ArticlePDF AvailableLiterature Review

Abstract

Since the mass production of plastics began in the 1940s, microplastic contamination of the marine environment has been a growing problem. Here, a review of the literature has been conducted with the following objectives: (1) to summarise the properties, nomenclature and sources of microplastics; (2) to discuss the routes by which microplastics enter the marine environment; (3) to evaluate the methods by which microplastics are detected in the marine environment; (4) to assess spatial and temporal trends of microplastic abundance; and (5) to discuss the environmental impact of microplastics. Microplastics are both abundant and widespread within the marine environment, found in their highest concentrations along coastlines and within mid-ocean gyres. Ingestion of microplastics has been demonstrated in a range of marine organisms, a process which may facilitate the transfer of chemical additives or hydrophobic waterborne pollutants to biota. We conclude by highlighting key future research areas for scientists and policymakers.
Review
Microplastics as contaminants in the marine environment: A review
Matthew Cole
a,
, Pennie Lindeque
a
, Claudia Halsband
b
, Tamara S. Galloway
c
a
Plymouth Marine Laboratory, Prospect Place, The Hoe, Plymouth PL1 3DH, UK
b
Akvaplan-niva AS, FRAM – High North Research Centre for Climate and the Environment, N-9296 Tromsø, Norway
c
College of Life and Environmental Sciences: Biosciences, Hatherly Laboratories, University of Exeter, Prince of Wales Road, Exeter EX4 4PS, UK
article info
Keywords:
Microplastics
Marine litter
Plastic debris
Priority pollutant
abstract
Since the mass production of plastics began in the 1940s, microplastic contamination of the marine envi-
ronment has been a growing problem. Here, a review of the literature has been conducted with the follow-
ing objectives: (1) to summarise the properties, nomenclature and sources of microplastics; (2) to discuss
the routes by which microplastics enter the marine environment; (3) to evaluate the methods by which
microplastics are detected in the marine environment; (4) to assess spatial and temporal trends of micro-
plastic abundance; and (5) to discuss the environmental impact of microplastics. Microplastics are both
abundant and widespread within the marine environment, found in their highest concentrations along
coastlines and within mid-ocean gyres. Ingestion of microplastics has been demonstrated in a range of mar-
ine organisms, a process which may facilitate the transfer of chemical additives or hydrophobic waterborne
pollutants to biota. We conclude by highlighting key future research areas for scientists and policymakers.
Ó2011 Elsevier Ltd. All rights reserved.
Contents
1. Introduction ........................................................................................................ 2588
2. Microplastics . . . . . . . . . . . . . . . ........................................................................................ 2589
2.1. Primary microplastics . . . . . . . . . . . . . . . . ................................................ ........................... 2589
2.2. Secondary microplastics . . . . . . . . . . . . . . ................................ ........................................... 2589
3. Sources and transfer of microplastics into the marine environment . . . . . . . . . .................................................. 2590
4. Assessing microplastic abundance . . . . . . . . . . . . . . . . . ..................................................................... 2591
5. Spatial and temporal trends of microplastics in the marine environment . . . . . .................................................. 2592
5.1. Accumulation of microplastics . . . . . . . . . .......................................................... ................. 2592
5.2. Microplastics in the water column. . . . . . ................ ........................................................... 2592
5.3. Temporal changes in microplastic abundance within the marine environment . . . . . ........................................ 2593
6. Impact of microplastics on the marine environment. . . ..................................................................... 2593
6.1. Microplastic ingestion. . . . . . . . . . . . . . . . .................... ....................................................... 2593
6.2. Microplastics and plasticiser leachates . . ............................. .............................................. 2594
6.3. Microplastics and adhered pollutants . . . .......... ................................................................. 2595
7. Conclusions and recommendations for future work . . . ..................................................................... 2596
Acknowledgements . . . . . . . . . . ........................................................................................ 2596
References . ........................................................................................................ 2596
1. Introduction
Plastics are synthetic organic polymers, which are derived from
the polymerisation of monomers extracted from oil or gas (Derraik,
2002; Rios et al., 2007; Thompson et al., 2009b). Since the develop-
ment of the first modern plastic; ‘Bakelite’, in 1907, a number of
inexpensive manufacturing techniques have been optimised,
resulting in the mass production of a plethora of lightweight, dura-
ble, inert and corrosion-resistant plastics (PlasticsEurope, 2010).
These attributes have led to the extensive use of plastics in near
inexhaustible applications (Andrady, 2011). Since mass production
began in the 1940s, the amount of plastic being manufactured has
increased rapidly, with 230 million tonnes of plastic being
produced globally in 2009 (PlasticsEurope, 2010), accounting for
8% of global oil production (Thompson et al., 2009b).
0025-326X/$ - see front matter Ó2011 Elsevier Ltd. All rights reserved.
doi:10.1016/j.marpolbul.2011.09.025
Corresponding author. Tel.: +44 (0)1752 633450; fax: +44 (0)1752 633101.
E-mail address: mcol@pml.ac.uk (M. Cole).
Marine Pollution Bulletin 62 (2011) 2588–2597
Contents lists available at SciVerse ScienceDirect
Marine Pollution Bulletin
journal homepage: www.elsevier.com/locate/marpolbul
Whilst the societal benefits of plastic are far-reaching (Andrady
and Neal, 2009), this valuable commodity has been the subject of
increasing environmental concern. Primarily, the durability of plas-
tic that makes it such an attractive material to use also makes it
highly resistant to degradation, thus disposing of plastic waste is
problematic (Barnes et al., 2009; Sivan, 2011). Exacerbated by
the copious use of throw-away ‘‘user’’ plastics (e.g. packaging
material), the proportion of plastic contributing to municipal waste
constitutes 10% of waste generated worldwide (Barnes et al., 2009).
While some plastic waste is recycled, the majority ends up in land-
fill where it may take centuries for such material to breakdown and
decompose (Barnes et al., 2009; Moore, 2008). Of particular con-
cern are plastics that, through indiscriminate disposal, are entering
the marine environment (Gregory, 2009). Despite plastics being an
internationally recognised pollutant with legislation in place
aimed to curb the amount of plastic debris entering the marine
environment (Gregory, 2009; Lozano and Mouat, 2009), Thompson
(2006) estimates up to 10% of plastics produced end up in the
oceans, where they may persist and accumulate.
The impact that large plastic debris, known as ‘macroplastics’,
can have on the marine environment has long been the subject
of environmental research. The presence of macroplastics in the
marine environment presents an aesthetic issue, with economic
repercussions for the tourist industry, a hazard for numerous mar-
ine-industries (e.g. shipping, fishing, energy production, aquacul-
ture) as plastic may result in entanglement and damage of
equipment, and significant environmental concerns (Barnes et al.,
2009; Derraik, 2002; Sivan, 2011). The environmental impact of
macroplastics include: the injury and death of marine birds, mam-
mals, fish and reptiles resulting from plastic entanglement and
ingestion (Derraik, 2002; Gregory, 2009; Lozano and Mouat,
2009), the transport of non-native marine species (e.g. bryozoans)
to new habitats on floating plastic debris (Barnes, 2002; Derraik,
2002; Winston, 1982), and the smothering of the seabed, prevent-
ing gas-exchange and creating artificial hard-grounds, resulting
from sinking plastic debris (Gregory, 2009; Moore, 2008).
In recent years, there has been increasing environmental concern
about ‘microplastics’: tiny plastic granules used as scrubbers in cos-
metics and air-blasting, and small plastic fragments derived from
the breakdown of macroplastics (Derraik, 2002; Ryan et al., 2009;
Thompson et al., 2004). The presence of small plastic fragments in
the open ocean was first highlighted in the 1970s (Carpenter and
Smith, 1972), and a renewed scientific interest in microplastics over
the past decade has revealed that these contaminants are wide-
spread and ubiquitous within the marine environment, with the
potential to cause harm to biota (Rands et al., 2010; Sutherland
et al., 2010). Owing to their small size, microplastics are considered
bioavailable to organisms throughout the food-web. Their composi-
tion and relatively large surface area make them prone to adhering
waterborne organic pollutants and to the leaching of plasticisers
that are considered toxic. Ingestion of microplastics may therefore
be introducing toxins to the base of the food chain, from where there
is potential for bioaccumulation (Teuten et al., 2009).
The objectives of this review are: (1) to summarise the proper-
ties, nomenclature and sources of microplastics; (2) to discuss the
routes by which microplastics enter the marine environment; (3)
to evaluate the methods by which microplastics are detected in
the marine environment; (4) to ascertain spatial and temporal
trends of microplastic abundance; and (5) to determine the envi-
ronmental impact of microplastics.
2. Microplastics
Whilst macroplastic debris has been the focus of environmental
concern for some time, it is only since the turn of the century that
tiny plastic fragments, fibres and granules, collectively termed
‘‘microplastics’’, have been considered as a pollutant in their own
right (Ryan et al., 2009; Thompson et al., 2004). Microplastics have
been attributed with numerous size-ranges, varying from study to
study, with diameters of <10 mm (Graham and Thompson, 2009),
<5 mm (Barnes et al., 2009; Betts, 2008), 2–6 mm (Derraik, 2002),
<2 mm (Ryan et al., 2009) and <1 mm (Browne et al., 2007; Browne
et al., 2010; Claessens et al., 2011). This inconsistency is particu-
larly problematic when comparing data referring to microplastics,
making it increasingly important to create a scientific standard
(Claessens et al., 2011; Costa et al., 2010). Recently, Andrady
(2011) has suggested adding the term ‘‘mesoplastics’’ to scientific
nomenclature, to differentiate between small plastics visible to
the human eye, and those only discernible with use of microscopy.
2.1. Primary microplastics
Plastics that are manufactured to be of a microscopic size are
defined as primary microplastics. These plastics are typically used
in facial-cleansers and cosmetics (Zitko and Hanlon, 1991), or as
air-blasting media (Gregory, 1996), whilst their use in medicine
as vectors for drugs is increasingly reported (Patel et al., 2009). Un-
der the broader size definitions of a microplastic, virgin plastic pro-
duction pellets (typically 2–5 mm in diameter) can also be
considered as primary microplastics, although their inclusion
within this category has been criticised (Andrady, 2011; Costa
et al., 2010).
Microplastic ‘‘scrubbers’’, used in exfoliating hand cleansers and
facial scrubs, have replaced traditionally used natural ingredients,
including ground almonds, oatmeal and pumice (Derraik, 2002;
Fendall and Sewell, 2009). Since the patenting of microplastic
scrubbers within cosmetics in the 1980s, the use of exfoliating
cleansers containing plastics has risen dramatically (Fendall and
Sewell, 2009; Zitko and Hanlon, 1991). Typically marketed as ‘‘mi-
cro-beads’’ or ‘‘micro-exfoliates’’, these plastics can vary in shape,
size and composition depending upon the product (Fendall and
Sewell, 2009). For example, Gregory (1996) reported the presence
of polyethylene and polypropylene granules (<5 mm) and polysty-
rene spheres (<2 mm) in one cosmetic product. More recently, Fen-
dall and Sewell (2009) reported an abundance of irregularly shaped
microplastics, typically <0.5 mm in diameter with a mode size
<0.1 mm, in another cosmetic product.
Primary microplastics have also been produced for use in air-
blasting technology (Derraik, 2002; Gregory, 1996). This process
involves blasting acrylic, melamine or polyester microplastic
scrubbers at machinery, engines and boat hulls to remove rust
and paint (Browne et al., 2007; Derraik, 2002; Gregory, 1996). As
these scrubbers are used repeatedly until they diminish in size
and their cutting power is lost, they will often become contami-
nated with heavy metals (e.g. Cadmium, Chromium, Lead) (Derraik,
2002; Gregory, 1996).
2.2. Secondary microplastics
Secondary microplastics describe tiny plastic fragments derived
from the breakdown of larger plastic debris, both at sea and on
land (Ryan et al., 2009; Thompson et al., 2004). Over time a culmi-
nation of physical, biological and chemical processes can reduce
the structural integrity of plastic debris, resulting in fragmentation
(Browne et al., 2007).
Over prolonged periods, exposure to sunlight can result in
photo-degradation of plastics; ultraviolet (UV) radiation in sun-
light causes oxidation of the polymer matrix, leading to bond
cleavage (Andrady, 2011; Barnes et al., 2009; Browne et al.,
2007; Moore, 2008; Rios et al., 2007). Such degradation may result
in additives, designed to enhance durability and corrosion resis-
tance, leaching out of the plastics (Talsness et al., 2009). The cold,
M. Cole et al. / Marine Pollution Bulletin 62 (2011) 2588–2597 2589
haline conditions of the marine environment are likely to prohibit
this photo-oxidation; plastic debris on beaches, however, have
high oxygen availability and direct exposure to sunlight so will de-
grade rapidly, in time turning brittle, forming cracks and ‘‘yellow-
ing’’ (Andrady, 2011; Barnes et al., 2009; Moore, 2008). With a loss
of structural integrity, these plastics are increasingly susceptible to
fragmentation resulting from abrasion, wave-action and turbu-
lence (Barnes et al., 2009; Browne et al., 2007). This process is
ongoing, with fragments becoming smaller over time until they
become microplastic in size (Fendall and Sewell, 2009; Rios et al.,
2007; Ryan et al., 2009). It is considered that microplastics might
further degrade to be nanoplastic in size, although the smallest
microparticle reportedly detected in the oceans at present is
1.6
l
m in diameter (Galgani et al., 2010). The presence of nano-
plastics in the marine environment is likely to be of increasing sig-
nificance in the years to come, and researchers, including Andrady
(2011), have already begun to speculate on the impact that such a
pollutant might have on the base of the marine food web.
The development of biodegradable plastics is often seen as a
viable replacement for traditional plastics. However, they too
may be a source of microplastics (Thompson et al., 2004). Biode-
gradable plastics are typically composites of synthetic polymers
and starch, vegetable oils or specialist chemicals (e.g. TDPA™)
designed to accelerate degradation times (Derraik, 2002; O’Brine
and Thompson, 2010; Ryan et al., 2009; Thompson et al., 2004)
that, if disposed of appropriately, will decompose in industrial
composting plants under hot, humid and well-aerated conditions
(Moore, 2008; Thompson, 2006). However, this decomposition is
only partial: whilst the starch components of the bio-plastic will
decompose, an abundance of synthetic polymers will be left behind
(Andrady, 2011; Roy et al., 2011; Thompson et al., 2004). In the rel-
atively cold marine environment, in the absence of terrestrial mi-
crobes, decomposition times of even the degradable components
of bio-plastics will be prolonged, increasing the probability of the
plastic being fouled and subsequently reducing UV permeation
on which the degradation process relies (Andrady, 2011; Moore,
2008; O’Brine and Thompson, 2010). Once decomposition does fi-
nally occur, microplastics will be released into the marine environ-
ment (Roy et al., 2011).
3. Sources and transfer of microplastics into the marine
environment
Marine litter results from the indiscriminate disposal of waste
items that are either directly or indirectly transferred to our seas
and oceans (Lozano and Mouat, 2009; Ryan et al., 2009). In this sec-
tion, we look at several sources of plastic litter and discuss both di-
rect and indirect routes by which plastic can enter the marine
environment. Whilst the emphasis of this review is on microplas-
tics, in this section we also consider the indiscriminate disposal
of macroplastics, as, with time, they have the potential to degrade
into secondary microplastics.
Plastic litter with a terrestrial source contributes 80% of the
plastics found in marine litter (Andrady, 2011). Such plastics in-
clude primary microplastics used in cosmetics and air-blasting,
improperly disposed ‘‘user’’ plastics and plastic leachates from re-
fuse sites. With approximately half the world’s population residing
within fifty miles of the coast, these kinds of plastic have a high po-
tential to enter the marine environment via rivers and wastewater-
systems, or by being blown off-shore (Moore, 2008; Thompson,
2006). Microplastics used both in cosmetics and as air-blasting
media can enter waterways via domestic or industrial drainage
systems (Derraik, 2002); whilst waste-water treatment plants will
trap macroplastics and some small plastic debris within oxidation
ponds or sewage sludge, a large proportion of microplastics will
pass through such filtration systems (Browne et al., 2007; Fendall
and Sewell, 2009; Gregory, 1996). Plastics that enter river systems
– either directly or within waste-water effluent or in refuse site
leachates – will then be transported out to sea. A number of studies
have shown how the high unidirectional flow of freshwater sys-
tems drives the movement of plastic debris into the oceans
(Browne et al., 2010; Moore et al., 2002). Using water samples from
two Los Angeles (California, USA) rivers collected in 2004–2005,
Moore (2008) quantified the amount of plastic fragments present
that were <5 mm in diameter. Extrapolating the resultant data re-
vealed that these two rivers alone would release over 2 billion
plastic particles into the marine environment over a 3-day period.
Extreme weather, such as flash flooding or hurricanes, can exacer-
bate this transfer of terrestrial debris from land to sea (Barnes
et al., 2009; Thompson et al., 2005). Work conducted by Moore
et al. (2002) showed neustonic litter (small, surface plastic debris)
<4.75 mm in diameter in Californian waters near the mouth of a
modified Los Angeles stormwater conveyance system increased
from 10 plastic items/m
3
to 60 plastic items/m
3
following a storm.
The work further showed how increased water volume in the river,
due to the recent storm, resulted in litter being deposited at even
greater distances from the river mouth. Similarly, in a study by Lat-
tin et al. (2004), microplastic concentrations 0.8 km off the south-
ern Californian coast jumped from an average <1 item/m
3
,to18
items/m
3
following a storm.
Coastal tourism, recreational and commercial fishing, marine
vessels and marine-industries (e.g. aquaculture, oil-rigs) are all
sources of plastic that can directly enter the marine environment,
posing a risk to biota both as macroplastics, and as secondary
microplastics following long-term degradation. Tourism and recre-
ational activities account for an array of plastics being discarded
along beaches and coastal resorts (Derraik, 2002), although it is
worth noting that marine debris observed on beaches will also
arise from beaching of materials carried on in-shore- and ocean
currents (Thompson, 2006). Fishing gear is one of the most com-
monly noted plastic debris items with a marine source (Andrady,
2011). Discarded or lost fishing gear, including plastic monofila-
ment line and nylon netting, is typically neutrally buoyant and
can therefore drift at variable depths within the oceans. This is par-
ticularly problematic due to its inherent capacity for causing
entanglement of marine biota, known as ‘‘ghost fishing’’ (Lozano
and Mouat, 2009). Historically, marine vessels have been a signifi-
cant contributor to marine litter, with estimates indicating that
during the 1970s the global commercial fishing fleet dumped over
23,000 tons of plastic packaging materials (Pruter, 1987). In 1988,
an international agreement (MARPOL 73/78 Annex V) was imple-
mented banning marine vessels from disposing of plastic waste
at sea; however, it is widely considered that a lack of enforcement
and education has resulted in shipping remaining a dominant
source of plastic in the marine environment (Derraik, 2002; Lozano
and Mouat, 2009), contributing an estimated 6.5 million tons of
plastic to the oceans in the early 1990s (Derraik, 2002).
Another notable source of plastic debris stems from the manu-
facture of plastic products that use granules and small resin pellets,
known as ‘nibs’, as their raw material (Ivar do Sul et al., 2009; Mato
et al., 2001; Pruter, 1987). In the US alone, production rose from
2.9 million pellets in 1960 to 21.7 million pellets by 1987 (Pruter,
1987). Through accidental spillage during transport, both on land
and at sea, inappropriate use as packing materials and direct out-
flow from processing plants, these raw materials can enter aquatic
ecosystems. In an assessment of Swedish waters using an 80
l
m
mesh, KIMO Sweden found typical microplastic concentrations of
150–2, 400 microplastics/m
3
, but in a harbour adjacent to a plastic
production facility, the concentration was 102,000/m
3
(Lozano and
Mouat, 2009). However, resin pellets are by no means localised:
they have been identified in marine systems worldwide, including
mid-ocean islands with no local plastic production facilities (Ivar
2590 M. Cole et al. / Marine Pollution Bulletin 62 (2011) 2588–2597
do Sul et al., 2009; Pruter, 1987). Concentrations of these pellets can
also be highly variable: studies conducted in the 1970s and 1980s
revealed pellet concentrations of 18/km
2
off the New Zealand coast,
but 3, 500/km
2
in the Sargasso Sea (Pruter, 1987). In 1991, under
the auspices of the US Environmental Protection Agency, many
American plastic manufacturers voluntarily committed to prevent-
ing or recapturing spilled pellets, an agreement that may explain
significant decreases in quantities of resin pellets identified in the
North Atlantic between 1986 and 2008 (Law et al., 2010). More
recently, Operation Cleansweep (www.opcleansweep.org), a joint
initiative of the American Chemistry Council and Society of the
Plastics Industry, is aiming for industries to commit to zero pellet
loss during their operations.
4. Assessing microplastic abundance
Within the marine environment, plastic is widely considered the
primary constituent of ‘marine debris’, a category that includes
both anthropogenic litter (e.g. glass, metal, wood), and naturally
occurring flotsam (e.g. vegetation, pumice; Barnes et al., 2009;
Moore, 2008; Ryan et al., 2009; Thompson et al., 2004). However,
small plastic debris (<0.5 mm in diameter) is considered a widely
under-researched component of marine debris (Doyle et al., 2011)
due to the difficulties in assessing the abundance, density and dis-
tribution of this contaminant within the marine environment.
Quantifying the input of plastics into the marine environment is
precluded by the array of pathways by which plastics may enter
the oceans and would require accurate timescales of the length at
which plastics remain at sea prior to degradation (Ryan et al.,
2009). Meanwhile, quantifying debris that has already reached
the marine environment is complicated by the vastness of the
oceans compared to the size of the plastics being assessed. Spatial
and temporal variability owing to oceanic currents and seasonal
patterns further complicate this issue (Doyle et al., 2011; Ryan
et al., 2009). Nevertheless, a suite of sampling techniques has been
developed that allow the presence of small plastic debris to be
determined. These include: (1) beach combing; (2) sediment sam-
pling; (3) marine trawls; (4) marine observational surveys; and
(5) biological sampling.
Beach combing is considered the easiest of the available tech-
niques to conduct, requiring little logistical planning and relatively
low costs (MCS, 2010). Typically carried out by researchers and
environmental awareness groups, this technique involves collect-
ing and identifying all litter items, in a systematic approach, along
a specified stretch of coastline. By repeating the beach combing
process on a regular basis, accumulation of plastic debris can be
monitored over time (Ryan et al., 2009). This technique is particu-
larly useful for determining the presence of macroplastics and
plastic resin pellets, termed ‘Mermaid’s Tears’ by beach combers,
but microplastics, especially those too small to be observed by
the naked eye, are likely to go unnoticed using such a technique.
Furthermore, as plastic debris along a coastline will consist of both
litter left by recreational beach users and debris deposited by the
sea, it must be considered that beach combing data represents a
mix of terrestrial litter and marine debris, and therefore may not
provide an accurate indicator of plastic debris in the marine
environment itself (OSPAR, 2007).
Sediment sampling allows benthic material from beaches,
estuaries and the seafloor to be assessed for the presence of micro-
plastics (Claessens et al., 2011). To separate any plastics from the
benthic material, saline water or mineral salts can be added to
the sediment samples to increase water density, permitting
lower-density microplastics to be separated via flotation. Visible,
denser plastic fragments can be removed by hand under a micro-
scope (Andrady, 2011; Thompson et al., 2004). A lipophilic dye
(e.g. Nile Red) can then be used to stain the plastics to assist iden-
tification using a range of microscopy techniques (Andrady, 2011).
Using Fourier-Transform Infrared Spectroscopy (FT-IR), items of
interest can then be confirmed as plastic by comparing spectra of
the samples with that of known polymers (Barnes et al., 2009;
Thompson et al., 2004).
Microplastics within the water column can be collected by con-
ducting a trawl along a transect (i.e. manta trawls for sampling sur-
face water, bongo nets for collecting mid-water levels and benthic
trawls to assess the seabed) using fine meshes (Browne et al., 2010;
Ryan et al., 2009; Thompson et al., 2004). The presence of micro-
plastics can then be determined by examining the samples under
a microscope, or allowing evaporation of the seawater and investi-
gating the residue left behind (Andrady, 2011). Despite the heter-
ogeneous nature of plastics within the ocean, sufficient transects
and repeats allow for both spatial and temporal patterns in plastic
abundance to be determined in a variety of marine ecosystems
(Ryan et al., 2009). Typically, 330
l
m aperture meshes have been
used for many of the microplastic trawls documented in this re-
view, but it is important to note that using meshes with different
apertures can produce large variations in the quantity of micro-
plastics collected: by utilising 80
l
m meshes, KIMO Sweden found
microplastics at 100,000 times higher concentrations than when
using 450
l
m meshes (Lozano and Mouat, 2009). In contrast, an
Algalita Marine Research Foundation survey of the North Pacific
central gyre, conducted in 1999, identified 9,470 plastic fragments
with a 1 mm mesh, but decreasingly smaller quantities of finer
sized particles when using smaller-aperture meshes (4,646 micro-
plastics with a 0.5 mm mesh, and just 2,626 microplastics using a
0.3 mm mesh) (Moore, 2008). Long-term data from Continuous
Plankton Recorders (CPRs) are of particular benefit to determining
microplastic abundance in the open ocean. These are specialised
units designed to constantly sample plankton within 280
l
m silk-
screen-meshes, whilst being towed behind vessels along fixed
routes (Thompson et al., 2004). Archived CPR samples, held by
the Sir Alastair Hardy Foundation for Ocean Science (SAHFOS) have
helped evaluate the prevalence of microplastics in the Northwest
Atlantic throughout the past fifty years. The importance of CPR
data in assessing microplastic abundance has led SAHFOS to in-
clude the presence of microplastics in their analysis of all future
samples (Richardson et al., 2006; Thompson et al., 2004).
Marine observational surveys allow divers or observers on boats
and in submersibles to record the size, type and location of visible
plastic debris. While this technique is effective at detecting macro-
plastics over relatively large areas, microplastics will often go
undetected, and – as debris is not collected – the litter can undergo
no further assessment (Pruter, 1987; Ryan et al., 2009). Further-
more, the subjective nature of observational work leaves such cen-
suses open to bias (Ryan et al., 2009).
Finally, biological sampling involves examining plastic frag-
ments consumed by marine biota. A number of marine organisms
can mistake plastic debris for prey (Blight and Burger, 1997; Tour-
inho et al., 2010; van Franeker et al., 2011). By dissecting beached
marine animals, or by instigating regurgitation in some seabirds,
their gut contents can be analysed for the presence of plastics,
which can then be identified and quantified (van Franeker, 2010).
The Fulmar has routinely been used to assess the abundance of
plastic debris at sea for some time and the abundance of micro-
plastics within the stomachs of Fulmars has now become one of
the ecological quality assessment markers used by OSPAR to assess
the abundance of plastic debris at sea (van Franeker et al., 2011).
Whilst migration and movement of this ocean foraging seabird
precludes matching their plastic load with specific locales, regional
differences and trends over time have become apparent (Blight and
Burger, 1997; Tourinho et al., 2010; van Franeker, 2010).
M. Cole et al. / Marine Pollution Bulletin 62 (2011) 2588–2597 2591
5. Spatial and temporal trends of microplastics in the marine
environment
Plastic litter has permeated marine ecosystems across the globe
(Derraik, 2002; Lozano and Mouat, 2009; Ryan et al., 2009). Driven
by ocean currents, winds, river outflow and drift (Barnes et al.,
2009; Martinez et al., 2009; Ng and Obbard, 2006) plastic debris
can be transported vast distances to remote, otherwise pristine,
locations, including mid-ocean islands (Ivar do Sul et al., 2009), the
poles (Barnes et al., 2010) and the ocean depths (Lozano and Mouat,
2009). However, whilst plastic litter may be found throughout the
marine environment, the distribution of this debris is heterogeneous
(Martinez et al., 2009; Moore, 2008). In this section we discuss how
microplastics accumulate along coastlines and within mid-ocean
gyres, examine the variable position of microplastics within the
water column and consider microplastic abundance over time.
5.1. Accumulation of microplastics
Coastlines receive plastic litter from both terrestrial and marine
sources, terrestrial sources of litter will typically dominate close to
urban areas, sites of tourism and near river outflows, whilst marine
debris will be deposited along shorelines when caught in near-
shore currents (Ryan et al., 2009). Using sediment analysis, Thomp-
son et al. (2004) have found microplastics, consisting of nine differ-
ent polymers, in 23 of 30 estuarine, beach and sub-tidal sediment
samples taken around Plymouth, UK, including microscopic fibres
and fragments typically derived from clothing, packaging and rope.
Further work showed that microplastics were present in beach
sediments throughout the UK. Browne et al. (2010) used the same
methodology to quantify microplastics in sediment throughout the
Tamar estuary (Plymouth, UK), identifying 952 items in 30 sedi-
ment samples. An abundance of microplastics have also been
found in productive coastal ecosystems off Alaska and California,
where nutrient upwelling results in high densities of planktonic
organisms (Doyle et al., 2011). Using 505
l
m meshes during sur-
face plankton trawls for the National Oceanic and Atmospheric
Administration (NOAA), Doyle et al. (2011) found an abundance
of plastic fragments derived from the breakdown of larger plastic
debris, in addition to plastic fibres and pellets, although concentra-
tions were significantly lower than those found in the adjacent
North Pacific gyre. The source of this plastic debris was unable to
be verified, however, it was suggested that the high concentration
of plastics in southern Californian waters during winter was linked
to urban run-off from major conurbations, whilst a marine source
was more likely during the summer months when currents altered.
After conducting beach surveys throughout the remote mid-Atlan-
tic archipelago of Fernando de Noronha, Ivar do Sul et al. (2009)
identified plastic pre-production resin pellets on the windward
beaches of the archipelago – yet no plastic-production facilities ex-
ist in the region. Therefore, it was hypothesised that they were
brought to the remote location via trans-oceanic currents before
being trapped in in-shore currents and washed ashore. Similarly,
a survey of beaches on the island of Malta, in the Mediterranean
Sea, found an abundance of disc- and cylindrical-shaped plastic re-
sin pellets (1.9–5.6 mm in diameter) on all beaches surveyed
(Turner and Holmes, 2011). The highest concentrations of pellets,
in some cases in excess of 1,000 pellets/m
2
, were found along the
high-tide mark, the majority of the pellets were yellow or brown
in colour, caused by photo-oxidative damage indicative of their
longevity within the marine environment. The presence of so many
plastics on a shoreline can dramatically alter the physio-chemical
properties of the beach sediment. In a recent study, vertical sedi-
ment cores were taken from beaches in Hawaii and analysed
(Carson et al., 2011). The presence of plastic debris not only in-
creased the permeability of the sediment, but also decreased its
heat absorbance so that the sediment would reach lower maximal
temperatures than sediment without plastics present. Such differ-
ences could affect marine biota, for example, lower maximal tem-
peratures might affect sex-determination in turtle eggs, and
greater permeability will increase the probability of desiccation
in sediment-dwelling organisms.
Oceanographic modelling indicates a large proportion of float-
ing debris reaching the ocean will accumulate in gyres – the centre
of vast anti-cyclonic, sub-tropical ocean currents. Using satellite-
tracked ‘‘drifters’’ placed throughout the South Pacific ocean, Mar-
tinez et al. (2009) mapped the average trajectories of ocean cur-
rents, drift and eddies over time, the team found that, whilst
some trackers were caught in near-shore currents, the majority
fed into the south Pacific gyre from where they could not easily es-
cape (Law et al., 2010; Martinez et al., 2009). Lagrangian drifters
have also been used in a more recent study, indicating a high pro-
portion of floating marine debris will end up in ocean gyres (Max-
imenko et al., in press). Data accumulated from over 6,000
plankton tows conducted between 1986 and 2008 in the North
Atlantic Ocean and Caribbean Sea, found plastic in 60% of the sam-
ples (Law et al., 2010). Mapping the plastic concentrations of each
transect, Law et al. (2010) revealed distinct spatial patterns of plas-
tic in these areas, with highest concentrations (83% of total plastic
sampled) found in sub-tropical latitudes. The highest concentra-
tion was mapped to the North Atlantic gyre, with 20, 328
(±2,324) pieces/km
2
. Due to the concentrations of plastic found it
was impossible to determine the sources of such debris, but use
of trackers suggested much of the eastern seaboard of the US fed
into the gyre, taking debris 60 days on average to reach the gyre si-
ted over 1,000 km away. Even higher plastic concentrations have
been recorded in the North Pacific gyre: conducting 11 transects
using a 333
l
m manta-trawl, Moore et al. (2001) identified plastics
in the majority of their tows, with an average density of 334,271
plastic fragments/km
2
. Such work has led to significant media
attention, with the North Pacific gyre being described ‘‘plastic
soup’’ and coined as the ‘‘great Pacific garbage patch’’ (Kaiser,
2010).
5.2. Microplastics in the water column
Plastics consist of many different polymers and, depending on
their composition, density and shape, can be buoyant, neutrally-
buoyant or sink. As such, microplastics may be found throughout
the water column. Low-density microplastics are predominantly
found in the sea-surface microlayer, as documented by numerous
studies presenting data from surface trawls (Derraik, 2002; Greg-
ory, 1996). However, there is evidence that their position in the
water column can vary: in estuarine habitats, low-density plastics,
such as polypropylene and polyethylene, will be submerged if they
meet water fronts. Furthermore, there is growing evidence that the
attachment of fouling organisms can cause buoyant microplastics
to sink (Barnes et al., 2009; Browne et al., 2010; Derraik, 2002;
Thompson et al., 2004). Plastic debris in the marine environment
can rapidly accumulate microbial biofilms, which further permit
the colonisation of algae and invertebrates on the plastics’ surface,
thus increasing the density of the particle (Andrady, 2011). The
speed at which biofouling may occur was recently demonstrated
using polyethylene plastic bags submerged in seawater (16.2 °C)
in Plymouth harbour (UK), a biofilm was visible after just one
week, and analysis showed a significant increase in microbial den-
sity over the 3-week experiment (Lobelle and Cunliffe, 2011).
Notably, the plastic became less buoyant over time, and by the
end of the experiment the plastic moved away from the surface
and appeared neutrally buoyant. When assessing plastic litter in
the North Pacific gyre, Moore et al. (2001) randomly sampled deb-
ris for signs of fouling organisms. Only a small proportion (8.5%) of
2592 M. Cole et al. / Marine Pollution Bulletin 62 (2011) 2588–2597
surface debris was colonised, and fouling decreased with particle
size. However, at a depth of 10 m, a higher proportion of plastics
debris was fouled with algae and diatoms. More recently, an
analysis of microplastics (<1 mm) collected in surface tows from
the western North Atlantic Ocean between 1991 and 2007, has
shown evidence of fouling (Morét-Ferguson et al., 2010). The study
found low-density polymers (e.g. polypropylene and polyethylene)
with higher densities than the same polymer found on beaches,
concluding the increase in density resulted from biofouling at
sea. Despite increases of plastic debris entering the marine envi-
ronment throughout the last century, Law et al. (2010) found no
significant change in microplastic abundance in the Northwest
Atlantic over the past twenty years. To test whether new input of
microplastics was compensated for by sedimentation of biofouled
plastics to greater depths, they analysed material from sediment
traps deployed at 500 to 3,200 m depths close to the north Atlantic
gyre, but found no significant accumulation of plastic particles. The
fate of fouled microplastics in gyres has now become a key re-
search area for the 5 Gyres Project, in association with the Algalita
Marine Research Foundation (AMRF) (Eriksen and Cummins,
2010).
High-density microplastics, including polyvinylchloride, polyes-
ter and polyamide, are likely found in their largest quantities in the
benthos. However, determining the magnitude of microplastic deb-
ris on the seafloor is hindered by cost and difficulties of sampling
(Barnes et al., 2009). While ‘Fishing for Litter’ schemes, conducted
in the Netherlands and Scotland, and submersible video-recordings
can document the quantity of macroplastics present on the seafloor
(Lozano and Mouat, 2009; Watters et al., 2010), microplastics will
fall below the lower limits of detection of these sampling methods.
Therefore, quantification of microplastics in the benthos relies on
sediment-grabs and benthic trawls using fine meshes. A recent
study has found some of the highest microplastic concentrations
within sediment thus far. Microplastics, <1 mm in diameter, consist-
ing of fibres, granules, pellets and films, were found in all beach, har-
bour and sub-littoral sediment samples taken off the Belgian coast
(Claessens et al., 2011), the highest microplastic concentration
(391 microplastics/kg of dry sediment) was found in a harbour
sediment sample, probably due to the local anthropogenic activity,
river run-off and trapping of sediments. It has been documented that
high-density microplastics can be temporarily suspended within the
water-column in smaller numbers resulting from turbulence. High-
density microplastics can remain in suspension when entering the
sea through estuaries due to tidal fronts, high-flow rate or because
of a large-surface area (Browne et al., 2010). Only when momentum
is lost will these dense polymers inevitably sink (Barnes et al., 2009).
Microplastics on the seabed may also be re-suspended resulting
from turbulence: Lattin et al. (2004) quantified microplastic concen-
trations >333
l
m at varying depths, 0.8 and 4.5 km off the southern
Californian coast. At the off-shore site, microplastics were most
abundant close to the seafloor (6 items/m
3
), but were redistributed
throughout the water column after a storm (Lattin et al., 2004).
5.3. Temporal changes in microplastic abundance within the marine
environment
Since the 1940s, when the mass production of plastics began in
earnest, the volume of plastic produced has risen rapidly. With leg-
islation to curb the indiscriminate disposal of plastic waste emerg-
ing slowly, plastic debris entering the marine environment
increased in parallel with rates of production during this time
(Moore, 2008;Ryan et al., 2009; Barnes et al., 2009). Continuous
fragmentation of larger plastic debris and the rising popularity of
‘‘plastic scrubbers’’ appears to have increased the volume of micro-
plastic debris in the oceans, resulting in a decrease in the average
size of plastic litter over time (Barnes et al., 2009). This was high-
lighted by Thompson et al. (2004), who demonstrated that micro-
plastic concentrations in the 1980s and 1990s were significantly
greater than those in the 1960s and 1970s in an analysis of CPR
samples from the North Sea and Northwest Atlantic. Furthermore,
incidence of plastic ingestion by Fulmars (ocean-foraging seabirds),
washed ashore in the Netherlands, increased from 91% to 98% be-
tween the 1980s and 2000, whilst the average consumption dou-
bled from 15 to 30 plastic fragments per bird during this period
(van Franeker et al., 2011).
Concentration trends within the past decade are not overtly
apparent, and there is some debate as to whether levels of plastic
debris are still increasing or have stabilised. The study by Thomp-
son et al. (2004) indicated minimal change in microplastic con-
tamination between the 1980s and 1990s. Similarly, an
evaluation of >6, 100 surface trawls conducted throughout the
Northwest Atlantic Ocean found no significant difference in
microplastic abundance over a 22 year period (Law et al., 2010).
The average number of plastics debris items consumed by Ful-
mars, beached on the shores of the Netherlands, decreased
slightly from the mid-1990s, but has remained relatively stable
since the turn of the century, currently averaging 26 plastic frag-
ments per bird (van Franeker et al., 2011). In contrast, Claessens
et al. (2011) indicate that microplastic concentrations have stea-
dily increased over the past two decades. Analysis of sediment
cores taken along the Belgian coast indicates microplastic pollu-
tion tripled from 55 microplastics/kg of dry sediment (1993–
2000) to 156 microplastics/kg of dry sediment (2005–2008), in
line with global production rates. However, use of sediment cores
is a new technique, and bio-turbation from tourism or sediment-
dwelling biota might have affected this data.
Any further conclusions are hampered by both a lack of studies
that have specifically considered trends of microplastic abundance
over time. Meta-studies are difficult to develop due to varieties of
sampling methodologies, huge spatial variations in microplastic
abundance, and lack of standardised size definitions of microplas-
tics (Ryan et al., 2009; Barnes et al., 2009).
6. Impact of microplastics on the marine environment
Whilst it is apparent that microplastics have become both wide-
spread and ubiquitous, information on the biological impact of this
pollutant on organisms in the marine environment is only just
emerging (Barnes et al., 2009; Gregory, 1996; Ryan et al., 2009).
The possibility that microplastics pose a threat to biota, as their
small size makes them available to a wide range of marine organ-
isms, is of increasing scientific concern (Barnes et al., 2009; Der-
raik, 2002; Fendall and Sewell, 2009; Lozano and Mouat, 2009;
Ng and Obbard, 2006; Thompson et al., 2004). In addition to poten-
tial adverse effects from ingesting the microplastics themselves,
toxic responses could also result from (a) inherent contaminants
leaching from the microplastics, and (b) extraneous pollutants, ad-
hered to the microplastics, disassociating.
6.1. Microplastic ingestion
Owing to their small size and presence in both pelagic and ben-
thic ecosystems, microplastics have the potential to be ingested by
an array of marine biota (Betts, 2008; Thompson et al., 2009a).
Observing microplastic ingestion in the wild is methodologically
challenging (Browne et al., 2008), but an increasing number of
studies are reporting microplastic ingestion throughout the food-
chain.
Table 1 lists a number of laboratory experiments demonstrating
that marine organisms, including zooplankton, invertebrates and
echinoderm larvae, ingest microplastics (Bolton and Havenhand,
M. Cole et al. / Marine Pollution Bulletin 62 (2011) 2588–2597 2593
1998; Brillant and MacDonald, 2002; Hart, 1991; Wilson, 1973).
Furthermore, phagocytic uptake of nanoplastics in a heterotrophic
ciliate has been demonstrated using fluorescent nanospheres (Pace
and Bailiff, 1987). These lower-trophic level organisms are partic-
ularly susceptible to ingesting microplastics as many of them are
indiscriminate feeders with limited ability to differentiate between
plastic particles and food (Moore, 2008). A study investigating the
colour and size distribution of microplastics in the North Pacific
Ocean hypothesised that planktonic organisms will most com-
monly mistake white and lightly-coloured plastic fragments for
prey (Shaw and Day, 1994). As low-density ‘‘user’’ plastics (e.g.
polyethylene and polystyrene) are buoyant, microplastics are
abundant near the sea surface. Therefore, microplastics will be
widely available to a host of planktonic organisms, including the
larval stages of a variety of commercially important species that re-
side within the euphotic zone (Fendall and Sewell, 2009; Gregory,
1996). This contact between plankton and microplastics is hypo-
thetically exacerbated in gyres, as plankton populations are low
whilst microplastic concentrations are high, resulting from plastic
accumulation by ocean currents (Moore, 2008).
A range of marine biota, including seabirds, crustaceans and
fish, can ingest microplastics (Blight and Burger, 1997; Tourinho
et al., 2010). Plastic fragments were first identified in the guts of
sea birds in the 1960s, when global plastic production was less
than 25 million tonnes per annum (Ryan et al., 2009; Thompson
et al., 2009b). In 1982, a team in the Netherlands found 94% of ful-
mars sampled contained plastics, with an average of 34 plastic
fragments per individual. Since, incidence and number of frag-
ments consumed has remained high, although the mass of plastic
found in each bird has decreased significantly in recent years (Loz-
ano and Mouat, 2009; van Franeker, 2010). Dissection of planktiv-
orous mesopelagic fish, caught in the North Pacific central gyre,
revealed microplastics in the guts of 35% of the fish sampled
(Boerger et al., 2010). Plastic fibres, fragments and films were also
found in the stomachs of 13 of 141 mesopelagic fish caught in the
North Pacific gyre (Davison and Asch, 2011). In the Clyde Sea (Scot-
land), 83% of Nephrops sp. collected had ingested plastics. This
commercially important, omnivorous, benthic-dwelling crustacean
mainly ate sections of monofilament line and fragments of plastic
bags (Murray and Cowie, 2011). Plastic fibres found in the environ-
ment can be as small as 1
l
m in diameter, and 15
l
m in length,
making them available to minute planktonic species (Frias et al.,
2010). Such fibres may be particularly hazardous as they may
clump and knot, potentially preventing egestion (Murray and Cow-
ie, 2011). In all these examples, these animals might have ingested
microplastics voluntarily, which they confuse for their prey. Alter-
natively, microplastic ingestion may result from eating lower tro-
phic organisms that have themselves consumed microplastics
(Browne et al., 2008; Fendall and Sewell, 2009). This process was
recently demonstrated by providing small fish, which had previ-
ously eaten plastic fibres, to Nephrops sp., after a 24-h exposure
period, all the Nephrops sp. had plastic fibres in their guts from eat-
ing the fish (Murray and Cowie, 2011).
It is yet to be established whether the ingestion of non-polluted
microplastics have any significant adverse health effects on biota
(e.g. morbidity, mortality or reproductive success) (Zarfl et al.,
2011). Microplastics may present a mechanical hazard to small
animals once ingested, similar to the effects observed for macro-
plastics and larger animals (Barnes et al., 2009; Fendall and Sewell,
2009): plastic fragments might block feeding appendages or hinder
the passage of food through the intestinal tract (Tourinho et al.,
2010) or cause pseudo-satiation resulting in reduced food intake
(Derraik, 2002; Thompson, 2006). However, Thompson (2006)
and Andrady (2011) note that numerous marine organisms have
the ability to remove unwanted materials (e.g. sediment, natural
detritus and particulates) from their body without causing harm,
as demonstrated using polychaete worms, which ingested micro-
plastics from their surrounding sediment, then egested them in
their faecal casts (Thompson et al., 2004). Nevertheless, once in-
gested, there is the potential for microplastics to be absorbed into
the body upon passage through the digestive system via transloca-
tion. Translocation of polystyrene microspheres was first shown in
rodents and humans, and has also been demonstrated for mussels
using histological techniques and fluorescence microscopy
(Browne et al., 2008). Mytilus edulis were able to ingest 2 and
4
l
m microplastics via the inhalant siphon, which the gill filtered
out and transported to the labial palps for digestion or rejection.
Translocation was proven following the identification of 3 and
9.6
l
m fluorescently tagged microspheres in the mussels’ haemo-
lymph (circulatory fluid), 3 days after exposure. Microspheres were
present in the circulatory system for up to 48 days after exposure,
although there was no apparent sub-lethal impact (measured as
oxidative status and phagocytic ability of the haemocytes) (Browne
et al., 2008). However, Köhler (2010) describes a pronounced im-
mune response and granuloma formation in the digestive glands
of blue mussels exposed to microplastics.
6.2. Microplastics and plasticiser leachates
Although plastics are typically considered as biochemically in-
ert (Roy et al., 2011; Teuten et al., 2009), plastic additives, often
termed ‘‘plasticisers’’, may be incorporated into plastics during
manufacture to change their properties or extend the life of the
plastic by providing resistance to heat (e.g. polybrominateddiphe-
nyl ethers), oxidative damage (e.g. nonylphenol) and microbial
degradation (e.g. triclosan) (Browne et al., 2007; Thompson et al.,
2009b). These additives are an environmental concern since they
both extend the degradation times of plastic and may, in addition,
leach out, introducing potentially hazardous chemicals to biota
(Barnes et al., 2009; Lithner et al., 2011; Talsness et al., 2009).
Table 1
Laboratory studies showing uptake of microplastics in marine biota.
Organism(s) Microplastic
(
l
m)
Identification technique Publication
Copepods (Acartia tonsa) 7–70 Microscopy Wilson (1973)
Echinoderm larvae 10–20 Video observation Hart (1991)
Trochophore larvae (Galeolaria caespitosa) 3–10 Microscopy Bolton and Havenhand
(1998)
Scallop (Placopecten magellanicus) 16–18 Detection of
51
Cr labelled
particles
Brillant and MacDonald
(2002)
Amphipod (Orchestia gammarellus), Lugworm (Arenicola marina) & Barnacle
(Semibalanus balanoides)
20–2000 Dissection and wormcast
examination
Thompson et al. (2004)
Mussel (Mytilus edulis) 2–16 Dissection and fluorescence
microscopy
Browne et al. (2008)
Sea cucumbers Various Excrement analysis Graham and Thompson
(2009)
2594 M. Cole et al. / Marine Pollution Bulletin 62 (2011) 2588–2597
Incomplete polymerisation during the formation of plastics al-
lows additives to migrate away from the synthetic matrix of plas-
tic, the degree to which these additives leach from plastics is
dependent on the pore size of the polymer matrix, which varies
by polymer, the size and properties of the additive and environ-
mental conditions (e.g. weathering; Moore, 2008; Ng and Obbard,
2006; Teuten et al., 2009). For example, phthalates are emollients
that soften plastics by reducing the affinity between molecular
chains within the synthetic polymer matrix (Oehlmann et al.,
2009; Talsness et al., 2009). In PVC, phthalates can constitute up
to 50% of the plastic’s weight (Oehlmann et al., 2009). Meanwhile,
Bisphenol A is a constituent monomer in polycarbonate which is
widely used in food and beverage containers. Neither compound
is persistent, but their instability within plastic products facilitates
leaching and their high prevalence in aquatic environments has
been widely reported, particularly in landfill leachates (vom Saal
and Myers, 2008).
Due to the large surface-area-to-volume ratio of microplastics,
marine biota may be directly exposed to leached additives after
microplastics are ingested. Such additives and monomers may inter-
fere with biologically important processes, potentially resulting in
endocrine disruption, which in turn can impact upon mobility,
reproduction and development, and carcinogenesis (Barnes et al.,
2009; Lithner et al., 2009, 2011). Commonly used additives, includ-
ing polybrominated diphenyl ethers, phthalates and the constituent
monomer bisphenol A, are renowned for being endocrine-disrupt-
ing chemicals as they can mimic, compete with or disrupt the syn-
thesis of endogenous hormones (Talsness et al., 2009). Hormonal
imbalance can cause permanent morphological issues in organisms
in developmental stages, or sexual disruption in adults. Phthalates
have been associated with a range of molecular and whole-organism
effects in aquatic invertebrates and fish, including genotoxic dam-
age (micronuclei and apoptosis in mussel haemocytes), inhibited
locomotion in invertebrates and intersex conditions in fish (Oehl-
mann et al., 2009). Bisphenol A is both an oestrogen agonist and
an androgen antagonist that can differentially affect reproduction
and development depending on its concentration and the organism
affected, at concentrations in the region of
l
g/l, Bisphenol A can be
acutely toxic to both crustaceans and insects. Chronic and wide-
spread exposure of human populations to Bisphenol A has further
been associated with chronic health effects, including heart disease,
diabetes and alterations in circulating hormone levels (Galloway
et al., 2010; Lang et al., 2008). Although it has been shown that plas-
ticisers can induce negative biological effects within the ng/l–
l
g/l
range, Oehlmann et al. (2009) note there has been relatively little re-
search into the chronic effects of these additives in long-term expo-
sures to aquatic species.
6.3. Microplastics and adhered pollutants
Marine plastic debris, in particular microplastics with their
large surface area to volume ratio, are susceptible to contamination
by a number of waterborne-pollutants, including aqueous metals
(Betts, 2008; Ashton et al., 2010), endocrine disrupting chemicals
(Ng and Obbard, 2006) and persistent organic pollutants (POPs),
also referred to as hydrophobic organic contaminants (HOCs) (Rios
et al., 2007).
Such chemicals are typically found at their highest concentra-
tions in the sea-surface microlayer, where low-density microplas-
tics are most abundant as well (Ng and Obbard, 2006; Rios et al.,
2007; Teuten et al., 2009). POPs, which include polychlorinated
biphenyls (PCBs), PAHs and organochlorine pesticides (e.g. DDT,
DDE), are stable, lipophillic chemicals that will adhere and concen-
trate on the hydrophobic surface of plastics, with environmental
concentrations recorded in the ng/g–
l
g/g range (Teuten et al.,
2007, 2009; Barnes et al., 2009). Using equilibrium partitioning
modelling, the adsorption coefficients (Kd) of the priority pollutant
phenanthrene were calculated for a range of plastic polymers in
seawater and natural sediments (Teuten et al., 2007). Phenan-
threne readily sorbs to small plastics, preferentially adhering to
polyethylene, likely due to larger molecular cavities in this poly-
mer. In environmentally relevant conditions, phenanthrene was
more likely to adhere to plastics than to sediment. However, if
heavily polluted microplastics come into contact with non-con-
taminated sediments, the concentration gradient would permit
desorption of phenanthrene to organic matter in the sediment.
Evidence of microplastic contamination has been highlighted by
several studies conducted in recent years. Mato et al. (2001) identi-
fied PCBs, nonylphenol and DDE on polypropylene resin pellets col-
lected from Japanese waters at similar or higher concentrations than
those found in sediments. In a further experiment, virgin resin pel-
lets were shown to adsorb contaminants from seawater within a
6-day exposure period. Although adsorption was constant, maximal
concentrations were not reached in this time, indicating adsorption
is not a rapid process. Rios et al. (2007) used GC–MS to detect sorbed
contaminants on plastic pellets in Japanese waters, 4,4-DDE was
found on all samples, up to a concentration of 5,600 ng/g, and PCBs
were observed on all but four samples with concentrations of 39–
1, 200 ng/g. Teuten et al. (2007) observed PCBs at concentrations
10
6
higher on polystyrene pellets than in surrounding water. Micro-
plastics found on two Portuguese beaches contained PAH concentra-
tions ranging from 0.2 to 319.2 ng/g, and PCBs from 0.02 to 15.56 ng/
g(Frias et al., 2010). Analysis of plastic fragments (<10 mm) sampled
from pelagic and neritic stations, revealed a range of pollutants
including PCBs, PAHs, DDTs and its metabolites, PBDEs and bisphe-
nol A were adhered to the plastics’ surface at concentrations of 1–10,
000 ng/g (Hirai et al., 2011).
Microplastic debris coated with POPs may be transported across
oceans polluting otherwise pristine ecosystems (Zarfl and Mat-
thies, 2010), or be ingested by marine organisms, thus transferring
toxins from the environment to biota (i.e. a ‘‘Trojan horse’’ effect)
(Gregory, 1996; Thompson et al., 2005, 2004). Many POPs are con-
sidered toxic, inducing endocrine disruption, mutagenesis and/or
carcinogenesis, and may biomagnify in higher-trophic organisms.
However, until recent years it was unclear whether contaminants
adhered to plastic detritus would disassociate once ingested
Table 2
Key requirements to answer research gaps relating to microplastics in the marine environment.
1. Employ a clear and standardised size definition of a microplastic, with further size definitions for nano- and mesoplastics
2. Optimise and implement routine, high-throughput microplastic sampling methodologies to better compare the results from different study areas
3. Develop appropriate methods for detecting minute microplastics and nanoplastics within the water-column and sediment
4. Expand knowledge of the fate and behaviour of microplastics within the water-column, including the effects of fragmentation and bio-fouling
5. Develop methods for determining microplastic uptake by biota throughout the marine food-web and expand the use of sentinel species (e.g. Fulmars) in detecting
microplastic abundance
6. Determine the impact (i.e. mortality, morbidity and/or reproduction) of ingested microplastics on marine biota, and understand the transfer of this contaminant
within the food-chain
7. Determine the impact (i.e. mortality, morbidity and/or reproduction) of leached plastic additives and adsorbed waterborne pollutants to biota, transferred via
microplastics, on marine biota
M. Cole et al. / Marine Pollution Bulletin 62 (2011) 2588–2597 2595
(Thompson et al., 2004). To determine whether pollutants adhered
to microplastics could desorb and cause harm to biota, Teuten
et al. (2007) used a partitioning model to assess the disassociation
of phenanthrene on microplastics. The model indicated that con-
taminated microplastics ingested by Arenicola marina, a sediment-
dwelling polychaete worm, will sequester a proportion of the sorbed
contaminants to the organism. However, if inhabiting clean, organ-
ic-rich sediment, much of the contaminant was predicted to adhere
to the sediment rather than be taken up by the polychaete itself
(Teuten et al., 2007, 2009). Transfer of contaminants from plastic
to biota has since been demonstrated. Streaked shearwater chicks
were fed with a diet of fish and resin pellets, or fish alone (Betts,
2008; Teuten et al., 2009). Both pellets and fish were obtained from
Tokyo Bay and were contaminated with polychlorinated biphenyls
(PCBs), at concentrations of 51–562 ng/g for the plastics, and 0.3–
0.7 ng/g for fish. Analysis of preen gland oil, taken every week for
42 days, showed that PCB concentrations increased in both groups
of chicks. To determine the uptake of PCBs from the resin pellets
alone, lower chlorinated congener PCBs, which were abundant in
the resin pellets but in low concentrations in fish, were analysed.
Chicks eating plastic pellets showed a significant increase in low
congener PCBs, whilst those eating fish alone showed no change.
7. Conclusions and recommendations for future work
Over the past decade, increased scientific interest has produced
an expanding knowledge base for microplastics. Nevertheless, fun-
damental questions and issues remain unresolved. An evolving suite
of sampling techniques has revealed that microplastics are a ubiqui-
tous and widespread marine contaminant, present throughout the
water column. However, disparity in the size definitions of micro-
plastics and lack of comparability of microplastic sampling method-
ologies hinder our ability to cross-examine quantitative studies to
better determine spatial and temporal patterns of this contaminant.
The highest abundance of microplastics is typically associated with
coastlines and mid-ocean gyres, but the fate of these microplastics is
elusive. It is hypothesised that microplastics sink following biofoul-
ing, fragment into smaller and smaller polymer fragments and/or
are ingested by marine biota. Fully testing such hypotheses is im-
peded by the complexity of sampling the ocean depths and the dif-
ficulty of routinely sampling and detecting smaller-sized fractions
of microplastics (including nanoplastics). Laboratory and field-stud-
ies have shown the consumption of microplastics in a range of mar-
ine biota, although it remains unclear whether microplastic
ingestion alone will result in adverse health effects (e.g. mortality,
morbidity and reproductive success) or whether such a contaminant
can routinely be passed up the food chain. The transfer of toxic
chemicals to biota via microplastic ingestion is a significant concern.
However, few existing studies have conducted toxicity-studies
using microplastic vectors. Looking to the future, here we present
a list of knowledge gaps we believe deserve further attention from
the scientific community (Table 2).
Acknowledgements
Matthew Cole is supported by a NERC Ph.D. studentship. This
work was supported by Grant ME5413 from the Department of
the Environment, Fisheries and Rural Affairs, UK.
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In recent years, the environmental pollution of microplastics has attracted much attention. To date, there have been a lot of researches on microplastics and a series of studies published. In this study, by bibliometric analysis method to evaluated the development and evolution on microplastics research trends and hot spots. A total of 2872 literature information was collected from the Web of Science (2004–2020), which was used for bibliometric visual analysis by CiteSpace. It was possible to see the contributing countries, institutions, authors, keywords, and future study directions in the microplastics sectors by looking at the visual representation of the results. (1) Since 2004, scientific advancements in this sector have advanced significantly, with a significant increase in speed since 2012. (2) China and the United States are the world's leading researchers in microplastics. (3) The study of microplastics was multidisciplinary, comprising researchers from the fields of ecology, chemistry, molecular biology, environmental science, and oceanography. (4) In recent years, researchers have concentrated their attention on the distribution and toxicity of microplastics in the environment, as well as their coupled pollution with heavy metal contaminants. In conclusion microplastics study in environmental science has become increasingly popular in recent years. Topics include dispersion, toxicity, and coupled pollution with heavy metal pollutants. Researchers in a wide range of fields are involved in microplastics research. Furthermore, policies and regulations about microplastics in global were summarized, and membrane technology has potential to remove microplastics from water. The above findings help to clearly grasp the content and development trend of microplastics research, point out the future research direction for scholars, and promote microplastics research and pollution prevention and control.
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Plastic products can contain chemicals that are hazardous to human health and the environment. In this study, it was investigated if various plastic products emit hazardous chemical substances to water. Two leaching methods (batch and diffusion tests) were used and the leachates were tested for acute toxicity to Daphnia magna. Nine out of 32 tested plastic product leachates had Daphnia 48-h EC(50)s ranging from 5 to 80 g plastic material L(-1). For the remaining 23 products no effect on mobility was seen even at the highest test concentrations (70-100 g plastic material L(-1)). A compact disc (recordable) was the most toxic plastic product, but the toxicity was traced to the silver layer not the polycarbonate plastic material. The other products that displayed toxicity were made of either plasticized PVC (artificial leather, bath tub toy, inflatable bathing ring and table cloth) or polyurethane (artificial leather, floor coating and children's handbag). While the Toxicity Identification Evaluation (TIE) for compact discs using sodium thiosulfate addition showed that silver was causing the toxicity, the TIE for artificial leather using C18 cartridges showed that hydrophobic compounds were causing the toxicity. Acute toxicity tests of plastic product leachates were found to be useful for screening purposes for differentiating between toxic and non-toxic products.
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Beach collection of plastic litter washed ashore at Fort Pierce on the Atlantic coast of Florida showed that, while some of this material may have had a local source, much of the rest originated in the Caribbean. During its pelagic existence the plastic commonly became encrusted by marine organisms. In contrast to the diverse assemblage of organisms associated with pelagic Sargassum, however, the number of species encrusting drift plastic was limited, being dominated primarily by the colonies of the bryozoan, Electra tenella, an organism which may be increasing its abundance and distribution, due to increasing amounts of drift plastic substrata.