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<figure><img src="image_41.png" /></figure>
Chapter 8

Dietary Exposure to Additives and Sorbed
Contaminants from Ingested Microplastic
Particles Through the Consumption
of Fisheries and Aquaculture Products

<figure><img src="image_42.png" /></figure>

Check for
updates

Esther Garrido Gamarro and Violetta Costanzo

**Abstract** Microplastics and nanoplastics may be found in the gastrointestinal tract of some aquatic animals and could potentially be ingested by humans if consumed whole. Information on the toxicity of plastic particles, as well as co-contaminants such as plastic additives, remains scarce. This represents a serious challenge to perform realistic risk assessments. An exposure assessment of selected plastic additives and co-contaminants of known toxicity associated with microplastics was carried out for shellfish in this study, which builds on an exposure assessment of microplastic additives and a limited number of associated contaminants in mussels conducted by the FAO in 2017. This study evaluates possible impacts to food safety by examining a diverse additives and associated sorbed contaminants. The results suggest that the levels of certain microplastic additives and sorbed co-contaminants in target animals (shrimp, prawns, clams, oysters, and mussels) do not pose a food safety threat to consumers. To get to further conclusions, an exposure assessment from the whole diet should be carried out and the toxicity of some of the most common polymers and plastic additives, as well as their mixtures, needs to be carefully evaluated.

**8.1 Introduction**

Plastic production has been increasing exponentially since the 1950s and was esti-
mated to be 8300 million metric tons to date (Geyer et al. 2017). Since its first
development in the 1800s, the production of plastic materials has changed to meet
the needs of a variety of sectors and consumers and has enabled technological
improvements and solutions. Due to their functional properties ("cheap and

E. Garrido Gamarro (✉) · V. Costanzo

Fisheries and Aquaculture Division, Food and Agriculture Organization of the United Nations

(FAO), Rome, Italy

e-mail: esther.garridogamarro@fao.org
durable”), plastics have displaced many non-plastic materials, becoming the most utilized materials worldwide.

Plastics consist of a range of synthetic or semi-synthetic chemicals that are made of fossil resources and organic by-products. They are commonly divided into three categories: thermoplastics (polymers that can be re-melted), elastomers (elastic polymers that return to their original shape after being deformed), and thermosets (polymers that remain in a permanent solid state once hardened).

Depending on their specific use, polymers with different physical and chemical properties can be mixed and additives such as plasticizers, colourants, UV-stabilizers, flame retardants, and antioxidants can be added to improve the performance of the final product.

Recycling the complex mixtures of chemicals used for plastic production can be challenging, as well as the evaluation of their impact on the environment and human health. The extensive production of plastic requires efficient waste management systems, but most countries do not have the capacity to develop them.

Microplastic particles have been found in a variety of human food items, such as salt, beer, honey, and aquatic products, with seafood being the best-studied source of dietary intake of microplastics. Exposure to microplastic particles, their additives, and their sorbed co-contaminants depends on several factors, such as particle size, shape, chemical changes that occurred during processing and/or cooking of fisheries and aquaculture products, and consumption patterns.

A previous study on exposure to microplastics and associated co-contaminants suggests that exposure to this contaminant burden is typically less than 0.1% (FAO 2017). Microplastic contribution to the total dietary intake of additives and sorbed co-contaminants was estimated as very low, with maximum increases in BPA, PAHs, and PCBs load of less than 2%, 0.004%, and 0.006% respectively, after the ingestion of a portion of mussels (EFSA 2016). However, a recent study by Barboza et al. (2020) observed a clear correlation between microplastics intake in three species of wild-caught commercial fish and the levels of bisphenols in the muscle and liver. Higher microplastic concentrations in fish were correlated with higher levels of these compounds, whose concentration in the edible tissue exceeded the established limits for human safety set by the EFSA. Furthermore, a relation between the concentration of plastic-associated chemicals and ingested microplastics in marine organisms has already been hypothesized (Granby et al. 2018; Rochman et al. 2013; Teuten et al. 2009). These findings suggest that more investigations should be conducted on this subject to better identify the role of microplastics in the transfer of pollutants and which factors could influence the process. This chapter aims to provide an overview of the dietary exposure of microplastic particles, additives, and common microplastic co-contaminants through aquatic products using consumption data from the FAO/WHO database, while information on contamination levels of plastic pellets and microplastic ingestion by seafood are updated with the current literature. Moreover, four different groups of seafood were considered, to extend the exposure evaluation also to crustaceans and other bivalves. The final estimations are compared with the no observed effect levels (NOELs) and no observed adverse
effect levels (NOAELs) and can be useful to provide a better understanding of the potential impacts on food safety.

## 8.2 Sorption of Environmental Contaminants by Microplastics

A potential threat to human health deriving from the exposure to microplastics is that these materials can scavenge and thus concentrate pollutants already present in production waters. The ingestion of contaminated plastic could lead to higher exposure to toxic chemicals, with possible endocrine disruption and carcinogenicity. The main process leading to the interaction between microplastics and hydrophobic organic chemicals (HOCs) in the water column does not involve the formation of covalent bonds; thus, its reversible nature preserves the likelihood of chemical desorption from the matrix (Endo and Koelmans 2016). Besides, plastic polymers are also recognized as possible vectors of heavy metals in the marine environment (Holmes et al. 2012, 2014), being experimentally able to accumulate concentrations even 800-fold higher than in seawater (Brennecke et al. 2016). Many field studies such as the International Pellet Watch have reported the concentration levels of persistent organic pollutants (POPs) and metals sorbed on beached and marine pellets, in addition to plastic additives (Tables 8.1 and 8.2).

Sorption processes can be classified into adsorption and absorption, depending on the mechanism of interaction between the polymer and the chemicals. Absorption mainly occurs when the molecules of pollutants diffuse into the bulk matrix of the polymer and interact with it through weak van der Waals forces or hydrophobic interactions. This process is mainly driven by the preferential partitioning of the chemical on plastic compared to water, which is usually linearly and positively correlated to its octanol-water partition coefficient ($K_{ow}$), a parameter that measures the level of hydrophobicity (Lee et al. 2014; Li et al. 2018c; O'Connor et al. 2016). Adsorption refers to a process that results in the sorption of molecules that are confined to the surface of microplastics (Endo and Koelmans 2016).

Absorption mainly occurs onto rubbery polymers (i.e. PE and PP), where external molecules pass through and associate within their matrix (Hüffer and Hofmann 2016; Müller et al. 2018; Teuten et al. 2009; Wang et al. 2018a). These polymers are generally recognized as the ones concentrating the highest amounts of HOCs and are then possibly more dangerous for marine life (Endo et al. 2005; Fisner et al. 2017; Hirai et al. 2011; O'Connor et al. 2016; Wang and Wang 2018; Wang et al. 2018b) and possibly human health.

Some polymers present several functional groups on their surface, conferring a certain degree of polarity. These are mostly PS and PVC, whose glassy nature also yields the formation of nanovoids and pores on the surface, which are the sites of sorption. In this case, the process is mainly led by adsorption, a mechanism through which the chemicals more efficiently bind to the plastic polymer through ionic,
<table><caption>**Table 8.1** Maximum concentrations of persistent organic pollutants (POPs) and metals sorbed on plastic pellets from field studies</caption><thead><tr><th>Location</th><th>Sampling</th><th>Extraction</th><th>PCB (ng/g)</th><th>PAH (ng/g)</th><th>DDT (ng/g)</th><th>DDE (ng/g)</th><th>PBDE (ng/g)</th><th>HCH (ng/g)</th><th>Cd (ng/g)</th><th>Pb (µg/g)</th><th>References</th></tr></thead><tbody><tr><td rowspan="2">Portuguese Coast</td><td rowspan="2">Metal mesh sieving</td><td>n-Hexane, hexane-acetone</td><td>223</td><td>44,800</td><td>41</td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Greek beaches</td><td>N.S.</td><td>Hexane</td><td>290</td><td>500</td><td>42</td><td>15</td><td></td><td>3.5</td><td></td><td></td><td></td></tr><tr><td>South Devon, England</td><td>Plastic tweezers</td><td>Dilute aqua regia digestion</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>10</td><td>1.08</td><td>Karapanagioti et al. (2011)</td></tr><tr><td>South West England</td><td>Plastic tweezers</td><td>Dilute aqua regia digestion</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>76.7</td><td>1.64</td><td>Holmes et al. (2012)</td></tr><tr><td>Portuguese coast</td><td>N.S.</td><td>Hexane</td><td>310</td><td>24,000</td><td>49</td><td>n.a.</td><td>n.a.</td><td>3.3</td><td></td><td>n.a.</td><td>Mizukawa et al. (2013)</td></tr><tr><td rowspan="2">Japanese coast</td><td rowspan="2">Solvent rinsed stainless tweezers or by hand with PE gloves</td><td>Hexane</td><td>117</td><td>n.a.</td><td>n.a.</td><td>3.1</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Japanese beaches</td><td>Solvent rinsed stainless tweezers</td><td>n-Hexane</td><td>2300</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td></td><td></td><td></td></tr><tr><td>Santos Bay, Brazil</td><td>Tweezers</td><td>DCM: n-hexane</td><td>n.a.</td><td>39,763</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td></td><td></td><td></td></tr><tr><td>Portuguese coast</td><td>N.S.</td><td>Hexane, hexane-acetone</td><td>54</td><td>319.6</td><td>4.54</td><td>0.81</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td rowspan="2">Belgian coast</td><td rowspan="2">N.S.</td><td>Hexane: DCM</td><td>236</td><td>3007</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td></td><td></td><td></td></tr><tr><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Central Pacific Gyre, Pacific Ocean</td><td>Neuston net and solvent rinsed stainless steel tweezers</td><td>DCM</td><td>78</td><td>868</td><td>4.8</td><td>2</td><td>9909</td><td></td><td></td><td></td><td></td></tr></tbody></table>
<table><tr><td rowspan="2">Japanese beaches</td><td>Neuston net and solvent rinsed stainless steel tweezers</td><td>DCM</td><td>436</td><td>9297</td><td>198</td><td>198</td><td>230</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>Hirai et al. (2011)</td></tr><tr><td>Neuston net and solvent rinsed stainless steel tweezers</td><td>DCM</td><td>102</td><td>2024</td><td>108</td><td>42</td><td>412</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>Hirai et al. (2011)</td></tr><tr><td>Tonkin Bay, Vietnam</td><td>Neuston net and solvent rinsed stainless steel tweezers</td><td>DCM</td><td>399</td><td>656</td><td>8</td><td>7.6</td><td>41</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>Hirai et al. (2011)</td></tr><tr><td>Seal Beach, California</td><td>Neuston net and solvent rinsed stainless steel tweezers</td><td>DCM</td><td>61</td><td>284</td><td>124</td><td>10</td><td>180</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>Hirai et al. (2011)</td></tr><tr><td>Marbella Beach, Costa Rica</td><td>Neuston net and solvent rinsed stainless steel tweezers</td><td>DCM</td><td>980</td><td>500</td><td>22</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>Hirai et al. (2011)</td></tr><tr><td>Hawaii</td><td>Squeeze, scoops, or directly in glass jars</td><td>DCM</td><td>n.a.</td><td>640</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>Rios et al. (2007)</td></tr><tr><td>Mexico</td><td>Squeeze, scoops, or directly in glass jars</td><td>DCM</td><td>n.a.</td><td>640</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>Rios et al. (2007)</td></tr><tr><td>California</td><td>Squeeze, scoops, or directly in glass jars</td><td>DCM</td><td>730</td><td>6200</td><td>1100</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>Rios et al. (2007)</td></tr><tr><td>9 Japanese beaches</td><td>Solvent rinsed stainless tweezers</td><td>DCM</td><td>892</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>Takada et al. (2006)</td></tr><tr><td>Japanese coast</td><td>Neuston net</td><td>N.S.</td><td>254</td><td>9370</td><td>n.a.</td><td>276</td><td>2.1</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>Teuten et al. (2009)</td></tr><tr><td>Central Gyre</td><td>Neuston net</td><td>N.S.</td><td>23</td><td>959</td><td>n.a.</td><td>4.7</td><td>57</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>Teuten et al. (2009)</td></tr><tr><td>Japan and South East Asia</td><td>Soap rinsed fingers or stainless steel tweezers</td><td>Hexane</td><td>453</td><td>n.a.</td><td>163</td><td>11.5</td><td>n.a.</td><td>1.23</td><td>n.a.</td><td>n.a.</td><td>Ogata et al. (2009)</td></tr></table>

(continued)
<table><caption>**Table 8.1** (continued)</caption><thead><tr><th>Location</th><th>Sampling</th><th>Extraction</th><th>PCB (ng/g)</th><th>PAH (ng/g)</th><th>DDT (ng/g)</th><th>DDE (ng/g)</th><th>PBDE (ng/g)</th><th>HCH (ng/g)</th><th>Cd (ng/g)</th><th>Pb (µg/g)</th><th>References</th></tr></thead><tbody><tr><td>US. Coast</td><td>Soap rinsed fingers or stainless steel tweezers</td><td>Hexane</td><td>605</td><td>n.a.</td><td>267</td><td>128</td><td>n.a.</td><td>0.94</td><td>n.a.</td><td>n.a.</td><td>Ogata et al. (2009)</td></tr><tr><td>Africa</td><td>Soap rinsed fingers or stainless steel tweezers</td><td>Hexane</td><td>41</td><td>n.a.</td><td>4.49</td><td>0.91</td><td>n.a.</td><td>37.1</td><td>n.a.</td><td>n.a.</td><td>Ogata et al. (2009)</td></tr><tr><td>India</td><td>Soap rinsed fingers or stainless steel tweezers</td><td>Hexane</td><td>141</td><td>n.a.</td><td>29.8</td><td>15.8</td><td>n.a.</td><td>3.24</td><td>n.a.</td><td>n.a.</td><td>Ogata et al. (2009)</td></tr><tr><td>Australia</td><td>Soap rinsed fingers or stainless steel tweezers</td><td>Hexane</td><td>16</td><td>n.a.</td><td>6.69</td><td>0.57</td><td>n.a.</td><td>0.19</td><td>n.a.</td><td>n.a.</td><td>Ogata et al. (2009)</td></tr><tr><td>West Europe</td><td>Soap rinsed fingers or stainless steel tweezers</td><td>Hexane</td><td>169</td><td>n.a.</td><td>27.6</td><td>15.9</td><td>n.a.</td><td>1.12</td><td>n.a.</td><td>n.a.</td><td>Ogata et al. (2009)</td></tr><tr><td>South Atlantic Ocean</td><td>Manta trawl</td><td>DCM</td><td>590</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>4.6</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>Rochman et al. (2014)</td></tr><tr><td>Saint Helena British territory</td><td>Soap rinsed fingers or stainless steel tweezers</td><td>Hexane</td><td>7</td><td>n.a.</td><td>3.4</td><td>0.5</td><td>n.a.</td><td>19</td><td>n.a.</td><td>n.a.</td><td>Heskett et al. (2012)</td></tr><tr><td>Cocos Islands, Australia</td><td>Soap rinsed fingers or stainless steel tweezers</td><td>Hexane</td><td>6.5</td><td>n.a.</td><td>3.4</td><td>0.8</td><td>n.a.</td><td>1.7</td><td>n.a.</td><td>n.a.</td><td>Heskett et al. (2012)</td></tr><tr><td>Hawaii, USA</td><td>Soap rinsed fingers or stainless steel tweezers</td><td>Hexane</td><td>9.9</td><td>n.a.</td><td>3.4</td><td>0.5</td><td>n.a.</td><td>0.6</td><td>n.a.</td><td>n.a.</td><td>Heskett et al. (2012)</td></tr><tr><td>Island of Oahu, USA</td><td>Soap rinsed fingers or stainless steel tweezers</td><td>Hexane</td><td>1.5</td><td>n.a.</td><td>0.8</td><td>0.6</td><td>n.a.</td><td>&lt;LOQ</td><td>n.a.</td><td>n.a.</td><td>Heskett et al. (2012)</td></tr><tr><td>Barbados, USA</td><td>Soap rinsed fingers or stainless steel tweezers</td><td>Hexane</td><td>1.7</td><td>n.a.</td><td>3.1</td><td>0.3</td><td>n.a.</td><td>&lt;LOQ</td><td>n.a.</td><td>n.a.</td><td>Heskett et al. (2012)</td></tr><tr><td>Canary Islands, Spain</td><td>Soap-rinsed fingers or stainless steel tweezers</td><td>Hexane</td><td>9.0</td><td>n.a.</td><td>4.1</td><td>2.5</td><td>n.a.</td><td>0.6</td><td>n.a.</td><td>n.a.</td><td>Heskett et al. (2012)</td></tr><tr><td>North Pacific Central Gyre</td><td>Hand net, manta trawl</td><td>DCM</td><td>2856</td><td>14,459</td><td>336</td><td>126</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>Rios et al. (2010)</td></tr></tbody></table>
<table><caption>**Table 8.2** Maximum concentrations of selected plastic additives in plastic pellets from field studies</caption><thead><tr><th>Location</th><th>Sampling</th><th>Extraction</th><th>UV stabilizers (µg/g)</th><th>BPA (ng/g)</th><th>Antioxidants (µg/g)</th><th>NP (ng/g)</th><th>OP (ng/g)</th><th>References</th></tr></thead><tbody><tr><td>Japanese coasts</td><td>Solvent rinsed stainless tweezers or by hand with PE gloves</td><td>Hexane</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>16,000</td><td>n.a.</td><td>Mato et al. (2001)</td></tr><tr><td>Central Pacific Gyre, Pacific Ocean</td><td>Neuston net and solvent rinsed stainless steel tweezers</td><td>DCM</td><td>n.a.</td><td>283</td><td>n.a.</td><td>997</td><td>4</td><td>Hirai et al. (2011)</td></tr><tr><td>Japanese beaches</td><td>Neuston net and solvent rinsed stainless steel tweezers</td><td>DCM</td><td>n.a.</td><td>3.4</td><td>n.a.</td><td>1244</td><td>49</td><td>Hirai et al. (2011)</td></tr><tr><td>Tonkin Bay, Vietnam</td><td>Neuston net and solvent rinsed stainless steel tweezers</td><td>DCM</td><td>n.a.</td><td>263</td><td>n.a.</td><td>551</td><td>154</td><td>Hirai et al. (2011)</td></tr><tr><td>Seal Beach, California</td><td>Neuston net and solvent rinsed stainless steel tweezers</td><td>DCM</td><td>n.a.</td><td>26</td><td>n.a.</td><td>130</td><td>17</td><td>Hirai et al. (2011)</td></tr><tr><td>Marbella Beach, Costa Rica</td><td>Neuston net and solvent rinsed stainless steel tweezers</td><td>DCM</td><td>n.a.</td><td>730</td><td>n.a.</td><td>3936</td><td>14</td><td>Hirai et al. (2011)</td></tr><tr><td>9 Japanese beaches</td><td>Solvent rinsed stainless tweezers</td><td>DCM</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>17,000</td><td>n.a.</td><td>Takada et al. (2006)</td></tr><tr><td>Japanese coast</td><td>Neuston net</td><td>N.S.</td><td>n.a.</td><td>&lt;25</td><td>n.a.</td><td>n.s.</td><td>n.a.</td><td>Teuten et al. (2009)</td></tr><tr><td>Central Gyre</td><td>Neuston net</td><td>N.S.</td><td>n.a.</td><td>284</td><td>n.a.</td><td>2,660,000</td><td>n.a.</td><td>Teuten et al. (2009)</td></tr><tr><td>Island of Kauai, Hawaii, USA</td><td>Stainless steel tweezers</td><td>n-Hexane</td><td>1133</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>n.a.</td><td>Tanaka et al. (2020)</td></tr><tr><td>Geoje, South Korea</td><td>N.S.</td><td>DCM</td><td>82</td><td>n.a.</td><td>1620</td><td>n.a.</td><td>n.a.</td><td>Rani et al. (2017)</td></tr><tr><td>South Atlantic Ocean</td><td>Manta trawl</td><td>DCM</td><td>n.a.</td><td>4.9</td><td>n.a.</td><td>n.s.</td><td>n.s.</td><td>Rochman et al. (2014)</td></tr></tbody></table>
steric, non-covalent, or covalent bonds (i.e. π-π interactions) (Brennecke et al. 2016; Hüffer and Hofmann 2016; Velzeboer et al. 2014; Wang et al. 2019).

Sorption is also largely influenced by the surface area to volume ratio of a plastic particle, which increases as the size decreases (Brennecke et al. 2016; Li et al. 2019; Ma et al. 2016; Teuten et al. 2009; Zhan et al. 2016; Zhang et al. 2018). When exposed to weathering, the plastic surface can be subjected to embrittlement and fragmentation, steps that increase the surface/volume and provide more space, a larger contact area, and new sorption sites for external molecules (Napper et al. 2015; Wang et al. 2018b). UV rays-induced weathering, or photo-oxidation, can also lead to chemical alterations and loss of hydrophobicity through the creation of new oxygen-rich functional groups (i.e. carbonyl moieties). Salinity and pH can also play a role in sorption mechanisms. When pH is above the point of zero charge (PZC) of the plastic polymer, it assumes a negative charge that could result in electrostatic repulsion between its surface and other anionic chemicals (Holmes et al. 2014; Li et al. 2019; Wang et al. 2015). Salinity, on the other hand, can either increase the partitioning of nonpolar compounds (salting out) or decrease that of polar molecules due to competition in the adsorption sites (Karapanagioti and Klontza 2008; Llorca et al. 2018; Wang et al. 2015; Zhan et al. 2016; Zuo et al. 2019). This can result in a difference in sorption capacities between freshwater and seawater environments. Exposure time, chain length, and temperature have also been observed as influential factors in hydrophobic partitioning (Engler 2012; Llorca et al. 2018; Mato et al. 2001; Takada et al. 2006; Zhan et al. 2016; Zhang et al. 2018).

Finally, it must be noted that new kinds of biodegradable plastic polymers are being designed and are expected to be more easily and fully degraded over a short time, thus reducing their potential harm. Despite this assumption, Zuo et al. (2019) recently indicated that those highly rubbery MP could become even stronger vectors of organic chemicals. Evaluation of toxicity of the alternative materials and experimental studies are needed to clarify the possible harm.

## 8.2.1 Polycyclic Aromatic Hydrocarbons (PAHs)

The presence of polycyclic aromatic hydrocarbons (PAHs) in the environment can be the result of three different processes: the incomplete combustion of organic material (pyrolytic origin), spillage discharge of crude oil (petrogenic origin), or the post-depositional transformation of biogenic precursors (diagenetic origin). The contribution of petrogenic over pyrogenic sources (and vice versa) can be manifested through the calculation of the ratio between lighter (2–3 rings) and heavier congeners (4–6 rings), with higher values of this parameter indicating a major contribution of fossil sources (low molecular weight congeners). High molecular weight PAHs are generally the ones detected at higher concentrations on plastic pellets in
the environment, implying combustion processes to be the main source of con-
tamination (Gauquie et al. 2015; Rios et al. 2007). In a recent study, low molecular
weight PAHs were mostly found on clearer materials, while high molecular
weight PAHs were mostly detected on darker materials (Fisner et al. 2017). The
colour of microplastics, along with their size and smell, is an important factor
since some organisms may selectively feed on those pellets which resemble their
prey (Chagnon et al. 2018; Hipfner et al. 2018; Ory et al. 2017; Savoca et al. 2016).

8.2.2 Polychlorinated Biphenyls (PCBs)

Polychlorinated biphenyls (PCBs) are a group of 209 lipophilic chemicals of a com-
pletely anthropogenic origin, which have been widely used from the 1930s to 1970s
until they were banned because of their harmful nature.

The concentration of these pollutants in plastic pellets nowadays is mostly
related to the presence of legacy PCBs in those industrialized countries that for-
merly used high amounts of them, which are still present in the environment due
to their persistent nature (Ogata et al. 2009). When lower chlorinated congeners
are detected inside an organism, their presence can be generally linked to the
ingestion of contaminated MPs, since they would be more easily depleted along
the trophic chain (Teuten et al. 2009). For this, only those highly substituted con-
geners would be more prone to be transferred and biomagnify in the trophic chain,
and their exposure would then be mainly caused by prey ingestion (Yamashita
et al. 2011). However, due to the high K<sub>ow</sub> of highly chlorinated PCBs, they are
more efficiently bound to the plastic and then less likely released from polymers
(Colabuono et al. 2010). Their concentration is typically expressed as ICES-7
(ΣPCBs 28, 52, 101, 118, 138, 153, 180), which corresponds to the sum of seven
indicator non-dioxin-like (NDL) congeners presenting the highest concentrations
in technical mixtures and the environment (Webster et al. 2013). They are gener-
ally calculated as the sum of different congeners. The ones typically showing the
highest concentrations are CB 118, 138, 153, 170, and 180 (Antunes et al. 2013;
Colabuono et al. 2010; Gauquie et al. 2015; Mato et al. 2001; Yamashita et al.
2011). Congeners 138, 153, and 180 are the ones mostly detected in human serum
and tissues (JECFA 2016).

Moreover, among PCBs, 12 congeners have a coplanar conformation (dioxin-
like PCBs) that enables them to interact with xenobiotic receptors in the cell (i.e.
aryl hydrocarbon receptor, AhR) and makes them able to interfere with the endo-
crine system, causing reproductive disorders among others (JECFA 2016; Pocar
et al. 2005).
### 8.2.3 Dichlorodiphenyltrichloroethane (DDT)

Dichlorodiphenyltrichloroethane (DDT) is an organochlorine compound used as an insecticide. Total DDT concentration (tDDT) is usually expressed as the sum of DDT and its metabolites DDD and DDE (Antunes et al. 2013; Hirai et al. 2011; Ogata et al. 2009; Rios et al. 2007). When DDE/DDT ratio is very small, it indicates recent contamination. High levels of DDT in the environment and on plastics can be ascribed to their high production levels, especially by the USA, their intensive use as pesticides in agriculture, and as insecticides (JECFA 1961). While DDT production has been banned in some countries such as the USA since 1972, DDT and its derivatives are still in use as insecticides in some countries to prevent the spread of malaria.

### 8.2.4 Polybrominated Diphenyl Ethers (PBDEs)

Polybrominated diphenyl ethers (PBDEs) are a group of 209 anthropogenic chemicals produced and added to materials in order to improve their resistance to fire. They make up one of the major classes of brominated flame retardants (BFR). They are part of the group of lipophilic, persistent organic contaminants, can be accumulated through the food chain, and converted from one congener to another through metabolization. Their concentration in commercial products and, thus, their occurrence in the environment are not as high as those of PCBs. Industrial products are mostly made up of few congeners, so environmental data are available only for penta-hepta-, octa- and deca-BDEs, among which BDE-47 is the most abundant together with BDE-99, BDE-100, and BDE-209 (Darnerud et al. 2001). Their exposure is generally associated with hypothyroidism, since they can bind to thyroid hormone transporters, act as agonists/antagonists, or displace the bound hormones (Darnerud et al. 2001).

The congeners most commonly analysed to check contamination in feed and food are BDE-28, BDE-47, BDE-99, BDE-100, BDE-153, BDE-154, BDE-183, and BDE-209 (EFSA 2011). The log K<sub>ow</sub> of these highly hydrophobic congeners ranges from 5.94 to more than 8 (Braekevelt et al. 2003). This class of POPs has among the greatest potential to cause harm to biological systems even at low concentrations (Abdelouahab et al. 2009; Carlson 1980; Fair et al. 2012).

### 8.2.5 Hexachlorocyclohexanes (HCHs)

Hexachlorocyclohexanes (HCHs) are long-range transported organochlorine pesticides that were mostly used either in agriculture or as insecticides. Several isomers can be present in the commercial mixtures, and when performing environmental
analyses, their concentration is expressed as ΣHCH (α-HCH, β-HCH, γ-HCH, and δ-HCH). γ-HCH, the major component of pesticides, was found as the most abundant isomer in plastic samples from Mozambique and South Africa (Ogata et al. 2009). Log K<sub>ow</sub> (γ-HCH = 3.8) for HCHs is lower than that for PCBs and DDTs, and they are thus supposed to partition the plastic pellets less than the other more hydrophobic chemicals (Mizukawa et al. 2013).

## 8.3 Desorption of Environmental Pollutants from Microplastics

New evidence implies a possible transfer of persistent, bioaccumulative, and toxic (PBT) pollutants into organisms from contaminated plastics (Avio et al. 2015; Browne et al. 2013; Chua et al. 2014; Engler 2012; Ryan et al. 1988; Tanaka et al. 2013; Wardrop et al. 2016; Yamashita et al. 2011). This would lead to an increased exposure to xenobiotic molecules once the plastic item is ingested.

Considering that the bond between a co-contaminant and the plastic substrate can be reversible, sorbed POPs will tend to desorb from the polymer into the water until equilibrium is finally attained (Andrady 2011). The desorption behaviour can be enhanced by the presence of some chemical surfactants, which are molecules able to increase the solubility of hydrophobic substances. These molecules possess an amphipathic nature, containing both hydrophilic and hydrophobic moieties, enabling them to interact with the two phases. Their mechanism of action consists of the lowering of aqueous phase polarity and the formation of micelles to include and contain the nonpolar chemicals. Among them, humic acids and linear alkyl benzene sulphonate, which can be found in the environment, have been found to significantly affect PCDD/DFs and PCBs leaching in some shredder residues and municipal solid waste (Sakai et al. 2000). In the gut, some digestive detergents are present and, through their surfactant properties, could induce the release of toxic substances from the ingested MPs, in a process that is negatively correlated with lipophilicity (Ahrens et al. 2001; Heinrich and Braunbeck 2019). This could suggest a higher degree of bioavailability.

### 8.3.1 Leaching of Additives from Microplastics

Many of the plastic additives (plasticizers, PBDEs flame retardants, antioxidants, and stabilizers) are not chemically bound to the polymer and can thus more easily migrate from the material. Only some reactive organic additives, such as some flame retardants, are polymerized with the plastic molecules becoming part of the polymer chain. All these chemicals are intentionally added during plastic manufacture in order to give plastics some specific features and improve their functional
characteristics (e.g. UV and thermal resistance, flexibility). These compound are mixed within the bulk matrix, and their leaching behaviour can be influenced and enhanced by external factors, such as changes in temperature or pH, plastic ageing, and the presence of surfactants (Suhrhoff and Scholz-Böttcher 2016; US EPA 1992; Wei et al. 2019).

## 8.4 Microplastics and Nanoplastics Occurrence in Foods

Microplastic pollution has become a potential food safety threat that is especially relevant for fishery products as well as other seafood products including table salt. Although microplastics have been reported in products such as honey and sugar (Liebezeit and Liebezeit 2013) or beer (Liebezeit and Liebezeit 2014), aquatic products and water seem to be the best-studied source of dietary intake of microplastics.

### 8.4.1 Microplastics and Nanoplastics Occurrence in Fisheries and Aquaculture Products

Ingestion of microplastics by aquatic organisms have been reported in numerous studies; indeed, microplastics have been found in 12 out of the 25 most important species and genera that contribute to global marine fisheries (FAO 2017).

Small concentrations of microplastics of around one to two particles per fish have been observed in many important commercial species such as sardine, mackerel, anchovy, herring, and sprat from the Pacific, Atlantic, and Indian oceans, as well as the Mediterranean Sea (GESAMP 2016; Lusher et al. 2016). Other aquatic species of local and regional relevance from marine and freshwater environments have been found to contain microplastic particles too (FAO 2017).

Most microplastics have been observed in the gastrointestinal tract of fish, where most particles seem to concentrate after ingestion. However, several studies in marine organisms have shown that smaller microplastics and nanoplastics could be translocated in other organs such as the liver, although the translocation pathways are not well understood yet (Collard et al. 2017).

Bivalve molluscs have also been found to contain microplastic particles. One of the best-studied are mussels, where their occurrence has been reported in Europe, North America, Brazil, and China. The lowest level of microplastic concentration in mussels (less than 0.5 particle/g) was observed in Europe and the highest concentration of microplastics in mussels was reported in China, amounting to 4 particles/g (EFSA 2016).

Microplastics have also been observed in crustaceans, such as shrimps and in 83% of Norway lobsters in coastal waters of the North Sea and the Irish Sea, with
average concentrations ranging from 0.03 to 1.92 particles/g in shrimp, mostly found in the digestive tract (Devriese et al. 2015; Murray and Cowie 2011).

An emerging threat to the consumption of fisheries and aquaculture products derives from the ability of microplastics to sorb environmental contaminants as persistent organic pollutants (POPs), metals, and pathogens, where its concentration can be several folds higher than in the water column. Humans can be exposed to these contaminants through the consumption of fisheries products especially bivalves and crustaceans, where the gastrointestinal tract is not removed. Another route of exposure might be the consumption of farmed species fed with contaminated fish or fishmeal (GESAMP 2016). Information on microplastics presence in different marine organisms (including commercial seafood species), sampling method/instrument, polymer type, and particle size are summarized in Table 8.3.

## 8.4.2 Microplastics and Nanoplastics Occurrence in Salt

Microplastics may contaminate table salt, as other sea products, and their occurrence has been reported in different countries such as Italy, Croatia, and Spain. The average microplastic content in Italy fluctuated from 1.57 (high-quality table salt) to 8.23 (low-quality table salt) particles/g, with a size range from 4 µm to 2100 µm. The results in Croatia fluctuated from and 27.13 (high-quality table salt) to 31.68 (low-quality table salt) particles/g with an average size range from 15 to 4628 µm (Renzi and Blašković 2018).

In Spain, microplastics were found in commercial table salt. The content went from 50 to 280 particles/kg, being PET the most frequently reported polymer, followed by PP and PE (Iñiguez et al. 2017).

## 8.4.3 Microplastics and Nanoplastics Occurrence in Water

The presence of microplastics has been reported in raw and treated drinking water. Depending on the water treatment, microplastic particles are found in different concentrations in water samples. Their average abundance was described from 1473 ± 34 to 3605 ± 497 particles/L in raw water and from 338 ± 76 to 628 ± 28 particles/L in treated water, depending on the treatment (Pivokonsky et al. 2018). A relevant finding is that some microplastic particles were reported to be down to the size of 1 µm and around 12 different microplastics compounds were identified, being PET, PP, and PE the most prominent polymers.
<table><caption>**Table 8.3** Field studies on microplastic occurrence in fisheries and aquaculture products</caption><thead><tr><th>Species</th><th>Particle amount</th><th>Sampling</th><th>Particle size</th><th>Polymers</th><th>Location</th><th>References</th></tr></thead><tbody><tr><td>5 mesopelagic and 1 epipelagic fish species</td><td>1–83 items/fish (2.1 ± 5.78 items/fish)</td><td>Manta trawl</td><td>1–2.79 mm</td><td></td><td>North Pacific Central Gyre</td><td>Boerger et al. (2010)</td></tr><tr><td>10 species of fish (5 pelagic, 5 demersal)</td><td>1–15 items/fish (1.90 ± 0.10 items/fish)</td><td>Standard haul trawl</td><td>0.13 mm–14.3 mm</td><td>PA, LDPE, PS, RY, PEST</td><td>English Channel</td><td>Lusher et al. (2013)</td></tr><tr><td><i>Mytilus edulis</i><br><i>Crassostrea gigas</i></td><td>0.36 ± 0.07 items/g ww<br>0.47 ± 0.16 items/g ww</td><td>Mussel farm, supermarket</td><td>5–10 µm<br>11–20 µm</td><td></td><td>North Sea, Atlantic Ocean</td><td>Van Cauwenberghes and Janssen (2014)</td></tr><tr><td><i>Mytilus edulis</i></td><td>Wild: 34 items/ind<br>106–126 items/ind<br>Farmed: 75 items/ind, 178 items/ind</td><td>Collection at low tide, grocery store</td><td></td><td></td><td>Newfoundland, Nova Scotia, Canada</td><td>Mathalon and Hill (2014)</td></tr><tr><td>26 species of commercial fish</td><td>0–20 items/fish<br>1.40 ± 0.66 items/fish or<br>0.27 ± 0.63 items/fish</td><td>Stern trawlers</td><td>0.217–4.81 mm</td><td>Alkyd resin, PE, PP, RY, PA-6, PEST, acrylic.</td><td>Portuguese coast</td><td>Neves et al. (2015)</td></tr><tr><td>9 species of commercial bivalves</td><td>4.3–57.2 items/ind<br>2.1–10.5 items/g ww</td><td>Shanghai fishery market</td><td>5 µm–5 mm</td><td>PE, PET, PA</td><td>Coastal waters of China</td><td>Li et al. (2015)</td></tr></tbody></table>
<table><tbody><tr><td>Indonesia: 11 fish species<br>USA: 12 fish species and Crassostrea gigas</td><td>Indonesia: 0–21 items/fish (1.4 ± 3.7 items/fish)<br>USA: 0–10/fish 0–2/ oyster (0.5 ± 1.4 items/ind)</td><td>Fish markets, fishermen</td><td>Indonesia: 3.5 ± 1.1 mm USA: 6.3 ± 6.7 mm (fish) 5.5 ± 5.8 mm (oysters)</td><td>Sulawesi (Indonesia) and California (USA)</td><td>Rochman et al. (2015)</td></tr><tr><td>Crangon crangon</td><td>0.68 ± 0.55 items/g ww<br>Max 1.92 ± 0.61 items/g ww<br>1.23 ± 0.99 items/ind</td><td>Shrimp trawl and beam trawl net</td><td>200–1000 µm</td><td>Channel area and southern part of the North Sea (France, Belgium, Netherlands, UK)</td><td>Devriese et al. (2015)</td></tr><tr><td>Mytilus edulis<br>Arenicola marina</td><td>0.2 ± 0.3 items/g ww<br>Max 1.1 items/g ww<br>1.2 ± 2.8 items/g ww<br>Max 11.3 items/g ww</td><td>Randomly on breakwaters<br>Bait-pump or shovel</td><td>20–90 µm 15–100 µm</td><td>LDPE, HDPE, PS<br>French, Belgian, and Dutch North Sea coast</td><td>Van Cauwenberghe et al. (2015)</td></tr><tr><td>5 fish species (3 demersal and 2 pelagic)</td><td>0.03 ± 0.18 items/fish<br>0.19 ± 0.61 items/fish (pelagic)</td><td>Bottom/pelagic trawling</td><td>180 µm–50 cm</td><td>PE, PA, PP, PS, PET, PEST, PU</td><td>North and Baltic seas</td></tr></tbody></table>

(continued)
<table><caption>Table 8.3 (continued)</caption><thead><tr><th>Species</th><th>Particle amount</th><th>Sampling</th><th>Particle size</th><th>Polymers</th><th>Location</th><th>References</th></tr></thead><tbody><tr><td><i>Mytilus edulis</i></td><td>1.5–7.6 items/ind<br>(4 items/ind)<br>0.9–4.6 items/gww<br>(2.2 items/g ww)</td><td>Tweezers at low tide, underwater with fishermen</td><td>&lt; 250 µm Fibres:<br>33 µm–4.7 mm</td><td>CP, PET, PEST</td><td>Coastline of China</td><td>Li et al. (2016)</td></tr><tr><td><i>Perna perna</i></td><td>75% mussels contaminated with MPs</td><td>Bysus cut and removal</td><td>&lt; 5000 µm</td><td></td><td>Santos estuary, São Paulo, Brazil</td><td>Santana et al. (2016)</td></tr><tr><td><i>Saccostrea cucullata</i></td><td>1.4–7.0 items/ind<br>1.5–7.2 items/gww</td><td>Intertidal zone</td><td>20–5000 µm (mostly &lt;100 µm)</td><td>PET, PP, PE, PS, CP, PVC, PA, EPS</td><td>Pearl River estuary, South China</td><td>Li et al. (2018a)</td></tr><tr><td><i>Penaeus monodon</i><br><i>Metapenaeus monoceros</i></td><td>6.60 ± 2 items/ind<br>1.55–4.84 items/g ww<br>(3.40 ± 1.23 items/g ww)<br>7.80 ± 2 items/ind<br>2.17–4.88 items/g ww<br>(3.87 ± 1.05 items/g ww)</td><td>Offshore shrimp trawlers and bag net fishing</td><td>1–5 mm<br>250–500 µm</td><td>PA-6, RY</td><td>Northern Bay of Bengal, Bangladesh</td><td>Hossain et al. (2020)</td></tr></tbody></table>
<table><tr><td>Venerupis philippinarum</td><td>0.07–5.47 items/g ww<br>Wild: 8.4 ± 8.5 items/ind<br>Max 12.7 ± 13.0 items/ind<br>0.9 ± 0.9 items/g Farmed:<br>11.3 ± 6.6 items/ind<br>Max 15.4 ± 6.3 items/ind<br>1.7 ± 1.2 items/g Max 2.2 ± 0.8 items/g</td><td>0.5 m × 0.5 m quadrat, from farmed sites</td><td></td><td>Baynes sound, British Columbia</td><td>Davidson and Dudas (2016)</td></tr><tr><td>Emerita analoga</td><td>0.65 ± 1.64 items/ind<br>Max 16 items/ind</td><td>Shovel, sand-coring tool</td><td></td><td>PP, polyacrylate, PEST, PE</td><td>Pacific coast of California</td></tr><tr><td>Lithodes santolla</td><td>1–3 items/ind</td><td>Trap gears</td><td>3–>20 mm</td><td></td><td>Nassau Bay, Cape Horn, Chile</td></tr></table>

(continued)
<table><caption>Table 8.3 (continued)</caption><thead><tr><th>Species</th><th>Particle amount</th><th>Sampling</th><th>Particle size</th><th>Polymers</th><th>Location</th><th>References</th></tr></thead><tbody><tr><td><em>Saccostrea forskalii</em><br><em>Littoraria sp.</em><br><em>Balanus amphitrite</em></td><td>0.37 ± 0.03–<br>0.57 ± 0.22<br>items/g ww<br>0.17 ± 0.08–<br>0.23 ± 0.02<br>items/g ww<br>0.23 ± 0.10–<br>0.43 ± 0.33<br>items/g ww</td><td>Intertidal zone at<br>low tide</td><td></td><td>PS, PET, PA</td><td>Upper gulf of<br>Thailand</td><td>Thushari et al.<br>(2017)</td></tr><tr><td><em>Crassostrea gigas</em><br><em>Gammarus sp.</em><br><em>Mytilus edulis</em><br><em>Littorina littorea</em></td><td>87 items/g dw<br>11 items/g dw<br>105 items/g dw<br>20 items/g dw</td><td>From littoral<br>zones</td><td>1–5000 µm (Mostly<br>1–300 µm)</td><td></td><td></td><td>Leslie et al.<br>(2013)</td></tr><tr><td><em>Mytilus edulis</em></td><td>0.04–0.81<br>items/g ww<br>(0.37 items/g<br>ww)</td><td>Department<br>stores, manual<br>sampling at<br>groyne and<br>quayside</td><td>200 µm–1500 µm</td><td></td><td>Belgium,<br>Netherlands</td><td>De Witte et al.<br>(2014)</td></tr><tr><td><em>Chelon aurata</em><br><em>Rutilus kutum</em></td><td>2.95 ± 1.98<br>items/fish<br>1.66 ± 1.23<br>items/fish</td><td></td><td>1.94 ± 0.71 mm<br>1.77 ± 0.53 mm</td><td>PA</td><td>Southern Caspian<br>Sea</td><td>Zakeri et al.<br>(2020)</td></tr><tr><td><em>Platichthys flesus</em><br><em>Diplodus vulgaris</em><br><em>Dicentrarchus labrax</em></td><td>0.18 ± 0.55<br>items/fish<br>3.14 ± 3.25<br>items/fish<br>0.30 ± 0.61<br>items/fish</td><td>Beam trawl</td><td></td><td>PE, PP, PEST, PA-6, RY,<br>PAN</td><td>Mondego estuary,<br>Portugal</td><td>Bessa et al.<br>(2018)</td></tr></tbody></table>
<table><tr><td>21 species of marine fish, 6 species of freshwater fish</td><td>1.1 ± 0.3–7.2 ± 2.8 items/fish<br>0.2 ± 0.1–17.2 ± 9.7 items/g</td><td>Fishery markets, fishermen</td><td>0.04–5 mm</td><td>CP, PET, PEST</td><td>Yangtze estuary, East China and South China Sea, Taihu lake</td><td>Jabeen et al. (2017)</td></tr><tr><td>Boops boops</td><td>2.47 ± 0.23–4.89 ± 0.45 items/fish</td><td>Bottom trawl nets and purse seine</td><td>1 mm–5 mm</td><td></td><td>Balearic Islands</td><td>Nadal et al. (2016)</td></tr><tr><td>6 marine fish species</td><td>1.62 ± 1.58–2.96 ± 5.21 items/fish (0–32 items/fish)</td><td>Hook and line</td><td>&lt;5 mm</td><td></td><td>Texas Gulf Coast</td><td>Peters et al. (2017)</td></tr><tr><td>28 marine fish species</td><td>2.36 items/fish (1–35 items/fish)</td><td>Trawl net</td><td>656 ± 803 µm (range: 9.07–12074.11 µm)</td><td>Copolymers, alloys, PA, LDPE, PP, rubber</td><td>Mediterranean coast of Turkey</td><td>Güven et al. (2017)</td></tr><tr><td>26 marine fish species</td><td>0–3 items/fish</td><td>Tucker nets, fishermen</td><td>2.39 ± 0.28 mm (range: 1–3 mm)</td><td>PP, PE, PS, PVC, PAN</td><td>Saudi Arabian Red Sea coast</td><td>Baalkhuyur et al. (2018)</td></tr><tr><td>Hoplosternum littorale</td><td>3.6 items/fish (1–24 items/fish)</td><td>Fishermen</td><td>&lt;1 mm–12 mm</td><td></td><td>Pajerū River, Northeast Brazil</td><td>Silva-Cavalcanti et al. (2017)</td></tr><tr><td>Scomberomorus cavalla<br>Rhizoprionodon lalandi</td><td>2–6 items/fish<br>1–3 items/fish</td><td>Fishermen</td><td>2–5 mm<br>1–3 mm</td><td></td><td>Salvador, Northeastern Brazil</td><td>Miranda and de Carvalho-Souza (2016)</td></tr><tr><td>Rastrelliger kanagurta<br>Stolephorus waitei<br>Chelon subviridis<br>Johnnius belangerii</td><td>0–3 items/fish</td><td>Local markets</td><td></td><td>PP, PE, PS, PET, PA-6</td><td>Malaysia</td><td>Karami et al. (2017a)</td></tr></table>

(continued)
<table><caption>**Table 8.3** (continued)</caption><thead><tr><th>Species</th><th>Particle amount</th><th>Sampling</th><th>Particle size</th><th>Polymers</th><th>Location</th><th>References</th></tr></thead><tbody><tr><td><i>Alepes djedaba</i></td><td>0.80 ± 0.12</td><td rowspan="5">Fishmongers</td><td rowspan="5">&lt;300 µm</td><td rowspan="5"></td><td rowspan="5">Persian Gulf</td><td rowspan="5">Akhbarizadeh et al. (2018)</td></tr><tr><td><i>Epinephelus coioides</i></td><td>items/g muscle</td></tr><tr><td><i>Sphyraena jello</i></td><td>0.78 ± 0.22</td></tr><tr><td rowspan="3"><i>Platycephalus indicus</i></td><td>items/g muscle</td></tr><tr><td>0.57 ± 0.17</td></tr><tr><td>items/g muscle</td></tr><tr><td></td><td>1.85 ± 0.46</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td><i>Mvitus galloprovincialis</i></td><td>1.7 ± 0.2-</td><td rowspan="10">Hand collection, trawling</td><td rowspan="10">0.1-0.5 mm</td><td rowspan="10">PE, PP, PET, PS, PTFE</td><td rowspan="10">Northern Ionian Sea</td><td rowspan="10">Digka et al. (2018)</td></tr><tr><td><i>Sardina pilchardus</i></td><td>2.0 ± 0.2 items/ind</td></tr><tr><td><i>Pagellus erythrinus</i></td><td>(1.9 ± 0.2 items/ind)</td></tr><tr><td rowspan="5"><i>Mullus barbatus</i></td><td>2.5 ± 0.3-</td></tr><tr><td>5.3 ± 0.5 items/gww</td></tr><tr><td>1.8 ± 0.2 items/fish</td></tr><tr><td>34.9 ± 7.9</td></tr><tr><td>items/g ww</td></tr><tr><td></td><td>1.9 ± 0.2 items/fish</td></tr><tr><td></td><td>27.8 ± 24.6</td></tr><tr><td></td><td>items/g ww</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td>1.5 ± 0.3 items/fish</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td>11.2 ± 2.8</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td>items/g ww</td><td></td><td></td><td></td><td></td><td></td></tr></tbody></table>
<table><tr><td>Engraulis japonicus</td><td>2.3 ± 2.5 items/fish (0–15 items/fish)</td><td>Fishing</td><td>783 ± 1020 µm (150–6830 µm)</td><td>PE, PP, PS, ethylene/ propylene copolymer, ethylene/propylene/diene terpolymer</td><td>Tokyo Bay</td><td>Tanaka and Takada (2016)</td></tr><tr><td>Mugil cephalus</td><td>Wild: 4.3 ± 14.5 items/fish<br>Captive:<br>0.20 ± 0.48 items/fish</td><td>Fish markets, fish ponds</td><td>0.1–12 mm</td><td>PP, PE, PEST, PET, PA, PTFE</td><td>East coast of Hong Kong</td><td>Cheung et al. (2018)</td></tr><tr><td>Platycephalus indicus, Saurida tumbil, Sillago sihama, Cynoglossus abbreviatus, Penaeus semisulcatus</td><td>0.16–1.5 items/g</td><td>Trawl net</td><td>&lt; 100 µm-&gt; 1000 µm</td><td></td><td>Musa Estuary, Persian Gulf</td><td>Abbasi et al. (2018)</td></tr><tr><td>Crassostrea gigas</td><td>0.07 ± 0.06 items/g (0–0.19 items/g)</td><td>Fishery markets</td><td>100–200 µm</td><td>PE, PP, PS PEST, silicon, PA, PEVA, PET, PU, acrylic, PTFE, PVC, PPS, PVA, PBT, polyethyl acrylate styrene, polyepoxides, styrene/acrylonitrile, and polystyrene ethylene butylene styrene</td><td>Seoul, Gwangju, Busan</td><td>Cho et al. (2019)</td></tr><tr><td>Mvitilus edulis</td><td>0.12 ± 0.11 items/g (0–0.35 items/g)</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Tapes philippinarum</td><td>0.34 ± 0.31 items/g</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>Patinopecten</td><td>(0.03–1.08 items/g)</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td>yessoensis</td><td>0.08 ± 0.08 items/g</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td>(0.01–0.17 items/g)</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td>Overall: 0–1.08 items/g ww</td><td></td><td></td><td></td><td></td><td></td></tr><tr><td></td><td>0–2.8 items/ind</td><td></td><td></td><td></td><td></td><td></td></tr></table>

(continued)
<table><caption>**Table 8.3** (continued)</caption><thead><tr><th>Species</th><th>Particle amount</th><th>Sampling</th><th>Particle size</th><th>Polymers</th><th>Location</th><th>References</th></tr></thead><tbody><tr><td><em>Cerithidea cingulata</em><br><em>Thais mutabilis</em><br><em>Amiantis umbonella</em><br><em>Amiantis purpuratus</em><br><em>Pincitada radiata</em></td><td>Clams and<br>oyster: 0.2–2.2<br>items/g ww<br>3.9–6.9 items/<br>ind<br>Snails:<br>12.8–20.0<br>items/g ww<br>3.7–17.7 items/<br>ind</td><td>Intertidal zone at<br>low tide</td><td>10–> 5000 µm</td><td>PE, PET, PA</td><td>Persian Gulf</td><td>Naji et al. (2018)</td></tr><tr><td><em>Mytilus edulis</em></td><td>0.23 ± 0.09<br>items/g</td><td>Supermarket,<br>collection from<br>the wild</td><td>30–200 µm</td><td>PP, PE, PEST, ABS</td><td>Region Pays de la<br>Loire, market in<br>Nantes</td><td>Phuong et al.<br>(2018)</td></tr><tr><td><em>Mytilus galloprovincialis</em></td><td>8.33 ± 3.58<br>items/g ww<br>3.6–12.4 items/<br>ind<br>4.4–11.4 items/g<br>ww</td><td>Local markets,<br>collection from<br>rocky bottom</td><td>0.75–6.00 mm<br>(average:<br>1.15–2.29 mm)</td><td></td><td>Mediterranean Sea</td><td>Renzi et al.<br>(2018)</td></tr><tr><td><em>Modiolus modiolus</em><br><em>Mytilus</em> spp.</td><td>0.086 ± 0.031<br>items/g ww<br>3.5 ± 1.29 items/<br>ind<br>3.0 ± 0.9 items/g<br>ww<br>3.2 ± 0.52 items/<br>ind</td><td>Scuba diving,<br>collection at low<br>tide</td><td>0.2–> 2 mm</td><td>PET, PEST, poly<br>(etherurethane)</td><td>Scottish coast</td><td>Catarino et al.<br>(2018)</td></tr></tbody></table>
<table><tr><td>Mytilus edulis</td><td>0.7–2.9 items/g ww<br>1.1–6.4 items/ind</td><td>Collection from the coast</td><td>8 µm–4.7 mm</td><td>PEST, PP, PE, RY</td><td>UK</td><td>Jiana Li et al. (2018b)</td></tr><tr><td>Mytilus galloprovincialis<br>Mytilus edulis</td><td>0.04 ± 0.09–<br>0.34 ± 0.33 items/g ww<br>(0.13 ± 0.14 items/g ww)</td><td>Hand collection, department stores</td><td></td><td></td><td>Europe</td><td>Vandermeersch et al. (2015)</td></tr><tr><td>Pyropia spp.</td><td>Packaged:<br>0.9–3.0 items/g dw<br>(1.8 ± 0.7 items/g dw)<br>Processed:<br>1.0–2.8 items/g dw<br>(1.8 ± 0.6 items/g dw)</td><td>Local markets</td><td>Packaged:<br>0.11–4.97 mm<br>Processed:<br>0.07–4.74 mm</td><td>PEST, RY, PP, PA, CP</td><td>China, Yellow Sea</td><td>Li et al. (2020)</td></tr><tr><td>Mytilus edulis<br>Perna viridis</td><td>1.52–5.36 items/g ww<br>0.77–8.22 items/ind</td><td>Tweezers at low tide</td><td>0.25–1 mm</td><td>PET, PE, RY, PVC, PP</td><td>Coastal waters of China</td><td>Qu et al. (2018)</td></tr></table>

(continued)
<table><caption>**Table 8.3** (continued)</caption><thead><tr><th>Species</th><th>Particle amount</th><th>Sampling</th><th>Particle size</th><th>Polymers</th><th>Location</th><th>References</th></tr></thead><tbody><tr><td rowspan="2"><i>Mytilus</i> spp.</td><td>0–6.9 items/ind (1.5 ± 2.3 items/ind)</td><td rowspan="2">Hand collection, snorkelling, metal rake with net</td><td rowspan="2">70 µm–3870 µm</td><td rowspan="2">PET, PP, PE, PA, epoxy resin, PVC, PAN, styrene acrylonitrile, EVA, soloprene plastomer</td><td rowspan="2">Norway</td><td rowspan="2">Bräte et al. (2018)</td></tr><tr><td>0–7.9 items/g ww (0.97 ± 2.61 items/g ww)</td></tr><tr><td rowspan="4"><i>Mytilus galloprovincialis</i><br><i>Chlamys farreri</i></td><td>1.9–9.6 items/ind</td><td rowspan="4">Fishery markets, coastal sampling</td><td rowspan="4">25 µm–5 mm</td><td rowspan="4">CP, PP, PTFE</td><td rowspan="4">China</td><td rowspan="4">DING et al. (2018)</td></tr><tr><td>2.0–12.8 items/g ww</td></tr><tr><td>5.2–19.4 items/ind</td></tr><tr><td>3.2–7.1 items/g ww</td></tr></tbody></table>
## 8.5 Risk Profiling of Microplastics in Fisheries and Aquaculture Products

### 8.5.1 Microplastics Dietary Intake

The dietary intake of microplastics in foods depends on consumption habits, especially when dealing with fisheries and aquaculture products that mainly accumulate microplastics in the gastrointestinal tract, which minimizes the direct exposure to these particles when aquatic products are degutted. Previous studies and reviews have provided some estimates of human intake of microplastic particles through the consumption of different food commodities, and their results are illustrated in Table 8.4.

Exposure can be higher through consumption of small aquatic species such as crustaceans, echinoderms, bivalves, and small-sized fish that are commonly eaten whole. Microplastics have been also found in the muscle of canned, dried, and fresh commercial fish species, suggesting that evisceration may not be a completely efficient process, but the number of particles found in muscle is relatively low compared to the number of particles in the gastrointestinal tract (Abbasi et al. 2018; Akhbarizadeh et al. 2018; Karami et al. 2018, 2017a). In crustaceans such as shrimps, peeling practices that remove most of the digestive tract, the head, and the gills will reduce the exposure to plastics, as these parts are estimated to contain 90 percent of the microplastic particles (Devriese et al. 2015).

In Europe, human exposure to microplastics resulting from the consumption of bivalves may account for 1800 to 11,000 particles/year (Van Cauwenberghe and Janssen 2014). Mussels are among the most consumed bivalve molluscs, and the occurrence of microplastics in these products has been reported in several studies (EFSA 2016).

### 8.5.2 Microplastics Uptake and Toxicity

The physical characteristics of microplastic particles (size, shape), as well as the chemical characteristics, in addition to leaching of additives and pollutants or transport of pathogens, are the main factors for estimating the impact of these particles on food safety, but literature at this respect is indeed limited, and evidence of a possible transfer through the diet are lacking, as are the possible consequences.

Some studies have shown that small plastic particles are able to cross human placenta (Ragusa et al. 2021; Wick et al. 2010), which suggests a possible systemic increase in exposure as their size decreases (De Jong et al. 2008). The most frequent route of uptake would be absorption from the gut epithelium into the lymphatic system (Hussain et al. 2001). Absorption from the intestine is known to be very limited, and it should occur only for those microparticles <150 µm (EFSA 2016; FAO 2017). According to their size, particles are supposed to face different densities
<table><caption>**Table 8.4** Previous estimates of exposure to microplastics through the consumption of seafood, water, and salt</caption><thead><tr><th>Food</th><th>Reference intake</th><th>Derived MP intake</th><th>Country</th><th>References</th></tr></thead><tbody><tr><td>Molluscs</td><td>72.1 g/day (top consumers)<br>11.8 g/day (minor consumers)</td><td>11,000 MP/year (top consumers)<br>1800 MP/year (minor consumers)</td><td>Europe</td><td>Van Cauwenberghe and Janssen (2014)</td></tr><tr><td>Fish muscle</td><td>300 g/week (adults)<br>50 g/week (children)</td><td>169–555 MP/week (adults)<br>28–92 MP/week (children)</td><td>Iran</td><td>Akhbarizadeh et al. (2018)</td></tr><tr><td>Bivalves</td><td>3.01 g/day</td><td>212 MP/year</td><td>South Korea</td><td>Cho et al. (2019)</td></tr><tr><td>Shellfish</td><td>4.03 g/day</td><td>283 MP/year</td><td>South Korea</td><td>Cho et al. (2019)</td></tr><tr><td>Mussels</td><td>82 g/year</td><td>123 MP/year</td><td>UK</td><td>Catarino et al. (2018)</td></tr><tr><td>Mussels</td><td>3.08 kg/year</td><td>4620 MP/year</td><td>Spain/ France/ Belgium</td><td>Catarino et al. (2018)</td></tr><tr><td>Mussels</td><td>225 g</td><td>7 µg<br>0.1 µg/kg bw/ day</td><td>Globally</td><td>EFSA (2016) and FAO (2017)</td></tr><tr><td>Fish Crustaceans<br>Molluscs</td><td>15.21 kg/year<br>2.06 kg/year<br>2.65 kg/year</td><td>31–8323 MP/year<br>206–17,716 MP/year<br>0–27,825 MP/year</td><td>Globally<br>Globally<br>Globally</td><td>Danopoulos et al. (2020), Danopoulos et al. (2020) and Danopoulos et al. (2020)</td></tr><tr><td>Water</td><td>2 l</td><td>85 µg/day<br>1.4 µg/kg bw/ day</td><td>Globally</td><td>WHO (2019)</td></tr><tr><td>Water</td><td>2.2 l/day (women)<br>3 l/day (men)</td><td>4400 MP/year (women)<br>5800 MP/year (men)</td><td>Globally</td><td>Kosuth et al. (2018)</td></tr><tr><td>Salt</td><td>5 g/day</td><td>1000 MP/year</td><td>China</td><td>Yang et al. (2015)</td></tr><tr><td>Salt</td><td>3.95 g/day</td><td>37 MP/year</td><td>Globally</td><td>Karami et al. (2017b)</td></tr><tr><td>Salt</td><td>5 g/day</td><td>40.6–1085.2 MP/year</td><td>Italy</td><td>Renzi and Blašković (2018)</td></tr><tr><td>Salt</td><td>14.8–18.01 g/day</td><td>64–302 MP/year</td><td>Turkey</td><td>Gündoğdu (2018)</td></tr><tr><td>Salt</td><td>2.3 g/day</td><td>40–680 MP/year</td><td>Globally</td><td>Kosuth et al. (2018)</td></tr><tr><td>Salt</td><td>10.06 g/day</td><td>0–42,600 MP/year (average 3000)</td><td>Globally</td><td>Kim et al. (2018)</td></tr><tr><td>Salt</td><td>5 g/day</td><td>510 MP/year</td><td>Spain</td><td>Iñiguez et al. (2017)</td></tr><tr><td>Salt</td><td>5 g/day</td><td>117 µg/year</td><td>India</td><td>Seth and Shriwastav (2018)</td></tr></tbody></table>
inside the body, being either phagocyted by macrophage (> 0.5 µm), endocyted
(<0.5 µm), cleared through splenic filtration (>0.2 µm), or eliminated via kidney
filtration (<10 nm) (Monti et al. 2015; Yoo et al. 2011). Nevertheless, it has to be
underlined that microplastics in the lymph will be excreted mostly through the
spleen or with faeces through bile clearance in the liver if present in the blood (Yoo
et al. 2011).

Among the possible adverse effects caused by microplastic exposure, oxidative stress and alteration of the immune function, possibly leading to immune depression, are the most likely to occur (Petit et al. 2002; Schirinzi et al. 2017). This is because some of the particles can be taken up in the lymphatic system by phagocytic cells, such as macrophages, as mentioned before.

Furthermore, it has to be underlined that PVC could cause additional toxicity because of the possible leaching of vinyl chloride, an extremely toxic monomer, classified as carcinogenic, that is used in the production process of this material and makes up 50 to 100% of the polymer by weight (Lithner et al. 2011). The toxicity of plastic is likely due to the unreacted residual monomers that constitute this material, as they may induce genotoxic effects. Lithner et al. (2011) have elaborated a hazard ranking of monomers and additives that are present in plastic polymers, indicating the hazard for human health. PVC is also the type of polymer that requires the highest amount of additives followed by PP, PE, and styrenics.

In addition to possible adverse health consequences of the polymers itself, the combined exposure to plastic additives and associated co-contaminants adsorbed by the surrounding environment must be taken into consideration. Many additives, such as BPA, phthalates, nonylphenols, and PBDE, are known to have an impact in the organism they enter in contact with, mostly through a mechanism of endocrine disruption. Besides, the main harm derives from the fact that these molecules, which give plastic some specific characteristics, are not strongly bound to the polymer matrix and tend to easily leach from it.

On the other hand, persistent organic pollutants (POPs) that interact and become associated with microplastic particles could not only impact and alter endocrine functions in the organism (i.e. PCBs, PBDEs) but also promote carcinogenicity (PAH). These outcomes mainly arise from the interaction with intracellular receptors and gene expression induction (JECFA 2016; Pocar et al. 2005). In consonance with the strength and type of interaction they have with the polymer's matrix, they could be released at different degrees.

The toxicity of dioxins and dioxin-like PCBs can be expressed through their corresponding toxic equivalency factor (TEF), implemented in the 1980s (Barnes 1991; Safe et al. 1985; Safe 1986) and later revised in 2005 (Van den Berg et al. 2006). This parameter expresses the relative toxicity of a compound in respect of a standard compound of known toxicity, the 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD, TEF = 1). The toxic equivalency quotient (TEQ) is a value resulting from the sum of the weighted concentrations of each chemical multiplied by their TEF. Maximum uptake levels have been set for dioxin-like PCBs and dioxins in 2006 (European Commission 2006).
Allowable concentrations are estimated as provisional maximum tolerable daily intake (PTDI), provisional tolerable weekly intake (PTWI), and provisional tolerable monthly intake (PTMI) and can be expressed as WHO-TEQ. Their values are established during the hazard characterization step, by evaluating the no observed adverse effect level (NOAEL), and indicate the amount of a chemical that can be ingested daily, weekly, or monthly over a lifetime without adverse health consequences (JECFA 1995). In 2017, the FAO conducted an exposure exercise that considered mussels as the target species of greatest interest because of their high consumption and because it is consumed with the viscera, where microplastics tend to be located. The exposure assessment took into consideration the highest concentration of microplastics in mussels, which was reported at 4 particles/g in China (EFSA 2016). According to CIFOCOss, the highest reported consumption of mussels corresponded to the Belgian elderly (P95) and was estimated to be 250 g per day per person. Considering that the highest concentration of microplastics reported in mussels was 4 particles/g and the highest consumption was 250 g of mussels, it was estimated that a portion of mussels could contain up to 1000 microplastic particles, i.e. around 9 µg, depending on the volume and density of the particles.

The exposure assessment on microplastics in mussels confirmed that the intake of microplastics per day was 0.15 µg/kg for a person of 60 kg. These values were selected in order to describe the worst-case scenario and cover all the populations, including those at the highest risk.

Although the exposure assessment for mussels was carried out, the tolerable daily intake (TDI) for microplastic particles and most of its compounds has not yet been established; therefore, it is not possible to determine if this level of exposure is safe. However, the TDI for some plastic additives such as phthalates, BPA, alkyl-phenols, and brominated flame retardants, as well as associated sorbed contaminants such as PCBs, PAHs, and DDT, has been established, and the exposure assessment estimation shows that the level of the compounds present in microplastics from mussels was significantly lower than the TDI. Therefore, based on these assumptions, the intake of associated chemicals from ingested plastics via seafood consumption is minor compared to the total intake from the diet (FAO 2017).

As a complement, an exposure assessment exercise of microplastics in shellfish has been carried out with updated information from the current literature and also taking into account other contaminants that were not included in the previous estimates (e.g. HCHs). Procedures and outcomes are described in the following paragraph.
### 8.5.3 Case Study: Exposure Assessment of Microplastic Additives and Associated Sorbed Contaminants via Shellfish Consumption

Since plastic particles are more likely found in the intestinal tract, the consumption of all those seafood organisms whose GI is not removed can be the main route of exposure. Almost all shellfish are eaten whole, and in the present study, the attention is focused on mussels, shrimps, prawns, clams, and oysters. Data on daily seafood consumption were taken from the Chronic Individual Food Consumption Database (CIFOCOss) of the WHO. In order to perform an exposure assessment, countries reporting the highest consumption levels of shellfish were selected. These countries were China and Finland (mussels, clams, oysters) and the Netherlands (shrimps and prawns). Only the P95 consumers, meaning top consumers of these products from each country, were considered. The adults and elderly category presented the highest daily consumption levels (g/day).

Information about microplastic load in shellfish was taken from scientific papers, only considering the highest detected amount, described in Table 8.3. Plastic materials were assumed to be spherical with a diameter of 25 µm, which was among the most common plastic sizes found in a study by Van Cauwenberghe and Janssen (2014) in mussels. With this information, it was possible to calculate the volume of plastic particles. The density of polymers was taken from scientific papers and reviews, and only maximum values were considered (Andrady 2017; Avio et al. 2017). By knowing the volume and the density, it was then possible to determine the weight of each plastic polymer.

Data on microplastic contamination in bivalve molluscs were taken for clams (*Scapharca subcrenata*), mussels (*Mytilus galloprovincialis*), and oysters (*Saccostrea cucullata*) from three studies on commercial species (Table 8.3). For these species, the highest plastic load was estimated to be 10.5, 12.8, and 7.2 particles/g wet weight, respectively (DING et al. 2018; Li et al. 2015, 2018a). In addition to these, information on shrimp contamination were taken from a recent study, where up to 4.88 particles/g tissue (wet weight) were found in commercial brown shrimp (*Metapenaeus monoceros*) (Hossain et al. 2020).

Indirect exposure to microplastic-bound pollutants was estimated by using the highest reported contamination levels of plastic particles found in field studies (Tables 8.1 and 8.2). These data were used to provide an estimate of exposure to microplastic-bound contaminants through the consumption of shellfish.

The exposure assessment estimation consists of several steps. First, the maximum load of microplastics in shellfish (particles/g) was multiplied by the daily dietary intake of each commodity, or consumer P95 (g/day), thus measuring the total number of particles ingested every day. This result was then multiplied by the weight of each polymeric particle derived before, in order to obtain the estimated consumption of plastic per day (g/day). Then, this value was multiplied by the maximum concentration of each contaminant (ng/g) reported in field studies. With these estimations, it was possible to establish the daily intake of pollutants (ng/day).
which, converted in picograms (pg) and divided by the average body weight of an adult (60 kg), would finally provide information on the daily dietary exposure to environmental contaminants mediated by microplastics (pg/ kg bw/day). This information is presented by commodity and polymer type in Table 8.5. The calculations were made based on the assumption that these chemicals were completely released from the microplastic particles.

The highest exposure to environmental contaminants associated with microplastics and plastic additives, as estimated in the present case study, could derive from the consumption of oysters, followed by mussels, clams, and finally shrimps and prawns (Table 8.5). This could cause concern because the three most contaminated shellfish groups are also the ones that are eaten whole, whereas shrimps are peeled, and most microplastics are removed.

Overall, nonylphenols are the group of xenobiotics that presented the highest microplastic-mediated exposure concentrations, which ranged from 0.25 ng/kg bw/day in PP and LDPE-contaminated shrimps to 2.33 ng/kg bw/day in PVC-contaminated mussels. These high levels can most likely be related to the use of this additive in the manufacture of plastics. After NPs, PAHs are the class of environmental pollutants that could bring out the highest harm through shellfish consumption, with a concentration ranging from 4.15 pg/kg bw/day in PP-contaminated shrimps to 39.26 pg/kg bw/day in PVC-contaminated mussels. Contrariwise, according to the present results, HCHs could be classified as the compounds whose sorbed concentration levels on microplastics could pose the least concern. In fact, the daily dietary exposure to HCHs varied from 0.00 pg/kg bw/day to 0.03 pg/kg bw/day, at maximum.

The results of the exercise demonstrate that polypropylene (PP) is the polymer that might raise the least concern when analysing microplastic-mediated exposure to xenobiotics, while PVC might be the most hazardous. Despite this, the dietary exposure to environmental contaminants on ingested plastics, as calculated in the present case study, can be considered negligible, compared to other sources. Also, there is still not enough clarity on tissue transfer dynamics and concentration of chemicals associated with microplastics, which can be influenced by factors as a fugacity gradient, so this can be considered as an estimate. In conclusion, when considering the outcomes of the present exercise, it is important to also keep into account the physical-chemical properties of each polymer type, as they could also influence the sorption/desorption of chemicals and then their concentration on the microplastics.

In addition to the previous analysis, China and Finland seafood consumption values were used to perform an estimate of the total dietary exposure to organic pollutants through shellfish consumption (Table 8.6). These two countries were chosen because they represent the ones with the highest shellfish consumption levels.

When only these two countries are considered in the analysis, the exposure levels to microplastic-associated contaminants seem to be higher. PVC is again the plastic polymer that can apparently cause the most significant exposure to environmental contaminants in organisms after consumption, followed by PET, PA, PS, HDPE, LDPE, and finally PP. Anyway, PVC is generally reported to sorb pollutants to a
<table><caption>**Table 8.5** Overall maximum dietary exposure (pg/kg/day) of MP additives and associated sorbed pollutants resulting from the consumption of the four shellfish groups in the Netherlands, Finland, and China (age class: adults and elderly). Results are categorized by polymer type</caption><thead><tr><th>Polymer</th><th>Seafood</th><th>MP-PAH<br>(pg/kg bw/day)</th><th>MP-PCB<br>(pg/kg bw/day)</th><th>MP-DDT<br>(pg/kg bw/day)</th><th>MP-PBDE<br>(pg/kg bw/day)</th><th>MP-HCH<br>(pg/kg bw/day)</th><th>MP-BPA<br>(pg/kg bw/day)</th><th>MP-NP<br>(ng/kg bw/day)</th></tr></thead><tbody><tr><td rowspan="3">HDPE</td><td>Shrimps and prawns</td><td>4.42</td><td>0.28</td><td>0.11</td><td>0.98</td><td>0.00</td><td>0.07</td><td>0.26</td></tr><tr><td>Mussels</td><td>26.26</td><td>1.67</td><td>0.64</td><td>5.81</td><td>0.02</td><td>0.43</td><td>1.56</td></tr><tr><td>Oysters</td><td>19.20</td><td>1.22</td><td>0.47</td><td>4.25</td><td>0.02</td><td>0.31</td><td>1.14</td></tr><tr><td rowspan="3">LDPE</td><td>Shrimps and prawns</td><td>4.24</td><td>0.27</td><td>0.10</td><td>0.94</td><td>0.00</td><td>0.07</td><td>0.25</td></tr><tr><td>Mussels</td><td>25.18</td><td>1.61</td><td>0.62</td><td>5.57</td><td>0.02</td><td>0.41</td><td>1.50</td></tr><tr><td>Oysters</td><td>18.41</td><td>1.17</td><td>0.45</td><td>4.07</td><td>0.02</td><td>0.30</td><td>1.09</td></tr><tr><td rowspan="3">PA</td><td>Shrimps and prawns</td><td>5.24</td><td>0.33</td><td>0.13</td><td>1.16</td><td>0.00</td><td>0.09</td><td>0.31</td></tr><tr><td>Mussels</td><td>31.14</td><td>1.98</td><td>0.76</td><td>6.89</td><td>0.03</td><td>0.51</td><td>1.85</td></tr><tr><td>Oysters</td><td>22.76</td><td>1.45</td><td>0.56</td><td>5.03</td><td>0.02</td><td>0.37</td><td>1.35</td></tr><tr><td rowspan="3">PS</td><td>Shrimps and prawns</td><td>5.06</td><td>0.32</td><td>0.12</td><td>1.12</td><td>0.00</td><td>0.08</td><td>0.30</td></tr><tr><td>Mussels</td><td>30.05</td><td>1.92</td><td>0.74</td><td>6.65</td><td>0.02</td><td>0.49</td><td>1.78</td></tr><tr><td>Oysters</td><td>21.97</td><td>1.40</td><td>0.54</td><td>4.86</td><td>0.02</td><td>0.36</td><td>1.30</td></tr><tr><td rowspan="3">PP</td><td>Shrimps and prawns</td><td>4.15</td><td>0.26</td><td>0.10</td><td>0.92</td><td>0.00</td><td>0.07</td><td>0.25</td></tr><tr><td>Mussels</td><td>26.64</td><td>1.57</td><td>0.60</td><td>5.45</td><td>0.02</td><td>0.40</td><td>1.46</td></tr><tr><td>Oysters</td><td>18.01</td><td>1.15</td><td>0.44</td><td>3.98</td><td>0.01</td><td>0.29</td><td>1.07</td></tr><tr><td rowspan="3">PVC</td><td>Shrimps and prawns</td><td>6.61</td><td>0.42</td><td>0.16</td><td>1.46</td><td>0.01</td><td>0.11</td><td>0.39</td></tr><tr><td>Mussels</td><td>39.26</td><td>2.50</td><td>0.96</td><td>8.68</td><td>0.03</td><td>0.64</td><td>2.33</td></tr><tr><td>Oysters</td><td>28.70</td><td>1.83</td><td>0.70</td><td>6.35</td><td>0.02</td><td>0.47</td><td>1.70</td></tr><tr><td></td><td>Clams</td><td>10.37</td><td>0.66</td><td>0.25</td><td>2.29</td><td>0.01</td><td>0.17</td><td>0.62</td></tr></tbody></table>

(continued)
<table><caption>**Table 8.5** (continued)</caption><thead><tr><th>Polymer</th><th>Seafood</th><th>MP-PAH<br>(pg/kg bw/day)</th><th>MP-PCB<br>(pg/kg bw/day)</th><th>MP-DDT<br>(pg/kg bw/day)</th><th>MP-PBDE<br>(pg/kg bw/day)</th><th>MP-HCH<br>(pg/kg bw/day)</th><th>MP-BPA<br>(pg/kg bw/day)</th><th>MP-NP<br>(ng/ kg bw/day)</th></tr></thead><tbody><tr><td rowspan="4">PET</td><td>Shrimps and prawns</td><td>6.38</td><td>0.41</td><td>0.16</td><td>1.41</td><td>0.01</td><td>0.10</td><td>0.38</td></tr><tr><td>Mussels</td><td>37.90</td><td>2.42</td><td>0.93</td><td>8.38</td><td>0.03</td><td>0.62</td><td>2.25</td></tr><tr><td>Oysters</td><td>27.71</td><td>1.77</td><td>0.68</td><td>6.13</td><td>0.02</td><td>0.45</td><td>1.65</td></tr><tr><td>Clams</td><td>10.01</td><td>0.64</td><td>0.25</td><td>2.21</td><td>0.01</td><td>0.16</td><td>0.59</td></tr></tbody></table>

<table><caption>**Table 8.6** Overall dietary exposure to MP additives and associated sorbed pollutants in Finland and China (age class: adults and elderly) resulting from shellfish consumption. Data on food consumption were taken from WHO CIFOCOss. Results are presented as overall exposure concentrations resulting from the combination of the four shellfish groups (mussels, oysters, clams, and shrimps and prawns), categorized by polymer type</caption><thead><tr><th>Polymer</th><th>Country</th><th>MP-PAH<br>(pg/kg bw/day)</th><th>MP-PCB<br>(pg/kg bw/day)</th><th>MP-DDT<br>(pg/kg bw/day)</th><th>MP-PBDE<br>(pg/kg bw/day)</th><th>MP-HCH<br>(pg/kg bw/day)</th><th>MP-BPA<br>(pg/kg bw/day)</th><th>MP-NP<br>(ng/ kg bw/day)</th></tr></thead><tbody><tr><td rowspan="2">HDPE</td><td>China</td><td>41.52</td><td>2.65</td><td>1.02</td><td>9.18</td><td>0.03</td><td>0.68</td><td>2.47</td></tr><tr><td>Finland</td><td>22.43</td><td>1.43</td><td>0.55</td><td>4.96</td><td>0.02</td><td>0.37</td><td>1.33</td></tr><tr><td rowspan="2">LDPE</td><td>China</td><td>39.81</td><td>2.54</td><td>0.98</td><td>8.80</td><td>0.03</td><td>0.65</td><td>2.36</td></tr><tr><td>Finland</td><td>21.51</td><td>1.37</td><td>0.53</td><td>4.76</td><td>0.02</td><td>0.35</td><td>1.28</td></tr><tr><td rowspan="2">PA</td><td>China</td><td>49.22</td><td>3.14</td><td>1.21</td><td>10.89</td><td>0.04</td><td>0.80</td><td>2.92</td></tr><tr><td>Finland</td><td>26.60</td><td>1.70</td><td>0.65</td><td>5.88</td><td>0.02</td><td>0.43</td><td>1.58</td></tr><tr><td rowspan="2">PS</td><td>China</td><td>47.51</td><td>3.03</td><td>1.17</td><td>10.51</td><td>0.04</td><td>0.77</td><td>2.82</td></tr><tr><td>Finland</td><td>25.67</td><td>1.64</td><td>0.63</td><td>5.68</td><td>0.02</td><td>0.42</td><td>1.52</td></tr><tr><td rowspan="2">PP</td><td>China</td><td>38.95</td><td>2.48</td><td>0.96</td><td>8.62</td><td>0.03</td><td>0.63</td><td>2.31</td></tr><tr><td>Finland</td><td>21.05</td><td>1.34</td><td>0.52</td><td>4.65</td><td>0.02</td><td>0.34</td><td>1.25</td></tr><tr><td rowspan="2">PVC</td><td>China</td><td>62.06</td><td>3.96</td><td>1.52</td><td>13.73</td><td>0.05</td><td>1.01</td><td>3.68</td></tr><tr><td>Finland</td><td>33.53</td><td>2.14</td><td>0.82</td><td>7.42</td><td>0.03</td><td>0.55</td><td>1.99</td></tr><tr><td rowspan="2">PET</td><td>China</td><td>59.92</td><td>3.82</td><td>1.47</td><td>13.25</td><td>0.05</td><td>0.98</td><td>3.56</td></tr><tr><td>Finland</td><td>32.38</td><td>2.06</td><td>0.79</td><td>7.16</td><td>0.03</td><td>0.53</td><td>1.92</td></tr></tbody></table>

lower amount compared to rubbery plastics and especially to PE. Our calculations, in fact, analyse the sorptive capacities of plastic materials by only taking into consideration their densities and not their physico-chemical properties. With respect to this, it should be noted that cooking and food processing can sometimes lead to physicochemical changes in the plastic particles. A recent study has measured a reduction of approximately 14% of microplastics in cooked mussels, with also a
possible size reduction of the particles (Renzi et al. 2018). Moreover, high temperatures could also enhance the release of chemical compounds from microplastics (Bach et al. 2013).

Finally, it is now possible to compare the results of the exposure assessment of associated chemicals from ingested plastics via seafood consumption (Table 8.5; Table 8.6) with the no observed effect levels (NOELs) and no observed adverse effect levels (NOAELs) established by international expert committees such as the Joint FAO/WHO Expert Committee on Food Additives (JECFA) or EFSA (Table 8.7). The purpose of this last step is to check whether the highest load of microplastic-bound pollutants could lead to a significant threat to humans after the consumption of shellfish or not.

The NOEL values are not indicative of the tolerable daily intake (TDI) of contaminants but can provide useful information on the threshold of toxicological concern. In the present case study, the dietary exposures to MP-bound pollutant concentrations were in the order of nanograms (ng/kg bw/day), well below the NOELs set by the literature.

PAHs can be metabolized, resulting in genotoxic effects and carcinogenicity, especially the ones presenting a higher number of aromatic rings (Scientific Committee on Food 2002b). For this reason, no TDI can be estimated. Mean and high daily intake in adults have been estimated to be 4 and 10 ng/kg bw, with children exposure being more than twofolds higher (JECFA 2006a). Benzo(a)pyrene can be used as a marker of PAH contamination in food, but the sum of benzo[a]pyrene and chrysene (PAH2); the sum of benzo[a]pyrene, chrysene, benz[a]anthracene, and benzo[b]fluoranthene (PAH4); and the sum of benzo[a]pyrene, benz[a]anthracene, benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[ghi]perylene, chrysene, dibenz[a,h]anthracene, and indeno[1,2,3-cd]pyrene (PAH8) have been suggested as

<table><caption>**Table 8.7** No observed effect levels (NOELs), benchmark dose lower confidence limit (BMDL), no observed adverse effect levels (NOAELs), allowable daily intake (ADI), and tolerable daily intake (TDI) values established by international authorities</caption><thead><tr><td>Chemical</td><td>NOEL/NOAEL/BMDL</td><td>ADI/TDI</td><td>Source</td></tr></thead><tbody><tr><td>PCBs</td><td>0.04 mg/kg bw/day</td><td>20 ng/kg bw</td><td>Faroon et al. (2003) and JECFA (1990)</td></tr><tr><td>PAHs</td><td>0.17 (PAH2), 0.34 (PAH4),0.49 (PAH8) mg/kg bw</td><td>Not established, carcinogen</td><td>EFSA (2008)</td></tr><tr><td>Benzo(a) pyrene</td><td>100 µg/kg bw</td><td>Not established, carcinogen</td><td>JECFA (2006b)</td></tr><tr><td>γ-HCH</td><td>0.47 mg/kg bw/day</td><td>0–0.005 mg/kg bw</td><td>JECFA (2002)</td></tr><tr><td>DDT</td><td>1 mg/kg bw/day</td><td>0.01 mg/kg bw</td><td>JECFA (2001)</td></tr><tr><td>PBDEs</td><td>Not established</td><td>Not established</td><td>JECFA (2006c)</td></tr><tr><td>BPA</td><td>5 mg/kg bw/day</td><td>4 µg/kg bw</td><td>EFSA (2015) and Scientific Committee on Food (2002a)</td></tr><tr><td>NP</td><td>15 mg/kg/day</td><td>5 µg/kg bw</td><td>Bontje et al. (2004) and Nielsen et al. (2000)</td></tr></tbody></table>
more suitable alternatives, with the last two parameters being the most appropriate markers for genotoxic and carcinogenic PAHs in food commodities (EFSA 2008).

For DDT, the no observed effect level (NOEL) and provisional tolerable daily intake (PTDI) are 1 mg/kg bw/day and 0-0.01 mg/kg bw/day, respectively, with the highest observed average intake of 0.68 mg/man/day (JECFA 1961).

The tolerable daily intake (TDI) temporarily established for BPA in foodstuff was set as 4 µg/kg bw/day (EFSA 2015), and the daily intake in adults in Europe has been estimated up to 1.5 µg/kg bw (Scientific Committee on Food 2002a). As it concerns intake for nonylphenol, the TDI was estimated at 5 µg/kg bw/day by the Danish Institute of Safety and Toxicology (Nielsen et al. 2000). It can be concluded that the levels of microplastic-bound contaminants found in the selected commodities (shrimp, prawns, clams, oyster, and mussels) are below the ones reported in Table 8.7. Based on these estimation values, it could be assumed that the transfer of environmental pollutants and additives mediated by microplastic particles in shell-fish is negligible. In fact, the contribution made by other sources such as the ingestion of contaminated prey items and subsequent transfer of POPs from food to organism supplies most of the contaminant burden (FAO 2017). However, emerging additives from MP should be further explored, and plasticizer additives such as phthalates and organophosphate flame retardants have not yet been investigated.

### 8.5.4 Limitations for Food Safety Risk Assessment

The fate of plastic in the human body and its possible food safety impact are unknown. Although it is thought that the particles below 1.5 µm can penetrate into the capillaries of the organs, while larger particles will be excreted (Yoo et al. 2011), there are many knowledge gaps such as toxicological data of commonly ingested plastics and its compounds.

The best-studied dietary sources of microplastics are fisheries and aquaculture products, which are important food commodities in certain areas. However, the toxicity of most plastic monomers, polymers, and additives present in microplastics has never been evaluated by relevant international expert scientific committees such as the JECFA. International expert committees such as the JECFA are key to evaluate the potential toxicity, considering newly generated scientific data and establishing the basis for the risk analysis exercises (risk assessment, risk management, and risk communication (Fig. 8.1)).

In order to perform a proper exposure and then risk assessment of plastic particles, and plastic as a vector of additives and associated sorbed pollutants, researchers should develop new techniques to better understand toxicity and transfer mechanisms. In addition to that, improvement should be made to detect smaller plastic fragments, especially those in the range of the nanoparticles, which are not much studied. These are, in fact, the ones that could mostly enhance negative consequences on the organism both because of their ability to cross biological barriers and their higher sorption capacity.
<figure><img src="image_43.png" /><figcaption>Fig. 8.1 FAO/WHO risk analysis framework. (Adapted from FAO 2017)</figcaption></figure>

## 8.6 Research Gaps

Although most microplastics are found in the gastrointestinal tract of aquatic animals, there is a limited understanding of the presence of microplastics and specially nanoplastics in edible parts of fisheries and aquaculture products. This mainly occurs because there are no standard methodologies to analyse plastic particles in foods. While many researchers are developing analytical methods, there are still research gaps:

* Standard tissue digestion and polymer identification and quantification protocols should be determined.

* Future research should focus more on nanoplastics, which may also be the ones eliciting the highest exposure due to their size.

* As some particles have been found in the edible tissues of fish, future studies should also include muscle analysis.

* More studies should focus their attention on other food commodities that could be contaminated by microplastics, such as table salt or seaweed.

* More investigations on additive leaching and contaminant desorption processes under gut conditions should be carried out, to understand the exposure to xenobiotic chemicals through plastic ingestion.

Data that would allow a food safety risk assessment is limited and the consequences of microplastic exposure through diet are poorly understood. Most studies have investigated the effects of the inhalation of plastic particles, and very few of them analysed the consequences of a dietary uptake. As it concerns the risk assessment of microplastics, the following research gaps were identified:
* Chronic exposure to microplastics and nanoplastics in humans via ingestion, inhalation and, skin contact should be carried out.

* Environmentally relevant concentrations of plastics, their additives, and sorbed pollutants should be used in dose-response assays for a better understanding of their toxicity.

* The exposure to microorganisms present on plastic debris should be evaluated, as some pathogens might also be found on them.

* Consequences of the dietary intake of different concentrations of plastic additives, monomers, and contaminants should be clarified, as many of these chemicals might trigger response mechanisms even at low doses.

* Changes in the structure and chemical composition of microplastics and nanoplastics in foods while processing or cooking, as well as their interaction with the food matrix, should be further studied.

* Once the consequences of microplastic exposure through diets are clarified, mechanisms of control should be put in place.

## 8.7 Conclusions

Micro- and nanoscale plastic particles are widely distributed in the aquatic environment, and aquatic animals are exposed to them, which results in the presence of microplastics and nanoplastics in fisheries and aquaculture commodities.

Most microplastics and nanoplastics are found in the gastrointestinal tract of aquatic species, and evisceration could lead to a substantial decrease in the exposure to plastics in the final consumer. Nevertheless, there are fisheries and aquaculture commodities such as small fish, crustaceans, and bivalve molluscs that are commonly consumed whole, so the exposure through these products is higher.

There is evidence of a sorptive behaviour of microplastics and nanoplastics, which has been observed both in the laboratory and in the field. Because of their hydrophobicity and high surface/volume, plastic polymers can concentrate organic contaminants by several folds compared to the water column and might also host biofilm-forming microorganisms on their surface.

Microplastic contaminants and additives added during the manufacturing process to add specific features of the final product have been found to have adverse effects on humans and animals, mostly related to carcinogenicity and reproductive toxicity.

Data on human dietary intake of microplastics are very scarce. An exposure assessment was carried out to evaluate human exposure to environmental contaminants and plastic additives through microplastic-contaminated seafood. The worst-case estimate of the exposure to contaminants through seafood indicates the contribution of microplastic contaminants, and additives through fisheries and aquaculture products are negligible compared to other sources. Besides, the exposure scenario assumed the complete release of contaminants and additives once in the
gut, which is unlikely. Moreover, even in this case, the levels of xenobiotics were
several orders of magnitude lower than threshold values (NOAEL) indicated by
international authorities.

Microplastic dietary intake for other food commodities such as salt or water should be carried out to understand the overall exposure through food and open the possibility for food safety risk assessment.

Literature on the actual effect of microplastics and nanoplastics in humans is extremely poor, and more research is needed to understand the toxicity of the most common polymers and plastic additives, as well as their mixture.

8.8 Glossary

8.8.1 *Microplastics and Nanoplastics Definition*

Recently, many studies have focused their attention on the possible harm caused by micro- and nano-sized plastic particles on human and animal health. The size of the particles seems to be one of the main aspects to possibly pose a food safety threat.

There is an ongoing debate about how to define micro- and nanoplastic particles. One of the most accepted definition describes them as plastic particles consisting of a heterogeneous mixture of different shaped materials in the range from 0.1 µm to 5000 µm in their longest dimensions (EFSA 2016; Lusher et al. 2017), while nano-plastics are defined as plastic particles whose size ranges from 0.001 µm to 0.1 µm (Klaine et al. 2012).

Plastics can also be classified according to the process that generated them. Primary microplastics are intentionally produced (e.g. plastic and/or cosmetics manufacture), while secondary microplastics are the result of fragmentation of larger materials are discharged into the environment (GESAMP 2015).

8.8.2 *Microplastics and Nanoplastics Composition*

Depending on the intended use, polymers with different physical and chemical
properties can be mixed. Additionally, additives such as plasticizers, flame retar-
dants, colourants, or antioxidants are normally included in various percentages to
improve their performance. When plastics reach the environment, they can also sorb
and accumulate many hydrophobic environmental contaminants, being a potential
vector for additives and sorbed contaminants to the organisms. Among the sorbed
hydrophobic pollutants are polychlorinated biphenyls (PCBs), polycyclic aromatic
hydrocarbons (PAHs), and chlorinated pesticides, all of them belonging to the group
of persistent organic pollutants (POPs), as they are persistent, bioaccumulative, and
toxic (PBT) substances. Trace metals and microorganisms such as pathogenic bacteria or viruses might also sorb on microplastics (FAO 2017; GESAMP 2016).

### 8.8.2.1 Monomers and Polymers

Monomers such as ethylene, propylene, and styrene are the building blocks of polymers that lead to the production of a variety of materials. The most common polymers are acrylonitrile butadiene styrene (ABS), acrylic, epoxy resin, expanded polystyrene (EPS), polyethylene high density (HDPE), polyethylene low density (LDPE), polycarbonate, polycaprolactone, polyethylene (PE), polyethylene terephthalate (PET), poly (glycolic) acid, poly(lactide), poly(methyl methacrylate), polypropylene (PP), polystyrene (PS), polyurethane, polyethylene linear low density, polyamide (Nylon) 4, 6, 11, 66 (PA), polyvinyl alcohol, polyvinyl chloride (PVC), styrene-butadiene rubber, and thermoplastic polyurethane (FAO 2017). All these monomers and polymers can be expected to be part of microplastic particles present in the environment and therefore enter different food value chains.

### 8.8.2.2 Flame Retardants

Today, there are more than 175 chemicals classified as flame retardants (FRs) (Alaee et al. 2003). Some of these compounds are commonly added to polymers to reduce their flammability. Polybrominated diphenyl ethers (PBDEs) and hexabromocyclododecanes (HBCDs) are the most utilized brominated FRs in plastic manufacture and are commonly added to polystyrene, polyesters, polyolefins, polyamides, epoxies, and ABS. Some HBCDs and PBDEs are simply blended with the polymers and therefore are more likely to leach out of the products, while others can be incorporated into the polymers (Hutzinger and Thoma 1987), a consequence that poses an environmental and food safety concern. PBDEs and HBCD are listed by the Stockholm Convention as Persistent Organic Pollutants (POPs) and are associated with hepatotoxicity, kidney toxicity, endocrine-disrupting effects, and teratogenicity (Muirhead et al. 2006; Yogui and Sericano 2009). Nowadays, organophosphorus flame retardants (OPFRs) are extensively used as additives in order to replace the brominated ones. These compounds have been seen to induce toxicity in vitro and in vivo up to a certain degree (e.g. oxidative stress, cytotoxicity, endocrine disruption), but information on their accumulation and biomagnification in the food chain is still scarce (reviewed by Du et al. 2019).

### 8.8.2.3 Plasticizers

Substances such as phthalates and BP are used to enhance flexibility and softness and to reduce brittleness. These chemicals are usually added to synthetic polymers used in food packaging such as polycarbonate and epoxy resins but also PE, PP, and/
or PVC (EFSA 2007; FAO/WHO 2009). Both phthalates and BPA have been found to act as endocrine disruptors, causing fertility problems, cardiovascular diseases, development disorders, and reproductive cancers, while the toxicity of other BPs either remains unknown or information is not sufficient (Chen et al. 2016; Ma et al. 2019; Rochester 2013).

### 8.8.2.4 Antioxidants and Stabilizers

Nonylphenols (NPs) are a group of organic compounds belonging to the family of alkylphenols and are extensively used as a stabilizer in food packaging and as antioxidants in polymers such as rubber, vinyl, polyolefins, polystyrenes, and PVC (GESAMP 2016; USEPA 2010). NPs are known to be endocrine disruptors and have been reported to exert synergistic effects following their co-occurrence with other compounds (Soares et al. 2008; Vethaak et al. 2005).

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Chapter 9
# The Microplastic-Antibiotic Resistance Connection

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Check for
updates

Nachiket P. Marathe and Michael S. Bank

**Abstract** Microplastic pollution is a big and rapidly growing environmental problem. Although the direct effects of microplastic pollution are increasingly studied, the indirect effects are hardly investigated, especially in the context of spreading of disease and antibiotic resistance genes, posing an apparent hazard for human health. Microplastic particles provide a hydrophobic surface that provides substrate for attachment of microorganisms and readily supports formation of microbial biofilms. Pathogenic bacteria such as fish pathogens *Aeromonas* spp., *Vibrio* spp., and opportunistic human pathogens like *Escherichia coli* are present in these biofilms. Moreover, some of these pathogens are shown to be multidrug resistant. The presence of microplastics is known to enhance horizontal gene transfer in bacteria and thus, may contribute to dissemination of antibiotic resistance. Microplastics can also adsorb toxic chemicals like antibiotics and heavy metals, which are known to select for antibiotic resistance. Microplastics may, thus, serve as vectors for transport of pathogens and antibiotic resistance genes in the aquatic environment. In this book chapter, we provide background information on microplastic biofouling ("plastispHERE concept"), discuss the relationship between microplastic and antibiotic resistance, and identify knowledge gaps and directions for future research.

**9.1 Introduction**

Microplastic (<5 mm, GESAMP 2019) pollution is a widespread and global environmental problem that is projected to increase in upcoming decades creating significant challenges for its management and prevention (Borrelle et al. 2020; Jambeck

N. P. Marathe (✉)
Institute of Marine Research, Bergen, Norway
e-mail: Nachiket.Marathe@hi.no

M. S. Bank
Institute of Marine Research, Bergen, Norway

University of Massachusetts Amherst, Amherst, MA, USA
e-mail: Michael.Bank@hi.no
et al. 2015). Transport of microplastic from land via headwater streams and large rivers to the ocean (Hurley et al. 2018; Jambeck et al. 2015; van Wijnen et al. 2019) is an important component of the microplastic pollution cycle, and plastic particles can now be found globally throughout all ecosystem components including the atmosphere, terrestrial landscapes, aquatic freshwater and marine environments, and all types of biota including seafood species commonly consumed by humans (Bank and Hansson 2019).

Microplastics represent a novel substrate for marine bacteria including both fish and human pathogens (Dang and Lovell 2016; McCormick et al. 2014; Zettler et al. 2013) and are also a reservoir for metal resistance and antibiotic resistance genes. The role of microplastics in the spread of antibiotic resistance is a relatively new research topic that has garnered significant interest by scientists (Bank et al. 2020; Bowley et al. 2021; Guo et al. 2020; Parthasarathy et al. 2019; Radisic et al. 2020). The indirect effects of microplastics have not been well studied especially in the context of seafood safety and global food security, and these effects may pose a significant hazard for human health regarding the spread of disease (Bank et al. 2020; Guo et al. 2020). The specific objectives of this chapter were to (1) provide background information on microplastic biofouling and describe the concept of the “plastisphere” (Zettler et al. 2013), (2) discuss the relationship of microplastic and antibiotic resistance, and (3) identify knowledge gaps and directions for future research.

## 9.2 The Plastisphere Concept

One of the critical mechanisms of the microplastic antibiotic resistance connection is the “plastisphere” concept. This concept was originally presented in the seminal paper by Zettler et al. (2013) who reported that microbial communities attached to plastic debris were diverse and composed of heterotrophs, autotrophs, predators, and symbionts and were distinct from the surrounding marine waters. These plastic particle surfaces represented a novel substrate and/or ecological habitat within the water column and on the surface of the open ocean (Amaral-Zettler et al. 2015, 2020; Bowley et al. 2021; Oberbeckmann et al. 2018; Wright et al. 2020; Zettler et al. 2013). Microplastic particles have hydrophobic surfaces, with no net charge, upon entering the ocean as virgin artificial materials; however, they can quickly become colonized by microbial biofilms (Bowley et al. 2021; Wright et al. 2020; Zettler et al. 2013). The development of this concept was important for forming scientific questions regarding the overall direct and indirect impacts of microplastic pollution primarily because of the long residence time of microplastic in the environment and the potential for long-range transport and the associated risks of transfer of pathogens and disease (Bowley et al. 2021). Pathogenic microbes such as *Vibrio* spp. have been reported in high abundance within the plastisphere (Amaral-Zettler et al. 2020; Bowley et al. 2021; Zettler et al. 2013; Zhang et al. 2020) and although not all *vibrios* are pathogenic, they often prefer lower salinity found in
coastal and estuarine areas, thus highlighting the importance of the plastisphere regarding its distribution, abundance, fate, and transport (Bowley et al. 2021; Thompson et al. 2004). These identified risks and the processes related to microplastic and microbe interactions are complex and are influenced by ocean currents (Hale et al. 2020), sources, fate and transport dynamics, trophic transfer and food web complexity, horizontal gene transfer and attachment properties (Arias-Andres et al. 2019), buoyancy and sinking properties of microplastics, variation, and uptake by farmed (Sun et al. 2020a) and wild seafood taxa, leading to subsequent human exposures (Bowley et al. 2021; Zhou et al. 2020).

## 9.3 Antibiotic Resistance

The introduction of antibiotics for the treatment of infectious disease is one of the most important advances in healthcare. The global spread of resistance in bacteria, particularly in human pathogens, presents major challenges for treatment and preventing the spread of infections (Ventola 2015). Annually, in the European Union/ European Economic Area, an estimated more than 33,000 deaths and more than 800,000 cases of “impacted life-years” are attributable to infections caused by antibiotic-resistant pathogens, with direct and indirect estimated costs of more than 1.5 B€ (Cassini et al. 2019). The World Health Organization (WHO) has predicted the advent of infectious diseases for which no antibiotic treatment will be available (WHO 2019).

Antibiotic resistance is a natural phenomenon. Misuse and over use of antibiotics has led to the development, selection, and global spread of antibiotic resistance (Roberts and Zembower 2020). Selection pressure from the presence of antibiotics or other antimicrobial compounds like heavy metals and biocides leads to the enrichment of antibiotic-resistant bacteria and antibiotic resistance genes (ARGs) in bacteria from humans, animals, and the environment (Francino 2016; Gullberg et al. 2014; Marathe et al. 2013; Seiler and Berendonk 2012). Horizontal gene transfer (HGT) is a fundamental force driving bacterial evolution and contributes to the dis simulation of ARGs in both clinical and environmental bacteria (Boto 2010; Emamalipour et al. 2020; Jain et al. 2003). Antimicrobial compounds like antibiot ics, biocides, and heavy metals can drive the development of antibiotic resistance and stimulate horizontal transfer of antibiotic resistance genes (Andersson and Hughes 2014; Jutkina et al. 2018; Zhang et al. 2018), thus aiding selection and dis semination of antibiotic resistance.
## 9.4 Microplastics and Antibiotic Resistance

Microorganisms attach themselves to surfaces forming a complex matrix of bio-polymers and microbial cells known as biofilm (Dang and Lovell 2016). Formation of biofilms protect bacteria from unfavorable conditions in the environment (Donlan 2002). Biofilms are ubiquitous in aquatic environments and play an important role in various biological and ecological processes (Guo et al. 2018). Aquatic biofilms serve as a sink for various contaminants, like heavy metals, and antibiotics that are known to select for antibiotic resistance and stimulate horizontal transfer of antibiotic resistance genes (Gullberg et al. 2014; Guo et al. 2018; Jutkina et al. 2018; Richard et al. 2019). Accordingly, antibiotic resistance genes have been detected in natural aquatic biofilms (Balcázar et al. 2015; Guo et al. 2018).

Microplastic particles provide a hydrophobic surface that readily supports formation of microbial biofilms, where environmental conditions are the main drivers of biofilm formation (Oberbeckmann et al. 2018; Rummel et al. 2017). Pathogenic bacteria such as fish pathogens *Aeromonas* spp., *Vibrio* spp., and opportunistic human pathogens like *E. coli* can invariably be present in these biofilms (Kirstein et al. 2016; Silva et al. 2019; Viršek et al. 2017). Microplastics can selectively enrich both antibiotics and antibiotic-resistant bacteria on their surfaces in landfill leachates, freshwater, as well as in sea water (Su et al. 2020; Sun et al. 2020b; Wang et al. 2020; Wu et al. 2019). Thus, microplastics may serve as a vector for transport of pathogens in the aquatic environment.

Several methods have been used for detecting and quantifying ARGs associated with marine plastics including selective isolation of resistant bacteria and phenotypic antibiotic sensitivity testing, whole genome sequencing, shotgun metagenomics, and quantitative polymerase chain reaction (qPCR). Culture-based methods involving isolation of bacteria on a culture media followed by antibiotic sensitivity testing is a traditional approach used for studying antibiotic resistance (Khan et al. 2019). Zhang et al. (2020) carried out isolation and characterization of antibiotic-resistant marine bacteria from microplastic particles collected from marine aquaculture sites using a combination of seven antibiotics and a non-selective media. They showed presence of several multidrug-resistant marine bacteria including pathogenic *Vibrio* species on microplastics (Zhang et al. 2020). In contrast, other studies carried out selective isolation of pathogens like *Vibrio* spp. (Laverty et al. 2020) and *E. coli* (Song et al. 2020) showing presence of multidrug-resistant pathogens on marine microplastics. Recently, a study reported whole genome sequences (WGS) of antibiotic-resistant fish pathogens isolated from marine plastics (Radisic et al. 2020). With the advent of next-generation sequencing technology, WGS analysis of pathogens has become common and affordable tool for genotyping and epidemiology in clinics (Quainoo et al. 2017). WGS analyses are effective in elucidating the total metabolic potential of microorganisms and understanding the genetic basis of antibiotic resistance (Grevskott et al. 2020; Hendriksen et al. 2019). Although this is true, WGS data on microplastic-associated bacteria is largely lacking.
Only a small proportion of bacteria present in an environment can be cultivated in the lab. This limits detection and quantification of antibiotic resistance genes present in uncultivable bacteria using traditional methods (Lloyd et al. 2018; Stewart 2012). Methods like qPCR analysis or shotgun metagenomics, that use the total genomic DNA extracted from a given sample, partly overcome this limitation. Using qPCR, Wang et al. (2020) showed enrichment of ARGs like *sul1*, *tetA*, *tetC*, *tetX*, and *ermE* on plastic particles in both freshwater and sea water (Wang et al. 2020), while another study showed selective enrichment of *strB*, *bla*<sub>TEM</sub>, *ermB*, *tetM*, and *tetQ* on microplastic particles in landfill leachates (Shi et al. 2020). These studies selected a limited number of ARGs for their analysis. In contrast, using recently developed high-throughput qPCR screening that can analyze more than 200 ARGs, Lu et al. (2020) showed presence of between 34 and 43 different ARGs on the surface of microplastic particles collected from vegetable soil (Lu et al. 2020).

Shotgun metagenomics gives an overview of the total bacteria and associated genes present in a given sample (Simon and Daniel 2011). Using this method Yang et al. (2019) found a total of 64 ARG subtypes that provide resistance against 13 different classes of antibiotics on macroplastics and microplastics collected from the North Pacific Gyre. Along with enrichment of ARGs, the study also found enrichment of metal resistance genes on microplastics (Yang et al. 2019). This study and several of the earlier described studies show presence of clinically important ARGs, like *sul1*, *tetA*, *tetC*, *tetX*, *ermE*, *aac(3)*, *macB*, and *bla*<sub>TEM</sub>, that are invariably found in human pathogens, on microplastic particles (Alcock et al. 2020), thus suggesting that microplastics in the environment act as reservoirs for clinically important antibiotic resistance genes.

Microplastics originate from a variety of processes and invariably ends up in the marine environment via streams and large rivers (Hurley et al. 2018; Jambeck et al. 2015). High levels of microplastics reach the wastewater treatment plants (WWTP) (Dris et al. 2015). Although most of the microplastics are removed during primary and secondary waste treatment, smaller microplastics may still be present in the treated effluents (Talvitie et al. 2017). Treated effluents have low concentrations of microplastic particles but the high volume of effluents released may leads to considerable contamination of the aquatic ecosystem (Murphy et al. 2016; Talvitie et al. 2017). WWTPs receive municipal and/or hospital waste which invariably contains both human pathogens and clinically important antibiotic resistance genes (Le et al. 2016; Marathe et al. 2017, 2018, 2019; Rizzo et al. 2013). Treated effluents from waste water treatment plants are recognized as one of the major sources of environmental pollution with antibiotic resistance genes and resistant pathogens (Alexander et al. 2020; Czekalski et al. 2014; Karkman et al. 2019). The presence of microplastic particles in waste water effluents, thus, presents opportunities for antibiotic-resistant pathogens to colonize and form biofilms on plastic particles. This may lead to further dissemination of resistance in the marine environment via microplastics. Although this is true, there is limited knowledge on the impact of microplastics from treated effluents from WWTP on dissemination of ARGs in the aquatic environment.
Microplastic particles adsorb several chemicals like antibiotics, biocides, and heavy metals (Chen et al. 2020; Godoy et al. 2019; Mammo et al. 2020; Wang et al. 2020). The presence of antibiotics and active metabolites from such agents in the environment leads to selection of multidrug resistance among both clinical and environmental bacteria. Similarly, biocides and heavy metals like copper and mercury are known to co-select for antibiotic resistance (Francino 2016; Gullberg et al. 2011, 2014; Imran et al. 2019; Marathe et al. 2013; Seiler and Berendonk 2012). Adsorption of these chemicals on plastic surfaces containing microbial biofilm may lead to selection pressure in the plastisphere, resulting in active selection of antibiotic resistance on microplastic surfaces. In accordance, Imran et al. (2019) has concluded that co-contamination with microplastics and heavy metals results in development and spread of multiple drug-resistant human pathogens through co-selection mechanisms (Imran et al. 2019). Studies have shown that very low levels of antibiotics and biocides not only can select for antibiotic resistance but also can induce horizontal transfer of ARGs (Gullberg et al. 2011; Jutkina et al. 2018; Zhang et al. 2018). Moreover, bacteria in biofilms are more efficient in horizontal gene transfer compared to planktonic bacteria (Abe et al. 2020). Accordingly, studies have shown increased horizontal gene transfer in presence of microplastics via conjugation (Arias-Andres et al. 2018, 2019). Although extensive research on selection of resistance and promotion of horizontal gene transfer by antimicrobial compounds has been carried out, there is limited knowledge on the effect of adsorbed chemicals on plastisphere bacteria, especially, with reference to selection and transfer of antibiotic resistance genes on microplastic particles.

## 9.5 Conclusions and Directions for Future Research

Microplastics are emerging pollutants that have been detected in a range of environments. With the current trend of plastic consumption and its global production, the environmental pollution and related environmental effects caused by microplastics are expected to increase (Borrelle et al. 2020). Microplastics provide surfaces for the microorganisms to form biofilms (plastisphere) (Zettler et al. 2013). The processes and mechanisms involved in biofilm formation on microplastics are largely unclear. In-depth studies on deciphering the succession of microbes and understanding the effect of different factors that may influence biofilm formation on microplastic particles, such as the environmental conditions and the age of microplastic particles are needed (Su et al. 2020; Yang et al. 2020). Moreover, there are a limited number of studies reporting WGS of bacteria associated with microplastics (Li et al. 2019; Radisic et al. 2021). Bacteria associated with microplastics may play different ecological roles and could also be useful for bioremediation (Debroas et al. 2017). Hence, understanding the metabolic potential of bacteria in plastisphere using WGS is necessary.

Studies have investigated the composition of biofilms on microplastics and shown presence of both fish and human pathogens as well as clinically important
antibiotic resistance genes (Dong et al. 2021). However, the risks associated with
presence of pathogens in terms of human or fish exposure and the ability of
microplastic-associated pathogens for causing infections is not fully understood.
In-depth risk assessment studies on the effect of pathogen carrying microplastics on
fish and human health are thus warranted. Microplastics originating in different
compartments like WWTPs or aquaculture sites may carry different microbiota.
WWTPs and aquaculture sites usually have presence of both antibiotic-resistant
pathogens and microplastics (Cabello et al. 2016; Rodriguez-Mozaz et al. 2015).
There is invariably selection pressure due to presence of antibiotics or biocides
along with presence of resistant bacteria in these sites (Cabello et al. 2016; Edo
et al. 2020; Yang et al. 2014). These environments could play an important role in
enrichment and dispersal of pathogenic bacteria and ARGs to the marine ecosystem
via microplastics. Although microplastics have been shown to increase HGT (Arias-
Andres et al. 2018, 2019), the impact of microplastics on evolution and dissemina-
tion of antibiotic resistance in pathogens and environmental bacteria is largely
unknown. In order to understand the indirect effects of microplastics, the relation-
ship and interactions between microplastics and ARGs, as well as the impact of
their association on aquatic environment especially on marine environment and sea
food safety, needs to be further assessed. Holistic multidisciplinary studies on fate,
migration, and potential environmental risks posed by microplastics through dis-
semination and evolution of antibiotic resistance are needed in the future, for better
understanding the indirect effects of microplastic pollution.

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