EPA
United States EPA-822-R-16-009
Environmental Protection December 2016
Agency
State of the Science White Paper
A Summary of Literature on the Chemical Toxicity of
Plastics Pollution to Aquatic Life and
Aquatic-Dependent Wildlife
Photographs courtesy of the National Oceanic and
Atmospheric Administration (http://marinedebris.noaa.gov/)
U.S. Environmental Protection Agency
Office of Water
Office of Science and Technology
Health and Ecological Criteria Division
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Authors (alphabetical)
Joe Beaman, Office of Water, Office of Science and Technology, Washington DC (co-lead)
Christine Bergeron, Office of Water, Office of Science and Technology, Washington DC (co-lead)
Robert Benson, Office of Water, Office of Watersheds Oceans and Wetlands, Washington DC
Anna-Marie Cook, EPA Region 9, Superfund, San Francisco, CA
Kathryn Gallagher, Office of Water, Office of Science and Technology, Washington DC
Kay Ho, Office of Research and Development, Atlantic Ecology Division, Narragansett, Rl
Dale Hoff, Office of Research and Development, Mid Continent Ecology Division, Duluth MN
Susan Laessig, on detail to the Office of Water. Home office: Office of Chemical Safety and Pollution
Prevention, EPA HQ, Washington DC
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Notices
This document provides a summary of the state of the science on the potential chemical toxicity of
ingested plastic and associated chemicals on aquatic organisms and aquatic-dependent wildlife. While
this document reflects EPA's assessment of the best available science for plastics pollution, it is not a
regulation and does not impose legally binding requirements on EPA, states, tribes, or the regulated
community, and might not apply to a particular situation based upon the circumstances. EPA may change
this document in the future. This document has undergone contractor-led external peer review as well
as a review process within the EPA. Final review by EPA's Office of Science and Technology, Health and
Ecological Criteria Division has been completed and the document has been approved for publication.
Mention of trade names or commercial products does not constitute endorsement or recommendation
for use. This document can be downloaded from: https://www.epa.gov/wqc/aquatic-life-ambient-water-
qualitv-criteria.
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Table of Contents
Figures 6
Tables 7
Acronyms 8
Executive Summary 10
1 Introduction 13
2 Background on the Chemicals Associated with Plastics 14
2.1 Common Plasticizers and Polymer Additives 14
2.2 Properties of Plasticizers and Other Polymer Additives 15
2.3 Sorption of Chemicals to Plastics in the Aquatic Environment 17
3 Sources and Transport of Plastics to the Aquatic Environment 22
3.1 Types, Quantities, and Abundance of Plastics in the Aquatic Environment 22
3.2 Key Sources of Plastics 25
3.3 Transport Mechanisms 28
3.4 General Occurrence and Accumulation in the Aquatic Environment 28
4 Toxicological Impacts of Chemicals Associated with Plastics on Aquatic Organisms and Aquatic-
Dependent Wildlife 31
4.1 Routes of Exposure 31
4.2 Bioaccumulation and toxicological effects of chemicals associated with plastics 32
5 Conclusions and Research Needs 36
6 References 39
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Figures
Figure 1. Concentrations of polychlorinated biphenyls (PCBs) in plastic resin pellets on beaches (ng/g -
pellet). Reprinted with permission from Dr. Takada (IPW) 20
Figure 2. Concentrations of dichlorodiphenyltrichloroethane (DDT) and its metabolites (DDE:
Dichlorodiphenyldichloroethylene, and DDD: Dichlorodiphenyldichloroethane) in plastic resin pellets on
beaches (ng/g - pellet). Reprinted with permission from Dr. Takada (IPW) 20
Figure 3. Concentration of total polycyclic aromatic hydrocarbons (PAHs) in plastic resin pellets on beaches
(ng/g - pellet). Values with an asterisk (*) are less than the limit of quantitation. Values with a double
asterisk (**) are a median of only one pool of 200 pellets as opposed to the median of five pools. Reprinted
with permission from Dr. Takada (IPW) 21
Figure 4. Types of plastics commonly found in the aquatic environment (Pruter, 1987; U.S. EPA, 1993;
Andrady, 2011) 23
Figure 5. Key sources of plastics and their modes of transport to the aquatic environment. (U.S. EPA, 1993;
Rayne, 2008; Gregory, 2009; Cole et al., 2011; Lambert et al., 2014) 26
Figure 6. Illustration of the North Pacific Ocean central gyre (NOAA, 2014) 30
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Tables
Table 1. Selected physical properties of several plastic additives 16
Table 2. Selected physical properties of several persistent, hydrophobic organic contaminants known to
sorb to plastics 18
Table 3. The number of species with records of ingestion as documented in the reviews by Laist (1997)
and Gall and Thompson (2015) 32
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Acronyms
Acronym Definition
BBP Benzylbutylphthalate
BPA Bisphenol A
CHAP Chronic Hazard Advisory Panel on Phthalates and Phthalate Alternatives
CPR Continuous Plankton Recorder
DBP Dibutylphthalate
DDE Dichlorodiphenyldichloroethylene
DDT Dichlorodiphenyltrichloroethane
DEHP Diethylhexylphthalate
DEP Diethylphthalate
DMP Dimethylphthalate
GESAMP Joint Group of Experts on the Scientific Aspects of Marine Environmental
Protection
IMDCC Interagency Marine Debris Coordinating Committee
IPW International Pellet Watch
JRC-IHCP Joint Research Centre, Institute for Health and Consumer Protection
(European Commission)
Kow Octonol-Water Partitioning Coefficient
Kpw Polymer- Water Partitioning Coefficient
NOAA National Oceanic and Atmospheric Administration
NP Nonylphenol
NPE Nonylphenol exthoxylate
OCP Organochlorine Pesticide
PAH Polycyclic aromatic hydrocarbon
PBDE Polybrominated diphenyl ether
PBT Persistent, Bioaccumulative, and Toxic
PCB Polychlorinated biphenyl
PET Polyethylene terephthalate
PFOA Perfluorooctanoic acid
PFOS Perfluorooctanesulfonic acid
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Acronym Definition
PTFE Polytetrafluoroethylene
PUR Polyurethane
PVC Polyvinyl chloride
SCBD Secretariat of the Convention on Biological Diversity
UNEP United Nations Environmental Programme
U.S. EPA United States Environmental Protection Agency
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Executive Summary
The purpose of this report is to synthesize the state of the science on the potential chemical toxicity of
ingested plastic and associated chemicals on aquatic organisms and aquatic-dependent wildlife. The focus
of this document is primarily on marine systems, with data provided on the Great Lakes and other
freshwater systems, where available.
Since mass-production of plastics began in the 1940s and 1950s, the amount of plastic debris entering
marine and freshwater ecosystems has increased by several orders of magnitude (Cole et al., 2011).
However, recently the accumulation and potential impacts of plastic pollution has been recognized as an
emerging environmental issue (GESAMP 2015; UNEP 2016). Recent estimates suggest that 4.8 to 12.7
million metric tons of plastic waste entered the global marine environment in 2010 (Jambeck et al., 2015).
Plastics, including bags, packing materials, water bottles, and fishing line and their breakdown products
are now found throughout marine ecosystems and in the Great Lakes and other freshwater ecosystems,
including near urban and remote beaches, in the open ocean, in sediments, within the water column, and
in Arctic Sea ice (Derraik, 2002; Law & Thompson 2014; Eerkes-Medrano et al., 2015). Plastic particles are
generally the most abundant type of debris encountered in the marine environment with estimates
suggesting that plastics comprise between 60% and 80% of total marine debris (Derraik, 2002).
Plastics found in the aquatic environment are generally categorized as macroplastics (i.e., items > 5 mm
diameter, such as disposable cups, bottles, and shipping pallets) and microplastics (i.e., items < 5 mm
diameter, such as microbeads and fishing line fragments) (Lambert et al., 2014). As a subcategory of
microplastics, nanoplastics (<100 nm size range, as defined in Koelmans et al., 2015) are likely to occur
from both primary and secondary sources; methods do not currently exist to detect nanoplastics in the
natural aquatic environment (Koelmans et al., 2015). The abundance of microplastics (compared to
macroplastics) in the marine environment is increasing, according to a state-of-the-science report
conducted by Canada's Secretariat of the Convention on Biological Diversity (SCBD) (SCBD, 2012). Recent
investigations have found the majority (90%) of plastic debris found in the pelagic environment (i.e., open
ocean) is generally less than 5 mm in diameter (Eriksen et al., 2013; Browne et al., 2010; Thompson et al.,
2004; Rochman et al., 2014b).
The harmful physical impacts (i.e., entanglement, smothering, or physical effects of ingestion) of plastic
to aquatic invertebrates, fish, seabirds, sea turtles, and marine mammals has been well-documented
(Kiihn et al., 2015). In the aquatic environment, the ingestion of plastics also establishes a potential
exposure pathway for other chemical contaminants including metals, and persistent, bioaccumulative,
and toxic contaminants that may be sorbed from the water column to plastic or incorporated into the
plastics during manufacture (Engler 2012). Given the potential for plastics to be a source of contaminants,
from both the chemical constituents of the manufactured plastic itself and contaminants sorbed to
plastics in the aquatic environment, there is growing concern about the toxicological impacts of chemicals
associated with plastics on aquatic organisms, as well as, aquatic-dependent wildlife, such as seabirds
(Teuten et al., 2009).
Because plastics have become pervasive in oceans, coasts, and inland watersheds, and there are concerns
about the potential toxicological impacts of chemicals associated with plastics on aquatic organisms and
aquatic-dependent wildlife, the United States Environmental Protection Agency's (EPA) Office of Water
produced a state-of-the-science review that summarizes available scientific information on the effects of
chemicals associated with plastic pollution and the potential impact of these chemical on aquatic life and
aquatic-dependent wildlife. Furthermore, EPA's Office of Water is committed to reducing and preventing
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aquatic debris from entering U.S. waterways and the ocean. The Agency works toward this goal through
the Trash Free Waters Program, https://www.epa.Rov/trash-free-waters. which builds upon the
foundation of earlier Agency work through the Marine Debris Program. The Trash Free Waters program
has been designed with a strong emphasis on working with government, business, and citizen
stakeholders to reduce and prevent plastic trash and debris from entering both freshwater and coastal
ecosystems. The EPA's Trash Free Waters Program is intended to be a catalyst for proactive trash
prevention and reduction, using tools including research, public outreach and education, public/private
partnerships, support for state and local trash prevention efforts, and engagement with international
trash free ocean programs.
Scientific research including field studies (e.g., Yamashita et al, 2011; Lavers et al., 2014; Rochman et al.,
2014b) and laboratory studies (e.g., Teuten et al., 2009; Besseling et al., 2013; Rochman et al., 2013a)
suggests that several groups of aquatic or aquatic-dependent organisms (invertebrates, fish, and birds)
can accumulate chemicals associated with plastics once ingested. Experimental studies investigating the
effects of chemicals associated with plastics on invertebrates and fish indicate that there are negative
sublethal effects on these organisms from chemicals associated with plastics as well as the plastic itself
(e.g., Rochman et al., 2013a, 2014c; Avio et al., 2015). However, some bioaccumulation modeling
approaches attempting to simulate environmentally realistic scenarios of exposure provide indirect
evidence that the role of plastics in contributing to body burdens and effects of chemical pollutants may
be relatively small compared with other exposure pathways, such as direct chemical exposure via water,
sediment, or ingestion of contaminated prey (Koelmans et al., 2016; Bakir et al., 2016, Ziccardi et al.,
2016). As identified in this report, there are significant opportunities for research to further our
understanding of the actual impacts of the potential chemical toxicity of plastic ingestion throughout the
food web.
Further research is needed to gain knowledge of the extent to which plastics transfer contaminants to
organisms compared to other sources, as well as the toxicological impacts of plastic ingestion compared
to other environmental stressors, in particular the following:
1. Studies to gain a better understanding of the fate of chemicals both sorbed to and in plastics
under differing environment conditions and within an organism after ingestion.
2. Research on the relative role plastics play in chemical contaminant transfer to the tissues of
organisms compared to other exposure pathways (aqueous dermal exposure and ingestion from
natural prey).
3. Research is needed to understand the relative impacts of physical and chemical effects of ingested
plastic particles on a wide range of organisms.
4. Research to determine whether the relatively high surface area of nanoplastics compared to
microplastics and their potential to permeate membranes with increased retention time may
increase their toxicological risk to organisms (Koelmans et al., 2015).
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1 Introduction
Modern plastics were first introduced in the early 20th century, with the production of polyvinyl chloride
(PVC) and polyethylene starting in the 1920s and 1930s. Mass-production of plastics began in the 1940s
and 1950s as part of the post-World War II increase in consumerism to include many of the plastics
commonly used today (e.g., polypropylene, polystyrene, polyethylene terephthalate (PET), polyurethane
(PUR), polycarbonate, and polytetrafluoroethylene (PTFE - Teflon®) coatings) (Barnes et al., 2009; Cole
et al., 2011; PlasticsEurope, 2008; Andrady and Neal, 2009; Lambert et al., 2014). While uses in the 1920s
and 1930s were more limited (e.g., a limited number of consumer goods, automobile parts, and military
equipment (Freinkel, 2011)), plastics are used in the 21st century in a wide array of products, including
packaging materials, water bottles, mobile telephones, computers, building insulation, medical devices,
protective clothing, piping systems for drinking water, and artificial limbs and joints (PlasticsEurope,
2008). World plastic production has increased dramatically from an estimated 1.7 million tons in 1950 to
311 million tons in 2014 (PlasticsEurope, 2013 and 2015). Coinciding with an increase in production, the
amount of plastics in the aquatic environment has been steadily increasing, and plastics and plastic
particles (i.e., microplastics, items < 5 mm diameter) are now commonly found in freshwater and marine
systems around the globe (Derraik, 2002; Eerkes-Medrano et al., 2015; Law & Thompson 2014; Wagner
et al., 2014).
Recent estimates suggest that 4.8 to 12.7 million metric tons of plastic waste entered the global marine
environment in 2010 (Jambeck et al., 2015). Areas of accumulation of plastic debris include enclosed
basins, ocean gyres, and bottom sediments, including the deep ocean (Moore et al., 2001; Ryan et al.,
2009; Collignon et al., 2012; Shimanaga & Yanagi, 2016). Plastics in the aquatic environment primarily
originate from land-based sources such as littering, improper or ineffective solid waste management, and
wind-blown debris, though plastic debris from fishing activities may be a key source in some areas (Galgani
et al., 2000; Andrady, 2011; Lambert et al., 2014; Unger & Harrison, 2016). Plastic particles are generally
the most abundant type of debris encountered in the marine environment, with estimates suggesting that
60% to 80% of marine debris is plastic, and that more than 90% of all floating debris particles is plastic
(Derraik, 2002; Gordon, 2006).
Efforts to address the issues of debris (including plastics) in the marine environment in the U.S. began in
the mid-1980s (NOAA, 2008). Since that time, numerous federal initiatives and committees have been
formed to address the problem of debris in marine and freshwater aquatic environments. For example,
the EPA's Office of Water has been committed to reducing and preventing trash, litter, and debris from
entering U.S. waterways and the ocean through the current Trash Free Waters Program which builds upon
the foundation of earlier Agency work through the Marine Debris Program. The Trash Free Waters
program facilitates research, outreach, and education for state and local governments, the private sector,
and the general public on ways to reduce and prevent plastic and other anthropogenic debris from
entering the aquatic environment (U.S. EPA, 2016). The Interagency Marine Debris Coordinating
Committee (IMDCC), formed in 2004, is a multi-agency committee with the National Oceanic and
Atmospheric Administration (NOAA) as chair and EPA as vice-chair. The committee makes
recommendations for research priorities, monitoring techniques, educational programs, and regulatory
actions for reducing the sources and ecological impacts of debris (NOAA, 2008). In addition, NOAA's
Marine Debris Program is authorized by Congress to work on debris through the Marine Debris Act, signed
into law in 2006 and amended in 2012, and focuses on research, removal and prevention efforts including
education and outreach such as an online podcast about the "garbage patch" in the Pacific Ocean (NOAA,
2014) and a series of online videos called "Trash Talk" (NOAA, 2015a).
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Recently the accumulation and potential impacts of plastic pollution has been recognized as an emerging
environmental issue (GESAMP 2015; UNEP 2016). There is growing concern regarding the potential
toxicological impacts caused by chemicals, in addition to the known physical impacts (i.e., entanglement,
smothering, or physical effects of ingestion) of plastic debris on aquatic life and ecosystems (e.g., Moore
et al., 2001; Arthur et al., 2009; Cole et al., 2011; Lavers et al., 2014). The purpose of this report is to
synthesize the state of the science on the potential chemical toxicity of ingested plastic and associated
chemicals on aquatic organisms and aquatic-dependent wildlife. The focus of this document is primarily
on marine systems, with data provided on the Great Lakes and other freshwater systems where available.
Recognizing that a variety of organic chemicals and metals found in the water column, sediments, and
atmosphere may sorb to plastic debris, this paper focuses on the most prevalent and potentially
hazardous contaminants currently known to be associated with plastics. This paper is organized as follows:
• Section 2 provides background information on the chemicals associated with plastics and their
properties.
• Section 3 describes key sources of plastics and their modes of transport to and within the aquatic
environment. This section also presents information on the general occurrence and abundance of
plastic debris in the aquatic environment.
• Section 4 describes the bioaccumulation and toxicological effects that chemicals associated with
plastics (i.e., their additives, as well as sorbed contaminants) may have on aquatic organisms and
aquatic-dependent wildlife such as seabirds.
• Section 5 provides conclusions and recommendations for future research.
2 Background on the Chemicals Associated with Plastics
Aside from the physical impacts plastics have on aquatic organisms (reviewed in Kiihn et al., 2015), plastics
may play a role in transporting chemicals that are associated with plastics within the aquatic environment
into the food chain. Key contaminants detected in plastics in the aquatic environment include: phthalates,
polychlorinated biphenyls (PCBs), polycyclic aromatic hydrocarbons (PAHs), organochlorine pesticides
(OCPs), polybrominated diphenyl ethers (PBDEs), alkylphenols, bisphenol A (BPA), and metals (e.g.,
cadmium, zinc, aluminum). Because plastic debris can persist in the aquatic environment for a long time,
plastic may be a source of chemicals to the aquatic environment if the ingredients used in making the
plastic (e.g., monomers and additives) leach into the surrounding waters and may be a sink for chemicals
that may accumulate on plastics from the surrounding aquatic environment (e.g., persistent,
bioaccumulative, and toxic (PBT) contaminants) (Engler 2012). These two categories of chemicals
associated with plastics are described in more detail below.
2.1 Common Plasticizers and Polymer Additives
While plastics are named for their primary monomeric ingredients, plastics typically contain additives that
modify the properties of the pure polymers to increase pliability, resist ultraviolet radiation, reduce
flammability or degradation, or impart other preferred physical characteristics to the finished product
(Andrady and Neal, 2009; Lambert et al., 2014). Additives can leach from the plastic to the surrounding
environment and as the plastic fragments and weathers, more chemicals are able to leach (Engler, 2012).
Due to the large numbers of plasticizers and flame retardants associated with plastics, only the common
and well-studied groups of these compounds are evaluated in this paper and discussed individually below.
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Plasticizers are widely used to impart pliability to prevent shattering, with phthalates and BPA among the
most common plasticizers. A variety of phthalates are used in plastics (See review by Staples et al. (1997)
regarding the properties of 18 phthalates, most or all of which may be found in plastics). Some of the
phthalates that were widely used in plastics in the past include dibutylphthalate (DBP),
diethylhexylphthalate (DEHP), dimethylphthalate (DMP), and benzylbutylphthalate (BBP) (Oehlmann et
al., 2009). Phthalates often comprise a substantial portion of polymeric materials; for example, phthalates
are reported to be present at up to 50% of the mass of finished PVC plastics (Oehlmann et al., 2009). BPA
is used as a monomer in the production of polycarbonate plastics and is also used as an antioxidant,
inhibitor, and stabilizer in PVC and other plastics (Andrady and Neal, 2009; Oehlmann et al., 2009;
Thompson et al., 2009; JRC-IHCP, 2010; Lambert et al., 2014). As a monomer in polycarbonates, BPA is a
primary component in this type of plastic. Both phthalates and BPA have been reported to accumulate in
organisms and to affect development and reproduction in a wide range of species, often at concentrations
found in the environment (Oehlmann et al., 2009). As a result, some phthalates have been banned from
use in the U.S. in children's toys and child care products as described in the Chronic Hazard Advisory Panel
on Phthalates and Phthalate Alternatives (CHAP) report (CHAP, 2014), but these banned phthalates
remain in use for other plastic applications.
Flame retardants, such as PBDEs, are commonly used in plastics (Ueno et al., 2004). PBDEs have been
reported to comprise 5-30% of the polymers, resins, and similar substrates in which they are used
(Darnerud et al., 2001). PBDEs are typically grouped by the number of bromine atoms they contain. There
are 209 possible congeners of PBDEs (i.e., various substitution patterns of varying numbers of bromine
atoms on the diphenyl ether structure); however, commercial products used as flame retardants primarily
contain penta-, hepta-, octa-, and decabromodiphenyl ethers (Darnerud et al., 2001). Toxicological studies
with both animals and humans have demonstrated that PBDEs are potential carcinogens, neurotoxins,
and endocrine disruptors (Akortia et al., 2016). Primarily due to their extreme persistence in the
environment and their potential human toxicity, EPA has phased out production of pentabromodiphenyl
ethers and octabromodiphenyl ethers (U.S. EPA, 2006) and decabromodiphenyl ethers (U.S. EPA, 2010).
Alkylphenols, such as nonylphenol (NP), are also commonly added to plastics. NP is a degradation product
of nonylphenol ethoxylates (NPEs) which are produced in large volumes and used in a wide variety of
industrial applications and consumer products (Engler, 2012; Soares et al., 2008). NP is associated with
plastic debris as a result of both the manufacturing process and from re-adsorption of the chemical in the
environment (Mato et al., 2001). NP is persistent in the aquatic environment, moderately
bioaccumulative, and has been reported to be an endocrine disrupter that is toxic to aquatic organisms
(Soares et al., 2008).
2.2 Properties of Plasticizers and Other Polymer Additives
The octanol-water partitioning coefficient (log Kow) represents the ratio of how much of a compound
partitions to the organic solvent octanol relative to water in an octanol-water system, and is a useful
parameter for assessing the potential biological partitioning of certain chemicals to lipid-rich tissues.
Lipophilic/hydrophobic compounds have a greater affinity for octanol relative to water. Larger log Kow
values thus indicate a greater affinity for non-polar materials (e.g., plastics and lipid membranes) and less
of an affinity for water.
Water solubility and log Kow data for several plastic additives are summarized in Table 1, and those for
several known PBT compounds are summarized in Table 2 in the next section. It should be noted that a
review of physical property data in the literature reveals wide ranges of values for a single compound or
mixture of compounds. The solubility and log Kow values included in Table 1 and 2 were selected as
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generally representative of these compounds. While not intended as definitive values for water solubility
and log Kow, the values in Table 1 and 2 are intended to highlight the general relationship between water
solubility and log Kow. Also, reference temperatures are not available for most of the water solubilities
listed in Table 1 and 2, which adds some uncertainty to how the data represent the relationship of water
solubility and log Kow because this property is temperature-dependent.
Comparing the log Kow values in Table 1 to those for known PBT compounds in Table 2 (in particular,
polychlorinated biphenyls (PCBs), polycyclic aromatic hydrocarbons (PAHs), and
dichlorodiphenyltrichloroethane (DDT)) indicates that many polymer additives have log Kow values that
are similar to those of known PBT compounds.
Table 1. Selected physical properties of several plastic additives.
Chemical
Water Solubility (|ig/L)
(Reference temperature noted where specified)
lOgKow
(dimensionless)
Reference
Dimethylphthalate
4,200,000
1.61
1
di-n-Butylphthalate
11,200
4.45
1
di-n-Octylphthalate
0.5
8.06
1
b/s-(2-Ethylhexyl)phthalate
3
7.50
1
Bisphenol A
120,000 - 300,000*
3.40
2
Tetrabromodiphenyl
ethers
Not reported
5.9-6.2
3
Pentabromodiphenyl
ethers
0.0009 @20°C
6.5-7.0
3
Octabromodiphenyl ethers
Not reported
8.4-8.9
3
Decabromodiphenyl ether
Data considered unreliable
10
3
Perfluorooctanoic acid
(PFOA)
9.50 x 10s @ 25 °C
Not measurable
4
Perfluorooctane sulfonic
acid (PFOS)
680,000
Not measurable
5
References:
1. Staples et al. (1997). Staples et al.-recommended values from ranges presented are used in Table 1.
2. Staples etal. (1998)
3. Darnerud et al. (2001)
4. ATSDR (2015)
5. ATSDR (2015), OECD (2002)
Notes:
* The water solubility of BPA is reported to increase with increasing pH due to two ionizable hydroxyl groups (Staples et al., 1998)
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2.3 Sorption of Chemicals to Plastics in the Aquatic Environment
Plastics can accumulate chemical contaminants from the surrounding aquatic environment due to their
physical properties, such as size, shape and surface area, and the chemical properties of the specific
polymer (GESAMP, 2015). Adsorbed chemicals have been found on plastic debris samples globally and
include metals as well as compounds categorized as PBT by the EPA, including PCBs, PAHs, and DDT (Mato
et al., 2001; Teuten et al., 2007, 2009; Rochman et al., 2013c; Holmes et al., 2012; Engler, 2012). These
chemicals resist degradation and persist in the aquatic environment. The longer a plastic particle remains
in the aquatic environment, the more concentrated the contaminants can become as they accumulate on
the particle surface over time (Carpenter et al., 1972; Mato et al., 2001; Lavers et al., 2014). However, the
process of chemical contaminant sorption and desorption is dynamic and depends on various factors, the
most important being the presence of a concentration gradient (Koelmans et al., 2013), but also other
factors including temperature, pH, and salinity (Engler, 2012). In addition, fragmentation, degradation,
and weathering, which occur over time in the aquatic environment, increase the surface area-to-volume
ratio of the plastic (Bernstein and Woods, 2009). These processes can result in an increase in the relative
concentration of sorbed contaminants and also allow for desorption of accumulated contaminants into
the water (Bernstein and Woods, 2009; Engler, 2012). For example, tests conducted on PVC, polyethylene,
and polypropylene indicate that exposure of the plastics to ultraviolet light, simulating weathering
conditions, increases the sorbtive capacity of these plastics relative to virgin plastics (Teuten, et al., 2009).
2.3.1 Hydrophobic Compounds in the Environment that May Sorb to Plastics
Hydrophobic compounds (e.g., PCBs, PAHs, OCPs) tend to have low water solubility (Table 2) and partition
from the water column to other matrices with similar hydrophobic properties (Engler, 2012). For this
reason, when in the aquatic environment, they tend to partition to organic matrices or to plastic debris
(Engler, 2012). The ability of hydrophobic compounds to partition to plastics is well documented; because
of their sorptive capacity some plastics have been developed as passive sampling devices to assess the
concentrations of hydrophobic compounds in different environmental media (Lohmann, 2012). In
addition, a recent review by Ziccardi et al. (2016) compiled polymer-water portioning coefficients (Kpw)
published for plastics and chemical combinations (see Table 1 in Ziccardi et al., 2016) and noted that
previous studies have reported that Kpw values are correlated with Kow values.
Table 2 summarizes water solubility and log Kow values for several persistent and hydrophobic organic
environmental contaminants known to sorb to plastics (Rochman et al, 2013a; Rochman et al, 2013b;
Rochman et al, 2014a; Mato et al., 2001; Engler, 2012). Other properties affecting sorption and
persistence may be of relevance (e.g., photolysis, hydrolysis and biodegradation); however, these other
properties and their relationships to persistence of a chemical in the environment require further
research.
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Table 2. Selected physical properties of several persistent, hydrophobic organic contaminants known
to sorb to plastics.
Chemical
Water Solubility (|ig/L)
@ 25°C unless noted
lOgKow
(dimensionless)
Reference
Aroclor 1016 (PCBs)*
225-250
5.6
1
Aroclor 1242 (PCBs)
240
5.6
1
Aroclor 1248 (PCBs)
54
6.2
1
Aroclor 1254 (PCBs)
12
Not reported
1
Aroclor 1260 (PCBs)
27
6.8
1
Benzo(a)pyrene
3.8
6.11
1
Phenanthrene
600+100 @22°C
4.55
1
p,p'-Dichlorodiphenyltrichloroethane
(DDT)
3.1-3.4
6.83
1
4,600 @ pH 5.0
Nonylphenol (branched and straight-chain)
6,237 @ pH 7.0
11,897 @ pH 9.0
3.80-4.77
2
References:
1. U.S. EPA (2000)
2. U.S. EPA (2005a)
Notes:
* Aroclors contain a four-digit numeric code that denotes the 12 carbons contained in the biphenyl molecule followed by two
additional digits that reflect the average weight percentage of chlorine in the congeners in the aroclor group (the exception
being aroclor 1016, whose congeners contain, on average, 41% chlorine) (ATSDR, 2000).
The quantity and rates at which persistent compounds sorb to plastics differ by many factors including
plastic type and length of time in the aquatic environment. A long-term (12-month) study on contaminant
sorption to plastics suggests that PCBs and PAHs sorbed more readily to high- and low-density
polyethylene and polypropylene compared to PET and PVC (Rochman et al., 2013b). In addition, a six-day
field experiment on polyethylene plastic resin pellets in marine water off the coast of Japan revealed a
significant increase in PCB and dichlorodiphenyldichloroethylene (DDE) accumulation in the pellets over
this short period of time via sorption from seawater (Mato et al., 2001).
The types and concentrations of hydrophobic chemicals detected in and on plastics collected from the
aquatic environment can differ across geographic locations. For example, several PCB congeners, PAHs,
DDT and its metabolites, PBDEs, and alkylphenols were detected at concentrations ranging from 1 to
10,000 ng/g in samples collected in the open ocean, as well as at remote and urban beaches (Hirai et al.,
2011). PCBs were detected in all plastics sampled, with concentrations ranging from 1 to 436 ng/g, with
generally higher concentrations found for polyethylene compared to polypropylene debris (Hirai et al.,
2011). Levels of PCBs and PAHs were higher for plastic debris found on urban beaches compared to those
found in the open ocean and remote beaches, as may be expected due to the impacts of local land-based
sources of those contaminants (Hirai et al., 2011). In a study conducted along the Portuguese coastline,
levels of PCBs and PAHs were orders of magnitude higher in plastic pellets collected near urban coastal
areas compared to rural sites, suggesting that analysis of these contaminants in pellets should be
conducted across regional scales to understand spatial trends in accumulation and source (Mizukawa et
al., 2013). Measured concentrations of PCBs ranged from 10.5 to 307 ng/g-pellet, and PAH levels were
Page 18 of 50
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generally within the 100 to 300 ng/g-pellet range (Mizukawa et al., 2013). In plastic samples collected in
the North Pacific Ocean gyre, PCBs, pesticides, and PAHs were prevalent and were measured in more than
50%, 40%, and nearly 80% of the samples, respectively (Rios et al., 2010). PCBs were produced in large
quantities after the Second World War, and while PCB production and new uses have largely been phased
out in developed countries, legacy uses and contamination remains an issue (Tanabe, 2002) and these
hydrophobic compounds are still used in the U.S. and some developing countries. Their persistence and
wide distribution in the environment add to the likelihood that marine ecosystems are the primary sink
for PCBs (Tanabe, 2002). Another class of compounds, alkylphenols, which may be used as additives during
the plastics manufacturing process are also often detected in wastewater discharges and can also sorb to
plastics in the water column (Hirai et al., 2011).
To monitor the global pattern of hydrophobic organic contaminants found on beaches around the world,
International Pellet Watch (IPW) was launched in 2005. Volunteers collect pellets along beaches and send
them to Tokyo University of Agriculture and Technology for analysis. Results from analyses are shown for
PCBs, DDT, and PAHs in Figure 1-3 respectively (reprinted with permission); methods are presented in
Mato et al. (2001). In the IPW's results, concentrations of PCBs on beached plastic resin pellets were
reported to be highest along the coasts of France (2,970 ng/g-pellet), followed by beaches in Los Angeles,
California (602 ng/g-pellet), the shoreline of Lake Erie (502 ng/g-pellet), the Japanese coastline (453-499
ng/g- pellet), and Boston, Massachusetts (405 ng/g- pellet) (Figure 1). Total PCB values represent the sum
of specific tetra, penta-, hexa-, hepta-, and nonachlorobiphenyl congeners (congeners 66, 101, 110, 149,
118, 105, 153, 138, 128, 187, 180, 170 and 206). This list includes two co-planar (i.e., dioxin-like) PCBs,
congeners 105 and 118, and 11 non-co-planar PCBs. Levels of DDT and its metabolites were reported to
be highest in beached pellets in Albania (1,061 ng/g-pellet), Brazil (777 ng/g-pellet), Vietnam (558 ng/g-
pellet), and Japan (299 ng/g-pellet) (Figure 2). The higher levels of DDT found along Vietnam coastlines
may be attributed to the continued use of DDT as a pesticide (Teuten et al., 2009). Concentrations of total
PAHs were highest in beached pellets in Europe (24,364 ng/g-pellet and 15,608 ng/g-pellet reported on
beaches in Portugal and the United Kingdom, respectively), and in New Zealand (13,278 ng/g-pellet)
(Figure 3). Importantly, these results are based on analyses of pellets submitted by volunteers around the
world. Therefore, the reported numbers of pellets across beaches is largely a function of how many pellets
are submitted for analysis, and the ability to accurately reflect geographic contaminant trends based on
the Tokyo University analyses is also closely tied to the extent of volunteer participation in a geographical
region.
Hydrophobic compounds can readily sorb to plastic debris in the aquatic environment, but it is also
important to compare concentrations on plastics to other environmental matrices. Studies have found
hydrophobic compounds can accumulate on plastic debris at concentrations up to six orders of magnitude
greater than the surrounding water (Ogata et al., 2009) and up to two orders of magnitude greater than
sediment and suspended particles (Mato et al., 2001; Teuten et al., 2007). However, assuming equilibrium,
a thermodynamically-based model predicted that with the current concentrations of plastic debris in the
ocean, the fraction of hydrophobic compounds sorbed to the plastic is of limited importance relative to
other exposure media, such as direct chemical exposure via water, sediment, or ingestion of contaminated
prey (Gouin et al., 2011; Koelmans et al., 2016).
Page 19 of 50
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2970
France
281
[[Netherlands
Belgium
Albania
Portugal
Ghana
453
499
502
Turkey
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53
139 i?1
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17 13i?,
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62
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r
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64
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Zealand
0.01
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.344
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p.oPzil
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Figure 1. Concentrations of polychlorinated biphenyls (PCBs) in plastic resin pellets on beaches (ng/g-
pellet). Reprinted with permission from Dr. Takada (IPW,
http://www.tuat.ac.jp/~gaia/ipw/en/map.html).
Albania1^61
\ 262
86
I
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|!6
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1 , |
iweden ¦
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¦ « T| 11 HK ~ "
.. 5 11 China Taiwan
lia , /Vietnam ~
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Kenya Malaysia Wl_Philippines
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9 j 72
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New Zealand
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55
Iff
Baltimore
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l? Boston
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£ 3
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Figure 2. Concentrations of dichlorodiphenyltrichloroethane (DDT) and its metabolites (DDE:
Dichlorodiphenyldichloroethylene, and DDD: Dichlorodiphenyldichloroethane) in plastic resin pellets
on beaches (ng/g- pellet). Reprinted with permission from Dr. Takada (IPW
http://www.tuat.ac.jp/~gaia/ipw/en/map.html).
Page 20 of 50
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15608
243m[
U.K.
312 Portugal
F364 4^165
m 2004 Gre^
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:
93*^ Mozambique
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0
Australia
New Zealand
Figure 3. Concentration of total polycyclic aromatic hydrocarbons (PAHs) in plastic resin pellets on
beaches (ng/g - pellet). Values with an asterisk (*) are less than the limit of quantitation. Values with
a double asterisk (**) are a median of only one pool of 200 pellets as opposed to the median of five
pools. Reprinted with permission from Dr. Takada (IPW
http://www.tuat.ac.jp/~gaia/ipw/en/map.html).
2.3.2 Metals
Metals can be added to plastics during the manufacturing process (Nakashima et al., 2012). However,
metals are also found in the water column and sediments from a variety of other sources, and they can
also sorb to plastic debris. While much research has been conducted on the sorption of persistent,
hydrophobic compounds to plastics, the metal accumulation on plastics and potential impacts is a newer
area of research (e.g., Ashton et al., 2010; Holmes et al., 2012; Nakashima et al., 2012; Rochman et al.,
2014a). Metal ions, such as aluminum, antimony, cadmium, chromium, cobalt, copper, iron, lead,
manganese, molybdenum, tin, uranium, and zinc, have also been shown to sorb to plastics in the aquatic
environment (Ashton et al., 2010; Holmes et al., 2012; Rochman et al., 2014a; Turner, 2016; Turner & Lau,
2016). Similar to hydrophobic compounds, metals may sorb more readily to plastic particles as the
particles become more degraded and porous in the aquatic environment (Ashton et al., 2010; Rochman
et al., 2014a; Rochman et al., 2014b). For example, following a year-long study in San Diego Bay on the
sorption of several metals on five different types of plastic, the authors reported that sorption of these
metals increased with time and that similar concentrations of each metal were observed in the differing
plastic types (Rochman et al., 2014a). In addition, concentrations of metals on plastics and in nearby
sediment have been found to be similar (Holmes et al., 2012).
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3 Sources and Transport of Plastics to the Aquatic Environment
3.1 Types, Quantities, and Abundance of Plastics in the Aquatic Environment
Plastic debris is ubiquitous in the aquatic environment (Thompson et al., 2009). While plastics in the
marine environment have received the most attention to date, investigations also indicate that plastics
readily accumulate in freshwater environments (Eerkes-Mdrano et al., 2015; Baldwin et al. 2016). Plastics
found in the aquatic environment are generally categorized as macroplastics (i.e., items > 5 mm diameter,
such as disposable cups, bottles, and shipping pallets) and microplastics (i.e., items < 5 mm diameter, such
as microbeads and fishing line fragments) (Lambert et al., 2014). Microplastics are further categorized
into primary and secondary sources where primary sources include manufactured products and secondary
sources result from the breakdown of macroplastics in the environment (GESAMP, 2015). As a
subcategory of microplastics, nanoplastics (<100 nm size range) are likely to occur from both primary and
secondary sources; methods do not currently exist to detect nanoplastics in the environment (Koelmans
et al., 2015). The abundance of microplastics (compared to macroplastics) in the marine environment is
increasing, according to a state-of-the-science report conducted by Canada's Secretariat of the
Convention on Biological Diversity (SCBD) (SCBD, 2012). Recent investigations have found the majority
(90%) of plastic debris found in the pelagic environment is generally less than 5 mm in diameter (Eriksen
et al., 2013; Browne et al., 2010; Thompson et al., 2004; Rochman et al., 2014b). Examples of some of the
common types of plastics typically found in the aquatic environment are shown in Figure 4.
Page 22 of 50
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Polyethylene
Microbeads/
pellets Buckets
Netting
Userjpags Construction
sheeting &
packaging film
Bubble wrap
Polystyrene
Apparel fibers
LI
Shipping pallets
Six-pack rings
Disposable cups
e
I lotation devices
Plastic utensils &
food containers_
Packaging peanuts
tern
Drinking straws
Polypropylene
Fishing line
Microbeads/
pellets
Bottle caps
Apparel fibers
Polyvinyl chloride (PVC)
Plumbing pipes
Electrical cable
insulation
Polyethylene
terephthalate (PET)
Bottles & carboys
.I^S
Polycarbonate
materials
Carboys
FT
Compact disks
Tt
Figure 4. Types of plastics commonly found in the aquatic environment (Pruter, 1987; U.S. EPA, 1993;
Andrady, 2011).
Page 23 of 50
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Plastic production continues to increase across the globe, and the volume of plastics entering the marine
environment has increased as much as three to four orders of magnitude since its mass production began
in the 1940s and 1950s (Thompson et al., 2004; Barnes et al., 2009; Ryan et al., 2009; Cole et al., 2011).
The presence of plastic debris in the ocean has been reported since the early 1970s, although recent
analyses of archived plankton samples collected around the United Kingdom from as early as the 1960s
revealed various types of plastic including polyethylene, polyester, and polypropylene, among others
(Thompson et al., 2004). Investigations conducted in US waters in the 1970s reported plastic debris in 62%
of surface plankton samples collected in the Atlantic Ocean from Cape Cod to the Caribbean (Colton et
al., 1974), as well as an abundance of polystyrene spherules (around one spherule per 0.03 cubic meters
on average) along the coastal waters of New England in 1971 (Carpenter et al., 1972). In the past five
decades, numerous studies have reported plastic debris in marine environments (see Table 3 in Lambert
et al., 2014) and the Great Lakes (e.g., Eriksen et al., 2013). In more recent surveys, Barnes et al., 2009
noted that average size of plastic particles in the environment seems to be decreasing, correlated to
breakdown of macroplastic debris and the increasing use of plastic scrubbers in cosmetics, personal
hygiene products, and cleaning products, and with the abundance and global distribution of microplastic
fragments increasing over the last few decades. Overall, approximately 60% to 80% of marine debris is
estimated to be comprised of plastics (Derraik, 2002; Browne et al., 2010; SCBD, 2012), and because
plastics are so frequently encountered in the ocean, the Continuous Plankton Recorder (CPR) Survey (the
longest running plankton monitoring program in the North Sea and North Atlantic) added 'plastics' as its
first non-biological entity to their program in 2004 (Richardson et al., 2006).
The research on plastics in the marine environment is more advanced than the freshwater environment
(Wagner et al., 2014) despite the fact that a majority of marine plastics are thought to originate from land-
based sources (see Section 3.2). Current studies suggest that plastics are also pervasive in freshwater
environments (Eerkes-Medrano et al., 2015). For example, a recent survey of floating plastic debris in 29
Great Lakes tributaries found plastics in all 107 samples with 98% of the plastics considered microplastics.
Most types of plastics (except fibers, the most frequently detected particle type) were positively
correlated with urban-related watershed influences (Baldwin et al., 2016). In addition, some studies have
found concentrations of microplastics in lakes and rivers similar to or higher than in oceanic gyres (e.g.,
Yonkos et al, 2014; Lechner et al., 2014; Mani et al., 2015). Freshwater environments may be similar to
marine environments in terms of sources and transport mechanisms, prevalence, and potential for
impacts to organisms, but the differences in waterbody size and the proximity to point sources may result
in different plastic composition (Eerkes-Medrano et al., 2015).
Estimating the abundance of plastic debris in the aquatic environment is complicated by a variety of
factors, including the vastness of the ocean compared to the often micro-sized plastic debris that may be
floating along the surface, mixed within the water column, or incorporated in the sediment or beach sands
(Ryan et al., 2009; Cole et al., 2011). Rates of degradation among plastic debris vary, with thicker types of
plastic expected to persist in the aquatic environment for potentially hundreds of years (Barnes et al.,
2009; Ryan et al., 2009). Ocean currents, weather patterns, and other transport mechanisms (see Section
3.3) alter the spatial distribution of plastic debris; therefore, a study conducted a decade ago may not
reflect the abundance or types of plastics that would currently be observed in the same location. Also, the
type of investigation and the methods employed (e.g., trawls vs. observational studies) can influence the
study results. For example, studies investigating plastics along the ocean surface likely provide
underestimates of the true abundance of plastic debris in the ocean, because the specific gravity of most
plastics is higher than that of seawater, causing them to sink (Ryan et al., 2009; Andrady, 2011; Woodall
et al., 2014). The buoyancy of plastic is dependent on the density of the material and presence of trapped
air; after some time, floating plastic may become fouled with organisms that increase the density and
Page 24 of 50
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cause it to sink (Andrady, 2011). In addition, current collection methods often use nets with a mesh size
of ~330 pirn, which is larger than many microplastic particles (Andrady, 2015). Also, open ocean sampling
necessitates large sample sizes to allow for statistical power to detect temporal and spatial changes in
abundance (Ryan et al., 2009). Reports from beach and shoreline clean-ups may also provide
underestimates of the abundance of plastics because many small microplastic fragments are not easily
discernable by the naked eye and can be the size of a grain of sand or planktonic organism (Ryan et al.,
2009; Andrady, 2011; Cole et al., 2011).
3.2 Key Sources of Plastics
3.2.1 Land-Based Sources
Plastic debris originates primarily from land-based activities that include intentional and accidental
disposal (Lambert et al., 2014; Figure 5). Plastics can enter the environment through many pathways
including littering of bags, bottles, and other plastic items, particularly at large public gatherings and
events, or from coastal tourism (Gregory, 2009). Plastic debris from uncovered trash receptacles or
inadequately lined or covered landfills can serve as sources (Rayne, 2008; Teuten et al., 2009). Sewage
treatment facilities generally do not have the equipment needed to adequately screen for wastewater
discharges of microplastic debris such as polyethylene beads in exfoliating scrubs, abrasive plastic
fragments in cleaning agents, or acrylic shed from clothing during washing (Thompson et al., 2004; Zubris
and Richards, 2005; Fendall and Sewell, 2009). Therefore, wastewater effluent discharges serve as sources
of plastics to surface waters. A recent study by Browne et al. (2011) found that a single piece of clothing
can produce more than 1,900 plastic fibers during one wash, which may not be completely removed
during wastewater treatment and screening. Another example of a land-based source is air blasting
technologies, where microplastic scrubbers such as acrylic, polyester, or melamine are blasted at boat
hulls, engines, and machinery to remove rust or paint, and can contribute plastics to the environment
through wastewater effluent discharges (Derraik, 2002; Cole et al., 2011). Air blasting of boat hulls is
performed at marinas that can be in close proximity to water and the microplastic scrubbers may enter
the water through stormwater runoff as well. In addition, low-density polyethylene film fragments on
farmlands from hay sleeves and silage bags, bunker silo covers, silage wrap, and films used on
greenhouses or as weed barriers can be windblown or degrade into the soil (Xu et al., 2006). Lastly, plastic
pellets, which are the 1-5 mm diameter raw material melted to create plastic products, can be discharged
or accidentally spilled into the environment throughout their lifecycle, including during their production,
transport, and waste disposal (Pruter, 1987; U.S. EPA, 1993; Thompson et al., 2004; Ryan et al., 2009).
Page 25 of 50
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Land Based Sources
Sea Based Sources
Cosmetics
and
personal
hygiene
products
Clothing
(e.g.,
shedding
during
washing)
Air
blasting
Plastic films
used in
agriculture
Littering
Commercial
and
fishing
Pellets and
debris from
military, and
recreational
boating
Oil and
gas
platforms
Li
TSs.
¦+--X
X
/ \
/ \
/ y
\
\
/
/
\
/
¦//
Oomestic and
industrial
waste water
discharges
W
Storm water
runoff
-4L
Festivals
coastal
Aquaculture
tourism
Modes of Transport to the
Aquatic Environment
Wind-blown
Direct disposal
{intentional and
accidental)
Landfill
weather
leachate
events
t-, ir
Figure 5. Key sources of plastics and their modes of transport to the aquatic environment. (U.S. EPA, 1993; Rayne, 2008; Gregory, 2009; Cole
et al., 2011; Lambert et al., 2014).
Page 26 of 50
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3.2.2 Sea-Based Sources
Sea-based sources of plastics include commercial and recreational fishing vessels, barges, recreational
boaters, military activities, oil and gas platforms, and aquaculture farms (Pruter, 1987; Cole et al., 2011;
Lambert et al., 2014) (Figure 5). Fishing gear, such as nylon netting, buoys, and plastic monofilament line
are used in fishing operations. Contributions of plastics to the sea from marine fishing vessels have
increased from an estimated 23,000 tons during the 1970s (Pruter, 1987) to an estimated 6.5 million tons
during the 1990s (Derraik, 2002). Pre-production resin polyethylene and polypropylene pellets have been
used on ship decks to facilitate the movement of large objects, and can be washed from the deck
accidentally or due to improper storage or handling (Tharpes, 1989; U.S. EPA, 1993). Plastic debris from
ship decks can be windblown or lost/discarded accidentally or during improper loading and unloading of
equipment or because of inadequate storage (Lambert et al., 2014). General litter from recreational
boaters is also a source of plastics, such as bags and bottles. Industrial plastic netting and tubing used in
ocean-based aquaculture operations can directly enter the marine environment via improper disposal
(Cole et al., 2011). While a number of studies have investigated incidences of plastic ingestion by
terrestrial and aquatic life (e.g., Lattin et al., 2004; van Franeker et al., 2011) (see also Section 4), minimal
research has been conducted on plastic ingestion as a mode of transport of plastics to and within the
aquatic environment, particularly by migrating birds and ocean-dwelling animals. However, plastic debris
can be transported when seabirds regurgitate indigestible materials (i.e., a bolus) (U.S. EPA 2014a).
3.2.3 Relative Importance of Plastic Sources
Primary sources of plastics are land-based, contributing an estimated 80% of the total plastic debris to the
environment (Andrady, 2011) with littering and improper waste disposal as the major route for plastics
to enter the environment (Galgani et al., 2000; Lambert et al., 2014). Recently, Jambeck et al. (2015)
estimated that 1.7 to 4.6% (4.8 to 12.7 million metric tons) of the total plastic waste generated in coastal
countries in 2010 entered the ocean. Fishing-related activities can also be a key source in certain areas
(Galgani et al., 2000; Andrady, 2011; Lambert et al., 2014; Unger & Harrison, 2016). The proportion of
plastics originating from ships tends to increase with distance from shore, with much of the plastic debris
eventually ending up in ocean gyres (see Section 3.4) and bottom sediments (Moore et al., 2001; Ryan et
al., 2009). An investigation of debris along the Southern California Bight found that the majority of
anthropogenic debris was from ocean vessels and fishing activities (Moore and Allen, 2000), which may
account for approximately 18% of all marine plastic debris found globally (Andrady, 2011). A recent
oceanic survey conducted by NOAA on the abundance and characteristics of plastics in plankton samples
collected in the Southeast Bering Sea and off the coast of southern California found that the majority of
plastic particles were product fragments (secondary microplastics) typically less than 2.5 mm in size (Doyle
et al., 2011). As noted by the authors, the prevalence of product fragments may help demonstrate the
persistence of plastics in the marine environment (Doyle et al., 2011). Of the particles that were greater
than 5 mm in size, the majority were fishing net and line fibers (Doyle et al., 2011). A key factor that
influences the quantity of plastics in the aquatic environment is the proximity of the source, in particular
urban areas, to shorelines as evidenced by the prevalence of plastics in samples collected in Lake Erie,
which is the most populated of the three Great Lakes (Eriksen et al., 2013), and the Mediterranean Sea
(Barnes et al., 2009; Pasquini et al., 2016) (see also discussion in Sections 3.4).
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3.3 Transport Mechanisms
Once released to the environment, plastics and other debris can be transported to freshwater and marine
environments via streams and rivers, stormwater and wastewater discharges, littering and disposal along
shorelines, wind, and weather events. The distances plastics travel in the environment vary because
plastics vary in size, weight, density, and shape. Lightweight plastic products and particles (e.g., bags, films,
clothing fibers, pellets and plastic bottles) are more easily transported greater distances by wind, weather
events, stormwater, effluent discharges, and inputs from freshwater systems, compared to more dense
and larger plastic items. A plastic litter survey conducted by the Algalita Marine Research Foundation
reported up to 81 grams per cubic meter of small plastic debris in stormwater discharges during rain
events in southern California (Ryan et al., 2009). Wastewater effluent discharges are an important source
and transport mechanism of synthetic clothing fibers. One study reported an average of one fiber of
polyester, acrylic, or polyamide per liter of effluent from two Australian wastewater treatment plants
(Browne et al., 2011; Dris et al., 2016). Airborne plastic fibers released from residential and commercial
clothes dryers have not been characterized for the amount and types of plastic particles released.
Both natural and human activities influence the lateral and vertical distribution of plastics within the
marine environment. Ocean currents carry plastic debris to accumulation areas such as enclosed basins
(Collignon et al., 2012) and ocean gyres (Barnes et al., 2009). The distribution of plastics near river
outflows can mimic the patterns of sediment deposition, distributing loads of plastic debris to estuaries
and open water areas (Galgani et al., 1996; Williams and Simmons, 1997; Baldwin et al., 2016; Sutton et
al. 2016; Wessel et al., 2016). Fouling organisms such as algae, biofilms, and invertebrates often colonize
plastic particles, causing the debris to sink to the seafloor and mix with the bottom sediment (Derraik,
2002; Barnes et al., 2009; Cole et al., 2011). In addition, organisms that colonize plastic debris may limit
degradation by photolysis because they prevent the plastic surface from being exposed to ultraviolet light
(Barnes et al., 2009).
The net transport of plastic to and within the ocean (and the Great Lakes) can be exacerbated by extreme
weather events such as floods, tsunamis, hurricanes, and tornados (Barnes et al., 2009). For example, a
9.0 magnitude earthquake off the coast of a highly urbanized region in Japan in 2011 caused a tsunami
that transported an estimated 5 million tons of debris, including plastics, into the marine environment
(NOAA, 2015b). Based on ocean current data, the debris is expected to be transported through the North
Pacific Current and California Current before looping back towards the Hawaiian Islands and eventually
accumulating in the North Pacific Gyre (refer to Section 3.4 and Figure 6) (Bagulayan et al., 2012). There
have been reports of fishing nets and floats, as well as other tsunami debris along the coast of Alaska,
British Columbia, Washington, Oregon, and Hawaii (Bagulayan et al., 2012; NOAA, 2013; Kubota, 2014).
While not reviewed in this paper, plastic debris can be a vehicle for invasive species that can colonize
particles and be transported great distances, and such concerns have already been raised with the
tsunami debris that has recently washed ashore in North America (Bagulayan et al., 2012; NOAA, 2013).
For further information on the colonization and transport of non-native species and pathogens via plastic
debris, refer to Barnes (2002), Gregory (2009), SCBD (2012), and Keswani et al. (2016).
3.4 General Occurrence and Accumulation in the Aquatic Environment
Plastics are ubiquitous in the aquatic environment and particles may be found along shorelines, floating
at the water surface, mixed within the water column, or settled within the sediments and submerged
aquatic vegetation (Colton et al., 1974; Thompson et al., 2004; Barnes et al., 2009; Ryan et al., 2009; Van
Cauwenberghe et al., 2013; Eriksen et al., 2014). Plastic debris also accumulates in Arctic Sea ice where 1
Page 28 of 50
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to 7 pieces of plastic/ft3 ice were found in ice cores (Obbard et al., 2014; Lusher et al., 2015). Plastic
particles have been reported in the ocean near urban areas, offshore of remote islands, in the Great Lakes,
and even in Antarctica (see reviews by Cole et al., (2011) and Lambert et al., (2014)). Sediments are an
ultimate sink of plastic debris; microplastics were recently found in surface sediment samples collected in
three of four remote deep sea areas in the Atlantic and Mediterranean Sea (Van Cauwenberghe et al.,
2013). Higher concentrations of microplastics in sediment may exist near harbors and densely populated
areas (Besseling et al., 2013), although further research is needed in this area. Reported occurrence of
plastics in the aquatic environment is a function of the geographical areas studied. As evidenced by a
review of the literature on the impacts of marine debris on aquatic life and seabirds, the majority of
research to date has been conducted on the east and west coasts of North America, Europe, and Australia,
while fewer studies have been conducted in Asia, Africa, the Arctic, and Antarctic (SCBD, 2012). However,
a recent study reports an estimate of the total number of plastic particles and their weight floating in the
world's oceans across all five sub-tropical gyres, coastal Australia, Bay of Bengal and the Mediterranean
Sea (Eriksen et al., 2014).
The occurrence and abundance of plastics at sea is dependent on a variety of factors, including: the density
and buoyancy of the plastic particle; currents and water flow; wind and weather conditions; oceanic
geographical characteristics (e.g., shallow vs. deep areas; confined areas vs. open ocean); the presence of
large rivers; proximity to urban areas as well as industrial and wastewater treatment discharges; proximity
to trade routes and shipping channels; and the prevalence of fishing activities (Moore and Allen, 2000;
Barnes et al., 2009; Eriksen et al., 2013). Plastic debris can accumulate in bay areas that have relatively
low water circulation compared to the open sea (Hess et al., 1999; Stefatos et al., 1999; Collignon et al.,
2012), although ocean gyres are an exception to this generalization as discussed further in this section.
Proximity of urbanized areas influence the abundance of plastic found in an area: for example,
approximately 90% of the plastics encountered during expeditions in the Great Lakes were found in Lake
Erie, which is the most densely populated of the three lakes sampled (Eriksen et al., 2013). Large rivers
typically carry plastic debris farther into the ocean, thereby leading to generally lower concentrations of
plastic debris along continental shelves (Moore and Allen, 2000). Deep, confined marine areas, such as
coastal canyons, have been reported to have high densities of plastics in the sediments, particularly in the
Northwest Mediterranean Sea, Celtic Sea, North Sea, Baltic Sea, Adriatic Sea, and East-Corsica (Galagani
et al., 2000). Sites in the Mediterranean tend to have a higher density of plastic debris compared to other
seas due to densely populated coastlines, limited water circulation, and shipping trade routes (Barnes et
al., 2009; Pasquini et al., 2016). Based on current knowledge, macroplastics generally tend to accumulate
at higher densities in semi-enclosed, deep, and confined marine areas (particularly near urban centers),
as well as at oceanic fronts in the Northern Hemisphere; lower densities of macroplastics are generally
encountered in remote marine areas, along the bottom of continental shelves, and in the Southern Ocean
(Barnes et al., 2009). In addition, gyres, high-pressure systems with rotating ocean currents, facilitate the
accumulation, concentration, and retention of marine debris, including plastics (Moore et al., 2001;
Moore, 2008; Barnes et al., 2009). Eriksen et al. (2014) found surface plastics in all ocean regions,
converging in gyres in both the northern hemisphere and southern hemisphere. Although coastal
population density is much lower in the southern hemisphere, the total amount of plastics was within the
same range for oceans in both northern and southern hemispheres. The authors also modeled a loss of
microplastics from the sea surface and estimated plastic pollution on the sea surface is 0.1% of the world
annual production (Eriksen et al., 2014).
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Kuroshio
Western Garbage Patch
North Pacific
Subtropical
Convergence Zone
L
California
Eastern Garbage Patch or
N. Pacific Subtropical High
North Equatorial
Figure 6. Illustration of the North Pacific Ocean central gyre (NOAA, 2014).
As a case study, Tern Island, part of the Hawaiian Islands National Wildlife Refuge, has been a subject of
ecological concern given the significant quantity of marine plastics in the area (Friedlander et al., 2009)
and the known release of hazardous substances associated with past activities on the Island (U.S. EPA,
2014a). Tern Island and the surrounding French Frigate Shoals is designated as critical habitat for the
endangered Hawaiian monk seal (Monachus schauinslandi) and the threatened Hawaiian green sea turtle
(Chelonia mydas) and a number of other endemic species. The Hawaiian Monk Seal, the nation's most
endangered marine mammal, is in precipitous decline with the species facing possible extinction. Among
the stressors potentially impacting the survival of the Monk Seals, are high levels of PCBs found in their
blood and blubber. Tern Island/French Frigate Shoals is located in the North Pacific Subtropical
Convergence Zone where a high concentration of marine debris has been observed to accumulate (shown
in Figure 6), and studies have documented plastics in all 18 seabird species found nesting on Tern (U.S.
EPA, 2014a). In September 2014, the EPA completed a Preliminary Assessment of the impacts of marine
debris, including plastics, on the threatened and endangered species on and around Tern Island (U.S. EPA,
2014a). PCBs and lead have been detected in sediment, soils, ground water, surface water, and biological
samples collected on Tern Island (U.S. EPA, 2014a). Landfilled military waste (e.g., scrap metal, cable,
batteries, and electronic equipment) is recognized as a key source of these hazardous chemicals on Tern
Island, and the Preliminary Assessment specifically identifies plastic debris as a potential exposure
pathway for PCBs and other chemicals (U.S. EPA, 2014b). EPA postulates that ingested plastic particles
may be acting as a transport mechanism to carry chemical contaminants such as PCBs into the food chain.
Trends in macro marine debris were analyzed on Tern Island from 1990-2006. During that 16-year
timeframe, over 52,000 marine debris items were deposited on the beaches of Tern Island, 71% of which
were plastics (U.S. EPA, 2014a).
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4 Toxlcologlcal Impacts of Chemicals Associated with Plastics on
Aquatic Organisms and Aquatic-Dependent Wildlife
The adverse physical impacts to organisms from plastic debris in the aquatic environments, including
ingestion, entanglement, and smothering, have been well documented (reviewed by Kiihn et al., 2015).
Ingestion of plastic is less visible than entanglement, but may lead to direct mortality or indirect mortality
due to poor nutrition or dehydration (Browne et al., 2015; Kiihn et al., 2015). However, the ingestion of
plastics also establishes an exposure pathway between chemical additives or chemicals sorbed to plastics
and the organism which has ingested this plastic. This section reviews the literature in terms of: (1) how
aquatic organisms and aquatic-dependent wildlife may be exposed to plastics and associated
contaminants (Section 4.1) and (2) the potential for chemicals associated with plastics to bioaccumulate,
biomagnify, and cause toxic effects in aquatic life (Section 4.2).
4.1 Routes of Exposure
As plastic production grows, more enters the aquatic environment as debris. Several recent reviews of
the literature on the lethal and sublethal interaction of marine debris and aquatic life and aquatic-
dependent wildlife have found that at least 693 species are impacted by debris in the aquatic environment
(e.g., entanglement or ingestion; Gall & Thompson, 2015; Kiihn et al., 2015; SCBD, 2012), with 92% of all
encounters occurring between individual organisms and debris reported to be with plastic (Gall &
Thompson, 2015). Reports of plastic ingestion have been made for 13,110 individuals from 208 species
and reports of entanglement have been made for 30,896 individuals from 243 species (Gall & Thompson,
2015). The most commonly reported plastic items causing impacts to organisms included rope and netting
(24%); fragments (20%)1; packaging (17%) including plastic bags; fishing debris (16%); and microplastics
(11%) (SCBD, 2012). It is possible that the higher incidence of impacts from rope and netting compared to
other plastic types is related to a higher number of studies reporting on entanglement as opposed to
ingestion impacts (see Ivar do Sul and Costa, (2014) and Lavers et al. (2014)) as the detection of ingestion
is typically less obvious and requires a necropsy to confirm (Gall & Thompson, 2015). However, in all
species groups, the number of species reported to have ingested marine debris (much of which is plastic)
has increased since 1997 (Table 3) (Gall & Thompson, 2015; Kiihn et al., 2015; SCBD, 2012). Importantly,
our understanding of the types of species that have been impacted by ingested debris reflects only a small
fraction of the biodiversity of aquatic life, and therefore does not reflect the true extent of plastic debris
ingestion.
The processes by which organisms may be exposed to the chemicals associated with plastics and
associated contaminants include direct ingestion (i.e., if the animal mistakes the plastic as prey or food or
inadvertently consumes plastics while feeding), indirect ingestion (i.e., consumption of prey that ingested
the plastic), and dermal exposure (GESAMP, 2015). In addition, the size of the plastic pellet or fragment
strongly affects the rate at which sorbed chemicals may be subsequently desorbed into organisms after
ingestion. The smaller the particle or fragment, the larger the relative surface area-to-volume ratio, and
the greater relative adsorption and desorption opportunity (Teuten et al., 2009). Additionally, the
1 SCBD (2012) did not specify how "fragments" were defined and how they were differentiated from microplastics.
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retention time and exposure to the organism's digestive fluids will affect the extent and rate of
desorption.
Aquatic organisms and aquatic-dependent wildlife can be selective in the types, forms, colors, and sizes
of plastics they ingest depending on their foraging technique and diet (Derraik, 2002). For example,
Carpenter et al. (1972) was one of the first studies to report plastic ingested by fish, and found only white
plastic spherules in the digestive tracts of various fish off the coast of southern New England. In a study
of loggerhead sea turtles (Caretta caretta) in the Mediterranean Sea in the late 1980s, the turtles
appeared to only ingest white plastic particles, including PVC and extruded polystyrene (Gramentz, 1988).
A 14-year survey of Western North Atlantic seabirds found that the seabirds studied were selective in the
shapes and colors of plastics they ingested as well (Moser and Lee, 1992). Rochman et al. (2015) were the
first to assess fish and shellfish from markets in Indonesia and California, USA for the presence of
anthropogenic debris for aquatic life sold for human consumption. Anthropogenic debris was observed
by visual interpretation in 28% of fish in Indonesia, and 25% of fish from California, as well as 33% of
shellfish evaluated overall. All debris in fish from Indonesia were plastic fragments, whereas
anthropogenic debris recovered from fish in California markets were primarily fibers.
Table 3. The number of species with records of ingestion as documented in the reviews by Laist (1997)
and Gall and Thompson (2015).
(Modified version of Table 1 in Gall and Thompson, 2015).
# of Species with Ingestion Records
% Increase in the # of
Species Group
Known Species
Laist (1997)
Gall & Thompson
(2015)
Species with Ingestion
Records (1997-2015)
Fish
16,754
33 (0.20%)
50 (0.30%)
52%
Seabirds
312
111 (36%)
122 (39%)
10%
Sea turtles
7
6 (86%)
6 (86%)
0%
Marine Mammals
115
26 (23%)
30 (26%)
15%
4.2 Bioaccumulation and toxicological effects of chemicals associated with
plastics
Numerous aquatic organisms ingest plastics (see Section 4.1) and because they can act as a source and
sink for chemicals incorporated into plastics as part of the manufacturing process and chemicals that
adsorb onto their surface from the aquatic environment (Rochman, 2015), there is the potential for these
chemicals to be transferred to organisms (Teuten et al., 2007, 2009). However, whether organisms can
bioaccumulate chemicals from ingested plastic and the relative importance of this exposure route
compared to other pathways has been a recent topic of discussion (Koelmans et al., 2016). Plastic is a
concern as a vector for chemicals if it increases exposure levels, and if that exposure brings the chemical's
body burden to higher levels than effect thresholds; whether plastics play a role as vectors depends on
the gradient between the plastic and the organism (Koelmans, 2015; Bakir et al., 2016; Ziccardi et al.,
2016). The following sections describe (1) empirical field and laboratory research on bioaccumulation of
chemicals associated with plastics, as well as recent modelling studies to address this issue, and (2) studies
focusing on the toxicological effects of chemicals associated with plastics.
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4.2.1 Bioaccumulation
Field Studies
Several studies provide evidence from the field that concentrations of chemicals are positively correlated
with ingested plastic densities or chemical concentrations. Fossi et al. (2014) suggested that the presence
of chemicals, specifically phthalates and organochlorines, in basking sharks (Cetorhinus maximus) and fin
whales (Balaenoptera physalus) might be evidence of microplastic ingestion. There have been a number
of field studies on birds. For example, PCBs found in the abdominal adipose tissues of dead great
shearwaters (Puffinus gravis) in Australia were correlated with plastic loads of ingested particles (Ryan et
al., 1988). Yamashita et al. (2011) investigated plastic ingestion in 12 short-tailed shearwaters (Puffinus
tenuirostris) in the North Pacific Ocean and found total PCB concentrations and higher-chlorinated
congeners in abdominal adipose tissue was not correlated with the mass of ingested plastic, but was
positively correlated with lower-chlorinated congeners. Since the natural prey of the shearwaters
concentrate higher-chlorinated congeners, but plastics have been found to retain lower-chlorinated
congeners, the authors conclude that the PCBs are likely to have been accumulated from plastic. Tanaka
et al. (2013) also studied short-tailed shearwaters from the North Pacific Gyre and reported that three in
12 had detectable PBDE congeners (BDE209 and BDE183) in their fatty tissues. These PBDE congeners
were not found in the prey of short-tailed shearwaters but were found in the plastic particles within the
bird's digestive tracts, suggesting that ingested plastic was the source of PBDEs. Lavers et al. (2014) found
high concentrations of chromium and silver in fledgling flesh-footed shearwaters (Puffinus carneipes), and
the level for both metals was positively correlated with the mass of plastic ingested by the birds.
Fish have also been found to ingest plastic and have elevated tissue concentrations of chemicals
associated with plastics. Gassel et al. (2013) investigated plastic ingestion and tissue concentrations of
persistent organic pollutants and nonionic surfactants in juvenile yellowtail (Seriola lalandi) from the
North Pacific Central Gyre. They found evidence of synthetic debris in 10% of the fish sampled and
nonylphenol was found in one-third of their samples. Nonylphenol was previously measured in plastic
from the gyre and because it is associated with anthropogenic sources, the authors concluded that long-
range transport was unlikely and exposure to the additive is most likely from plastic debris. Rochman et
al. (2014b) found a significant relationship between the density of plastic debris and the concentration of
higher-brominated PBDE congeners typical of those found in plastic in myctophid fish (lanternfish) tissues
in the same region of the South Atlantic Ocean. However, the challenge with field studies is that it is
difficult to definitively link bioaccumulation of chemicals to the ingestion of plastics versus other uptake
pathways (Koelmans et al., 2016).
Laboratory Studies
Some experimental studies provide evidence that chemicals can be transferred from plastics to aquatic
organisms. In the laboratory, invertebrates have been found to accumulate both chemicals associated
with plastic manufacturing and chemicals adsorbed to the plastic. Besseling et al. (2013) found that
concentrations of PCBs in tissue increased by a factor of 1.1 to 1.5 in lugworms (Arenicola marina) exposed
to polystyrene microplastics mixed with PCB-contaminated sediment compared to contaminated
sediment alone. However, there was a decrease in bioaccumulation with increasing plastic concentration
that the authors attributed to the physical effects of plastic ingestion. Browne et al. (2013) found that
both additive chemicals and chemicals sorbed from the surrounding environment desorb from PVC and
can accumulate in lugworms after ingestion. PBDEs were found to bioaccumulate in amphipods
(Allorchestes compressa) exposed in the presence and absence of polyethylene particles (Chua et al.,
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2014). PBDE concentrations were measured in amphipod tissues after a 72-hr exposure indicating transfer
to tissues. However, the authors observed that the presence of microplastics reduced the uptake of PBDE
compared to controls, suggesting the PBDEs were retained within the microplastic and were less
bioavailable than unadsorbed PBDEs. A study conducted with fish (Japanese medaka; Oryzias latipes) fed
polyethylene pellets at 10% of their diet reported that mean concentrations of hydrophobic organic
compounds in body tissues were 1.2 to 2.4 times greater than those in control fish that were not fed the
pellets (Rochman et al., 2013a). In a field experiment, Teuten et al. (2009) fed eight 40-day old streaked
shearwater (Calonectris leucomelas) chicks plastic resin pellets contaminated with PCBs mixed with
natural fish prey. Within one week, PCB concentrations, especially lower-chlorinated congeners which
were relatively enriched in the plastic compared with the fish, increased in preen gland oil, suggesting
that plastic ingestion may facilitate PCB accumulation. Other researchers have found that polystyrene
beads amended with fluoranthene did not directly increase the amount of flouranthene in mussels,
however the beads themselves caused cellular and tissue damage (Paul-Pont et al., 2016). While these
studies do confirm the potential for chemicals associated with plastics to bioaccumulate in aquatic
organisms once ingested, most of these studies were not environmentally realistic in that a large chemical
concentration gradient existed favoring transfer by using unexposed organisms and feeding them high
levels of plastics or plastics with high concentrations of chemicals (Koelmans et al., 2016). Indeed, some
experimental studies have shown plastic ingestion may reduce bioaccumulation by sorbing chemicals to
ingested plastic (Teuten et al., 2007; Gouin et al., 2011; Koelmans et al., 2013; Chua et al., 2014).
Modeling Studies
The relative contributions of plastic- and trophic-derived contaminants can be difficult to quantify in both
field observational studies and experimental studies (Bakir et al., 2014; Holmes et al. 2012; Ryan et al.,
1988), and therefore several researchers have applied equilibrium-partitioning modeling to provide
indirect evidence of the transfer of chemicals associated with plastics into organisms compared to other
exposure pathways. In general, experimental studies and modeling studies both predict an increase in
bioaccumulation of up to a factor of two to three if plastic is the only source of uptake. However,
considering more environmentally relevant situations, such as accounting for all potential routes of
exposure, the role of plastics may be relatively small compared with other pathways (Koelmans, 2015).
Food-web models developed by Gouin et al. (2011) predicted reductions in bioaccumulation for PBT
chemicals with log kow values between 5.5 and 6.5 in piscivorous fish with polyethylene as 10% of their
diet due to the high adsorption affinity of the chemicals for plastic. Koelmans et al. (2013) developed a
bioaccumulation model for hydrophobic organics which was evaluated using recent laboratory research
on the lugworm and accumulation of PCBs from polystyrene microplastic (Besseling et al., 2013) using
several environmentally realistic scenarios. The model results indicated a negligible contribution of plastic
to the accumulation of PCBs compared to other exposure pathways. In a follow up study using the same
model, Koelmans et al. (2014a) investigated leaching of the plastic additives nonylphenol and BPA after
plastic ingestion to lugworms and cod (Gadus morhua). The authors argue that additives may be more
relevant than adsorbed chemicals because the plastic itself would be the source, particularly for larger
organisms with longer gut retention times. However, the model results suggest that bioaccumulation of
nonylphenol and BPA through plastic ingestion by lugworms and cod would be negligible. Recently, Herzke
et al. (2016) investigated the bioaccumulation of hydrophobic organic chemicals (PCBs, PBDEs, and DDTs)
in northern fulmars (Fulmarus glacialis) off the coast of Norway. Using fugacity calculations and
bioaccumulation modeling, the authors conclude that plastic is more likely to act as a passive sampler
than as a vector of these chemicals and ingested plastic reflects the chemical profiles of ingested natural
prey. The modeling studies, using invertebrates, fish, and seabirds, demonstrate that while plastics have
a high capacity to sorb chemicals, the desorption of the chemical once ingested may not be a major factor
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because chemicals bioaccumulated from natural prey may overwhelm the chemical dose from the plastic
(Koelmans et al., 2016; Ziccardi et al., 2016).
4.2.2 Toxlcologlcal Effects
Currently, the research investigating harmful toxicological effects of chemicals associated with plastics are
limited to several experimental studies and may not adequately reflect environmental concentrations of
either the chemicals associated with the plastics or realistic exposure scenarios via various routes (Gall &
Thompson, 2015; Koelmans et al., 2014). However, there are some studies that indicate that additional
research is needed to clarify the scope of the potential issue. For example, Lavers et al. (2014) found that
body condition is negatively influenced by the amount of ingested plastic in flesh-footed shearwaters, and
that the shearwater contaminant load (chromium and silver) was positively related to the amount of
ingested plastic.
There have been several laboratory toxicology tests using invertebrates. The toxicity of chemicals
desorbed from 32 plastic products to waterfleas (Daphnia) was performed via 24-hr batched and 3-day
diffusion tests (Lithner et al., 2009). The study revealed that Daphnia were most affected by the silver
within a compact recordable disc, followed by leachate from plastics made of plasticized PVC (e.g.,
artificial leather, a bath tub toy, and a table cloth), and polyurethane (e.g., floor coating, a child's handbag,
and artificial leather). Daphnia 48 hour EC5o values ranged from 5 - 80 g plastic material/L for nine of the
tested plastic products; leachate from the remaining plastic products did not demonstrate toxicity to
Daphnia. Browne et al. (2013) found that nonylphenol, phenanthrene, and PBDE-47 sorbed to PVC can
transfer to the tissues of uncontaminated lugworms following ingestion. Their research found non-
significant trends to suggest that the lugworms may have accumulated enough pollutants by ingesting the
microplastic to show potential patterns of reduced feeding (PBDE), compromised immunity
(nonylphenol), and reduced antioxidant capacity (PVC)(Browne et al., 2013). Another laboratory study
involving the lugworm investigated the effect of microplastic uptake on weight loss and the potential for
bioconcentration of PCBs via bioassays. According to Besseling et al. (2013), higher concentrations of
polystyrene microplastics within sediment were positively correlated with weight loss in A. marina. The
authors cite other research indicating that microplastics have been found in marine sediments at
concentrations as high as 81 mg/kg, which is approximately three orders of magnitude below the author's
observed effects concentration (Besseling et al., 2013).
Avio et al. (2015) conducted a 7-day study where polyethylene and polystyrene microplastics with and
without adsorbed pyrene were fed to mussels (Mytilus gallaprovincialis) as their sole diet. Pyrene
accumulation was found in the gills and digestive glands of the mussels at concentrations greater than
those measured on the contaminated microplastics. Biochemical and cellular biomarkers were assessed
and all microplastic treatment groups were found to alter immunological responses, lysosomal membrane
stability, peroxisomal proliferation, antioxidant response, neurotoxic effects, and genotoxicity. The study
found that most effects were not influenced by the type of polymer or contamination except for genotoxic
responses where an increased frequency of micronuclei was found after exposure to pyrene-
contaminated polystyrene. The authors suggest that, in the short-term, energy resources in the mussels
may have been directed towards the physical impacts rather than the chemical impacts of the
microplastics, but recognized that pyrene-contaminated plastics could be a potential risk for the condition
of the mussels with long-term, chronic exposure. Another mussel study (Paul-Pont et al., 2016) showed
that microplastics amended with fluoranthene did not change fluoranthene bioaccumulation in marine
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mussels, but that microplastic exposure alone increased hemocyte mortality, changed oxidative and
energetic processes in mussels, and combined microplastic and fluoranthene exposure led to highest
tissue alterations and anti-oxidant marker levels.
Exposure to PCBs and PBDEs has been linked to toxicological impacts in fish, and the transfer of these
toxicants to fish via plastic ingestion has been documented (Rochman et al., 2013a). In a unique study,
Rochman et al. (2013a) deployed pieces of low-density polyethylene in San Diego Bay, California. After
three months, the levels of total PBDEs, PAHs, and PCBs on the polyethylene were measured at 1.4, 4, and
15 times greater than measured amounts on virgin low-density polyethylene, demonstrating that these
contaminants sorbed to the plastics from the sea water. The marine-exposed and virgin plastics were
ground and fed to uncontaminated Japanese medaka as 10% of their diet via the water column for two
months to simulate likely exposure in the wild. The study found that PBDEs and some PCBs
bioaccumulated in the fish and induced liver toxicity (i.e., glycogen depletion, fatty vacuolation, and single
cell necrosis), but the authors concluded that the observed toxicity resulted from both the sorbed
contaminants and the plastic material itself (Rochman et al., 2013a). In a follow-up study using the same
fish, Rochman et al. (2014c) investigated whether the chemicals associated with plastics produced
endocrine-disrupting effects in the fish. They found altered gene expression in male fish exposed to the
marine-deployed plastic (down-regulation of choriogenin) and in female fish exposed to both the marine-
deployed plastic (down-regulation of vitellogenin, choriogenin, and the estrogen receptor) and virgin
plastic (down-regulation of choriogenin). The changes in gene expression observed in females fed the
virgin plastic suggests that the chemicals within the plastic may induce endocrine-disrupting effects.
While biomarkers of exposure and physiological signs of stress have been observed in test organisms after
ingestion of chemicals associated with plastics, few studies (with the exception of Browne et al., 2013)
have found effects on endpoints which are expected to directly affect populations such as growth,
survival, and reproduction (Ziccardi et al., 2016). In addition, in two of these studies (Browne et al., 2013
and Rochman et al., 2013c), plastics alone induced adverse effects, but the effects were greater when
exposure to both the plastic and the chemical sorbed to the plastic was assessed. In light of the modeling
studies discussed above, the physical impact associated with exposure to plastic may be as concerning as
the potential chemical impacts (Koelmans et al. 2014).
5 Conclusions and Research Needs
Plastic particles are ubiquitous in the aquatic environment and are routinely found along beaches, in
sediment, within the water column, and at the water surface (Thompson et al., 2009). Plastic debris can
be found in freshwater and marine environments ranging from coastlines near densely populated areas
to the remote open ocean and along remote island shorelines (Cole et al., 2011; Lambert et al., 2014;
Eerkes-Medrano et al., 2015). The very characteristics that make plastics useful (e.g., durability and
longevity) allow plastics to persist in the aquatic environment (Engler, 2012). Most plastics (by particle
number) encountered in the aquatic environment are microplastics, which can be roughly the size of a
grain of sand or a planktonic organism (Browne et al., 2008; Eriksen et al., 2014; Ivar do Sul and Costa,
2014). Methods to extract, isolate and identify these microplastics exist, but standardization to enumerate
and identify these very small particles need to be standardized. Numerous research studies demonstrate
that plastics are ingested (either directly or via prey) by aquatic invertebrates, fish, seabirds, sea turtles,
and marine mammals (Laist, 1997; SCBD, 2012; Goldstein and Goodwin, 2013; Setala et al., 2014; Kiihn et
al., 2015).
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Plastics in aquatic systems contain chemicals originating from the plastic material, chemicals added during
the manufacturing process, as well as organic chemicals, metals, and other contaminants sorbed from the
water column (Mato et al., 2001; Hirai et al., 2011; Lithner et al., 2011; Rochman et al., 2013). Given that
many of these chemicals have been found to have harmful effects once in the aquatic environment, the
potential toxicological impacts of these chemicals associated with plastic once ingested by aquatic
organisms and aquatic-dependent wildlife is an area of concern (Mato et al., 2001; Oehlmann et al., 2009).
However, aquatic organisms and seabirds face a multitude of environmental stressors and attributing
toxicological impacts directly to the ingestion of plastics and associated contaminants is challenging
because organisms are exposed to metals, and organic chemicals including PBT contaminants from
wastewater discharges, atmospheric deposition, and other sources in addition to plastics. The extent to
which plastics are a relative source of metals and other chemicals to aquatic organisms and aquatic-
dependent wildlife is a recent area of study.
There is evidence that aquatic organisms and aquatic-dependent wildlife accumulate chemicals from
ingested plastics. Field studies have observed correlations of plastic densities or chemical concentrations
in plastic with chemical concentrations in organisms and laboratory experiments document transfer of
chemicals from plastic to organisms when there is a concentration gradient favoring transfer (i.e., high
concentrations of chemicals on plastics and/or unexposed experimental organisms; see Section 4.2.1).
Limited modeling approaches have been used to attempt to mimic environmentally realistic scenarios,
and the models generally show a small to negligible contribution of plastic to the bioaccumulation of
associated chemicals to aquatic organisms and seabirds, relative to other sources (reviewed in Koelmans
et al., 2016). Some recent reviews suggest that, the bioaccumulation of chemicals associated with plastics
is most likely overwhelmed by uptake through other pathways; however, this does not imply that plastics
do not have negative effects on aquatic organisms (GESAMP, 2015; Koelmans et al., 2016). A limited
number of toxicological studies have been performed, mainly in the laboratory, investigating the effects
of chemicals associated with plastics. Similar to the laboratory bioaccumulation studies, many of the
toxicological studies were conducted under environmentally unrealistic situations that favored
accumulation of the chemicals from the plastic. While negative sublethal effects were observed in
treatments with chemically contaminated plastics and effects were often greater than treatments with
the plastic alone, adverse effects were also demonstrated in organisms exposed to the plastic alone (see
Section 4.2.2).
Based on the review of the literature performed to support the development of this document, a number
of research needs have been identified:
1. Studies have demonstrated that plastic debris can contain or sorb chemicals of concern, but
chemical exchange kinetics under conditions of weathering, degradation, and biofilm formation
are poorly understood (Koelmans et al., 2015). Further research to gain a better understanding of
the fate of chemicals both sorbed to and in plastics under differing environment conditions and
within an organism after ingestion is needed.
2. Laboratory experiments and modeling approaches confirm that chemicals can transfer from
plastic to organisms, but because organisms in the environment can accumulate the same classes
of chemicals from other sources, further research on the relative role plastics play in chemical
contaminant transfer to the tissues of organisms compared to other exposure pathways (aqueous
dermal exposure and ingestion from natural prey) is needed.
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3. Distinguishing the chemical from physical effects of plastics in the field is a challenge. Laboratory
experiments have shown toxicological effects from both the chemicals associated with plastics
and the plastic itself. Further research is needed to understand the relative impacts of physical
and chemical effects of ingested plastic particles.
4. There is relatively little known about nanoplastics compared to other plastic size classes, and they
are difficult to study due to the lack of detection methods (Koelmans et al., 2015). However,
research is needed because the relatively high surface area of nanoplastics may result in higher
concentrations per unit weight than microplastics. Nanoplastics may also have additional impacts
and potentially long retention times if these particles are able to cross tissue and cellular
membranes, potentially increasing their risk (Koelmans et al., 2015).
In addition to highlighting these research needs, EPA's Trash Free Waters program is a catalyst for the
development of innovative strategies to keep trash from entering U.S. rivers, lakes, and coastal waters,
thereby reducing the country's contribution to the ever-increasing volume of trash (particularly plastic
trash) in the world's oceans. Trash Free Waters program activities support trash prevention efforts by many
public and private stakeholders. Given the land-based origins of the trash problem, the program has placed
a strong emphasis on helping states, municipalities, and businesses work together to define more effective
ways to reduce litter, prevent trash entry into water, and minimize packaging waste. The Trash Free Waters
program uses an inclusive approach of multi-stakeholder consultation, strategic planning, innovative pilot
initiatives, and public/private collaboration, providing a forum for all constituencies to contribute solutions
to this non-point source challenge.
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