Marine Microbial Assemblages On Microplastics: Diversity, Adaptation, and Role in Degradation

MA12CH08_Labrenz ARjats.
cls November 16, 2019 9:59
Annual Review of Marine Science

Marine Microbial Assemblages
on Microplastics: Diversity,
Adaptation, and Role
in Degradation
Sonja Oberbeckmann and Matthias Labrenz
Department of Biological Oceanography, Leibniz Institute for Baltic Sea Research Warnemünde
(IOW), D-18119 Rostock, Germany; email: matthias.labrenz@io-warnemuende.de
Annu. Rev. Mar. Sci. 2020. 12:209–32 Keywords

First published as a Review in Advance on
ocean, microplastics, biofilms, plastisphere, biodegradation, evolution
June 21, 2019
The Annual Review of Marine Science is online at Abstract

marine.annualreviews.org
We have known for more than 45 years that microplastics in the ocean are
https://doi.org/10.1146/annurev-marine-010419-
carriers of microbially dominated assemblages. However, only recently has
010633
the role of microbial interactions with microplastics in marine ecosystems
Copyright © 2020 by Annual Reviews.
been investigated in detail. Research in this field has focused on three main
All rights reserved
areas: (a) the establishment of plastic-specific biofilms (the so-called plasti-
sphere); (b) enrichment of pathogenic bacteria, particularly members of the
genus Vibrio, coupled to a vector function of microplastics; and (c) the micro-
bial degradation of microplastics in the marine environment. Nevertheless,
the relationships between marine microorganisms and microplastics remain
unclear. In this review, we deduce from the current literature, new compara-
tive analyses, and considerations of microbial adaptation concerning plastic
degradation that interactions between microorganisms and microplastic
particles should have rather limited effects on the ocean ecosystems. The
majority of microorganisms growing on microplastics seem to belong to
opportunistic colonists that do not distinguish between natural and artificial
surfaces. Thus, microplastics do not pose a higher risk than natural particles
to higher life forms by potentially harboring pathogenic bacteria. On the
other hand, microplastics in the ocean represent recalcitrant substances for
microorganisms that are insufficient to support prokaryotic metabolism and
will probably not be microbially degraded in any period of time relevant
to human society. Because we cannot remove microplastics from the ocean,
proactive action regarding research on plastic alternatives and strategies to
prevent plastic entering the environment should be taken promptly.
209
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1. THE GENESIS OF A NONDEGRADABLE MULTIPURPOSE PRODUCT

AND ITS ENTRY INTO THE MARINE ENVIRONMENT
Since the Paleolithic era, a hallmark of the human experience has been the development of tools
essential for survival. Biopolymers of high molecular mass, such as wood, wool, flax, hemp, bones,
resin, and latex, have served to make tools, clothes, and residences, but their benefit is limited
by their restricted formability, modifiability, consistency, and vulnerability to degradation. As a
consequence, humans tried very early on to modify biopolymers and create materials of better
quality for their daily needs. Dating back to the Middle Pleistocene, the distillation of birch bark
and modification of polymers led to the generation of birch-bark tar that has been used as an
adhesive (Kozowyk et al. 2017). Ancient Mesoamericans harvested latex [high-molecular-weight
poly(cis-1,4-isoprene)] from Castilla elastica, processed it using Ipomoea alba juice, and fashioned
rubber tools (Tarkanian & Hosler 2011), a process that dates back to at least 1600 BC. Around
3,000 years later, in 1839, Charles Goodyear developed vulcanization, a major breakthrough in the
development of new and synthetic polymers. Originally, vulcanization involved the cross-linking
of linear poly(cis-1,4-isoprene) chains by sulfur bridges, leading to superior physical properties
(Akiba & Hashim 1997). The first synthetic polymer with industrial relevance was Bakelite; in
the early years of the twentieth century, Leo Hendrik Baekeland discovered and published a tech-
nique to produce this phenol- and formaldehyde-based polymer (Baekeland 1909), which could
be pressed and hardened into shape, resisting mechanical damage, heat, and acids. Since then, and
especially after the 1950s, the development of new synthetic polymers with various societal bene-
fits increased significantly, making plastics an easily producible and almost indispensable product
(Andrady & Neal 2009, Thompson et al. 2009). Plastic is now crucial in such diverse areas as
medicine, building and construction, packaging, electronics, and aeronautics.
More than 348 million tons of plastic was produced worldwide in 2017 (PlasticsEurope 2018).
The European demand for polymers was highest for low- and high-density polyethylene (17.5%
and 12.3%, respectively), polypropylene (19.3%), polyvinyl chloride (10.2%), polyurethane
(7.7%), polystyrene (7.4%), and polyethylene terephthalate (7.4%). Of the remaining ∼20%,
polyamide, acrylonitrile butadiene styrene, and polycarbonate represent the industrially impor-
tant polymers (PlasticsEurope 2018). A certain portion of this plastic ends up in the global oceans
(Barnes et al. 2009, Galgani et al. 2013, Law 2017), and studies have estimated that floating marine
plastic totals somewhere between 70,000 and 270,000 tons (Cozar et al. 2014, Eriksen et al. 2014,
van Sebille et al. 2015). On the other hand, estimates suggest that 4.8–12.7 million tons of plastic
litter was introduced into marine systems in 2010 alone ( Jambeck et al. 2015). This means that
only 1% of plastic introduced in marine systems is recovered as floating debris, indicating that the
particles only remain on the ocean surface for a limited period of time (Cozar et al. 2014, Eriksen
et al. 2014). Thus, because the fate and behavior of plastic in the marine system are still largely
unclear, there is a large gap between knowledge of plastic entry into the ocean and knowledge of
its retention.
Human-generated litter enters the oceans from land and offshore (Law 2017, Sheavly &
Register 2007). Approximately 80% of the marine plastic debris originates from land-based
sources (Andrady 2011), whether discarded directly into the environment by beach-related
tourism, improperly managed farther inland and blown into the sea, or introduced by rivers, mu-
nicipal drainage systems, or sewage effluents (Andrady 2011, Auta et al. 2017a, Barnes et al. 2009,
Browne et al. 2011, Derraik 2002, Pruter 1987). The most substantial offshore source is the world’s
fishing fleet (Andrady 2011), followed by marine aquaculture (Hinojosa & Thiel 2009); illegally
discarded litter from vessels or offshore platforms (Sheavly & Register 2007) and the content of
lost cargo containers also contribute (Derraik 2002). The relative prevalence of macro- and mi-
croplastics in the ocean is still unknown and is the subject of current research.
210 Oberbeckmann • Labrenz

Microplastics (particles less than 5 mm in size) are divided into two groups according to their
origin. So-called primary microplastics are directly synthesized for consumer products, such as
hand and facial cleansers, shower gels, or toothpaste (Fendall & Sewell 2009, Gregory 2009).
Along with clothing fibers rinsed out by washing machines, these microplastics can reach the sea
via sewage effluents (Browne et al. 2011). Primary microplastics also include abrasive materials
from the air-blasting industry and virgin preproduction resin pellets lost during transport (Ogata
et al. 2009). Hence, primary microplastics are already on a millimeter size scale when they reach
aquatic environments. Secondary microplastics, on the other hand, are formed from larger floating
plastic fragments as a result of fragmentation in the environment (Arias-Villamizar & Vazquez-
Morillas 2018, Cooper & Corcoran 2010); fragmentation is a main topic in the second part of this
review and is discussed in more detail in Section 4.
The ubiquitous presence of microplastics in the marine environment has been demonstrated.
These plastics not only concentrate in the large ocean gyres but are also found everywhere from
the polar regions (e.g., Peeken et al. 2018) to the equator, from densely populated areas to re-
mote islands (Ivar do Sul et al. 2009), and from beaches (Claessens et al. 2011) down to the deep
sea (Van Cauwenberghe et al. 2013). Plastic particles are found floating at the sea surface, sus-
pended in the water column, and contained in sediments, depending on their density relative
to seawater. Most consumer polymers are less dense than seawater, but the density of a virgin
polymer particle is altered when the end product is manufactured (e.g., increased due to fillers
or decreased by foaming), as well as through aging and biofouling (Harrison et al. 2011, Kaiser
et al. 2017). Consequently, the polymer composition found in environmental samples depends
on the sampling depth. Most of the microplastic particles found in neustonic samples are low-
and high-density polyethylene, polypropylene, and expanded polystyrene (Moret-Ferguson et al.
2010, Reisser et al. 2014). Sediment studies have also reported the accumulation of polymers such
as polyamide, solid polystyrene, polyvinyl chloride, polyurethane, polyester, polyethylene tereph-
thalate, acrylic polyoxymethylene, polyvinyl alcohol, polymethyl methacrylate, and alkyd (Browne
et al. 2010, Hidalgo-Ruz et al. 2012, Moret-Ferguson et al. 2010). Beaches serve as intermediate
environments, connecting land-based debris with aquatic ecosystems, and their sediments can ac-
cumulate all polymer densities. The highest concentration of microplastic particles, at 1.2 × 107
particles per cubic meter, was recently detected in a core taken from the pack ice of Fram Strait
(Peeken et al. 2018).
Dissolved organic pollutants in the ocean interact with microplastics, and this interaction is
dependent on the physicochemical properties of the organic compounds. Microplastics have a
large surface-to-volume ratio when compared with relatively larger plastics and thus are more
likely to accumulate hydrophobic organic substances (Engler 2012, Mato et al. 2001), including
persistent organic pollutants such as polychlorinated biphenyls, polycyclic aromatic hydrocarbons,
and organic chlorine compounds such as DDT. In addition to the polymer, plastics usually contain
potentially toxic plasticizers or additives, which leach into the environment. Therefore, research
is being carried out to determine whether microplastics cause a significant accumulation of toxic
substances in the marine food web. Currently, however, the ecological consequences cannot be
assessed.
2. EXISTING KNOWLEDGE AND RESEARCH GAPS FOR MARINE

MICROPLASTIC BIOFILMS
Besides accumulating organic pollutants, microplastic surfaces serve as colonization grounds for
diverse microbial communities in aquatic habitats (De Tender et al. 2015, Dussud et al. 2018,
Hoellein et al. 2017, McCormick et al. 2014, Oberbeckmann et al. 2014, Zettler et al. 2013).
www.annualreviews.org • Marine Microbial Assemblages on Microplastics 211

The collection of microbial communities inhabiting plastic debris is commonly termed the
plastisphere.
Marine microplastic biofilms are shaped primarily by biogeographical and environmental fac-
tors, such as salinity and nutrient concentration (Amaral-Zettler et al. 2015, Oberbeckmann et al.
2018). Although not as strong a contributor, the microplastic surfaces themselves also influence
the colonization processes, and there has been discussion of whether specific overrepresented and
potentially hydrocarbonoclastic members of marine microplastic biofilms could use plastics as
an energy source because of their ability to degrade highly complex biopolymers such as lignin
and petroleum derivatives (Oberbeckmann et al. 2016, 2018; Ogonowski et al. 2018; Zettler et al.
2013).
In recent years, increasing concern has been raised about the hazard potential of microplastic-
associated microbial communities. Potential pathogens might be distributed into formerly unaf-
fected ecosystems while hitchhiking on microplastics that originated, for instance, from sewage
treatment plants or animal guts (Oberbeckmann et al. 2015). Studies have suggested that members
of the genus Vibrio are particularly enriched on microplastics (Frere et al. 2018, Zettler et al. 2013),
but others have disputed the preferential colonization of microplastics by Vibrio spp. (Bryant et al.
2016, Oberbeckmann et al. 2018, Schmidt et al. 2014). Besides colonization with pathogens, the
role of microplastics as carriers for antibiotic resistance genes has been discussed. In this context,
two freshwater studies compared microplastic assemblages with their corresponding water com-
munities and demonstrated that the microplastic-associated assemblages had an increased transfer
frequency of a plasmid coding for trimethoprim resistance (Arias-Andres et al. 2018) and higher
abundance of the gene int1, a proxy for anthropogenic pollution (Eckert et al. 2018).
Knowledge of the microbial composition of biofilms associated with microplastics has in-
creased in the last five years (Ivar do Sul et al. 2018). The first groundbreaking studies on marine
biofilms on microplastics showed that these communities can differ significantly from those of
the surrounding water (Amaral-Zettler et al. 2015, Bryant et al. 2016, Debroas et al. 2017, Frere
et al. 2018, Oberbeckmann et al. 2014, Zettler et al. 2013). This can be expected, however, since
microbial community compositions on natural particles usually differ from free-living microor-
ganisms (Crespo et al. 2013, Rieck et al. 2015) due to their dissimilar lifestyles as sessile organisms.
Thus, until it is established whether there is a true microplastic-indicative plastisphere, as com-
pared with natural-particle-associated assemblages (e.g., those on wood, cellulose, or glass), the
role of biofilms on microplastics will remain obscure.
The aim of this review is to critically evaluate the potential role of marine microorganisms in
relation to ocean-polluting microplastics. We present the current state of knowledge of marine mi-
crobial assemblages on microplastics, with a focus on pathogenic bacteria, and attempt to deduce
their impact on marine ecosystems and potential risk for humans. We evaluate whether plastic-
specific microbial communities may indicate microplastic degradation, or at least the possibility
of adaptation of microorganisms to its degradation. Based on this evaluation, we critically assess
whether marine bacteria have the potential to remediate the ocean from plastic pollution in the
future and conclude with recommendations for further action in scientific and societal contexts.
3. AN EVALUATION OF THE PLASTISPHERE

3.1. Data Selection for Meta-Analysis
Our approach to evaluating the plastisphere was to merge the relevant worldwide data and re-
analyze the microbial communities in a comparative way. We focused on bacterial 16S rRNA
gene databases because of the large number of comparable (harmonized) data sets available.
When selecting sequences for the meta-analysis, we focused on polyethylene samples as a model

polymer surface because it represents the most common plastic pollutant in the ocean. Additional
sequences came from polystyrene, unknown polymers, natural (i.e., control) surfaces (wood, cellu-
lose, and glass), sediment, and, if available, the particle-attached water fraction (>3 μm) or other-
wise the whole surrounding water. In November 2018 we performed independent searches in Web
of Science using the following keywords: microplastic(s) and biofilm(s), microplastic(s) and micro-
bial/bacterial assemblages, microplastic(s) and colonization/colonisation, and microplastic(s) and
plastisphere. This led to a list of 61 peer-reviewed publications, of which 16 represented com-
munity analyses regarding microplastics. We selected studies for the meta-analysis that met the
following criteria: (a) The study had been carried out in marine or brackish waters, (b) the study
used Illumina technology for sequencing, (c) the targeted 16S rRNA fragment contained the V4
region, (d) the data were deposited at the National Center for Biotechnology Information’s Se-
quence Read Archive, and (e) the description of the sampling source in the database was sufficient
(e.g., clear differentiation between seawater and plastic samples). Five studies met all of these cri-
teria: one from the North Sea (De Tender et al. 2015), three from the Baltic Sea (Kesy et al. 2019,
Oberbeckmann et al. 2018, Ogonowski et al. 2018), and one from the Yangtze Estuary ( Jiang et al.
2018). Supplemental Table 1 lists all of the data used in the meta-analysis, and the Supplemental
Appendix describes the methods applied for data processing and statistical analyses.
3.2. Geographical Factors Play a Larger Role than the Particle Surface in
Shaping Biofilms
Our comparative analyses revealed that the plastic itself was a minor factor in determining
microplastic-associated biofilms. Instead, the first-order determinant shaping the bacterial assem-
blages was the sampling area (i.e., geographical region), which discriminated the communities into
distinct clusters, as supported by pairwise PERMANOVA (p = 0.001, North Sea versus Baltic Sea
versus Yangtze Estuary) (Figure 1).
The average similarity between communities from the Baltic Sea and Yangtze Estuary—both
ecosystems strongly influenced by rivers—was 12%, slightly higher than the similarity between
communities from the Baltic and North Seas (7%). Within the area of the Baltic Sea, salinity (and
potentially other factors, which were not available for all studies) seemed to play an important
role in the community composition, as indicated by nonmetric multidimensional scaling ordina-
tion (Figure 1). Overall, the results of this reanalysis are in accordance with previous studies that
showed major differences among microplastic communities from distinct geographical regions
(e.g., Amaral-Zettler et al. 2015). Here, we can expand this statement and show that the influence
of a geographical region is greater than the influence of surface characteristics when comparing
plastic polymers with natural particle surfaces.
Experimental setups, laboratory handling, and extraction methods probably led to distinct bac-
terial sequencing results in different studies, indicating that the study procedures themselves likely
contributed to the differences between samples. This, however, cannot be proven here and is be-
yond the scope of this review.
3.3. Microbial Life on Microplastics and Natural Particles

No agreement has been reached on whether microplastic-associated communities display an in-
creased or decreased diversity compared with their counterparts on natural particles and in the wa-
ter. While some studies from aquatic ecosystems have reported similar or even higher α-diversities
on microplastics (De Tender et al. 2015, Debroas et al. 2017, Dussud et al. 2018, Frere et al.
2018), other studies have postulated the opposite (Hoellein et al. 2017, McCormick et al. 2014,

North Sea
a Baltic Sea
Yangtze Estuary
3.5 (Ogonowski et al. 2018)
6.5 (Oberbeckmann et al. 2018)
4.4–8.7 (Kesy et al. 2019)
14.7 (Oberbeckmann
et al. 2018)
Polyethylene
b Polystyrene
Unknown polymer
Control surface
(cellulose, glass, or wood)
Particle-attached
water fraction
Whole water fraction
Sediment
Figure 1
Nonmetric multidimensional scaling based on a Bray–Curtis similarity matrix of square-root-transformed
relative abundances for the five studies analyzed, with stress = 0.13. (a) Ordination based on the sampling
areas in the North Sea (De Tender et al. 2015), the Baltic Sea (Kesy et al. 2019, Oberbeckmann et al. 2018,
Ogonowski et al. 2018), and the Yangtze Estuary ( Jiang et al. 2018). The salinities of the three studies from
the Baltic Sea are also given. (b) Ordination based on the different sample types.
Zettler et al. 2013). Ogonowski et al. (2018) detected similar α-diversities among all tested biofilm
communities but higher values for water communities and explained this result with an overall
substrate-driven selection. Kettner et al. (2017) found that the fungal α-diversity of microplastic
assemblages was lower than or analogous to those of wood and water assemblages, depending on
spatial factors.
When investigating the β-diversity in our meta-analysis, bacterial communities associated with
different polymer types did not show significant differences (pairwise PERMANOVA, p > 0.01,

polyethylene versus polystyrene versus unknown; see Supplemental Table 2). Furthermore,
polystyrene-colonizing communities did not significantly differ from communities on natural con-
trol surfaces (p = 0.127). Communities from all other sample types were significantly different.
Some of these differences, however, were also identified as significant by PERMDISP compar-
isons (e.g., polyethylene versus control surface; Supplemental Table 2), hinting at dispersion
effects. This means that differences in variability within the sample data sets, rather than the ac-
tual sample characteristics, may have led to the significant result. The influence of the particle
surface on its colonization can be shaped by characteristics such as degradability, hydrophobicity,
electric charge, or roughness, or indirectly via the formation of a conditioning film over the par-
ticle. This probably explains why some studies have reported differences between communities
associated with microplastics and natural particles such as cellulose (Ogonowski et al. 2018), the
particle-attached water fraction (Dussud et al. 2018), or sediment (De Tender et al. 2015). Also,
an incubation experiment in the Baltic Sea demonstrated a differentiation between assemblages
on polyethylene and polystyrene from assemblages on wood (as a model of a natural particle),
but only under certain environmental conditions (Kettner et al. 2017, Oberbeckmann et al. 2018),
highlighting the importance of the sampling area in the development of the microbial biofilm.
Like other bacteria that prefer an attached over a free-living lifestyle (e.g., Nesse & Simm
2018), we can assume that overall most microplastic-biofilm members are opportunistic general
colonizers. Even early colonizers might be attracted not by the polymer surface itself but rather by
the conditioning film, which increases (for instance) their access to nutrients. This is the case for all
particles in suspension in the oceans and does not represent a microplastic-specific phenomenon
(e.g., Witt et al. 2011). The family Rhodobacteraceae, for example, is well known for its early
and abundant colonization of a broad range of particle surfaces (Dang & Lovell 2016, Elifantz
et al. 2013, Mata et al. 2017, Moura et al. 2018) and also colonizes polyethylene microplastics, as
revealed by our reanalysis (see Section 3.4).
Other opportunistic colonizers seem to be particularly successful in occupying microplastics as
their niche. Several studies have reported that members of the family Hyphomonadaceae thrive on
microplastics (Bryant et al. 2016, Oberbeckmann et al. 2018, Zettler et al. 2013), probably because
they are able to adhere firmly to the smooth plastic surface by forming the polysaccharide holdfast
and because their prosthecae enable more efficient nutrient uptake compared with other biofilm
members.
When analyzing all operational taxonomic units (OTUs) from our meta-analysis that reached
a mean relative abundance of more than 2% in at least one sample type, we found a high similarity
among bacterial communities associated with polyethylene, polystyrene, and natural particles
(Figure 2). However, some bacteria appeared to prefer polyethylene (OTU 0002, unclassified
Flavobacteriaceae; OTU 0006, unclassified γ-proteobacteria), polystyrene (OTU 0005, Hy-
drogenophaga; OTU 0011, Blastomonas; OTU 0027, Pseudomonas; and OTU 0048, Marinomonas),
or unknown polymer types (OTU 0017, Erythrobacter). Abundant OTUs associated with unknown
polymers were in several cases classified as members of the Sphingomonadaceae family, which the
core analysis showed play an important role in polyethylene-associated communities (see Section
3.4). We can deduce that, first, those unknown polymers are in fact polyethylene microplastics
or that, second, members of these family are particularly associated with microplastics, such as
polyethylene and polystyrene.
After the genera of the formerly independent family Erythrobacteraceae had been assigned to
the family Sphingomonadaceae, according to release 132 of the SILVA database, this family be-
came an important one—if not the most important one—associated with microplastic-associated
biofilms. Two characteristics of this bacterial family might lead to its dominance in microplas-
tic biofilms: its ability to degrade hydrocarbons and its formation of carotenoids. Due to their

Particle-associated
Control surface
water fraction
Polyethylene
Whole water
Polystyrene
Unknown
Sediment
polymer
fraction
OTU 0001 α-Proteobacteria; Rhodobacteraceae_uncl 0.4
OTU 0002 Bacteroidia; Flavobacteriaceae_uncl
Square root transformed

OTU 0003 Bacteroidia; Flectobacillus
0.3
relative abundance
OTU 0004 Planctomycetacia; Fuerstia
OTU 0005 γ-Proteobacteria; Hydrogenophaga
OTU 0006 γ-Proteobacteria; γ-Proteobacteria_uncl
0.2
OTU 0007 Bacteroidia; Flavobacteriaceae_uncl
OTU 0009 γ-Proteobacteria; Shewanella
OTU 0010 γ-Proteobacteria; Neptunomonas
OTU 0011 α-Proteobacteria; Blastomonas 0.1
OTU 0013 γ-Proteobacteria; Pseudomonas
OTU 0014 γ-Proteobacteria; Paraglaciecola
OTU 0017 α-Proteobacteria; Erythrobacter 0
OTU 0020 α-Proteobacteria; Hyphomonas
OTU 0023 Verrucomicrobiae; Prosthecobacter
OTU 0027 γ-Proteobacteria; Pseudomonas
OTU 0029 γ-Proteobacteria; Pseudoalteromonas
OTU 0030 α-Proteobacteria; Amylibacter
OTU 0046 Oxyphotobacteria; Snowella_0TU37S04
OTU 0048 γ-Proteobacteria; Marinomonas
OTU 0052 α-Proteobacteria; SAR11_clade_Ia
OTU 0065 α-Proteobacteria; Planktomarina
OTU 0073 γ-Proteobacteria; B2M28_ge
OTU 0075 Bacteroidia; Tenacibaculum
OTU 0106 Verrucomicrobiae; Persicirhabdus
OTU 0131 γ-Proteobacteria; Woeseia
OTU 0133 γ-Proteobacteria; γ-Proteobacteria_uncl
OTU 0166 Atribacteria; JS1_ge
OTU 0167 γ-Proteobacteria; OM43_clade
OTU 0176 δ-Proteobacteria; uncultured
OTU 0197 γ-Proteobacteria; SAR86 clade_ge
OTU 0210 α-Proteobacteria; Altererythrobacter
OTU 0230 α-Proteobacteria; SAR11_clade_II
OTU 0502 Bacteroidia; Lewinella
OTU 0676 Bacteroidia; Lewinella
OTU 0866 γ-Proteobacteria; Psychrosphaera
OTU 1031 α-Proteobacteria; Sphingomonas
Figure 2
Shade plot illustrating the relative abundances (square root transformed) of operational taxonomic units (OTUs) with a mean relative
abundance of more than 2% in at least one sample type. The displayed sample types are polyethylene (n = 52), polystyrene (n = 12),
unknown polymer (n = 4), control surface (cellulose, glass, or wood; n = 42), particle-attached water fraction (n = 34), whole water
fraction (n = 7), and sediment (n = 18). The hierarchical cluster of sample types on top is based on a Bray–Curtis similarity matrix using
square-root-transformed mean relative abundances of all OTUs from the meta-analysis (including the ones with an abundance of less
than 2%). The class and genus of the OTUs are given according to classification with release 132 of the nonredundant SILVA database.

ability to degrade aromatic or halogenated hydrocarbons, including petroleum and pesticides,

several members of the Sphingomonadaceae are already being used for remediation purposes
(Kertesz et al. 2017, Rosenberg et al. 2014). This ability also makes them prime candidates to
potentially degrade pollutants associated with microplastics, which either leach out of the plastics
or accumulate on them because of their hydrophobic surfaces (as is the case, e.g., for polycyclic
aromatic hydrocarbons). The formation of carotenoid pigments, on the other hand, protects bac-
terial cells from oxidative stress caused by UV light in the ocean surface, giving the members of
Sphingomonadaceae an advantage over nonpigmented bacteria. Therefore, this family could ei-
ther use the plastic or associated pollutants as an energy source (see Section 4.2) or take advantage
of the carotenoid pigments to more efficiently resist the UV light, as has been reported for bacteria
from the genus Erythrobacter (Matallana-Surget et al. 2012).
Studies have suggested that members of the genus Vibrio use plastic as a favorable transport
platform (Frere et al. 2018, Kirstein et al. 2016, Zettler et al. 2013), but Arcobacter spp., Colwellia
spp., Escherichia spp., Pseudomonas spp., and other taxa have also been discussed as potentially
pathogenic plastic colonizers (Curren & Leong 2019, Harrison et al. 2014, Keswani et al.
2016). In our reanalysis, no OTUs classified as Enterococcus sp. were found in the data set. In all
sample types (microplastics or natural particles), the mean relative abundances were very low
for members of Aeromonas (<0.23%), Colwellia (<0.08%), the taxonomic unit Escherichia–Shigella
(<0.01%), and the family Enterobacteriaceae (<0.03%). Members of the genera Arcobacter,
Pseudomonas, Shewanella, and Vibrio, which contain potentially pathogenic species, were indeed
associated with microplastics (Figure 3), but their median relative abundances remained below
those from communities associated with natural control surfaces (wood, cellulose, or glass) and/or
a b
Sediment
Particle-attached
water fraction
Control surface
Plastic
Sediment
c d
Particle-attached
water fraction
Control surface
Plastic
0.0001 0.001 0.01 0.1 1 10 100 0.0001 0.001 0.01 0.1 1 10 100
Relative abundances (%, log scale)

Figure 3
Box plots displaying the distribution of relative abundances (log scale) within the 5th–95th percentile of (a) Arcobacter, (b) Pseudomonas,
(c) Shewanella, and (d) Vibrio associated with different surfaces in marine and brackish waters. The different sample types comprise
plastic particles (polyethylene, polystyrene, and unknown polymers; n = 68), control surfaces (cellulose, glass, and wood; n = 42),
particle-attached water fractions (n = 34), and sediment (n = 18).

the particle-attached water fraction. The reanalysis reveals that microplastics do not per se
represent a higher risk to transport or enrich marine pathogenic microorganisms when compared
with natural particles. Nonetheless, the high durability of plastics, potentially enabling associated
microorganisms to travel longer distances horizontally as well as vertically in the oceans, can
be significant when compared with many natural particles, which biodegrade in shorter time
periods. Considering the sampling area as a major determinant on the community composition
(Amaral-Zettler et al. 2015, Oberbeckmann et al. 2018; see Section 3.2), we can assume that
microplastic-associated bacterial communities will rapidly adapt their composition to changing
environments rather than remaining stable over long distances. Once again, the nature of the
particle will have a small role in the spread of pathogens over large areas.
Another concern related to potential pathogens transported by plastics (including microplas-
tics) is the introduction of invasive species, especially microbial eukaryotes (Barnes & Fraser 2003,
Goldstein et al. 2014, Tutman et al. 2017). In particular, the 2011 tsunami in Japan raised concerns
about the distribution of living organisms via plastic debris (Miller et al. 2018). For example, Maso
et al. (2003) reported the colonization of floating plastic by harmful dinoflagellates (Ostreopsis sp.,
Coolia sp., and Alexandrium taylori) when sampling during a bloom of A. taylori. Likewise, poten-
tially harmful diatom species were found to raft on plastic fragments in Mediterranean coastal
waters (Maso et al. 2016). The specific processes regarding the potential spread of invasive mi-
croorganisms, including harmful microalgae, need to be further investigated.
3.4. The Core Bacterial Community Associated with Polyethylene Microplastics

Remains to Be Established
Four of the data sets in our meta-analysis contained sequences clearly attributed to polyethylene,
our model plastic chosen to screen for a core of associated bacteria across different sampling areas.
While most OTUs occurred in just one of the data sets, 45 OTUs were detected in all of them
(Figure 4, Supplemental Table 3). Of these, 27 were classified as α-proteobacteria, in partic-
ular within the families Rhodobacteraceae (11 OTUs) and Sphingomonadaceae (8 OTUs). The
relatively higher abundances of these families within microplastic biofilms have been reported by
several additional studies (Bryant et al. 2016, Curren & Leong 2019, Debroas et al. 2017, Dussud
et al. 2018) that did not meet all of the criteria for inclusion in our reanalysis (see Section 3.1) and
were discussed above (see Section 3.3). A mean relative abundance of greater than 1% across all
polyethylene samples was reached by five of the core OTUs, which were classified as members
of the families Rhodobacteraceae and Flavobacteriaceae (unclassified at the genus level) and the
genera Albirhodobacter, Methylotenera, and Hydrogenophaga.
By excluding bacterial communities that were also associated with polystyrene, unknown poly-
mers, and natural particles, we were able to identify 13 of the core 45 OTUs specifically associated
with polyethylene, including members of the Microbacteriaceae (3 OTUs) and Sphingomon-
adaceae (3 OTUs). Most of these polyethylene-specific OTUs, however, had low abundances
(<0.1%) within the polyethylene data set, with only OTU 0040 (classified as Erythrobacter) ex-
ceeding this threshold, with a mean relative abundance of 0.47%. These first results indicate that
very few bacterial communities are associated exclusively with polyethylene microplastics.
3.5. Microplastics Potentially Impact Ecological Processes in the Oceans

Microplastics are relatively recently introduced particles in the ocean and are increasing in amount
over time; consequently, the ratio of attached to free-living communities and the microbial
biomass colonizing those microplastics are also increasing. Therefore, regardless of the properties

Oberbeckmann et al. 2018 Ogonowski et al. 2018
245 370
Kesy et al. 2019 De Tender et al. 2015
51
100 9
47 8
180 1,109
45
47 390
6 112
64
Figure 4
Venn diagram displaying both study-specific and overlapping core operational taxonomic units (OTUs) in
polyethylene-associated microbial communities sampled in studies in the North Sea (De Tender et al. 2015)
and Baltic Sea (Kesy et al. 2019, Oberbeckmann et al. 2018, Ogonowski et al. 2018). Most OTUs occurred
in just one data set, but 45 were common to all of them (for taxonomy and mean relative abundances, see
Supplemental Table 3).
of the new material (i.e., the plastic), the large quantities of the newly introduced plastic particles
and their biofilms can influence the ecosystem processes in their surroundings, along with other
natural particles. Studies have already demonstrated the immense regulatory potential of biofilms,
particularly in streams, rivers, and intertidal systems (e.g., Battin et al. 2016, Decho 2000, Sabater
et al. 2002). On the other hand, the plastic leachate can significantly increase the dissolved organic
carbon in the oceans (Romera-Castillo et al. 2018), which, in turn, has the potential to increase the
marine microbial activity and its biomass in the oceans. Studies on the functional impact of mi-
croplastic biofilms on ecosystems are scarce and are mostly lacking natural controls to account for
general particle effects. Michels et al. (2018), however, showed that microplastic biofilms promote
the aggregation of plastic and biogenic particles.
One may also speculate that the introduction of microplastics as new surfaces into the ocean
will selectively enrich previously resting or less active members of the marine rare biosphere,
which are most likely subjected to ecological processes such as selection (Galand et al. 2009). If
so, microplastics would increase the number of metabolically active microbial species in the ocean
and potentially affect fluxes of dissolved organic matter.
4. MICROBIAL DEGRADATION OF PLASTICS

As discussed above, biogeography is overall the most important determinant for microbial as-
semblages on microplastics. However, microplastics themselves can also affect biofilm formation,
especially in nutrient-poor conditions (Oberbeckmann et al. 2018), and can selectively enrich
hydrocarbon-degrading bacteria, such as members of the family Sphingomonadaceae. It is un-
clear, however, whether this is due to microplastic surface properties or if microplastics could be

actively biodegraded. Because biodegradation is the only way of finally remediating the plastic
pollution in the oceans, we discuss the potential of microbial plastic degradation in more detail.
4.1. Definition of Biodegradation

The term biodegradation is one of the most misleading in the field of potential plastic degrada-
tion. Lucas et al. (2008) and Harrison et al. (2018) suggested three principal successive stages for
the process of biodegradation of synthetic polymers: biodeterioration, biofragmentation, and as-
similation. Andrady (2017), on the other hand, differentiates weathering from fragmentation and
defines, from an ecological point of view, degradation as the complete mineralization of synthetic
polymers. According to his definition, weathering describes a biotic or abiotic superficial degrada-
tion that modifies the mechanical, physical, and chemical properties of a polymer. Weathering can
be verified by determining the accumulation of oxidized moieties, especially carbonyl function-
alities, via Fourier transform infrared spectroscopy, changes in crystallinity, and the mechanical
properties of plastics (Andrady 2017). The eventual fragmentation into smaller components is
likely due to mechanical forces such as friction and shearing during wave movements or abrasion,
as well as surface ablation (concerning surface ablation, see figure 8 in Andrady 2017). These pro-
cesses enhance the polymer surface and reduce its molecular weight, which is a precondition for
further microbial cleavage and degradation on a molecular level. The total mineralization of poly-
mers into CO2 , H2 O, and salts, accompanied by the generation of new biomass, is the final part of
the assimilation step. We assume that, for biologists, the term biodegradation is often defined as
mineralization, or is at least differentiated from degradation and assimilation (Debroas et al. 2017).
But other perspectives exist, such as those of polymer scientists or the plastic industry. From their
perspective, plastic fragmentation during managed degradation processes, such as composting, is
sufficient to propagate biodegradability. In this case, biodegradation is being equated with the
eventual formation of microplastics from larger plastic items, without significantly reducing the
quantity of plastic in the environment. For society, these differentiations might not be transpar-
ent, and it might be assumed that a label of “biodegradable” or “biocompostable” is automatically
associated with remediation. Several reviews have covered these discrepancies (e.g., Gewert et al.
2015, Harrison et al. 2018, Lambert & Wagner 2017), and there is no need to go into further
detail here. It is important, however, to understand that the term biodegradation, when associated
with plastic pollution, as it is currently used in the literature and by the public, does not necessarily
include plastic remediation from polluted oceans. For this review, we define the biodegradation
of synthetic polymers as its assimilation and mineralization (and, therefore, its eventual removal)
in the natural environment.
4.2. Biodegradation of Microplastics in the Ocean

The weathering and consequent fragmentation of larger plastics can be attributed to UV radia-
tion, leaching of additives (and thus the loss of stabilizing properties), biofouling, and mechanical
processes ( Jahnke et al. 2017). Photodegradation probably plays the most important role (Andrady
2017), especially in combination with wave action in the beach zone. Other factors include the salts
in saline water (Da Costa et al. 2018) and potentially microorganisms themselves, as it has been
suggested that their metabolic products can indirectly influence, for instance, the discoloration
of plastics in the environment (Ghosh et al. 2013), probably due to a loss in surface properties
(Andrady 2015). As a result of the fragmentation of larger plastic items, secondary microplas-
tics have a significantly increased surface-to-volume ratio, offering larger surfaces for microbial
colonization and potential microbial attacks on the polymer. However, instead of degrading the

Important environmental Potential plastic degradation Number of marine research papers

factors for plastic degradation steps in the ocean (in situ or experiment) in 2018
Weathering 0 25
MACROSCALE
Photodegradation Leaching Weathering studies
Mechanical Fragmentation Fragmentation studies

fragmentation
MICROSCALE
Salt
NANOSCALE
Microbial
degradation
Assimilation
?
MOLECULAR
SCALE
Mineralization
Figure 5
Important environmental factors that may potentially catalyze different plastic degradation steps in the ocean. The areas below the
horizontal dashed line indicate steps that can be extrapolated from optimized laboratory experiments (left of the arrow) but still require
confirmation for the ocean (right of the arrow).
plastics, microorganisms can also increase their stability. Biofilms formed on microplastics can
protect the particles from photodegradation in the ocean surface (Weinstein et al. 2016) or in-
crease their relative density to above seawater density, which leads to microplastic sedimentation
and consequently protection from photodegradation ( Jahnke et al. 2017). In combination, these
changes in the plastic properties shift the factors that predominantly determine the potential plas-
tic degradation from physicochemical forces to microbial activity. As a consequence, the highest
remineralization rates of synthetic polymers are expected in the size range of smaller microplastics
(Figure 5).
Synthetic polymers are energy rich and theoretically represent a good source of energy and
carbon for microorganisms. For instance, the maximum usable energy for the complete oxidation
of polyethylene would be between −422 and −425 kJ per mole of O2 , similar to that of glucose
(−479 kJ per mole of O2 ), which is a well-known bacterial substrate. In terms of oxygen, extend-
ing the chain by more CH2 units hardly makes a difference. The complexity level of synthetic
polymer depolymerization is determined instead by the polymer’s hydrolyzability. Nonhydrolyz-
able plastics, such as polyethylene and polypropylene, consist of C-C-bond backbones where the
polymer must be cleaved into smaller molecules by redox reactions before its assimilation by cells
(Gewert et al. 2015, Krueger et al. 2015). By contrast, hydrolyzable plastics such as polyethylene
terephthalate and polyamide, which contain well-degradable structural elements like amide or

ester bonds, could be cleaved enzymatically or via hydrolysis, analogously to natural substrates
such as lignin or cellulose (Gewert et al. 2015, Krueger et al. 2015). However, the accessibility of
the bonds by the crystalline structure of the plastic surface can be as complex as, for example, that of
lignocellulose. Hydrolysis of lignocellulose, which is a natural substrate, depends on extracellular
lignin-modifying enzymes, including manganese peroxidase, versatile peroxidase, lignin peroxi-
dase, and multicopper oxidase laccase to initiate cometabolic biodegradation of lignin (Krueger
et al. 2015). Hydrolases (lipases and cutinases) have been described for the polymer polyethylene
terephthalate, a member of the polyester family that is formed by the monomers ethylene glycol
and terephthalic acid. The bacterium Ideonella sakaiensis, which has been isolated outside a plastic
bottle recycling facility, was able to cleave polyethylene terephthalate using two hydrolases and
thus biodegrade their monomers completely. The result was bacterial growth based exclusively on
polyethylene terephthalate energy sources (Yoshida et al. 2016).
Although polyethylene terephthalate biodegradation by Ideonella sakaiensis so far represents
the only example of complete plastic mineralization by bacteria, it was reached under optimal
laboratory conditions that do not represent natural environments. In the marine environment,
where conditions are more complex, the potential degradation of synthetic polymers follows the
biodegradation decalogue of Alexander (1975), a famous terrestrial soil microbiologist who stud-
ied the microbial decomposition of xenobiotic chemicals. Alexander’s biodegradation decalogue
specified under which conditions microorganisms should not metabolize a substrate. In the context
of microplastic pollution in the marine environment, three of Alexander’s (1975) commandments
describe especially well why highly dense and hydrophobic polymers should have low biological
degradability: A compound should not be degraded if the molecule is too large to penetrate the
cell (commandment 5), the compound concentration in aqueous solution is extremely low (com-
mandment 6), or the cleavage sites of the compound are hard to access (commandment 10). In
reference to these commandments, we can deduce that, in the marine environment, it is not the
energy content of synthetic polymers but factors such as the extremely low bioavailability and high
chemical stability that will determine the substrate quality (i.e., propensity to biodegradation) of
microplastics.
To explore the scientific literature from Alexander’s decalogue in 1975 to the most recent pa-
pers, we performed a literature search in Web of Science with a set of keywords related to both
(micro)plastics and degradation in marine systems (Table 1). The retrieved papers reflect the
limited findings regarding the biodegradability of synthetic polymers in the oceans. Of the 185
retrieved papers, 69 matched with the topic under analysis. Of these, 46 were related to the oc-
currence or quantification of fragmented plastics in the environment, as well as to (micro)plastic
degradation experiments under laboratory conditions (Table 1). The other 23 papers (one-third
of the 69 that matched the topic) were literature review papers. There seems to be a high ratio of
reviews to research papers when compared with other research topics, which indicates that (mi-
cro)plastics have been a hot topic of interest in the last few decades that has been explored by
scientists of different disciplines.
The search revealed that the weathering of plastics and microplastics, according to the defini-
tion by Andrady (2017), has been experimentally proven and observed in situ. Also, although the
fragmentation process is difficult to observe in the marine environment, and a global mass inven-
tory of ocean plastics still depends on educated guesses (Koelmans et al. 2017, Thompson et al.
2004), secondary microplastics, which are the direct result of plastic fragmentation, have been de-
tected in all marine habitats (see examples in Table 1) and represent the ubiquity of (micro)plastics
in the world oceans.
By contrast, the potential biodegradation of (micro)plastics catalyzed by marine microorgan-
isms in marine environments has only been assumed based on the weight loss of plastics during

Table 1 Research papers resulting from literature searches in Web of Science (topics),a which reflect the fact that the
final biodegradability of synthetic polymers in the ocean remains to be demonstrated
Study Typeb Degradation stepc Size classificationd Plastic typee
Ioakeimidis et al. 2016 ◦ W Macro PET
Lobelle & Cunliffe 2011 ♦ W Macro PE
Arias-Villamizar & ♦ W Meso HDPE
Vazquez-Morillas 2018
Artham et al. 2009 ♦ W∗ Meso LDPE, HDPE, PC, PP
Balasubramanian et al. 2010 ♦ W∗ Not described PE
Devi et al. 2015 ♦ W∗ Not described HDPE
Karlsson et al. 2018 ♦ W Not described PE
Khaled et al. 2018 ♦ W Not described PS
Muthukumar et al. 2011 ♦ W∗ Meso CFRP, GFRP, PET, PUR,
SR, SF
Welden & Cowie 2017 ♦ W∗ Meso PA, PE, PP
Sudhakar et al. 2008 ♦ W∗ Meso LDPE, HDPE
Nauendorf et al. 2016 ♦ W∗ Meso PE
Syranidou et al. 2017a ♦ W∗ Meso PS
Syranidou et al. 2017b ♦ W∗ Meso PE
Mohanrasu et al. 2018 ♦ W∗ Meso HDPE
Dussud et al. 2018 ♦ W Meso PE
Fotopoulou & ◦ W Micro PE, PP
Karapanagioti 2012
Paco et al. 2017 ♦ W∗ Micro PE
Auta et al. 2017b ♦ W∗ Micro PE, PET, PP, PS
Auta et al. 2018 ♦ W∗ Micro PP
Cai et al. 2018 ♦ W Micro PE, PP, PS
Da Costa et al. 2018 ♦ W Micro PE
Costa et al. 2011 ◦ F All sizes Marine debris
Alshawafi et al. 2017 ◦ F All sizes Marine debris
Cozar et al. 2014 ◦ F All sizes Marine debris
Cozar et al. 2017 ◦ F All sizes Marine debris
Eriksen et al. 2014 ◦ F All sizes Marine debris
Fok & Cheung 2015 ◦ F All sizes Marine debris
Thornton & Jackson 1998 ◦ F All sizes Marine debris
Tsiota et al. 2018 ♦ F Microplastics generation HDPE
Weinstein et al. 2016 ♦ F Microplastics generation HDPE, PP, PS
Hodgson et al. 2018 ♦ F Microplastics generation HDPE
Song et al. 2017 ♦ F Microplastics generation EPS, PE, PP
Fok et al. 2017 ◦ F Micro/meso Marine debris
Palombini et al. 2018 ◦ F Micro/meso Marine debris
Reisser et al. 2014 ◦ F Micro/meso Marine debris
Debroas et al. 2017 ◦ F Micro/meso Marine debris
Jang et al. 2018 ◦ ♦ F Micro PS
Jiang et al. 2018 ◦ F Micro PE, PP, PS
Jungnickel et al. 2016 ♦ F Micro PE
(Continued)

Table 1 (Continued)
Study Typeb Degradation stepc Size classificationd Plastic typee
Lenz et al. 2015 ◦ F Micro Marine debris
Naidu et al. 2018 ◦ F Micro Marine debris
Sagawa et al. 2018 ◦ F Micro Marine debris
Frias et al. 2016 ◦ F Micro Marine debris
Acosta-Coley & ◦ F Micro Microplastic resin pellets
Olivero-Verbel 2015
Chubarenko et al. 2018 ◦ F Micro Marine debris
a
The keywords for the searches were entered as follows (with individual searches separated by semicolons): microplastic & microbial biodegradation;
microplastic & fragmentation; plastic & microbial degradation & marine; synthetic polymer & microbial degradation & marine; plastic & fragmentation
& marine; synthetic polymer & fragmentation & marine; plastic & biodegradation & marine; synthetic polymer & biodegradation & marine; microplastic
& biodegradation; microplastic & biodegradation.
b
◦, environmental study; ♦, experimental study.
c
W, weathering; F, fragmentation; ∗, weight loss was determined, but the degradation of leachates or mineralization of plastics was unproven.
d
Macro, >20 cm; meso, 0.5–20 cm; micro, <0.5 cm.
e
CFRP, carbon-fiber-reinforced plastic; EPS, extracellular polymeric substance; GFRP, glass-fiber-reinforced polymer; HDPE, high-density
polyethylene; LDPE, low-density polyethylene; PA, polyamide; PC, polycarbonate; PE, polyethylene; PET, polyethylene terephthalate; PP, polypropylene;
PS, polystyrene; PUR, polyurethane; SF, syntactic foams; SR, silicone rubber.
in situ or laboratory experiments, and there have been fewer papers on this topic than on plastic
weathering or fragmentation processes. For instance, a 12-month experiment in the Bay of Bengal,
India, revealed a weight loss of 0.65–1.9% for low- and high-density polyethylene, polycarbonate,
and polypropylene (Artham et al. 2009). Relatively higher weight losses have been determined
in vitro when using microbial strains isolated from plastics in the marine environment. Bacte-
ria from genera such as Bacillus, Rhodococcus (Auta et al. 2017a,b), Arthrobacter, and Pseudomonas
(Balasubramanian et al. 2010) were related to weight losses of up to 7.4% for polyethylene mi-
croplastic particles (Auta et al. 2017b) and 15% for relatively larger high-density polyethylene
films (Balasubramanian et al. 2010). Analogous degradation studies with fungi reported weight
losses of up to 8.5% for high-density polyethylene degraded by Aspergillus sp. after 30 days of in-
cubation (Devi et al. 2015) and more than 43% for polyethylene degraded by Zalerion maritimum
after 14 days of incubation (Paco et al. 2017).
Thus, plastic weight loss can reach promisingly high values in comparatively short time frames,
but the fact remains that this approach does not cover all aspects of the biodegradation process.
Weight loss alone, even in combination with microbial growth, cannot discriminate between the
degradation of polymers and the degradation of additives or monomers. Due to incomplete poly-
merization processes, the latter may also constitute a substantial proportion of the final plastic
product and are often easily microbially mineralized, thus potentially leading to falsification of
the polymer degradation rates even in additive-free polymers (Klaeger et al. 2019). A simple fur-
ther fragmentation of microplastics into nanoplastics (particles less than 100 nm in size) has also
been observed in the laboratory—for instance, through digestive fragmentation experiments in
Antarctic krill (Dawson et al. 2018, Jahnke et al. 2017)—but so far remains unproven for the marine
environment. Thus, in order to evaluate the retention times of microplastics in the ocean, further
research on their biodegradation should focus on the final steps: assimilation and mineralization.
However, because of the biodegradation decalogue (Alexander 1975), studying these steps remains
challenging for microplastics. Highly sensitive approaches, such as mineralization experiments of
14
C-labeled synthetic polymers in marine in situ conditions [comparable to the 14 C-polystyrene
polymer in vitro cultivation experiment using the fungus Penicillium variabile (Tian et al. 2017)],
are essential to recognize the lifetime of microplastics in the ocean.

To summarize, biodegradation of microplastics has not yet been detected in the marine en-
vironment. Due to the low bioavailability of plastics, their degradation is determined mainly by
physicochemical forces, and the degradation ends (based on our current knowledge) with the en-
richment of microplastics and nanoplastics in the marine system. Consequently, as has already been
assumed, microplastics likely remain unmineralized in the oceans for hundreds of years (Barnes
et al. 2009) or even longer (Andrady 2015). The question remains of whether marine microorgan-
isms may adapt evolutionarily to plastic degradation in the future.
5. MICROBIAL ADAPTATION TO MICROPLASTIC DEGRADATION

Whether marine microorganisms will adapt, evolve, and assist in cleaning up the ocean through
plastic degradation is environmentally relevant but a complex scientific question. The low bioavail-
ability of plastics is apparently unfavorable for the evolution of productive and significant degra-
dation pathways (Krueger et al. 2015). Furthermore, it should be considered that the large variety
of plastic polymers would necessitate an equivalently large range of degradation pathways. How-
ever, in reference to Darwin’s theory on the evolution of species (Darwin 1868), we asked instead
whether a potential microbial adaptation to plastic degradation will increase microbial fitness,
significantly enabling the growth and reproduction of cells in the ocean.
From the literature, we know that a natural dissolved organic carbon concentration below
30.7 μmol L−1 would be insufficient to support prokaryotic metabolism (Arrieta et al. 2015), and
this value is at the lower limit of the estimated concentration of dissolved organic carbon in the
open ocean, approximately 34–80 μmol kg−1 (Hansell et al. 2009). Degradation pathways should,
at least in theory, evolve and result in the hydrolysis of synthetic polymers into monomers or
chemical fragments, with these biodegradable substances then released at concentrations above
approximately 30 μmol C L−1 in the ocean. The polyethylene terephthalate degrader Ideonella
sakaiensis isolated from outside a recycling facility (Yoshida et al. 2016) is an example of this pro-
cess and indicates that relatively higher concentrations of plastics can lead to the evolution of
plastic degradation in bacteria. A metagenomic study covering putative biodegradation pathways
generated from marine microorganisms living on plastics indeed demonstrated that putative xeno-
biotic biodegradation exists in the ocean (Bryant et al. 2016). However, and not surprisingly, the
global distribution of polyethylene-terephthalate-degrading bacteria and their respective genes is
not significant (Danso et al. 2018).
We therefore deduce that it is unlikely that bioavailable synthetic polymers will reach suffi-
ciently higher levels in the oceans to promote the evolution of bacteria that can effectively de-
grade microplastics. As research on marine microplastics is still growing, potential hot spots of
marine plastics may remain unidentified, which would eventually change the scenario analyzed
here. Most important, however, is to prevent these (micro)plastics thresholds from being reached
in the world’s oceans.
6. CONCLUSION AND RECOMMENDATIONS

For some decades now, the burden of microplastics has been increasing. Microplastics are an es-
tablished potential microbial substrate and colonization surface in the oceans, with members of
the family Sphingomonadaceae in particular selectively colonizing microplastic polymers. Po-
tentially more relevant for society, the microplastic-microbial biofilm has been associated with
(a) concerns related to the role of microplastics as a long-living vector for microorganisms, espe-
cially pathogenic species, and (b) the potential for microorganisms to degrade (micro)plastics in
the long term and thus contribute to cleaning plastics from the oceans. Based on our reanalysis

and on the critical review of the available literature, we have concluded that, at present, (a) there
has been no increase in the accumulation of pathogens colonizing microplastics, and (b) marine
microorganisms play a negligible role in the biodegradation of microplastics. Because of the low
bioavailability of microplastics in the oceans, microorganisms will not be able to adapt to signifi-
cantly degrade plastics, at least on a human timescale, although this may depend on the concen-
tration of plastics and the generation of plastics hot spots in the future.
The extremely diverse properties and benefits of plastics cannot be ignored in our everyday
lives. However, due to improper treatment, microplastics are ubiquitous in the marine environ-
ment, and the available studies indicate that this contaminant is triggering an ecological distur-
bance. However, the extent of the potential impacts associated with marine microplastics on both
spatial and temporal scales has not yet been determined. Tagg & Labrenz (2018) have discussed
the need for proactive regulations for microplastics. To ensure environmental compatibility and
sustainability, multiple actions are needed at the same time: (a) further research on microbial path-
ways potentially linked to plastic degradation, in order to improve knowledge of how to develop
in situ biodegradable materials (as proposed in Quero & Luna 2017); (b) a significant reduction
of all packaging materials or typical products of our twenty-first century society (from plastic
foils to plastic toys); and (c) the development of an efficient recycling system that can be applied
easily and cost neutrally worldwide. Such a recycling system should involve plastic-producing,
plastic-recycling, and plastic-using industries alike, as the current recycling regime lags behind its
potential. We further predict that research on the microbial remediation of plastics will increase
in the future and lead to optimized biotechnological concepts that can be applied in vitro.
DISCLOSURE STATEMENT
The authors are not aware of any affiliations, memberships, funding, or financial holdings that
might be perceived as affecting the objectivity of this review.
ACKNOWLEDGMENTS
We greatly appreciate critical reviews of earlier versions of this article by Juliana Ivar do Sul,
Brittan S. Scales, and Alexander S. Tagg from the Leibniz Institute for Baltic Sea Research
Warnemünde. We are grateful to Friedrich Widdel from the Max Planck Institute for Marine
Microbiology for sharing his expertise on energy calculations. This work resulted from the
MikrOMIK project funded by the Leibniz Association (SAW-2014-IOW-2); the BONUS MI-
CROPOLL project supported by BONUS (Article 185), funded jointly by the European Union
and the German Federal Ministry of Education and Research (BMBF) (03F0775A); and the
BMBF project MicroCatch_Balt (03F0788A).
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MA12CH08_Labrenz ARjats.

cls November 16, 2019 9:59

Annual Review of Marine Science

Annu. Rev. Mar. Sci. 2020. 12:209–32 Keywords

The Annual Review of Marine Science is online at Abstract

1. THE GENESIS OF A NONDEGRADABLE MULTIPURPOSE PRODUCT

210 Oberbeckmann • Labrenz

2. EXISTING KNOWLEDGE AND RESEARCH GAPS FOR MARINE

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3. AN EVALUATION OF THE PLASTISPHERE

212 Oberbeckmann • Labrenz

3.3. Microbial Life on Microplastics and Natural Particles

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6.5 (Oberbeckmann et al. 2018)

4.4–8.7 (Kesy et al. 2019)

214 Oberbeckmann • Labrenz

www.annualreviews.org • Marine Microbial Assemblages on Microplastics 215

Square root transformed

216 Oberbeckmann • Labrenz

ability to degrade aromatic or halogenated hydrocarbons, including petroleum and pesticides,

Relative abundances (%, log scale)

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3.4. The Core Bacterial Community Associated with Polyethylene Microplastics

3.5. Microplastics Potentially Impact Ecological Processes in the Oceans

218 Oberbeckmann • Labrenz

Oberbeckmann et al. 2018 Ogonowski et al. 2018

4. MICROBIAL DEGRADATION OF PLASTICS

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4.1. Definition of Biodegradation

4.2. Biodegradation of Microplastics in the Ocean

220 Oberbeckmann • Labrenz

Important environmental Potential plastic degradation Number of marine research papers

Photodegradation Leaching Weathering studies

Mechanical Fragmentation Fragmentation studies

www.annualreviews.org • Marine Microbial Assemblages on Microplastics 221

222 Oberbeckmann • Labrenz

www.annualreviews.org • Marine Microbial Assemblages on Microplastics 223

224 Oberbeckmann • Labrenz

5. MICROBIAL ADAPTATION TO MICROPLASTIC DEGRADATION

6. CONCLUSION AND RECOMMENDATIONS

www.annualreviews.org • Marine Microbial Assemblages on Microplastics 225

226 Oberbeckmann • Labrenz

www.annualreviews.org • Marine Microbial Assemblages on Microplastics 227

228 Oberbeckmann • Labrenz

www.annualreviews.org • Marine Microbial Assemblages on Microplastics 229

230 Oberbeckmann • Labrenz

www.annualreviews.org • Marine Microbial Assemblages on Microplastics 231

232 Oberbeckmann • Labrenz

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