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Microplastics in the aquatic environment : Evidence for or against adverse impacts and major knowledge gaps

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Abstract and Figures

There is increasing scientific and public concern over the presence of microplastics (MPs) in the natural environment. Here, we present the results of a systematic review of the literature to assess the weight of evidence for MPs causing environmental harm. We conclude that MPs do occur in surface water and sediments. Fragments and fibers predominate with beads making up only a small proportion of the detected MP types. Concentrations detected are orders of magnitude lower than those reported to affect molecular level endpoints, feeding, reproduction, growth, tissue inflammation and mortality in organisms. The evidence for MPs acting as a vector for hydrophobic organic compounds (HOC) to accumulate in organisms is also weak. The available data therefore suggest that these materials are not causing harm to the environment. There is however a mismatch between the particle types, size ranges, and concentrations of MPs used in laboratory tests and those measured in the environment. Select environmental compartments have also received limited attention. There is an urgent need for studies that address this mismatch by performing better quality and more holistic monitoring studies alongside more environmentally realistic effects studies. Only then will we be able to fully characterize risks of MPs to the environment in order to support the introduction of regulatory controls that can make a real positive difference to environmental quality.
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Critical Review
Microplastics in the Aquatic Environment: Evidence for or
Against Adverse Impacts and Major Knowledge Gaps
Emily E. Burns and Alistair B.A. Boxall*
Environment Department, University of York, Heslington, United Kingdom
Abstract: There is increasing scientific and public concern over the presence of microplastics in the natural environment. We
present the results of a systematic review of the literature to assess the weight of evidence for microplastics causing
environmental harm. We conclude that microplastics do occur in surface water and sediments. Fragments and fibers
predominate, with beads making up only a small proportion of the detected microplastic types. Concentrations detected
are orders of magnitude lower than those reported to affect endpoints such as biochemistry, feeding, reproduction, growth,
tissue inflammation and mortality in organisms. The evidence for microplastics acting as a vector for hydrophobic organic
compounds to accumulate in organisms is also weak. The available data therefore suggest that these materials are not causing
harm to the environment. There is, however, a mismatch between the particle types, size ranges, and concentrations of
microplastics used in laboratory tests and those measured in the environment. Select environmental compartments have also
received limited attention. There is an urgent need for studies that address this mismatch by performing high quality and more
holistic monitoring studies alongside more environmentally realistic effects studies. Only then will we be able to fully
characterize risks of microplastics to the environment to support the introduction of regulatory controls that can make a real
positive difference to environmental quality. Environ Toxicol Chem 2018;37:2776–2796.
Keywords: Microplastics; Species sensitivity distribution; Risk; Persistent organic pollutants
Over the past decade there has been increasing scientific,
public, and regulatory interest in the occurrence and impacts of
microplastic in the environment, which have been defined as
plastic particles <5 mm in size (Hidalgo-Ruz et al. 2013) arising
from a number of sources including cosmetics, abrasion of larger
items through use (such as tire fragments), and the fragmenta-
tion of larger items of plastic (Sundt et al. 2014). In 2010, fewer
than 10 peer-reviewed articles contained the word “micro-
plastic,” although this number had risen to approximately 306 in
2017. Alongside this, there have been significant policy and
regulatory developments around the use and emissions of
microplastics, for example, in the United States, the Microbead
Free Water Act of 2015 and the Environmental Protection
(Microbeads; England) Regulations 2017, which announced a
ban on the use of microbeads in all wash-off cosmetic products.
These regulatory interests are being driven by the increasing
evidence that microplastics occur in the environment (Lusher
2015), are taken up into organisms (Eerkes-Medrano et al. 2015),
and the perception among many that the materials are adversely
affecting marine life and that they may pose a risk to human
health (Cole et al. 2011; Wright et al. 2013; Ivar do Sul et al. 2014;
Van Cauwenberghe et al. 2015; Eerkes-Medrano et al. 2015;
Galloway 2015; Koelmans et al. 2015; Lassen et al. 2015; Lusher
2015; Oberbeckmann et al. 2015; Duis and Coors 2016; Auta
et al. 2017; Anbumani and Kakkar 2018). However, various
researchers have raised concerns over the quality of some of the
studies (Song et al. 2014; Phuong et al. 2016; Connors et al.
2017; Lusher et al. 2017), and little effort has been made to put
the findings from different studies of the environmental
occurrence and effects of microplastics into a risk context
(Koelmans et al. 2017).
Therefore, we present the results from a systematic review of
the published literature to attempt to answer the following
question: do existing data on the occurrence and effects of
microplastics in the environment indicate that these materials
are causing harm? In answering this question, we explore the
evidence base for a number of assertions made by the broader
community concerning microplastics in the environment,
This article includes online-only Supplemental Data.
*Address correspondence to
Published online 17 October 2018 in Wiley Online Library
DOI: 10.1002/etc.4268
Environmental Toxicology and Chemistry—Volume 37, Number 11—pp. 2776–2796, 2018
2776 Received: 3 May 2018
Revised: 30 May 2018
Accepted: 4 September 2018
including “wastewater treatment processes are unable to
remove microplastics,” “microplastics occur in waters and
sediment,” “microplastics are taken up by organisms,” “micro-
plastics can act as vectors of persistent organic pollutants into
organisms and through food chains,” and “microplastics are
adversely affecting organisms in the environment.” We also
identify major knowledge gaps that need to be addressed to
establish the extent of microplastic environmental impacts. It is
our hope that the results of the analysis will help to focus future
research efforts on the impacts of microplastics in the
A systematic review was conducted of papers published up to
the end of 2017, which were identified by the search engines
Scopus and Web of Science. The search terms “microplastic”
and “environment” were used and 320 peer-reviewed research
articles identified. Further targeted searching was conducted
when cited literature yielded relevant peer-reviewed articles and
applicable reports published by government agencies that were
missed by the search engine.
To allow comparison of data from different sources,
ecotoxicity studies reporting concentrations in mass per liter
were converted to particles per liter according to the method of
Connors et al. (2017). Aquatic measured environmental con-
centrations (MECs; freshwater and marine) were converted from
particles per cubic meter to particles per liter by dividing by a
factor of 1000. The MECs that were reported in particles per
square meter were not converted to particles per liter and
subsequently not included in assessments. A species sensitivity
distribution (SSD) was created using the US Environmental
Protection Agency’s CADDIS Species Sensitivity Distribution
Generator (US Environmental Protection Agency 2014), also
used in a recent study to build SSDs for engineered nano-
particles (Garner et al. 2015). Ecotoxicity and occurrence data for
marine and freshwater species/environments were included.
Ecotoxicity endpoints included were limited to mortality,
growth, and reproduction (Connors et al. 2017), and both no-
observed-effect concentrations (NOECs) and lowest-observed-
effect concentrations (LOECs) were included. Ecotoxicity data
included were limited to 10- to 5000-mm particle size exposures
because this reflects the smallest size fraction identified in
environmental samples with commonly used spectrometric
methods (L
oder and Gerdts 2015; Song et al. 2015) and the
upper microplastic size limit. Ecotoxicity data used to build the
SSD are listed in the Supplemental Data.
In the present review, we use the definition of “plastic”
described by the Joint Group of Experts on the Scientific Aspects
of Marine Environmental Protection (2015), which defines a
plastic as a synthetic water-insoluble polymer, generally of
petrochemical origin, that can be molded on heating and
manipulated into various shapes designed to be maintained
during use (see also Lassen et al. 2015). This includes both
thermoplastics, such as polyethylene and polypropylene, and
thermosets (i.e., cannot be remolded after successive heating),
for example, polyurethane foams and epoxy resins (Joint Group
of Experts on the Scientific Aspects of Marine Environmental
Protection 2015). A microplastic is any solid plastic particle
5 mm in size (Eerkes-Medrano et al. 2015). Agreement on
the higher end of the microplastic range (5 mm) is consistent in
the literature; however, various authors have proposed differing
lower limits (Hidalgo-Ruz et al. 2013; Joint Group of Experts on
the Scientific Aspects of Marine Environmental Protection 2015;
Lassen et al. 2015). This generally coincides with particle
sampling size limitations (Barrows et al. 2017) or analytical limits
of detection (L
oder and Gerdts 2015; Shim et al. 2017). For
example, the Joint Group of Experts on the Scientific Aspects of
Marine Environmental Protection (2015) set the lower limit of the
microplastic size range to 1 nm, whereas Lassen et al. (2015)
limited the lower end of the range to 1 mm. Standardization of
the microplastic size range would be useful, as would agreement
on subclassifications of particle size. For example, as particles
become smaller, especially in the nanometer size range, they are
expected to behave differently from their larger counterparts,
which can influence environmental transport or fate (Besseling
et al. 2017) and potentially increase the likelihood of adverse
effects on exposed organisms (Jeong et al. 2016).
In the environment, microplastics are classified as either
primary or secondary, depending on their source. Primary
microplastics are used intentionally in the 5-mm size range and
include cosmetic beads that are used in scrubs and shampoos,
particles used for sandblasting and preproduction resin pellets
(Joint Group of Experts on the Scientific Aspects of Marine
Environmental Protection 2015; Duis and Coors 2016). Second-
ary microplastics are fragments of larger plastic materials
degraded through either use (e.g., release of fibers from
washing clothing or textiles), waste management, or fragmenta-
tion of larger plastic in the natural environment (e.g., plastic bags
or bottles; Lassen et al. 2015).
Little is known about the emission rates of these microplastic
sources to the environment, and a detailed analysis of the
current knowledge in this area is beyond the scope of the
present review. Briefly, the focus thus far has been on primary
microplastics (Lebreton et al. 2017). This is likely because usage/
sales volume multiplied by microplastic content enables a rough
emission estimation for down-the-drain microplastics, which are
expected to enter the environment through wastewater
treatment plants (WWTPs; Sundt et al. 2014). Less is known
about the formation rate of secondary microplastics because this
is influenced collectively by several factors such as polymer type
and environmental exposure conditions (Song et al. 2017).
Fragmentation can be aided by biotic activity, for example,
microbial degradation or animal activity (Sundt et al. 2014),
although photodegradation will also fragment plastic particles
at variable rates depending on the surrounding environment
(e.g., temperature, water depth) and mechanical weathering is
also possible (Cooper and Corcoran 2010). How these factors
operate together is poorly understood, making exposure
assessments of secondary microplastics difficult (Ter Halle
Microplastics in the environment—Environmental Toxicology and Chemistry, 2018;37:2776–2796 2777
et al. 2017). In contrast, microplastics in cosmetic products have
received more attention (Gouin et al. 2015). Sundt et al. (2014)
attempted a detailed assessment of all primary and secondary
microplastic emissions for Norway and concluded that tire dust
was expected to be the largest contributor to microplastic
concentrations in the Baltic Sea, whereas consumer products
were expected to have the smallest contribution. A similar
conclusion was drawn for emission estimations of microplastics
in Denmark: 0.9% of the total microplastic emission to the
aquatic environment was expected to be primary microplastics
(0.1% cosmetic products), while tire dust was expected to
contribute 60% of the total microplastic emission to the aquatic
environment (Lassen et al. 2015). Eunomia (2016) also came to a
similar conclusion, where land-based microplastic emissions to
the marine environment were dominated by tire dust. Eunomia
(2016) also reported the relative contribution of inland, coastal,
and at-sea activities on total plastic entering the marine
environment as 0.5, 9, and 1.75 million tonnes, respectively.
As these emission estimates develop for both primary and
secondary microplastics to marine, freshwater, and terrestrial
systems, they can be paired with models (Besseling et al. 2017;
Horton et al. 2017) that can estimate how particle size and source
(e.g., wastewater effluent) impact microplastic environmental
fate and occurrence.
Microplastic environmental occurrence
We identified 109 studies reporting MECs of microplastics in
the environment. These studies focused on sampling freshwater,
marine water, and sediment. Data for terrestrial soils are virtually
nonexistent (Lwanga et al. 2017), despite agricultural micro-
plastic sources or spreading of WWTP biosolids for agriculture,
as well as land-based waste disposal being potential sources of
microplastics in agricultural soils (Wagner et al. 2014). In the
present section, we summarize the analytical methods used and
the results obtained in terms of microplastic concentrations and
Methods of microplastic sampling and analysis. The
majority of monitoring studies (42%) employed solely visual
identification methods (i.e., naked eye or dissecting micro-
scopes), with 43% of those studies published in 2016 and 2017
(Figure 1). Visual identification only permits identification down
to 500 mm(L
oder and Gerdts 2015). Although visual confirma-
tion techniques are inexpensive in terms of time and cost,
misidentification of natural particles such as coal ash or coal fly
(Eriksen et al. 2013), quartz or calcium carbonate (Ballent et al.
2016), or steric acid and castor oil (Ziajahromi et al. 2017b) is
possible. Several authors have therefore concluded that the
visual identification error rate for identifying natural particles as
microplastics is unacceptably high, ranging from 33 to 70%
(Hidalgo-Ruz et al. 2013; Dekiff et al. 2014; Lenz et al. 2015;
Lusher et al. 2015; Ballent et al. 2016; Clunies-Ross et al. 2016;
Fischer et al. 2016; Horton et al. 2017; Imhof et al. 2017; Kanhai
et al. 2017). Studies not using appropriate analytical confirma-
tion techniques are likely overestimating environmental con-
centrations of relevant size fractions (Lusher et al. 2017). This is
especially true for fibers, where visual analysis alone cannot
differentiate between cotton or other natural fibrous materials
and those of synthetic origin (Fischer et al. 2016). It is also
evident from Figure 1 that the total microplastic particle count
ranges substantially among studies, 17 to over 100 000 pieces,
which is likely the result of sampling location, effort, and method.
Advanced analytical confirmation methods (some form of
Raman scattering or [m]-Fourier transform infrared spectroscopy
[FTIR]), which allow particles to be characterized in terms of their
chemical makeup and hence to distinguish from natural particles
and identify polymer type, were used in 58% of the studies. The
use of various Raman and FTIR spectroscopy techniques can also
lower the particle size detection limit to 1 and 10 mm,
respectively (L
oder and Gerdts 2015; Song et al. 2015; Duis
and Coors 2016); however, confidence in detection is decreased
at <131 mm (Fr
ere et al. 2017). In 64% of the studies involving
confirmation methods, confirmation was performed on <50% of
particles sampled. A further 13% used a chemical identification
technique to identify >50% of particles sampled, whereas 23%
confirmed 100% of suspected microplastics (Figure 1). Confir-
mation of >50% of suspect microplastics was not limited to
studies with low total particle counts (e.g., <500) despite the
additional cost and effort for sample analysis. Similar to the
FIGURE 1: Percentage of suspected microplastics per study subject to
polymer identification using analytical techniques such as Raman and
Fourier transform infrared spectroscopy, with 0% indicating that only
visual analysis techniques were used to identify microplastics. The total
particle count for studies in each category is also provided as the range
(i.e., the study identifying the fewest and greatest numbers of
microplastic particles) and average. In addition, 63, 60, 38, and 53%
of studies did not report a total number of particles found in the 0, <50,
>50, and 100% polymer identification categories, respectively.
2778 Environmental Toxicology and Chemistry, 2018;37:2776–2796—E.E. Burns and A.B.A. Boxall
studies using visual techniques, MECs from any study where
<50% of suspect microplastics have been confirmed, should be
treated with caution.
Problems can also be encountered in microplastic detection
when using appropriate analytical confirmation methods
because of difficulties pertaining to particle brittleness (breaking
apart in the sample preparation stage), biofouling of particles
(interfering with the signal), or the particle size being too small to
be adequately analyzed (Leslie et al. 2017; Shim et al. 2017).
Occurrence in surface water. Surface water monitoring for
microplastics has been performed on all continents (Figure 2).
The majority of studies that have monitored microplastics in the
water column have focused on oceans and seas (n¼58), with
only a handful focusing on freshwater (n¼10; Figure 2). The
studies report results in different units of items per square meter
and items per cubic meter, which are incompatible, and a
conversion between the 2 is not straightforward (Isobe et al.
2015). Reporting in items per square meter diminishes the
usefulness of occurrence data because all ecotoxicity tests are
reported in terms of mass or particles per volume (Duis and
Coors 2016). Despite this, studies still report only items per
square meter (Ruiz-Orejon et al. 2016; Sutton et al. 2016; Imhof
et al. 2017; Nel et al. 2017), highlighting the fact there is still a
need for standardization of reporting. The sampling methods
employed will also affect results (Lusher et al. 2014). A study
comparing several commonly used sampling methodologies
found that concentrations differed by orders of magnitude
FIGURE 2: Global distribution of marine and freshwater aquatic measured environmental concentrations (MECs) from the reviewed literature (see
Supplemental Data for references). Reported units were not converted, andtherefore, relevantMECs are reportedin 2 separate maps:items per cubic
meter (top) and items per square meter (bottom). Black dots represent concentrations reported using the other unit (e.g., in the items/m
map (top)
black dots are where concentrations in items/m
have been reported).
Microplastics in the environment—Environmental Toxicology and Chemistry, 2018;37:2776–2796 2779
depending on the method used (Song et al. 2014). This is
attributable to the depth at which sampling was focused or
particle size sampling limitations imposed by net mesh sizes. In
contrast, methods that collect whole water samples (such as grab
sampling) will not discriminate based on particle size (Barrows
et al. 2017). Again, standardization is needed to produce
repeatable and comparable monitoring results (Hidalgo-Ruz
et al. 2013).
There is reasonable global coverage from the 16 yr of
occurrence data we have reviewed (Figure 2). Highest concen-
trations have been reported near heavily urbanized and
industrialized coastal areas and in rivers, with the highest
MECs being reported in the canals of Amsterdam (100 000
[100 items/L]; Leslie et al. 2017) and off the South
Korean coast (16 727 items/m
[16.7 items/L]; Song et al. 2015).
This observation is supported by a recent extensive modeling
exercise which identified rivers passing through heavily industri-
alized areas in Asia as one of the largest freshwater contributors
to oceanic microplastic loads (Lebreton et al. 2017).
Occurrence in sediment. Fifty monitoring studies quantified
microplastics in marine/coastal sediments, with only 10 studies
investigating occurrence in freshwater sediments (Figure 3).
Similarly to the water column monitoring studies, these
investigations report occurrence in different units (i.e., items
per kilogram, items per square meter, and items per cubic
meter; Figure 3). Although it can be possible to convert between
units, the methodological details to achieve this are not always
reported (Van Cauwenberghe et al. 2015). The majority of
reported sediment monitoring studies were performed in
Europe, and similar to the aqueous occurrence studies, a
greater focus has been on beach and nearshore sediment
(Figure 3). Freshwater sediment samples came mainly from lakes
(66%). Highest concentrations were reported in the Taiwan Strait
(42 560 items/m
; Kunz et al. 2016).
Microplastic type and chemical characterization. Sample
morphological composition was reported for sediment and
the water column (marine and freshwater) in terms of sample
concentration or percentage in 83% of the occurrence studies
reviewed (Figure 4A and B). The overall average sample
composition in the water column was 52% fibers, followed by
29% fragments, with other particle morphologies including
beads/spherules, films, foams, and others making up only a
small proportion of the overall microplastics detected. A similar
trend was observed in sediment where fibers made up 45% of
the particles followed by fragments which made up 33%
(Figure 4B). In terms of polymer type, trends were also similar
in the water column and sediment, with the greatest proportion
of particles comprised of polyethylene, followed by polyethyl-
ene terephthalate and polyacrylamide in water (Figure 4C) and
polypropylene in sediment (Figure 4B). Distributions of percent-
age compositions for different particle types seen in the
sediment and water column monitoring data are summarized
in Figure 4E and F.
Microparticles do occur in surface waters and sediments
around the world. It is, however, difficult to define the precise
degree of exposure in different regions and environmental
matrices because of variability/challenges in sampling techni-
ques (Song et al. 2014), differences in the microplastic detection
methods used (L
oder and Gerdts 2015), ways in which micro-
plastics in samples have been categorized (Helm 2017), differ-
ences in sampling design (Underwood et al. 2017), reporting
units, and surveyed particle sizes (Phuong et al. 2016; Barrows
et al. 2017). Standardization is imperative in the future to allow
comparison of results across monitoring studies and also with
data from effects studies.
Where microplastic characterization has been done, the
majority of microplastics detected in monitoring studies are
believed to be of secondary origin (i.e., fragments of larger
plastic items that have degraded or fibers unintentionally
released from clothing), which indicates that sources of
secondary microplastics will be important to understand if
policy or mitigation measures to reduce microplastics in the
environment are to be effective. A great deal of regulatory focus
has been placed on primary microplastics, which, in terms of
occurrence, appear to be less significant based on the present
results. Therefore, reducing or banning (e.g., cosmetic microbe-
mental microplastic loads, a conclusion also drawn by Gouin
et al. (2015). Tracing the source of secondary microplastics is
more complex than that of primary microplastics, which may
be why they have evaded focus so far. Therefore, reporting
sample composition is important to help identify which
particles are of highest priority for ecotoxicity testing and
evaluation of their sources and pathways. The majority of
data plotted in Figure 4 pertains to the marine environment
(water column and sediment); however, the environmental
distribution of microplastics (polymer type and morphology)
will vary based on microplastic and environmental character-
istics. Therefore, as more data become available for other
compartments, such as freshwater and the sea surface layer,
these data should be presented separately to better
characterize microplastic distribution and exposure in various
environmental compartments.
Are WWTPs significant sources of microplastics?
It is believed that WWTPs are a significant contributor of
microplastics to the environment, and it has been suggested that
they remove little or none of the microplastics that are emitted to
the wastewater system because of their small size (Browne et al.
2011; McCormick et al. 2014). Detecting microplastics and
estimating WWTP removal presents many challenges; for
example, biofouling is highly likely, and many cellulosic fibers
(e.g., toilet paper) are present, resulting in the possibility of a
high percentage of misidentifications (e.g., Ziajahromi et al.
2017b). Ideally, therefore, when performing monitoring of
microplastics in WWTPs, all suspected microplastics should be
subject to analytical confirmation (Tagg et al. 2015; Dyachenko
et al. 2017). Furthermore, robust sampling approaches are
needed to capture daily variations in flow and WWTP residence
times because significant differences have been found in
samples taken throughout the day (Leslie et al. 2017).
2780 Environmental Toxicology and Chemistry, 2018;37:2776–2796—E.E. Burns and A.B.A. Boxall
A number of studies have been performed that have
quantified the removal of microplastics in different wastewater
treatment processes (Table 1). Primary treatment alone can
remove an average of 65% of the total microplastic influent load,
whereas secondary and tertiary treatment options can remove
an average of 94% of the total influent load (Table 1). A study of
Danish WWTPs predicted environmental emission rates of 0.3%
of the incoming microplastic mass (Volertsen and Hansen 2017).
The majority of microplastics that have been detected in WWTP
effluent are plastic fibers and fragments, with only a small
proportion comprising microbeads, even though microbeads
are the focus of regulatory concern. The observed removal of
microplastics is explained by the fact that even though they can
move through the exclusion meshes, many are likely to float
because of their density and be subsequently removed in the
grease layer (Murphy et al. 2016) by skimmers (Carr et al. 2016) in
the primary treatment process. If the microplastic is not floating,
it is likely fouled and will either sink to the bottom of a settling
FIGURE 3: Global distribution of marine and freshwater sediment measured environmental concentrations (MECs) from the reviewed literature
(see Supplemental Data for references). Reported units were not converted, and therefore, relevant MECs are reported in 3 separate maps, items per
square meter (top), items per cubic meter (middle), and items per kilogram (bottom). Black dots represent concentrations reported using the other units
(e.g., in the items/m
map (top) black dots are where concentrations in items/m
or items/kg have been reported).
Microplastics in the environment—Environmental Toxicology and Chemistry, 2018;37:2776–2796 2781
tank or associate with flocculants and subsequently be removed
(Gouin et al. 2015; Carr et al. 2016). In either case, it is unlikely
that a large fraction of the microplastic load will remain in the
aqueous phase of the treatment process and subsequently be
released with effluents to the environment. Volertsen and
Hansen (2017) estimated that WWTP effluent contributes only
3% of the total microplastic load reaching the environment. In
addition, a recent fate modeling exercise predicted that effluent
receiving rivers will efficiently retain many microplastics prior to
reaching the ocean, including the most dominant of the
microplastic size fractions found in WWTP effluent (Besseling
et al. 2017), suggesting that freshwater sediments are the most
FIGURE 4: Measured environmental concentration sample summary characteristics. Average polymer composition per study in the water column
(A) and sediment (B) and overall shape/morphology averages in the water column (C) and sediment (D). Sample percentages reported in reviewed
studies of fibers, fragments, and beads were ranked and plotted to give 3 distributions reflecting sample shape morphology trends in the water column
(E) and sediment (F). Studies were only included which intended to quantify all microplastic shape morphologies. Both freshwater and marine studies
were included. PA ¼polyamide; PE ¼polyethylene; PET ¼polyethylene terephthalate; PP ¼polypropylene; PS ¼polystyrene; PU ¼polyurethane;
PVA ¼polyvinyl alcohol; PVC ¼polyvinyl chloride.
TABLE 1: Summary of wastewater treatment plant removals and effluent composition for specific treatment types reported in the literature
Effluent composition
Treatment type Reported removal Fiber Fragment Bead/spherule Reference
Primary 50, 78% Mainly Fibers 1, 2
Secondary 98, 99, 96% 36–48% 46–67% 0–9% 1, 3, 4
Tertiary 98, 97, 90, 99.9% 8.8% 91% Not reported 2, 4–6
Membrane bioreactor 72, 99% 61–84% 11–33% 0% 4, 7
Full table references are reported in the Supplemental Data.
References are as follows: 1 ¼Murphy et al. (2016); 2 ¼Talvitie et al. (2015); 3¼Magnusson and Nor
en (2014); 4¼Michielssen et al. (2016); 5¼Ziajahromi et al. (2017b);
6¼Carr et al. (2016); 7 ¼Leslie et al. (2017).
2782 Environmental Toxicology and Chemistry, 2018;37:2776–2796—E.E. Burns and A.B.A. Boxall
relevant compartment when considering exposure to micro-
plastics released through WWTP effluent discharge to rivers.
The available data indicate that a significant proportion of
microplastics will be removed in WWTPs and, of those emitted in
effluent, only a small proportion will be microbeads. Results thus
far (i.e., removals) indicate that a far greater fraction of
microplastics entering wastewater will be directed to sewage
sludge instead of effluent. This suggests that spreading sewage
sludge for agricultural applications may be a more pertinent
exposure pathway to explore for microplastics released to
wastewater systems (Nizzetto et al. 2016).
Are microplastics ingested by organisms?
Several field studies have documented the ingestion of
microplastics in many species from multiple trophic levels and
geographic areas (Table 2). We direct the interested reader to
Lusher (2015) for an extensive review of animal ingestion of
microplastics in the field. Microplastics have been detected in
fish, invertebrates, and avian species (Table 2). Consistent with
water and sediment microplastic occurrence data, the greatest
proportion of microplastics detected in tissues is made up of
fibers and fragments, with only a small proportion being beads.
A recent study of 400 fish from the North Sea, employing strict
quality control criteria, yielded only 2 microplastics in a single
fish (Hermsen et al. 2017). Furthermore, fish and plankton
sampled over the past 30 yr in the Baltic Sea showed no
significant increases of internal microplastic concentration over
time. Approximately 20% of the fish sampled contained
microplastics, and 93% of these microplastics were fibers
(Beer et al. 2017). Fiber abundance could be higher for 2
reasons: internal organism concentrations reflect aquatic and
sediment MEC sample composition or fibers are not egested as
efficiently as harder particles (Murray and Cowie 2011). Fish tend
to have the lowest internal concentration, which may be
attributable to reduced exposure (e.g., feeding strategy;
Wagner et al. 2014); however, field studies have demonstrated
that higher internal microplastic concentrations were correlated
with higher surrounding microplastic concentrations and not
related to feeding mode, length, or weight for both deep water
invertebrates and fish species (Courtene-Jones et al. 2017; Pazos
et al. 2017; Steer et al. 2017). This connection was possible to
establish because the authors also quantified microplastics in
the surrounding water. This is not common practice but greatly
aids in the interpretation of results. Therefore, field uptake
studies can be improved by reporting microplastic concen-
trations both internally and externally.
In the laboratory, many studies have demonstrated uptake
of microplastics into organisms. Scherer et al. (2017) found that
microplastics co-exposed with algae significantly reduced
microplastic ingestion by Daphnia magna, which is similar to a
previous conclusion drawn by Ayukai (1987), where Acartia clausi
demonstrated preferential feeding when exposed to algae and
microplastic spheres. Weber et al. (2018) found that the
microplastic body burden of Gammarus pulex depended on
dose and age, whereas experiments conducted by Marn-Mag
and Ca~
navate (1995) linked preferential ingestion to life stage
in Penaeus japonicus. When quantifying microplastic ingestion
rates, it is important to consider test conditions because the
presence of food or the type of food could impact results, in
addition to the feeding mode and life stage of the test species
(Connors et al. 2017).
The ingestion of microplastics needs to be considered
concomitantly with egestion rates to provide meaningful
interpretation of the presence of microplastics in organisms.
Laboratory microplastic exposure studies on fish and inverte-
brate species are numerous; however, few examine the question
of whether microplastic ingestion affects egestion rates,
particularly at concentrations similar to those found in the
environment (Chua et al. 2014; Au et al. 2015; Scherer et al.
2017). There is evidence of efficient gut clearance in goldfish of
both bead-shaped microplastics and fibers (Grigorakis et al.
2017). Furthermore, Mazurais et al. (2015) observed complete
egestion of bead-shaped microplastics (10–45 mm) from Dicen-
trarchus labrax larvae after a 48-h depuration period. Significant
microplastic egestion has also been demonstrated in inverte-
brates, despite concern that egestion could be impeded by their
smaller size. Irregular particles (11–700 mm) were egested within
TABLE 2: Average and internal concentration range as well as microplastic sample composition reported in reviewed studies from the literature
Trophic group Concentration range (mean) Sample composition % mean (range) Reference
Fish 0–19 (1.4) items/fish 38% (0–100) Fiber 1–17
27% (0–94) Fragment
2%(0–24) Bead
Invertebrate 0.47–11.2 (2.8) 91% (65–100) Fiber 15–24
items/organism 13% (0–13) Fragment
0.36–11 (3.05) particles/g 5.3% Bead
Bird 14.2 items/bird 74% (55–100) Fiber 25, 26
7.7% (0–7.7) Fragment
0% Bead
Full references are reported in the Supplemental Data.
References are as follows: 1 ¼Bellas et al. (2016); 2 ¼Silva-Cavalcanti et al. (2017); 3 ¼Rochman et al. (2015) 4 ¼Nadal et al. (2016) ; 5 ¼Neves et al. (2015); 6 ¼McGoran
et al. (2017); 7 ¼Tanaka and Takada (2016); 8 ¼Wesch et al. (2016); 9 ¼G
uven et al. (2017); 10 ¼Boerger et al. (2010); 11 ¼Davison and Asch (2011 ); 12 ¼Ory et al. (2017);
13 ¼Collignon et al. (2014); 14 ¼Alomar and Deudero (2017); 15 ¼Jabeen et al. (2017); 16 ¼Rummel et al. (2016); 17 ¼Pazos et al. (2017); 18 ¼Li et al. (2016);
19 ¼Davidson and Dudas (2016); 20 ¼Remy et al. (2015); 21 ¼De Witte et al. (2014); 22 ¼Leslie et al. (2017); 23 ¼Van Cauwenberghe and Janssen (2014); 24 ¼Courtene-
Jones et al. (2017); 25 ¼Zhao et al. (2016); 26 ¼Amelineau et al. (2016).
Microplastics in the environment—Environmental Toxicology and Chemistry, 2018;37:2776–2796 2783
36 h by Allorchestes compressa (Chua et al. 2014), and complete
egestion of fibers was observed in 4 h by Gammarus fossarum
(Blarer and Burkhardt-Holm, 2016); efficient gut clearance of
beads and fragments (10–106 mm) by D. magna within 24 h,
though fragments were slower to egest than beads (Frydkjaer
et al. 2017); and complete egestion of a mixture of beads, fibers,
and fragments ingested by Idotea emarginata (H
amer et al.
2014). Au et al. (2015) reported slower egestion of fibers than
bead-shaped microplastics (which was equivalent to food
egestion) in Hyalella azteca; however, complete egestion did
occur in both exposures. Finally, field observations of Atlantic
cod identified that the vast majority of stomachs found with
microplastics were also full of organic content (Bra
˚te et al. 2016).
The authors proposed that microplastic gut clearance was
therefore similar to food. These findings suggest that micro-
plastic egestion will be significant in both fish and invertebrates
and may be influenced by species and microplastic morphology;
this information is important from a risk-assessment point of view
and should be reported with all microplastic exposure studies.
Although many studies suggest that microplastic egestion is
significant, there are also a few observations of particles
translocating from the digestive tract. For example, D. magna
exposed to 1-mm spheres exhibited translocation across the gut
epithelial barrier (Rosenkranz et al. 2009). Crabs exposed to 0.5-
mm spheres also demonstrated translocation to the hemolymph,
gills, and ovary (Farrell and Nelson 2013). Tissue translocation of
microplastics from the gut to the circulatory system has also
been demonstrated in mussels exposed to <10-mm particles
(Browne et al. 2008); however, repetition of this experiment in
Pacific oyster did not result in translocation (Sussarellu et al.
2016). Von Moos et al. (2012) provided evidence of microplastic
(<80 mm) uptake into the digestive gland of blue mussels,
causing an inflammatory response at the cellular level. Lu et al.
(2016) exposed zebra fish to 20- and 5-mm as well as 70-nm
microplastics and found 5-mm and 70-nm particles in the gills,
liver, and gut, whereas 20-mm particles were found only in the
gills and gut. The mechanisms for translocation from the gut to
the circulatory system and then to the liver are not well
understood. The translocation of particles 5 to 150 mmis
thought to be attributable to persorption, a phenomenon which
occurs in vertebrate species where particles passively and
infrequently pass from the gut to the circulatory system after
ingestion (Volkheimer 1977). Interestingly, particles greater than
the 150-mm size limit (which is the persorption threshold
associated with humans) have been found in the fish liver (up
to 600 mm; Avio et al. 2015). It may be possible, albeit unlikely,
that the persorption threshold in fish is higher, allowing >150-
mm particles to infrequently pass into the circulatory system
c 2017; Jovanovi
c et al. 2018), or another currently
unknown mechanism could be occurring. Collard et al. (2017)
reported translocation of mainly 323-mm microplastics in
anchovies and suggested 2 possible translocation theories: 1)
the agglomeration of smaller pieces that were taken up, or
2) passage through the intestinal barrier. However, the authors
state that methodological limitations prevent the precise
localization of microplastics. There is also the possibility that
studies demonstrating translocation of particles >150 mm could
be subject to contamination because follow-up research to
define a possible mechanism for this translocation has yet to be
undertaken (Avio et al. 2015; Jovanovi
c et al. 2018). What is
known is that translocation can occur and seems to be size-
dependent, but is not consistently observed after every
exposure. Particles <5mm can enter the circulatory system
more easily (e.g., nanoplastics), but smaller particles can also be
removed more easily than larger particles (Jovanovi
c 2017). It
should be highlighted that methodological limitations and small
study sizes prevent the precise localization of microplastics,
making robust conclusions difficult to draw; furthermore, these
studies, although useful, do not provide advancement toward
understanding the mechanisms behind translocation. Thus, the
mechanism behind the translocation of various particle sizes
from the gut to the circulatory system and liver and the frequency
of these events are important knowledge gaps that need to be
addressed. With a better understanding of the relationship
between translocation mechanisms, frequency, and particle size,
evaluation of the risks that microplastic translocation may pose
will become possible.
Similar to WWTP samples, analytical confirmation of the
presence of microplastics presents significant challenges in the
tissues of organisms (Vandermeersch et al. 2015; Hermsen et al.
2017), and caution should be exercised when interpreting results
from studies only using visual identification methods (Rochman
et al. 2015; Bellas et al. 2016; Davidson and Dudas 2016; Zhao
et al. 2016; Silva-Cavalcanti et al. 2017). Close attention should
also be paid to sample extraction and digestion methods
because some are inefficient, potentially degrade, or color
plastics in a sample, such as methods using nitric acid (Dehaut
et al. 2016).
Trophic transfer of microplastics
The trophic transfer of microplastics has been suggested as
an important biomagnification pathway for predators owing to
their similarity to prey and small size, resulting in availability to
lower trophic organisms (Andrady 2011). This could both
impede feeding and permit microplastics to be passed to
predators, which, after prolonged periods of feeding, may result
in biomagnification (Wright et al. 2013). Trophic transfer of
microplastics has been demonstrated in the laboratory (Farrell
and Nelson 2013; Set
a et al. 2014; Tosetto et al. 2017);
however, the circumstances of these conclusions are important
to consider. Firstly, in these studies invertebrates have been
limited to a diet of only microplastics, which could influence
uptake (Scherer et al. 2017); secondly, invertebrates are then fed
to predators prior to a depuration period; and thirdly, micro-
plastic occurrence in predators is quantified prior to depuration,
despite the high microplastic egestion rates reported in the
literature for species in both trophic levels. It is important to note
that these artificial conditions are poorly representative of
environmental conditions and thus results should be interpreted
with caution. The trophic transfer of microplastics has yet to be
shown in the field, although a recent study reported that neither
fish mass nor trophic level was related to microplastic ingestion,
leading the authors to conclude that observed microplastic
2784 Environmental Toxicology and Chemistry, 2018;37:2776–2796—E.E. Burns and A.B.A. Boxall
presence is ephemeral, suggesting low biomagnification
potential because of significant gut clearance (G
uven et al.
2017). This agrees with laboratory studies demonstrating low
microplastic gut retention times in fish (Mazurais et al. 2015;
Grigorakis et al. 2017) and invertebrates (Ugolini et al. 2013;
amer et al. 2014; Blarer and Burkhardt-Holm 2016), providing
further evidence that accumulation will be minimal; however,
available data do demonstrate that microplastics can be taken
up by organisms in the environment.
Effect studies with microplastics have explored a range of
endpoints including survival, growth, reproduction, moulting,
and biochemical endpoints. In the present section we review the
types of tests that have been employed and the results obtained.
Study conditions
A variety of experimental designs have been used to evaluate
the impacts of microplastics on freshwater and marine organ-
isms. The most common test material is polystyrene, despite
polyethylene being reported as the most common polymer in
environmental samples (Figures 4A and B and 5). The majority of
studies (95%) have worked with smaller particle sizes than those
that can be confidently detected in the environment (e.g.,
<131 mm; Figure 5). The majority of studies focus on spherical
particles, with only a handful testing fibers (Au et al. 2015) or
fragments (Imhof and Laforsch 2016), despite the prevalence of
fragments and fibers in environmental samples, an issue also
identified by a recent review on the subject (Phuong et al. 2016).
The majority of test species used in the studies are from the
primary consumer group (e.g., invertebrates), which is expected
for ethical reasons (Figure 5), with the majority of studies
investigating effects of microplastic exposure on marine
organisms, suggesting a data gap for ecotoxicity pertaining to
freshwater and terrestrial species.
Distribution of ecotoxicity endpoints
In Figure 6, NOECs and LOECs in terms of particles per liter
from each of the ecotoxicity studies reviewed are presented. The
endpoints have been separated according to the particle size
ranges studied because this is thought to impact the likelihood
of ingestion and therefore the effect (Jeong et al. 2016).
Immediately it is clear that the particle sizes tested are much
smaller than those that have been documented with confidence
as occurring in the natural environment. Micro- and nano-
particles are able to be studied in laboratory-based effects
studies because they can be labeled to ease analytical detection,
for example, with a fluorescent label (Kaposi et al. 2014). In most
of the studies, spherical particles that were either precleaned or
obtained straight from the manufacturer were used, whereas
only 5 studies tested the effects of exposure to fibers (H
et al. 2014; Au et al. 2015) and weathered fragments (Rochman
et al. 2013b; Imhof and Laforsch 2016; Ogonowski et al. 2016).
The ecotoxicity endpoint distributions (Figure 6) give a broad
overview of our current understanding of the potential effects of
microplastics. They include nonstandard and standard end-
points from both acute and chronic tests, regardless of whether
or not the test followed established guidelines such as those
recommended by the Organisation for Economic Co-operation
and Development (OECD). The majority of tests have resulted in
a NOEC; however, in many cases this refers to the highest
exposure concentration tested (Browne et al. 2008; Blarer and
Burkhardt-Holm 2016; Watts et al. 2016; Chen et al. 2017). This
would indicate that the true NOEC could actually be greater.
Fragments are thought to have a higher potential to cause
internal abrasion because of jagged or sharp edges; however,
there are limited experimental data to confirm this. A single
study thus far has reported a fragment effect concentration for
50% of the studied population (EC50), 8.6 10
particles/L for
D. magna (Ogonowski et al. 2016). The tested particle size was
approximately 1 mm, which is a relevant size in terms of reported
microplastic MECs; however, the EC50 is orders of magnitude
greater than the maximal MEC (e.g., 16.7–100 particles/L).
Lethal doses for 50% of the tested population have also been
reported for fibers, 71 430 fibers/L for the amphipod H. azteca
(Au et al. 2015) and 13 000 fibers/L for the zooplankton
Ceriodaphnia dubia (Ziajahromi et al. 2017a), which again is
an order of magnitude greater than the highest reported MECs.
Endpoints presented in Figure 6 only pertain to reviewed
studies where either particles per liter was reported or a
conversion from mass per liter using particle size and density
according to the methodology of Connors et al. (2017) was
possible. In several cases, studies reported the exposure as mass
or percentage of diet and without the necessary particle
characteristics to enable a particle per liter conversion (i.e.,
FIGURE 5: Summary of the test characteristics (particle types and sizes
and test species) used in the identified effects studies for microplastics.
Pie charts are presented for exposure particle size and polymer as well
as test species trophic level. Test species are initially reported by
trophic level, followed by the percentage of those studies that used
either marine, freshwater, or terrestrial species. PA¼polyamide; PE ¼
polyethylene; PET ¼polyethy lene terepht halate; PS ¼polystyrene;
PVC ¼polyvinyl chloride.
Microplastics in the environment—Environmental Toxicology and Chemistry, 2018;37:2776–2796 2785
Cedervall et al. 2012; H
amer et al. 2014; Imhof and Laforsch
2016; Tosetto et al. 2016; Lwanga et al. 2017). Several studies
have tested multiple particle sizes but based the exposure on
mass per liter; therefore, smaller microplastic sizes had particle
per liter counts orders of magnitude greater than the larger
particle sizes tested (Jeong et al. 2017, 2016; Lu et al. 2016;
Chen et al. 2017). In these cases, it is not possible to evaluate
whether smaller particles sizes are more harmful than larger
particle sizes. Reporting in particles per liter is preferable
because it is directly comparable with environmental occurrence
data and a better option to encompass the diversity of
microplastic particle sizes.
Effects from molecular or biomarker endpoints can be
difficult to scale up to effects in the environment; however, we
report these endpoints in the interest of canvasing the breadth of
reported effects to date. Unfortunately, not all studies could be
included, for example, Rochman et al. (2013b). Important
biomarker responses related potentially to lack of nutrition
were reported; however, a conversion to particles per liter
was not possible because the authors did not report the size
distribution of microplastics used in the study. In addition,
the study, similar to others (Paul-Pont et al. 2016), lacked a
negative control. For example, part of the diet was replaced by
plastic; therefore, effects seen in treatment fish could be
attributable to reduced diet, not the addition of the plastic
(Duis and Coors 2016). A more realistic approach would be the
addition of plastic to food without replacement (Imhof and
Laforsch 2016) or including a negative control (Karami et al.
2016; Watts et al. 2016). A similar issue is observed in
invertebrate studies where effects are attributed to microplastic
intake without consideration of effects experienced from a lack
of, or an inappropriate, food source (Huntley et al. 1987; Scherer
et al. 2017).
The usefulness of the ecotoxicity testing strategies em-
ployed in many of the studies, in terms of environmental
relevance, has been questioned (Phuong et al. 2016). Study
design limitations include the lack of environmental relevance
pertaining to the size, shape, and concentration of tested
microplastics; lack of detailed test particle characterization
such as the size distribution, density, and assessment of
chemicals potentially already sorbed prior to exposure
(Connors et al. 2017); variability in reporting units (e.g., mass
per liter or particles per liter, percentage of diet); the use of
nonstandard endpoints or biomarkers (Karami 2017); and lack
of appropriate controls (e.g., negative controls; Duis and
Coors 2016). In conclusion, data from laboratory-based studies
indicate that some microplastics have the potential to
adversely affect organisms when exposed at very high
concentrations (e.g., EC 50 of 8.6 10
particles/L; Ogonowski
et al. 2016). However, there is a mismatch between the size,
FIGURE 6: Microplastic cumulative ecotoxicity endpoint distributions for tests using particlesizes of 0.01 to 0.1mm(A), 0.1 to 1 mm(B), 1 to 10 mm(C),
and >10 mm(D). Red and black symbols represent lowest-observed-effect concentrations (LOECs) and no-observed-effect concentrations (NOECs),
respectively. A cumulative distribution can be interpreted as where along the X-axis a NOEC/LOEC is likely to fall. For example in (A) the 25th percentile
of LOECs/NOECs is approximately 10
particles/L, whereas the 75th is approximately 10
particles/L. Endpoints include a range of acute, sublethal,
and standard and nonstandard endpoints identified by the present review (see Supplemental Data for references). EC50¼median effect
2786 Environmental Toxicology and Chemistry, 2018;37:2776–2796—E.E. Burns and A.B.A. Boxall
morphology, and concentration of microplastics investigated
in the effects studies and those monitored in the environment.
Furthermore, environmental microplastics exist as a mixture,
and this should be reflected in ecotoxicity studies; for example,
testing fibers, fragments, and beads simultaneously in the
appropriate proportions would be useful (H
amer et al. 2014;
Ziajahromi et al. 2017a), in addition to investigating lesser
studied particle types, such as films, fibers, and fragments
because evidence suggests that these morphologies could be
more harmful than beads (Gray and Weinstein 2017; Hodson
et al. 2017; Ziajahromi et al. 2017a). As a result, there are
significant data gaps pertaining to microplastic ecotoxicity
(Phuong et al. 2016; Connors et al. 2017; Karami 2017), and
standardized testing which can generate EC50 data would be
useful for regulatory risk assessment. Study designs should
incorporate adequate controls and follow, when appropriate,
OECD test guidelines. Most importantly, there is an urgent
need for both monitoring and effect studies to report in
concentrations that are comparable. In the field of particle
toxicology, units of particle number per volume, surface area
per volume, and mass per volume have been used. In
performing these studies, it may be appropriate to use a
number of standardized units. The key is that authors fully
characterize the test particles used in ecotoxicity studies and
report these data to enable conversions between the various
units, which allows comparison to exposure data.
Do microplastics act as vectors of persistent
organic pollutants directly and through food
It has been claimed that, because of their physicochemical
properties, microplastics adsorb significant loads of hydropho-
bic organic contaminants (HOCs) and that when these micro-
plastics are ingested, they can act as a vector for the transport of
HOCs into the organism (Cole et al. 2011; Wright et al. 2013).
This is sometimes referred to as the “Trojan horse effect.” We
therefore examined literature that has discussed the potential for
microplastics to act as vectors of HOCs to 1) identify the most
influential papers cited as evidence of this phenomenon, and
2) determine whether the influential studies do indeed provide
evidence of the Trojan horse effect.
Plastic is efficient at sorbing HOCs, mainly because of its
hydrophobicity; and this been demonstrated in both the
laboratory (Bakir et al. 2012, 2014b) and the field (Mato et al.
2001; Rios et al. 2010). The amount of and the particular HOCs
adsorbed will be dependent on the polymer type and available
surface area (Rochman et al. 2013a). The amount of time HOCs
take to reach an equilibrium between the plastic and the
surrounding environment has been shown to take months to
years (Endo et al. 2013; Rochman et al. 2013a; Koelmans et al.
2016), whereas desorption half-lives for some compounds range
from 14 d to hundreds of years (Endo et al. 2013). This, in
conjunction with recent modeling evidence (Gouin et al. 2011;
Bakir et al. 2016; Koelmans et al. 2016; Lee et al. 2017), has led
many to conclude that microplastics in the environment are
expected to act as sinks for HOCs and not sources to organisms
postingestion (Herzke et al. 2016; Kwon et al. 2017). Conversely,
it has been suggested that internal gut conditions will facilitate
HOC desorption (Teuten et al. 2007; Bakir et al. 2014a), and
many studies published in 2017 suggest that this contaminant
exposure pathway is highly relevant, indicating that the debate is
still ongoing.
It is difficult to test the Trojan horse effect, and studies that
have attempted it have almost exclusively been limited to
laboratory experiments. Modeling studies have also been
employed to determine whether the effect is possible based
on theory. An analysis of the different studies that have
explored the effects of ingested microplastics on HOC uptake
is provided in Table 3. Correlations of HOCs in wild species
with environmental microplastics (Ryan et al. 1988; Tanaka
et al. 2013) provide little proof that plastics are responsible for
observed contamination of organisms. Laboratory studies that
have employed environmentally unrealistic test gradients
using either clean exposure media (sand or water), clean
animals, or unrealistically high HOC concentrations also only
provide limited proof of the effect (Ziccardi et al. 2016). It is not
surprising that a transfer under these laboratory conditions can
be shown (Browne et al. 2008; Chua et al. 2014; Wardrop et al.
2016); however, these results need to be put into an
environmental context. For example, several authors have
observed less transfer from plastics than other more abundant
and naturally occurring particles (e.g., sediment), suggesting
that the transfer of contaminants from plastic is not significant
(Browne et al. 2008; Beckingham and Ghosh 2017). Further-
more, studies with the polycyclic aromatic hydrocarbon
phenanthrene indicate that greater sorption occurs to plankton
than to plastic, suggesting that normal food sources may be a
more important uptake pathway for certain HOCs than plastic
(Frydkjaer et al. 2017). Another important component to
consider is the desorption half-life from plastic. Several
laboratory studies have reported complete egestion of
microplastics (in unrealistically high exposures) in 24 to 48h
(Grigorakis et al. 2017). This, in addition to the low internal
concentrations of microplastics in wild animals (Table 2), would
suggest that plastic does not accumulate in the gut long
enough to facilitate desorption, even if gut surfactants did
slightly enhance the thermodynamic favorability of HOC
To demonstrate the inconclusive categorization for studies
seemingly providing evidence of the Trojan horse hypothesis,
we use a study where Oryzias latipes were exposed to
microplastics associated with a concentration of HOCs mea-
sured in the marine environment (Rochman et al. 2013b). Fish
were kept in clean water that was refreshed regularly, with
contaminated plastics sprinkled into the water with food
(Rochman et al. 2013b). This study design is not actually testing
the Trojan horse hypothesis because it is impossible to
differentiate whether microplastics were ingested and HOCs
subsequently desorbed internally or whether the unrealistic
gradient between the clean water and microplastics sorbed with
HOCs caused the HOCs to leach directly into the water and
subsequently associate with the fish (i.e., bioconcentration
instead of bioaccumulation).
Microplastics in the environment—Environmental Toxicology and Chemistry, 2018;37:2776–2796 2787
A mass balance calculation was undertaken to determine the
theoretical maximum concentration that the HOC-associated,
microplastic-exposed fish could have based on the reported
concentrations on the pellets (Figure 7). Contamination of the
control fish with HOCs is evident (Figure 7) and may be a result of
the use of cod oil in the diet (Rochman et al. 2013b). To
demonstrate the transfer of HOCs from plastic, the reported
concentrations (black dots) would need to fall somewhere along
the blue bars; this only occurs for fluoranthrene, pyrene, PCB 123
and PCB 187, none of which were reported as significantly
different between control and treatment fish. There could be
many reasons why the experimental results do not match the
TABLE 3: Evaluation of evidence for ingestion and subsequent desorption of contaminants from microplastics as a significant exposure pathway
Evidence category
Study type Demonstrated Inconclusive Not supported Reasoning Reference
Field UCorrelation between PCBs and mass of ingested
plastic, correlation causation.
Ryan et al. (1988)
Field UHigh degree of PCB and other contaminant
absorption to polyethylene in seawater.
Mato et al. (2001)
Model / laboratory UPresence of plastic will increase sediment organism
exposure, observed enhanced desorption rates in
synthetic gut surfactant; model limited: no
biofouling or transport from organics for
Teuten et al. (2007)
Laboratory UChicks fed resin pellets, total PCB load not
significant, but lower chlorinated congeners
significantly different; small sample size and large
variability among replicates.
Teuten et al. (2009)
Model U“MP as a vector for PBT substances may be relatively
small compared to other exposure pathways.”
Gouin et al. (2011)
Field UPBDE composition found in seabirds similar to
plastic in stomach, prey samples taken 7 yr later,
>1000 km away did not contain similar PBDEs.
Tanaka et al. (2013)
Laboratory UTransfer from plastic demonstrated to worms,
determined impact of plastic on PCB transfer
Besseling et al. (2013)
Laboratory UTransfer demonstrated (high plastic, contaminant
concentration) but 250% less than transferred from
sediment (lower concentration than plastic).
Browne et al. (2013)
Laboratory UExperimental design cannot differentiate between
desorption in water and subsequent uptake or
via internal gut releases (Trojan horse); unrealistic
contaminant gradient between pellets and
exposure water.
Rochman et al. (2013b)
Laboratory USignificance at 10 times environmentally relevant
concentrations; at environmentally relevant
concentrations, uptake into amphipods was less
than sediment.
Chua et al. (2014)
Model UMicroplastic could be a substantial exposure
pathway to worms; however, conditions required
unlikely in environment; pathway for fish appears
Koelmans et al. (2014)
Field/model UPOP concentration in seabirds not correlated with
plastic ingestion; modeling suggests more likely
to act as passive sampler.
Herzke et al. (2016)
Laboratory UDemonstrated uptake in worms; however, plastic
76% less than sediment; concluded transfer
dominated by natural particles.
Beckingham and Ghosh
Model UModeled existing empirical data, flux of HOCs
bioaccumulated from natural prey >flux from
Koelmans et al. (2016)
Model UPlastic is not a quantitatively important pathway for
transfer of adsorbed chemicals.
Bakir et al. (2016)
Model URole of plastic as a vector to transfer to organisms
minimal (PAHs, fugacity).
Lee et al. (2017)
Laboratory UNo elevation from sedimented microplastics to larval
fish in unrealistically high exposures.
Sleight et al. (2017)
Laboratory UIngestion of microplastics is unlikely to increase
worm exposure to zinc.
Hodson et al. (2017)
Studies conducted prior to 2016 are most commonly cited as evidence/support for the phenomenon.
HOC ¼hydrophobic organic contaminant; MP ¼microplastic; PBDE ¼polybrominated diphenyl ether; PBT ¼persistent, bioaccumulative, toxic; PCB ¼polychlorinated
biphenyl; POP ¼persistent organic pollutant.
2788 Environmental Toxicology and Chemistry, 2018;37:2776–2796—E.E. Burns and A.B.A. Boxall
mass balance calculations. What we can say is that the added
contaminant contribution from pellets in most cases is substan-
tially less than what the control fish were exposed to. When
pellet HOC concentrations were much greater than those found
in the control fish concentrations (pyrene, phenanthrene,
fluoranthrene), a corresponding concentration spike in micro-
plastic treatment fish was not observed, suggesting that the
microplastics retained these compounds. Therefore, this study
and those using similar experimental designs are inconclusive
and cannot be used in support of the Trojan horse hypothesis. A
better design, for example, would be to use marine fish and keep
them in tanks of relevant seawater, then subsequently introduce
the presorbed microplastics, as well as controls without micro-
plastics in seawater and clean HOC-free water. In conclusion, the
available evidence either does not support that microplastics
can act as a vector of HOCs into organisms or is inconclusive. We
were not able to find a study where uptake of HOCs could truly
be attributed to transport into the organisms by microplastics.
A major component missing thus far from microplastic
environmental research is putting the effects and occurrence
studies into the context of risk. In a word analysis of abstracts
from all reviewed literature, “risk” was determined to be the
188th ranked word, whereas “concentration” and “effect”
ranked 10th and 11th, respectively. Risk assessment provides a
starting point for determining the particular microplastic
shapes, sizes, or polymers that are most likely to be harmful
in the real world and to identify geographical regions at
greatest risk. This information would help to focus future
research efforts on microplastics of greatest concern and help
inform which, if any, mitigation strategies should be intro-
duced and where they should be introduced. Therefore, in the
next sections, we bring together the results of the monitoring
and the effects studies to determine whether, based on the
current evidence base, there is a likelihood for negative
impacts in the natural environment.
Comparison of MECs with effects endpoints
To put the data from the effects studies into context, we
initially compare the MEC distributions with effect concentration
distributions (Figure 8). This comparison is limited to effects
endpoints pertaining to growth, mortality, and reproduction
because these are the standard endpoints used in the regulatory
risk assessment of chemicals (Connors et al. 2017). The lowest
LOECs/NOECs (obtained for particles in the >1-mm size range)
from the effect studies were more than 2 orders of magnitude
greater than the highest MEC (Figure 8). Based on these data,
there is therefore little evidence that concentrations of micro-
plastics seen thus far in the environment have a negative effect
on organisms, particularly given that many of the monitoring
studies are thought to have overestimated concentrations
because of limitations in the identification methodologies that
we have described.
FIGURE 7: Calculated theoretical maximum lipid concentrations in
marine plastic–exposed fish (blue bars) based on a mass balance analysis
of reported initial marine pellet concentrations from Rochman et al.
(2013b). Reported control fish (dotted bar) and marine plastic–exposed
fish (black dots) lipid concentrations are also plotted. Fish were
assumed to be 300 mg and the lipid content ranged from 2.1 to 6.2%
(C. Rochman, University of Toronto, Toronto, Ontario, Canada personal
FIGURE 8: Cumulative species endpoint distribution plotted with the
measured environmental concentration distribution (marine and
freshwater). Three separate endpoint distributions are plotted which
contain both no-observed-effect concentrations and lowest-observed-
effect concentrations from acute and chronic tests from fish,
invertebrates, and algae. Only endpoints related to growth, mortality,
and reproduction are plotted. Ecotoxicity endpoints are divided into
3 distributions based on test particle size: 0.01 to 0.1 mm, 0.1 to 1 mm,
and >1mm. LOEC ¼lowest-observed-effect concentration;
MEC ¼measured environmental concentration; NOEC ¼no-observed-
effect concentration.
Microplastics in the environment—Environmental Toxicology and Chemistry, 2018;37:2776–2796 2789
Species Sensitivity Distribution
Comparison of MECs with the ecotoxicity endpoint distribu-
tions is useful for gauging the overall trends between micro-
plastic particle size–related effects and MECs; however, there
are also enough data available in the literature to take the first
steps toward creating SSDs for microplastics and performing a
probabilistic assessment of risks. An SSD is a cumulative
probability function based on ecotoxicity tests from multiple
species representing a range of taxa (Posthuma et al. 2002).
When these endpoints are combined into a distribution (log-
normal), predictions of percentage of species affected can be
made (Newman et al. 2000). Therefore, community-level risk can
be estimated by extrapolating this statistical distribution from
individual species toxicity (Garner et al. 2015). Different
environments (e.g., freshwater, marine, and terrestrial) contain
specialized species that employ a variety of feeding strategies
(e.g., filter feeders) or life history characteristics that can increase
microplastic exposure. Equally they could be particularly
sensitive to microplastic ingestion because of body size or the
inability to egest microplastics (Wright et al. 2013). The SSD
captures this interspecies variability to a stressor (e.g., micro-
plastics), which can then be used to derive key risk-assessment
components such as predicted-no-effect concentrations
(European Chemicals Bureau 2003) or a 5% hazard concentra-
tion (HC5). The HC5 is a key regulatory parameter used to derive
legally binding environmental quality standards and translates
to the concentration where 5% of species in an ecosystem would
be harmed (Wheeler et al. 2002). Also, SSDs can be used to
derive maximum acceptable concentrations from a limited set of
laboratory data (Silva et al. 2014).
An SSD was built using the Species Sensitivity Distribution
Generator (US Environmental Protection Agency 2014). There
are several assumptions and criteria required to build a
representative SSD (Posthuma et al. 2002), and the authors
recognize that there are several limitations with the distribution
presented in Figure 9. The usefulness of an SSD depends on the
data it is created from; therefore, an important caveat to
consider for the SSD presented in Figure 9 is that both NOECs
and LOECs were used so that a range of species (n¼9) could be
included, covering key taxa (e.g., fish species, isopods,
copepods echinoderms, and crustaceans; see Supplemental
Data for references). Only mortality, reproduction, and growth
endpoints (Connors et al. 2017) from the largest particle size
class (10–5000 mm) of ecotoxicity studies were considered
because this size fraction is most relevant to particle sizes
measured in the environment and consequently most represen-
tative based on current approaches. It should be noted,
however, that only a single ecotoxicity study where a particles
per liter value could be calculated used a >100-mm particle size
exposure. If no significant effect was reported or the concentra-
tion below the LOEC was reported, this was considered the
NOEC, whereas LOECs were the lowest concentration that had a
significant effect. Endpoints were included that may not have
adhered to high-quality tests that are desirable for SSDs
(Wheeler et al. 2002). Marine and freshwater data were
combined in the SSD presented to increase statistical power
because alone not enough data are yet available to build an SSD
for the freshwater or marine environment singly. Freshwater- and
marine-specific SSDs are presented in the Supplemental Data.
We present the first attempt to build an SSD for the risk
assessment of microplastics, which in itself cannot provide
regulatory guidance; however, it provides a starting point for
what the SSD will look like and should be updated as more
relevant data become available.
The confidence intervals of the 95% MEC and the 5% effect
concentration do not overlap; the HC5 is 6.4 10
3 orders of magnitude greater than the 95% MEC, 8.5 particles/
L, which, based on current data, indicates that risks are limited.
Knowledge gaps do, however, need to be addressed to improve
the quality and relevance of the SSDs and enable sound
probabilistic risk assessment of microplastics in the environment
(Koelmans et al. 2017). This includes ecotoxicity testing of
relevant particle size and shape fractions, standardized testing,
improved reporting of methods and results, and a greater focus
on freshwater and terrestrial compartments. We have provided a
starting point to be refined as research progresses, which,
despite the caveats, does likely provide a general idea of what a
refined SSD will look like. The MEC distribution could begin to
overlap with the SSD when methods to measure smaller particle
sizes in the environment emerge. This would be useful for
putting the vast majority of current ecotoxicity studies in an
environmental context and should be considered a research
priority. On the other hand, ecotoxicity data from the 10- to
5000-mm size fraction were nearest the concentrations reported
in the environment (Figure 8).
Overall, the comparison of MECs with effects endpoints does
not support the claim of some that microplastics are negatively
impacting the health of organisms in the environment. Concen-
trations of microplastics seen to cause effects on organisms are
orders of magnitude higher than concentrations of microplastics
measured in the environment. There are several limitations to
keep in mind with regard to this comparison. We know that
approximately half of the reported MECs have fiber contents
FIGURE 9: Species sensitivity distribution plotted with the 95%
confidence interval (CI; red) based on no-observed-effect
concentrations and lowest-observed-effect concentrations from
studies of particles in the size range of 10 to 5000 mm (most relevant
to environmental size distributions). The measured environmental
concentration (MEC) cumulative distribution is also plotted (marine
and freshwater MECs) with the 95% CI (green).
2790 Environmental Toxicology and Chemistry, 2018;37:2776–2796—E.E. Burns and A.B.A. Boxall
>50%, followed by fragments, neither of which are well
represented in the effects studies, which tend to focus on
beads/spheres (H
amer et al. 2014). The effects studies have also
focused on particle sizes much smaller than those typically
monitored in the environment. To answer the question of
whether microplastics negatively impact organisms in the
environment, the size range of microplastics needs to be clearly
defined, monitoring studies need to characterize the complete
size range of microplastics that occur in the environment, and
effects studies need to work with test materials (plastic types,
sizes, and shapes) that are consistent with those found in the
environment. Only then will we be able to come to any
conclusion as to whether microplastics negatively impact the
environment or not.
Microparticles do occur in the environment, but based on our
analysis, there is currently limited evidence to suggest that they
are causing significant adverse impacts or that they are
increasing the uptake of hydrophobic organic compounds into
organisms. This conclusion is in line with conclusions from
others, calling into question the claims around risks posed by
microplastics (Koelmans et al. 2017; Burton 2017). However,
based on the current evidence, it is impossible to conclude that
microplastics do or do not cause harm to the environment. This is
attributable to the fact that monitoring efforts tend to focus on
only a fraction of the microplastic size range that could occur in
the environment and that effects studies tend to work with
materials which are not the ones currently being monitored. Only
limited data are available for freshwater environments, with even
less for terrestrial systems, even though exposures in these
environments could be greater than those in the marine
environment. To determine whether microplastics cause harm
in the environment, work therefore needs to focus on the
following 3 aspects.
First, exposure of the environment to microplastics. Higher-
quality occurrence data are needed in a broader range of
compartments (i.e., including freshwater and terrestrial systems).
This monitoring needs to determine concentrations of the
complete size range of microplastics that occur in the
environment. Concentrations need to be expressed in mean-
ingful units that can be compared to effects study data. Accurate
classification and chemical characterization of particles are
essential. Monitoring of sources, such as diffuse (e.g., tire ware,
paints, coatings) and point sources (e.g., industrial emissions
and WWTPs) is needed to establish what the major sources are of
microplastics in the environment. This will likely require the
development of new sampling and analytical methodologies
with lower concentration and size detection limits which are able
to detect all microplastics and their transformation products,
such as nanoplastics, in the natural environment. The lessons
learned from other fields, such as nanoparticles, and interdisci-
plinary work involving analytical chemists and physicists could be
valuable to help tackle these analytical challenges. The use of
exposure modeling approaches, such as that used by Lambert
et al. (2013) to characterize environmental exposure to latex and
its degradation products, will also help to characterize real-world
exposures. Exposure modeling may be particularly useful in
situations where detection of a material is not possible because
of limitations in current analytical methodologies and can
provide information at greater spatial and temporal resolution
than monitoring studies and help to identify major sources of
exposure. To inform this exposure modeling, better information
is needed on the types of macro- and microplastics in use, the
amounts used, and the usage patterns, as well as information on
the fate and behavior of these materials from laboratory and
semi-field simulation studies.
Second, effects characterization. Effects studies are needed
on the types of microplastics that actually occur in the
environment and on their transformation products, such as
nanoplastics. In particular, more work is needed on the effects of
fragments and fibers of the size ranges currently being observed
in the environment and on the effects of secondary micro-
plastics. Studies need to characterize potential effects on not
only marine organisms, but also freshwater and terrestrial
species. Although studies should explore potential impacts on
nonstandard organisms that could, because of their traits, be
vulnerable to microplastic exposures, they should focus on
ecologically relevant endpoints (e.g., mortality, growth, and
reproduction) that are used in the assessment of risks of standard
chemicals. For secondary microplastics, where the environment
will likely be exposed to a complex mixture of particles of
different sizes and shapes (Lambert et al. 2013), the use of semi-
field environmental degradation studies on microplastics (e.g.,
Lambert and Wagner 2016) followed by effects testing on the
resulting materials (e.g., Lambert et al. 2013) might help to
determine whether these materials are causing harm or not.
Third, assessment of microplastic risks. The discussion around
microplastics in the environment needs to be risk-based because
occurrence does not always equate to impact and just because
an effect is seen in the laboratory does not mean that the effect
will occur in the real environment. Better design of monitoring
and effect studies so that they yield data that inform risk
assessment will mean that it will be possible to establish the
degree of risk in different regions of the world and to identify
activities and practices contributing most to the risk. This will
mean that policies can be informed by sound science and that
they will then actually have impact on the health of the
We have presented the first detailed risk assessment of
microplastics in the environment, using both a probabilistic
method (SSD) and an ecotoxicity endpoint distribution to include
nonstandard endpoints to demonstrate that current ecotoxicity is
not comparable with MECs in terms of particle size; however,
initial assessment provides littleevidence of microplastics causing
harm in the real environment. We have also demonstrated that
significant evidence for microplastics acting as a vector for HOCs
into organisms has yet to be proven and that recent laboratory
and modeling evidence suggests that the impact of this exposure
pathway is minimal. There is currently limited evidence to
suggest that adverse environmental impacts are caused by
microplastics; however, there are major knowledge gaps that
urgently need to be addressed to confirm or disprove this.
Microplastics in the environment—Environmental Toxicology and Chemistry, 2018;37:2776–2796 2791
Supplemental Data—The Supplemental Data are available on
the Wiley Online Library at DOI: 10.1002/etc.4268
Acknowledgment—The present study was funded by the
Personal Care Products Council. The authors would like to
thank S. Dyer, K. Connors, and I. Davies for their comments
which greatly improved the manuscript.
Data Accessibility—Data, associated metadata, and calculation
tools are available by contacting the corresponding author
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... 152,153 Therefore, laboratory studies should focus on assessing intake and egestion rates during extended periods of exposure. 154 information on dose-dependent effects and ingestion. 153 The surface characteristics of MPs, together with an organism's physiology, may substantially affect the uptake and residence of MP particles. ...
... 171,172 Laboratory investigations should therefore seek to quantify egestion and ingestion rates over prolonged exposure times. 154 Zhang et al. 173 178 reported that biofouling (aging) behavior supported Ag adsorption onto PE microbeads from cosmetic products and affected its subsequent leaching. Subsequently, aged microbeads with absorbed Ag significantly increased the combined toxicity to aquatic organisms, reduced growth rates and root length of duckweed Lemna minor, and completely inhibited daphnid motility. ...
... So far, only one study has reported a fragment EC 50 , namely 8.6 Â 10 7 particles l À1 for D. magna.143 To clarify the existence of MPs in organisms, the intake of MPs must be examined with egestion rates.154 When calculating MP ingestion and egestion rates in freshwater daphnids, test conditions must be taken into account because, in addition to the life stage of the test species and feeding mode, the presence of food or the type of food could influence the results.152 ...
Full-text available
The environmental impacts of plastic pollution have recently attracted universal attention, especially in the aquatic environment. However, research has mostly been focused on marine ecosystems, even though freshwater ecosystems are equally if not more polluted by plastics. In addition, the mechanism and extent to which plastic pollution affects aquatic biota and the rates of transfer to organisms through food webs eventually reaching humans are poorly understood, especially considering leaching hazardous chemicals. Several studies have demonstrated extreme toxicity in freshwater organisms such Daphnia. When such keystone species are affected by ambient pollution, entire food webs are destabilized and biodiversity is threatened. The unremitting increase in plastic contaminants in freshwater environments would cause impairments in ecosystem functions and structure, leading to various kinds of negative ecological consequences. As various studies have reported the effects on daphnids, a consolidation of this literature is critical to discuss the limitations and knowledge gaps and to evaluate the risk posed to the aquatic environment. This review was undertaken due to the evident need to evaluate this threat. The aims were to provide a meaningful overview of the literature relevant to the potential impact of plastic pollution and associated contaminants on freshwater daphnids as primary consumers. A critical evaluation of research gaps and perspectives is conducted to provide a comprehensive risk assessment of microplastic as a hazard to aquatic environments. We outlined the challenges and limitations to microplastic research in hampering better‐focused investigations that could support the development of new plastic materials and/or establishment of new regulations.
... Recently, a greater number of studies have begun to report concentrations as both mass and counts (Fig. 2). This has been repeatedly recommended to provide maximum utility when comparing concentrations to other toxicity studies or environmental concentrations, which are typically reported as particle count per volume or surface area of water [21,25,26]. ...
... Furthermore, the literature continues to be dominated by a handful of polymer types such as PS (51% of studies in the ToMEx database) and polyethylene (35% of studies in the ToMEx database). Though these metrics only represent data within the ToMEx database, previous literature reviews have also reported similar findings [25]. The importance of polymer type in influencing toxicity is uncertain, with hypothesized significant differences based on monomer toxicities [40] and statistically significant differences based on ecotoxicological effect studies [41]. ...
Full-text available
There is definitive evidence that microplastics, defined as plastic particles less than 5 mm in size, are ubiquitous in the environment and can cause harm to aquatic organisms. These findings have prompted legislators and environmental regulators to seek out strategies for managing risk. However, microplastics are also an incredibly diverse contaminant suite, comprising a complex mixture of physical and chemical characteristics (e.g., sizes, morphologies, polymer types, chemical additives, sorbed chemicals, and impurities), making it challenging to identify which particle characteristics might influence the associated hazards to aquatic life. In addition, there is a lack of consensus on how microplastic concentrations should be reported. This not only makes it difficult to compare concentrations across studies, but it also begs the question as to which concentration metric may be most informative for hazard characterization. Thus, an international panel of experts was convened to identify 1) which concentration metrics (e.g., mass or count per unit of volume or mass) are most informative for the development of health-based thresholds and risk assessment and 2) which microplastic characteristics best inform toxicological concerns. Based on existing knowledge, it is recommended that microplastic concentrations in toxicity tests are calculated from both mass and count at minimum, though ideally researchers should report additional metrics, such as volume and surface area, which may be more informative for specific toxicity mechanisms. Regarding particle characteristics, there is sufficient evidence to conclude that particle size is a critical determinant of toxicological outcomes, particularly for the mechanisms of food dilution and tissue translocation .
... Sources of MPs are vast and vary considerably with regard to environment; for example, atmospheric MPs are generally attributed to production and wear of textiles, abrasion of tires, incinerations, and exhaust emission (Akdogan & Guven, 2019;Burns & Boxall, 2018;Mishra et al., 2019). In rivers, the sources of MPs are a result of discarded waste, effluent from gray and black water and land runoff, while in the ocean, MPs are thought to be predominantly from weathering of larger plastics, runoff from rivers, and atmospheric deposition (Campanale et al., 2020;Xu et al., 2020). ...
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Microplastics (MPs) have become frequent topics of research within Pacific Islands (PIs) in recent years; however, within PI freshwater aquaculture systems, MPs have not yet been quantified. As such this study is aimed at quantifying and characterizing the MP load from across a freshwater aquaculture system within Fiji. Water, sediment, and fish samples were collected from various stages between water source and drainage channels of an aquaculture facility in Navua, Fiji. MPs were extracted using established protocols and analyzed for abundance, form type, size, and polymer composition. Results show no significant difference in MP abundance between sampling sites for, water (average: 3.2 ± 1.14 MP/L), sediment (average: 2.3 ± 0.7 MP/100 g DW), and fish (average: 2.7 ± 1.4 MP/fish). Fibers were the most frequent form type in all three elements (average: 2.9 ± 0.2 MP/L in water, 2.1 ± 0.75 MP/100 g DW, 2.8 ± 0.14 MP/fish); however, the difference across sites was significant within water samples only. In water and sediments, smaller MPs (< 1.4 mm) were the most frequent comprising > 35% in all three elements; however, the difference was not significant between sites. Polymer analysis found that polypropylene, polyurethane, and nylon were the most abundant polymers, which coupled with observed form type and size characteristics suggest a common sources of MPs across sites.
... First, the concentrations at which those effects manifest in biota are not well understood. That uncertainty arises because of shortcomings in existing toxicological studies [16], with fewer than half of the studies conducted to date having included more than two exposure concentrations and many of those exposures at extreme concentrations well beyond what is typically encountered in the natural environment [9]. Although this testing provides useful insights into potential effects and mechanisms of toxicity, testing at multiple relevant concentrations to establish a dose-response relationship is necessary to quantitatively characterize risk. ...
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Assessing microplastics risk to aquatic ecosystems has been limited by lack of holistic exposure data and poor understanding of biological response thresholds. Here we take advantage of two recent advances, a toxicological meta-analysis that produced biotic response thresholds and a method to quantitatively correct exposure data for sampling methodology biases, to assess microplastic exposure risk in San Francisco Bay, California, USA. Using compartment-specific particle size abundance data, we rescaled empirical surface water monitoring data obtained from manta trawls (> 333 μm) to a broader size (1 to 5000 μm) range, corrected for biases in fiber undercounting and spectroscopic subsampling, and assessed the introduced uncertainty using probabilistic methods. We then compared these rescaled concentrations to four risk thresholds developed to inform risk management for California for each of two effect categories/mechanisms - tissue translocation-mediated effects and food dilution - each aligned to ecologically relevant dose metrics of surface area and volume, respectively. More than three-quarters of samples exceeded the most conservative food dilution threshold, which rose to 85% when considering just the Central Bay. Within the Central Bay, 38% of the samples exceeded a higher threshold associated with management planning, which was statistically significant at the 95% confidence interval. For tissue translocation-mediated effects, no samples exceeded any threshold with statistical significance. The risk associated with food dilution is higher than that found in other systems, which likely reflects this study having been conducted for an enclosed water body. A sensitivity analysis indicated that the largest contributor to assessment variability was associated with estimation of ambient concentration exposure due to correcting for fiber undercounting. Even after compensating for biases associated with fibers and other small particles, concentrations from the trawl samples were still significantly lower than the 1-L grab samples taken at the same time, suggesting our SFB risk estimates are an underestimate. We chose to rely on the trawl data because the 1-L grab sample volume was too small to provide accurate spatial representation, but future risk characterization studies would be improved by using in-line filtration pumps that sample larger volumes while capturing a fuller range of particle size than a towed net.
The first stranded macrodebris study on a national scale in Indonesia was conducted on 18 beaches from February 2018 to December 2019. The average weight and abundance of beach debris were higher between October and February (rainy season). The highest stranded macrodebris was located in Ambon, Manado, Takalar, and Padang. Plastic (46.38 %) was the most prevalent type of debris across all macrodebris categories, with single-use plastics such as plastic sachets, plastic bags, and plastic bottles being the dominant macroplastic debris (64.64 %). Based on CCI, HII, and BGI, 18 beaches are “moderately clean,” with few hazardous items observed, and “Good.” This anthropogenic macrodebris is thought to be more localized (55 %) than transboundary macrodebris. Litter control and environmental quality of this Indonesian coastal region should be improved through a proactive and flexible approach. Finally, extensive stranded beach debris monitoring is recommended to better understand the distribution of macrodebris in the region.
The prevalence of microplastics (MPs) in global aquatic environments has received considerable attention. Currently, concerns have been raised regarding reports that the adverse effect of MPs on aquatic animals in the exposure phase may not be (completely) reversed in the depuration phase. In order to provide insights into the legacy effect (i.e., post-exposure effect) of MPs from the depuration phase, this study evaluated the kinetic characteristics and recovery potential of aquatic animals after the exposure to MPs. More specifically, a total of 68 depuration kinetic curves were highly fitted to estimate the retention time of MPs. It was shown that the retention time ranged from 1.26 to 3.01 days, corresponding to the egestion of 90% to 99% of ingested MPs. The retention time decreased nonlinearly with the increased retention rate. Furthermore, variables potentially affecting the retention time were ranked by the decision tree-based eXtreme Gradient Boosting (XGBoost) algorithm, suggesting that the particle size and tested species were of great importance for explaining the difference in retention time of MPs. Moreover, a biomarker profile was recompiled to determine the toxic changes. Results indicated that the MPs-induced toxicity significantly reduced in the depuration phase, evidenced by the recovery of energy reserves and metabolism, hepatotoxicity, immunotoxicity, hematological parameters, neurotoxicity and oxidative stress. However, the continuous detoxification and remarkable genotoxicity implied that the toxicity was not completely alleviated. In addition, the current knowledge gaps are also highlighted, with recommendations proposed for future research.
The interactions of plastics and soil organisms are complex and inconsistent observations on the effects of plastics have been made in published studies. In this study, we assessed the effects of plastic exposure on plants, fauna and microbial communities, with a meta-analysis. Using a total of 2936 observations from 140 publications, we analysed how responses in plants, soil fauna and microorganisms depended on the plastic concentration, size, type, species and exposure media. We found that overall plastics caused substantial detrimental effects to plants and fauna, but less so to microbial diversity and richness. Plastic concentration was one of the most important factors explaining variations in plant and faunal responses. Larger plastics (>1 μm) caused unfavourable changes to plant growth, germination and oxidative stress, while nanoplastics (NPs; ≤ 1 μm) only increased oxidative stress. On the contrary, there was a clear trend showing that small plastics adversely affected fauna reproduction, survival and locomotion than large plastics. Plant responses were indifferent to plastic type, with most studies conducted using polyethylene (PE) and polystyrene (PS) plastics, but soil fauna were frequently more sensitive to PS than to PE exposure. Plant species played a vital role in some parameters, with the effects of plastics being considerably greater on vegetable plants than on cereal plants.
Studies in the oceans and The Great Lakes have found several orders of magnitude less plastic in surface samples than predicted by input estimates. Some plastic likely sinks after entering the water because it is naturally more dense than freshwater. For less dense particles, it has been proposed that biofouling, or the buildup of organic materials on the plastic, can cause them to become more dense and ultimately induce sinking. In this work we compare two different functional biofouling models: one basic algal growth population model and one model that assumes photosensitive defouling. We investigate the effects within the scope of a large-scale hydrodynamic model that includes advection, vertical diffusion, and sediment deposition applied to both Lake Erie and Lake Ontario. We find that deposition rates are dependent on the fouling method and lake depth. Lastly, we use the model to develop a first pass mass estimate for the sediment deposition rate in Lake Ontario.
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Microplastics (MPs) are environmental pollutants of growing concern, and awareness of MPs pollution in marine and freshwater environments has increased in recent years. However, knowledge of MPs contamination in riverine sediments in Ireland is limited. To address this, we collected and analysed sediment samples from 16 selected sites along the River Barrow. Microplastics were extracted through a density separation method, after which their size, colour, and shape were analysed under a stereo microscope (Optica SZM-2). Attenuated total reflection Fourier transform infrared (ATR-FTIR) spectroscopy was used to identify polymer types. A total of 690 MPs were recovered from the 16 sites, with fibres as the dominant MP type. The highest concentration of MPs was 155 MP fibres kg⁻¹ wet sediment found in samples collected from Graiguenamanagh, Co. Kilkenny (GK). The majority of the recovered MPs were polyethylene (PE), polypropylene (PP), nylon, and cellulose acetate (CA) fibres. Overall, this study highlighted the presence of MPs in Irish river sediments and provided a baseline for future studies on MPs pollution. Further research is needed to better understand sources, distribution, and effects of MPs in freshwater ecosystems.
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The ubiquitous presence of microplastics in the environment has drawn the attention of ecotoxicologists on its safety and toxicity. Sources of microplastics in the environment include disintegration of larger plastic items (secondary microplastics), personal care products like liquid soap, exfoliating scrubbers, and cleaning supplies etc. Indiscriminate usage of plastics and its poor waste disposal management pose serious concern on ecosystem quality at global level. The present review focused on the ecological impact of microplastics on biota at different trophic levels, its uptake, accumulation, and excretion etc., and its plausible mechanistic toxicity with risk assessment approaches. Existing scientific evidence shows that microplastics exposure triggers a wide variety of toxic insult from feeding disruption to reproductive performance, physical ingestion, disturbances in energy metabolism, changes in liver physiology, synergistic and/ or antagonistic action of other hydrophobic organic contaminants etc. from lower to higher trophics. Thus, microplastic accumulation and its associated adverse effects make it mandatory to go in for risk assessment and legislative action. Subsequent research priorities, agenda, and key issues to be addressed are also acknowledged in the present review.
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Microscopic plastic items (microplastics) are ubiquitously present in aquatic ecosystems. With decreasing size their availability and potential to accumulate throughout food webs increase. However, little is known on the uptake of microplastics by freshwater invertebrates. To address this, we exposed species with different feeding strategies to 1, 10 and 90 µm fluorescent polystyrene spheres (3–3 000 particles mL⁻¹). Additionally, we investigated how developmental stages and a co-exposure to natural particles (e.g., food) modulate microplastic ingestion. All species ingested microplastics in a concentration-dependent manner with Daphnia magna consuming up to 6 180 particles h⁻¹, followed by Chironomus riparius (226 particles h⁻¹), Physella acuta (118 particles h⁻¹), Gammarus pulex (10 particles h⁻¹) and Lumbriculus variegatus (8 particles h⁻¹). D. magna did not ingest 90 µm microplastics whereas the other species preferred larger microplastics over 1 µm in size. In C. riparius and D. magna, size preference depended on the life stage with larger specimens ingesting more and larger microplastics. The presence of natural particles generally reduced the microplastics uptake. Our results demonstrate that freshwater invertebrates have the capacity to ingest microplastics. However, the quantity of uptake depends on their feeding type and morphology as well as on the availability of microplastics.
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The presence of microplastics in aquatic ecosystems is of increasing global concern. This study investigated ingestion, egestion and acute effects of polyethylene microplastics in Daphnia magna. Fate of regular shaped microplastic beads (10–106 µm) were compared with irregular shaped microplastic fragments (10–75 µm). Daphnia magna ingested regular and irregular microplastic with uptake between 0.7 and 50 plastic particles/animal/day when exposed to microplastic concentrations of 0.0001–10 g/L. Egestion of irregular fragments was slower than that of microplastic beads. The EC50 for irregular microplastic was 0.065 g/L whereas microplastic beads were less inhibitory. The potential of microplastic to act as vector for hydrophobic pollutants was examined using [14C]phenanthrene as tracer. Polyethylene microplastic sorbed less [14C]phenanthrene compared to natural plankton organisms (bacteria, algae, yeast). As microplastics are much less abundant in most aquatic ecosystems compared to plankton organisms this suggests a limited role as vector for hydrophobic pollutants under current environmental conditions.
In spite of the growing importance of Species Sensitivity Distribution models (SSDs) in ecological risk assessments, the conceptual basis, strengths, and weaknesses of using them have not been comprehensively reviewed. This book fills that need. Written by a panel of international experts, Species Sensitivity Distributions in Ecotoxicology reviews the current SSD methods from all angles, compiling for the first time the variety of contemporary applications of SSD-based methods. Beginning with an introduction to SSDs, the chapter authors review the issues surrounding SSDs, synthesizing the positions of advocates and critics with their own analysis of each issue. Finally, they discuss the prospects for future development, paving the way for improved future uses. In sum, this book defines the field of SSD modeling and application. It reveals a lively field, with SSD-applications extending beyond legally adopted quality criteria to other applications such as Life-Cycle Analysis. For anyone developing or revising environmental criteria or standards, this book explores the pros and cons of using the SSD approach. For anyone who needs to apply and interpret SSD-based criteria or standards, the book explains the basis for the numbers, thereby making it possible to correctly apply and defend them. For anyone performing ecological risk assessments, the book covers when and how to use SSDs including alternative assumptions, data treatments, computational methods, and available resources. Species Sensitivity Distributions in Ecotoxicology provides you with a clear picture of these standard models for estimating ecological risks from laboratory toxicity data.
Among aquatic organisms, fish are particularly susceptible to ingesting microplastic particles due to their attractive coloration, buoyancy, and resemblance to food. However, in previous experimental setups, fish were usually exposed to unrealistically high concentrations of microplastics, or the microplastics were deliberately contaminated with persistent organic chemicals; also, in many experiments, the fish were exposed only during the larval stages. The present study investigated the effects of virgin microplastics in gilt-head seabream (Sparus aurata) after 45 days' exposure at 0.1 g kg−1 bodyweight day−1 to 6 common types of microplastics. The overall growth, biochemical analyses of the blood, histopathology, and the potential of the microplastics to accumulate in gastrointestinal organs or translocate to the liver and muscles were monitored and recorded. The results revealed that ingestion of virgin microplastics does not cause imminent harm to the adult gilt-head seabream during 45 days of exposure and an additional 30 days of depuration. The retention of virgin microplastics in the gastrointestinal tract was fairly low, indicating effective elimination of microplastics from the body of the fish and no significant accumulation after successive meals. Therefore, both the short- and the long-term retention potential of microplastics in the gastrointestinal tract of fish is close to zero. However, some large particles remained trapped in the liver, and 5.3% of all the livers analyzed contained at least one microplastic particle. In conclusion, the dietary exposure of S. aurata to 6 common types of virgin microplastics did not induce stress, alter the growth rate, cause pathology, or cause the microplastics to accumulate in the gastrointestinal tract of the fish.
Over the past decade, microscopic plastic debris, known as microplastics, emerged as a contaminant of concern in marine and freshwater ecosystems. Although regularly detected in aquatic environments, the toxicity of those synthetic particles is not well understood. To address this, we investigated whether the exposure to microplastics adversely affects the amphipod Gammarus pulex, a key freshwater invertebrate. Juvenile (6-9 mm) and adult (12-17 mm) individuals were exposed to irregular, fluorescent polyethylene terephthalate fragments (PET, 10-150 μm; 0.8-4,000 particles mL-1) for 24 h. Results show that body burden after 24 h depends on the dose and age of G. pulex with juveniles ingesting more microplastics than adults. After chronic exposure over 48 d, microplastics did not significantly affect survival, development (molting), metabolism (glycogen, lipid storage) and feeding activity of G. pulex. This demonstrates that even high concentrations of PET particles did not negatively interfere with the analyzed endpoints. These results contradict previous research on marine crustaceans. Differences may result from variations in the exposure regimes (e.g., duration, particle concentrations), plastic characteristics (e.g., type, size, shape, additives) as well as the species-specific morphological, physiological and behavioral traits. As a detritivorous shredder G. pulex is adapted to feed on non-digestible materials and might, therefore, be less sensitive towards exposure to synthetic particles. Accordingly, we argue that the autecology needs to be taken into account and that research should focus on identifying traits that render species susceptible to microplastic exposure.
There is limited knowledge regarding the adverse effects of wastewater-derived microplastics, particularly fibers, on aquatic biota. In this study, we examined the acute (48 h) and chronic (8 d) effects of microplastic polyester fibers and polyethylene (PE) beads on freshwater zooplankton Ceriodaphnia dubia. We also assessed the acute response of C. dubia to a binary mixture of microplastic beads and fibers for the first time. Acute exposure to fibers and PE beads both showed a dose-dependent effect on survival. An equitoxic binary mixture of beads and fibers resulted in a toxic unit of 1.85 indicating less than additive effects. Chronic exposure to lower concentrations did not significantly affect survival of C. dubia, but a dose-dependent effect on growth and reproduction was observed. Fibers showed greater adverse effects than PE beads. While ingestion of fibers was not observed, scanning electron microscopy showed carapace and antenna deformities after exposure to fibers, with no deformities observed after exposure to PE beads. While much of the current research has focused on microplastic beads, our study shows that microplastic fibers pose a greater risk to C. dubia, with reduced reproductive output observed at concentrations within an order of magnitude of reported environmental levels.
Microplastic is considered a potential threat to marine life as it is ingested by a wide variety of species. Most studies on microplastic ingestion are short-term investigations and little is currently known about how this potential threat has developed over the last decades where global plastic production has increased exponentially. Here we present the first long-term study on microplastic in the marine environment, covering three decades from 1987 to 2015, based on a unique sample set originally collected and conserved for food web studies. We investigated the microplastic concentration in plankton samples and in digestive tracts of two economically and ecologically important planktivorous forage fish species, Atlantic herring (Clupea harengus) and European sprat (Sprattus sprattus), in the Baltic Sea, an ecosystem which is under high anthropogenic pressure and has undergone considerable changes over the past decades. Surprisingly, neither the concentration of microplastic in the plankton samples nor in the digestive tracts changed significantly over the investigated time period. Average microplastic concentration in the plankton samples was 0.21±0.15particlesm(-3). Of 814 fish examined, 20% contained plastic particles, of which 95% were characterized as microplastic (<5mm) and of these 93% were fibres. There were no significant differences in the plastic content between species, locations, or time of day the fish were caught. However, fish size and microplastic in the digestive tracts were positively correlated, and the fish contained more plastic during summer than during spring, which may be explained by increased food uptake with size and seasonal differences in feeding activity. This study highlights that even though microplastic has been present in the Baltic environment and the digestive tracts of fishes for decades, the levels have not changed in this period. This underscores the need for greater understanding of how plastic is cycled through marine ecosystems. The stability of plastic concentration and contamination over time observed here indicates that the type and level of microplastic pollution may be more closely correlated to specific human activities in a region than to global plastic production and utilization as such.