Huerta Lwanga E, Gertsen H, Gooren H, Peters P, Salánki T, van der Ploeg M et al (2017) Incorporation of microplastics from litter into burrows of Lumbricus terrestris. Environ Pollut 220:523-531. https://doi.org/10.1016/j.envpol.2016.09.096

Hurley RR, Nizzetto L (2018) Fate and occurrence of micro(nano)plastics in soils: knowledge gaps and possible risks. Curr Opin Environ Sci Health 1:6-11. https://doi.org/10.1016/j. coesh.2017.10.006

Hurley RR, Woodward JC, Rothwell JJ (2017) Ingestion of microplastics by freshwater Tubifex worms. Environ Sci Technol 51(21):12844-12851. https://doi.org/10.1021/acs.est.7b03567

Hurley RR, Woodward JC, Rothwell JJ (2018) Microplastic contamination of river beds significantly reduced by catchment-wide flooding. Nat Geosci. https://doi.org/10.1038/s41561-018-0080-1

Imhof HK, Ivleva NP, Schmid J, Niessner R, Laforsch C (2013) Contamination of beach sediments of a subalpine lake with microplastic particles. Curr Biol 23(19):R867-R868. https://doi.org/10.1016/j.cub.2013.09.001

Imhof HK, Wiesheu AC, Anger PM, Niessner R, Ivleva NP, Laforsch C (2018) Variation in plastic abundance at different lake beach zones-A case study. Sci Total Environ 613:530-537

Isobe A, Buenaventura NT, Chastain S, Chavanich S, Cózar A, DeLorenzo M et al (2019) An interlaboratory comparison exercise for the determination of microplastics in standard sample bottles. Mar Pollut Bull 146:831-837. https://doi.org/10.1016/j.marpolbul.2019.07.033

Jabeen K, Su L, Li J, Yang D, Tong C, Mu J, Shi H (2017) Microplastics and mesoplastics in fish from coastal and fresh waters of China. Environ Pollut 221:141-149. https://doi.org/10.1016/j. envpol.2016.11.055

Jambeck JR, Geyer R, Wilcox C, Siegler TR, Perryman M, Andrady A et al (2015) Plastic waste inputs from land into the ocean. Science 347(6223):768-771. https://doi.org/10.1126/science.1260352

Jiang C, Yin L, Li Z, Wen X, Luo X, Hu S et al (2019) Microplastic pollution in the rivers of the Tibet Plateau. Environ Pollut 249:91-98. https://doi.org/10.1016/j.envpol.2019.03.022

Johnsen JP, Bye NH (2019) Assessment of tire wear emission in a road tunnel, using benzothiazoles as tracer in tunnel wash water 79. Retrieved from https://nmbu.brage.unit.no/nmbu-xmlui/handle/11250/2612289

Kamata H, Masuda K, Yamada J, Takada H (2000) Water-Particle distribution of hydrophobic micro pollutants in storm water runoff. Polycycl Aromat Compd 20(1-4):39-54. https://doi.org/10.1080/10406630008034774

Kay P, Hiscoe R, Moberley I, Bajic L, McKenna N (2018) Wastewater treatment plants as a source of microplastics in river catchments. Environ Sci Pollut Res 25(20):20264-20267. https://doi.org/10.1007/s11356-018-2070-7

Keller AS, Jimenez-Martinez J, Mitrano DM (2020) Transport of nano- and Microplastic through unsaturated porous media from sewage sludge application. Environ Sci Technol 54(2):911-920. https://doi.org/10.1021/acs.est.9b06483

Khan FR, Halle LL, Palmqvist A (2019) Acute and long-term toxicity of micronized car tire wear particles to Hyalella azteca. Aquat Toxicol 213:105216. https://doi.org/10.1016/j.aquatox.2019.05.018

Klein M, Fischer EK (2019) Microplastic abundance in atmospheric deposition within the Metropolitan area of Hamburg, Germany. Sci Total Environ 685:96-103. https://doi.org/10.1016/j.scitotenv.2019.05.405

Klöckner P, Reemtsma T, Eisentraut P, Braun U, Ruhl AS, Wagner S (2019) Tire and road wear particles in road environment – Quantification and assessment of particle dynamics by Zn determination after density separation. Chemosphere 222:714-721. https://doi.org/10.1016/j.chemosphere.2019.01.176

Koelmans AA, Mohamed Nor NH, Hermsen E, Kooi M, Mintenig SM, De France J (2019) Microplastics in freshwaters and drinking water: critical review and assessment of data quality. Water Res 155:410-422. https://doi.org/10.1016/j.watres.2019.02.054

Kole PJ, Löhr AJ, Van Belleghem FGAJ, Ragas AMJ (2017) Wear and tear of tyres: a stealthy source of microplastics in the environment. Int J Environ Res Public Health 14(10):1265. https://doi.org/10.3390/ijerph14101265
Kum V, Sharp A, Harnpornchai N (2005) Improving the solid waste management in Phnom Penh city: a strategic approach. Waste Manag 25(1):101–109. https://doi.org/10.1016/j.wasman.2004.09.004

Lambert S, Wagner M (2018) Microplastics are contaminants of emerging concern in freshwater environments: an overview. In: Freshwater microplastics. Springer, Cham, pp 1–23. https://doi.org/10.1007/978-3-319-61615-5_1

Lassen, C., Hansen, S. F., Magnusson, K., Hartmann, N. B., Jensen, P. R., Nielsen, T. G., & Brinch, A. (2015). Microplastics: occurrence, effects and sources of releases to the environment in Denmark. Danish Environmental Protection Agency. Retrieved from https://orbit.dtu.dk/en/publications/microplastics-occurrence-effects-and-sources-of-releases-to-the-e

Lebreton LCM, van der Zwet J, Damsteeg J-W, Slat B, Andrady A, Reisser J (2017) River plastic emissions to the world's oceans. Nat Commun 8:15611. https://doi.org/10.1038/ncomms15611

Lecce SA, Pavlowsky RT (1997) Storage of mining-related zinc in floodplain sediments, Blue River, Wisconsin. Phys Geogr 18(5):424–439. https://doi.org/10.1080/02723646.1997.10642628

Lee J, Hong S, Song YK, Hong SH, Jang YC, Jang M et al (2013) Relationships among the abundances of plastic debris in different size classes on beaches in South Korea. Mar Pollut Bull 77(1–2):349–354. https://doi.org/10.1016/j.marpolbul.2013.08.013

Lenaker PL, Baldwin AK, Corsi SR, Mason SA, Reneau PC, Scott JW (2019) Vertical distribution of microplastics in the water column and surficial sediment from the Milwaukee River Basin to Lake Michigan. Environ Sci Technol. https://doi.org/10.1021/acs.est.9b03850

Li X, Chen L, Mei Q, Dong B, Dai X, Ding G, Zeng EY (2018) Microplastics in sewage sludge from the wastewater treatment plants in China. Water Res 142:75–85. https://doi.org/10.1016/j.watres.2018.05.034

Li C, Busquets R, Campos LC (2020) Assessment of microplastics in freshwater systems: a review. Sci Total Environ 707:135578. https://doi.org/10.1016/j.scitotenv.2019.135578

Liu EK, He WQ, Yan CR (2014) ‘White revolution’ to ‘white pollution’—agricultural plastic film mulch in China. Environ Res Lett 9(9):091001. https://doi.org/10.1088/1748-9326/9/9/091001

Liu M, Lu S, Song Y, Lei L, Hu J, Lv W et al (2018) Microplastic and mesoplastic pollution in farmland soils in suburbs of Shanghai, China. Environ Pollut 242:855–862. https://doi.org/10.1016/j.envpol.2018.07.051

Liu F, Vianello A, Vollertsen J (2019a) Retention of microplastics in sediments of urban and highway stormwater retention ponds. Environ Pollut 113335. https://doi.org/10.1016/j.envpol.2019.113335

Liu K, Wu T, Wang X, Song Z, Zong C, Wei N, Li D (2019b) Consistent transport of terrestrial microplastics to the ocean through atmosphere. Environ Sci Technol 53(18):10612–10619. https://doi.org/10.1021/acs.est.9b03427

Liu S, Jian M, Zhou L, Li W (2019c) Distribution and characteristics of microplastics in the sediments of Poyang Lake, China. Water Sci Technol 79(10):1868–1877. https://doi.org/10.2166/wst.2019.185

Liu X, Yuan W, Di M, Li Z, Wang J (2019d) Transfer and fate of microplastics during the conventional activated sludge process in one wastewater treatment plant of China. Chem Eng J 362:176–182. https://doi.org/10.1016/j.cej.2019.01.033

Liu K, Wang X, Song Z, Wei N, Li D (2020a) Terrestrial plants as a potential temporary sink of atmospheric microplastics during transport. Sci Total Environ 742:140523. https://doi.org/10.1016/j.scitotenv.2020.140523

Liu K, Wang X, Song Z, Wei N, Ye H, Cong, X. ... Li, D. (2020b) Global inventory of atmospheric fibrous microplastics input into the ocean: an implication from the indoor origin. J Hazard Mater 400:123223. https://doi.org/10.1016/j.jhazmat.2020.123223

Lorenz K, Kandeler E (2005) Biochemical characterization of urban soil profiles from Stuttgart, Germany. Soil Biol Biochem 37(7):1373–1385. https://doi.org/10.1016/j.soilbio.2004.12.009

Lusher AL, Hurley RR, Vogelsang C, Nizzetto L, Olsen M (2018) Mapping microplastics in sludge, No. M907, p 55. Retrieved from http://www.miljodirektoratet.no/Documents/publikasjoner/M907/M907.pdf
Lusher AL, Hurley RR, Vogelsang C (2019) Microplastics in sewage sludge: captured but released? In: Karapanagioti HK, Kalavrouziotis IK (eds) Microplastics in water and wastewater. IWA Publishing, London, pp 85–100

Lv W, Zhou W, Lu S, Huang W, Yuan Q, Tian M et al (2019a) Microplastic pollution in rice-fish co-culture system: a report of three farmland stations in Shanghai, China. Sci Total Environ 652:1209–1218. https://doi.org/10.1016/j.scitotenv.2018.10.321

Lv X, Dong Q, Zuo Z, Liu Y, Huang X, Wu W-M (2019b) Microplastics in a municipal wastewater treatment plant: fate, dynamic distribution, removal efficiencies, and control strategies. Retrieved from https://pubag.nal.usda.gov/catalog/6367168

Maaß S, Daphi D, Lehmann A, Rillig MC (2017) Transport of microplastics by two collembolan species. Environ Pollut 225:456–459. https://doi.org/10.1016/j.envpol.2017.03.009

Magnusson K, Norén F (2014) Screening of microplastic particles in and down-stream a wastewater treatment plant, No. C 55, p 22. Retrieved from http://www.diva-portal.org/smash/record.jsf?pid=diva2:773505

Magnusson K, Eliasson K, Fråne A, Haikonen K, Hultén J, Olshammar M et al (2016) Swedish sources and pathways for microplastics to the marine environment. A Review of Existing Data IVL, C, 183. Retrieved from https://www.naturvardsverket.se/upload/miljoarbete-i-samhallet/miljoarbete-i-sverige/regeringssuppdrag/2016/mikroplaster/swedish-sources-and-pathways-for-microplastics-to-marine%20environment-ivl-c183.pdf

Mahon AM, O'Connell B, Healy MG, O'Connor I, Officer R, Nash R, Morrison L (2017) Microplastics in sewage sludge: effects of treatment. Environ Sci Technol 51(2):810–818. https://doi.org/10.1021/acs.est.6b04048

Mai L, Bao L-J, Wong CS, Zeng EY (2018) Microplastics in the terrestrial environment. In: Zeng EY (ed) Microplastic contamination in aquatic environments; an emerging matter of environmental urgency. Elsevier, pp 365–378. https://doi.org/10.1016/B978-0-12-813747-5.00012-6

Mai L, You S-N, He H, Bao L-J, Liu L, Zeng EY (2019) Riverine microplastic pollution in the Pearl River Delta, China: are modeled estimates accurate? Environ Sci Technol. https://doi.org/10.1021/acs.est.9b04838

Mani T, Hauk A, Walter U, Burkhardt-Holm P (2015) Microplastics profile along the Rhine River. Sci Rep 5:17988. https://doi.org/10.1038/srep17988

Mantecca P, Gualtieri M, Andrioletti M, Bacchetta R, Vismara C, Vailati G, Camatini M (2007) Tire debris organic extract affects Xenopus development. Environ Int 33(5):642–648. https://doi.org/10.1016/j.envint.2007.01.007

Marais M, Armitage N, Pithey S (2001) A study of the litter loadings in urban drainage systems – methodology and objectives. Water Sci Technol 44(6):99–108. https://doi.org/10.2166/wst.2001.0350

Marais M, Armitage N, Wise C (2004) The measurement and reduction of urban litter entering stormwater drainage systems: paper I – quantifying the problem using the City of Cape Town as a case study [Text]. Retrieved February 24, 2020, from https://www.ingentaconnect.com/content/sabinet/waters/2004/00000030/00000004/art00006

Mason SA, Garneau D, Sutton R, Chu Y, Ehmann K, Barnes J et al (2016a) Microplastic pollution is widely detected in US municipal wastewater treatment plant effluent. Environ Pollut 218:1045–1054. https://doi.org/10.1016/j.envpol.2016.08.056

Mason SA, Kammin L, Eriksen M, Aleid G, Wilson S, Box C, Williamson N, Riley A (2016b) Pelagic plastic pollution within the surface waters of Lake Michigan, USA. J Great Lakes Res 42(4):753–759. https://doi.org/10.1016/j.jglr.2016.05.009

McCormick AR, Hoellein TJ, London MG, Hittie J, Scott JW, Kelly JJ (2016) Microplastic in surface waters of urban rivers: concentration, sources, and associated bacterial assemblages. Ecosphere 7(11):e01556. https://doi.org/10.1002/ecs2.1556

McGoran AR, Clark PF, Morritt D (2017) Presence of microplastic in the digestive tracts of European flounder, Platichthys flesus, and European smelt, Osmerus eperlanus, from the River Thames. Environ Pollut 220:744–751. https://doi.org/10.1016/j.envpol.2016.09.078

Mehlhart G, Blepp M (2012) Study on land-sourced litter (LSL) in the marine environment: review of sources and literature. Öko-Institut. https://www.oeko.de/oekodoc/1487/2012-058-en.pdf
Meland S, Rødland E (2018) Forurensning i tunnelvaskevann –en studie av 34 veitunneler i Norge. VANN 1:54–65. Retrieved from: https://vannforeningen.no/wp-content/uploads/2018/07/Meland.pdf

Meland S, Borgstrøm R, Heier LS, Rosseland BO, Lindholm O, Salbu B (2010a) Chemical and ecological effects of contaminated tunnel wash water runoff to a small Norwegian stream. Sci Total Environ 408(19):4107–4117. https://doi.org/10.1016/j.scitotenv.2010.05.034

Meland S, Heier LS, Salbu B, Tollefsen KE, Farmen E, Rosseland BO (2010b) Exposure of brown trout (Salmo trutta L.) to tunnel wash water runoff—chemical characterisation and biological impact. Sci Total Environ 408(13):2646–2656. https://doi.org/10.1016/j.scitotenv.2010.03.025

Migwi FK, Ogunah JA, Kiratu JM (2020) Occurrence and spatial distribution of microplastics in the surface waters of Lake Naivasha, Kenya. Environ Toxicol Chem n/a(n/a). https://doi.org/10.1002/etc.4677

Muñoz-Cadena CE, Lina-Manjarrez P, Estrada-Izquierdo I, Ramón-Gallegos E (2012) An approach to litter generation and littering practices in a Mexico City neighborhood. Sustainability 4(8):1733–1754. https://doi.org/10.3390/su4081733

Nan B, Su L, Kellar C, Craig NJ, Keough MJ, Pettigrove V (2020) Identification of micro-plastics in surface water and Australian freshwater shrimp Paratyra australiensis in Victoria, Australia. Environ Pollut 259:113865. https://doi.org/10.1016/j.envpol.2019.113865

Nel HA, Dalu T, Wasserman RJ (2018) Sinks and sources: assessing microplastic abundance in river sediment and deposit feeders in an Austral temperate urban river system. Sci Total Environ 612:950–956. https://doi.org/10.1016/j.scitotenv.2017.08.298

Ngo PL, Pramanik BK, Shah K, Roychand R (2019) Pathway, classification and removal efficiency of microplastics in wastewater treatment plants. Environ Pollut 255(2):113326. https://doi.org/10.1016/j.envpol.2019.113326

Nizzetto L, Futter M, Langaas S (2016) Are agricultural Soils dumps for microplastics of urban origin? Environ Sci Technol 50(20):10777–10779. https://doi.org/10.1021/acs.est.6b04140

O'Connor D, Pan S, Shen Z, Song Y, Jin Y, Wu W-M, Hou D (2019) Microplastics undergo accelerated vertical migration in sand soil due to small size and wet-dry cycles. Environ Pollut. https://doi.org/10.1016/j.envpol.2019.03.092

Ockelford A, Cundy A, Ebdon JE (2020) Storm response of fluvial sedimentary microplastics. Sci Rep 10(1):1–10. https://doi.org/10.1038/s41598-020-58765-2

Olesen KB, Stephansen DA, van Alst N, Vollertsen J (2019) Microplastics in a stormwater pond. Water 11(7):1466. https://doi.org/10.3390/w11071466

Panebianco A, Nalbone L, Giarratana F, Ziino G (2019) First discoveries of microplastics in terrestrial snails. Food Control 106:106722. https://doi.org/10.1016/j.foodcont.2019.106722

Panko JM, Kreider ML, McAtee BL, Marwood C (2013) Chronic toxicity of tire and road wear particles to water- and sediment-dwelling organisms. Ecotoxicology 22(1):13–21. https://doi.org/10.1007/s10646-012-0998-9

Phillips MB, Bonner TH (2015) Occurrence and amount of microplastic ingested by fishes in watersheds of the Gulf of Mexico. Mar Pollut Bull 100(1):264–269. https://doi.org/10.1016/j.marpolbul.2015.08.041

Piehl S, Leibner A, Löder MGJ, Dris R, Bogner C, Laforsch C (2018) Identification and quantification of macro- and microplastics on an agricultural farmland. Sci Rep 8(1):17950. https://doi.org/10.1038/s41598-018-36172-y

Qin Y, Wang Z, Li W, Chang X, Yang J, Yang F (2019) Microplastics in the sediment of Lake Ulansuhai of Yellow River Basin, China. Water Environ Res 92(6). https://doi.org/10.1002/wer.1275

Ranneklev SB, Hurley R, Bråte ILN, Vogelsang C (2019) Plast i landbruket: kilder, massebalanse og spredning til lokale vannforekomster (Plastland). Norsk insitutt for vannforskning. Retrieved from https://niva.brage.unit.no/niva-xmlui/handle/11250/2632595

Raupach MR, Lu H (2004) Representation of land-surface processes in aeolian transport models. Environ Model Softw 19(2):93–112. https://doi.org/10.1016/S1364-8152(03)00113-0
Redondo-Hasselerharm PE, de Ruijter VN, Mintenig SM, Verschoor A, Koelmans AA (2018) Ingestion and chronic effects of car tire tread particles on freshwater benthic macroinvertebrates. *Environ Sci Technol* 52(23):13986–13994. https://doi.org/10.1021/acs.est.8b05035

Rezaei M, Riksen MJ, Sirjani E, Sameni A, Geissen V (2019) Wind erosion as a driver for transport of light density microplastics. *Sci Total Environ*. https://doi.org/10.1016/j.scitotenv.2019.02.382

Roch S, Walter T, Ittner LD, Friedrich C, Brinker A (2019) A systematic study of the microplastic burden in freshwater fishes of south-western Germany – Are we searching at the right scale? *Sci Total Environ* 689:1001–1011. https://doi.org/10.1016/j.scitotenv.2019.06.404

Rochman CM (2018) Microplastics research—from sink to source. *Science* 360(6384):28–29. https://doi.org/10.1126/science.aar7734

Rødland E (2019) Ecotoxic potential of road-associated microplastic particles (RAMP). VANN 3:166–183. Retrieved from: https://vannforeningen.no/wp-content/uploads/2019/12/R%C3%B8dland.pdf

Rødland E, Helgadottir D (2018) Prioriteringsverktøy for vurdering av forurensingsbelastning fra urenset tunnelvaskevann. VANN 4:367–376. Retrieved from: https://vannforeningen.no/wp-content/uploads/2019/04/R%C3%B8dland.pdf

Rødland E, Okoffo ED, Rauert C, Heier LS, Lind OC, Reid M, Thomas KV, Meland S (2020) Road de-icing salt: assessment of a potential new source and pathway of microplastics particles from roads. *Sci Total Environ* 738:139352. https://doi.org/10.1016/j.scitotenv.2020.139352

Rodrigues MO, Abrantes N, Gonçalves FJM, Nogueira H, Marques JC, Gonçalves AMM (2018) Spatial and temporal distribution of microplastics in water and sediments of a freshwater system (Antuā River, Portugal). *Sci Total Environ* 633:1549–1559. https://doi.org/10.1016/j.scitotenv.2018.03.233

Ruan Y, Zhang K, Wu C, Wu R, Lam PKS (2019) A preliminary screening of HBCD enantiomers transported by microplastics in wastewater treatment plants. *Sci Total Environ* 674:171–178. https://doi.org/10.1016/j.scitotenv.2019.04.007

Sanchez W, Bender C, Porcher J-M (2014) Wild gudgeons (Gobio gobio) from French rivers are contaminated by microplastics: Preliminary study and first evidence. *Environ Res* 128:98–100. https://doi.org/10.1016/j.envres.2013.11.004

Scarascia-Mugnozza G, Sica C, Russo G (2012) Plastic materials in European agriculture: actual use and perspectives. *J Agric Eng* 42(3):15–28. https://doi.org/10.4081/jae.2011.3.15

Scherer C, Brennholt N, Reifferscheid G, Wagner M (2017) Feeding type and development drive the ingestion of microplastics by freshwater invertebrates. *Sci Rep* 7(1):17006. https://doi.org/10.1038/s41598-017-17191-7

Scherer C, Weber A, Lambert S, Wagner M (2018) Interactions of microplastics with freshwater biota. In: Freshwater microplastics. Springer, Cham, pp 153–180. https://doi.org/10.1007/978-3-319-61615-5_8

Scherer C, Weber A, Stock F, Vurusic S, Egerci H, Kochleus, C. ... Reifferscheid, G. (2020) Comparative assessment of microplastics in water and sediment of a large European river. *Sci Total Environ* 738:139866. https://doi.org/10.1016/j.scitotenv.2020.139866

Scheurer M, Bigalke M (2018) Microplastics in Swiss floodplain soils. *Environ Sci Technol* 52(6):3591–3598. https://doi.org/10.1021/acs.est.7b06003

Schmidt C, Krauth T. Wagner S (2017) Export of plastic debris by rivers into the sea. *Environ Sci Technol* 51(21):12246–12253. https://doi.org/10.1021/acs.est.7b02368

Scopetani C, Chelazzi D, Cincinelli A, Esterhuizen-Londt M (2019) Assessment of microplastic pollution: occurrence and characterisation in Vesijärvi lake and Pikku Vesijärvi pond, Finland. *Environ Monit Assess* 191(11):652. https://doi.org/10.1007/s10661-019-7843-z

Shiber JG (1979) Plastic pellets on the coast of Lebanon. *Mar Pollut Bull* 10(1):28–30. https://doi.org/10.1016/0025-326X(79)90321-7

Shruti VC, Jonathan MP, Rodriguez-Espinosa PF, Rodríguez-González F (2019) Microplastics in freshwater sediments of Atoyac River basin, Puebla City, Mexico. *Sci Total Environ* 654:154–163. https://doi.org/10.1016/j.scitotenv.2018.11.054
Siegfried M, Koelmans AA, Besseling E, Kroeze C (2017) Export of microplastics from land to sea. A modelling approach. Water Res 127(Supplement C):249–257. https://doi.org/10.1016/j.watres.2017.10.011

Sighicelli M, Pietrelli L, Lecce F, Iannilli V, Falconieri M, Coscia L et al (2018) Microplastic pol- lution in the surface waters of Italian Subalpine Lakes. Environ Pollut 236:645–651. https:// doi.org/10.1016/j.envpol.2018.02.008

Silva-Cavalcanti JS, Silva JDB, de França EJ, de Araújo MCB, Gusmão F (2017) Microplastics ingestion by a common tropical freshwater fishing resource. Environ Pollut 221:218–226. https://doi.org/10.1016/j.envpol.2016.11.068

Simon M, van Alst N, Vollertsen J (2018) Quantification of microplastic mass and removal rates at wastewater treatment plants applying Focal Plane Array (FPA)-based Fourier Transform Infrared (FT-IR) imaging. Water Res 142:1–9. https://doi.org/10.1016/j.watres.2018.05.019

Sinha T, Bhagwatwar P, Krishnamoorthy C, Chidambaram R (2019) Polymer based micro- and nanoencapsulation of agrochemicals. In: Gutiérrez TJ (ed) Polymers for agri-food applications. Springer, Cham, pp 5–28. https://doi.org/10.1007/978-3-030-19416-1_2

Sommer F, Dietze V, Baum A, Sauer J, Gilge S, Maschowski C, Gieré R (2018) Tire abrasion as a major source of microplastics in the environment. Aerosol Air Qual Res 18(8):2014–2028. https://doi.org/10.4209/aaqr.2018.03.0099

Sruthy S, Ramasamy EV (2017) Microplastic pollution in Vembanad Lake, Kerala, India: the first report of microplastics in lake and estuarine sediments in India. Environ Pollut 222:315–322. https://doi.org/10.1016/j.envpol.2016.12.038

Stanton T, Johnson M, Nathanail P, MacNaughton M, Gomes RL (2020) Freshwater microplastic concentrations vary through both space and time. Environ Pollut 263(B):114481. https://doi. org/10.1016/j.envpol.2020.114481

Steinmetz Z, Wollmann C, Schaefer M, Buchmann C, David J, Tröger J et al (2016) Plastic mulch- ing in agriculture. Trading short-term agronomic benefits for long-term soil degradation? Sci Total Environ 550:690–705. https://doi.org/10.1016/j.scitotenv.2016.01.153

Su L, Xue Y, Li L, Yang D, Kolandhasamy P, Li D, Shi H (2016) Microplastics in Taihu Lake, China. Environ Pollut 216(Supplement C):711–719. https://doi.org/10.1016/j.envpol.2016.06.036

Su L, Cai H, Kolandhasamy P, Wu C, Rochman CM, Shi H (2018) Using the Asian clam as an indi- cator of microplastic pollution in freshwater ecosystems. Environ Pollut 234:347–355. https:// doi.org/10.1016/j.envpol.2017.11.075

Su Y, Zhang Z, Wu D, Zhan L, Shi H, Xie B (2019) Occurrence of microplastics in landfill systems and their fate with landfill age. Water Res 164:114968. https://doi.org/10.1016/j. watres.2019.114968

Sun J, Dai X, Wang Q, van Loosdrecht MCM, Ni B-J (2019) Microplastics in wastewater treatment plants: detection, occurrence and removal. Water Res 152:21–37. https://doi.org/10.1016/j. watres.2018.12.050

Sundt P, Schulze P-E, Syversen F (2014) Sources of microplastics pollution to the marine environ- ment, No. M-321 | 2015. MEPEX fpr the Norwegian Environment Agency, p 86. Retrieved from http://www.miljodirektoratet.no/Documents/publikasjoner/M321/M321.pdf

Sundt P, Syversen F, Skogesal O, Schulze P-E (2016) Primary microplastic pollution: measures and reduction potentials in Norway, No. M-545. Norwegian Environment Agency, p 117. Retrieved from http://www.miljodirektoratet.no/Documents/publikasjoner/M545/M545.pdf

Townsend AK, Barker CM (2014) Plastic and the nest entanglement of urban and agricultural crows. PLoS One 9(1):e88006. https://doi.org/10.1371/journal.pone.0088006

Turner S, Horton AA, Rose NL, Hall C (2019) A temporal sediment record of microplastics in an urban lake, London. UK J Paleolimnol:1–14. https://doi.org/10.1007/s10933-019-00071-7

UNEP (2009) Marine litter: a global challenge. Regional Seas, United Nations Environment Programme, Nairobi. Retrieved from: http://wedocs.unep.org/handle/20.500.11822/7787

UNESCO (2017) Waste water: the untapped resources. The United Nations World Water Development Report. Retrieved from: https://www.unido.org/sites/default/files/2017-03/UN_ World_Water_Development_Report_-_Full_0.pdf
Unice KM, Kreider ML, Panko JM (2013) Comparison of tire and road wear particle concentrations in sediment for watersheds in France, Japan, and the United States by quantitative pyrolysis GC/MS Analysis. *Environ Sci Technol* 47(15):8138–8147. https://doi.org/10.1021/es400871j

Uurasjärvi E, Hartikainen S, Setälä O, Lehtiniemi M, Koistinen A (2020) Microplastic concentrations, size distribution, and polymer types in the surface waters of a northern European lake. *Water Environ Res* 92(1):149–156. https://doi.org/10.1002/wer.1229

van den Berg P, Huerta-Lwanga E, Corradini F, Geissen V (2020) Sewage sludge application as a vehicle for microplastics in eastern Spanish agricultural soils. *Environ Pollut* 261:114198. https://doi.org/10.1016/j.envpol.2020.114198

van Emmerik T, Loozen M, van Oeveren K, Buschman F, Prinsen G (2019) Riverine plastic emission from Jakarta into the ocean. *Environ Res Lett* 14(8):084033. https://doi.org/10.1088/1748-9326/ab30e8

van Wijnen J, Ragas AMJ, Kroeze C (2019) Modelling global river export of microplastics to the marine environment: Sources and future trends. *Sci Total Environ* 673:392–401. https://doi.org/10.1016/j.scitotenv.2019.04.078

Vaughan R, Turner SD, Rose NL (2017) Microplastics in the sediments of a UK urban lake. *Environ Pollut* 229:10–18. https://doi.org/10.1139/facets-2016-0070

Veerasingam S, Mugilarasan M, Venkatachalapathy R, Vethamony P (2016) Influence of 2015 flood on the distribution and occurrence of microplastic pellets along the Chennai coast, India. *Mar Pollut Bull* 109(1):196–204. https://doi.org/10.1016/j.marpolbul.2016.05.082

Vegkart (2020) Road statistics from the road map applications Vegkart https://vegkart.atlas.vegvesen.no/. Accessed 8th May 2020

Vermaire JC, Pomeroy C, Herczegh SM, Haggart O, Murphy M (2017) Microplastic abundance and distribution in the open water and sediment of the Ottawa River, Canada, and its tributaries. *FACETS* 2:201–314. https://doi.org/10.1139/facets-2016-0070

Villena OC, Terry I, Iwata K, Landa ER, LaDeau SL, Leisnham PT (2017) Effects of tire leachate on the invasive mosquito Aedes albopictus and the native congener Aedes triseriatus. *PeerJ* 5:e3756. https://doi.org/10.7717/peerj.3756

Vogelsang C, Lusher AL, Dadkhah ME, Sundvor I, Umar M, Ranneklev SB et al (2019) Microplastics in road dust – characteristics, pathways and measures. Norsk institutt for vannforskning. Retrieved from https://brage.bibsys.no/xmlui/handle/11250/2493537

Vollertsen J, Hansen AA (2017) Microplastic in Danish wastewater: Sources, occurrences and fate, Environmental Project No. 1906. Danish Environmental Protection Agency, pp 1–55. Retrieved from https://www2.mst.dk/Udgiv/publications/2017/03/978-87-93529-44-1.pdf

Wagner S, Hüffer T, Klöckner P, Wehrhahn M, Hofmann T, Reemtsma T (2018) Tire wear particles in the aquatic environment – a review on generation, analysis, occurrence, fate and effects. *Water Res* 139:83–100. https://doi.org/10.1016/j.watres.2018.03.051

Walling DE, Webb BW, Woodward JC (1992) Some sampling considerations in the design of effective strategies for monitoring sediment-associated transport. Erosion and Sediment Transport Monitoring Programmes in River Basins (Proceedings of the Oslo Symposium, August 1992), IAHS Publication no. 210, pp 279–288. Retrieved from: https://pdfs.semanticscholar.org/c3db/f67dbe41df9f7194ebe55d3748015fd816ad.pdf?_ga=2.165832254.18612053.1583661455-2018662021.1583059758

Walling DE, Owens P, Carter J, Leeks GJ, Lewis S, Meharg A, Wright J (2003) Storage of sediment-associated nutrients and contaminants in river channel and floodplain systems. *Appl Geochem* 18(2):195–220. https://doi.org/10.1016/S0883-2927(02)00121-X

Wang W, Nnungu AW, Li Z, Wang J (2017) Microplastics pollution in inland freshwaters of China: a case study in urban surface waters of Wuhan, China. *Sci Total Environ* 575:1369–1374. https://doi.org/10.1016/j.scitotenv.2016.09.213

Wang W, Yuan W, Chen Y, Wang J (2018) Microplastics in surface waters of Dongting Lake and Hong Lake, China. *Sci Total Environ* 633:539–545. https://doi.org/10.1016/j.scitotenv.2018.03.211
Wang T, Li B, Zou X, Wang Y, Li Y, Xu Y et al (2019a) Emission of primary microplastics in mainland China: Invisible but not negligible. Water Res 162:214–224. https://doi.org/10.1016/j.watres.2019.06.042

Wang Z, Qin Y, Li W, Yang W, Meng Q, Yang J (2019b) Microplastic contamination in freshwater: first observation in Lake Ulansuhai, Yellow River Basin, China. Environ Chem Lett 17(4):1821–1830. https://doi.org/10.1007/s10311-019-00888-8

Watkins L, Sullivan PJ, Walter MT (2019) A case study investigating temporal factors that influence microplastic concentration in streams under different treatment regimes. Environ Sci Pollut Res 26(21):21797–21807. https://doi.org/10.1007/s11356-019-04663-8

Wei S, Luo H, Zou J, Chen J, Pan X, Rousseau DPL, Li J (2020) Characteristics and removal of microplastics in rural domestic wastewater treatment facilities of China. Sci Total Environ 739:139935. https://doi.org/10.1016/j.scitotenv.2020.139935

Wen X, Du C, Xu P, Zeng G, Huang D, Yin L, Yin Q, Hu L, Wan J, Zhang J, Tan S, Deng R (2018) Microplastic pollution in surface sediments of urban water areas in Changsha, China: Abundance, composition, surface textures. Mar Pollut Bull 136:414–423. https://doi.org/10.1016/j.marpolbul.2018.09.043

Wik A, Dave G (2005) Environmental labeling of car tires—toxicity to Daphnia magna can be used as a screening method. Chemosphere 58(5):645–651. https://doi.org/10.1016/j.chemosphere.2004.08.103

Wik A, Dave G (2006) Acute toxicity of leachates of tire wear material to Daphnia magna—variability and toxic components. Chemosphere 64(10):1777–1784. https://doi.org/10.1016/j.chemosphere.2005.12.045

Wik A, Dave G (2009) Occurrence and effects of tire wear particles in the environment – a critical review and an initial risk assessment. Environ Pollut 157(1):1–11. https://doi.org/10.1016/j.envpol.2008.09.028

Wik A, Nilsson E, Källqvist T, Tobiesen A, Dave G (2009) Toxicity assessment of sequential leachates of tire powder using a battery of toxicity tests and toxicity identification evaluations. Chemosphere 77(7):922–927. https://doi.org/10.1016/j.chemosphere.2009.08.034

Windsor FM, Tilley RM, Tyler CR, Ormerod SJ (2019a) Microplastic ingestion by riverine macropinvertebrates. Sci Total Environ 646:68–74. https://doi.org/10.1016/j.scitotenv.2018.07.271

Windsor FM, Durance I, Horton AA, Thompson RC, Tyler CR, Ormerod SJ (2019b) A catchment-scale perspective of plastic pollution. Glob Chang Biol 25:1207–1221

Winter LT, Foster IDL, Charlesworth SM, Lees JA (2001) Floodplain lakes as sinks for sediment-associated contaminants—a new source of proxy hydrological data? Sci Total Environ 266(1–3):187–194. https://doi.org/10.1016/s0048-9697(00)00745-2

Wright SL, Ulke J, Font A, Chan KLA, Kelly FJ (2020) Atmospheric microplastic deposition in an urban environment and an evaluation of transport. Environ Int 136:105411. https://doi.org/10.1016/j.envint.2019.105411

Xiong X, Zhang K, Chen X, Shi H, Luo Z, Wu C (2018) Sources and distribution of microplastics in China’s largest inland lake – Qinghai Lake. Environ Pollut 235:899–906. https://doi.org/10.1016/j.envpol.2017.12.081

Xu Q, Gao Y, Xu L, Shi W, Wang F, LeBlanc GA et al (2020) Investigation of the microplastics profile in sludge from China’s largest Water reclamation plant using a feasible isolation device. J Hazard Mater 388:122067. https://doi.org/10.1016/j.jhazmat.2020.122067

Yang L, Li K, Cui S, Kang Y, An L, Lei K (2019) Removal of microplastics in municipal sewage from China’s largest water reclamation plant. Water Res 155:175–181. https://doi.org/10.1016/j.watres.2019.02.046

Yin L, Jiang C, Wen X, Du C, Zhong W, Feng Z, Long Y, Ma Y (2019) Microplastic pollution in surface water of urban lakes in Changsha, China. Int J Environ Res Public Health 16(9):1650. https://doi.org/10.3390/ijerph16091650

Yu M, van der Ploeg M, Lwanga EH, Yang X, Zhang S, Ma X et al (2019) Leaching of microplastics by preferential flow in earthworm (Lumbricus terrestris) burrows. Environ Chem 16(1):31–40. https://doi.org/10.1071/EN18161
Yuan W, Liu X, Wang W, Di M, Wang J (2019) Microplastic abundance, distribution and composition in water, sediments, and wild fish from Poyang Lake, China. *Ecotoxicol Environ Saf* 170:180–187. https://doi.org/10.1016/j.ecoenv.2018.11.126

Zhang Z, Chen Y (2020) Effects of microplastics on wastewater and sewage sludge treatment and their removal: a review. *Chem Eng J* 382:122955. https://doi.org/10.1016/j.cej.2019.122955

Zhang GS, Liu YF (2018) The distribution of microplastics in soil aggregate fractions in southwestern China. *Sci Total Environ* 642:12–20. https://doi.org/10.1016/j.scitotenv.2018.06.004

Zhang K, Su J, Xiong X, Wu X, Wu C, Liu J (2016a) Microplastic pollution of lakeshore sediments from remote lakes in Tibet plateau, China. *Environ Pollut* 219:450–455. https://doi.org/10.1016/j.envpol.2016.05.048

Zhang K, Su J, Xiong X, Wu X, Wu C, Liu J (2016b) Microplastic pollution of lakeshore sediments from remote lakes in Tibet plateau, China. *Environ Pollut* 219:450–455. https://doi.org/10.1016/j.envpol.2016.05.048

Zhang Y, Gao T, Kang S, Sillanpää M (2019) Importance of atmospheric transport for microplastics deposited in remote areas. *Environ Pollut* 254:112953. https://doi.org/10.1016/j.envpol.2019.07.121

Zhao S, Zhu L, Li D (2016) Microscopic anthropogenic litter in terrestrial birds from Shanghai, China: Not only plastics but also natural fibers. *Sci Total Environ* 550:1110–1115. https://doi.org/10.1016/j.scitotenv.2016.01.112

Zhou P, Huang C, Fang H, Cai W, Li D, Li X, Yu H (2011) The abundance, composition and sources of marine debris in coastal seawaters or beaches around the northern South China Sea (China). *Mar Pollut Bull* 62(9):1998–2007. https://doi.org/10.1016/j.marpolbul.2011.06.018

Zhou Q, Tian C, Luo Y (2017) Various forms and deposition fluxes of microplastics identified in the coastal urban atmosphere. *Chin Sci Bull* 62(33):3902–3909. https://doi.org/10.1360/N972017-00956

Zhou B, Wang J, Zhang H, Shi H, Fei Y, Huang S et al (2019) Microplastics in agricultural soils on the coastal plain of Hangzhou Bay, east China: multiple sources other than plastic mulching film. *J Hazard Mater* 121814. https://doi.org/10.1016/j.jhazmat.2019.121814

**Open Access** This chapter is licensed under the terms of the Creative Commons Attribution 4.0 International License (http://creativecommons.org/licenses/by/4.0/), which permits use, sharing, adaptation, distribution and reproduction in any medium or format, as long as you give appropriate credit to the original author(s) and the source, provide a link to the Creative Commons license and indicate if changes were made.

The images or other third party material in this chapter are included in the chapter's Creative Commons license, unless indicated otherwise in a credit line to the material. If material is not included in the chapter's Creative Commons license and your intended use is not permitted by statutory regulation or exceeds the permitted use, you will need to obtain permission directly from the copyright holder.

<figure><img src="image_26.png" /></figure>
**Chapter 5**

**Marine Microplastics and Seafood:**
**Implications for Food Security**

<figure><img src="image_27.png" /></figure>

Check for
updates

Anne-Katrine Lundebye, Amy L. Lusher, and Michael S. Bank

**Abstract** Seafood is an important food source, and this chapter addresses the food safety concerns related to plastic particles in different seafood. Here we focus on those species which are commonly consumed by humans, such as bivalves, cephalopods, echinoderms, crustaceans, and finfish. The objectives of this chapter are to (1) outline the major sources, fate, and transport dynamics of microplastics in marine ecosystems, (2) provide a critical assessment and synthesis of microplastics in seafood taxa commonly consumed by humans, (3) discuss the implications of microplastics with regard to human health risk assessments, and (4) suggest future research priorities and recommendations for assessing microplastics in marine ecosystems in the context of global food security and ocean and human health.

**5.1 Introduction**

Seafood is an important food source – with fisheries and aquaculture production predicted to increase by about 17.5% from 171 million tonnes in 2016 to approximately 201 million tonnes in 2030 (FAO 2018). It is a necessity that these marine-based foods are carefully managed and are safe for human consumption. Food

A.-K. Lundebye (✉)
Institute of Marine Research, Bergen, Norway
e-mail: Anne-Katrine.Lundebye@hi.no

A. L. Lusher
Norwegian Institute for Water Research, Oslo, Norway
University of Bergen, Bergen, Norway
e-mail: Amy.Lusher@niva.no

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

University of Massachusetts Amherst, Amherst, MA, USA
e-mail: Michael.Bank@hi.no; mbank@eco.umass.edu
security is defined by the Food and Agriculture Organization (FAO) as “a situation that exists when all people, at all times, have physical, social and economic access to sufficient, safe and nutritious food that meets their dietary needs and food preferences for an active and healthy life” (FAO 2017). In this chapter, the food safety concerns related to plastic particles in seafood will be addressed.

Global annual production of plastics is estimated to be approximately 300 million tonnes (Galloway 2015) and is still increasing steadily. Most plastic polymers are resistant to complete degradation and pose a potential risk to both human and environmental health. Of particular concern are microplastics, which are defined as particles <5 mm (GESAMP 2019) and which are the focus of this chapter.

Microplastics occur in different shapes and sizes and are formed from different polymers as well as additives, which reflects the diversity of sources and emissions to the environment (Rochman et al. 2019). The dominant microplastic polymers which are detected in the marine environment include polyethylene (PE), polypropylene (PP), polystyrene (PS), polyethylene terephthalate (PET), polyamide, and polyvinylchloride (PVC) (Hantoro et al. 2019). Primary microplastics are manufactured intentionally for a range of commercial uses (e.g., microbeads) whereas secondary microplastics originate from parent material such as textiles and discarded plastic items and are either generated through the use of plastic products or fragmentation following their loss to different environmental compartments. Plastic debris can enter the ocean from ships and fishing gear, as well as from atmospheric deposition, river transport, stormwater, sewage effluents, etc. (Browne et al. 2011; Napper and Thompson 2016; Lebreton et al. 2017; Allen et al. 2019). Plastics and microplastics have been identified in the oceans, from coastal zones to offshore areas, such as oceanic gyres (Eriksen et al. 2014; Jiang et al. 2020;), as well as in remote areas including the Arctic (e.g., Cózar et al. 2017). The ubiquitous nature of plastics in the ocean is an obvious concern for marine ecosystems and their inhabitants. In particular, the progression from macro- to microplastics at sea is a result of physical erosion and UV action and increases the bioavailability of smaller-sized particles to a wide array of marine organisms (Browne et al. 2008; Wright et al. 2013). Plastics have long been reported associated with marine organisms, from the first study of plastic ingestion by fish (Carpenter et al. 1972) to mariculture sites where boring worms facilitate the generation of microplastics from polystyrene buoys (Jang et al. 2018). Many investigations have been conducted to further understand the interaction between marine organisms and microplastics with several studies focusing on microplastic uptake, ingestion, exposure, and metabolic dynamics (Roch et al. 2020). Fibers are routinely identified as the most common microplastic type reported in fish, accounting for 58–87% of the plastic morphologies observed (Walkinshaw et al. 2020). Fragments, films, and fibers are also frequently found in fish and shellfish while microplastics in the forms of spheres are less common. The physical impacts of microplastic ingestion on marine organisms can include oxidative stress, inflammation, and potentially starvation, while less is known regarding the chemical effects of ingestion. The bioavailability and potential toxicity of microplastics are size dependent, with smaller particles able to penetrate further into an organism (Browne et al. 2008), with the potential of the release of associated
co-contaminants (Bakir et al. 2016; Batel et al. 2016). It is widely accepted that ingestion is the main route of microplastics uptake in marine biota; however, it has recently also been demonstrated that surface scavenging appears to be an alternative route, as demonstrated in mussels (Kolandhasamy et al. 2018). Microplastics can also be taken up through respiration via the gills (Watts et al. 2016; Franzellitti et al. 2019) and have additionally been demonstrated to be maternally transferred to eggs in zebra fish (Pitt et al. 2018).

Plastics in aquatic environments have been shown to affect an organism's health (such as behavioral changes and reduced growth rates); however, there is limited information on the effects of microplastics in seafood on human health. Lusher et al. (2017a) reported that more than 220 species of marine organisms including zooplankton, bivalves, crustaceans, fish, marine mammals, sea turtles, and seabirds had been shown to have ingested plastics, and more recently this number of species has increased to 690 (Carbery et al. 2018). Here we focus on those species which are commonly consumed by humans, such as bivalves, echinoderms, gastropods, cephalopods, crustaceans, and finfish. The specific objectives of this chapter are to (1) outline the major sources, fate, and transport dynamics of microplastics in marine ecosystems, (2) provide a critical assessment and synthesis of microplastics in seafood taxa commonly consumed by humans, (3) discuss the implications of microplastics with regard to human health risk assessments, and (4) suggest future research priorities and recommendations for assessing microplastics in marine ecosystems in the context of global food security and ocean and human health.

## 5.2 Fate and Transport of Microplastics in Marine Ecosystems.

The fate and transport of microplastics in the context of physical and biological oceanography has recently been reviewed by van Sebille et al. (2020), as well as by Thushari and Senevirathna (2020) with older reviews and critical papers developed by Andrady (2011), Wright et al. (2013), Galloway et al. (2017), and Wieczorek et al. (2019). The ocean can be both a source and sink for microplastics (Allen et al. 2020), and important themes within the cycling and degradation of microplastic particles (Weinstein et al. 2016) include the importance of transport from land via rivers (Lebreton et al. 2017; Hurley et al. 2018), the role of seafloor ocean circulation patterns as a driver of microplastic hotspots (Kane et al. 2020), and the concept of marine snow which has been identified as an important mechanism for transporting microplastic particles from the water column to the sediment (Porter et al. 2018). Moreover, fishing gear and other sources of macroplastics can degrade into microplastics via biological, chemical, and physical processes (Davidson 2012). Although settling of microplastic particles to the ocean floor is well-documented, recent research has shown that episodic events such as flooding (Hurley et al. 2018) and
typhoons (Wang et al. 2019) are important drivers regarding the distribution and abundance of microplastics in coastal marine ecosystems.

## 5.3 Microplastic in Bivalves

Bivalves are by far the most investigated seafood species (Smith et al. 2018; Walkinshaw et al. 2020). Much of the investigations were performed for the purpose of uptake of microplastics from the environment, as filtering puts bivalves at an increased risk of microplastic intake from the water column (Li et al. 2019). Early investigations focused on blue mussels (*Mytilus* spp.), with wild and market brought samples presenting contamination levels of up to 7.2 microplastics per gram (Abidli et al. 2019; Bråte et al. 2018; Cho et al. 2019; De Witte et al. 2014; Li et al. 2015, 2016; Renzi et al. 2018a; van Cauwenberghe and Janssen 2014; van Cauwenberghe et al. 2015; Vandermeersch et al. 2015). Other bivalve species which have been investigated for microplastic uptake include clams (*Venerupis philippinarum*), oysters (*Crassostrea gigas*), and scallops (*Patinoppecten yessoensis*) (Abidli et al. 2019; Cho et al. 2019, 2020; Davidson and Dudas 2016; Li et al. 2015; Rochman et al. 2015; van Cauwenberghe and Janssen 2014).

Microplastic fibers are often the most dominant morphology reported in bivalves. For example, fibers accounted for 80% of microplastics in mussels (*Mytilus edulis*, *Perna viridis*) from China (Qu et al. 2018), 90% of microplastics in Manila clams (*V. philippinarum*) from British Columbia (Davidson and Dudas 2016), and 99% in razor clams (*Siliqua patula*) and Pacific oyster (*Crassostrea gigas*) from Oregon, USA (Baechler et al. 2020a). One of the hypotheses behind the observed high abundance of this type of microplastic is that fibers are likely harder to remove from digestive tracts. Ward et al. (2019) reported that larger spheres are rejected at higher numbers (98%) than smaller spheres (10–30%). Fragments were most common in blue mussels and Pacific oyster from the French Atlantic coast (Phuong et al. 2018) as well as those from Korea, where EPS fragments likely originated from the high abundance of aquaculture facilities in the region (Cho et al. 2020). De Witte et al. (2014) reported that there was a high prevalence of fibers in blue mussels collected from quaysides related to fishing activities.

Microplastics in bivalves are likely dependent on several factors including, but not limited to, culture conditions and contamination levels in the environment, depuration procedures, filtration capabilities, as well as the tissues targeted for investigation. Some investigations, in distinct parts of the world, have found that bivalves sampled from highly contaminated areas or within the vicinity of urban sources of microplastics contained higher numbers of microplastics (Bråte et al. 2018; Qu et al. 2018; Cho et al. 2020). However, conversely, some investigations have reported no difference in microplastic exposure in bivalves related to sources (Covernton et al. 2019; Phuong et al. 2018).

There have been some reported differences between the occurrence of microplastics in market purchased (80%) and wild-caught individual bivalves (40%)
(Ding et al. 2018). Similarly, farmed mussels displayed higher concentrations of microplastics than wild mussels (75 items and 34 items per mussel, respectively) (Mathalon and Hill 2014), although no difference was observed for wild and cultured Manila clams (*Venerupis philippinarum*) in British Columbia (Davidson and Dudas 2016). The use of depurations procedures appears to reduce the number of microplastics identified in bivalve species (van Cauwenberghe and Janssen 2014 – *Mytilus edulis* and *Crassostrea gigas*; Birnstiel et al. 2019 – *Perna perna*), which would hold significance in the preparation of mussels for consumption. The seasonality of sampling could also play a role in observed microplastic concentrations in marine biota. A significant seasonal variation was observed during summer for oyster samples which contained more microplastics; however, this trend was not detected for razor clams (Baechler et al. 2020c).

Particle selection by bivalves, related to size and morphology, will influence which particles are internalized both pre- and post-ingestion (Ward et al. 2018). Gut retention times, which are known to vary between bivalve species and the age of individuals, have shown, in general, that as particle size decreases, accumulation increases (Browne et al. 2008; Ward and Kach 2009; Ward et al. 2019). Much of the work performed on bivalves is based on the sampling and processing of whole organisms, with no differentiation between and among tissue types; this makes it impossible to determine whether microplastics were internalized by individuals, had migrated from gills and guts to visceral tissue, or were in the process of being egested (e.g., as pseudofeces). Kolandhasamy et al. (2018) reported that microplastic fibers can accumulate on the foot and mantle of blue mussels.

Consequences of microplastic intake/uptake by bivalves indicate that microplastics can directly affect bivalve physiology but also indirectly change the structure of their habitats, impairing food resources and facilitate the efficient transfer of organic pollutants (Zhang et al. 2019a). Other observed implications include negative effects on filtration activity (Green et al. 2019; Xu et al. 2017), feeding behavior (Wegner et al. 2012), and reproduction (Sussarellu et al. 2016; Gardon et al. 2018). It is important to highlight that effects are mostly studied using uniform particles, mostly spheres so these may not be truly representative of environmentally relevant microplastic exposure regimes (see Gomes et al. 2021, Chap. 7, this volume).

## 5.4 Microplastics in Echinoderms

Sea urchins and sea cucumbers are the main echinoderms consumed as food item, and few studies have been conducted on the abundance of microplastics in these marine organisms. Of the heart urchins (*Brissopsis lyrifera*) analyzed, 40% were found to contain microplastics in their soft tissue, primarily in the form of flakes (90%, the remaining 10% as fibers). In most cases the number of particles present was 1/individual (Bour et al. 2018). It is noteworthy that this study was conducted for an ecological assessment of the influence of habitat, feeding mode, and trophic level on microplastic abundance in benthic and epibenthic organism and that this
species is not commonly consumed. Feng et al. (2020) reported a higher prevalence of microplastics (in 90% of the individuals) in four species of sea urchins (*Strongylocentrotus intermedius*, *Temnopleurus hardwickii*, *Temnopleurus reevesii*, and *Hemicentrotus pulcherrimus*) harvested from 12 sites along the northern China coast. The average abundance of microplastics (predominantly as fibers) in soft tissue from sea urchins from all sites was 5 particles/individual (1.1 particles/g), considerably higher than reported in heart urchins from the Oslofjord, Norway (Bour et al. 2018). Higher detection rates and abundances were found in sea urchins from Dalian, China (Feng et al. 2020). The tissue of relevance in urchins with regards to seafood safety is the gonads, and while whole soft tissue of heart urchins was analyzed for microplastics (Bour et al. 2018), the abundance in urchins from the Yellow Sea was assessed in gonads, coelomic fluid, and the gut. Gonads and coelomic fluid contained significantly lower number of particles/individual than the gut in all four species of urchin; however, this difference was not evident when normalized to wet weight in three of the species, and it only remained significantly lower in *S. intermedius* (Feng et al. 2020).

Microplastic ingestion has been reported in several species of sea cucumber including *Holothuria grisea*, *Cucumaria frondosa*, *Holothuria floridana*, *Thyonella gemmata* (Graham and Thompson 2009), *Holothuria tubulosa* (Renzi et al. 2018b), *Holothuria mexicana*, *Actinopyga agassizi* (Plee and Pomory 2020), and *Apostichopus japonicus* (Mohsen et al. 2019). Sea cucumbers are commonly eaten in Asia, and farming is widespread to meet consumer demand. The body wall of sea cucumbers is typically eaten raw in Japan and boiled, pickled, or salted in China, and the internal organs (gonads, respiratory trees, and intestines) are also edible (Kiew and Don 2011). Iwalaye et al. (2020) reported microplastic particles in the intestines, coelomic fluid, and respiratory trees of the *Holothuria cinerascens* and that uptake was both via the feeding tentacles and the respiratory trees. The most abundant microplastics found in farmed sea cucumbers (*Apostichopus japonicus*) from eight locations in the Bohai Sea and Yellow Sea in China were cellophane microfibers (Mohsen et al. 2019).

## 5.5 Microplastics in Gastropods

Limited research has been carried out on microplastics in marine gastropods. Xu et al. (2020) analyzed nine species of gastropods from shores in Hong Kong for microplastics, with the highest abundance found in sea snails (*Batillaria multifor-mis*, 5.4 ± 1.2 particles/g wet weight) and the lowest observed in Chameleon nerite snails (*Nerita chamaeleon*, 1.50 ± 0.2 particles/g wet weight). The common peri-winkle (*Littorina littorea*) sampled from four different locations in Galway Bay, Ireland, contained between 0.6 and 2.8 microplastics/g wet weight of soft tissue, and commercial common periwinkles, intended for human consumption, contained on average 2.2 microplastic s/g wet weight soft tissue (Doyle et al. 2019). Most of the microplastics (97%) recorded in periwinkles were fibers. Similarly, fibers
accounted for more than half of the total microplastics present in the girdled horn shell (*Cerithidea cingulata*), whereas film was the most abundant microplastic (approximately 44%) in *Thais mutabilis* from the Persian Gulf region. The mean number of total microplastics was 13 and 20 particles/g wet soft tissue weight for *C. cingulata* and *T. mutabilis*, respectively (Naji et al. 2018). Lower levels of microplastic contamination were reported in periwinkles (*Littorina* spp.) from two sites on the eastern coast of Thailand with an average of 0.17 particles/g wet weight and 0.23 particles/g wet weight, with no contamination observed in periwinkles from Bangasaen, the third site investigated (Thushari et al. 2017).

## 5.6 Microplastics in Cephalopods

Cephalopods are the seafood phylum which have received the least focus with regard to microplastic contamination. Oliveira et al. (2020) investigated the levels of microplastics in the stomach, caecum/intestine, and digestive gland of cuttlefish (*Sepia officinalis*); however, tissue relevant for consumer exposure was not included in this study. Microplastic contamination in Indian squid (*Uroteuthis duvaucelii*) was found in 18% of the individuals examined, with an average of 0.2 microplastic particles/individual and 0.008 microplastic particles/g wet weight of edible tissue (Peng et al. 2020).

## 5.7 Microplastics in Crustaceans

Most biota-based studies have examined microplastics in the organisms' gut, which is not generally an organ consumed directly by humans. However, shellfish including crustaceans and mollusks are an exception since these are frequently eaten either whole or with their gut removed. The risk of ingesting microplastics from other tissues, such as muscle, depends on the ability to cross the intestinal barrier and subsequent accumulation (Zeytin et al. 2020).

To date, most literature on crustaceans which are commonly harvested for human consumption has focused on wild individuals, rather than those that are farmed. Investigations generally have not focused on crustaceans in the context of seafood safety but rather from an environmental contaminant perspective. For example, there have been numerous investigations into langoustine, *Nephrops norvegicus*, which are also commercially exploited. *N. norvegicus*, sampled from the Clyde Sea area, were shown to contain more microplastic fibers in their gut than individuals from the North Sea and Minch where only a small percentage of individuals contained microplastic, predominantly as single-strand fibers (Welden and Cowie 2016). Other commercially relevant species, such as spinous spider crabs (*Maja squinado*), shrimps, and prawns, have been observed to contain microplastics (Welden et al. 2018; Devriese et al. 2015; Zhang et al. 2019b; Cau et al. 2019).
Many crustaceans are harvested from coastal environments, which may be close to sources of microplastic contamination, including the influence of terrestrial plastic sources. As an example, shrimp (*A. antennatus*) from the Mediterranean had an average occurrence of microplastics equal to 39.2%; however, those in close vicinity to urban areas had 100% presence of microplastics (Carreras-Colom et al. 2018). The same % occurrence trend was observed between remote populations (<40%) of *N. norvegicus* compared to those sampled near Glasgow (84%) in the Clyde Sea (Welden and Cowie 2016).

Additionally, no spatial pattern was observed in a similar study of *N. norvegicus* in Irish waters (Hara et al. 2020). Both *N. norvegicus* and *Aristeus antennatus* were investigated in the Mediterranean Sea from depths between 270 and 660 meters (Cau et al. 2019). The authors reported a significant difference in the size and composition of microplastics identified between the two species and suggested that the nonselective feeding strategy of *N. norvegicus* likely led to a higher degree of exposure to microplastics and hence a higher measured abundance. Nonselective feeding is an example of direct uptake of microplastics from the environment. Organisms can also ingest microplastics which have been internalized by prey species, a concept commonly referred to as trophic transfer. Laboratory studies on this topic performed with shore crabs (*Carcinus maenas*) fed mussels which had been exposed to microplastics showed that polystyrene microspheres could accumulate in the foregut of the crabs (Watts et al. 2015).

Fibers and fragments are the most often reported particle type in crustaceans sampled from the wild, with fiber bundles reported across many species (Welden and Cowie 2016; Cau et al. 2019; McGoran et al. 2020). In most studies, stomachs were often the target organ of microplastics investigations, but other tissues are starting to be considered further, as these may have relevance for human exposure, especially when stomachs are removed prior to cooking and consumption. As an example, microplastics have been found in different tissues of wild-caught *Portunus gracilimanus* and *P. trituberculatus* (Zhang et al. 2019b).

## 5.8 Microplastics in Finfish

Evaluating microplastic occurrence and abundance in finfish is fundamental to understanding how plastics and their associated chemical compounds affect and potentially impact wild fisheries that are relied upon by humans as an important source of food and nutrition (Rochman et al. 2015; Barboza et al. 2018; FAO 2020; Lusher and Welden 2020). The topic of microplastics in the marine environment, including information on finfish, has been reviewed by several authors (Andrady 2011; Cole et al. 2011; Hidalgo-Ruz et al. 2012; Wright et al. 2013; Gall and Thompson 2015; Galloway et al. 2017; Baechler et al. 2020a, b; Thushari and Senevirathna 2020; Wang et al. 2020; Walkinshaw et al. 2020). Microplastics exposure in finfish is largely a result of plastics being mistaken for natural prey items, via ingestion of contaminated prey items or by passive uptake through gills (Lusher
et al. 2016; Watts et al. 2015; Nelms et al. 2018; Roch et al. 2020). Trophic transfer of microplastics may also expose predaceous fish to microplastics (Farrell and Nelson 2013; Setälä et al. 2014; Lusher et al. 2016; Baechler et al. 2020a), and microplastics have frequently been detected in finfish gastrointestinal tracts (e.g., Lusher et al. 2017b,). The methodological challenges with identifying particles within fillet muscle tissue have limited the number of published studies thus far, although they have been identified albeit at extremely low concentrations (Zeytin et al. 2020).

Many species of edible demersal, pelagic, and reef fish, sampled from across the globe, have been found to contain microplastics (e.g., Bellas et al. 2016; Rummel et al. 2016; Bråte et al. 2016; Lusher et al. 2013; Tanaka and Takada 2016; Rochman et al. 2015; Neves et al. 2015; Critchell and Hoogenboom 2018; Abbasi et al. 2018; Su et al. 2018). The percentages of different fish species which have been found to contain microplastics in their gut vary greatly: 0.9% Peruvian anchovy, 2.8% Atlantic cod, 8.8% Atlantic herring, 9.4% Skipjack tuna, 24.5% Jack and Horse mackerel, 23.3% Pacific chub mackerel, 23.4% Yellowfin tuna, and 76.6% Japanese anchovy (Neves et al. 2015; Lusher et al. 2013; Güven et al. 2017; Ogonowski et al. 2017; Rummel et al. 2016; Hermsen et al. 2018; Rochman et al. 2015; Choy and Drazen 2013; Markic et al. 2018; Bråte et al. 2016; Liboiron et al. 2016; Tanaka and Takada 2016). Several studies have examined the microplastic particle prevalence in fish with different feeding ecology (Foekema et al. 2013; Lusher et al. 2013). Lusher et al. (2013) did not find any significant difference between the abundance of plastic ingested by pelagic and demersal fish. Of the 24 fish species examined from the Beibu Gulf, one of the world's largest fishing grounds, in the South China Sea, 12 species contained microplastics (Koongolla et al. 2020). The abundance of microplastics varied from 0.027 to 1 item per individual, and most was present in fish stomach (57.7%) and less in intestines and gills (34.6% and 7.7%, respectively). Nine of the 11 fish species sampled from Zhoushan fishing grounds in the East China Sea were found to contain microplastics, with 23 different polymer types identified, and the highest number of items was 8 in a single individual (Zhang et al. 2019a). It is challenging to compare all the studies listed above, as many different methods have been utilized by researchers to determine the presence or absence of microplastics across these species. Some trends in the methods used have previously been described, with visually searching the most common method (Lusher et al. 2017b); however, the lack of standards and incomplete reporting of data, and quality control procedures have also been highlighted (Hermsen et al. 2018). Differences in sampling and analytical methods may lead to different values being observed and are important to consider when evaluating trends across regions, ecosystem types, and species.

The microplastic content of wild fish has been more widely studied than aquaculture species. A recent review of microplastics in seafood found that data were lacking for four of the ten most cultured aquatic food species, namely, grass carp, whiteleg shrimp, bighead carp, and catla (Walkinshaw et al. 2020).
## 5.9 Co-contaminants Associated with Microplastics in Seafood

The role of marine microplastics as vectors for major ocean pollutants was recently reviewed by Ziccardi et al. (2016), Santillo et al. (2017), and Amelia et al. (2021). Plastics are inherently complex in size, morphology, and polymer composition and may contain a range of additives, including plasticizers, stabilizers, pigments, fillers, and flame retardants which may leach out into the environment including air, water, and food, and in general, microplastics are now considered to represent a suite of co-contaminants (Rochman et al. 2019). More than 50% of plastics are associated with hazardous monomers, additives, and chemical byproducts (Lithner et al. 2011). Plastics have been shown to accumulate various organic and inorganic co-contaminants from the surrounding water column (Rochman et al. 2013, 2015). The high surface area to volume ratio of small particles and hydrophobic nature facilitate the sorption of chemicals on the plastic surface, forming a complex mixture of contaminants available to marine organisms (Rochman et al. 2013). Laboratory studies have demonstrated that continuous exposure to contaminated plastics can lead to the accumulation of plastic-associated co-contaminants in fish (Rochman et al. 2013; Wardrop et al. 2016).

Both field and modeling studies suggest that transfer of environmental pollutants through microplastics are negligible compared to other routes of uptake (Gouin et al. 2011; Bakir et al. 2016; Espinosa et al. 2018; Koelmans et al. 2016; Ziccardi et al. 2016; Lohmann 2017; Smith et al. 2018). Nonetheless, caution is warranted as many of the chemicals sorbed onto microplastics are known to be potent toxicants to humans and marine biota, triggering adverse effects such as endocrine disruption, neurological disorders, and reduced reproductive success (GESAMP 2016). An example of this is the investigation by Barboza et al. (2020a) who reported significantly higher concentrations of bisphenols in fish with microplastics compared to individuals with no microplastics. However, none of the fish species investigated (European seabass *Dicentrarchus labrax*, Atlantic horse mackerel *Trachurus trachurus*, and Atlantic chub mackerel *Scomber colias*) contained bisphenol A levels which would lead to an exceedance of the Tolerable daily Intake established by the European Food Safety Authority (EFSA) (Barboza et al. 2020a).

## 5.10 Microplastic Uptake and Toxicity in Humans

The uptake of microplastics is dependent on size, morphology, solubility, and surface charge and chemistry. Microplastics <130 µm in diameter can potentially translocate into human tissue (EFSA 2016), and particles sized 1.5 µm and below can penetrate capillaries (Yoo et al. 2011). Proposed mechanisms for uptake of microplastics include endocytotic and paracellular transfer across epithelial tissues (Wright and Kelly 2017). It is estimated that 90% of ingested microplastics are
excreted from the body (EFSA 2016); however, the remaining microplastics may be detrimental to human health, and further research is required to develop a more comprehensive understanding regarding public health aspects of microplastic pollution.

Oxidative stress and subsequent inflammation are both thought to be the main mechanisms of particle toxicity (Feng et al. 2016). Other potential biological responses to microplastic exposure include genotoxicity, apoptosis, and necrosis, which could ultimately lead to tissue damage, fibrosis, and carcinogenesis (Wright and Kelly 2017). The extent of potential adverse effects is dependent on particle size, and nanoparticles have been found to generate more reactive oxygen species (ROS) than larger particles and are more likely to be translocated (Stone et al. 2007). Consequently, potential health effects of microplastics largely depend on particle characteristics, and it is envisaged that nanoplastics are likely more deleterious than microplastics (Feng et al. 2016).

## 5.11 Consequences of Microplastics in Marine Animals

More than 690 marine species from different trophic levels have been reported to contain plastic debris; however, the transfer of microplastics and associated co-contaminants, from seafood to humans, and the implications for seafood safety have received limited attention to date (Carbery et al. 2018; Lusher et al. 2017a; Walkinshaw et al. 2020). Most studies conducted have considered the environmental rather than the potential human health impacts of micro- and nanoplastics. Effects of micro- and nanoplastic exposure reported in marine organisms include reduced growth, impacted energy metabolism, feeding behavior, and locomotion, effects on the immune system, and hormonal regulation, physiological stress, oxidative stress, inflammation, aberrant development, cell death, general toxicity, and altered lipid metabolism (Kögel et al. 2019). In humans, it is evidenced that consumers may be exposed to microplastics from seafood consumption; however, the risks remain unclear (Smith et al. 2018; VKM 2019).

Shellfish and small fish that are consumed whole are the seafoods which are likely to give the highest exposure risk since the gastrointestinal tract, which generally contains the highest microplastic concentrations, are consumed (van Raamsdonk et al. 2020). In contrast most fish species are filleted, and most crustaceans have their digestive tracts removed before consumption, thereby reducing microplastic exposure. Similarly, bivalves, shellfish, and other lower trophically positioned marine organisms are probably the most important seafood source of dietary exposure to microplastics (Walkinshaw et al. 2020). It has been estimated that the average European shellfish consumer may ingest up to 11,000 microplastics per year based on levels in mussels and oysters (van Cauwenberghe and Janssen 2014). A systematic review and meta-analysis of microplastic contamination in seafood reported a maximum annual consumption of 55,000 microplastic particles, with mollusks from Asia being the most heavily contaminated (Danopoulos et al. 2020).
The presence of several types of microplastics in human stool from different countries has been reported, with PP and PET being the most abundant types (Schwabl et al. 2019), indicating human exposure. However, there is currently no indisputable evidence on the effects of microplastics on human health (Toussaint et al. 2019). Potential health impacts can result directly such as tissue damage but also indirectly from environmental contaminants associated with microplastics or associated microorganisms (Oberbeckmann et al. 2015).

While the focus of the scientific literature has primarily been on human exposure to microplastics from seafood consumption, much less data is available on the occurrence of microplastics in other food groups, so their relative contribution is unknown which is important from a risk assessment perspective (Wright and Kelly 2017). Data on microplastics in foods (Kwon et al. 2020) other than seafood include sugar, salt, honey, and drinking water and beer (Karbalaei et al. 2018), whereas there are significant data gaps for plant- and terrestrial animal-derived foods (van Raamsdonk et al. 2020). To date there are also limited data on microplastic levels in freshwater fish (Collard et al. 2019) and terrestrial foods (e.g., vegetables, poultry). In addition to food and drinking water, inhalation is a potential route of exposure, and atmospheric fallout is thus also an important source of microplastic exposure (Dris et al. 2016; Allen et al. 2019). Catarino et al. (2018) concluded that the potential for microplastic ingestion from shellfish consumption was minimal, especially when compared to general air exposure from household dust (123–4620 particles/year/capita and 13,731–68,415 particles/year/capita from food versus dust, respectively). Similarly, Rist et al. (2018) highlighted that food and beverages likely only constitute a minor exposure pathway to human microplastic exposure. Based on the current knowledge on microplastics in seafood, there is no evidence that food safety is compromised (Gamarro et al. 2020).

The extent to which microplastics present in foods contribute to human exposure is not well understood, especially as studies evaluating microplastics and associated chemical exposure to humans are not consistent (Rist et al. 2018; Barboza et al. 2020a, b). Human exposure estimates in the USA to microplastics in food (seafood, sugars, salts, honey), drinking water, alcohol, and air found that inhalation was the main route of exposure for adults whereas drinking water was the main source for children (Cox et al. 2019). However, this study did not include major food groups such as meats, grains, and vegetables due to a lack of empirical data.

## 5.12 Challenges and Priorities in Marine Microplastic Research

Risk characterization including information on the particle size-dependent toxico-
kinetics and dynamics of microplastics is needed to calculate evidence-based guid-
ance or tolerable weekly intakes to support realistic human health and exposure risk
assessments. Discrepancies exist in the sampling, extraction, identification, and quantification of microplastics (Collard et al. 2019), and there is a need for harmonization of current procedures (Hartmann et al. 2019; Cowger et al. 2020). An effective risk assessment of the human health effects of microplastics requires reliable human exposure data which is currently limited (Toussaint et al. 2019). Knowledge gaps regarding the uptake and potential human health effects of microplastic exposure have been highlighted (EFSA 2016; Wright and Kelly 2017; van Raamsdonk et al. 2020).

Importantly, there is currently still a lack of harmonized and proven methods for microplastics which can compromise the level to which microplastic contamination in seafood species (and other foods) can be compared. Some recommendations have been presented which focus on the methods that – thus far – have proven efficient at isolating microplastics from biota tissues (e.g., Dehaut et al. 2019; Lusher et al. 2020; Ribeiro et al. 2020). The field of microplastic research has been moving very rapidly, with several advancements in methods emerging in parallel. It is therefore of great urgency to coordinate an effort to compare the field and laboratory-based methods to one another to determine the level of comparability and overall effectiveness. This is easier said than done. Currently, laboratory comparisons have been limited to scientific approaches conducted by individual research groups (e.g., Catarino et al. 2017; Karlsson et al. 2017; Thiele et al. 2019; Yu et al. 2019; Jaafar et al. 2020; Ribeiro et al. 2020), rather than between different institutions. Some interlaboratory efforts have been made, but these have generally focused on clean water samples, rather than complex matrices such as seafood material and biological tissues (e.g., ongoing EU-JRC and SCCWRP intercalibration exercises). Similarly, there are different reporting criteria that have been applied to the study of microplastics in biota, and the quantification of the microplastics is not standardized which presents some important challenges to this sub-discipline of environmental chemistry. Different measurement units are often used (e.g., numbers per weight or per individual) highlighting the need for harmonization and standardization.

Moving forward, methods will need to be adopted that are truly reproducible and that can be validated and compared using standard reference materials. This requires that validation and feasibility assessments are undertaken while also supporting initiatives that promote scientific discovery and method development. There are several methods that are promising, and utilizing these novel approaches will allow for the development of more robust and comparable methods across different sectors/regions within the sphere of microplastic research. Unfortunately, several methods are focused on the larger fraction of microplastics (e.g., >100 µm), and method development is still required for accurately detecting smaller microplastics (<20 µm) and nanoplastics (<1 µm). Methods that focus on smaller fractions are needed to support risk characterization and exposure assessments in marine biota and humans.
## 5.13 Future Recommendations and Conclusions

Microplastics are ubiquitously found in seafood, and the importance of standardized and harmonized methods for the effective biomonitoring of farmed and wild seafood species including bivalves and finfish has become evident (Lusher et al. 2017b; Hartmann et al. 2019; Ribeiro et al. 2020). Lower trophically positioned organisms may be at the highest risk of contamination from microplastics, and currently there is insufficient evidence to conduct realistic and meaningful human health risk assessments. Moreover, several seafood species from wild fisheries and aquaculture are not well studied in the context of global food security including commonly consumed taxa (Lusher et al. 2017a; Walkinshaw et al. 2020). Microplastic pollution and exposure to plastics and their associated co-contaminants via seafood consumption will likely serve as effective themes to help link the IOC-UNESCO's Decade of Ocean Science for Sustainable Development (2021–2030) with the UN Decade of Action on Nutrition (2016–2025) and to gather critical stakeholders and develop important sustainable development strategies to support ocean and human health. In conclusion, the effects of microplastics on food security are still largely unknown, and further research and robust biomonitoring efforts on seafood are required to elucidate potential impacts.

## References

* Abbasi S, Soltani N, Keshavarzi B, Moore F, Turner A, Hassanaghaei M (2018) Microplastics in different tissues of fish and prawn from the Musa Estuary, Persian Gulf. Chemosphere 205:80–87

* Abidli S, Lahbib Y, El Menif NT (2019) Microplastics in commercial molluscs from the lagoon of Bizerte (Northern Tunisia). Mar Pollut Bull 142:243–252

* Allen S, Allen D, Phoenix VR (2019) Atmospheric transport and deposition of microplastics in a remote mountain catchment. Nat Geosci 12:339–344

* Allen S, Allen D, Moss K, Le Roux G, Phoenix VR, Sonke JE (2020) Examination of the ocean as a source for atmospheric microplastics. PLoS One 15(5):e0232746. https://doi.org/10.1371/journal.pone.0232746

* Amelia TSM, Khalik WMAWM, Ong MC, Shao YT, Pan H-J, Bhubalan K (2021) Marine microplastics as vectors of major ocean pollutants and its hazards to the marine ecosystem and humans. Prog Earth Planet Sci 8:12

* Andrady AL (2011) Microplastics in the marine environment. Mar Pollut Bull 62:1596–1605

* Baechler BR, Granek EF, Hunter MV, Conn KE (2020a) Microplastic concentrations in two Oregon bivalve species: spatial, temporal, and species variability. Limnol Oceanogr Lett 5:54–65

* Baechler BR, Stienbarger CD, Horn DA, Joseph J, Taylor AR, Granek EF, Brander SM (2020b) Microplastic occurrence and effects in commercially harvested North American finfish and shellfish: current knowledge and future directions. Limnol Oceanogr Lett 5:113–136

* Baechler BR, Granek EF, Mazzone SJ, Nielsen-Pincus M, Brander SM (2020c) Microplastic exposure by razor clam in recreational harvester-consumers along a sparsely populated coastline. Front Mar Sci 7:980
Bakir A, O'Connor IA, Rowland SJ, Hendriks AJ, Thompson RC (2016) Relative importance of microplastics as a pathway for the transfer of hydrophobic organic chemicals to marine life. *Environ Pollut* 219:56–65. https://doi.org/10.1016/j.envpol.2016.09.046

Barboza LGA, Vethaak AD, Lavorante BRBO, Lundebye A-K, Guilhermino L (2018) Marine microplastic debris: an emerging issue for food security, food safety and human health. *Mar Pollut Bull* 133:336–348

Barboza LGA, Cunha SC, Monteiro C, Fernandes JO, Guilhermino L (2020a) Bisphenol A and its analogs in muscle and liver of fish from the North East Atlantic Ocean in relation to micro-plastic contamination. Exposure and risk to human consumers. *J Hazard Mater* 393:122419

Barboza L, Lopes C, Oliveira P, Bessa F, Otero V, Henriques B, Raimundo J, Caetano M, Vale C, Guilhermino L (2020b) Microplastics in wild fish from North East Atlantic Ocean and its potential for causing neurotoxic effects, lipid oxidative damage, and human health risks associated with ingestion exposure. *Sci Tot Environ* 717:134625

Batel A, Linti F, Scherer M, Erdinger L, Braunbeck T (2016) Transfer of benzo[a]pyrene from microplastics to Artemia nauplii and further to zebrafish via a trophic food web experiment: CYP1A induction and visual tracking of persistent organic pollutants. *Environ Toxicol Chem* 35(7):1656–1666

Bellas J, Martínez-Armental J, Martínez-Cámara A, Besada V, Martínez-Gómez C (2016) Ingestion of microplastics by demersal fish from the Spanish Atlantic and Mediterranean coasts. *Mar Pollut Bull* 109:55–60. https://doi.org/10.1016/j.marpolbul.2016.06.026

Birnstiel S, Soares-Gomes A, da Gama BA (2019) Depuration reduces microplastic content in wild and farmed mussels. *Mar Pollut Bull* 140:241–247

Bour A, Avio CG, Gorbi S, Regoli F, Hylland K (2018) Presence of microplastics in benthic and epibenthic organisms: influence of habitat, feeding mode and trophic level. *Environ Pollut* 243:1217–1225. https://doi.org/10.1016/j.envpol.2018.09.115

Bráte ILN, Eidsvoll DP, Steindal CC, Thomas KV (2016) Plastic ingestion by Atlantic cod (Gadus morhua) from the Norwegian coast. *Mar Pollut Bull* 112:105–110. https://doi.org/10.1016/j.marpolbul.2016.08.034

Bráte ILN, Hurley R, Iversen K, Beyer J, Thomas KV, Steindal CC, Green NW, Olsen M, Lusher A (2018) Mytilus spp. as sentinels for monitoring microplastic pollution in Norwegian coastal waters: a qualitative and quantitative study. *Environ Pollut* 243:383–393

Browne MA, Dissanayake A, Galloway TS, Lowe DM, Thompson RC (2008) Ingested microscopic plastic translocates to the circulatory system of the mussel, Mytilus edulis (L). *Environ Sci Technol* 42:5026–5031

Browne M, Crump P, Niven SJ, Teuten E, Tonkin A, Galloway T, Thompson R (2011) Accumulation of microplastic on shorelines worldwide: sources and sinks. *Environ Sci Technol* 45. https://doi.org/10.1021/es201811s

Carbery M, O'Connor W, Palanisami T (2018) Trophic transfer of microplastics and mixed contaminants in the marine food web and implications for human health. *Environ Int* 115:4. https://doi.org/10.1016/j.envint.2018.03.007

Carpenter EJ, Anderson SJ, Harvey GR, Miklas HP, Peck BB (1972) Polystyrene spherules in coastal waters. *Science* 178(4062):749–750. https://doi.org/10.1126/science.178.4062.749

Carreras-Colom E, Constenla M, Soler-Membrives A, Cartes JE, Baeza M, Padrós F, Carrassón M (2018) Spatial occurrence and effects of microplastic ingestion on the deep-water shrimp Aristeus antennatus. *Mar Pollut Bull* 133:44–52

Catarino AI, Thompson R, Sanderson W, Henry TB (2017) Development and optimization of a standard method for extraction of microplastics in mussels by enzyme digestion of soft tissues. *Environ Toxicol Chem* 36(4):947–951

Catarino AI, Macchia V, Sanderson WG, Thompson RC, Henry TB (2018) Low levels of microplastics (MP) in wild mussels indicate that MP ingestion by humans is minimal compared to exposure via household fibres fallout during a meal. *Environ Pollut* 237:675–684

Cau A, Avio CG, Dessi C, Follesa MC, Moccia D, Regoli F, Pusceddu A (2019) Microplastics in the crustaceans Nephrops norvegicus and Aristeus antennatus: flagship species for deep-sea environments? *Environ Pollut* 255:113107
Cho Y, Shim WJ, Jang M, Han GM, Hong SH (2019) Abundance and characteristics of microplastics in market bivalves from South Korea. *Environ Pollut* 245:1107–1116

Cho Y, Shim WJ, Jang M, Han GM, Hong SH (2020) Nationwide monitoring of microplastics in bivalves from the coastal environment of Korea. *Environ Pollut* 270:116175

Choy CA, Drazen JC (2013) Plastic for dinner? Observations of frequent debris ingestion by pelagic predatory fishes from the central North Pacific. *Mar Ecol Prog Ser* 485:155–163

Cole M, Lindeque P, Halsband C, Galloway SC (2011) Microplastics as contaminants in the marine environment: a review. *Mar Pollut Bull* 62:2588–2597

Collard F, Gasperi J, Gabrielsen GW, Tassin B (2019) Particle ingestion by wild freshwater fish: a critical review. *Environ Sci Technol* 53:12974–12988. https://doi.org/10.1021/acs.est.9b03083

Covernton GA, Collicutt B, Gurney-Smith HJ, Pearce CM, Dower JF, Ross PS, Dudas SE (2019) Microplastics in bivalves and their habitat in relation to shellfish aquaculture proximity in coastal British Columbia, Canada. *Aquac Environ Interact* 11:357–374. https://doi.org/10.3354/aei00316

Cowger W, Booth AM, Hamilton BM, Thaysen C, Primpke S, Munno K, Lusher AL, Dehaut A, Vaz VP, Liboiron M, Devriese LI, Hermabessiere L, Rochman C, Athey SN, Lynch JM, De Frond H, Gray A, Jones OAH, Brander B, Steele C, Moore S, Sanchez A, Nel H (2020) Reporting guidelines to increase the reproducibility and comparability of research on microplastics. *Appl Spectrosc* 74:1066–1077. https://doi.org/10.1177/0003702820930292

Cox KD, Covernton GA, Davies HL, Dower JF, Juanes F, Dudas SE (2019) Human consumption of microplastics. *Environ Sci Technol* 53:7068. https://doi.org/10.1021/acs.est.9b01517

Cózar A, Martí E, Duarte CM, García-de-Lomas J, van Sebille E, Ballatore TJ, Eguíluz VM, González-Gordillo JI, Pedrotti ML, Echevarría F, Troublè R, Irigoien X (2017) The Arctic Ocean as a dead end for floating plastics in the North Atlantic branch of the thermohaline circulation. *Sci Adv* 3(4):e1600582

Critchell K, Hoogenboom MO (2018) Effects of microplastic exposure on the body condition and behaviour of planktivorous reef fish (Acanthochromis polyacanthus). *PLoS One* 13(3):e0193308. https://doi.org/10.1371/journal.pone.0193308

Danopoulos E, Jenner LC, Twiddy M, Rotchell JM (2020) Microplastic contamination of seafood intended for human consumption: a systematic review and meta-analysis. *Environ H Persp* 126002-1. https://doi.org/10.1289/EHP7171

Davidson TM (2012) Boring crustaceans damage polystyrene floats under docks polluting marine waters with microplastic. *Mar Pollut Bull* 64(9):1821–1828

Davidson K, Dudas SE (2016) Microplastic ingestion by wild and cultured Manila clams (Venerupis philippinarum) from Baynes Sound, British Columbia. *Arch Environ Contam Toxicol* 71(2):147–156

De Witte B, Devriese L, Bekaert K, Hoffman S, Vandermeersch G, Cooreman K, Robbens J (2014) Quality assessment of the blue mussel (Mytilus edulis): comparison between commercial and wild types. *Mar Pollut Bull* 85(1):146–155

Dehaut A, Hermabessiere L, Duflos G (2019) Current frontiers and recommendations for the study of microplastics in seafood. *Trends Anal Chem* 116:346–359

Devriese LI, Van der Meulen MD, Maes T, Bekaert K, Paul-Pont I, Frère L, Vethaak AD (2015) Microplastic contamination in brown shrimp (Crangon crangon, Linnaeus 1758) from coastal waters of the southern North Sea and channel area. *Mar Pollut Bull* 98(1–2):179–187

Ding JF, Li JX, Sun CJ, He CF, Jiang FH, Gao FL, Zheng L (2018) Separation and identification of microplastics in digestive system of bivalves. *Chin J Anal Chem* 46(5):690–697

Doyle D, Gammell M, Frias J, Griffin G, Nash R (2019) Low levels of microplastics recorded from the common periwinkle, Littorina littorea on the west coast of Ireland. *Mar Pollut Bull* 149:110645. https://doi.org/10.1016/j.marpolbul.2019.110645

Dris R, Gasperi J, Saad M, Mirande C, Tassin B (2016) Synthetic fibres in atmospheric fall-out: a source of microplastics in the environment? *Mar Pollut Bull* 104:290–293. https://doi.org/10.1016/j.marpolbul.2016.01.006
EFSA (2016) Presence of microplastics and nanoplastics in food, with particular focus on sea-food. Panel on contaminants in the food chain. EFSA J 14:e04501. https://doi.org/10.2903/j.efsa.2016.4501

Eriksen M, Lebreton LC, Carson HS, Thiel M, Moore CJ, Borerro JC, Galgani F, Ryan PG, Reisser J (2014) Plastic pollution in the world's oceans: more than 5 trillion plastic pieces weighing over 250,000 tons afloat at sea. PLoS One 9(12):e111913

Espinosa C, García Beltrán JM, Esteban MA, Cuesta A (2018) In vitro effects of virgin micro-plastics on fish head-kidney leucocyte activities. Environ Pollut 235:30–38. https://doi.org/10.1016/j.envpol.2017.12.054

FAO (2017) The state of food security and nutrition in the world 2017. Building resilience for peace and food security. FAO, Rome. https://doi.org/10.1080/15226514.2012.751351

FAO (2018) The state of world fisheries and aquaculture meeting the sustainable development goals. Food and Agriculture Organization of the United Nations. https://doi.org/10.1364/OE.17.003331

FAO (2020) The state of world fisheries and aquaculture 2020. Sustainability in action. Rome. doi: https://doi.org/10.4060/ca9229en

Farrell P, Nelson K (2013) Trophic level transfer of microplastic: Mytilus edulis (L.) to Carcinus maenas (L.). Environ Pollut 177:1–3

Feng S, Gao D, Liao F, Zhou F, Wang X (2016) The health effects of ambient PM2.5 and potential mechanisms. Ecotoxicol Environ Saf 128:67–74. https://doi.org/10.1016/j.ecoenv.2016.01.030

Feng Z, Wang R, Zhang T, Wang J, Huang W, Li J, Xu J, Gao G (2020) Microplastics in specific tissues of wild sea urchins along the coastal areas of northern China. Sci Total Environ 728:138660. https://doi.org/10.1016/j.scitotenv.2020.138660

Foekema EM, De Gruijter C, Mergia MT, van Franeker JA, Murk AJ, Koelmans AA (2013) Plastic in north sea fish. Environ Sci Technol 47(15):8818–8824. https://doi.org/10.1021/es400931b

Franzellitti S, Canesi L, Auguste M, Wathsala RHGR, Fabbri E (2019) Microplastic exposure and effects in aquatic organisms: a physiological perspective. Environ Toxicol Pharmacol 68:37–51

Gall SC, Thompson RC (2015) The impact of debris on marine life. Mar Pollut Bull 92:170–179

Galloway TS (2015) Micro- and nano-plastics and human health. In: Marine Anthropogenic Litter. https://doi.org/10.1007/978-3-319-16510-3_13

Galloway TS, Cole M, Lewis C (2017) Interactions of microplastic debris throughout the marine ecosystem. Nat Ecol Evol 1:1–8

Gamarro EG, Ryder J, Elvevoll EO, Olsen RL (2020) Microplastics in fish and shellfish – a threat to seafood safety? J Aquat Food Product Technol 29:417–425. https://doi.org/10.1080/10498850.2020.1739793

Gardon T, Reisser C, Soyez C, Quillien V, Le Moullac G (2018) Microplastics affect energy balance and gametogenesis in the pearl oyster Pinctada margaritifera. Environ Sci Technol 52(9):5277–5286

GESAMP (2016) Sources, fate and effects of microplastics in the marine environment: part two of a global assessment. In: Kershaw PJ, Rochman CM (eds) Joint Group of Experts on the Scientific Aspects of Marine Environmental Protection, Rep. Stud. GESAMP No. 93. 220 p

GESAMP (2019) Guidelines or the monitoring and assessment of plastic litter and microplastics in the ocean. In: Kershaw PJ, Turra A, Galgani F (eds) IMO/FAO/UNESCO-IOC/UNIDO/WMO/IAEA/UN/UNEP/UNDP/ISA Joint Group of Experts on the Scientific Aspects of Marine Environmental Protection, Rep. Stud. GESAMP No. 99. 130p

Gouin T, Roche N, Lohmann R, Hodges G (2011) A thermodynamic approach for assessing the environmental exposure of chemicals absorbed to microplastic. Environ Sci Technol 45:1466–1472

Graham ER, Thompson JT (2009) Deposit- and suspension-feeding sea cucumbers (Echinodermata) ingest plastic fragments. J Exp Mar Biol Ecol 368:22–29. https://doi.org/10.1016/j.jembe.2008.09.007
Green DS, Colgan TJ, Thompson RC, Carolan JC (2019) Exposure to microplastics reduces attachment strength and alters the haemolymph proteome of blue mussels (Mytilus edulis). *Environ Pollut* 246:423–434

Güven O, Gökdağ K, Jovanović B, Kıdeys AE (2017) Microplastic litter composition of the Turkish territorial waters of the Mediterranean Sea, and its occurrence in the gastrointestinal tract of fish. *Environ Pollut* 223:286–294. https://doi.org/10.1016/j.envpol.2017.01.025

Hantoro I, Löhr AJ, Van Belleghem F, Widianarko B, Ragas A (2019) Microplastics in coastal areas and seafood: implications for food safety. Food Addit Contam 36(5):674–711

Hara J, Frias J, Nash R (2020) Quantification of microplastic ingestion by the decapod crustacean Nephrops norvegicus from Irish waters. Mar Pollut Bull 152:110905

Hartmann NB, Hüffer T, Thompson RC, Hassellöv M, Verschoor A, Daugaard AE, Rist S, Karlsson T, Brennholt N, Cole M, Herrling MP, Hess MC, Ivleva NP, Lusher AL, Wagner M (2019) Are we speaking the same language? Recommendations for a definition and categorization framework for plastic debris. *Environ Sci Technol* 53(3):1039–1047

Hermsen E, Mintenig SM, Besseling E, Koelmans AA (2018) Quality criteria for the analysis of microplastic in biota samples: a critical review. *Environ Sci Technol* 52(18):10230–10240

Hidalgo-Ruz V, Gutow L, Thompson RC, Thiel M (2012) Microplastics in the marine environment: a review of the methods used for identification and quantification. *Environ Sci Technol* 46:3060–3075

Hurley R, Woodward J, Rothwell JJ (2018) Microplastic contamination of river beds significantly reduced by catchment-wide flooding. *Nat Geosci* 11:251–257. https://doi.org/10.1038/s41561-018-0080-1

Iwalaye OA, Moodley GK, Robertson-Andersson DV (2020) The possible routes of microplastics uptake in sea cucumber Holothuria cinerascens (Brandt, 1835). *Environ Pollut* 264:114644. https://doi.org/10.1016/j.envpol.2020.114644

Jaafar N, Musa SM, Afzaralariff A, Mohamed M, Yusoff AH, Lazim AM (2020) Improving the efficiency of post-digestion method in extracting microplastics from gastrointestinal tract and gills of fish. *Chemosphere* 260:127649

Jang M, Shim WJ, Han GM, Song YK, Hong SH (2018) Formation of microplastics by poly-chaetes (Marphysa sanguinea) inhabiting expanded polystyrene marine debris. *Mar Pollut Bull* 131:365–369

Jiang Y, Yang F, Zhao Y, Wang J (2020) Greenland Sea Gyre increases microplastic pollution in the surface waters of the Nordic Seas. *Sci Total Environ* 712:136484

Kane IA, Clare MA, Miramontes E, Wogelius R, Rothwell JJ, Garreau P, Pohl F (2020) Seafloor microplastic hotspots controlled by deep-sea circulation. *Science* 368:1140–1145

Karbalaei S, Hanachi P, Walker TR, Cole M (2018) Occurrence, sources, human health impacts and mitigation of microplastic pollution. *Environ Sci Pollut Res Int* 25:36046–36063. https://doi.org/10.1007/s11356-018-3508-7

Karlsson TM, Vethaak AD, Almroth BC, Ariese F, van Velzen M, Hassellöv M, Leslie HA (2017) Screening for microplastics in sediment, water, marine invertebrates and fish: method development and microplastic accumulation. *Mar Pollut Bull* 122(1–2):403–408

Kiew PL, Don MM (2011) Jewel of the seabed: sea cucumbers as nutritional and drug candidates. *Int J Food Sci Nutr* 63(5):616–636. https://doi.org/10.3109/09637486.2011.641944

Koelmans AA, Bakir A, Burton GA, Janssen CR (2016) Microplastic as a vector for chemicals in the aquatic environment: critical review and model-supported reinterpretation of empirical studies. *Environ Sci Technol* 50(7):3315–3326. https://doi.org/10.1021/acs.est.5b06069

Kögel T, Bjorøy Ø, Toto B, Bienfait AM, Sanden M (2019) Micro- and nanoplastic toxicity on aquatic life: determining factors. *Sci Total Environ* 709:136050. https://doi.org/10.1016/j.scitotenv.2019.136050

Kolandhasamy P, Su L, Li J, Qu X, Jabeen K, Shi H (2018) Adherence of microplastics to soft tissue of mussels: a novel way to uptake microplastics beyond ingestion. *Sci Total Environ* 610-611:635–640. https://doi.org/10.1016/j.scitotenv.2017.08.053
Koongolla JB, Lin L, Pan Y-F, Yang C-P, Sun D-R, Liu S, Xu X-R, Maharana D, Huang J-S, Li H-X (2020) Occurrence of microplastics in gastrointestinal tracts and gills of fish from Beibu Gulf, South China Sea. *Environ Pollut* 258:113734. https://doi.org/10.1016/j.envpol.2019.113734

Kwon J-H, Kim J-W, Pham TD, Tarafdar A, Hong S, Chun S-H, Lee S-H, Kang D-Y, Kim J-Y, Kim S-B, Jung J (2020) Microplastics in food: a review on analytical methods and challenges. *Int J Environ Res Public Health* 17(18):6710. https://doi.org/10.3390/ijerph17186710

Lebreton L, van der Zwet J, Damsteeg J, Slat B, Andrady A, Reisser J (2017) River plastic emissions to the world's oceans. *Nat Commun* 8:15611. https://doi.org/10.1038/ncomms15611

Li J, Yang D, Li L, Jabeen K, Shi H (2015) Microplastics in commercial bivalves from China. *Environ Pollut* 207:190-195

Li J, Qu X, Su L, Zhang W, Yang D, Kolandhasamy P, Shi H (2016) Microplastics in mussels along the coastal waters of China. *Environ Pollut* 214:177-184

Li J, Lusher AL, Rotchell JM, Deudero S, Turra A, Bråte ILN, Sun C, Hossain MS, Li Q, Kolandhasamy P, Shi H (2019) Using mussel as a global bioindicator of coastal microplastic pollution. *Environ Pollut* 244:522-533

Liboiron M, Liboiron F, Wells E, Richárd N, Zahara A, Mather C, Bradshaw H, Murichi J (2016) Low plastic ingestion rate in Atlantic cod (Gadus morhua) from Newfoundland destined for human consumption collected through citizen science methods. *Mar Pollut Bull* 113(1-2):428-437. https://doi.org/10.1016/j.marpolbul.2016.10.043

Lithner D, Larsson A, Dave G (2011) Environmental and health hazard ranking and assessement of plastic polymers based on chemical composition. *Sci Total Environ* 409:3309-3324

Lohmann R (2017) Microplastics are not important for the cycling and bioaccumulation of organic pollutants in the oceans-but should microplastics be considered POPs themselves? *Integr Environ Assess Manag* 13(3):460-465. https://doi.org/10.1002/ieam.1914

Lusher A, Welden NAC (2020) Microplastic impact in fisheries and aquaculture. In: Rocha-Santos T et al (eds) Handbook of microplastics in the environment. Springer, Cham

Lusher AL, McHugh M, Thompson RC (2013) Occurrence of microplastics in the gastrointestinal tract of pelagic and demersal fish from the English Channel. *Mar Pollut Bull* 67:94-99. https://doi.org/10.1016/j.marpolbul.2012.11.028

Lusher AL, O'Donnell C, Officer R, O'Connor I (2016) Microplastic interactions with North Atlantic mesopelagic fish. *ICES J Mar Sci* 73:1214-1225

Lusher A, Hollman P, Mendoza-Hill J (2017a) Microplastics in fisheries and aquaculture – status of knowledge on their occurrence and implications for aquatic organisms and food safety. *FAO Fisheries and Aquaculture Technical Paper*, 615

Lusher AL, Welden NA, Sobral P, Cole M (2017b) Sampling, isolating and identifying microplastics ingested by fish and invertebrates. *Anal Methods* 9(9):1346-1360

Lusher AL, Munno K, Hermabessiere L, Carr S (2020) Isolation and extraction of microplastics from environmental samples: an evaluation of practical approaches and recommendations for further harmonization. *Appl Spectrosc* 74(9):1049-1065

Markic A, Niemand C, Bridson JH, Mazouni-Gaertner N, Gaertner JC, Eriksen M, Bowen M (2018) Double trouble in the South Pacific subtropical gyre: increased plastic ingestion by fish in the oceanic accumulation zone. *Mar Pollut Bull* 136:547-564. https://doi.org/10.1016/j.marpolbul.2018.09.031

Mathalon A, Hill P (2014) Microplastic fibers in the intertidal ecosystem surrounding Halifax Harbor, Nova Scotia. *Mar Pollut Bull* 81(1):69-79

McGoran AR, Clark PF, Smith BD, Morritt D (2020) High prevalence of plastic ingestion by Eriocheir sinensis and Carcinus maenas (Crustacea: Decapoda: Brachyura) in the Thames estuary. *Environ Pollut* 265:114972

Mohsen M, Wang Q, Zhang L, Sun L, Lin C, Yang H (2019) Microplastic ingestion by the farmed sea cucumber Apostichopus japonicus in China. *Environ Pollut* 245:1071-1078. https://doi.org/10.1016/j.envpol.2018.11.083

Naji A, Nuri M, Vethaak AD (2018) Microplastics contamination in molluscs from the northern part of the Persian Gulf. *Environ Pollut* 235:113-120. https://doi.org/10.1016/j.envpol.2017.12.046
Napper IE, Thompson RC (2016) Release of synthetic microplastic plasticfibres fromdomestic washing machines: effects of fabric type and washing conditions. Mar Pollut Bull 112:39–45. https://doi.org/10.1016/j.marpolbul.2016.09.025

Nelms SE, Galloway TS, Godley BJ, Jarvis DS, Lindeque PK (2018) Investigating microplastic trophic transfer in marine top predators. Environ Pollut 238:999–1007. https://doi.org/10.1016/j.envpol.2018.02.016

Neves D, Sobral P, Ferreira JL, Pereira T (2015) Ingestion of microplastics by commercial fish off the Portuguese coast. Mar Pollut Bull 101(1):119–126. https://doi.org/10.1016/j.marpolbul.2015.11.008

Oberbeckmann SL, Martin GJ, Labrenz M (2015) Marine microplastic-associated biofilms – a review. Environ Chem 12(5):551–562. https://doi.org/10.1071/EN15069

Ogonowski M, Wenman V, Danielsson S Gorokhova E (2017) Ingested microplastic is not correlated to HOC concentrations in Baltic Sea herring. In: Proceedings of the 15th international conference on environmental science and technology, Rhodes, Greece, 31 August–2 September

Oliveira AR, Sardinha-Silva A, Andrews PLR, Green D, Cooke GM, Hall S, Blackburn K, Sykes AV (2020) Microplastics presence in cultured and wild-caught cuttlefish, Sepia officinalis. Mar Pollut Bull 160:111553. https://doi.org/10.1016/j.marpolbul.2020.111553

Peng L, Fu D, Qi H, Lan CQ, Yu H, Ge C (2020) Micro- and nano-plastics in marine environment: source, distribution and threats—a review. Sci Total Environ 698:134254. https://doi.org/10.1016/j.scitotenv.2019.134254

Phuong NN, Poirier L, Pham QT, Lagarde F, Zalouk-Vergnoux A (2018) Factors influencing the microplastic contamination of bivalves from the French Atlantic coast: location, season and/or mode of life? Mar Pollut Bull 129(2):664–674

Pitt JA, Trevisan R, Massarsky A, Kozal JS, Levin ED, Di Giulio RT (2018) Maternal transfer of nanoplastics to offspring in zebrafish (*Danio rerio*): a case study with nanopolystyrene. Sci Total Environ 643:324–334. https://doi.org/10.1016/j.scitotenv.2018.06.186

Plee TA, Pomory CM (2020) Microplastics in sandy environments in the Florida keys and the panhandle of Florida, and the ingestion by sea cucumbers (Echinodermata: Holothuroidea) and sand dollars (Echinodermata: Echinoidea). Mar Pollut Bull 158:111437. https://doi.org/10.1016/j.marpolbul.2020.111437

Porter A, Lyons BP, Galloway TS, Lewis C (2018) Role of marine snows in microplastic fate and bioavailability. Environ Sci Technol 52:7111–7119

Qu X, Su L, Li H, Liang M, Shi H (2018) Assessing the relationship between the abundance and properties of microplastics in water and in mussels. Sci Total Environ 621:679–686

Renzi M, Guerranti C, Blašković A (2018a) Microplastic contents from maricated and natural mussels. Mar Pollut Bull 131:248–251

Renzi M, Blašković A, Bernardi G, Russo GF (2018b) Plastic litter transfer from sediments towards marine trophic webs: a case study on holothurians. Mar Pollut Bull 135:376–385. https://doi.org/10.1016/j.marpolbul.2018.07.038

Ribeiro F, Okoffo ED, O'Brien JW, Fraissinet-Tachet S, O'Brien S, Gallen M, Samanipour S, Kaserzon S, Mueller JF, Galloway T, Thomas KV (2020) Quantitative analysis of selected plastics in high-commercial-value Australian seafood by pyrolysis gas chromatography mass spectrometry. Environ Sci Technol 54:9408–9417

Rist S, Almroth BC, Hartmann NB, Karlsson TM (2018) A critical perspective on early communications concerning human health aspects of microplastics. Sci Total Environ 626:720–726

Roch S, Friedrich C, Brinker A (2020) Uptake routes of microplastics in fishes: practical and theoretical approaches to test existing theories. Sci Rep 10:3896

Rochman CM, Hoh E, Kurobe T, The SJ (2013) Ingested plastic transfers hazardous chemicals to fish and induces hepatic stress. Sci Rep 3:3263. https://doi.org/10.1038/srep03263

Rochman CM, Tahir A, Williams SL, Baxa DV, Lam R, Miller JT, The F-C, Werorilangi S, Teh SJ (2015) Anthropogenic debris in seafood: plastic debris and fibers from textiles in fish and bivalves sold for human consumption. Sci Rep 5:14340
Rochman CM, Brookson C, Bikker J, Djuric N, Earn A, Bucci K, Athey S, Huntington A, McIlwraith H, Munno K, De Frond H, Kolomijeca A, Erdle L, Grbic J, Bayoumi M, Borrelle SB, Wu T, Santoro S, Werbowski LM, Zhu X, Giles RK, Hamilton BM, Thaysen C, Kaura A, Klasios N, Ead L, Kim J, Sherlock C, Ho A, Hung C (2019) Rethinking microplastics as a diverse contaminant suite. Environ Toxicol Chem 38:703–711

Rummel CD, Löder MG, Fricke NF, Lang T, Griebeler EM, Janke M, Gerdts G (2016) Plastic ingestion by pelagic and demersal fish from the North Sea and Baltic Sea. Mar Pollut Bull 102(1):134–141. https://doi.org/10.1016/j.marpolbul.2015.11.043

Santillo D, Miller K, Johnston P (2017) Microplastics as contaminants in commercially important seafood species. Integr Environ Assess Manag 13:516–521

Schwabl P, Koppel S, Königshofer P, Bucsics T, Trauner M, Reiberger T, Liebmann B (2019) Detection of various microplastics in human stool: a prospective case series. Ann Intern Med. https://doi.org/10.7326/M19-0618

Setälä O, Fleming-Lehtinen V, Lehtiniemi M (2014) Ingestion and transfer of microplastics in the planktonic food web. Environ Pollut 185:77–83

Smith M, Love DC, Rochman CM, Neff RA (2018) Microplastics in seafood and the implications for human health. Curr Environ Health Rep 5:375. https://doi.org/10.1007/s40572-018-0206-z

Stone V, Johnston H, Clift MJD (2007) Air pollution, ultrafine and nanoparticle toxicology: cellular and molecular interactions. IEEE Trans Nanobioscience 6(4):331–340. https://doi.org/10.1109/TNB.2007.909005

Su L, Deng H, Li B, Chen Q, Pettigrove V, Wu C, Shi H (2018) The occurrence of microplastic in specific organs in commercially caught fishes from coast and estuary area of East China. J Hazard Mater. https://doi.org/10.1016/j.jhazmat.2018.11.024

Sussarellu R, Suquet M, Thomas Y, Lambert C, Fabioux C, Pernet ME, Le Goïc N, Quillien V, Mingant C, Epelboin Y, Corporeau C, Guyomarch J, Robbens J, Paul-Pont I, Soudant P, Huvet A (2016) Oyster reproduction is affected by exposure to polystyrene microplastics. Proc Natl Acad Sci U S A 113(9):2430–2435. https://doi.org/10.1073/pnas.1519019113

Tanaka K, Takada H (2016) Microplastic fragments and microbeads in digestive tracts of planktivorous fish from urban coastal waters. Sci Rep 6:34351. https://doi.org/10.1038/srep34351

Thiele CJ, Hudson MD, Russell AE (2019) Evaluation of existing methods to extract microplastics from bivalve tissue: adapted KOH digestion protocol improves filtration at single-digit pore size. Mar Pollut Bull 142:384–393

Thushari G, Senevirathna J (2020) Plastic pollution in the marine environment. Heliyon 6(8):e04709. https://doi.org/10.1016/j.heliyon.2020.e04709

Thushari GGN, Senevirathna JDM, Yakupitiyage A, Chavanich S (2017) Effects of microplastics on sessile invertebrates in the eastern coast of Thailand: an approach to coastal zone conservation. Mar Pollut Bull 124:349–355. https://doi.org/10.1016/j.marpolbul.2017.06.010

Toussaint B, Raffael B, Angers-Loustau A, Gilliland D, Kestens V, Petrillo M, Rio-Echevarria IM, Van den Eede G (2019) Review of micro- and nanoplastic contamination in the food chain. Food Addit Contam A 36:639. https://doi.org/10.1080/19440049.2019.1583381

van Cauwenberghe L, Janssen CR (2014) Microplastics in bivalves cultured for human consumption. Environ Pollut 193:65–70

van Cauwenberghe L, Claessens M, Vandegehuchte MB, Janssen CR (2015) Microplastics are taken up by mussels (Mytilus edulis) and lugworms (Arenicola marina) living in natural habitats. Environ Pollut 199:10–17

van Raamsdonk LWD, van der Zande M, Koelmans AA, Hoogenboom RLAP, Peters RJB, Groot MJ, Peijnenburg ACM, Weesepoel YJA (2020) Current insights into monitoring, bioaccumulation, and potential health effects of microplastics present in the food chain. Foods. https://doi.org/10.3390/foods9010072

van Sebille E, Aliani S, Law KL, Maximenko N, Alsina JM, Bagaev A, Bergmann M, Chapron B, Chubarenko I, Cózar A, Delandmeter P, Egger M, Fox-Kemper B, Garaba SP, Goddijn-Murphy L, Hardesty BD, Hoffman MJ, Isobe A, Jongedijk CE, Kaandorp MLA, Khatmullina L, Koelmans AA, Kukulka T, Laufkötter C, Lebreton L, Lobelle D, Maes C, Martinez-
Vicente V, Morales Maqueda MA, Poulain-Zarcos M, Rodríguez E, Ryan PG, Shanks AL, Shim WJ, Suaria G, Thiel M, van den Bremer TS, Wichmann D (2020) The physical oceanography of the transport of floating marine debris. Environ Res Lett 15:023003. https://doi.org/10.1088/1748-9326/ab6d7d

Vandermeersch G, Van Cauwenberghe L, Janssen CR, Marques A, Granby K, Fait G, Kotterman MJ, Diogène J, Bekaert K, Robbins J, Devriese L (2015) A critical view on microplastic quantification in aquatic organisms. Environ Res 143:46–55. https://doi.org/10.1016/j.envres.2015.07.016

VKM (2019) Microplastics; occurrence, levels and implications for environment and human health related to food. Scientific opinion of the Scientific Steering Committee of the Norwegian Scientific Committee for Food and Environment (No. 2019:16). Norwegian Scientific Committee for Food and Environment (VKM), Oslo, Norway

Walkinshaw C, Lindeque PK, Thompson R, Tolhurst T, Cole M (2020) Microplastics and seafood: lower trophic organisms at highest risk of contamination. Ecotoxicol. Environ Saf 190. https://doi.org/10.1016/j.ecoenv.2019.110066

Wang J, Lu L, Wang M, Jiang T, Liu X, Ru S (2019) Typhoons increase the abundance of microplastics in the marine environment and cultured organisms: a case study in Sanggou Bay, China. Sci Total Environ 667:1–8. https://doi.org/10.1016/j.scitotenv.2019.02.367

Wang W, Ge J, Yu X (2020) Bioavailability and toxicity of microplastics to fish species: a review. Ecotoxicol Environ Saf 189:109913. https://doi.org/10.1016/j.ecoenv.2019.109913

Ward JE, Kach DJ (2009) Marine aggregates facilitate ingestion of nanoparticles by suspension-feeding bivalves. Mar Environ Res 68(3):137–142

Ward JE, Rosa M, Shumway SE (2018) Capture, ingestion, and egestion of microplastics by suspension-feeding bivalves: a 40-year history. Mar Microplast Pollut Contr 2(1):39–49. https://doi.org/10.1139/anc-2018-0027@anc-mmpc issue1

Ward JE, Zhao S, Holohan BA, Mladinich KM, Griffin TW, Wozniak J, Shumway SE (2019) Selective ingestion and egestion of plastic particles by the blue mussel (Mytilus edulis) and eastern oyster (Crassostrea virginica): implications for using bivalves as bioindicators of microplastic pollution. Environ Sci Technol 53(15):8776–8784

Wardrop P, Shimeta J, Nugegoda D, Morrison PD, Miranda A, Tang M, Clarke BO (2016) Chemical pollutants sorbed to ingested microbeads from personal care products accumulate in fish. Environ Sci Technol 50:4037–4044. https://doi.org/10.1021/acs.est.5b06280

Watts AJR, Urbina MA, Corr S, Lewis C, Galloway T (2015) Ingestion of plastic microfibers by the crab Carcinus maenas and its effect on food consumption and energy balance. Environ Sci Technol 49(24):14597–14604. https://doi.org/10.1021/acs.est.5b04026

Watts AJR, Urbina MA, Goodhead R, Moger J, Lewis C, Galloway TS (2016) Effect of microplastic on the gills of the shore crab Carcinus maenas. Environ Sci Technol 50:5364–5369. https://doi.org/10.1021/acs.est.6b01187

Wegner A, Besseling E, Foekema EM, Kamermans P, Koelmans AA (2012) Effects of nanopoly-styrene on the feeding behavior of the blue mussel (Mytilus edulis L.). Environ Toxicol Chem 31(11):2490–2497

Weinstein JE, Crocker BK, Gray AD (2016) From macroplastic to microplastic: degradation of high-density polyethylene, polypropylene, and polystyrene in a salt marsh habitat. Environ Toxicol Chem 35:1632–1640. https://doi.org/10.1002/etc.3432

Welden NA, Cowie PR (2016) Environment and gut morphology influence microplastic retention in langoustine, Nephrops norvegicus. Environ Pollut 214:859–865

Welden NA, Abylkhani B, Howarth LM (2018) The effects of trophic transfer and environmental factors on microplastic uptake by plaice, Pleuronectes plastessa, and spider crab, Maja squinado. Environ Pollut 239:351–358

Wieczorek AM, Croot PL, Lombard F, Sheahan JN, Doyle TK (2019) Microplastic ingestion by gelatinous zooplankton may lower efficiency of the biological pump. Environ Sci Technol 53(9):5387–53955. https://doi.org/10.1021/acs.est.8b07174
Wright SL, Kelly FJ (2017) Plastic and human health: a micro issue? *Environ Sci Technol* 51(12):6634–6647. https://doi.org/10.1021/acs.est.7b00423

Wright SL, Thompson RC, Galloway TS (2013) The physical impacts of microplastics on marine organisms: a review. *Environ Pollut* 178:483–492

Xu XY, Lee WT, Chan AKY, Lo HS, Shin PKS, Cheung SG (2017) Microplastic ingestion reduces energy intake in the clam *Actactodea striata*. *Mar Pollut Bull* 124(2):798–802

Xu X, Wong CY, Tam NFY, Lo H-S, Cheung S-G (2020) Microplastics in invertebrates on soft shores in Hong Kong: influence of habitat, taxa and feeding mode. *Sci Tot Environ* 715:13699. https://doi.org/10.1016/j.scitotenv.2020.136999

Yoo JW, Doshi N, Mitragotri S (2011) Adaptive micro and nanoparticles: temporal control over carrier properties to facilitate drug delivery. *Adv Drug Deliv Rev* 63(14–15):1247–1256

Yu Z, Peng B, Liu LY, Wong CS, Zeng EY (2019) Development and validation of an efficient method for processing microplastics in biota samples. *Environ Toxicol Chem* 38(7):1400–1408

Zeytin S, Wagner G, Mackay-Roberts N, Gerdts G, Schuirmann E, Klockmann S, Slater M (2020) Quantifying microplastic translocation from feed to the fillet in European sea bass *Dicentrarchus labrax*. *Mar Pollut Bull* 156:111210

Zhang F, Man YB, Mo WY, Man KY, Wong MH (2019a) Direct and indirect effects of microplastics on bivalves, with a focus on edible species: A mini-review. *Crit Rev Environ Sci Technol* 1–35. https://doi.org/10.1080/10643389.2019.1700752

Zhang F, Wang X, Xu J, Zhu L, Peng G, Xu P, Li D (2019b) Food-web transfer of microplastics between wild caught fish and crustaceans in East China Sea. *Mar Pollut Bull* 146:173–182. https://doi.org/10.1016/j.marpolbul.2019.05.061

Ziccardi LM, Edgington A, Hentz K, Kulacki KJ, Driscoll SK (2016) Microplastics as vectors for bioaccumulation of hydrophobic organic chemicals in the marine environment: a state-of-the-science review. *Environ Toxicol Chem* 35:1667–1676

**Open Access** This chapter is licensed under the terms of the Creative Commons Attribution 4.0 International License (http://creativecommons.org/licenses/by/4.0/), which permits use, sharing, adaptation, distribution and reproduction in any medium or format, as long as you give appropriate credit to the original author(s) and the source, provide a link to the Creative Commons license and indicate if changes were made.

The images or other third party material in this chapter are included in the chapter's Creative Commons license, unless indicated otherwise in a credit line to the material. If material is not included in the chapter's Creative Commons license and your intended use is not permitted by statutory regulation or exceeds the permitted use, you will need to obtain permission directly from the copyright holder.

<figure><img src="image_28.png" /></figure>
<figure><img src="image_29.png" /></figure>

Check for
updates

**Chapter 6**

**Weight of Evidence for the Microplastic**
**Vector Effect in the Context of Chemical**
**Risk Assessment**

**Albert A. Koelmans, Noël J. Diepens, and Nur Hazimah Mohamed Nor**

**Abstract** The concern that in nature, ingestion of microplastic (MP) increases exposure of organisms to plastic-associated chemicals (the ‘MP vector effect’) plays an important role in the current picture of the risks of microplastic for the environment and human health. An increasing number of studies on this topic have been conducted using a wide variety of approaches and techniques. At present, the MP vector effect is usually framed as ‘complex’, ‘under debate’ or ‘controversial’. Studies that critically discuss the approaches and techniques used to study the MP vector effect, and that provide suggestions for the harmonization needed to advance this debate, are scarce. Furthermore, only a few studies have strived at interpreting study outcomes in the light of environmentally relevant conditions. This constitutes a major research gap, because these are the conditions that are most relevant when informing risk assessment and management decisions. Based on a review of 61 publications, we propose evaluation criteria and guidance for MP vector studies and discuss current study designs using these criteria. The criteria are designed such that studies, which fulfil them, will be relevant to inform risk assessment. By critically reviewing the existing literature in the light of these criteria, a weight of evidence assessment is provided. We demonstrate that several studies did not meet the standards for their conclusions on the MP vector effect to stand, whereas others provided overwhelming evidence that the vector effect is unlikely to affect chemical risks under present natural conditions.

A. A. Koelmans (✉) · N. J. Diepens · N. H. Mohamed Nor

Aquatic Ecology and Water Quality Management Group, Wageningen University & Research,
Wageningen, The Netherlands

e-mail: bart.koelmans@wur.nl; noel.diepens@wur.nl; Hazimah.mohamednor@wur.nl
## 6.1 Introduction

Plastics in the environment contain mixtures of chemicals. These chemicals stem from deliberate additions during manufacture of the plastic or from the ambient water via absorption or both. Plastic particles are known to be ingested by organisms, including humans (World Health Organization 2019), but not all particles can be ingested by all organisms. Exposure, bioavailability, feeding behaviour and size of the plastic item in relation to size of the mouth opening determine whether a plastic item can be taken up (Jâms et al. 2020; Koelmans et al. 2020). Plastic is persistent, and ingestion of sufficiently high particle concentrations has been demonstrated to cause effects on small organisms like some zooplankton and invertebrates, most likely due to the reduced caloric value of the ingested material (Gerdes et al. 2019a; de Ruijter et al. 2020). Besides this mechanism of reduced caloric value due to dilution of ingested material, concerns have been raised due to the fact that chemicals on the plastic are ingested with the plastic. This has led to an increasing number of studies that investigate the potential of microplastic to increase exposure of organisms to plastic-associated chemicals. Henceforth, we will refer to these studies as MP vector studies. MP vector studies have used a wide variety of approaches and techniques and have been summarized in several reviews (Rochman 2015, 2019; Koelmans et al. 2016; Ziccardi et al. 2016; Hartmann et al. 2017; Burns and Boxall 2018). However, none of these have focused on quality assurance and study design criteria.

In the recent literature, MP vector studies are often framed as ‘controversial’, ‘complex’ or ‘under debate’ (Gassel and Rochman 2019). This suggests that apparently there is no consensus in the scientific community on the nature or the relevance of the MP vector effect. We argue that there are three main reasons why this could be the case, reasons that may explain why it takes so long before consensus is reached.

The first reason relates to confusion about when a study is to be considered relevant in this context, a question which in the literature on chemical risks often is referred to as the ‘so what’ question. To date, studies have addressed detailed mechanisms, specific exposure scenarios in the lab, have either measured effects on uptake or on biological endpoints in the lab and have evaluated uptake under natural conditions either by field studies or by model scenario analysis. Studies mostly addressed whether chemicals *can (potentially)* be taken up from plastic under some specific conditions, but not if they *will* be taken up under natural conditions. Few studies have strived at interpreting results in the context of environmentally relevant conditions, and none of them addressed to what extent ingestion of MP would actually increase the chemical risks for organisms. An increase in chemical risk would exist if, due to the MP vector effect, exposure to chemicals (e.g., predicted exposure concentrations (PEC)) would exceed the toxicity thresholds that are known for these chemicals (predicted no effect concentration (PNEC)). Putting study results in the context of such actual risks is relevant, because if a detected vector effect would not increase risks of the chemicals, then it might be less important to address it and
there would be less reason for concern. With no studies actually demonstrating the
occurrence of an MP vector effect on chemical risks under natural conditions, the
evidence base is remarkably thin. This constitutes a research gap, because these are
the conditions that would be most relevant when informing risk assessment and
management decisions. We thus propose relevance for chemical risk assessment as
an overarching umbrella criterion to evaluate the setup of MP vector studies.

The second reason is that the discussion of the topic may have lacked a common
understanding of the processes at play, especially in the earlier studies. Most of
them address a part of the exposure conditions that are relevant to the occurrence of
MP vector effects and their possible implications for effects on risks in the environ-
ment (e.g. Rochman et al. 2013; Chua et al. 2014; Wardrop et al. 2016; Granby et al.
2018; Kühn et al. 2020). Only few studies use theoretical frameworks to add greater
depth to the interpretation of study outcomes. Such frameworks allow for inter- and
extrapolation across chemical properties, microplastic characteristics, biological
traits and environmental conditions (Koelmans et al. 2016; Bakir et al. 2016;
Rochman et al. 2017; Lee et al. 2019; Mohamed Nor and Koelmans 2019; Wang
et al. 2020a, b). As these frameworks often come in the form of mathematical mod-
els, they are not automatically adopted by researchers that use laboratory or field
observational studies as their primary research tools.

The third reason relates to the general quality of some of the microplastic
research, which has been framed as limiting in recent literature (Lenz et al. 2016;
Hermsen et al. 2018; Koelmans et al. 2019; Markic et al. 2019; de Ruijter et al.
2020; Provencher et al. 2020). Microplastic research is a young and fast-growing
field in the environmental sciences, and this is reflected in the wide variety of
approaches used when sampling and analysing microplastic and in effect testing
methods. The same diversity and therefore incomparability of approaches apply to
MP vector studies.

In this review, we discuss the potential of studies and study designs to inform risk
assessment of plastic-associated chemicals, emphasizing organic chemicals. Heavy
metals bind to microplastic particles as well, but their sorption affinity is limited
compared to that for sorption to natural particles like sediment (e.g. Besson et al.
2020). First, we propose evaluation criteria for MP vector studies. The criteria are
designed such that studies that fulfil them can be considered relevant to inform risk
assessment. Second, we critically review the existing literature in the light of these
criteria, thereby providing a weight of evidence assessment. Finally, we discuss
several key references from the literature as examples of how they can be used to
inform plastic-associated chemical risk assessment given the criteria and
recommendations.

Literature was selected through reference as well as cited reference searches for
the aforementioned six existing MP vector effect reviews. This method assumes that
all relevant MP vector studies either cite or are cited (in) at least one of these six
reviews, published between 2015 and 2019.

We emphasize that we reviewed the existing literature only for the aim set for the
present review: to assess the weight of evidence for the chemical vector effect to
occur under environmentally relevant conditions and with respect to implications
for chemical risks. Part of the reviewed papers may not have had this specific aim. They often were performed to verify the validity of underlying mechanisms, or they were meant to study potential effects rather than to demonstrate actual chemical risks under field conditions. Therefore, the present retrospective assessment does not necessarily disqualify studies as such, as they could have had other aims than mimicking the environmental realism with respect to the chemical risks stemming from the MP vector effect that we aim to pursue here.

## 6.2 Guidance for Microplastic Vector Studies in the Context of Chemical Risk Assessment

Several criteria can be used when interpreting results from studies investigating the MP vector effect with respect to their relevance for conditions occurring in nature (Table 6.1). The criteria can be placed into two broad categories: (a) characteristics of a study in providing evidence for the occurrence of an MP vector effect in nature and (b) relevance of the setup and outcomes for risk assessment of chemically contaminated MP. The difference is that the latter category would need reflection on whether effect thresholds for chemical toxicity are exceeded. For instance, if an MP vector effect is detected, concentrations of environmentally relevant chemical mixtures still may be too low to exceed such toxicity threshold concentrations. Furthermore, even if ingestion of plastic would increase actual risks because the plastic acts as a source for some plastic-associated chemicals, the same ingested plastic could at the same time act as a sink ('cleaning agent') for chemicals other than these plastic-associated ones, i.e. chemicals which are present in the gastrointestinal tract from other sources (Koelmans 2015). An increasing number of studies has demonstrated that this is possible, from empirical evidence as well as through model scenario analysis (Koelmans et al. 2013b, 2016; Devriese et al. 2017; Scopetani et al. 2018; Mohamed Nor and Koelmans 2019; Heinrich and Braunbeck 2020; Thaysen et al. 2020). In terms of risks of the overall chemical mixture that animals are exposed to, the cleaning phenomenon could compensate for the increased exposure of plastic-associated chemicals, possibly leading to net zero or even less effects of the chemical mixture. Below, we further detail these criteria for laboratory studies, field studies, in vitro studies and modelling studies.

### 6.2.1 Criteria for In Vivo Laboratory Studies

An often used setup of laboratory MP vector studies is that MP is first contaminated with chemicals, after which test organisms are exposed to these contaminated MPs and to controls without MP for comparison. An MP vector effect is then indicated if body burdens are higher in the MP treatments than in the control, whereas a risk is
<table><caption>**Table 6.1** Quality criteria for studies investigating the microplastic vector effect in the context of chemical risk assessment</caption><thead><tr><th>Criterion</th><th>Guidance</th><th>Type of study for which the criterion is relevant <sup>a</sup></th></tr></thead><tbody><tr><td>1 Preparation of microplastic with associated chemicals</td><td>Use long-term field-contaminated microplastic, or laboratory spiking with equilibration time of at least a month, and/or corrections for non-equilibrium by means of kinetic sorption modelling</td><td>L</td></tr><tr><td>2 Resemblance of natural exposure pathways</td><td>Chemical exposure via all pathways that are relevant to the organism under consideration is assessed, either through measurement or modelling, preferably through both</td><td>L, F, M</td></tr><tr><td>3 Verification of chemical exposure</td><td>Chemical exposure should be assessed for all exposure pathways, either through measurement or modelling</td><td>L, F</td></tr><tr><td>4 Concentration gradient</td><td>No or limited concentration gradient; if the study aims to mimic natural conditions<br>Maximum gradient; if the study aims to assess sorption kinetic parameters</td><td>L, M</td></tr><tr><td>5 Ingestion</td><td>Ingestion should be demonstrated</td><td>L, F</td></tr><tr><td>6 Evidence from correlations</td><td>Rule out multiple causation<br>Assess correlations on the level of individual chemicals<br>Account for measurement error</td><td>F</td></tr><tr><td>7 Reversibility of chemical transfer</td><td>Parameter fitting and data interpretation should account for bidirectional chemical transfer between ingested plastic and biota tissue</td><td>L, F, M</td></tr><tr><td>8 Model validity</td><td>The model needs to be consistent with empirical data, with current knowledge, and with design criteria</td><td>M</td></tr><tr><td>9 Threshold effect concentration</td><td>If a microplastic vector effect is detected, it should be assessed whether it leads to exceedance of a chemical threshold effect concentration</td><td>L, F, M</td></tr><tr><td>10 Mixture toxicity</td><td>If a microplastic vector effect is detected, it should be assessed whether the increased ('vector effect') and decreased ('cleaning effect') chemical exposures due to microplastic ingestion still lead to a net exceedance of a chemical threshold effect concentration for the chemical mixture</td><td>L, F, M</td></tr></tbody></table>

indicated if the increased exposure is the reason for exceeding an effect threshold. Several factors define the weight of evidence from such studies (Table 6.1).

**Preparation of MP with Associated Chemicals** Several studies have spiked chemicals, e.g. persistent organic pollutants (POPs) to MP, in order to mimic environmentally relevant MP. This should be done for long enough time, because in the environment the far majority of microplastic particles has an age in time scale of

<footnote><sup>a</sup>L laboratory study, F field study, M model study</footnote>
more than months. For instance, Rochman et al. (2013) deployed MP pellets in San Diego Bay for 3 months, which yields environmentally relevant concentrations that are in (near-)equilibrium with the ambient water. Equilibrium has been argued to be the most relevant state for POPs on MP (Koelmans et al. 2016; Lohmann 2017; De Frond et al. 2019; Seidensticker et al. 2019) and implies that the chemicals are relatively well diffused inside the polymer phase, rendering desorption to be slower compared to non-equilibrium situations. After all, when POPs are pre-sorbed for short times only, they desorb from outer sorption domains, which is faster. This means that any detected vector effect would be overestimated. Sorption times as short as 72 h have been applied in some studies (e.g. Kleinteich et al. (2018), Chua et al. (2014)) which thus would lead to higher desorption rates than in nature. In principle, such non-equilibrium artefacts can be corrected via modelling, where the parameters obtained from a non-equilibrium setup can be used in a scenario analysis for equilibrium conditions (Mohamed Nor and Koelmans 2019; Seidensticker et al. 2019).

Sometimes it is argued that additives can have higher than equilibrium concentrations, suggesting that the aforementioned long sorption times would not be needed for such studies. However, microplastic is, by definition, a mixture of the smaller plastic in the environment, largely originating from slow fragmentation, with time scales longer than chemical desorption time scales. The default state for additives thus also is chemical equilibrium with ambient water (Koelmans et al. 2016; Lohmann 2017; De Frond et al. 2019; Seidensticker et al. 2019), a condition which is not always met in laboratory studies mimicking scenarios with additives. For a credible study, we advise an equilibration time of a month or longer, or corrections by means of kinetic sorption modelling when shorter times are used.

**Resemblance of Natural Exposure Pathways** In the environment, exposure of organisms to chemicals occurs via multiple pathways, simultaneously (Fig. 6.1). For aquatic organisms, the most important pathways are dermal uptake via skin and/or gills, and water and food ingestion, whereas for air-breathing organisms, inhalation can be relevant. If MP is present, ingestion of MP might contribute to total uptake. If this occurs, it does not necessarily imply that it is a significant vector. It has been demonstrated that transfer of chemicals via ingestion of MP often is negligible compared to the sum of the uptake via the other pathways (Koelmans et al. 2013b; Bakir et al. 2016). Recent experimental studies have confirmed this experimentally, e.g. for lugworms (Besseling et al. 2017), seabirds (Herzke et al. 2016), daphnids (Horton et al. 2018), marine phytoplankton and zooplankton (Beiras et al. 2018, 2019; Beiras and Tato 2019; Sørensen et al. 2020). These observations comply with theory as was shown in several studies where experimental and natural conditions were simulated using numerical modelling (Bakir et al. 2016; Lee et al. 2019). In contrast, some of the key studies in the literature did not address other exposure pathways (e.g. Browne et al. (2013); Rochman et al. (2013); Chua et al. (2014); Wardrop et al. (2016)). These studies have only one possible outcome, namely, that the ingestion of microplastic is important for chemical uptake because other exposure mechanisms that would occur in nature have been disabled. These
<figure><img src="image_30.png" /><figcaption>**Fig. 6.1** Diagram illustrating the uptake and elimination pathways for chemicals by and from fish. PCB is used to represent chemicals in general. Processes A, B, C and D represent the concomitant terms in Eq. (6.1)</figcaption></figure>

studies have provided support for the MP vector effect paradigm, even though they did not necessarily seek relevance with respect to natural conditions. While the data from these studies are relevant under the conditions used, their deviation from natural conditions makes them less useful for risk assessment related to the MP vector chemical effect. In summary, experimental designs should ideally include, quantify and discuss all pathways relevant in nature, either based on measurement or on measurement in combination with modelling.

**Control, Measurement and Assessment of Exposure** Multiple pathways thus should be included in MP vector study designs, and these should be quantified in order to assess the relative importance of the MP vector effect. One approach is to assess this via measurement. This requires assessment of chemical concentrations in plastic, water and ingested food throughout the experiment (e.g., Besseling et al. 2017; Rehse et al. 2018; Gerdes et al. 2019b). Many studies have interpreted data based on nominal concentrations of chemicals on plastic, which however is not fully reliable. After all, when chemicals are spiked on the plastic as described in the previous section, or when they also are present in added food, they can (partly) desorb or re-equilibrate before MP ingestion has taken place. This implies that the original, nominal concentrations on plastic do not apply anymore, and actual exposure is unknown. This hampers data interpretation and has led to uncertainty with respect to the applicability of the results of some studies. This especially occurs for studies that used gradient between MP sorbed with chemicals and clean water (Rochman et al. 2013; Chua et al. 2014; Wardrop et al. 2016; Beckingham and Ghosh 2017). In a recent review, Burns and Boxall (2018) re-analysed exposure in a study where *Oryzias latipes* were exposed to MP associated with organic contaminants sorbed
from San Diego Bay (Rochman et al. 2013) and assessed that there was insufficient chemical mass on the MP to explain mass measured in fish for most chemicals, suggesting that the cod oil in the diet also was a source. MP vector studies also have suggested that chemicals evaporated from their experimental systems, further limiting data interpretation (Rochman et al. 2017). There are established approaches to measure aqueous phase concentrations, such as with passive samplers, and these have already been applied in some MP vector studies (Besseling et al. 2017; Horton et al. 2018; Wang et al. 2019).

The second approach to assess the relative magnitude of the MP vector effect compared to competing uptake pathways is by calculation. Several studies have demonstrated how results from experiments that did not as such fully comply to natural conditions can still be used in model scenario analysis that were environmentally relevant (Koelmans et al. 2016; Rochman et al. 2017). An example of such an assessment is provided as a case study in Section 4 of this review.

**Resemblance of Chemical Concentration Gradients as They Would Occur in Nature** Several studies have mentioned non-equilibrium exposure as a limitation in the interpretation of their data (e.g. Beckingham and Ghosh 2017; Sleight et al. 2017; Rochman et al. 2017). Furthermore, it has been argued that chemical equilibrium is the most likely state for chemicals on small microplastic (Koelmans et al. 2016; Diepens and Koelmans 2018; De Frond et al. 2019; Seidensticker et al. 2019). Microplastic is small by definition and ‘old’ for the far majority of the particles, because its main process of formation is the slow process of embrittlement, erosion and fragmentation. Smaller and older particles are closer to equilibrium, due to shorter intrapolymer diffusion path lengths and longer sorption time scales (Seidensticker et al. 2019). For microplastic, sorption kinetic time scales range from weeks to months. Residence times of the particles in the environment are long (90% being older than 2 years in the ocean (reviewed in Koelmans et al. (2016)), which is one of the main reasons for the concerns surrounding microplastic. In combination, these factors cause equilibrium to be the default state for plastic-associated chemicals. The same applies to other small particles in the ocean, like detritus, black carbon and sediment or suspended solid organic matter, which all would have sorption half-lives of days to weeks at most (Schwarzenbach et al. 2005). This however also applies to organisms at the base of the food chain, as well as, for instance, fish eggs and larvae. In later stages of their life cycle, fish consume contaminated prey keeping the organism body burden close to equilibrium or even higher due to biomagnification. This implies that for chemicals in ingested microplastics, there is no concentration gradient that would drive transfer to the organism, to begin with. For MP vector studies to be environmentally relevant, this means that plastic, organisms, water and other compartments (food, sediment) should all include a scenario where chemicals are contaminated to the same extent (equal fugacity), ideally by pre-equilibration of the chemicals prior to exposure to microplastic. Although second best, studies could still use a large enough gradient, in order to be able detect an effect in the first place. However, in such a case, results should be calculated back to a natural conditions scenario. Alternatively, the bias caused by testing chemical
uptake from contaminated MP by clean organisms can be balanced by also testing the opposite treatment, i.e. by using (relatively) clean MP ingested by contaminated test organisms (Koelmans et al. 2013b; Rummel et al. 2016; Scopetani et al. 2018; Wang et al. 2020b). Studies that make an effort to use pre-equilibrated exposure media and organisms (e.g. Besseling et al. (2013); Besseling et al. (2017)) or that otherwise account for vector as well as cleaning mechanisms combined, score higher, whereas evidence from studies that, for instance, force chemical transfer by using clean test animals would receive lower weight.

**Assessment of Ingestion** Laboratory MP vector studies have been performed with a wide variety of test organisms, covering functional groups such as phytoplankton, zooplankton, (benthic) invertebrates and fish. The essence of the MP vector hypothesis relates to chemical transfer upon ingestion of the contaminated particles. Therefore, we argue that MP vector studies should ideally provide evidence that the test animals actually ingested the particles and preferably also at which rate.

## 6.2.2 Criteria for Field Studies

Studies have attempted to use observed co-occurrences or correlations between field data on plastic densities or of chemical concentrations in plastic, with chemical concentrations in organisms, as evidence supporting the MP vector hypothesis. The main challenge here is that, whereas in laboratory experiments typically all conditions are kept the same except for the one being researched, in nature, (a) multiple mechanisms are at play simultaneously, (b) there is no control over them, and (c) they may remain partly unknown. Given this, several factors define the weight of evidence from such studies (Table 6.1).

**Resemblance of Natural Exposure Pathways** Various researchers and expert groups have defined multiple processes and uptake pathways that can explain the occurrence of chemicals (e.g. POPs, PBTs, additives) in organisms, one of which is by ingestion of MP and subsequent chemical desorption inside the organism (GESAMP 2015; Rochman 2015; Koelmans et al. 2016). This means that studies should motivate why the uptake flux via these combined other parallel processes is small compared to the chemical uptake flux via microplastic ingestion. Pathways should be assessed such that it is relevant for the organism studied. However, studies have often neglected to consider such parallel pathways. For instance, the effect of microplastic ingestion on bioaccumulation of phthalates by the fin whale was speculated from detection of plastic in plankton samples and from phthalates being detected in the same plankton samples and fin whale (Fossi et al. 2012). However, the plankton to microplastic number concentration ratios were higher than 1000. Given these high ratios and residence times of microplastic in the oceans, causing chemical desorption from microplastic and subsequent uptake by plankton, it is most likely that dermal absorption and ingestion of phthalate-contaminated plank-
ton were the dominant pathways, rather than from direct ingestion of the microplastic. A follow-up study analysed the case using theoretical modelling and indeed found that 99.2 % of phthalate uptake was by plankton ingestion and only 0.8 % by microplastic ingestion (Panti et al. 2016). Another example is the study by Tanaka et al. (2013), who found some chemicals in seabirds that were absent in prey and thus concluded that the ingested plastic must have been the source. However, the prey organisms in which the chemicals were not detected were collected 7 years after the year of collection of the seabirds, rendering the comparison inconclusive (Burns and Boxall 2018; Tanaka et al. 2013). There are few field studies that quantify the fluxes of the parallel pathways based on field data (Koelmans et al. 2014; Herzke et al. 2016; Panti et al. 2017), which all conclude a negligible relevance of the MP vector effect. As an opposite example, for a ‘hot spot’ location in the North Pacific accumulation zone, Chen et al. (2017) calculated MP mass to biomass ratios and determined that MPs outweigh prey in the same size range (0.5–5 mm) by 40 times (and by 180 times if all buoyant plastic and biota > 0.5 mm are considered). They thereby suggested that MP at the surface may make up a significant dietary contribution. A further example of such an analysis is provided as a case study in Section 4. In summary, like empirical studies, field studies should account for all exposure pathways in order to assess the relative importance of the MP ingestion vector effect, either based on measurement or on measurement in combination with modelling.

**Evidence from Correlations** Some studies have found evidence for the MP vector effect in observed correlations between field data on plastic densities or of chemical concentrations in plastic, with chemical concentrations in organisms (Rochman et al. 2014; Panti et al. 2017; Gassel and Rochman 2019). There are a few pitfalls with the use of such correlations. First, correlations show that variables are related, but they do not reveal the causality of the relationship. For example, in the oceans, high plastic concentrations with high chemical concentrations will inevitably lead to high chemical concentrations in the seawater and thus in biota, via phase partitioning. These concentrations thus will always be correlated. However, this also occurs without ingestion, thus minimizing the weight of evidence from such correlations. Given that (a) MP ingestion is not often verified (Koelmans et al. 2016), (b) studies that assessed ingested MP reveal that MP levels are very low (Foekema et al. 2013; Hermsen et al. 2017; Markic et al. 2019), and (c) if they would be ingested, a gradient for transfer would be lacking (see above), chemical accumulation via the water or ingestion of zooplankton is a far more plausible explanation. Furthermore, bioaccumulation is driven by many factors and not by plastic density alone. For instance, phytoplankton blooms or dissolved organic carbon (DOC) concentrations are highly dynamic and have been demonstrated to affect chemical concentrations in the oceans (Dachs et al. 2002; Jurado et al. 2004). Bioaccumulation thus is multifactorial and without considering all factors at play, i.e. by using univariate correlations, no conclusions on implications of MP can be drawn. Co-occurrence or correlations of high chemical concentrations in fish with higher plastic density are difficult to interpret if plastic densities apply to the surface, whereas the fish species
reside (for considerable time) in another habitat, for instance, below the surface, like for myctophids (e.g. Gassel and Rochman 2019). Another pitfall relates to the use of summed concentrations of the investigated chemicals (e.g. ΣPCBs; Rochman et al. 2014; Gassel and Rochman 2019). The sums all have a different proportion of each of the individual chemicals in the mixture. Therefore, each data point in the correlation resembles something different, another variable, rendering the correlation to be unclear. Instead, we argue that for an unambiguous measurement of chemical trends, such correlations should only use concentrations of the same one chemical. A final evaluation criterion relates to accounting for error and uncertainty in the data used in the correlation. As mentioned, the dependent variable, which is often the summed chemical concentration, suffers from uncertainty related to the variability in congener composition which is not accounted for. However, regression analysis, as used in these types of studies, considers the independent variable (e.g. MP density) as if it contained no error itself. This is problematic when these data points actually are prone to error, which is often the case. For instance, Gassel and Rochman (2019) correlated sum chemical concentrations with plastic concentrations interpolated by modelling, which can be assumed to have considerable uncertainty. In these cases, it is recommended that weighted regression procedures are applied. In studies that use such correlations, such pitfalls are typically not discussed. In summary, we recommend that studies that provide evidence based on correlations, to rule out parallel causal explanations for the observations, use analyses of individual chemicals and take measurement error into account.

**Assessment of Ingestion** Similar to laboratory MP vector studies, field studies that conclude that ingestion of MP increase chemical concentration in biota should provide evidence that the test animals actually ingested the particles and preferably also at which rate.

### 6.2.3 Criteria for In Vitro Studies

Several scientists have studied chemical transfer kinetics between plastic particles and water or plastic particles and gut fluids or stomach oil (e.g. Teuten et al. (2007); Bakir et al. (2014); Beckingham and Ghosh (2017); Lee et al. (2019); Martin and Turner (2019); Mohamed Nor and Koelmans et al. (2019); Kühn et al. (2020)). Besides the common quality assurance (QA) criteria that apply to any study that needs the analysis of chemical concentrations on plastic, there are several criteria that need to be fulfilled in order to make a study relevant with respect to risks caused by the MP vector effect (Table 6.1). Chemical desorption kinetics in artificial gut fluids have been assessed in order to show the MP vector effect. Such studies should either use long-term adsorption prior to desorption in order to not overestimate the desorption rates, as mentioned in Section 2.1, or the level of non-equilibrium should be accounted for in the desorption kinetic parameter estimation (Mohamed Nor and Koelmans 2019). Because sorption involves a dynamic equilibrium, desorption
studies should also account for backward sorption in their setup and data analysis. Kinetic modelling to acquire the rate constants may need to take intrapolymer diffusive mass transfer and/or biphasic behaviour into account (Seidensticker et al. 2017, 2019; Town et al. 2018; Lee et al. 2019). Kinetic parameters should be provided with significance level and/or error estimate, preferably using multiple time points in the desorption curve allowing for a rigorous estimation of parameter values (Mohamed Nor and Koelmans 2019). Furthermore, besides exchange fluxes between MP and gut fluid, the fluxes from food should be assessed in order to assure that the role of ingested MP is not negligible as compared to the parallel exposure pathways. For instance, Kühn et al. (2020) observed leaching of five chemicals from MP into stomach oil sampled from seabirds. However, the uptake of chemicals from normal food and the concentrations of background chemicals in the oil were not measured or modelled, nor were kinetic parameters estimated, precluding conclusions regarding the relevance of the vector effect under environmentally realistic conditions. Furthermore, the in vitro transfer data cannot be assumed relevant for natural conditions in vivo where removal processes such as metabolization and elimination would reduce the gradient for transfer and where bioaccumulation by seabirds is likely to be at steady state. The kinetic data thus have to be put in the context of all relevant processes that govern the uptake and elimination of the chemicals of interest. Models have been developed that allow for such a context-dependent evaluation using the kinetic parameters obtained from in vitro desorption studies (Bakir et al. 2016; Koelmans et al. 2013b, 2016). Studies that account for all these aspects would be most relevant for assessing the relevance of the MP vector effect in the context of chemical risk assessment.

### 6.2.4 Criteria for Model Scenario Studies

Several studies have applied models to calibrate parameters relevant for the MP vector effect (Koelmans et al. 2016; Mohamed Nor and Koelmans 2019) or to evaluate the importance of the MP vector effect based on empirical data (Gouin et al. 2011; Koelmans et al. 2013b, 2014, 2016; Bakir et al. 2016; Herzke et al. 2016; Rochman et al. 2017; Wang et al. 2019; Wang et al. 2020a, b). As a starting point for the evaluation of studies that used models, we adopt the criteria for a valid model suggested by Rykiel Jr (1996): 'Validation is establishing the truth of a model in the sense of (a) consistency with data, (b) accordance with current knowledge, (c) conformance with design criteria'. Consistency with data means that MP vector models should be demonstrated to concur with observations. This implies that calibrated parameters comply with theory or that parameter sets purely obtained from theory comply with observations. Model frameworks that lack such consistency would score lower. Accordance with current knowledge means that the model application or parts of it should not be in conflict with established and widely accepted evidence-based theoretical concepts or that no such concept is overlooked in case it would be relevant. For instance, model scenario studies informing chemical risk assessments about the
relevance of the MP vector effect should account for reversible exchange in the gut, should take into account what is the gradient occurring under natural conditions and should account for all pathways that are relevant under natural conditions. Conformance with design criteria means that a model's complexity, approach and output comply with what a model is meant to do in a certain context. For instance, a model should not be over-parameterized or should it be too simple with respect to the dominant processes that drive the behaviour of the system under consideration.

## 6.3 Weight of Evidence Supporting the Microplastic Vector Hypothesis in the Context of Chemical Risk

The literature provides varying types of evidence for the MP vector effect and its implications on chemical risks. In 2016, we provided a first evaluation of 13 key laboratory, field and model studies that were available at that time (Koelmans et al. 2016). The evaluation was based on whether studies used realistic MP concentrations or MP fractions in the diet, whether MP ingestion was confirmed and whether all environmentally relevant uptake processes were accounted for. In 2018, a similar analysis was published by Burns and Boxall (2018), who evaluated 18 studies and ranked them into 3 evidence categories: 'demonstrated', 'inconclusive' and 'not supported'.

The review in this section can be seen as an extended and detailed version of these earlier evaluations. Here, we evaluate the weight of evidence for the MP vector effect from 61 studies, based on 2 criteria: (a) is the evidence conclusive enough, given the recommendations discussed in the previous section, and if so, (b) does the study provide evidence for an MP vector effect that would affect chemical risk in nature. Following Burns and Boxall (2018), the same three weight of evidence categories were distinguished.

### 6.3.1 Weight of Evidence from In Vivo Laboratory Studies

We reviewed 30 studies providing evidence with respect to the MP vector effect (Table 6.2). None of them explicitly discussed the results with respect to implications for chemical risks as defined by the PEC/PNEC approach (Koelmans et al. 2017). Twenty-one out of thirty studies were evaluated as inconclusive with respect to detection of a chemical vector effect caused by microplastic ingestion. However, there appears to be a trend among studies over time. Earlier studies more often neglect environmentally relevant exposure conditions by focusing on chemical uptake from microplastic alone, resulting in a low (i.e. 'inconclusive') weight of evidence score. Later studies more often include co-exposure from water or also from food or other background particles empirically and/or via modelling, which makes them more relevant for environmental conditions, resulting in a higher weight of evidence (conclusive) score. All of the latter studies, however, conclude that a
evidence, also because many of these simulations were done with validated models or backed up with experimental data provided in the same study. Most of the studies concluded that a vector effect is not supported, due to lack of gradient or due to other pathways being more important. Some studies might have overlooked some

<table><caption>**Table 6.2** Overview of laboratory in vivo studies addressing the role of microplastic ingestion by organisms on bioaccumulation of plastic-associated chemicals</caption><thead><tr><td>Study</td><td>Demonstrated</td><td>Inconclusive</td><td>Not supported</td><td>Comments</td></tr></thead><tbody><tr><td>Besseling et al. (2013)</td><td></td><td>X</td><td></td><td>Measured environmentally relevant exposure of PCBs to A. marina in the lab, including all pathways. Treatments with MP showed higher bioaccumulation. However, aqueous exposure and organism lipids were not measured and could have differed among treatments, which limits interpretation with regard to the MP vector effect</td></tr><tr><td>Browne et al. (2013)</td><td></td><td>X</td><td></td><td>Exposed initially clean A. marina to MP or sand spiked with phenanthrene, nonylphenol, triclosan or PBDE-47. Co-exposure from water and food was not considered. Considering the experimental design, uptake from water could have occurred as well. Found more uptake from sand than from ingested plastic. This may be explained from the plastic being a more efficient cause for elimination from the test organisms, via particle egestion/defaecation, compared to sand</td></tr><tr><td>Rochman et al. (2013)</td><td></td><td>X</td><td></td><td>Test organisms Oryzias latipes were exposed to MPs that were enriched with environmental contaminants sorbed from San Diego Bay. Uptake from water was not accounted for and not measured. Test organisms were not at equilibrium at the start of the experiment. Not clear which part of uptake was from contaminated food or from water</td></tr><tr><td>Chua et al. (2014)</td><td></td><td>X</td><td></td><td>Amphipods Allorchestes compressa were exposed to PBDEs in the presence or absence of MP. Co-exposure from water and food was not considered, and initially clean organisms were used. Considering the experimental design, uptake from water could have occurred as well. Used unrealistically high plastic concentrations</td></tr></tbody></table>

(continued)
<table><caption>**Table 6.2 (continued)**</caption><thead><tr><th>Study</th><th>Demonstrated</th><th>Inconclusive</th><th>Not supported</th><th>Comments</th></tr></thead><tbody><tr><td>Avio et al. (2015)</td><td></td><td>X</td><td></td><td>Mussels (Mytilus galloprovincialis) were exposed to MP with and without sorbed pyrene. Uptake from water and natural food was not considered, and initially clean organisms were used. Considering the experimental design, uptake from water could have occurred as well. Used very high plastic concentrations</td></tr><tr><td>Wardrop et al. (2016)</td><td></td><td>X</td><td></td><td>Rainbow fish (Melanotaenia fluviatilis) were exposed to MP with and without sorbed PBDEs. Co-exposure from water and food was not considered, and initially clean organisms were used. Considering the experimental design, uptake from water could have occurred as well. Used very high plastic concentrations</td></tr><tr><td>Rummel et al. (2016)</td><td></td><td>X</td><td></td><td>Studied ‘cleaning’ of Oncorhynchus mykiss by microplastic ingestion (reversed vector effect), by feeding PCB-contaminated fish a diet with uncontaminated PE. No observable effect of plastic ingestion on the rate of depuration of PCBs was reported. However, due to the wide confidence intervals in the depuration rate constants, the study had a low sensitivity to detect treatment effects. Hence, the observations may not have been sufficient to make the assumption that the cleaning effect was not present</td></tr><tr><td>Besseling et al. (2017)</td><td></td><td></td><td>X</td><td>Mimicked environmentally relevant exposures of A. marina to PCBs in sediment, MP and water after 6-week equilibration, including all pathways, and assessed ingestion. Aqueous exposure was measured using passive samplers, and organism lipids were measured. Data interpretation was aided by biodynamic modelling based on Koelmans et al. (2013b) (Table 6.4). Bioaccumulation factors did not differ between MP and no-MP treatments, providing evidence for the absence of a vector effect, which was supported by model simulations</td></tr></tbody></table>

(continued)
<table><caption>**Table 6.2** (continued)</caption><thead><tr><td>Study</td><td>Demonstrated</td><td>Inconclusive</td><td>Not supported</td><td>Comments</td></tr></thead><tbody><tr><td>Beckingham and Ghosh (2017)</td><td></td><td>X</td><td></td><td>In vivo study on bioaccumulation of PCBs with <i>Lumbriculus variegatus</i>. Did not consider leaching of PCBs from polypropylene into sediment and water. Did not quantify the relevance of the polypropylene as vector compared to the indirect exposure through desorption to the surrounding media. Did not assess ingestion of microplastic in the worms</td></tr><tr><td>Devriese et al. (2017)</td><td></td><td>X</td><td></td><td>Norway lobster <i>Nephrops norvegicus</i> fed with (short-term spiked) PCB-loaded MPs showed limited additional PCB uptake after 3-week ingestion of PE MP, but not of PS MP. PCBs were MP surface spiked, gelatine food contained PCBs as well, and uptake from water could have occurred as well, impeding differentiation between accumulation from food, water and plastic</td></tr><tr><td>Sleight et al. (2017)</td><td></td><td>X</td><td></td><td>Tested the hypothesis that MP can act as a vector of chemicals from pelagic to benthic habitats. MP bound chemical bioavailability for zebrafish larvae <i>Danio rerio</i> was assessed via gene expression. Aqueous exposure was not assessed. Sorption equilibrium was assumed in data interpretation, yet not experimentally verified and unlikely given the short (5d) exposure time. MP and chemical concentrations were considerably higher than environmentally relevant concentrations</td></tr><tr><td>Rochman et al. (2017)</td><td></td><td>X</td><td></td><td>Tested how the interaction between microplastics and PCBs could affect a prey species <i>Corbicula fluminea</i> and its predator <i>Acipenser transmontanus</i>. Exposure was to initially clean organisms, via plastic ingestion alone, but not quantified. PCBs could have desorbed to the clean water followed by volatilization of PCBs from microplastics from aeration in the tanks. PCBs were not detected in test animals, so treatment effects on bioaccumulation could not be assessed</td></tr></tbody></table>

(continued)
<table><caption>**Table 6.2 (continued)**</caption><thead><tr><th>Study</th><th>Demonstrated</th><th>Inconclusive</th><th>Not supported</th><th>Comments</th></tr></thead><tbody><tr><td>Horton et al. (2018)</td><td></td><td>X</td><td></td><td>Environmentally relevant co-exposure of <i>Daphnia magna</i> to chemicals in MP and water. However, the systems were not at equilibrium partly due to addition of the test organisms. MP concentrations were much higher than in the environment. No effects of MP on chemical toxicity were found</td></tr><tr><td>Guven et al. (2018)</td><td></td><td>X</td><td></td><td>Tested acute effect of pyrene and MP on swimming and predatory performance of a tropical fish (<i>Lates calcarifer</i>). Not clear if test organisms, MP and pyrene were in equilibrium as effects were recorded between 1 and 24 h. fish were not fed, eliminating the realism of this exposure pathway. Not clear how pyrene was measured. The authors confirm that short test duration may have restricted pyrene absorption to the MP</td></tr><tr><td>Batel et al. (2018)</td><td></td><td>X</td><td></td><td>Tested and confirmed transfer of BaP from MP by simple attachment to clean adult zebrafish (<i>Danio rerio</i>) gills and zebrafish embryos. BaP equilibration with MP was for only 24 h. BaP may have desorbed to the water during exposure. Role of waterborne exposure not assessed. No environmentally realistic concentration gradient due to use of clean organisms. Very high MP and BaP levels</td></tr><tr><td>Beiras et al. (2018)</td><td></td><td></td><td>X</td><td>Tested ingestion and contact with PE MP; did not find acute toxicity on marine zooplankton (<i>Brachionus plicatilis</i>, <i>Tigriopus fulvus</i>, <i>Acartia clausi</i>, <i>Mytilus galloprovincialis</i>, <i>Paracentrotus lividus</i>, <i>Oryzias melastigma</i>). Spiking of MP was for 2 d only. Contaminated MP was dosed to clean organisms, and no other parallel chemical exposure pathways were taken into account. Despite these conditions of artificially favouring an effect of MP, no effect was found. This means the results can be considered as not supporting the occurrence of an MP vector effect for the species tested</td></tr></tbody></table>

(continued)
<table><caption>**Table 6.2** (continued)</caption><thead><tr><td>Study</td><td>Demonstrated</td><td>Inconclusive</td><td>Not supported</td><td>Comments</td></tr></thead><tbody><tr><td>Scopetani et al. (2018)</td><td></td><td></td><td>X</td><td>Clean <i>Talitrus saltator</i> were fed with uncontaminated fish food mixed with (short term) contaminated MPs, or contaminated <i>Talitrus saltator</i> was fed with contaminated food in combination with clean MP. The first treatment showed a vector effect, whereas the second treatment showed that MP cleaned the organism. Spiking of MP was short term and desorption in the water could have occurred. Given the detection of opposite direction chemical transport, the conclusion of a limited relevance for a vector effect in the environment is considered correct</td></tr><tr><td>Barboza et al. (2018)</td><td></td><td>X</td><td></td><td>Effects of <i>Dicentrarchus labrax</i> juveniles upon exposure to MP and HgCl<sub>2</sub>. MP polymer type and associated chemicals not known. Clean fish were used and no food was added, causing the hg speciation and exposure to be different from natural conditions, leaving uncertainty with respect to how MP would make a difference in reality</td></tr><tr><td>Rehse et al. (2018)</td><td></td><td></td><td>X</td><td>Analysed how the presence of polyamide MP modifies effects of bisphenol A (BPA) on <i>Daphnia magna</i>. BPA exposure was via water only or via water plus ingested MP. EC<sub>50</sub> values were the same, which demonstrates that an MP vector effect did not occur</td></tr><tr><td>Beiras and Tato (2019)</td><td></td><td></td><td>X</td><td>MP and nonylphenol effects on urchin larvae (<i>Paracentrotus lividus</i>). Conditions aimed at maximizing the relevance of the particulate phase for chemical uptake in the test species: Natural particles were not present, and MP loads tested were far above the environmental levels ever found in marine waters. With MP and food present, there were no differences in chemical effect thresholds as compared to when no MPs were present, indicating that no vector effect occurred</td></tr></tbody></table>

(continued)
<table><caption>**Table 6.2** (continued)</caption><thead><tr><th>Study</th><th>Demonstrated</th><th>Inconclusive</th><th>Not supported</th><th>Comments</th></tr></thead><tbody><tr><td>Wang et al. (2019)</td><td></td><td></td><td>X</td><td>Earthworm <i>Eisenia fetida</i> exposed to PE and PS particles in agricultural soil; PCB-, PAH- and MP-contaminated soil; and PCB/PAH-contaminated soil. Data interpretation was aided by biodynamic modelling based on Koelmans et al. (2013b) (Table 6.4). No MP vector effect was found</td></tr><tr><td>Beiras et al. (2019)</td><td></td><td></td><td>X</td><td>Sea urchin pluteus and copepod nauplius larvae actively ingest MP particles. MP did not increase accumulation of organic chemicals in sea urchin larvae. MP did not increase the toxicity or 4-n-NP or 4-MBC to zooplankton. PE microplastics did not act as vectors of hydrophobic chemicals to zooplankton</td></tr><tr><td>Sørensen et al. (2020)</td><td></td><td></td><td>X</td><td><i>Acartia tonsa</i> and <i>Calanus finmarchicus</i> were exposed to ingestible and non-ingestible PE microbeads, spiked with fluoranthene and phenanthrene, which also were present in the water phase. Under this co-exposure scenario, bioaccumulation factors were the same for systems with versus without MP, and for systems with ingestible versus non-ingestible MP, indicating that no MP vector effect occurred</td></tr><tr><td>Thaysen et al. (2020)</td><td></td><td>X</td><td></td><td>Studied the direction of transfer between ingested plastic and biota lipids of brominated flame retardants in <i>Larus delawarensis</i>. It was found that the concentration gradients were opposite for different chemicals, and thus would lead to bidirectional transfer dependent on the chemical, yet with transfer from bird to plastic ('cleaning') occurring for a higher number of compounds</td></tr></tbody></table>

(continued)
<table><caption>**Table 6.2** (continued)</caption><thead><tr><th>Study</th><th>Demonstrated</th><th>Inconclusive</th><th>Not supported</th><th>Comments</th></tr></thead><tbody><tr><td>Bartonitz et al. (2020)</td><td></td><td>X</td><td></td><td><i>Gammarus roeseli</i> were exposed to phenanthrene in water, as well as in the presence of MP or sediment. Due to their high concentrations, MP (and sediment) particles reduced phenanthrene toxicity due to sorption to the particles, and absence of a vector effect was concluded. In nature, however, MP would not reduce aqueous phase concentrations at realistic MP concentrations, rendering the results inconclusive</td></tr><tr><td>Xia et al. (2020)</td><td></td><td>X</td><td></td><td>Exposed <i>Chlamys farreri</i> to BDE209 with versus without presence of PS MP. Bioconcentration factors (BCF) were the same; however, depuration was faster in the presence of MP, demonstrating the 'cleaning' effect. The conditions of MP and BDE209 pre-equilibration were not fully clear; co-exposure from the water was not quantified; water was renewed and scallops removed every day, which was not accounted for in the modelling; clean test organisms were used; and no food was present during exposure</td></tr><tr><td>Coffin et al. (2020)</td><td></td><td></td><td>X</td><td>Exposed <i>Atractoscion nobilis</i> to environmentally relevant MP (0.32 particles/L) and benzo(a)pyrene-sorbed MP concentrations. Co-exposure via food or water was not included. No effects of the presence of MP were observed in 5-d exposure, indicating that no MP vector effect occurred</td></tr><tr><td>Tanaka et al. (2020)</td><td></td><td>X</td><td></td><td>Feeding experiment under environmentally relevant conditions, in which PE pellets contaminated with 5 additives were fed to <i>Calonectris leucomelas</i> chicks. The MP vector effect was demonstrated under these experimental conditions. However test animals were clean at start, and the chemicals were not present in the diet, whereas co-exposure via the diet is plausible in nature where most chemicals are ubiquitous</td></tr></tbody></table>

(continued)
<table><caption>**Table 6.2** (continued)</caption><thead><tr><td>Study</td><td>Demonstrated</td><td>Inconclusive</td><td>Not supported</td><td>Comments</td></tr></thead><tbody><tr><td>Wang et al. (2020a)</td><td></td><td>X</td><td></td><td>Earthworm <i>Eisenia fetida</i> exposed to five MP polymer types, spiked with PCB and PAH, over 28 d. Chemical transfer to the worms was observed with dermal exposure generally dominating overexposure via ingested plastic ('vector'). Addition of clean MP reduced transfer ('cleaning'). Data interpretation was aided by biodynamic modelling based on Koelmans et al (2013b) (Table 6.4). It was concluded that MP can act as a source and a sink but that predictions based on short-term non-equilibrium conditions may both be representative of natural conditions</td></tr><tr><td>Wang et al. (2020b)</td><td></td><td>X</td><td></td><td>Earthworm <i>Eisenia fetida</i> exposed to PCB-contaminated soil and clean MP (PE of three sizes) or to clean soil and PCB-contaminated MP, for 28 d. Both treatments demonstrated less uptake, i.e. 'cleaning' compared to non-MP controls. Data interpretation was aided by biodynamic modelling based on Koelmans et al (2013b) (Table 6.4). Chemical uptake via MP ingestion was smaller than via the parallel pathways. It was concluded that MP can act as a source and a sink but that any effect in nature would be small due to low MP concentration</td></tr></tbody></table>

General approach, type of evidence for an effect on chemical risk and environmental realism are summarized

vector effect is of little relevance under natural conditions (Table 6.2). Four studies explicitly aimed for finding empirical evidence for the ‘cleaning’ effect by micro-plastic ingestion (Rummel et al. 2016; Scopetani et al. 2018; Thaysen et al. 2020; Heinrich and Braunbeck 2020), as predicted based on first principles (Gouin et al. 2011; Koelmans et al. 2013b). A cleaning effect was, however, also suggested to occur in the study by Devriese et al. (2017). Browne et al. (2013) found lower chemical body burdens in *A. marina* after chemical uptake via ingested microplastic as compared to uptake via sand, suggesting more efficient ‘cleaning’ via the plastic as compared to sand. Recently, Xia et al. (2020) reported a statistically significant 30% to 45% faster depuration of BDE-209 from marine scallops in treatments with MP. Bioconcentration factors (BCF) were 6 to 14% higher in the presence of MP, differences which, however, were not statistically significant. An MP ‘cleaning’ effect was thus demonstrated, in contrast to an MP vector effect.
### 6.3.2 Weight of Evidence from Field Studies

We evaluated six field studies, of which only one was considered to provide conclusive evidence (Herzke et al. 2016), evidence which in this case was not supportive of a chemical vector effect (Table 6.3). The other studies all suffered from the pitfalls of causality, i.e. did not address alternative yet possibly occurring mechanisms that could also explain the observations, such as uptake via water or food.

<table><caption>**Table 6.3** Overview of field studies addressing the role of microplastic ingestion by organisms on bioaccumulation of plastic-associated chemicals</caption><thead><tr><td>Study</td><td>Demonstrated</td><td>Inconclusive</td><td>Not supported</td><td>Comments</td></tr></thead><tbody><tr><td>Fossi et al. (2012)</td><td></td><td>X</td><td></td><td>Effect of microplastic ingestion on bioaccumulation of phthalates was speculated from detection of plastic in plankton samples, and phthalates detected in the same plankton samples and in fin whale (Balaenoptera physalus). The plankton to microplastic number concentration ratio was 1600 (Ligurian Sea) to 18000 (Sardinian Sea). Given these ratios and ageing of plastic in the oceans causing chemical desorption and uptake by plankton, uptake via plankton is more likely to occur than via plastic ingestion</td></tr><tr><td>Gassel et al. (2013)</td><td></td><td>X</td><td></td><td>Ingestion of plastic was speculated to best explain the detection of nonylphenol in 6 out of 19 fish individuals (Seriola lalandi), given the detection of two plastic particles in 2 out of the 19 fish individuals. Given the data provided, the study shows that plastic may have been the source of the nonylphenol. However, fish (n=19) and plastic (n=2) sample sizes were very low, and parallel pathways may have contributed to the uptake</td></tr><tr><td>Tanaka et al. (2013)</td><td></td><td>X</td><td></td><td>PBDE found in seabirds (Puffinus tenuirostris) were also present in plastic in the stomach, but not in prey, suggesting that the ingested plastic was the source of the PBDEs. However, the prey samples were taken 7 yrs later and >1000km away. The authors acknowledged these caveats</td></tr></tbody></table>

(continued)
<table><caption>**Table 6.3** (continued)</caption><thead><tr><td>Study</td><td>Demonstrated</td><td>Inconclusive</td><td>Not supported</td><td>Comments</td></tr></thead><tbody><tr><td>Rochman et al. (2014)</td><td></td><td>X</td><td></td><td>Myctophid sampled at stations with greater plastic densities had larger concentrations of BDEs # 183–209 in their tissues suggesting that these chemicals are indicative of plastic contamination in the marine environment. Plastic was not measured in the fish, and the BDEs might as well have accumulated from water or the plankton diet. No strong conclusion on the role of ingestion was drawn</td></tr><tr><td>Herzke et al. (2016)</td><td></td><td></td><td>X</td><td>Combined three lines of evidence: (a) correlations among POP concentrations, differences in tissue concentrations of POPs between plastic ingestion subgroups, (b) fugacity calculations and (c) bioaccumulation modelling, to show that MP did not act as a vector of POPs for Fulmarus glacialis</td></tr><tr><td>Gassel and Rochman (2019)</td><td></td><td>X</td><td></td><td>Examined relationships among chemical contaminants and MP in lanternfish (Myctophidae). Lower chlorinated PCBs, higher in gyre fish, correlated with higher modelled plastic density. Data normality and uncertainty in modelled MP densities data potentially affecting significance levels were not taken into account in regression analysis. Exposure pathways other than MP that could also explain the observed differences were not considered</td></tr></tbody></table>

General approach, type of evidence for an effect on chemical risk and environmental realism are summarized

### 6.3.3 Weight of Evidence from In Vitro Studies

Thirteen in vitro studies reported on the chemical release kinetics from microplastic in the context of the chemical vector effect (Table 6.4). In general, most in vitro studies only reported the amount bioavailable or leached from plastic after exposure in the simulated gut fluid or stomach oil. These bioavailability percentages which are reported generally are not relevant in nature as they also depend on external factors such as fugacities in other compartments in the gut, which would be different in nature. Exchange fluxes with other components in the gut, like food organic matter or micelles, are also usually neglected. Generally, analysis of experiment results with models that describe the data are lacking. This limits the applicability of the findings as they cannot be extrapolated to the actual environment. All studies except two (Mohamed Nor and Koelmans 2019; Kühn et al. 2020) had used a clean gut
<table><caption>**Table 6.4** Overview of laboratory in vitro studies aiming to assess the role of microplastic ingestion by organisms on bioaccumulation of plastic-associated chemicals</caption><thead><tr><td>Study</td><td>Demonstrated</td><td>Inconclusive</td><td>Not supported</td><td>Comments</td></tr></thead><tbody><tr><td>Teuten et al. (2007)</td><td></td><td>X</td><td></td><td>Compared desorption rates of sorbed phenanthrene from plastic and sediment in seawater and sodium taurocholate that simulated the gut fluid of <i>Arenicola marina</i>. Model analysis was limited to pseudo first-order rate analysis to determine initial desorption. Considered theoretical calculations using equilibrium partitioning method with different compartments to understand their respective relative contributions. This analysis overestimates the amount bioaccessible to benthic organisms as the environment is more dynamic and it also neglected other pathways</td></tr><tr><td>Bakir et al. (2014)</td><td></td><td>X</td><td></td><td>Investigated the desorption rates of chemicals from PVC and PE under simulated gut conditions. Model analysis was limited as only first-order rate kinetics was considered, neglecting backward sorption and bimodal behaviour. Chemicals were sorbed for a short time, but non-equilibrium desorption kinetics were not considered. Assessment with maximum concentration gradient is not relevant in nature. Other pathways in the gut are neglected</td></tr><tr><td>Turner and Lau (2016)</td><td></td><td>X</td><td></td><td>Evaluated bioaccessibility of elements indicative of halogenated flame retardants in a simulated digestion fluid of seabirds. Model analysis was limited and obtained rate constants using a diffusion-controlled and parabolic model. Assessment with maximum concentration gradient is not relevant in nature. Evaluation of the results in the context of the actual environment is based on total percentage leached from plastic which is not adequate as the gut of the seabirds are already contaminated from other sources</td></tr></tbody></table>

(continued)
<table><caption>**Table 6.4 (continued)**</caption><thead><tr><td>Study</td><td>Demonstrated</td><td>Inconclusive</td><td>Not supported</td><td>Comments</td></tr></thead><tbody><tr><td>Beckingham and Ghosh (2017)</td><td></td><td>X</td><td></td><td>Compared gut fluid solubilization between polypropylene and other natural and anthropogenic organic particles. The model remained limited and data was analysed assuming that equilibrium was reached for all congeners in the system after 4h. Maximum concentration gradient between contaminated particles and gut fluid was not environmentally relevant</td></tr><tr><td>Massos and Turner (2017)</td><td></td><td>X</td><td></td><td>Evaluated bioaccessibility of Cd, Pb and Br from beached microplastics in a physiologically based extraction test. Considered only percentage bioaccessible (over total available in the plastic) which does not reflect what happens in nature as the results are only applicable for the case when the organism do not have any accumulated element. Compared bioaccessibility of the elements through plastic against dietary concentrations and did not consider that the percentages may change in the presence of food</td></tr><tr><td>Turner (2018)</td><td></td><td>X</td><td></td><td>Evaluated bioaccessibility of hazardous elements in simulated digestion fluid of northern fulmar. Model analysis was limited and did not consider accounting for individual exchange fluxes as it uses the pseudo first-order diffusion model. Did not consider the scenario in which the elements have bioaccumulated in the fulmar. Maximum concentration gradients and evaluation of bioaccessibility percentage based on maximum leached (equilibrium concentrations) as in the experiment are not relevant in nature</td></tr><tr><td>Lee et al. (2019)</td><td></td><td>X</td><td></td><td>Evaluated relevance of ingested microplastics to overall transfer of chemicals into fish using simulated intestinal fluid and model analysis. Model analysis did not include error estimates and did not reflect experimental results as it was not calibrated. Did not evaluate the results in the context of the actual environment</td></tr></tbody></table>

(continued)
<table><caption>**Table 6.4 (continued)**</caption><thead><tr><td>Study</td><td>Demonstrated</td><td>Inconclusive</td><td>Not supported</td><td>Comments</td></tr></thead><tbody><tr><td>Coffin et al. (2019)</td><td></td><td>X</td><td></td><td>Evaluated leaching of plastic additives from commonly ingested plastic items in gut mimic models for fish and seabirds. Used virgin plastic items, maximizing the concentration gradient between plastic and gut fluid, which is not relevant in nature. Single time point statistical analysis of the concentration leaching out from plastic and its effects on estrogenicity only reflects the vector effect under such experimental conditions. Results were not evaluated in the context of other relevant processes</td></tr><tr><td>Guo et al. (2019)</td><td></td><td>X</td><td></td><td>Investigated leaching of flame retardants from acrylonitrile-butadiene-styrene (ABS) in simulated avian digestive systems and the effect of co-ingesting sediment. Did not consider further evaluation of results with models or calculations with chemical characteristics (K<sub>OW</sub>). Results indicated chemicals transferred from plastic to sediment. However, the study did not consider sorption capacities of the components in the system and its exchange fluxes</td></tr><tr><td>Martin and Turner (2019)</td><td></td><td>X</td><td></td><td>Evaluated the mobilization of cadmium under simulated digestive conditions over 6h. Model was limited and did not consider the other phases and exchange fluxes. Applicability of the rate constants obtained by fitting a second-order diffusion model to the experimental data was not discussed. Evaluation of environmental implication is not relevant in nature as there is no clean sediment in nature. Sediment used in experiment was collected from a protected and unpolluted location which is not representative of other areas</td></tr></tbody></table>

(continued)
<table><caption>**Table 6.4 (continued)**</caption><thead><tr><td>Study</td><td>Demonstrated</td><td>Inconclusive</td><td>Not supported</td><td>Comments</td></tr></thead><tbody><tr><td>Mohamed Nor and Koelmans (2019)</td><td></td><td></td><td>X</td><td>Evaluated chemicals exchange of PCBs in gut fluid mimic systems considering different scenarios representing different fugacity levels between plastic and organism. Considered all exchange fluxes in the system and bimodal behaviour in model analysis, showing error estimates for parameters. Considered exchange fluxes with other components in the gut such as food. Discussed bioavailability from ingested plastic to organisms under different scenarios</td></tr><tr><td>Heinrich and Braunbeck (2020)</td><td></td><td>X</td><td></td><td>Studied the effect of PE MP addition to rainbow trout RTL-W1 cells dosed with 7-ethoxyresorufin-O-deethylase (EROD) inducers by addition. The addition of MP reduced EROD activity. The authors concluded that the presence of MP can reduce the amount of bioavailable pollutants in situ ('cleaning' effect) but that it remains unclear to what extent this mechanism occurs under natural conditions</td></tr><tr><td>Kühn et al. (2020)</td><td></td><td>X</td><td></td><td>Observed leaching of 5 (out of 15) chemicals to stomach oil sampled from fulmars. Results of different experiments were said to be inconsistent. Concentrations in the oil could have decreased due to biodegradation, increasing the gradient for transfer. Biomagnification from regular food was not addressed. Chemical transfer from the oil to the plastic was not considered. It was acknowledged that the experimental data should not be compared with models simulating natural conditions</td></tr></tbody></table>

General approach, type of evidence and environmental realism are summarized

system, which is unrealistic. This may result in an overestimation of the amount leached from plastic. Kühn et al. (2020) used natural contaminated stomach oil and indeed only found limited transfer for five and no transfer for the other ten chemicals studied. Only few studies relate experimental observations with the chemical and sorbent characteristics. Most studies did not provide error estimates or significance levels for their parameter estimations. Due to these limitations, 12 out of 13 studies were evaluated as inconclusive. The one study providing conclusive evidence demonstrated that the occurrence of a vector effect was context-dependent
(Mohamed Nor and Koelmans 2019). They showed that the boundary conditions, i.e. where either the MP or the food was initially contaminated, or both, determined whether transfer occurred and in which direction.

### 6.3.4 Weight of Evidence from Modelling Studies

Twelve studies used models to investigate the microplastic vector effect (Table 6.5). Most of them were explicitly applied to simulate exposure to plastic-associated chemicals under natural conditions, taking multiple exposure pathways into account, or to simulate realistic experiments that also had multiple exposure pathways covered in their experimental design. These studies generally provided conclusive

<table><caption>**Table 6.5** Overview of studies that used simulation models to address the role of microplastic ingestion by organisms on bioaccumulation of plastic-associated chemicals</caption><thead><tr><td>Study</td><td>Demonstrated</td><td>Inconclusive</td><td>Not supported</td><td>Comments</td></tr></thead><tbody><tr><td>Teuten et al. (2007)</td><td></td><td>X<sup>a</sup></td><td></td><td>An equilibrium model scenario predicted a cleaning effect of plastics in reducing contaminant concentrations in benthic organisms. Another scenario predicted enrichment of chemical concentrations on plastic in the surface microlayer (SML) leading to higher exposure to benthic organisms. The acclaimed SML enrichment, however, has been argued to be based on a misinterpretation, as discussed in Koelmans (2015). Desorption during microplastic settling and burial was not accounted for. Uptake by sediment or food ingestion was not considered</td></tr><tr><td>Gouin et al. (2011)</td><td></td><td></td><td>X<sup>a</sup></td><td>Considered all known accumulation pathways in order to quantitatively assess the relative importance of plastic ingestion to total bioaccumulation. A worst case was considered by assuming concentrations in plastic and tissue to be at steady state. Provided mechanistic evidence based on first principles</td></tr><tr><td>Koelmans et al. (2013b)</td><td></td><td></td><td>X</td><td>Considered all known accumulation pathways in order to quantitatively assess the relative importance of plastic ingestion to total bioaccumulation for a previously published dataset. Provided mechanistic evidence based on first principles, as well as causal evidence on the treatment level by validation with empirical data</td></tr></tbody></table>

(continued)
<table><caption>**Table 6.5** (continued)</caption><thead><tr><th>Study</th><th>Demonstrated</th><th>Inconclusive</th><th>Not supported</th><th>Comments</th></tr></thead><tbody><tr><td>Koelmans et al. (2014)</td><td></td><td></td><td>X<sup>a</sup></td><td>Compared bioaccumulation due to plastic ingestion only, with total observed bioaccumulation in the field. Provided mechanistic evidence based on first principles. The model was validated elsewhere (Koelmans et al. 2013b, 2016)</td></tr><tr><td>Bakir et al. (2016)</td><td></td><td></td><td>X</td><td>Considered all known accumulation pathways in order to quantitatively assess the relative importance of plastic ingestion to total bioaccumulation. Neglected backward sorption. The 50% scenario can be considered unrealistic or worst case. Provided mechanistic evidence based on first principles</td></tr><tr><td>Koelmans et al. (2016)</td><td></td><td></td><td>X<sup>a</sup></td><td>Simulated a series of published experiments using spiked plastic and clean organisms. Provided mechanistic model validation based on three lines of evidence: (a) intropolymer diffusion, (b) in vitro desorption kinetic data to artificial gut fluids and (c) evaluation against experimental data from three published datasets. Simulations representing natural exposure conditions demonstrated the vector effect to be negligible</td></tr><tr><td>Herzke et al. (2016)</td><td></td><td></td><td>X</td><td>Modelled chemical fluxes of microplastic ingested by seabirds, accounting for all uptake pathways. Revealed that plastic was more likely to act as a passive sampler than as a vector for chemicals</td></tr><tr><td>Rochman et al. (2017)</td><td></td><td>X</td><td></td><td>Used the model developed by Koelmans et al. (2013b) to simulate chemical concentrations from a dietary uptake experiment, which however could not be measured due to detection limit problems. Only uptake via plastic was addressed; uptake from food or water was not accounted for</td></tr><tr><td>Besseling et al. (2017)</td><td></td><td></td><td>X</td><td>Modelled experimental data on chemical uptake by lugworms accounting for all pathways (plastic water, food). Aqueous exposure was assessed using passive samplers. The ingestion vector effect was demonstrated to be irrelevant</td></tr></tbody></table>

(continued)
<table><caption>**Table 6.5** (continued)</caption><thead><tr><td>Study</td><td>Demonstrated</td><td>Inconclusive</td><td>Not supported</td><td>Comments</td></tr></thead><tbody><tr><td>Diepens and Koelmans (2018)</td><td></td><td>X (PAHs)</td><td>X (POP)</td><td>Reported a food web model (MICROWEB) based on the model by Koelmans et al. (2013b), which has been validated for POPs. For POPs no vector effect was predicted across all trophic levels of the food web (evidence for ‘non-supported’). For degrading compounds, a vector effect was predicted, which, however, has not yet been experimentally validated and therefore was rated ‘inconclusive’</td></tr><tr><td>Lee et al. (2019)</td><td></td><td></td><td>X</td><td>Considered all known pathways to assess the contribution of plastic ingestion to chemical bioaccumulation with an uncertainty analysis using Monte-Carlo simulation. Did not use results from in vitro experiments in model to obtain a more refined model</td></tr><tr><td>Wang et al. (2019)</td><td></td><td></td><td>X</td><td>Modelled experimental data on chemical uptake by earthworms (Eisenia fetida) accounting for all pathways (plastic, water, food), using the model by Koelmans et al. (2013b). Concentrations in soil (food) and water were measured. Modelled and empirical data agreed well and no vector effect was found</td></tr><tr><td>Wang et al. (2020a)</td><td></td><td>X</td><td></td><td>Modelled experimental data on chemical uptake by earthworms (Eisenia fetida) accounting for all pathways (plastic, water, food), using the model by Koelmans et al. (2013b) for contaminated plastic – clean worm scenarios (‘vector effect’). Chemical transfer to the worms was observed with dermal exposure generally dominating overexposure through MP ingestion. A minor vector effect thus was found, but it remained unclear to what extent this would apply under natural conditions</td></tr></tbody></table>

(continued)
<table><caption>**Table 6.5** (continued)</caption><thead><tr><td>Study</td><td>Demonstrated</td><td>Inconclusive</td><td>Not supported</td><td>Comments</td></tr></thead><tbody><tr><td>Wang et al. (2020b)</td><td></td><td>X</td><td></td><td>Modelled experimental data on chemical uptake by earthworms (Eisenia fetida) accounting for all pathways (plastic, water, food), using the model by Koelmans et al. (2013b) for a contaminated plastic – clean worm scenario (‘vector effect’), as well as the opposite (‘cleaning effect’). Both effects were found implying that the vector effect is context-dependent</td></tr></tbody></table>

General approach, type of evidence for an effect on chemical risk and environmental realism are summarized

aspects, such as release kinetics, gut retention times or reversible (backward) sorption, aspects that however would lead to an overestimation of the importance of the vector effect. Studies that considered such ‘worst-case’ conditions but still concluded that a vector effect was minor were evaluated as conclusive. Diepens and Koelmans (2018) provided theory suggesting that chemicals present at higher than equilibrium concentrations, like additives, in some cases can (over-)compensate for the simultaneously occurring cleaning effect, leading to a net MP vector effect if plastic mass concentrations are high. Furthermore, they calculated that degradable compounds such as PAH may also be prone to a vector effect due to the chemicals being less bioavailable for metabolization during gut passage. It must be noted that the credibility of their model applies to all scenarios modelled (equilibrium and ‘under’-equilibrium) and that the scenarios that represent equilibrium conditions for persistent (i.e. non-metabolizable) chemicals are more likely to be relevant (Koelmans et al. 2016; De Frond et al. 2019).

## 6.4 Risk Assessment of Plastic-Associated Chemicals: A Case Study Illustrating the Relevance of the MP Vector Effect for Risks of Plastic-Associated Chemicals in San Diego Bay

The previous sections have shown that studies addressing the MP vector effect often did not put their results in an environmentally realistic context or did not address results in the context of chemical risks. Here, the notion ‘risk’ is not meant as probability, hazard, harm or threat but as the extent to which exposure chemical concentrations exceed a known effect threshold for the species for which the risk is then assessed. We use this definition because it is objective and quantitative and

<footnote>ªSimilarly assessed earlier by Burns and Boxall (2018)</footnote>
because it is the metric that is used for risk assessment of chemicals in regulatory frameworks (Koelmans et al. 2017). As mentioned, this context of actual risks is relevant, because if a detected MP vector effect would not affect risks of the chemicals, then it might be less urgent to study it and there might be less reason for concern.

**Description of the Case** For the case study, we use the pioneering laboratory experiment by Rochman et al. (2013) where *Oryzias latipes* (Japanese medaka) were exposed to MP loaded with HOCs sorbed from San Diego Bay seawater. Note that this Asian species was used as test organisms and was not claimed to live in San Diego Bay. For the case study, we use the PCB data from their study and calculate back what medaka would have been exposed to, when they would actually reside in San Diego Bay. In the laboratory study, the observed increase in PCB bioaccumulation by medaka was ascribed to the ingestion of MP. The MP was equilibrated in San Diego Bay, which yielded field relevant PCB concentrations on the plastic and which also contributed to the environmental relevance of the study. The fact that the MPs acquired PCBs from the seawater implies that there were PCBs in the water. Therefore, in the bay, medaka would have been exposed to plastic-associated PCBs as well as to the same PCBs dissolved in the water. Additionally, fish would also be exposed to PCBs in food (prey), which, similar to the MPs, would have absorbed or bioconcentrated PCBs from the ambient water. In the laboratory experiment, however, aqueous phase concentrations were kept at zero. Furthermore, in the experiment, fish was fed PCB contaminated cod liver oil. PCB concentrations were known, but these were not equal to concentrations that would be present in natural food at equilibrium in the bay. Furthermore, in the laboratory experiment, fish were already contaminated with PCBs, but at lower than equilibrium concentrations, because uptake was observed, in the controls as well as in the plastic treatments. In contrast, in the bay, chemical concentrations in small fish like medaka would have been at steady state.

**Calculation of the Contribution of Plastic in the Laboratory Experiment** Koelmans et al. (2016) modelled the experiment published by Rochman et al. (2013) with aqueous concentration set to zero and ingestion rate and concentrations in food set at the values reported in the study. This model framework has been developed, applied and validated (Table 6.5) in a series of studies (Koelmans et al. 2013a, b, 2014, 2016; Koelmans 2015; Rochman et al. 2017; Wang et al. 2019, 2020a, b). A similar model has been applied by Bakir et al. (2016) to assess the relative contribution of chemical uptake pathways for marine worms, fish and seabirds. This modelling shows that in the experiment (Rochman et al. 2013), the contribution of the ingested MP to bioaccumulation by medaka in the experiment can be estimated to range from 3% to 100%, depending on the PCB congener (Table 6.6).