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EPA Document# EPA-740-D-24-025
December 2024
Office of Chemical Safety and
Pollution Prevention
xvEPA
United States
Environmental Protection Agency
Draft Physical Chemistry and Fate and Transport Assessment
for Dibutyl Phthalate (DBP)
Technical Support Document for the Draft Risk Evaluation
CASRN 84-74-2
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28 TABLE OF CONTENTS
29 ACKNOWLEDGEMENTS 6
30 SUMMARY 7
31 1 INTRODUCTION 9
32 2 APPROACH AND METHODOLOGY FOR PHYSICAL AND CHEMICAL
33 PROPERTY ASSESSMENT 10
34 2.1 Selected Physical and Chemical Property Values for DBP 10
35 2.2 Endpoint Assessments 11
36 2.2.1 Melting Point 11
37 2.2.2 Boiling Point 11
38 2.2.3 Density 11
39 2.2.4 Vapor Pressure 11
40 2.2.5 Vapor Density 12
41 2.2.6 Water Solubility 12
42 2.2.7 Octanol:Air Partition Coefficient (log Koa) 12
43 2.2.8 Octanol:Water Partition Coefficient (log Kow) 12
44 2.2.9 Henry's Law Constant 13
45 2.2.10 Flash Point 13
46 2.2.11 Autoflammability 13
47 2.2.12 Viscosity 13
48 2.3 Strengths, Limitations, Assumptions, and Key Sources of Uncertainty for the Physical and
49 Chemical Property Assessment 14
50 3 APPROACH AND METHODOLOGY FOR FATE AND TRANSPORT ASSESSMENT 15
51 3.1 Tier I Analysis 18
52 3.1.1 Soil, Sediment, andBiosolids 18
53 3.1.2 Air 18
54 3.1.3 Water 18
55 3.2 Tier II Analysis 18
56 3.2.1 Fugacity Modeling 19
57 4 TRANSFORMATION PROCESSES 22
58 4.1 Biodegradation 22
59 4.1.1 Aerobic Biodegradation in Water 22
60 4.1.2 Biodegradation in Sediment 23
61 4.1.3 Biodegradation in Soil 23
62 4.2 Hydrolysis 25
63 4.3 Photolysis 26
64 5 MEDIA ASSESSMENTS 27
65 5.1 Air and Atmosphere 27
66 5.1.1 Indoor Air and Dust 27
67 5.2 Aquatic Environments 28
68 5.2.1 Surface Water 28
69 5.2.2 Sediments 29
70 5.3 Terrestrial Environments 29
71 5.3.1 Soil 29
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72 5.3.2 Biosolids 30
73 5.3.3 Landfills 31
74 5.3.4 Groundwater 31
75 6 REMOVAL AND PERSISTENCE POTENTIAL OF DBP 32
76 6.1 Destruction and Removal Efficiency 32
77 6.2 Removal in Wastewater Treatment 32
78 6.3 Removal in Drinking Water Treatment 34
79 7 BIO ACCUMULATION POTENTIAL OF DBP 36
80 8 OVERALL FATE AND TRANSPORT OF DBP 41
81 9 Weight of the Scientific Evidence Conclusions for Fate and Transport 42
82 9.1 Strengths, Limitations, Assumptions, and Key Sources of Uncertainty for the Fate and
83 Transport Assessment 42
84 REFERENCES 43
85
86 List of Tables
87 Table 2-1. Selected Physical and Chemical Property Values for DBP 10
88 Table 3-1. Summary of DBP's Environmental Fate Information 15
89 Table 3-2. Summary of Key Environmental Pathways & Media Specific Evaluations 19
90 Table 3-3. DBP Half-Life Inputs Used in EPI Suite™ Level III Fugacity Modeling 20
91 Table 4-1. Summary of DBP Biodegradation Information 24
92 Table 6-1. Summary of DBP's WWTP Removal Information 34
93 Table 7-1. Summary of DBP's Bioaccumulation Information 37
94
95 List of Figures
96 Figure 3-1. EPI Suite™ Level III Fugacity Modeling for DBP 21
97
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KEY ABBREVIATIONS AND ACRONYMS
AT SDR
Agency for Toxic Substances and Disease Registry
Atm
Atmospheres
atmmVmol
Atmospheres - cubic meters per mole
BAF
Bioaccumulation factor
BCF
Bioconcentration factor
BMF
Biomagnification factor
BSAF
Biota-sediment accumulation factor
C
Celsius
CASRN
Chemical Abstract Service registry number
CP
Centipoise
DBP
Dibutyl phthalate
DOE
Department Of Energy
DOC
Dissolved organic carbon
dw
Dry weight
DW
Drinking water
ECHA
European Chemicals Agency
EC/HC
Environment Canada and Health Canada
EPA
Environmental Protection Agency
F
Fahrenheit (°F)
g/cm3
Grams per cubic centimeter
GC
Gas chromatography
HLC
Henry's Law constant
K
Kelvin
Kaw
Air-water partition coefficient
Koa
Octanol-air partition coefficient
Koc
Organic carbon-water partition coefficient
Kow
Octanol-water partition coefficient
M
Molarity (mol/L = moles per Liter)
mg/L
Milligrams per liter
mL/min
Milliliters per minute
mmHg
Millimeters of mercury
MBP
Mono butyl phthalate
mol
Mole
MS
Mass spectrometry
N/A
Not applicable
NCBI
National Center for Biotechnology Information
NIST
National Institute of Standards and Technology
NIOSH
National Institute for Occupational Safety and Heal
NLM
National Library of Medicine
nm
Nanometers
NR
Not reported
OH
Hydroxyl radical
Pa (hPa)
Pascals (hectopascals; 1 hPa =100 Pa)
PA
Phthalic Acid
PCF
Plant concentration factor
Pg/L
Picograms per liter
ppm
parts per million
QSAR
Quantitative structure activity relationship
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RSC
Royal Society of Chemistry
RSD
Relative standard deviation
SI
Supplemental information
STP
Sewage treatment plant
TSCA
Toxic Substances Control Act
TMF
Trophic magnification factor
U.S.
United States
UV (UV-Vis) Ultraviolet (visible) light
WHO
World Health Organization
WW
Wet weight
WWTP
Wastewater Treatment Plant
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ACKNOWLEDGEMENTS
This report was developed by the United States Environmental Protection Agency (U.S. EPA or the
Agency), Office of Chemical Safety and Pollution Prevention (OCSPP), Office of Pollution Prevention
and Toxics (OPPT).
Acknowledgements
The Assessment Team gratefully acknowledges the participation, review, and input from EPA OPPT
and OSCPP senior managers and science advisors. The Agency is also grateful for assistance from the
following EPA contractors for the preparation of this draft technical support document: ICF (Contract
Nos. 68HERC19D000, 68HERD22A0001, and 68HERC23D0007), and SRC, Inc. (Contract No.
68HERH19D0022).
As part of an intra-agency review, this technical support document was provided to multiple EPA
Program Offices for review. Comments were submitted by EPA's Office of Research and Development
(ORD).
Docket
Supporting information can be found in the public docket, Docket ID EPA-HQ-QPPT-2018-0503.
Disclaimer
Reference herein to any specific commercial products, process or service by trade name, trademark,
manufacturer, or otherwise does not constitute or imply its endorsement, recommendation, or favoring
by the United States Government.
Authors: Collin Beachum (Management Lead), Mark Myer (Assessment Lead), Jennifer Brennan
(Assessment Lead), Ryan Sullivan (Physical Chemistry and Fate Assessment Discipline Lead),
Aderonke Adegbule, Andrew Middleton, Juan Bezares-Cruz (Physical Chemistry and Fate Assessors)
Contributors: Marcella Card, Maggie Clark, Daniel DePasquale, Patricia Fontenot, Lauren Gates,
Grant Goedjen, Roger Kim, Jason Wight
Technical Support: Hillary Hollinger, S. Xiah Kragie
This draft technical support document was reviewed and cleared for release by OPPT and OCSPP
leadership.
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SUMMARY
This technical support document is in support of the TSCA Draft Risk Evaluation for Dibutyl Phthalate
(DBP) (U.S. EPA. 2024c). EPA gathered and evaluated physical and chemical property data and
information according to the process described in the Draft Risk Evaluation for Dibutyl Phthalate (DBP)
- Systematic Review Protocol (U.S. EPA. 2024d). During the evaluation of dibutyl phthalate (DBP),
EPA considered both measured and estimated physical and chemical property data and information
summarized in Table 2-1, as applicable. Information on the full, extracted data set is available in the file
Draft Risk Evaluation for Dibutyl Phthalate (DBP) - Systematic Review Supplemental File: Data
Quality Evaluation and Data Extraction Information for Physical and Chemical Properties (U.S. EPA.
2024b).
DBP - Physical Chemistry: Key Points
• DBP is a branched phthalate ester used as a plasticizer.
• Under standard environmental conditions, DBP is an oily liquid (O'Neil. 2013) with a melting
point around -35 °C (Rumble. 2018b).
• DBP has a water solubility of 11.2 mg/L at 24°C (Howard et al.. 1985) and a log Kow of 4.5
(NLM. 2024).
• With a vapor pressure of 2.1 x 10~5 mmHg at 25 °C (U.S. EPA. 2019) and a boiling point of 340
°C (Rumble. 2018b). DBP has the potential to be volatile from dry, non-adsorbing surfaces.
• The selected Henry's Law constant for DBP is 1.81 x 10~6 atmm3/mol at 25 °C (NLM. 2024).
DBP - Environmental Fate and Transport: Key Points
EPA evaluated the reasonably available information to characterize the environmental fate and transport
of DBP, the key points are summarized below. Given the consistent results from numerous high-quality
studies, there is robust evidence that DBP:
• Is expected to degrade rapidly via direct and indirect photolysis and will rapidly degrade in the
atmosphere (ti/2 =1.15 days) (Section 4.3);
• Is not expected to hydrolyze under environmental conditions (Section 4.2);
• Is expected to have an environmental biodegradation half-life in aerobic environments on the
order of days to weeks (Section 4.1);
• Is not expected to be subject to long range transport;
• Is expected to transform in the environment via biotic and abiotic processes to form phthalate
monoesters, then phthalic acid, and ultimately biodegrade to form CO2 and/or CH4 (Section 1);
• Is expected to show strong affinity and sorption potential for organic carbon in soil and sediment
(Section 3.2);
• Will be removed at rates between 68 to 98 percent in conventional wastewater treatment systems
(Section 6.2);
• When released to air, will mostly partition to soil and water, and remaining DBP fraction in air
will rapidly degrade in the atmosphere (Section 5.1); and
• Is likely to be found and accumulate in indoor dust (Section 5.1.1).
As a result of limited studies identified, there is moderate evidence that DBP:
• Is not expected to biodegrade under anoxic conditions and may be persistent in anaerobic soils
and sediments (Section 4.1).
• Is not bioaccumulative in fish that reside in the water column (Section 1).
• May be bioaccumulative in benthic organisms exposed to sediment with elevated concentrations
of DBP proximal to continual sources of release (Section 1).
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240 • Is expected to be partially removed in conventional drinking water treatment systems via
241 sorption to suspended organic matter and filtering media (Section 6.3).
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1 INTRODUCTION
DBP is produced by the esterification of phthalic anhydride with isobutyl alcohol in the presence of an
acid catalyst. DBP is a member of the phthalate class of chemicals that are widely used as adhesives and
sealants in the construction and automotive sectors. DBP is also commonly used in electronics,
children's toys, and plastic and rubber materials. DBP is considered ubiquitous in various environmental
media due to its presence in both point and non-point source discharges from industrial and conventional
wastewater treatment effluents, biosolids, sewage sludge, stormwater runoff, and landfill leachate (Net
et al.. 2015).
This assessment was used to determine which environmental pathways to assess further for DBP's risk
evaluation. Details on the environmental partitioning and media assessments can be found in Section 4.
Based on DBP's fate parameters, EPA anticipates DBP to predominantly be found in water, soil, and
sediment. DBP in water is mostly attributable to discharges from industrial and municipal wastewater
treatment plant effluent, surface water runoff, and, to a lesser degree, atmospheric deposition. Once in
water, DBP is expected to mostly partition to suspended organic matter and aquatic sediments. DBP in
soils is attributable to deposition from air and land application of biosolids.
EPA quantitatively assessed concentrations of DBP in surface water, sediment, and soil from air-to-soil
deposition. Ambient air concentrations were quantified for the purpose of estimating soil concentrations
from air deposition but were not used for the exposure assessment as DBP was not assumed to be
persistent in the air (ti/2 =1.15 days (Peterson and Staples. 2003)). In addition, partitioning analysis
showed DBP partitions primarily to soil and water when compared to air and sediment, including from
air releases. Soil concentrations of DBP from land applications were not quantitatively assessed in the
screening level analysis since DBP is expected to have limited persistence potential and mobility in soils
receiving biosolids.
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268 2 APPROACH AND METHODOLOGY FOR PHYSICAL AND
269 CHEMICAL PROPERTY ASSESSMENT
270 EPA completed a systematic review by conducting a literature search of available published articles
271 through 2019 to find the following physical and chemical property values. After physical and chemical
272 property data have been extracted and evaluated, values for the endpoints are selected for use in the risk
273 evaluation as described in the Draft Systematic Review Protocol for Dibutyl Phthalate (DBP) (U.S.
274 EPA. 2024d). Due to the large quantity of available data, only studies with an overall data quality
275 ranking of "high" were selected for use in this risk evaluation. Empirical data for the octanol:air
276 partition coefficient (log Koa) and the air:water partition coefficient (log Kaw) were not available, thus
277 EPI Suite™ (U.S. EPA. 2017) was used to estimate a value for each of these parameters.
278 2.1 Selected Physical and Chemical Property Values for DBP
279
280 Table 2-1. Selected Physical and Chemical Property Values for DBP
Property
Selected Value(s)
Reference(s)
Data Quality
Rating
Molecular formula
C16H22O4
NLM (2024)
High
Molecular weight
278.35 g/mol
Havnes(2014b)
High
Physical form
Oily liquid
O'Neil (2013)
High
Melting point
-35 °C
Rumble (2018b)
High
Boiling point
340 °C
O'Neil (2013)
High
Density
1.0465 g/cnr
Rumble (2018b)
High
Vapor pressure
2.01E-05 mmHg
U.S. EPA (2019)
High
Vapor density
9.58
NLM (2024)
High
Water solubility
11.2 mg/L
Howard et al. (1985)
High
Octanol: water partition
coefficient (log Kow)
4.5
NLM (2024)
High
Octanol:air partition coefficient
(log Koa)
8.63 (EPI Suite™)
U.S. EPA (2017)
High
Airwater partition coefficient
(log Kaw)
-4.131 (EPI Suite™)
U.S. EPA (2017)
High
Henry's Law constant
1.81E-06 atm m7mol at 25
°C
NLM (2024)
High
Flash point
157 °C
NLM (2024)
High
Autoflammability
402 °C
NLM (2024)
High
Viscosity
20.3 cP
NLM (2024)
High
281
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2.2 Endpoint Assessments
2.2.1 Melting Point
Melting point informs the chemical's physical state, environmental fate and transport, as well as the
chemical's potential bioavailability. EPA extracted and evaluated eight high-quality data sources
containing DBP melting point information. These sources reported DBP melting points ranging from -
40 to -35 °C (NLM. 2024; NIST. 2022; Elsevier. 2019; U.S. EPA. 2019; Rumble. 2018b; DOE. 2016;
ECHA. 2012; NIOSH. 2007; Wang and Richert. 2007; Park and Sheehan. 2000). The mean of the
reported melting point values within these sources is -35.57°C. Seven of these sources reported a DBP
melting point of-35°C (NLM. 2024; NIST. 2022; Elsevier. 2019; U.S. EPA. 2019; Rumble. 2018b;
DOE. 2016; NIOSH. 2007). while one source reported a DBP melting point of -40°C (Park and
Sheehan. 2000). EPA selected a melting point value of-3 5 °C (Rumble. 2018b) as a representative
melting point value since this value is consistent with the average of the identified information from the
overall high-quality data sources. The identified value is consistent with the value proposed in the Final
Scope for the Risk Evaluation ofDibutyl Phthalate (DBP) (U.S. EPA. 2020).
2.2.2 Boiling Point
Boiling point informs the chemical's physical state, environmental fate and transport, as well as the
chemical's potential bioavailability. EPA extracted and evaluated eleven high-quality data sources
containing DBP boiling point information. These sources reported DBP boiling points ranging from 338
to 340.7 °C (NLM. 2024; NIST. 2022; Elsevier. 2019; U.S. EPA. 2019; Rumble. 2018b. c; DOE. 2016;
O'Neil. 2013; ECHA. 2012; NIOSH. 2007; Wang and Richert. 2007; Park and Sheehan. 2000). The
mean of reported boiling point values within these sources is 340 °C. EPA selected the boiling point
value of 340 °C reported by O'Neil (2013) since this value is consistent with the mean of all boiling
points measured under standard environmental conditions. The identified value is consistent with the
value proposed in the Final Scope for the Risk Evaluation ofDibutyl Phthalate (DBP) (U.S. EPA. 2020).
2.2.3 Density
EPA extracted and evaluated nine high-quality data sources containing DBP density information. These
sources reported DBP density values of 1.042 to 1.0501 g/cm3 (NLM. 2024; Elsevier. 2019; Rumble.
2018b; DOE. 2016; O'Neil. 2013; ECHA. 2012; Cadogan and Howick. 2000; Park and Sheehan. 2000;
WHO. 1997). The mean of the reported density values is 1.0462 g/cm3. EPA selected a density of
1.0465 g/cm3 (Rumble. 2018b) to closely represent the mean of the density values obtained from the
available high-quality data sources. The identified value is consistent with the value proposed in the
Final Scope for the Risk Evaluation ofDibutyl Phthalate (DBP) (U.S. EPA. 2020).
2.2.4 Vapor Pressure
Vapor pressure indicates the chemical's potential to volatilize, undergo fugitive emissions and other
releases to the atmosphere, undergo long range transport, and be available for specific exposure
pathways. EPA extracted and evaluated eight high-quality data sources containing DBP vapor pressure
information. One of these sources reported DBP vapor pressure values of 1.2 to 2.5 x 10~4 mmHg at 25
°C (Elsevier. 2019). The remaining seven high-quality data sources reported DBP vapor pressure
ranging from 2.01 x 10~5 to 7.28x 10~5 mmHg at 25 °C (NLM. 2024; U.S. EPA. 2019; Ishaketal.. 2016;
ECHA. 2012; Lu. 2009; NIOSH. 2007; Howard et al.. 1985; Hamilton. 1980). The mean and mode of
these reported vapor pressure values are 4.38xl0~5 and 2.01 xl0~5 mmHg, respectively, at 25°C. EPA
selected the experimentally derived vapor pressure value of 2.01 x 10~5 mmHg (U.S. EPA. 2019) to best
represent the mode vapor pressure of DBP obtained from the overall high-quality data sources under
standard environmental conditions. The identified value is consistent with the value proposed in the
Final Scope for the Risk Evaluation ofDibutyl Phthalate (DBP) (U.S. EPA. 2020).
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2.2.5 Vapor Density
EPA extracted and evaluated one high-quality and one medium-quality data source containing DBP
vapor density information. These sources reported DBP vapor densities of 9.58 and 9.60 (NLM. 2024;
NIOSH. 1976). EPA selected the vapor density value of 9.58 from the one available high-quality data
source as a representative value for standard environmental conditions. The identified value is consistent
with the value proposed in the Final Scope for the Risk Evaluation of Dibutyl Phthalate (DBP) (U.S.
EPA. 2020).
2.2.6 Water Solubility
Water solubility informs many endpoints not only within the realm of fate and transport of DBP in the
environment, but also when modeling for industrial process, engineering, human and ecological hazard,
and exposure assessments. EPA extracted and evaluated twelve high-quality data sources containing
DBP water solubility information. These sources reported water solubility values from 1.5 to 14.6 mg/L
(NLM. 2024; Elsevier. 2019; U.S. EPA. 2019; Rumble. 2018a; EC/HC. 2017; ECHA. 2012; NIOSH.
2007; Mueller and Klein. 1992; Defoe et al.. 1990; Howard et al.. 1985; SRC. 1983b). EPA excluded
two of the reported values, 1.5 and 14.6 mg/L (Elsevier. 2019). as those were determined to be potential
outliers. These values were higher or lower than the upper (13.00 mg/L) and lower bounds (8.20 mg/L)
calculated using the interquartile range (1.2 mg/L) rule for potential outliers (U.S. EPA. 2006). The rest
of the available data sources reported DBP's water solubility values from 8.7 to 11.4 mg/L. The mean of
the reported water solubilities at near ambient temperature is 10.62 mg/L. A water solubility of 11.2
mg/L (Howard et al.. 1985) was selected as the empirical value obtained from the overall high-quality
data sources that best represents DBP's mean water solubility under standard environmental conditions.
The identified value is consistent with the value proposed in the Final Scope for the Risk Evaluation of
Dibutyl Phthalate (DBP) (U.S. EPA. 2020).
2.2.7 Octanol:Air Partition Coefficient (log Kqa)
The octanol-air partition coefficient (Koa) provides information on how the chemical will partition
between octanol (which represents the lipids or fats in biota) and air. Koa informs on how DBP is likely
to partition between air, aerosol particles, foliage, dust, dry surfaces, soil, and animal tissue. No Koa
data for DBP were identified in the initial data review for the Final Scope for the Risk Evaluation of
DBP (U.S. EPA. 2020). After the final scope was published, EPA extracted and evaluated DBP octanol-
air partitioning (Koa) data from a single medium quality source. This source reported a predicted log
Koa value of 8.45 obtained from a quantitative structure-property relationship (QSPR) model (Lu.
2009). The QSPR-derived estimate of 8.45 reasonably aligns with DBP's log Koa value of 8.63
estimated using EPI Suite™ (U.S. EPA. 2017). As such, EPA has selected the EPI Suite™ derived value
of 8.63 as the representative log Koa value for use in risk assessment (U.S. EPA. 2017). The EPI
Suite™ modeled value was selected because it closely aligns with the reported predicted value and EPI
Suite™ is considered a highly reliable model.
2.2.8 OctanolrWater Partition Coefficient (log Kow)
The octanol-water partition coefficient (Kow) provides information on how the chemical will partition
between octanol (which represents the lipids or fats in biota) and water. Kow informs on how the
chemical is likely to partition in biological organisms as well as for the estimation of other properties
including water solubility, bioconcentration, soil adsorption, and aquatic toxicity. EPA extracted and
evaluated ten high-quality data sources containing DBP log Kow information. These sources included
six new additional sources not available for the Final Scope for the Risk Evaluation of Dibutyl Phthalate
(DBP) (U.S. EPA. 2020). EPA excluded a reported Kow of 3.74 from two data sources (Howard et al..
1985; SRC. 1984) as it was determined to be a potential outlier. This value is higher or lower than the
upper (4.41) and lower (4.65) bounds calculated using the interquartile range (0.24) rule for potential
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outliers (U.S. EPA. 2006). With the exclusion of potential outliers, these sources reported log Kow
values ranging from 4.25 to 4.79 (NLM. 2024; Elsevier. 2019; U.S. EPA. 2019; EC/HC. 2017; Ishak et
al.. 2016; ECHA. 2012; Verbruggen et al.. 1999; Mueller and Klein. 1992; Howard et al.. 1985; SRC.
1984). The mean of the reported log Kow values (excluding outliers) is 4.5. EPA selected an
experimental log Kow value of 4.5 (NLM. 2024) as this is consistent with the mean value obtained from
the overall high-quality data sources under standard environmental conditions. The identified value
replaces the value proposed in the Final Scope for the Risk Evaluation ofDibutyl Phthalate (DBP) (U.S.
EPA. 2020).
2.2.9 Henry's Law Constant
Henry's Law constant (HLC) provides an indication of a chemical's volatility from water and gives an
indication of potential environmental partitioning, potential removal in sewage treatment plants during
air stripping, and possible routes of environmental exposure. EPA extracted and evaluated four high-
quality data sources containing DBP HLC information. These sources reported DBP HLC values
ranging from 8.83xl0~7to 1.81 xl0~6 atmm3/mol (NLM. 2024; Elsevier. 2019; U.S. EPA. 2019;
Cousins and Mackav. 2000). The mean of the reported HLC values is 1.5xl0~6 atmm3/mol. EPA
selected the HLC value of 1.81 xl0~6 atmm3/mol (NLM. 2024) as the value obtained from the overall
high-quality data sources that best represents DBP's mean HLC under standard environmental
conditions. The identified value is consistent with the value proposed in the Final Scope for the Risk
Evaluation of Dibutyl Phthalate (DBP) (U.S. EPA. 2020).
2.2.10 Flash Point
EPA extracted and evaluated five high-quality data sources containing DBP flash point information.
These sources reported a DBP flash point of 157 to 171 °C (NLM. 2024; Elsevier. 2019; Rumble.
2018c; O'Neil. 2013; NIOSH. 2007). The mean of the reported flash point values is 162 °C. EPA
selected a flash point value of 157 °C (NLM. 2024) as the value that best represents the mean flash point
value obtained from the available overall high-quality data sources under standard environmental
conditions. The identified value is consistent with the value proposed in the Final Scope for the Risk
Evaluation of Dibutyl Phthalate (DBP) (U.S. EPA. 2020).
2.2.11 Autoflammability
No autoflammability data for DBP were identified in the initial data review for the Final Scope for the
Risk Evaluation of DBP (U.S. EPA. 2020). After the final scope was published, two high-quality and
two medium-quality data sources were identified in the systematic review process. The autoflammability
values ranged from 402 to 403 °C (NLM. 2024; NCBI. 2020; Rumble. 2018c; NIOSH. 1976). The mean
of the reported autoflammability values is 402 °C. EPA selected an autoflammability value of 402 °C for
DBP (NLM. 2024) as the value that best represents the mean flashpoint value.
2.2.12 Viscosity
EPA extracted and evaluated three high-quality data sources containing DBP viscosity information.
These sources reported viscosity values ranging from 16.63 to 20.3 cP at 20 to 25 °C (NLM. 2024;
Elsevier. 2019; Rumble. 2018d). The mean of the reported values is 19.12 cP. EPA selected a value of
20.3 cP at 20 °C as the value that best represents the mean of reported viscosity values under standard
environmental conditions for this risk evaluation. The identified value is consistent with the value
proposed in the Final Scope for the Risk Evaluation ofDibutyl Phthalate (DBP) (U.S. EPA. 2020).
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2.3 Strengths, Limitations, Assumptions, and Key Sources of Uncertainty
for the Physical and Chemical Property Assessment
The representative physical and chemical property values were selected based on professional
judgement and the weight of the scientific evidence, including the overall data quality ranking of the
associated references. These physical and chemical property values are then used to inform chemical-
specific decisions and model inputs across other disciplines. High and medium quality data are preferred
when selecting physical and chemical properties. In some instances where no data were available,
models such as EPI Suite™ were used to estimate the value for the endpoint (i.e., octanol:air partitioning
coefficient) and cross-checked with reported data from systematic review. The number and overall
quality of the available data sources results in different confidence strength levels for the corresponding
selected physical and chemical property values (U.S. EPA. 2021).
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3 APPROACH AND METHODOLOGY FOR FATE AND
TRANSPORT ASSESSMENT
In assessing the environmental fate and transport of DBP, EPA considered 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).
The full range of results from data sources that were rated high- and medium-quality were considered
for fate endpoints.
Information on the full extracted data set is available in the file Draft Risk Evaluation for Dibutyl
Phthalate (DBP) - Systematic Review Supplemental File: Data Quality Evaluation and Data Extraction
Information for Environmental Fate and Transport (U.S. EPA. 2024a). When no measured data were
available from high- or medium-quality data sources, fate values were obtained from EPI Suite™ (U.S.
EPA. 2017). a predictive tool for physical and chemical properties and environmental fate estimation.
Information regarding the model inputs is available in Section 3.2.1.
Table 3-1 provides a summary of the selected data that EPA considered while assessing the
environmental fate of DBP and were updated after publication of Final Scope of the Risk Evaluation for
Dibutyl Phthalate (DBP) (U.S. EPA. 2020) with additional information identified through the systematic
review process.
Table 3-1. Summary of E
~BP's Environmental Fate Information
Property or Endpoint
Value(s)
Reference(s)
Hydrolysis
ti/2 = approximately 22 years at pH 7 and 25 °C
ATSDR (1999)
Kh = 1.0 ± 0.05E-02 M"1 sec1 atpH 10-12 and 30
°C
Wolfe etal. (1980)
ti/2 = 45.4 hours at pH 10 and 30 °C
Zhane et al. (2019)
ti/2 = 3.43 years at pH 7 and 25 °C (estimated);
ti/2 = 125 days at pH 8 and 25 °C (estimated)
U.S. EPA (2017)
Indirect photodegradation
ti/2 =1.15 days (estimated based on a 12-hour day
with 1.5E06 OH/cm3 and OH rate constant of
9.28E-12 OH/cm3 and OH cm7molecule-sec)
Peterson and Staples (2003)
ti/2 = 1.13 days (based on a 12-hour day with
1.5E06 OH/cm3 and OH rate constant of 9.47E-12
OH/cm3 and OH cm3/molecule-sec)
Lei et al. (2018)
Organic carbon:water
partition coefficient (log
Koc)
3.69 (average of 7 values ranging between 3.14 to
3.94)
Russell and Mcduffie (1986);
Xiane et al. (2019)
Aerobic primary
biodegradation in water
69% by BOD, 100% by UV-VIS, and 100% by
GC after 2 weeks at a concentration of 100 ppm
using an unspecified method (most likely
Japanese MITI)
NITE (2019)
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Property or Endpoint
Value(s)
Reference(s)
100% in 7 days based on loss of test substance in
a synthetic medium containing 5 mg yeast extract
Tabak et al. (1981)
68.3 to >99% (average: 89.8%) primary
biodegradation after 28 days using inoculum
prepared with soil, domestic, influent sewage
microorganisms with a 2-week acclimation period
SRC (1983a)
Aerobic ultimate
biodegradation in water
57.4% by theoretical CO2 (ThC02) evolution
after 28 days
SRC (1983a)
84.6 ± 2.1% (mean ± SD) after 14 days at 22 °C
based on primary biodegradation
Johnson et al. (1984)
Aerobic biodegradation
in sediment
16, 56, 73, and 86% after 7 days at 5, 12, 22, and
28 °C, respectively
ti/2 = 2.9 days in natural river sediment collected
from the Zhonggang, Keya, Erren, Gaoping,
Donggang, and Danshui Rivers in Taiwan.
Yuan et al. (2002)
Anaerobic biodegradation
in sediment
ti/2 = 14.4 days in natural river sediment collected
from the Zhonggang, Keya, Erren, Gaoping,
Donggang, and Danshui Rivers in Taiwan.
Yuan et al. (2002)
Aerobic biodegradation
in soil
88.1-97.2% after 200 days in Chalmers slit loam,
Plainfield sand, and Fincastle silt loam soils.
Inman et al. (1984)
101%, 128%, and 89% after 8 weeks by mean %
theoretical gas production in revised anaerobic
mineral medium (RAMM), American Society for
Testing Materials (ASTM), and supplemental
medium from Jackson, MI, respectively
Anaerobic biodegradation
46%, 59%, and 19% after 8 weeks by %
theoretical gas production in RAMM, ASTM, and
supplemental medium from Holt, MI, respectively
Union Carbide (1974)
in WWTP sludge
72%, 117%, and 77% after 8 weeks by %
theoretical gas production in RAMM, ASTM, and
supplemental medium from Ionia, MI,
respectively
ti/2 = 5.1-6.2 days in primary sludge from
Lundofte municipal wastewater treatment plant
acclimated to 10 mg/L di-ethylhexyl phthalate
(DEHP) in Lyngby, Denmark
Gavala et al. (2003)
Removal in wastewater
treatment
96.6% removal by degradation and decantation
based on GC-MS analysis in Fontenay-les-Briis
(Essonne-France) WWTP
Tran et al. (2014)
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Property or Endpoint
Value(s)
Reference(s)
Removal efficiency (approximate, based on
figure):
primary sedimentation: ca. -50%;
chemical enhanced primary treatment: ca. -
100%;
activated sludge: ca. 75%;
sand filtration: ca. 95%;
chlorination disinfection: ca. 20%
Wuetal. (2017)
Aquatic bioconcentration
factor (BCF)
2.9 ± 0.1 and 30.6 ± 3.4 in brown shrimp
{Penaens aztecus) at 100 and 500 ppb,
respectively
Wofford et al. (1981)
11.7 in sheepshead minnow (Cyprinodon
variegate) at 100 ppb
21.1 ± 9.3 and 41.6 ± 5.1 in American oyster
(Crassostrea virginica) at 100 and 500 ppb,
respectively
Aquatic bioaccumulation
factor (BAF)
100, 316, 251 and 1259 L/kg dry weight (dw) in
bluegill, bass, skygager, and crucian carp,
respectively.
Lee et al. (2019a)
159 (estimated; upper trophic)
U.S. EPA (2017)
Aquatic biota-sediment
accumulation factor
(BSAF)
Log BSAF: -1.6, -1.5, -1.5 and -1.4 kg/kg dw, in
bluegill, bass, skygager, and crucian carp,
respectively
Lee et al. (2019a)
BSAF: 0.2-2 (approximate range from figure) in
Oreochromis miloticus niloticus, Liza subviridis,
Acanthopagrus schlegeli, Zacco platypus and
Acrossocheilus paradoxus
Huana et al. (2008)
BSAF: 5.5 ± 4.8, 6.0 ± 2.3, and 11.8 ± 12.6 in
roach, chub, and perch, respectively
Teiletal. (2012)
Aquatic Trophic
Magnification Factor
(TMF)
0.70 in 18 marine species
Mackintosh et al. (2004)
Terrestrial Biota-Soil
Accumulation Factor
(BSAF)
0.242 to 0.460 for earthworms
Hu et al. (2005) and Ji and
Dena (2016)
Plant Concentration
Factor (PCF)
0.26 to 4.78 (Fruit and vegetables)
Sun et al. (2015)
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3.1 Tier I Analysis
To be able to understand and predict the behaviors and effects of DBP in the environment, a Tier I
analysis will determine whether an environmental compartment (e.g., air, water, etc.) will accumulate
DBP at significant concentrations (i.e., major compartment) or not (i.e., minor compartment). The first
step in identifying the major and minor compartments for DBP is to consider partitioning values (Table
3-1), which indicate the potential for a substance to favor one compartment over another. DBP does not
naturally occur in the environment; however, DBP has been detected in water, soil, and sediment in
environmental monitoring studies (NLM. 2024; EC/HC. 2017).
3.1.1 Soil, Sediment, and Biosolids
Based on the partitioning values shown in Table 3-1, DBP will favor organic carbon over water or air.
Because organic carbon is present in soil, biosolids, and sediment, they are all considered major
compartments for DBP. This is consistent with monitoring data where higher concentrations of DBP
were detected in sediment samples (20-698 ng/g) compared to water samples (114-2,116 ng/L)
collected from the Mersey Estuary in the United Kingdom (NLM. 2024).
3.1.2 Air
DBP is a liquid at standard environmental temperatures with a melting point of-35°C and a vapor
pressure of 2.01 x 10~5 mm Hg at 25 °C (NLM. 2024). DBP will exist both in the vapor (gaseous) phase
and particulate phase in the atmosphere (EC/HC. 1994). The mean concentration of DBP was 1.9 ± 1.3
ng/m3 in the vapor phase and 4.0 ± 2.2 ng/m3 in the particulate phase in air samples collected along the
Niagara River (EC/HC. 1994). In another monitoring study from Paris, France, higher concentrations of
DBP were detected in the vapor phase (2.9 to 59.3 ng/m3) compared to the particulate phase (0.6 to 4.6
ng/m3) {NLM 2024). The octanol:air partition coefficient (Koa) indicates that DBP will favor the
organic carbon present in airborne particles. Based on its physical and chemical properties and short
half-life in the atmosphere (ti/2 =1.15 days (U.S. EPA. 2017)). DBP in the vapor phase is assumed to
not be persistent in the air. The AERO WIN™ module in EPI Suite™ estimates that a fraction of DBP
may be sorbed to airborne particulates and these particulates may be resistant to atmospheric oxidation.
DBP has been detected in both indoor and outdoor air and settled house dust in the USA, Europe,
Canada and China (NLM. 2024; EC/HC. 2017; Kubwabo et al.. 2013; Wang etal.,2013).
3.1.3 Water
A log Kaw value of -4.131 indicates that DBP will favor water over air. With a water solubility of 11.2
mg/L at 25 °C, DBP is expected to be slightly soluble in water (Howard et al.. 1985). In water, DBP is
likely to partition to suspended organic material present in the water column based on DBP's water
solubility of 11.2 mg/L (Howard et al.. 1985)and organic carbon:water partition coefficient of 3.69
(Table 3-1). A monitoring study showed that total seawater DBP concentrations, in the False Creek
Harbor is a shallow marine inlet in Vancouver, ranged from 50 to 244 ng/L and the dissolved fraction
concentrations ranged from 34 tol65 ng/L, compared to the suspended particulate fraction concentration
which ranged from 9,320 to 63,900 ng/g dry weight (dw) (Mackintosh et al.. 2006). Although DBP has
low water solubility, surface water will be a major compartment for DBP since it is detected in the ng/L
range.
3.2 Tier II Analysis
A Tier II analysis involves reviewing environmental release information for DBP to determine if further
assessment of specific media is needed. The Toxics Release Inventory (TRI) reported the total on-site
releases for DBP in 2022 to be 130,800 pounds with 49,600 pounds released to air, 81,200 pounds
released to land, and none released to water. According to production data from the Chemical Data
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Reporting (CDR) 2020 reporting period, between one million and ten million pounds of DBP were
produced annually from 2016-2019 for use in commercial products, chemical substances or mixtures
sold to consumers, or at industrial sites. Environmental release data from the Discharge Monitoring
Reports (DMRs) reported total annual releases for DBP from watershed discharge to be 1,224 total lb
per year from 585 watersheds in 2021, 5,149 total lb per year from 588 watersheds in 2022, and 16,555
total lb per year from 568 watersheds in 2023.
DBP is used mainly as a plasticizer in polyvinyl emulsions and can be used in adhesives, paints and
coatings, building materials, printing inks, fabric and textiles, children's toys, and plastic and rubber
materials (EC/HC. 2017. 1994). Because DBP is not chemically bound to the polymer matrix and can
migrate from the surface of polymer products (EC/HC. 2017). DBP can easily be released to the
environment from polymer-based products during their use and disposal. Additionally, DBP may be
released to the environment from the disposal of wastewater, and liquid and solid wastes. After
undergoing wastewater treatment processes, effluent is released to receiving waters and biosolids
(treated sludge) may be landfilled, land-applied, or incinerated and these processes may indicate that
media-specific evaluations are necessary (Table 3-2). Releases from landfills and incinerators will occur
from the disposal of liquid and solid wastes and warrants media specific evaluations.
Table 3-2. Summary of Key Environmental Pathways & Media Specific Evaluations
Environmental Releases
Key Pathway
Media Specific Evaluations
Wastewater and liquid
waste treatment
Effluent discharge to water and land
application of biosolids
Air, water, sediment, soil,
groundwater, and biosolids
Disposal of liquids and
solids to landfills
Leachate discharge to water and biogas to
air
Air, water sediment, soil, and
groundwater
Incineration of liquid and
solids
Stack emissions to air and ash to landfill
Air, water, sediment, soil, and
groundwater
Urban/remote areas
Fugitive emissions to air
Air, water, sediment, soil, and
groundwater
Deposition
Water and soil
Partitioning
Water, sediment, soil, and
groundwater
3.2.1 Fugacity Modeling
The approach described by Mackay et al. (1996) using the Level III Fugacity model in EPI Suite™
V4.11 (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. 2017). Environmental degradation half-lives were taken from high and medium
quality studies that were identified through systematic review to reduce levels of uncertainties (Table
3-3).
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Table 3-3. DBP Half-Life Inputs Used in EPI Suite™ Level III Fugacity Modeling
Media
Half-Life (days)
Reference(s)
Air
1.15
Lei etal. (2018); SRC (1983a)
Water
10
Soil
20
Sediment
90
The following input parameters, taken from Table 2-1 and discussed in detail in Section 2.2, were used
in LEV3EPI™:
• Melting point =-35.00 °C
• Vapor pressure = 2.01 xl0~5 mm Hg
• Water solubility = 11.2 mg/L
• Log Kow = 4.5
• SMILES: 0=C(0CCCC)c(c(cccl)C(=0)0CCCC)cl (representative structure)
Based on DBP's environmental half-lives, partitioning characteristics, and the results of LEV3EPI™,
DBP is expected to be found predominantly in water and soil (Figure 3-1). The model suggests that,
under a continuous release scenario, 99.9 percent of releases to soil will remain in soil, 90 percent of
releases to water will remain in water with about 10 percent partitioning to sediments, while 58 percent
of releases to air will end up in soil with another 6 percent in water. The LEV3EPI™ results were
consistent with environmental monitoring data. DBP's partitioning behavior are further discussed in the
media specific assessment (Section 1).
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ion
90
80
70
W so
4—1
c
o SO
E ^
<
i/i
ft 4"
30
20
¦
10
n
¦
Li
100% Soil Release 100% Air Release 100% Water Release Equal Releases (User Koc of
4915)
Air ¦ Water bSoiI ¦ Sediment
538 Figure 3-1. EPI Suite™ Level III Fugacity Modeling for DBP
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4 TRANSFORMATION PROCESSES
When released to the environment, DBP will be transformed to the monoester form (monobutyl
phthalate via abiotic processes such as photolysis and hydrolysis of the carboxylic acid ester group (U.S.
EPA. 2023). Biodegradation pathways for the phthalates consist of primary biodegradation from
phthalate diesters to phthalate monoesters and then to phthalic acid (PA), and ultimately biodegradation
of phthalic acid to form carbon dioxide (CO2) and/or methane (CH4) (Huang et al.. 2013a; Wolfe et al.,
1980). Monobutyl phthalate is both more soluble and more bioavailable than DBP. It is also expected to
undergo biodegradation more rapidly than the diester form. EPA considered DBP transformation
products and degradants qualitatively but due to their lack of persistence, these byproducts are not
expected to substantially contribute to risk, thus EPA is not considering them further in this risk
evaluation. Both biotic and abiotic degradation routes for DBP are described in the sections below.
4.1 Biodegradation
DBP is expected to be readily biodegradable in most aquatic and terrestrial environments. EPA extracted
and evaluated 85 biodegradation studies during systematic review. Twenty-six of these studies were
extracted and evaluated as overall high-quality data sources (Table 4-1). For the purposes of the
following biodegradation analysis, due to the large number of available high-quality data sources, EPA
focused on studies that were given a high data quality rating.
4.1.1 Aerobic Biodegradation in Water
EPA extracted eight high-quality studies evaluating the primary aerobic biodegradation of DBP in water
(Table 3-1). Studies that used activated sludge inoculums reported primary biodegradation rates greater
than 80 percent over 40 days (Desai et al.. 1990). 100 percent over 14 days (Fuiita et al.. 2005). and 100
percent over 7 days (Tabak et al.. 1981). However, there was an additional study using activated sludge
as an inoculum with a degradation rate of 0 percent over 100 hours in an unacclimated inoculum, but a
rate of 100 percent over 100 hours using an inoculum that was acclimated to DBP over the course of 150
days using a DBP concentration of 100 mg/L (Jianlong. 2004). These findings suggest that DBP might
appear to be persistent when released to aquatic environments with microbial populations that require an
adaptation phase to the initial DBP exposure; however, once adapted to DBP exposure, biodegradation
of DBP is expected to occur. A study using a combination inoculum of soil, activated sludge, and raw
sewage reported a primary biodegradation rate ranging from 68.3 to greater than 99 percent over 28 days
(SRC. 1983a). Three of the eight identified studies used natural surface water inoculums. Studies
reported primary biodegradation rates of 100 percent over 2 days in river water (Cripe et al.. 1987). 100
percent over 14 days in river water (Fuiita et al.. 2005). and 100 percent over 14 days in pond water
(Fuiita et al.. 2005); the other reported a half-life of 1.7 to 13 days estuarine and freshwater sites
(Walker et al.. 1984).
Two studies evaluating the ultimate biodegradation of DBP in water were also extracted. One of the
studies reported rates of 50 to 70 percent, 40 to 60 percent, and 20 to 50 percent when using inoculums
of activated sludge, river water, and pond water, respectively (Fuiita et al.. 2005). The other reported
rates were 47.4 to 74.9 percent, with half-lives of 9.6 to 20.9 days when using a mixture of soil,
activated sludge, and raw sewage as the inoculum (SRC. 1983a). While the biodegradation rate of DBP
in water will depend on the microbial community and its previous exposures to DBP, most of the data
on primary and ultimate biodegradation rates suggest an aerobic half-life of less than approximately 30
days using a variety of different inoculums. Therefore, for the purposes of this evaluation, DBP is
readily biodegradable under aerobic conditions and will have a half-life on the order of days to weeks.
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4.1.2 Biodegradation in Sediment
EPA extracted twenty-three high-quality studies evaluating biodegradation of DBP in sediment. Out of
the twenty-three studies, EPA focused on the nine studies that did not use sediments amended with
external inoculums to best represent DBP's biodegradation under natural environmental conditions. One
of these studies reported 70.1 to 84.6 percent biodegradation of DBP in freshwater lake sediment under
aerobic conditions (Johnson et al.. 1984). The same study also evaluated the effects of temperature and
DBP initial concentration on DBP's aerobic degradation rates. The results showed 70.1 to 72.6 percent
loss at initial DBP concentrations of 0.082 to 8.2 mg/L in 14 days and 73 to 86 percent loss at 22 to 28
°C in 7 days. However, the study reported 16 to 56 percent loss of DBP at 5 to 12 °C. These findings
showed that temperature could have a significant effect on biodegradation. Three of the selected data
sources reported DBP's biodegradation half-lives of 2.7, 2.9, and 46 days in marine sediments under
aerobic conditions (Li et al.. 2015; Kickham et al.. 2012; Yuan et al.. 2010). The data source reporting a
DBP biodegradation half-life of 46 days in marine sediment, reported higher than expected
biodegradation half-lives for other phthalates as well. In general, DBP is expected to have a sediment
biodegradation half-life of 2.7 to 2.9 days under normal aerobic environmental conditions, but extended
half-lives might be possible.
In contrast to aerobic conditions, phthalate esters are expected to have extended half-lives in sediment
under anaerobic conditions. Five of the nine extracted studies reported DBP biodegradation information
in sediment under anaerobic conditions (Li et al.. 2015; Lertsirisopon et al.. 2006; Chang et al.. 2005;
Kao et al.. 2005; Yuan et al.. 2002). Chang et. al. (2005) reported 100 percent biodegradation in 28 days
at pH 7 and 30 °C in sediments with a nutrient content commonly used to support microbial growth. Kao
et. al. (2005) reported 24 percent biodegradation in 30 days at pH 7 and 30 °C in sediment samples with
water (not amended with nutrients for microbial growth). Additionally, half-lives in anaerobic sediments
have been reported to be 1.2 to 1.6 days in pond sediment (Lertsirisopon et al.. 2006). 3.6 days in
submerged marine sediment (Li et al.. 2015). and 5.1 to 12.7 days in river sediment (Yuan et al.. 2002).
Overall, the available data show that there is variability in the biodegradation rates and that rates will
depend on environmental conditions, such as temperature, redox conditions, and pre-exposure of the
microbial communities to DBP. However, most of the studies evaluated suggest that at ambient
temperatures DBP will have a half-life of less than one year in both aerobic and anaerobic sediments.
Therefore, for the purposes of this evaluation it is assumed that DBP will have a half-life in sediments
on the order of weeks to months.
4.1.3 Biodegradation in Soil
EPA extracted eight high-quality studies evaluating biodegradation of DBP under aerobic and anaerobic
conditions in soil. Studies conducted using aerobic conditions report degradation rates of 88 to 98.6
percent over 200 days in silty loam and sandy soils (Inman et al.. 1984); 100 percent over 72 hours in
soils of non-specified types from Broome County, New York (Russell et al.. 1985); 100 percent over 15
days (Shanker et al.. 1985) in an alluvial garden soil; and 66 percent over 30 days in soil taken from a
garden of an unspecified soil type (Wang et al.. 1997a). Additionally, studies have reported aerobic
biodegradation half-lives of 0.338 to 1.2 days in udic ferrosol and aquic cambisol soils (Cheng et al..
2018); 7.8 to 8.3 days in agricultural black soils (Xu et al.. 2008); 1.6 days in a sandy clay loam (Yuan et
al.. 2011); and 17.2 days in loam from a farm (Zhao et al.. 2016). Anaerobic biodegradation rates have
been reported to be 97.8 percent over 200 days in a silt loam (Inman et al.. 1984) and 66 percent over 30
days in soil taken from a garden of an unspecified soil type (Shanker et al.. 1985). Overall, the available
data show that there is variability in the biodegradation rates, depending on environmental conditions
such as temperature, redox conditions, and pre-exposure of the microbial communities to DBP.
However, the study with the slowest biodegradation rate in this evaluation suggests a half-life in soil of
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631 approximately 65 days (Inman et al.. 1984). Therefore, for the purposes of this evaluation, it is assumed
632 that DBP will have a half-life in soils on the order of weeks to months.
633
634 Table 4-1. Summary of DBP Biodegradation Information
Environmental
Conditions
Degradation Value
Half-life (days)
Reference
Overall Data
Quality
Ranking
100%/2 days
N/A
Cripe et al. (1987)
High
>80%/40 days
N/A
Desai et al. (1990)
High
100%/14 days
N/A
Fuiita et al. (2005)
High
Aerobic primary
biodegradation in
100% in acclimated
and 0% in
unacclimated
activated sludge/100
hours
N/A
Jianlone (2004)
High
water
68.3-99%/28 days
N/A
SRC (1983a)
High
100%/7 days
N/A
Tabak et al. (1981)
High
N/A
1.7-13 days
Walker et al. (1984)
High
N/A
45.3-47.5 hours
Wane et al. (1997b)
High
Aerobic ultimate
biodegradation in
water
50-70% in activated
sludge, 40-60%
river water, and 20-
50% in pond
water/14 days
N/A
Fuiita et al. (2005)
High
47.7-74.9%/28d
days
9.6-20.9 days
SRC (1983a)
High
70.9% at 0.082
mg/L, 70.1% at 0.82
mg/L, and 8.2% at
8.2 mg/L/14 days
N/A
Johnson et al. (1984)
High
Aerobic
biodegradation in
16% at 5 °C, 56% at
12 °C, 73% at 22 °C,
86% at 28 °C/7 days
N/A
Johnson et al. (1984)
High
sediment
N/A
46 days in
marine inlet
sediment
Kickham et al. (2012)
High
N/A
2.7 days in
surface marine
sediment
Li et al. (2015)
High
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Environmental
Conditions
Degradation Value
Half-life (days)
Reference
Overall Data
Quality
Ranking
N/A
14.6 days in river
sediment
Pens and Li (2012)
High
N/A
0.6-5.4 days in
river sediment
Yuan et al. (2002)
High
N/A
1.6-2.9 days in
mangrove
sediment
Yuan et al. (2010)
High
Anaerobic
biodegradation in
sediment
100%/28 days
9.4 days
Chans et al. (2005)
High
24%/30 days in river
sediment
N/A
Kao et al. (2005)
High
N/A
3.6 days in non-
surface layer
marine sediment
Li et al. (2015)
High
N/A
5.1-12.7 days
Yuan et al. (2002)
High
N/A
1.2-1.6 days in
pond sediment
Lertsirisopon et al.
(2006)
High
Aerobic
biodegradation in
soil
N/A
0.338-1.2 days
Chens et al. (2018)
High
88-98.6%/200 days
(CO2 evolution)
N/A
Inman et al. (1984)
High
100%/72 hours
N/A
Russell et al. (1985)
High
100%/15 days
N/A
Shanker et al. (1985)
High
66%/30 days
N/A
Wane et al. (1997a)
High
N/A
7.8-8.3 days
Xu et al. (2008)
High
N/A
1.6 days
Yuan et al. (2011)
High
N/A
17.2 days
Zhao et al. (2016)
High
Anaerobic
biodegradation in
soil
97.8%/200 days
(CO2 evolution)
N/A
Inman et al. (1984)
High
66%/30 days
N/A
Shanker et al. (1985)
High
635
636 4.2 Hydrolysis
637 The hydrolysis half-life of DBP at neutral pH and temperatures relevant to environmental waters is not
638 expected to be significant (Lei et al.. 2018; Huang et al.. 2013a; Wolfe et al.. 1980). The hydrolysis half-
639 life was reported to be approximately 22 years (ATSDR. 1999). Hydrolysis under acidic and alkaline
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conditions is expected to occur with alkaline hydrolysis being more rapid. Alkaline hydrolysis will yield
phthalic acid with the monoester as an intermediate (Zhang et al.. 2019; Huang et al.. 2013a; Wolfe et
al.. 1980). Zhang et al. (2019) evaluated the hydrolysis of DBP in aqueous alkaline solutions (pH 10) at
30 °C. The study reported hydrolysis to be rapid under the tested conditions, reporting a hydrolysis half-
life of 45.4 hours. Temperature has also been shown to impact hydrolysis rates with hydrolysis rates
increasing with an increase in temperature. The hydrolysis half-life for DBP was reported to be 280.2
hours in neutral solution at a temperature of 80 °C. Wolfe et al. (1980) evaluated the hydrolysis of DBP
in aqueous alkaline solutions at 30 °C. The study reported a hydrolysis rate constant of 1.0 ± 0,05/10 2
1VT1 sec-1 which corresponds to half-lives of 22 years at pH 7 and 8 days at pH 10. In addition, EPI
Suite™ estimated the hydrolysis half4ives of DBP to be 3.43 years at pH 7 and 25 °C, and 125 days at
pH 8 and 25 °C (U.S. EPA. 2017) indicating that hydrolysis of DBP is more likely under more caustic
conditions and supporting DBP's resistance to hydrolysis under standard environmental conditions.
When compared to other degradation pathways, hydrolysis it is not expected to be 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 DBP via hydrolysis (Huang et al.. 2013a). This is discussed further in
Section 5.3.3.
4.3 Photolysis
DBP contains chromophores that absorb light at greater than 290 nm wavelength (NLM. 2013).
therefore, direct photodegradation is a relevant but minor degradation pathway for DBP released to air.
The major degradation pathway for DBP in air is indirect photodegradation with a measured half4ife
of 1.13 days (27.1 hours) (calculated from a -OH rate constant of 9.47xl0~12 cm3 /molecule-second
and a 12-hour day with 1.5><106 OH/cm3) (Lei et al.. 2018). Similarly, Peterson and Staples (2003)
reported a calculated DBP photodegradation half-life of 1.15 days (~ 27.6 hours) (calculated from a
OHrate constant of 9.28xl0~12 cm3 /molecule-second and 1.5><106 OH/cm3). Indirect
photodegradation of DBP will yield MBP, PA, di-butyl 4-hydroxyphthalate (m-OH-DBP), and di-butyl
4-nitrophthalate (m-NCh-DBP) (Lei et al.. 2018). DBP photodegradation in water is expected to be
slower than air, due to the typical light attenuation in natural surface water. There is limited
information on the aquatic photodegradation of DBP. However, Lertsirisopon et al. (2009) reported
DBP aquatic direct photodegradation half-lives of 50, 66, 360, 94 and 57 days 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). Peterson and Staples (2003) also reported a half-life of 3 hours for aqueous
photolysis of DBP in natural sunlight when DBP was present in a surface microlayer on the water at
mg/L concentrations. The rate was noted to be stimulated by titanium dioxide and hydrogen peroxide.
These findings suggest DBP will be susceptible to photochemical decay in air but that photolysis is not
expected to be a significant degradation process in surface water.
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5 MEDIA ASSESSMENTS
DBP has been reported to be present in the atmosphere, aquatic environments, and terrestrial
environments. Once in the air, DBP will be most predominant in the organic matter present in airborne
particles and is expected to have a short half-life in the atmosphere. Based on its physical and chemical
properties, DBP is likely to partition to house dust and airborne particles in the indoor environment and
is expected to have a longer half-life in indoor air as compared to outdoor air. DBP present in surface
water is expected to partly partition to aquatic sediments and have an aerobic biodegradation half4ife
ranging from days to weeks. In terrestrial environments, DBP has the potential to be present in soils and
groundwater but is likely to be immobile in both media types. In soils, DBP is expected to be deposited
via air deposition and land application of biosolids. DBP in soils is expected to have a half4ife on the
order of weeks to months, and to have low bioaccumulation potential and biomagnification potential in
terrestrial organisms. DBP will be released to groundwater via infiltration from wastewater effluent and
landfill leachates but is not likely to be persistent in groundwater and/or subsurface environments unless
anoxic conditions exist.
5.1 Air and Atmosphere
DBP is a liquid at environmental temperatures with a melting point of-3 5 °C (Havnes. 2014a) (Rumble.
2018b) and a vapor pressure of 2.01 x 10~5 mmHg at 25 °C (NLM. 2024). Based on its physical and
chemical properties and short half4ife in the atmosphere (ti/2 =1.15 days (Peterson and Staples. 2003)).
DBP is not expected to be persistent in air. The AEROWIN™ module in EPI Suite™ estimated a log
KoAof 8.63, which suggests that a fraction of DBP may be sorbed to airborne particles and these
particulates may be more resistant to atmospheric oxidation. Thus, DBP has the potential to undergo dry
deposition and wet deposition into soils and surface water (Zeng et al.. 2010; Peters et al.. 2008; Xie et
al.. 2005; Parkerton and Staples. 2003; Atlas and Giam. 1981). Two studies reported a range of 33 to 46
percent of DBP concentration in the air to be associated with suspended particles (Xie et al.. 2007; Xie
et al.. 2005). A net deposition of DBP from ambient air into the North Sea was also measured (Xie et al..
2005). Based on DBP's short half-life in the atmosphere, it is not expected to be persistent in
atmospheric air under standard environmental conditions.
Three studies reported DBP to be detected in air at concentrations of greater than 0.002 to 3.4 ng/m3
over the North Sea (Xie et al.. 2005). 0.2 to 0.6 ng/m3 over the Arctic (Xie et al.. 2007). and 0.4 to 1.8
ng/m3 over the North Pacific Ocean (Atlas and Giam. 1981). Other studies measured concentrations of
DBP in ambient air ranging from 23.7 to 191 ng/m3 in the United States (Wilson et al.. 2003; Wilson et
al.. 2001) and 0.08 to 15 ng/m3 in Sweden (Cousins et al.. 2007).
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 al.. 2015). A log Koa
value of 8.63 suggests a strong affinity of DBP for organic matter in air particulates. DBP is expected to
be more persistent in indoor air than in outdoor air due to the lack of natural chemical removal
processes, such as solar photochemical degradation.
EPA identified several data sources reporting the presence of DBP in indoor air and dust in the United
States. These studies reported the presence of DBP at higher concentrations in indoor dust samples than
in indoor air, supporting DBP's strong affinity and partitioning to organic matter in dust. Wilson et al.
(2001) reported measured samples of indoor air and dust from ten daycare centers located in North
Carolina. DBP was detected in all air and dust samples with a mean concentration of 239 ng/m3 (108-
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404 ng/m3) in air samples and a mean concentration of 18.4 ppm (1.58-46.3 ppm) in dust samples. In a
second study, Wilson et al. (2003) reported measured samples of indoor air and dust from two other
daycare centers located in North Carolina with a mean concentration of 488 ng/m3 (222-786 ng/m3) in
air samples and a mean concentration of 1.87 ppm (0.058-5.85 ppm) in dust samples. Air and dust
samples were collected from residential and office buildings in Massachusetts with a 100 percent
detection frequency for DBP. Concentrations of DBP were found to be a mean of 0.251 |ig/m3 (0.101-
0.41 |ig/m3) in air and a mean of 27.4 |ig/g (11.1-59.4 |ig/g) in dust (Rudel et al.. 2001).
EPA also identified several data sources reporting the presence of DBP in indoor air and dust outside of
the United States. Das et al. (2014) explored the implications of industrial activities by comparing the
presence of phthalates in two different cities from India. The study analyzed indoor air and dust samples
from the Jawaharlal Nehru University campus (a city with low industrial activities) and Okhla (a city
with high industrial activities related to the use of phthalates), reporting a general tendency of higher
detectable concentrations of DBP in air and dust samples collected in the city of Okhla. This finding
suggests that higher concentrations of phthalates in air and dust could be expected near facilities with
high use and production of phthalates. Wormuth et al. (2006) determined the indoor air and indoor dust
concentrations DBP based on measured concentrations of phthalates in dust of European homes. The
study reported DBP mean concentrations of 1,153 ng/m3 and 98 mg/kg for indoor air and indoor dust,
respectively. In a study done in Sapporo, Japan, DBP was found to range from 79.6 to 740 ng/m3 in air
and 1.8 to 1,476 ng/m3 in indoor dust in residential houses (Kanazawa et al.. 2010). DBP was found to
be the dominating phthalate in a study which analyzed the phthalate content (DBP, BBP, dicyclohexyl
phthalate [DCHP], and di-ethylhexyl phthalate [DEHP]) of particulate matter in indoor spaces in
Norway (Rakkestad et al.. 2007).
5.2 Aquatic Environments
5.2.1 Surface Water
DBP 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. DBP has frequently been
detected in surface waters (Zeng et al.. 2008a; Tan. 1995; Preston and Al-Omran. 1989).
The principal properties governing the fate and transport of DBP in surface water are water solubility
(11.2 mg/L, Table 2-1), log Kaw (-4.131, Table 3-1), and log Koc (3.14-3.94, Table 3-1). Due to its
HLC (1.81xl0~6 atm m3/mol at 25 °C, Table 2-1), volatilization is not expected to be a significant
source of loss of DBP from surface water. A partitioning analysis estimates that about 10 percent of the
DBP released to water will partition to sediments and approximately 90 percent will remain in surface
water (see Section 3.2.1). However, based on its log Koc (3.14-3.94), DBP in water is expected to
partition to suspended particles and sediments. DBP is also expected to biodegrade rapidly in most
aquatic environments (Section 4.1.1) and thus is not expected to persist in surface water except at areas
of continuous release, such as a water body receiving discharge from a municipal wastewater treatment
plant, where rate of release exceeds the rate of biodegradation.
No monitoring data for DBP in surface water was readily available for the United States. Several studies
from outside the U.S. were examined. The available data sources reported the presence of DBP and
other phthalates in surface water samples collected from rivers and lakes. Preston and Al-Omran (1989)
explored the presence of phthalates within the River Mersey Estuary and reported the presence of DBP
freely dissolved in water at concentrations ranging from 0.541 to 1.805 |ig/L. Tan (1995) reported the
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presence of DBP in Klang River at concentrations of 0.8 to 4.8 |ig/L. Zeng et al. (2008a) reported the
presence of DBP in the dissolved aqueous phase of urban lakes in Guangzhou City at mean
concentrations of 2.03 |ig/L. Grigoriadou et al. (2008) reported the presence of DBP in lake water
samples collected near the industrial area of Kavala city at concentrations of 0.640 to 16 |ig/L. The total
seawater concentrations of DBP in False Creek Harbor, Vancouver ranged from 50 to 244 ng/L with the
dissolved fraction concentrations ranging from 34 to 165 ng/L. The bottom sediment concentrations
ranged from 57 to 182 ng/g dw. The concentration in suspended sediment ranged from 9,320 to 63,900
ng/g dw (Mackintosh et al.. 2006). These results show higher concentrations of DBP in the suspended
sediments than in the dissolved phase or the bottom sediment, which was not expected given the Koc
value and partitioning analysis results for DBP. This suggests that partitioning of DBP to sediments may
be much higher than what was predicted in the partitioning analysis and that the concentrations of DBP
in water may be mostly found in suspended sediment.
5.2.2 Sediments
Based on a log Koc range of 3.14 to 3.94, DBP will partition to the organic matter present in soils and
sediment when released into aquatic environments. Once in water, LEV3EPI™predicts that close to 90
percent of the DBP will remain in water (U.S. EPA. 2017) (see Section 3.2.1). However, some data
sources have documented higher concentrations of DBP in suspended solids than the dissolved phase
(Mackintosh et al.. 2006).
DBP is expected to biodegrade rapidly in aquatic sediments with a half4ife of weeks to months (see
Section 4.1.2). Due to its strong affinity to organic carbon (log Koc = 3.14- 3.94), DBP is expected to
partly partition to aquatic sediments. This is consistent with the monitoring data sources containing
information on the presence of DBP in river sediment samples. DBP concentrations in river sediment
ranged between 3 to 3,670 ng/g dw (Cheng et al.. 2019; Li et al.. 2017b; Li et al.. 2017a; Tang et al..
2017; Tan. 1995; Preston and Al-Omran. 1989).
No monitoring data for DBP in surface water was readily available for the United States. Several studies
from outside the U.S. were examined. Mackintosh (2006) reported higher concentrations of DBP in the
suspended particles than in deep sediment samples collected from the False Creek Harbor in Vancouver,
Canada. The study reported DBP mean concentrations of 103 and 22,400 ng/g in the deep sediment and
suspended particles, respectively.
In another study, Kim (2021) evaluated the presence of plasticizers in sediments from highly
industrialized bays of Korea. DBP was detected in 95 percent of the collected surface sediment samples
at a median concentration of 13.2 ng/g dw. 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 suggests that industrial activities are a major contributor of phthalates in
sediments within the area. It also suggests that DBP has the potential to accumulate in sediments at areas
of continuous release, such as a surface water body receiving discharge from a municipal wastewater
treatment plant.
5.3 Terrestrial Environments
5.3.1 Soil
DBP is expected to be deposited to soil via two primary routes: 1) application of biosolids and sewage
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sludge in agricultural applications or sludge drying applications; and 2) atmospheric deposition. Based
on DBP's HLC of 1.81xl0~6 atm m3/mol at 25 °C and vapor pressure of 2.01xl0~5 mmHg at 25 °C,
DBP is not likely to volatilize significantly from soils.
DBP is expected to show strong affinity for sorption to soil and its organic constituents based on a log
Koc of 3.14-3.94 (Xiang et al.. 2019: Russell and Mcduffie. 19861 and a log Kow of 4.5 (NLM. 2024).
Thus, DBP is expected to have slow migration potential in soil environments. In addition, DBP is
expected to biodegrade rapidly in soil with a half-life of weeks to months. In general, DBP 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.
Available data sources have reported the presence of DBP in soil samples. Concentrations ranging from
0.49 to 3.59 mg/kg dw were measured in soil and sediment samples in a vacant tract adjacent to the
Union Carbide Corporation's Bound Brook plant in New Jersey (ERM. 1988). Soil samples from waste
disposal sites in Taizhou, China were shown to be contaminated by DBP through improper disposal of
electronic waste. DBP and DEHP were two of the major phthalates in the study with DBP ranging from
1 to 5 mg/kg in the soil samples (Liu et al.. 2009). DBP, di-isobutyl phthalate (DIBP), and DEHP were
also found to be the main phthalates in agricultural soils in peri-urban areas around Guangzhou, China
with a 100 percent detection frequency. In New York, DBP was found in soil at concentrations ranging
from 0.009 to 2.74 |ig/g dw, which exceeds the recommended allowable soil concentrations for DBP set
by the state of New York (0.081 |ig/g). The study attributed the source of the phthalates to wastewater
irrigation, sewage sludge application, disposed plastics and atmospheric deposition (Zeng et al.. 2008b).
Similarly, another study found DBP in abundance in Chinese arable soil. Zeng et al. (2009) reported
DBP concentrations ranging from 0.206 to 30.1 |ig/g dw in soils from roadsides, residential areas, and
parks in Guangzhou, China.
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 meets the EPA pollutant and pathogen requirements
for land application and surface disposal and can be beneficially recycled (40 CFR Part 503) (U.S. EPA.
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.
As described in Section 6.2, DBP in wastewater has been reported to be mainly removed by particle
sorption and retained in the sewage sludge. Based on the STPWIN™ module in EPI Suite™, about 55
percent of DBP present in wastewater is expected to be accumulated in sewage sludge and discharged
into biosolids. The National Sewage Sludge Survey detected DBP in 1998 at a mean concentration of
11,200 |ig/kg, a standard deviation of 17,800 |ig/kg, a maximum concentration of 331,000 |ig/kg and a
4 percent detection frequency. Separately, DBP concentrations ranging from 1.7 to 1,260 ng/g dw were
measured in 20 municipal sewage sludge samples from publicly owned treatment works in the United
States (Ikonomou et al.. 2012).
Three studies have reported DBP's concentration in sludge in 71 Chinese WWTPs ranging from 0.0004
to 111 |ig/g dw (Zhu et al.. 2019; Meng et al.. 2014) and 0.58 to 59 |ig/g dw in 40 Korean WWTPs (Lee
et al.. 2019b). Two U.S. studies reported sludge concentrations ranging from 0.32 to 17 |ig/g dw
(Howie. 1991) and 966 |ig/L (ATSDR. 1999). When in biosolids, DBP may be transferred to soil during
land applications. DBP is likely to be more persistent in soil due to its strong sorption potential (Section
5.3.1). Land applied DBP is expected to be moderately mobile in the environment despite its strong
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sorption to soils. Disposal of sewage effluent has been reported to contaminate groundwater with DBP
concentrations up to 450 mg/L (ATSDR. 1999).
5.3.3 Landfills
For the purpose of this assessment, landfills will be divided into two zones: 1) an "upper-landfill" zone,
with standard environmental temperatures and pressures, where biotic processes are the predominant
route of degradation for DBP, and 2) a "lower-landfill" zone where elevated temperatures and pressures
exist, and abiotic degradation is the predominant route of degradation for DBP. 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 DBP. 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.. 2013a).
DBP is deposited in landfills continually and in high amounts from the disposal of consumer products
containing DBP. Some aerobic biodegradation may occur in upper-landfills. Similar to other phthalate
esters, under anaerobic conditions present in lower-landfills, DBP is likely to be persistent in landfills
due to the expected low rates of anaerobic biodegradation in lower-landfills. There is some evidence to
support that hydrolysis may be the main route of abiotic degradation of phthalate esters in lower-
landfills (Huang et al.. 2013a). Due to the expected persistence of DBP in landfills, it may dissolve into
leachate in small amounts based on a water solubility of 11.2 mg/L and may travel slowly to ground
water during infiltration of rainwater based on a log Koc of 3.14 to 3.94. For instance, several data
sources have reported the presence of DBP in landfill leachate. These sources have reported a rapid
decrease in DBP concentration from core to leachate samples (Norin and Strom vail. 2004; Jang and
Townsend. 2001; Oman and Hynning. 1993; DERS. 1991). These data sources reported DBP
concentrations ranging from 0.4 to 7.8 mg/kg and 1 to 17 |ig/L in landfill core and leachate samples,
respectively. The reported rapid decrease in DBP's concentration aligns with the expectation that DBP is
likely to sorb to organic matter in landfill soils.
5.3.4 Groundwater
There are several potential sources of DBP in groundwater, including wastewater effluents and landfill
leachates, which are discussed in Sections 5.3.3 and 6.2 . Furthermore, in environments where DBP is
found in surface water, it may enter groundwater through surface water/groundwater interactions.
Diffuse sources include storm water runoff and runoff from biosolids applied to agricultural land.
Even though DBP has a strong affinity to adsorb to organic matter present in soils and sediments (log
Koc = 3.14-3.94 (Xiang et al.. 2019; Russell and Mcduffie. 1986)). DBP partitioning to groundwater is
possible, though will be limited by DBP's low water solubility (11.2 mg/L). For instance, the presence
of DBP in groundwater has been reported at concentrations of 0.12 mg/L in Carson, California
(Geraghtv & Miller Inc. 1990). In cases where DBP 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 DBP in aerobic and
anaerobic environments (Section 4.1), and DBP is not likely to be persistent in groundwater or
subsurface environments unless anoxic conditions exist.
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6 REMOVAL AND PERSISTENCE POTENTIAL OF DBP
DBP is not expected to be persistent in the environment, as it is expected to degrade rapidly under most
environmental conditions, with lower biodegradation potential in low-oxygen media. In the atmosphere,
DBP is unlikely to remain for long periods of time as it is expected to undergo photolytic degradation
through reaction with atmospheric hydroxyl radicals, with an estimated half-life of 1.15 days. In aquatic
environments, DBP is predicted to hydrolyze slowly at ambient temperature, but it is not expected to
persist since it undergoes rapid aerobic biodegradation (Section 5.2.1). In soil and sediments, DBP has
the potential to remain for longer periods of time. Due to the rapid biodegradation under most aquatic
environments and its estimated BCF of 159.4 L/kg, DBP is expected to have low bioaccumulation
potential. Using LEV3EPI™ (Section 3.2.1), DBP's overall environmental half-life was estimated to be
approximately 14 days (U.S. EPA. 2017). Therefore, DBP 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. DBP is
classified as a hazardous substance (40CFR116.4) 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) (Federal Register. 1981).
Currently there is limited information available on the DRE of DBP. The available data sources reported
the presence of DBP in the ashes and exhaust gas from hazardous waste incinerators at concentrations of
0 to 1200 |ig/kg and 7.66 to 260 |ig/m3, respectively (Jay and Stieglitz. 1995; Nishikawa et al.. 1992;
Shane et al.. 1990). These findings suggest that incineration of DBP containing waste has the potential
to contribute to DBP concentrations in air. However, EPA estimated that highest waste incineration
stack emissions for DBP to be 0.03 tons per year, which corresponds to 0.058 percent of the reported
DBP TRI air releases in 1990 (Dempsev. 1993). This suggest that DBP present during incineration
processes will mainly be released with ash to landfills, with a small fraction released to air as stack
emissions. Based on its hydrophobicity and sorption potential, DBP released to landfills is expected to
partition to waste organic matter. Similarly, DBP released to air is expected to rapidly react via indirect
photochemical processes within hours (U.S. EPA. 2017) or partition to soil and sediments as described
in Section 3.2.1. DBP in sediments and soils is not expected to be bioavailable for uptake and is highly
biodegradable in its bioavailable form (Kickham et al.. 2012).
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) (U.S. EPA. 1988).
EPA selected twelve high-quality data sources reporting the removal of DBP in wastewater treatment
systems employing both aerobic and anaerobic processes. These sources reported a range of 38 to
greater than 99 percent removal of DBP in WWTPs employing secondary and/or tertiary treatment units
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such as activated sludge, secondary clarifiers, and sand filtration (Wu et al.. 2017; Huang et al.. 2013b;
Shao and Ma. 2009; Peterson and Staples. 2003) (Table 6-1). These studies reported that biodegradation
accounted for 27 to 58.9 percent of the overall DBP removal (Shao and Ma. 2009; Peterson and Staples.
2003). that the main removal mechanisms are sorption and biodegradation during the primary and
secondary treatment, respectively (Wu et al.. 2017; Huang et al.. 2013b). and that WWTPs employing
secondary and tertiary treatment achieve greater than 99 percent removal of DBP (Wu et al.. 2017). The
median removal of DBP has been reported to be 68 to 98 percent within 50 WWTPs in the United States
(U.S. EPA. 1982).
Based on the available information, the main mechanisms for the removal of DBP in conventional
municipal WWTPs are sorption to suspended organic matter, biodegradation during activated sludge
treatment, or a combination of sorption and biodegradation. For instance, recent studies have reported
greater than 93 percent removal of DBP in three conventional WWTPs with activated sludge treatment
in South Africa (Salaudeen et al.. 2018a. b). The studies reported that DBP is mainly removed by
sorption to suspended particles and sludge. Tran et al. (2014) reported similar findings in a WWTP in
France employing a combined decantation and activated sludge tank that achieved 96.6 percent removal
of DBP. The study reports that the evaluated phthalate esters (DIBP, DBP, BBP, DEHP, di-isononyl
phthalate [DINP], and di-isododecyl phthalate [DIDP]) were mainly removed by sorption to solids.
Other studies have reported biodegradation during the activated sludge treatment process to be the main
removal mechanism of DBP in two WWTPs in Denmark and India, achieving greater than 91 percent
removal of DBP (Saini et al.. 2016; Roslev et al.. 2007).
In contrast to higher molecular weight phthalate esters, DBP has been reported to be efficiently removed
during anoxic and anaerobic wastewater treatment processes (Table 6-1). Gani and Kazmi (2016)
evaluated the removal efficiency of DBP in three WWTPs employing anoxic, aerobic, and anaerobic
treatment units near the Ganga and Dhamola rivers in India. The wastewater treatment plants
investigated were designed as nutrient rem oval-based sequencing batch reactor (WWTP1-SBR) with
anoxic pretreatment zone followed by an activated sludge unit, a conventional activated sludge process
(WWTP2-ASP) and up-flow anaerobic sludge blanket (WWTP3-UASB) with a polishing pond. The
study reported that biotransformation processes accounted for 70, 67, and 61 percent of the overall DBP
removal in the SBR, ASP, UASB treatment plants, respectively. Sorption accounted for less than 5
percent of the overall removal of DBP. These findings suggest DBP to be biodegradable under anaerobic
conditions. This is supported by a study that explored the efficiency of anaerobic and aerobic sludge
post-treatment for the removal of phthalate esters (PAEs) and reported complete removal of DBP during
the anaerobic phase (Tomei et al.. 2019).
Unlike phthalate esters with longer carbon chains, DBP's water solubility (11.2 mg/L) and log Koc
(3.14-3.94) suggest partial removal in WWTP via sorption to sludge. This finding is supported by
STPWIN™, which predicted 56 percent of DBP to be removed during conventional wastewater
treatment by sorption to sludge with the potential for higher removal via rapid aerobic biodegradation
processes (U.S. EPA. 2017). In general, the available information suggests that aerobic processes have
the potential to help biodegrade DBP from wastewater, which is in agreement with the expected aerobic
biodegradation described in Section 3.1. Air stripping within the aeration tanks for activated sludge
processing is not expected to be a significant removal mechanism for DBP present in wastewater. In
general, based on the available measured and predicted information, WWTPs are expected to remove 65
to 98 percent of DBP present in wastewater.
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Table 6-1. Summary of I
>BP's WWTP Removal Information
Property
Selected Value(s)
Reference(s)
Data Quality
Rating
Removal by sorption
>93% removal; DBP removal in
three activated sludge WWTPs in
South Africa; main removal
mechanism: sorption
Salaudeen et al. (2018a);
Salaudeen et al. (2018b)
High
96.6% removal, main removal
mechanism: sorption
Tran et al. (2014)
High
Removal by
biodegradation
91% removal; biodegradation
during activated sludge process
Roslev et al. (2007)
High
92.67% removal; biodegradation
during activated sludge process
Saini et al. (2016)
High
Removal by
biodegradation and
sorption
90.10% removal; sorption and
biodegradation during the
primary and secondary treatment,
respectively
Huans et al. (2013b)
High
85.9% overall removal, 58.9%
biodegradation, and 11.3%
sorption to solids
Shao and Ma (2009)
N2F
38 to >99% removal;
Wu et al. (2017)
High
85 and 95%; biodegradation
accounted about 27 percent of the
overall removal
Peterson and Staples
(2003)
N2F
>57%
Wu et al. (2019)
High
70% (SBR),
67% (ASP),
61% (UASB);
<5% sorption, mainly
biodegradation
Gani and Kazmi (2016)
High
Anaerobic sludge post-
treatment
>99% removal, anaerobic sludge
post-treatment
Tomei et al. (2019)
High
WWTP = Wastewater treatment plant; SBR = Sequencing batch reactor; ASP = Activated sludge process; UASB =
Up-flow anaerobic sludge blanket
997 6.3 Removal in Drinking Water Treatment
998 Drinking water in the United States typically comes from surface water (i.e., lakes, rivers, reservoirs)
999 and groundwater. The source water flows to a treatment plant where it undergoes a series of water
1000 treatment steps before being dispersed to homes and communities. In the United States, public water
1001 systems often use conventional treatment processes that include coagulation, flocculation,
1002 sedimentation, filtration, and disinfection, as required by law.
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Limited information is available on the removal of DBP in drinking water treatment plants. A water
concentration of 100 ng/L was measured in the city of Philadelphia's drinking water (Roy F. Weston
Inc. 1980). Several available data sources reported concentrations of DBP in drinking water outside the
United States (1-1,830 ng/L) (Ding et al.. 2019; Li et al.. 2019; Kong et al.. 2017; Shan et al.. 2016; Das
et al.. 2014; Shi et al.. 2012). Kong et al. (2017) explored the presence and removal of phthalate esters in
a drinking water treatment system in east China employing coagulation, sedimentation, and filtration
treatment processes, and reported 64.5 percent removal of DBP from the treated effluent with a drinking
water concentration of 17.2 ng/L. Similarly, Shan et al. (2016) explored the removal of phthalate esters
in two drinking water treatment plants in east China. The first plant employs coagulation, sedimentation,
filtration, and disinfection treatment processes and reported 31 to 48 percent removal of DBP from the
treated effluent while the second plant reported 38 to 56 percent removal of DBP from the treated
effluent in a drinking water treatment system employing peroxidation, coagulation, combined
flocculation and sedimentation, filtration, and disinfection treatment processes. These findings suggest
that conventional drinking water treatment systems have the potential to partially remove DBP present
in source water via sorption to suspended organic matter and filtering media.
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7 BIOACCUMULATION POTENTIAL OF DBP
The presence of DBP in several marine aquatic species in North America suggest that the substance is
bioavailable in aquatic environments (Mackintosh et al.. 2004). However, DBP can be considered
readily biodegradable in most aquatic environments, and the estimated BCF of 159.4 L/kg (U.S. EPA.
2017) suggests that it is expected to have low bioaccumulation potential. EPA evaluated thirteen overall
high quality data sources reporting the aquatic bioconcentration, aquatic bioaccumulation, aquatic food
web magnification, and terrestrial bioconcentration of DBP (Table 7-1). The available data sources
discussed below suggest that DBP has low bioaccumulation potential in aquatic and terrestrial
organisms (Lee et al.. 2019a; U.S. EPA. 2017; Teil et al.. 2012). and no apparent biomagnification
across trophic levels in the aquatic food web (Mackintosh et al.. 2004).
Several overall high-quality data sources have reported the bioconcentration, bioaccumulation, and food
web magnification of DBP in aquatic species. One of these data sources reported DBP BCF values of
2.9 to 41.6 in sheepshead minnow, American oyster, and brown shrimp after a 24-hour exposure of DBP
(100-500 ppb) (Wofford et al.. 1981). suggesting low potential for bioconcentration in aquatic species.
This finding agrees with the predicted BCF values of 159.4 to 525 L/kg and monitored BCF values of
0.78 to 7.48 L/kg in fish, respectively (U.S. EPA. 2017; Adeogun et al.. 2015; Chemical Manufacturers
Association. 1984). BCF values of 1,500-5,000 have been reported in glass shrimp in a 3-day DBP
exposure experiment (Mayer Jr et al.. 1973); however, the DBP was rapidly excreted with a 75 percent
loss of DBP during a 7-day depuration period. A monitoring study reported BAF values of 110 to 1247
L/kg dwin crucian carp, skygager, bluegill, and bass samples collected from the Asan Lake in Korea
(Lee et al.. 2019a). The highest BAF value reported in Lee et al. (2019a) was 1,247 L/kg dw in crucian
carp. This species is a benthic feeder that generally tends to contain higher levels of phthalate esters due
to greater interaction with sediments. However, the available overall high-quality data sources
containing aquatic biota-sediment accumulation factors (BSAF), reported BSAF values of 0.35 to 11.8
giipid/goc for fish, and 130 giiPid/goc for oysters (Adeogun et al.. 2015; Teil et al.. 2012; Huang et al..
2008; Mcfall et al.. 1985). In addition, the available data sources reported aquatic trophic magnification
factor (TMF) values of 0.70-0.81 (Kim et al.. 2016; Mackintosh et al.. 2004). Despite the differences in
DBP biomonitoring values, DBP is expected to have low bioconcentration potential and low
biomagnification potential across trophic levels in the aquatic food web, but potentially result in higher
uptake by benthic organisms.
There is very limited information on the bioconcentration and bioaccumulation of DBP in terrestrial
environments. EPA extracted and evaluated nine high-quality data sources containing DBP terrestrial
plant concentration factors (PCFs) and biota-soil accumulation factor (BSAF) information for plants and
earthworms, respectively (Table 7-1). Based on DBP's expected strong affinity to organic matter and
rapid biodegradation (on the order of weeks to months in soil), DBP is expected to have limited
bioavailability in soils. This is supported by the reported low BSAF values of 0.242-0.460 in
earthworms (Eisenia foetida) (Ji and Deng. 2016; Hu et al.. 2005). Similarly, low PCF values have been
reported in the range of 0.02-9.60 for rice, fruits, vegetables, wheat and maize, pond weed, and wetland
grasses. These findings suggest that DBP has a low uptake potential for most edible fruits, vegetables,
grasses, and weeds from soil. Therefore, DBP is expected to have low bioaccumulation potential and
biomagnification potential in terrestrial organisms.
Overall, the available data suggest that DBP is expected to have low bioaccumulation potential and low
biomagnification potential in aquatic and terrestrial organisms.
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1067 Table 7-1. Summary of DBP's Bioaccumulation Information
Endpoint
Value(s)
Details
Reference(s)
Overall
Quality
Ranking
159.4 L/kg
(fish)
Estimated steady-state BCF;
Arnot-Gobas method, fish
upper trophic level.
U.S. EPA (2017)
High
0.78-7.48 L/kg
(fish)
Experimental monitoring
sample collection in Nigeria.
Tested organisms: Tilapia
zillii, Hepsetus odoe,
Parctchanna obscura and
Chrysichthys nigrodigitatus,
Mormyrus rume, and a
decapod crustacean (African
river prawn, Macrobrachium
vollenhovenii).
Adeoaun et al.
(2015)
High
Aquatic
bioconcentration
factor (BCF)
11.7 (minnow),
21.1-41.6 (oyster),
2.9-30.6 (shrimp)
Experimental laboratory
exposure. Organisms Type:
(Small Fish) Sheepshead
minnow, Cyprinodon
variegains: American oyster,
Crcissostreci virginica: brown
shrimp, Penaens ctztecus
Wofford et al.
(1981)
High
500-6,600 (aquatic
invertebrates)
Experimental laboratory
exposure; BCF of 3,500-
6,600 (Midge larvae, 1-7
days); 2,200-5000 (Water
flea, 1-7 days); 1,700-6,500
(Scud, 1-7 days); 500-1,900
(Mayfly, 1-7 days); 1,000-
2,700 (Damselfly, 1-7 days);
1,500-5,000 (Glass shrimp,
1-3 days). 75% DBP loss
during 7-day depuration.
Maver Jr et al.
(1973)
High
525 (fish)
Predicted fish BCF
calculated from actual Kow
determinations: log BCF =
(0.542 x log Kow) +
0.124
Chemical
Manufacturers
Association (1984)
High
Aquatic
bioaccumulation
factor
(BAF)
110-1,247 L/kg
dw (fish)
Experimental, monitoring
study lakes in Korea.
Average concentration in
fish: 3.3-37.4 (ig/kg dw. Log
BAF: 2.0-3.1 L/kg dw.
Organisms Type: crucian
carp, skygager, bluegill, and
bass.
Lee et al. (2019a)
High
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Endpoint
Value(s)
Details
Reference(s)
Overall
Quality
Ranking
0.56-6.11 (fish)
Experimental monitoring
sample collection in Nigeria.
Tested organisms (see above
for details).
Adcoeun et al.
(2015)
High
130 (oyster)
Experimental monitoring
Lake Pontchartrain in New
Orleans, Louisiana.
Calculated from
concentration in oysters
divided by concentration in
sediment. Average: 570 ng/g
ww in oyster, Crassotrea
virginica.
Mcfall et al. (1985)
High
Aquatic biota-
sediment
accumulation
factor (BSAF)
5.5 to 11.8
glipid/g0C
(fish)
Experimental monitoring
sample collection from the
Orge river in France. Roach:
5.5 ±4.8, Chub: 6.0 ±2.3,
and Perch: 11.8 ±12.6;
BSAF Cbiota (ng/g)/Csediment
(ng/g)
Teiletal. (2012)
High
0.35 to 1.35
giipid/goc (fish)
Experimental monitoring in
17 out of 21 principal rivers
of Taiwan. BSAF =
(phthalate in fish/lipid
content in fish) / (phthalate
in sediment/organic carbon
in sediment) Organism type:
Oreochromis niloticus, Liza
subviridis, Acanthopagrus
schlegeli, Zctcco platypus and
Acrossocheilus paradoxus.
Huana et al. (2008)
High
Aquatic trophic
magnification
factor (TMF)
0.70
95% confidence interval
(lower and upper interval
0.40-1.23) of the reported
TMF values in the False
Creek food web species
including 3 phytoplankton, 1
zooplankton, 10
invertebrates, and 10 fish.
Kim et al. (2016)
High
0.70-0.81
Food-web magnification
factor of 0.70 to 0.81 in 18
marine species in the False
Creek food web.
Mackintosh et al.
(2004)
High
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Endpoint
Value(s)
Details
Reference(s)
Overall
Quality
Ranking
Terrestrial biota-
soil
accumulation
factor (BSAF)
0.242-0.460
Earthworm from agricultural
field in China; 0.23-30 (soil
1) and 0.18-0.23 (soil 2);
BSAF = 0.460.
Hu et al. (2005);
Ji and Dene (2016)
High
Plant
Concentration
Factor (PCF)
0.02-0.495 (rice)
Approx. 0.105-0.4 (root),
0.02-0.14 (stem), 0.1-0.495
(leaf), and 0.005-0.255
(grain)
Cai et al. (2017)
High
0.16-0.19 (radish)
PCF Value: 0.19 (shoot),
0.16 (root)
Cai et al. (2008)
High
1.38-9.60
(pondweed)
BCF Value: 4.43-8.04 L/kg;
Study length: 10 days; root
bioconcentration: 9.60 ± 0.8
(control; lower conc. in
found sediment) 1.75 ± 0.2
(spiked; higher conc. found
in sediment); stems and
leaves bioconcentration: 7.40
±0.5 (control; lower conc. in
sed) 1.38 ±0.1 (spiked;
higher conc. found in
sediment)
Chi and Gao (2015);
Wane (2014)
High
0.33-1.03 (winter
wheat and summer
maize)
Winter wheat PCF: 0.89 and
0.42 (reclaimed water), 0.80
and 0.33 (mixed water), 0.91
and 0.43 (ground water);
Summer maize PCF: 1.03
(reclaimed water), 0.94
(mixed water), 1.01 (ground
water)
Li et al. (2018)
High
0.26-4.78 (fruit
and vegetables)
Mean PCF Value:
Lettuce leaf: 0.26 ± 0.01;
strawberry leaf: 0.34 ± 0.08;
carrot leaf: 1.09 ±0.21;
lettuce root: 0.77 ± 0.09;
strawberry root: 2.61 ± 0.42;
carrot root 4.78 ± 0.59;
purchased from the Certified
Plant Growers in Temecula,
CA; Study length: 28 days.
Sun et al. (2015)
High
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Endpoint
Value(s)
Details
Reference(s)
Overall
Quality
Ranking
2.11-9.32 (wetland
grasses)
Root bioconcentration: 2.11-
9.32 Organisms: P. australis
and Tvpha orientalise root
systems collected. Study
length: 17 days
Wane and Chi
(2012)
High
1068
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1081
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8 OVERALL FATE AND TRANSPORT OF DBP
The inherent physical and chemical properties of DBP govern its environmental fate and transport.
Based on DBP's aqueous solubility, slight tendency to volatilize, and strong tendency to adsorb to
organic carbon, this chemical substance will be preferentially sorbed to sediments, soils, and suspended
solids in wastewater treatment processes. Soil, sediment, and sludge/biosolids are predicted to be the
major receiving compartments for DBP as indicated by its physical, chemical, and fate properties 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 areas where continuous releases
of phthalates occur, higher levels of phthalates in surface water can be expected, trending downward
distally from the point of release. This also holds true for DBP concentrations in both suspended and
benthic sediments. While DBP undergoes relatively rapid aerobic biodegradation, it is persistent in
anoxic or anaerobic environments (i.e., sediment, landfills) and like other phthalates, it is expected to
slowly hydrolyze under standard environmental conditions.
When released directly to the atmosphere, DBP 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 DBP may be elevated proximal to
sites of releases. Off-gassing from landfills and volatilization from wastewater treatment processes are
expected to be negligible in terms of ecological or human exposure in the environment due to DBP's
low vapor pressure. DBP (not sorbed to suspended particles) released to air may undergo rapid
photodegradation and it is not expected to be a candidate chemical for long range transport.
Under indoor settings, DBP in the air is expected to partition to airborne particles and have an extended
lifetime as compared to airborne DBP in outdoor settings. The available information suggests that
DBP's indoor dust concentrations are associated with the presence of phthalate-containing articles and
proximity to the facilities producing them (Wang et al.. 2013; Abb et al.. 2009). as well as daily
consumer activities that might introduce DBP-containing products into indoor settings (Dodson et al..
2017).
DBP has a predicted average environmental half-life of 14 days. DBP is expected to degrade rapidly in
situations where aerobic conditions are predominant and be more persistent under anoxic or anaerobic
conditions (e.g., in some sediments, landfills, and soils). In anaerobic environments, such as deep
landfill zones, hydrolysis is expected to be the most prevalent process for the degradation of DBP.
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9 Weight of the 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 robust confidence that DBP:
• is not expected to undergo significant direct photolysis, but will undergo indirect
photodegradation by reacting with hydroxyl radicals in the atmosphere with a half4ife of 1.13 to
1.15 days (Section 4.3);
• will partition to organic carbon and particulate matter in air (Section 5.1);
• will not hydrolyze under standard environmental conditions, but its hydrolysis rate increases
with increased pH and temperature in deep-landfill environments (Sections 4.2 and 5.3.3);
• will biodegrade in aerobic surface water, soil, and wastewater treatment processes (Sections 4.1
and 6.2);
• will not biodegrade under anoxic conditions and may have high persistence in anaerobic soils
and sediments (Sections 4.1.2 and 4.1.3);
• will be removed with wastewater treatment and will sorb significantly to sludge, with a small
fraction being present in WWTP effluent (Section 6.2);
• has low bioaccumulation potential (Section 1);
• may be persistent in surface water and sediment proximal to continuous points of release
(Section 5.2); and
• is expected to transform to MBP, butanol, and phthalic acid in the environment (Section 1).
As a result of limited studies identified, there is moderate confidence that DBP:
• will 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 with a removal efficiency of 31 to 64.5 percent (Section 6.3).
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