.. I , H „oQ3b).
5.1,1 Indoor Air and Dust
In general, phthalate esters are ubiquitous in the atmosphere and indoor air. Their worldwide presence in
air has been documented in the gas phase, suspended particles and dust (Net et at.. 2015). Most of the
studies reported DEHP (di-ethylhexyl phthalate) to be the predominant phthalate esters in the
environment. Despite the limited information on the presence of DINP on the atmosphere, similar trends
to those reported for DEHP could be expected based on their similar vapor pressure (EC ).
Limited studies have reported the presence of particle bound DINP on indoor and outdoor settings
(Gupta and Gadi. 2018; Haseeawa. 2003; Hetmie et at.. 1990). Once in indoor air, DINP is expected to
partition to organic carbon present on indoor airborne particles. Under indoor environments, DINP is
expected to be more persistent in indoor air than in ambient (outdoor) air due to the lack of natural
chemical removal processes, such as solar photochemical degradation.
The available information suggests that the concentration of DINP in dust under indoor environments to
be higher than outdoors dust and to be associated with the presence of phthalate-containing articles and
the proximity to manufacturing facilities (Kubwabo et at.. 2013; Wane et at.. JO I <; et at.. 2009).
Kubwabo (2013) monitored the presence of 17 phthalate compounds in vacuum dust samples collected
in 126 urban single-family homes. The study reported that DEHP, DIDP, and DINP were detected in all
the collected dust samples accounting for 88 percent of the median total concentration of phthalates in
dust. Wang ( ) evaluated the presence of phthalates in dust samples collected from indoor and
outdoor settings in two major Chinese cities. The study reported the total phthalates concentration of the
collected indoor dust samples were 3.4 to 5.9 times higher than those collected outdoors. The aggregate
concentration of DEHP, DINP, and DIDP in indoor dust samples accounted for 91 to 94 percent of the
total phthalate's concentration. The study revealed that the aggregate concentration of phthalates was
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higher in the commercial and industrial areas with heavy production of textiles, costumes, and toys. Abb
(2009) evaluated the presence of phthalates in indoor dust samples collected from 30 households in
Germany. The study revealed the presence of DEHP, DIDP and DINP in all the collected samples.
Samples collected from households containing a high percentage of plastics (greater than 50 percent
plastic content) resulted with higher aggregate concentration of phthalates in dust. The aggregate
concentration of DEHP, DIDP and DINP accounted for 87 percent of the total phthalate concentration in
dust.
Similarly, recent studies monitoring the presence of phthalates in dust from USA households have
revealed DEHP and DINP to de detected in 96 to 100 percent of the collected samples (Ham m el et at..
2019; Dodson et al. 2017). Hammel ( ) and Dodson (2017) reported the presence of phthalate esters
on indoor air and dust samples collected in USA homes. Hammel ( ) reported that DINP accounted
for close to 83 percent of the total concentration of phthalates found in indoor dust. Dodson (2017)
evaluated the presence of phthalate esters in air samples of U.S. homes before and after occupancy
reporting increased presence of DINP after occupancy due to daily anthropogenic activities that might
introduce phthalate containing products into indoor settings. Increasing trends could be expected for
DINP with its increased use on household's construction materials or consumer products.
5.2 Aquatic Environments
5.2.1 Surface Water
DINP is expected to be released to surface water via industrial and municipal wastewater treatment plant
effluent, surface water runoff, and, to a lesser degree, atmospheric deposition. DINP and other phthalate
esters have been detected in surface waters, although at lower frequencies than some other phthalate
esters (Wen et al.. 2018). The principal properties governing the fate and transport of DINP in surface
water are water solubility, organic carbon partitioning coefficients, and volatility. Due to its Henry's
Law constant (9,14/10 5 atmm3/mol at 25 °C) of DINP, volatilization is not expected to be a significant
source of loss of DINP from surface water. A partitioning analysis of DINP released to the environment
is described in Section 4 above. The analysis estimates that during releases to surface water bodies,
greater than 92 percent of DINP released to surface water will partition to both suspended and benthic
sediments.
DINP has a low water solubility of 0.00061 mg/L, but is likely to form a colloidal suspension and may
be detected in surface water at higher concentrations (EC/HC. 2015b). Based on DINP's water solubility
and partitioning coefficients, DINP in water will partition to suspended organic material present in the
water column. DINP is expected to be readily biodegradable in water (Section 3.1). In addition, total
seawater samples concentrations of DINP measured in False Creek ranged from 61 to 135 ng/L; the
dissolved fraction concentrations ranged from 29 to 64 ng/L and the suspended particulate fraction
concentration ranged from 14,700 to 50,400 ng/g dry weight (dvv) (EC/HC. 2015a; Mackintosh et al..
2006). Concentrations of DINP above the aqueous solubility of 0.00061 mg/L are not uncommon in
monitoring studies proximal to releases of DINP to surface water (Wen et al.. 2018).
5.2.2 Sediments
Based on the water solubility (0.00061 mg/L) and affinity for sorption to organic matter (log Koc = 5.5
to 5.7), DINP will partition to the organic matter present in soils and sediment when released into the
environment. Once in water, DINP is expected to be readily biodegradable and the Level III Fugacity
Model in EPI Suite™ (U.S. EPA. ) predicts that greater than 92 percent of the DINP will partition
to and remain in sediments (Section 4). The available information suggests that DINP could persist
longer in subsurface sediments and soils than in water. In terrestrial and aquatic environments, DINP has
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potential to accumulate in sediments at areas of continuous release, such as a surface water body
receiving discharge from a municipal wastewater treatment plant.
Due to the strong sorption to organic carbon, DINP is expected to be found predominantly in sediments
near point sources, with a decreasing trend in sediment concentrations downstream. This is consistent
with monitoring information for phthalate esters from Sweden and Korea. One study reported the
presence of DINP in only one sediment sample near a point source in Sweden that recently have
replaced DEHP with DINP in their production processes (Parkinan and Remberg. 1995). The presence
of DINP has been documented in urban sediments at concentrations ranging between 130 to 3,200 |ig/kg
total solids with a 62 percent detection frequency (Cousins et ai. 2007). In a similar study, Kim (2021)
evaluated the presence of plasticizers in sediments from highly industrialized bays of Korea. DINP was
detected in all surface sediment samples. The study revealed a gradual decreasing trend in the overall
concentration of phthalates toward the outer region of the bays farther away from industrial activities.
The findings of this study suggest industrial activities to be the major contributor of phthalates in
sediments within the area.
Monitoring data from the Rhine River and the Neckar River in Germany detected DINP concentrations
in sediment samples of 30, 220, 650, and 1,460 ppb and 430, 570, 1,050 ppb (3 sites), respectively
(NCBI. 2020). DINP was also detected in sediment from 21 locations in the Netherlands at
concentrations up to 6.16 mg/kg dry weight (ECJRC. 2003b).
5.3 Terrestrial Environments
5.3.1 Soil
DINP is expected to be deposited to soil via two primary routes: application of biosolids and sewage
sludge in agricultural applications or sludge drying applications as well as atmospheric deposition.
Based on DINP's Henry's Law constant of 9.14/ 10 5 atnrmVmol at 25 °C and vapor pressure of
5.40xl0~7mmHg, DINP is not likely to volatilize from soils.
DINP shows an affinity for sorption to soil and its organic constituents (log Koc = 5.5-5.7; log Kd of
2.55-3.27 (Li et aL 2017b: Li etai. I * « « \ ;01:) and an estimated log Kow of 10.21 (U.S.
) Given that these properties indicate the likelihood of strong sorption to organic carbon
present in soil, DINP is expected to have low mobility in soil environments.
Under aerobic conditions, DINP is expected to have a half-life in soil of 20 days. This aerobic
biodegradation half-life for soil was estimated by doubling the experimentally derived half-life of DINP
in water as very limited soil biodegradation data for DINP identified in the systematic review process
(SRC. 1983).
Under anaerobic conditions that may be present in some soil profiles, there is very little evidence to
support that DINP appreciably biodegrades (ECJRC. 2003b: Eilertsson etai. 1996). One study found
that 0 percent degradation had occurred under anaerobic conditions after 100 days by CH4 evolution and
no transformation reported based on the concentrations of methane and test substance with gas
chromatographic analysis in municipal solid waste samples with an anaerobic microflora inoculum
(Eilertsson etai. 1996). Furthermore, another study reported less than 1 percent DINP degradation in
anaerobic sediments after 28 days (Johnson et ai. 1984).
In general, DINP is not expected to be persistent in soil as long as the rate of release does not exceed the
rate at which biodegradation can occur, but continuous exposure to DINP in soil proximal to points of
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releases may be possible if the rate of releases exceeds the rate of biodegradation under aerobic
conditions. Under anaerobic conditions in soil, DINP is assumed to be persistent, and continuous
exposure is likely.
5.3.2 Biosolids
Sludge is defined as the solid, semi-solid, or liquid residue generated by wastewater treatment processes.
The term "biosolids" refers to treated sludge that meet the EPA pollutant and pathogen requirements for
land application and surface disposal and can be beneficially recycled (40 CFR Part 503) (
1993). Typically, chemical substances with very low water solubility and high sorption potential are
expected to be sorbed to suspended solids and efficiently removed from wastewater via accumulation in
sewage sludge and biosolids.
There is limited information about the presence and biodegradation of DINP in biosolids. As described
in Section 6.2, DINP in wastewater has been reported to be mainly removed by particle sorption and
retained in the sewage sludge. In general, greater than 93 percent of the DINP present in wastewater is
expected to be accumulated in sewage sludge and discharged into biosolids. Once in biosolids, DINP
could be transferred to soil during land applications. DINP will be strongly sorbed to organic matter on
soils and to be more persistent in soil profiles with anaerobic conditions (ECJRC. 2003b). Due to its
strong sorption to soils, land applied DINP is not expected to be bioavailable; thus, exposures to
environmental organisms and people are negligible. In addition, based on the water solubility and
hydrophobicity of DINP, will have low bioaccumulation, biomagnification appear to be of minimal
concern. Additionally, terrestrial species have been reported to have the capacity to metabolize phthalate
substances (Bradlee and Thomas. 2003; Gob as et al. 200.'; Kjrron et ai. 1995) and DINP is expected to
have low bioaccumulation potential and biomagnification potential in terrestrial organisms (Section 7).
5.3.3 Landfills
For the purpose of this assessment, landfills will be considered to be divided into two zones: (1) an
"upper-landfill" zone, with normal environmental temperatures and pressures, where biotic processes
are the predominant route of degradation for DINP; and (2) a "lower-landfill" zone where elevated
temperatures and pressures exist, and abiotic degradation is the predominant route of degradation for
DINP. In the upper-landfill zone where oxygen may still be present in the subsurface, conditions may
still be favorable for aerobic biodegradation; however, photolysis and hydrolysis are not considered to
be significant sources of degradation in this zone. In the lower-landfill zone, conditions are assumed to
be anoxic and temperatures present in this zone are likely to inhibit biotic degradation of DINP.
Temperatures in lower-landfills may be as high as 70 °C. At temperatures at and above 60 °C, biotic
processes are significantly inhibited, and are likely to be completely irrelevant at 70 °C (Huang et al..
2013).
DINP is deposited in landfills continually and in high amounts from the disposal of consumer products
containing DINP. However due to its strong sorption to soils and low water solubility, small
concentration of DINP is likely to be present in landfill leachate. DINP is likely to be persistent in
landfills due to the apparent lack of anaerobic biodegradation and unfavorable conditions for
biodegradation in lower-landfills. Some aerobic biodegradation in upper-landfills may occur. In lower-
landfills, there is some evidence to support that hydrolysis may be the main route of abiotic degradation
of phthalate esters (Huang et al.. 2013).
Despite the expected persistence of DINP in landfills, it is not expected to be bioavailable and mainly
sorbed to organic matter in soils due to the low water solubility of DINP (0.00061 mg/L), and its high
sorption to organic carbon (log Koc = 5.5-5.7). Although DINP may be present at small concentrations
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in landfill leachate, it is unlikely to migrate to, or be mobile in groundwater proximal to landfills, and
would not be expected to be transported distally from landfills via groundwater.
5,3,4 Groundwater
There are several potential sources of DINP in groundwater, including wastewater effluents and landfill
leachates, which are discussed in Sections 5.3.3 and 6.2. Further, in environments where DINP is found
in surface water, it can enter groundwater through surface water/groundwater interactions. Diffuse
sources include storm water runoff and runoff from biosolids applied to agricultural land.
Given the strong affinity of DINP to adsorb to organic matter present in soils and sediments (log Koc =
5.5-5.7) ( ) DINP is expected to have low mobility in soil and groundwater
environments. Furthermore, due to the insoluble nature of DINP (0.00061 mg/L), high concentrations of
DINP in groundwater are unlikely. In instances where DINP could reasonably be expected to be present
in groundwater environments (proximal to landfills or agricultural land with a history of land applied
biosolids), limited persistence is expected based on rates of biodegradation of DINP in aerobic
environments, DINP is not likely to be persistent in groundwater/subsurface environments unless anoxic
conditions exist.
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6 PERSISTENCE POTENTIAL OF DINP
DINP is not expected to be persistent in the environment, as it is expected to degrade rapidly under most
environmental conditions, with delayed biodegradation in low-oxygen media. In the atmosphere, DINP
is unlikely to remain for long periods of time as it is expected to undergo photolytic degradation through
reaction with atmospheric hydroxyl radicals, with estimated half-lives of 5.36 hours. DINP is predicted
to hydrolyze slowly at ambient temperature, but it's not expected to persist in aquatic media as it
undergoes rapid aerobic biodegradation (Section 5.2.1). DINP have the potential to remain for longer
periods of time in soil and sediments, but due to the inherent hydrophobicity (log Kow = 8.8) and
sorption potential (log Koc = 5.5-5.7) DINP is not expected to be bioavailable for uptake. Using the
Level III Fugacity model in EPI Suite™ (LEV3EPITM) (Section 4), DINP's overall environmental half-
life was estimated to be approximately 34 days (U. c. < \ I J). Therefore, DINP is not expected to be
persistent in the atmosphere or aquatic and terrestrial environments.
6.1 Destruction and Removal Efficiency
Destruction and removal efficiency (DRE) is a percentage that represents the mass of a pollutant
removed or destroyed in a thermal incinerator relative to the mass that entered the system. DINP is
classified as a hazardous substance and EPA requires that hazardous waste incineration systems destroy
and remove at least 99.99 percent of each harmful chemical in the waste, including treated hazardous
waste (46 FR 7684) ( ?81).
Currently there is no information available on the DRE of DINP. However, the DEHP annual releases
from a Danish waste incineration facility were estimated to be 9 percent to air and 91 percent to
municipal land fill (ECB. 2008). These results suggest that DINP present during incineration processes
will be very likely to be released to landfills and the remaining small fraction released to air. Based on
its hydrophobicity and sorption potential, DINP released to landfills is expected to partition to waste
organic matter. Similarly, DINP released to air is expected to be rapidly react via indirect photochemical
processes within hours (U.S. EPA. 2017a) and partition to soil and sediments as described in Section 4.
DINP in sediments and soils is not expected to be bioavailable for uptake, and highly biodegradable in
its bioavailable form (Kickham et ai. ).
6.2 Removal in Wastewater Treatment
Wastewater treatment is performed to remove contaminants from wastewater using physical, biological,
and chemical processes. Generally, municipal wastewater treatment facilities apply primary and
secondary treatments. During the primary treatment, screens, grit chambers, and settling tanks are used
to remove solids from wastewater. After undergoing primary treatment, the wastewater undergoes a
secondary treatment. Secondary treatment processes can remove up to 90 percent of the organic matter
in wastewater using biological treatment processes such as trickling filters or activated sludge.
Sometimes an additional stage of treatment such as tertiary treatment is utilized to further clean water
for additional protection using advanced treatment techniques (e.g., ozonation, chlorination,
disinfection).
Limited information is available in the fate and transport of DINP in wastewater treatment systems. The
EPA selected two high quality sources reporting the removal of DINP in wastewater treatment systems
employing aerobic and anaerobic processes. One study reported 98.0 percent DINP removal efficiencies
in a municipal wastewater treatment facility in France, employing a combined decantation and activated
sludge tank (Tram et ai. 2014). Like other phthalates esters with long carbon chains and high log Kow,
DINP was reported to be mainly removed by particle sorption and retained in the sewage sludge. This
finding is supported by STPWIN™, an EPI Suite™ module that estimates chemical removal in sewage
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treatment plants. The model predicts greater than 93 percent removal of DINP during conventional
wastewater treatment by sorption to sludge with the potential of increased removal via rapid aerobic
biodegradation processes (U.S. EPA. ). In addition, the treatment of wastewater final solids via
aerobic digestions processes have been reported to achieve 41.1 and 85.9 percent reduction on DINP
concentration from the digestion effluents (Armstrong et al. 2018). In addition, the same study reported
anaerobic solids digestion to be not effective in the removal of DINP. In general, the available
information suggest that aerobic processes have the potential to help biodegrade DINP from wastewater
in agreement with the expected aerobic biodegradation described in Section 3.1.
Overall, DINP has a high log Kow and remains in suspended solids and efficiently removed from
wastewater via accumulation in sewage sludge (Tran et al.. 2014). partially removed during aerobic
solids digestion processes (Armstrong et al.. 2018). and ineffectively removed under anaerobic solids
digestion conditions (Armstrong et al.. 2018). Biodegradation and air stripping are not expected to be
significant wastewater removal processes. Therefore, greater than 93 percent of the DINP present in
wastewater is expected to be accumulated in sewage sludge and released with biosolids disposal or
application, with the remaining fraction sorbed to suspended solids in the wastewater treatment effluent
and discharged with surface water (Tran et al.. 2014; U.S. EPA. 2012).
6.3 Removal in Drinking Water Treatment
Drinking water in the United States typically comes from surface water {i.e., lakes, rivers, reservoirs)
and groundwater. The source water then flows to a treatment plant where it undergoes a series of water
treatment steps before being dispersed to homes and communities. In the U.S., public water systems
often use conventional treatment processes that include coagulation, flocculation, sedimentation,
filtration, and disinfection, as required by law.
Very limited information is available on the removal of DINP in drinking water treatment plants. No
data was identified by the EPA for DINP in drinking water. Based on the water solubility and Log Kow,
DINP in water it is expected to mainly partition to suspended solids present in water. This is supported
by the Level III Fugacity model in EPI Suite™ (Section 4) which predicts 92.7 percent of DINP released
to water partitioning to sediments ( ). The available information on the DEHP removal
efficiency of flocculants and filtering media, suggest that DINP could potentially be partially removed
during drinking water treatment by sorption into suspended organic matter. This data source reported
58.7 percent reduction on drinking water DEHP concentration from a conventional drinking water
treatment effluent in China and 78 to 86 percent loss of DINP during storage of treated drinking water
effluent after 48 hours in Taiwan using chlorine for disinfection prior to distribution (Kong et al.. 2017;
Yang et al.. 2014). These findings suggest that conventional drinking water treatment systems may have
the potential to partially remove DINP is present in drinking water sources via sorption to suspended
organic matter and filtering media and the use of disinfection technologies.
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7 BIOACCUMULATION POTENTIAL OF DINP
The presence of DINP in several marine aquatic species in North America suggest that the substance to
be bioavailable in aquatic environments (Mackintosh et al. 2004). However, based on the very low
water solubility and high hydrophobicity, DINP is expected to have low bioaccumulation potential,
biomagnification potential, and low potential for uptake. EPA selected three overall high quality data
sources and two overall medium quality data source reporting the aquatic bioconcentration, aquatic
bioaccumulation, aquatic food web magnification, terrestrial biota-sediment accumulation, and
terrestrial bioconcentration of DINP (Table 7-1). The available data sources discussed below, suggest
that DINP has low bioaccumulation potential in aquatic and terrestrial organisms (EC/HC. 2015a;
Solbakken et al.. 1985; Chemical Manufacturers. 1984). and no apparent biomagnification across trophic
levels in the aquatic food web (Mackintosh et al.. 2004).
Several studies have investigated the aquatic bioconcentration and food web magnification of DINP in
several marine species. Solbakken (1985) evaluated the bioconcentration of DINP in Area zebra in a 24-
hour exposure study, followed by a 14-day depuration period. The study reported DINP BCF values of
8.2, 183.8, 13.6, and 9.3 dpm/|iL during the 24-hour exposure period on Area zebra muscle,
hepatopancreas, gills, and blood respectively (Table 7-1). The study reported a 92 to greater than 99
percent decrease on BCF values during the 14-days depuration period. A similar study evaluating the
presence of phthalates on estuaries reported a mussel BAF of 68 and DINP content of 75 |ig/kg wet
weight (ECJRC. 2003b). A DINP exposure studies on rainbow trout have reported BCF lower than 3
L/kg wet weight and biomagnification factor lower than 0.1 (EC/HC. 2015a). The reported low BCF
values suggest that DINP has low potential to bioaccumulate in aquatic organisms. In the other hand, the
Chemical Manufacturers Association (1984) reported a higher predicted DINP aquatic BCF of 1,155
using a regression model based on the substance water solubility. Despite of the different range of
reported BCF values, the empirical rapid BCF decrease during a 14-day depuration period (Solbakken et
al.. 1985). the empirical aquatic TMF of 0.46 (Mackintosh et al.. 2004). modeled upper trophic BCF of
5.2 L/kg wet-wt and upper trophic BAF of 21 L/kg wet weight (1 c. I ^ \ JO I j) helps support that
DINP will have low bioconcentration potential and low biomagnification potential across trophic levels
in the aquatic food web.
There is very limited information on the bioconcentration and bioaccumulation of DINP in terrestrial
environments. Based on DINP's strong sorption organic matter (log Koc 5.5-5.7) ( 017a)
and water solubility (0.00061 mg/L) (Letinski et al.. 2002). DINP is not expected to be bioavailable in
soils. This is supported by the reported low BCF values of 0.1 to 0.2 on earthworms (Eisenia foetida)
(ECJRC. 2003b). Therefore, DINP is expected to have low bioaccumulation potential, biomagnification
potential in terrestrial organisms.
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604 Table 7-1. Summary of DINP's Bioaccumulation Information
Endpoint
Value
Details
Reference
Overall
Quality
Ranking
8.2 at day 0
4.6 at day 1
1.3 at day 4
0.03-0.01 at day 14
(dpm/mg)
Experimental; Muscle BCF; 14-C
DINP 24 hours exposure followed
by 14-day depuration period; Area
Zebra (mollusk)
183.8 at day 0
125.2 at day 1
64.5 at day 4
14.4 at day 14
(dpm/mg)
Experimental; Hepatopancreas
BCF; 14-C DINP 24 hours
exposure followed by 14-day
depuration period; Area Zebra
(mollusk)
13.6 at day 0
12.4 at day 1
6.5 at day 4
0.8 at day 14
(dpm/mg)
Experimental; Gills BCF; 14-C
DINP 24 hours exposure followed
by 14-day depuration period; Area
Zebra (mollusk)
(Solbakken et al..
1985)
High
Aquatic
Bioconcentration
(BCF)
9.3 at day 0
5.6 at day 1
4.4 at day 4
0.1 at day 14
(dpm/|iL)
Experimental; Blood BCF; 14-C
DINP 24 hours exposure followed
by 14-day depuration period; Area
Zebra (mollusk)
0.46 at day 0
0.45 at day 1
0.26 at day 4
0.13 at day 14
(dpm/mg)
Experimental; BCF; 14-C DINP 24
hours exposure followed by 14-day
depuration period; Diploria
Strigosa (coral)
1155
Predicted; log BCF=(0.542 x log
Kow)+0.124; calculated using Kow
values that were calculated from
water solubility; log Kow = 5.2-
0.68 x log (micromolar WS).
(Chemical
Manufacturers.
1984)
High
<3 L/kg wet
weight
experimental; rainbow trout;
Oncorhvnchiis my kiss: elimination
rate: 1.16/day; tissue elimination
half-life: <1 day; BMF: <0.1
(EC/HC. 2015a)
Medium
Aquatic
Bioaccumulation
(BAF)
68
Experimental; preliminary study;
Field study; Mussel; Collected from
Seine estuary, France; 75 ug/kg wet
weight in mussel from field study in
Seine estuary, France
(ECJRC. 2003b)
Medium
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Endpoint
Value
Details
Reference
Overall
Quality
Ranking
Aquatic Food
web
Magnification
Factor (FWMF)
0.46
experimental; 18 marine species,
representing four trophic levels;
trophic dilution, predominantly
absorbed via the diet and depurated
at a rate greater than the passive
elimination rate via fecal egestion
and respiratory ventilation, due to
metabolism; FWMF (food web
magnification factor) = 0.44;
(Mackintosh et
ah. 2004)
High
Terr.
Bioconcentration
(BCF)
0.01-0.02
Terrestrial BCF; Experimental;
earthworms (Eisenia fetida); steady
state may not have been achieved.;
14 days
(ECJRC. 2003 b)
Medium
Terr. Biota-
sediment
accumulation
factor (BSAF)
0.018
experimental; other: OECD Test
Guideline 207 (Earthworm, acute
toxicity; OECD 1984a); earthworm;
Eisenia fetida;
CEC/HC. 2015a)
Medium
605
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8 OVERALL FATE AND TRANSPORT OF DINP
The inherent physical chemical properties of DINP govern its environmental fate and transport. Based
on DINP's aqueous solubility, slight tendency to volatilize, and strong tendency to adsorb to organic
carbon this chemical substance will be preferentially sorbed into sediments, soils, and suspended solids
in wastewater treatment processes. Soil, sediment, and sludge/biosolids are predicted to be the major
receiving compartments for DINP as indicated by its physicochemical and fate properties, partitioning
analysis, and verified by monitoring studies. Surface water is predicted to be a minor pathway, and the
main receiving compartment for phthalates discharged via wastewater treatment processes. However,
phthalates in surface water will sorb strongly to suspended and benthic sediments. In area where
continuous releases of phthalates occur, higher levels of phthalates in surface water can be expected,
trending downward distally from the point of releases. This also holds true for DINP concentration in
both suspended and benthic sediments. While DINP undergoes relatively rapid aerobic biodegradation
but it is persistent in anoxic/anaerobic environments (sediment, landfills) and like other phthalates it is
expected to slowly hydrolyze under normal environmental conditions.
If released directly to the atmosphere, DINP is expected to adsorb to particulate matter. It is not
expected to undergo long-range transport facilitated by particulate matter due to the relatively rapid rates
of both direct and indirect photolysis. Atmospheric concentrations of DINP may be elevated proximal to
sites of releases. Off gassing from landfills and volatilization from wastewater treatment processes are
expected to be negligible releases in terms of ecological or human exposure in the environment due to
its low vapor pressure. DINP released to air may undergo rapid photodegradation and it is not expected
to be a candidate chemical for long range transport.
Under indoor settings, air released DINP is expected to partition to airborne particles at concentrations
three times higher than in vapor phase (ECIRC. 2003a) and is expected to have extended lifetime as
compared to outdoor settings. The available information suggests that DINP's indoor dust
concentrations to be associated with the presence of phthalate containing articles and the proximity to
the facilities producing them (Kubwabo et al.. 2013; Wane et at.. 2013; Abb et at.. 2009) as well as daily
anthropogenic activities that might introduce DINP containing products into indoor settings fDodson et
at.. 2017).
DINP has a predicted average environmental half-life of 35 days. In situations where aerobic conditions
are predominant, DINP is expected to degrade rapidly and be more persistent under anoxic/anaerobic
conditions. In some sediments, landfills, and soils, DINP may be persistent as it is resistant to anaerobic
biodegradation. In anerobic environments, such as deep landfill zones, hydrolysis is expected the most
prevalent process for the degradation of DINP.
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9 WEIGHT OF SCIENTIFIC EVIDENCE CONCLUSIONS FOR
FATE AND TRANSPORT
9.1 Strengths, Limitations, Assumptions, and Key Sources of Uncertainty
for the Fate and Transport Assessment
Given the consistent results from numerous high-quality studies, there is a robust confidence that DINP
• is expected to undergo significant direct photolysis (Section 3.3);
• will partition to organic carbon and particulate matter in air (Sections 4);
• will biodegrade in aerobic surface water, soil, and wastewater treatment processes (Sections
5.2.1, 5.3.1, and 6.2);
• does not biodegrade in anaerobic environments (Section 5.2 and 5.3);
• will be removed after undergoing wastewater treatment and will sorb to sludge at high fractions,
with a small fraction being present in effluent (Section 6.2);
• is not bioaccumulative (Section 7);
• is not expected to biodegrade under anoxic conditions and may have high persistence in
anaerobic soils and sediments (Sections 3.1, 5.2.2, and 5.3.2);
• may show persistence in surface water and sediment proximal to continuous points of release
(Sections 3.1, 5.2.2, and 5.3.2); and
• is expected to transform to monoisononyl phthalate, isononanol, and phthalic acid in the
environment (Section 3).
As a result of limited studies identified, there is a moderate confidence that DINP
• is expected to be removed in conventional drinking water treatment systems both in the treatment
process, and via reduction by chlorination and chlorination byproducts in post treatment storage
and drinking water conveyance (Section 6.3); and
• showed no significant degradation via hydrolysis under standard environmental conditions but
hydrolysis rate was seen to increase with increasing pH and temperature in deep-landfill
environments (Section 5.3.3).
Findings that were found to have a robust weight of evidence supporting them had one or more high-
quality studies that were largely in agreement with each other, findings that were said to have a
moderate weight of evidence were based on a mix of high and medium-quality studies that were largely
in agreement but varied in sample size and consistence of findings.
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