State of the Art

Plastic debris in ocean surface waters

Since 1950, 8300 Mio t of plastics have been produced and 4900 Mio t discarded (Geyer et al. 2017). While the global amounts of produced and discarded plastic are constrained relatively well, reliable estimates on the amount ending up in the environment are lacking. Currently,
5-13 Mio t of plastic debris are believed to enter the oceans every year, and cumulative plastic input is estimated to increase by more than double in the next decade (Jambeck et al., 2015). Approximately 50 - 60% of the plastic that is produced is less dense than seawater (Andrady 2011), and about 60% of the positively buoyant plastic debris that enters the oceans from land is likely transported offshore by JPI Oceans HOTMIC - ID: 87 12 of 44 surface currents and winds (Maximenko et al., 2012; Lebreton et al., 2012). Floating plastic debris therefore tends to accumulate within oceanic gyres and restricted coastal waters (Lebreton et al., 2012). During the time the debris remains at sea, large plastics items gradually fragment into smaller pieces – so called microplastics (MP; plastic particles < 5 mm) – under the combined effects of temperature, UV radiation, and actions by waves and organisms (Andrady, 2011).

Little is known about the mechanisms that transport plastic from the coast to accumulation areas in the centre of subtropical gyres. Coastal waters have a tendency to spread along the coast, making offshore transport unlikely to occur. Mesoscale eddies may provide an effective mechanism for offshore transport by trapping coastal water inside their core and carrying it into the interior ocean (Thomsen et al., 2016). The trapped coastal water masses and MP loads are released only when the eddies decay, which may take up to 5 years after generation and occurs thousands of km west of their generation site (Chelton et al., 2011). Indeed, increased concentrations of MP have been found in anti-cyclonic mesoscale eddies in the North Atlantic subtropical gyre (Brach et al., 2018).

Estimates of plastic input to the ocean indicate at least 1.71 Mio t should reach global ocean gyres annualy (Eriksen et al., 2014). Despite these theoretical reflections, it is still not known how much plastic debris actually floats at the ocean surface, or what mechanisms control plastic transport and fate from land to the open ocean. Empirical data from field observations suggest 0.27 Mio t (Eriksen et al., 2014), but more recent  measurements in the North Pacific suggest that global inventories are underestimated (Lebreton et al., 2018). However, Lebreton et al. (2018) also concluded that there remains a major discrepancy between the annual input of buoyant plastic (i.e. millions of tons) and the amount observed at the ocean surface (i.e. hundred thousands of tons). Furthermore, time series revealed no significant increase in surface ocean plastic debris since the 1980s (Law et al., 2010; Law et al., 2014).

This prompts a fundamental question: what has become of the missing plastic? Several, mutually non-exclusive mechanisms have been proposed to explain the apparent loss of plastic debris from surface waters. First, a part of the material could constantly be transferred into the ocean interior by biota-plastic interactions, namely ingestion by invertebrates and vertebrates with subsequent formation of sinking fecal pellets, and colonization by biofilms which decreases buoyancy. Another part may be fragmented into pieces too small to be captured by current monitoring schemes (Poulain et al., 2019), and/or it could be deposited along the shorelines (Andrady, 2011). Furthermore, MP may coagulate with suspended biogenic particles such as phytoplankton cells (Michels et al., 2018), increasing particle density of otherwise buoyant MP and accelerating removal by sinking (Galloway et al., 2017). Recent findings also indicate that gelatinous zooplankton could play a significant role in capturing and sinking marine particles including MP through the water
column (Katija et al., 2017; Macali et al., 2018), since the mucus that they release during reproduction, defense, or as stress response is highly efficient at trapping micro- and nano-sized particles (Patwa et al., 2015). While all of these processes are possible explanations,
quantitative data on these mechanisms and their relative contribution to plastic debris loss from ocean surface waters do not exist.

The North Atlantic is estimated to contain approximately 20% of the global amount of floating plastic debris (Cózar et al., 2014; Eriksen et al., 2014). Most of this material is concentrated in the inner accumulation zone of the North Atlantic subtropical gyre, which spans from the
Azores to Bermuda (Fig. 1). Plastic inventories averaged 400 g km-2 within the inner accumulation zone, with maxima as high as 2500 g km-2 (Cózar et al., 2014). For the same sea area Eriksen et al. (2014), reported plastic particle densities of up to 106 km-2. Plastic particles within the North Atlantic gyre have a wide range of compositions and sizes, including nanoplastics (ter Halle et al., 2017), and small size particles may comprise the major mass fraction in the region (Poulain et al., 2019).

Plastic debris in the water column

Most of the available data on the abundance and composition of floating plastic debris have been collected with sea surface trawls developed for the sampling of neuston (e.g. Manta trawl). These devices usually have a small aperture size (0.5 -1 m width, 0.15 – 1 m
depth) and can therefore cover only small areas and are restricted to the uppermost water column. Very few studies have quantified plastic debris at different water depths, which is a major knowledge gap in efforts to monitor and map plastic distributions in the ocean. A few studies indicated that MP concentrations are highest at the sea surface and in deep water (Lattin et al., 2004; Bagaev et al., 2017; Reisser et al., 2015). However, data about the abundance of plastic debris in the open ocean water column between the surface and the seafloor, in particular from beneath known accumulation zones, are missing.

Plastic debris in deep-sea sediments

A few studies have investigated the abundance of plastic debris and MP in coastal sediments (e.g., Van Cauwenberghe et al., 2015a), but data from the deep sea (water depths greater than 1000 m) are even scarcer. The lack of information on MP in deep-sea sediments
prevents assessment of MP effects on sensitive deep-sea ecosystems, and evaluation of this potential sink for land-sourced plastic debris.
Microplastics have been detected in sediment cores from the Arctic (Bergmann et al., 2017), the North Atlantic, South Atlantic, and Mediterranean Sea (Van Cauwenberghe et al., 2013; Woodall et al., 2014), and North-West Pacific (Fischer et al., 2015). Furthermore, there is increasing evidence of potential transport of MP to deeper layers in soft marine sediments by means of bioturbation (Nakki et al 2017), although field estimates are lacking. Collectively, this suggests that the deep sea is a sink for plastic debris and MP, although the magnitude of this sink and the processes that control it are very poorly constrained (van Sebille et al., 2015).

Weathering and alteration of MP in the ocean

Plastic debris in the ocean occurs in a wide range of size classes, which constitute different proportions of either the total mass of plastic or the total number of particles. Although larger particles constitute most of the mass of the debris, small MP account for the greatest number
of particles (Eriksen et al. 2014). The vast majority of marine plastic is between 0.3 and 5 mm, with a peak around 2 mm (Cózar et al., 2014). Surprisingly, below 1 mm a gap in the size distribution of particles has been observed and according to models that predict the size
class abundance of MP, sub-mm particles are 100-fold lower than expected. This suggests that particles of this size are selectively removed from the ocean surface by biological processes such as aggregation and settling, or by fragmentation into microand nano-sized particles at rates which are much higher than those for other size classes (Cózar et al., 2014). Small MP particles (i.e. < 0.3 mm) evade current methods used in MP surveys, despite their potential environmental importance.

Degradation of MP in seawater is evident in both chemical and physical characteristics. Plastic polymer degradation can occur via a variety of mechanisms, but photodegradation (UV), oxidation, leaching of polymers, and physical degradation are likely the most important (Gewert et al., 2015).Physical alteration also occurs during MP weathering, including cracking, pitting, erosion, and color change (Fotopoulou and Karapanagioti, 2015; ter Halle et al., 2017; Tang et al., 2018; Resmerita et al., 2018; Paluselli et al., 2019). Ingestion of plastics by biota drives further fragmentation and can produce small-size MP (Hodgson et al., 2018; Dawson et al., 2018).

Weathering leads to changes in chemical bond structure measured by FTIR and Raman spectroscopy (Fotopoulou and Karapanagioti, 2015; Brandon et al., 2016), increased crystallinity and shortening of polymer chains (ter Halle et al., 2017), and oxidation of MP surfaces (ter Halle et al., 2017; Tang et al 2018). Microbial colonization of plastic surfaces can also affect weathering, including accelerated leaching of plasticizers (Paluselli et al., 2019). However, signs of microbial degradation of macroplastic debris could not be identified in other field experiments (Nauendorf et al., 2016) or plastics recovered from the 4100-m deep Peru Basin after two decades of deposition (Krause et al., submitted). Chemical changes during weathering suggest that chemical fingerprinting methods could provide a sensitive method for evaluating the weathering state and age of MP in the ocean, including sub-micron-sized particles.

Biofouling and Bioshredding

Virtually all solid surfaces in the sea are colonized by sessile organisms in a process known as biofouling. Physico-chemical weathering can produce micro-sized, non-buoyant fragments, but biofouling is suspected to be a major agent in vertical transport processes that shift plastic from the surface to deeper parts of the oceans (Reisser et al., 2014). Biofouling changes the surface properties and facilitates the formation of aggregates consisting of inorganic and organic matter and (micro)biological constituents such as phytoplankton cells (Zhao et al., 2017). These aggregates have a higher density than dispersed particles and therefore sink rapidly through the water column. The formation of aggregates that include MP may constitute an important mechanism for transporting small-sized plastic debris from the ocean surface to deeper waters and ultimately to the seafloor (Michels et al., 2018).

The term ‘Plastisphere’ has been suggested by Zettler et al. (2013) as a term for the specific bacterial assemblages found on plastic particles. Microbial assemblages on plastic debris have also been recognized as being distinct from those on other marine substrata in the vicinity (Oberbeckmann et al. 2015; Nauendorf et al., 2016; Krause et al., submitted). Although microorganism association with plastic particles has long been recognized (Carpenter and Smith, 1972), few studies have investigated microbial communities associated with marine plastics at or near the ocean surface (Bryant et al., 2016; Oberbeckmann et al., 2014, 2016). Furthermore, only one study exists about bacterial communities colonizing plastics at the deep seafloor (Krause et al., submitted). Assessing microbial communities on plastic debris is of particular interest, since some bacterial enzymes are capable of degrading polyethylene (PE) and polyethylene terephthalate (PET) (Danso et al., 2018; Wei and Zimmermann, 2017; Yang et al., 2014).

Colonization of macroplastic surfaces can shield plastics from degradation by UV light (O'Brine and Thompson 2010), but it may also increase the palatability of the macroplastic to organisms and increase rates of ingestion. Hodgson et al. (2018) showed that biofouling of
plastic bags led to increased shredding by amphipods, and that this fragmentation depended more on the degree of fouling than the plastic composition. Digestive fragmentation of ingested MP into sub-micron particles has also been observed in Antarctic krill (Dawson et al.,

Uptake of MP by biota

Most plastic litter items within oceanic and coastal waters are MP (Eriksen et al., 2014), which tend to be more abundant than larger particles and which have, due to their small size, a higher potential for accumulation in biotic tissues (Browne et al., 2008). For the latter, MP
must be taken up into the intestinal tract and this has so far been documented for 227 different marine organisms (Lusher et al., 2017). This includes pelagic animals ranging from copepods (Cole et al. 2013), over salps (Moore et al., 2001) and tuna (Romeo et al., 2015), to whales (Besseling et al., 2015). Assuming that the ingestion of MP can have negative consequences for the health of the affected organism, this suggests that MP can be a threat to the integrity of the entire marine ecosystems.

Plastic-biota interactions in the open ocean comprise the ingestion of particles by pelagic or benthic organisms, as well as colonization by biofoulers (bacteria to metazoans), and entrapment in mucus released by gelatinous plankton.Gelatinous zooplankton can mediate the transport of plastic particles, for instance, by formation of a “mucus house” that grabs bits of food in the water column (e.g. Katija et al. 2017). Plastic particles captured during feeding are removed from the water column, and mucous aggregates may promote ingestion by other organisms (Davies and Hawkins, 1998). Some medusa also ingest plastic particles into their gastrovascular system, becoming a vector for the sinking of MP and transfer to upper trophic levels (Macali et al., 2018). Infaunal organisms can also promote MP burial via ingestion and bioturbation (Nakki et al., 2017). In addition to active MP uptake, biological receptors can be externally exposed to MP if the biological surfaces are sticky and sufficiently sensitive. Fish eggs and larvae may be susceptible to adhesion by MP with potential negative impacts. All of these processes presumably play a role in the transport of MP and eventual impact, but very few empirical data exist that document their action.

Additives and absorbed organic pollutants

From a chemical perspective, plastic debris constitutes mixtures of polymer, residual monomers, additives and absorbed chemical contaminants from the environment (Koelmans et al., 2016; Galloway et al., 2017; Gallo et al, 2018). Adsorbed environmental contaminants
include toxic, low molecular weight species that can be transferred to animal tissues if the contaminated plastic debris is ingested. Biologically-active adsorbed compounds such as endocrine disruptors have the potential to influence metabolic pathways as well as
reproductive success (Koelmans et al., 2016).

Plastic additives can leach into seawater (Teuten et al., 2009), both altering the chemical composition of the residual MP and leaving a chemical trace in the dissolved phase. Leaching effects by UV irradiation vary with different polymers, but microbial colonization of plastic surfaces leads to polymer-independent accelerated leaching of plasticizers (Paluselli et al., 2019). Leaching of plastic additives can lead to both direct and indirect exposure to marine biota (Teuten et al., 2009; Koelmans et al., 2014). The presence of these leachates has been shown in the southwest Baltic Sea (Erhardt and Derenbach, 1980) and the northwest Mediterranean Sea (Paluselli et al., 2018). It remains unknown how chemical signatures of MP leachates in the dissolved phase correspond to presence of MP in the ocean.

Sampling and analysis of MP in the ocean

Accurate chemical identification of MP is challenging due to the broad range of  polymer/copolymer structures that contribute to MP, which also contains residual organic additives and inorganic fillers. It is further complicated by the fact that biodegradation, photodegradation, biological uptake, transport processes and adsorption of organic contaminants alter the MP composition. This all needs to be accounted for when tracking and analyzing MP in environmental samples (Fahrenfeld et al., 2019, Hidalgo-Ruz et al., 2012; Silva et al., 2018). Analysis of microplastics incl. sampling, sample preparation, identification, quantification and characterization of microplastic particles in different size ranges is a challenging task, particularly if capture, detection and characterization is aimed also at size classes in the lower micrometer and sub-micrometer range (100 nm – 1 µm).

The most advanced pyrolysis-based analytical techniques such as Evolved Gas Analysis Mass Spectrometry (EGA-MS) and both flash and multi-shot Pyrolysis coupled with Gas Chromatography and Mass Spectrometry (Py-GC/MS) can provide in-depth structural information complementary to the FTIR and Raman spectroscopies typically used for MP identification, thus representing useful confirmatory techniques for validation and calibration of spectroscopic analyses. Thermal desorption has been demonstrated extremely promising in MP
characterization (Dümichen et al., 2017), and EGA-MS can also provide information on the additives and other hydrophobic organic chemicals and persistent organic pollutants (POPs) as well as on the polymer constituting the MP, in particular when combined in one
analytical run with Py-GC/MS for multi-shot Py-GC/MS.

Dissolved MP leachates in the water column are sampled using traditional oceanographic methods, although care is required to use nonplastic materials, such as metal or glass, whenever possible. Leachates are preconcentrated from seawater using solid-phase extraction
(Loghmani, 2018; Fauvelle et al., 2018), and the seawater volume can be adjusted to achieve necessary detection limits. The extracted compounds are eluted with an appropriate solvent, and measured by GC-MS or soft-ionizing uHPLC-ESI-MS (Loghmani, 2018; Fauvelle et al.,