Status Reports
Microplastics in Global Marine Waters and Biota: Effectiveness of Potential Bioindicators in Mirroring Local Pollution Levels
2023 Volume 3 Pages 43-68
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2023 Volume 3 Pages 43-68
The ever-increasing production, poor waste management, and persistent nature have caused plastic debris to become ubiquitous in the environment. Microplastics, usually defined as plastic particles less than 5 mm, have invaded our coasts, oceans, and seafood. Thus, there is a dire need to fully monitor and understand the global spatial distribution of microplastics in marine waters and biota, especially in commercial species. This review summarizes the microplastic abundance in the water column of both coastal and offshore areas, as well as microplastic concentrations in various biota globally. In addition, this review aims to discuss the effectiveness and limitations of various biota as bioindicators of microplastic pollution in marine environments. Coastal waters of South America, Asia, and Africa were reported to have very high abundances of microplastics, with abundances of 782,000, 19,000, and 1,200 particles/m3, respectively, possibly due to excessive input and mismanagement of plastic waste. Microplastic abundances in open oceans were significantly lower (up to three orders of magnitude) than those in coastal areas, apart from some convergence zones such as gyres. Microplastics were also detected in remote pristine areas, such as the Arctic and Antarctica, demonstrating the omnipresence and ability of microplastics to be transported long distances via atmospheric and oceanographical forces. Due to the lack of standardization, quantifiable particle size and sampling methodology were shown to greatly affect the results of microplastic abundance, making them important variables to note when comparing multiple studies. Most marine biota species were able to reflect local microplastic pollution levels to a certain extent. Filter feeders are popular bioindicator species, but their ingestion is limited to particles smaller than 1,500 μm. Although selective feeders have the capacity to ingest a larger size range of plastic particles (up to 5,000 μm or larger), their wide range of movements, ability to reject microplastics, and highly varied behaviors make them less reliable as bioindicators than filter feeders. Nonetheless, a greater variety of marine species should be monitored to ensure population health and food safety.
Plastic debris, especially in the marine environment, has been a long-standing issue, with floating plastic debris reported in open oceans as early as the 1970s (Carpenter and Smith, 1972). Since then, studies on plastic debris, including microplastics, have been steadily increasing. Despite reports and an increase in awareness regarding the negative impacts of plastics on the marine environment, global plastic production has continued to rise, with 367 million tonnes (Mt) of new plastics reported to be produced in 2020 alone (Plastics Europe, 2021). It has also been estimated that approximately 6,500 Mt of plastic waste was generated in 2015. From that, only about 9% was said to be recycled, 12% was incinerated, and the remaining 79% was merely discarded, mostly to landfills (Geyer et al., 2017). In addition, up to 28% of discarded plastic waste could end up in the marine environment owing to mismanagement (Jambeck et al., 2015). Once these plastic wastes enter the environment, they are exposed to the elements and start to break down into smaller particles over time, becoming microplastics. Although microplastics are usually defined as plastic particles smaller than 5 mm (GESAMP, 2010), micro- to nanometer-sized plastic particles have been observed in marine environments. The volume of microplastics in the environment is currently not only omnipresent and detrimental, but also ever-evolving because there are still constant inputs of plastic waste into the environment despite recent efforts to curb plastic pollution. Additionally, plastic particles have been detected in marine organisms throughout the food web, from the base, such as zooplankton (Frias et al., 2014; Setälä et al., 2014; Desforges et al., 2015; Zheng et al., 2020),up to top predators (Romeo et al., 2015; Nelms et al., 2018), suggesting that microplastics may potentially accumulate through the food chain. It has also been shown that microplastic particles are not only able to enter the human body but also translocate to sensitive organs, causing numerous health effects (Zarus et al., 2021). Apart from their obvious physical impacts, microplastics have also been shown to act as vectors for chemical pollutant transfer into the food web (Teuten et al., 2009; Yamashita et al., 2011; Tanaka et al., 2013, 2018). This could affect food safety, especially in seafood, and eventually pose risks to human health.
Apart from the obvious risks that microplastics pose to marine organisms, they have also been shown to negatively affect sensitive and rich marine habitats, such as coral reefs and mangroves (Chapron et al., 2018; Reichert et al., 2018; John et al., 2022). Thus, it is important to not only constantly monitor the spread of microplastics in the marine environment but also to have a holistic understanding of the current global distribution. Although there have been comprehensive reviews regarding microplastic in the marine environment (Andrady, 2011; Ivar do Sul and Costa, 2014; Tang et al., 2021), they have mainly focused on improvements in methodologies, identification of sources, and overall impacts. This review paper intends to focus not only on distribution of microplastics in the water column of seas and oceans globally, but also on microplastic abundance in marine organisms. Likewise, there have been reviews on the myriad methodologies used to quantify microplastics in biota (Wesch et al., 2016; Hermsen et al., 2018) and concerns about food safety and human health in coastal areas (Rezania et al., 2018; Hantoro et al., 2019). However, this study aims to take a more monitoring approach by reviewing the current state of spatial microplastic distribution in the marine environment and comprehend how microplastic abundance in the water column translates to microplastic concentrations in marine biota within the same area. Moreover, reviewing the correlation between abundance in environmental matrices and multiple types of biotas will aid in distinguishing possible bioindicators and their effectiveness in reflecting the surrounding microplastic pollution. Although there have already been regional case studies suggesting filter feeders, such as bivalves, as bioindicators for microplastic monitoring (Li et al., 2019; Ding et al., 2021; Staichak et al., 2021) and barnacles (Xu et al., 2020), this review intends to confirm if this phenomenon is shared over a wider scope. Thus, this review aims to examine the global spatial trends of microplastic distribution in both selective and nonselective feeders in relation to microplastic abundance in the water column.
Owing to the increasing number of publications on microplastic pollution in recent years, this review aims to focus on more recent papers that consider publications available up to 2022. An extensive search was performed for peer-reviewed original research articles pertaining to microplastic pollution in marine waters and selected marine organisms. Specific keywords were used when searching through various databases, such as microplastics, marine, water, ocean, sea, coastal, shellfish, mussel, oyster, filter feeders, scallops, predators, foragers, and fish. A total of 86 papers were selected for this review based on how closely the data could be compared to one another. Literature selection was based on three criteria. 1) Units used in the calculation of abundance for better comparison, 2) use of instruments for polymer identification for better accuracy of microplastic identification, and 3) survey locations for a wider and more comprehensive understanding of global status. For literature used in the discussion of microplastic abundance in global marine waters, only papers that reported in the unit of particles/m3 or had the necessary information needed to convert the abundances to particles/m3 were selected. For literature used in the discussion of microplastic abundance in biota, literature reporting in the units particles/individual (n/ind.) or particles/gram (N/g) or both were used. This is to ensure that comparisons between locations and literature are more accurate. In addition, priority in terms of literature selection for discussion in both microplastic abundances in marine waters and biota was given to literature that uses polymer identification instruments such as any form of Fourier-transform infrared spectroscopy (FTIR), Raman spectrometry, and near-infrared hyperspectral imaging. This was to minimize any overestimation or underestimation of microplastic abundance due to misidentification. Also, to gain a wider understanding of the global status of microplastic pollution, some studies reporting on less studied areas, such as the African and South American continents, were still selected despite not using any polymer identification instruments while still satisfying the criteria for standardized units. The selected papers were then divided into two main categories: 1) microplastics in global marine waters and 2) microplastics in biota. The literature will also be discussed according to geographical region for a more systematic review of global microplastic distribution.
The collected literature data on microplastic abundance in marine waters are presented in Tables 1 and 2. Although the units were standardized for comparison here, owing to the variations in methodologies used in each study, some margins of differences are expected. In addition, not all desired information and data were available from all papers. Since the early 2000s, an increase in research and publications on microplastic pollution in the marine environment has been observed (Andrady, 2011; Barboza and Gimenez, 2015), with most surveys focusing on coastal habitats (Browne et al., 2011). Most of these studies were conducted in the Asian region, whereas a very limited number of studies focused on the Middle East. A variety of microplastic particle sizes were quantified, with studies reporting particles as small as 10 μm. Obviously, the ability to quantify smaller microplastics particles and cover a larger range of particle sizes would result in more reported particles, as shown by Fernández Severini et al. (2019). Hence, it is also important to note the quantifiable particle size ranges when comparing microplastic concentrations between literature.
| Location | MP Abundance (particles/m3) | Identification Method | MP Size Range (μm) | Polymer Types | Sampling Year | Reference |
|---|---|---|---|---|---|---|
| Asia | ||||||
| Incheon coast, South Korea | 1.96 | FTIR | 330–5,000 | PES, PP, Others | 2015–2017 | Kwon et al. (2020) |
| Gwangyang Bay, South Korea | 1.65 | |||||
| Busan coast, South Korea | 1.35 | |||||
| Ulsan Bay, South Korea | 4.73 | |||||
| Youngil Bay, South Korea | 4.54 | |||||
| Cheonsu Bay, South Korea | 2.79 | |||||
| Hampyeon Bay, South Korea | 1.7 | |||||
| Deukryang Bay, South Korea | 1.12 | |||||
| Jinhae Bay, South Korea | 88±68 | FTIR | <50–2,000 | PP, PE, PS, PES, synthetic rubber | 2013 | Song et al. (2015) |
| Rural coast, South Korea | 560 | μ-FTIR | >20 | PP, PE, EVA | 2016–2017 | Song et al. (2018) |
| Urban Coast, South Korea | 1,051 | μ-FTIR | >20 | PP, PE, EVA | ||
| Coast of Korea | 1,400±560 | μ-FTIR | >20 | PP, PE, PS, PES, PET, PVC | 2017 | Cho et al. (2021) |
| Iyo Sea, Japan | 346 | FTIR | 300–5,000 | PP, PE, Others | 2010–2011 | Isobe et al. (2014) |
| Hyuga Sea, Japan | 90 | 2011 | ||||
| Tokyo Bay, Japan | 2.47–20.5 | FTIR–ATR | 350–5,000 | PE, PP, PS, PA, PET | 2020 | Xu et al. (2022) |
| Tokyo Bay, Japan | 3,060–3,220 | μ-FTIR | 50–350 | PE, PP, PA, PS, PET, PVC | ||
| Western Harbour, Xiamen, China | 695.8±735.0 | FTIR | 330–5,000 | PE, PP, PS, PES, PU | 2017 | Tang et al. (2018) |
| Hong Kong Coast | 0.61–271 | FTIR | 30–4,960 | PP, HDPE, LDPE, styrene acrylonitrile | 2015–2016 | Tsang et al. (2017) |
| Guangdong Coast, China | 3,000–19,000 | FTIR | 0.45–4,000 | na | 2020 | Zhang et al. (2020) |
| Beibu Gulf, Guangxi, China | 399–5,531 | μ-Raman spectrometer | 1.2–5,000 | PP, PE | 2018 | Li et al. (2020) |
| Jiaozhou Bay, China | 20–120 | μ-FTIR–ATR | 20–4,000 | PET, PP, PE | 2017 | Zheng et al. (2019) |
| Lamong Bay, Surabaya, Indonesia | 380–610 | FTIR–ATR | >450 | PS, PE, PP, PU, PET, PES | 2017 | Cordova et al. (2019) |
| Kenjeran Beach, Indonesia | 460–550 | |||||
| Wonorejo Beach Indonesia | 440–530 | |||||
| Thottappally, Kerala, India | 3.58 | FTIR–ATR | 300–5,000 | PE, PP | 2018 | Robin et al. (2020) |
| Thaikadappuran, Kerala, India | 0.22 | |||||
| Mumbai, India | 165–547 | Raman spectrometry | >333 | PP, PE, PS, PA, PES, nylon | 2019–2020 | Gurjar et al. (2022) |
| Terengganu Estuary, Malaysia | 421.8±110 | μ-FT–IR | >20 | PA, PE, PP | 2018 | Taha et al. (2021) |
| Lukut Estuary, Malaysia | 4,170 | FTIR | >25 | PES, PE | 2020 | Zainuddin et al. (2022) |
| Port Dickson Coast, Malaysia | 4,990 | |||||
| Inner Gulf of Thailand | 21.29±36.21 | μ-FTIR, FTIR–ATR | 100–5,000 | PP, PE | 2018 | Vibhatabandhu and Srithongouthai (2022) |
| Bandon Bay, Thailand | 0.28±0.07–0.63±0.13 | FTIR | 400–1,500 | PP, PE, PET, nylon, rayon | 2019 | Chinfak et al. (2021) |
| Rayong Coast, Thailand | 200–8,500 | FTIR–ATR | 1.2–5,000 | PP, PE, PS | 2019 | Prarat and Hongsawat (2022) |
| North America | ||||||
| West Coast Vancouver Island, Canada | 1,710±1,110 | Optical microscopy | 62–5,000 | na | 2012 | Desforges et al. (2014) |
| Queen Charlotte Sound, Canada | 7,630±1,410 | |||||
| Strait of Georgia, Canada | 3,210±628 | |||||
| Moss Landing Harbour, California, USA | 4–11 | Raman spectrometry | >100 | PET, PA, PC, PVC | 2017 | Choy et al. (2019) |
| Gulf of Maine, USA | 6.03±1.03 | FTIR–ATR | >100 | PE, PES | 2013 | Lindeque et al. (2020) |
| South America | ||||||
| Esmeraldas, Ecuador | 6–340 | Stereomicroscope Amscope | >63 | na | 2020 | Capparelli et al. (2021) |
| Río dela Plata estuary, Argentina | 139 | Stereomicroscope | >36 | na | 2016–2017 | Pazos et al. (2018) |
| Mangrove Creeks, Goiana Estuary, Brazil | 0.03 | Stereomicroscope | 300–5,000 | na | 2012–2013 | Lima et al. (2015) |
| Guanabara Bay, Río de Janeiro, Brazil | 1.40–21.3 | FTIR–ATR | 300–5,000 | PP, PE | 2016 | Olivatto et al. (2019) |
| Bahía Blanca Estuary, Argentina | 5,900–782,000 | Stereoscopic Microscope | 0.22–5,000 | na | 2018 | Fernández Severini et al. (2019) |
| Bahía Blanca Estuary, Argentina | 42.6–113.6 | 60–5,000 | ||||
| Argentinian continental shelf, Argentina | 0.14±0.08 | Stereomicroscope | 350–5,000 | na | 2018 | Ronda et al. (2019) |
| Brazilian Equatorial Margin, Brazil | 0.14±0.11 | Stereomicroscope | >120 | na | 2010 | Garcia et al. (2020) |
| Enseada Beach, Acaraí Lagoon, Brazil | 0.03±0.04 | Stereomicroscope | >500 | na | 2018 | Lorenzi et al. (2020) |
| Comtinental waters, Galápagos Islands | 0.26±0.08 | Stereomicroscope | 150–5,000 | na | 2017 | Alfaro-Núñez et al. (2021) |
| International water stations, Galápagos Islands | 0.36±0.10 | |||||
| Eastern Galápagos station | 0.24±0.09 | |||||
| Western Galápagos station | 0.22±0.08 | |||||
| Saigon Bay, Bocas del Toro, Panama | 107,000±25,000 | Fluorescent microscope | >10 | na | 2019 | Fallon and Freeman (2021) |
| Europe | ||||||
| Black Sea | 600–1,200 | Binocular microscope | 200–5,000 | na | 2014–2015 | Aytan et al. (2016) |
| Esposende, Portugal | 0.015±0.014 | FTIR | 500–5,000 | PE, PET, PA, PP, nylon | 2019 | Rodrigues et al. (2020) |
| Matosinhos Coast, Portugal | 2,748±2,617 | FTIR | 30–5,000 | PE | 2019 | |
| Porto of Leixões, Portugal | 4,028±1,878 | PE, PP, nylon | 2020 | |||
| English Channel, UK | 10.03±2.21 | FTIR–ATR | >100 | PES, PP, acrylic, PE, PA, PVC | 2015 | Lindeque et al. (2020) |
| Skagerrak, Sweden | 0.02–0.05 | Near-infrared hyperspectral imaging | >300 | PE, PP | 2014 | Schönlau et al. (2020) |
| Kattegat, Sweden | 0.01–0.02 | |||||
| Southern Baltic Proper | 0.03–0.16 | |||||
| Western Gotland Basin | 0.13–0.46 | |||||
| Bothnian Sea | 0.04–0.12 | |||||
| The Quark, Sweden | 0.02–0.04 | |||||
| Northern Baltic Proper | 0.00–0.04 | |||||
| Eastern Gotland Basin | 0 | |||||
| Aveiro, Portugal | 0.002±0.001 | μ-FTIR | >180 | PP, PA, PE | 2002 | Frias et al. (2014) |
| Lisboa, Portugal | 0.033±0.021 | 2005–2008 | ||||
| Costa Vicentina, Portugal | 0.036±0.027 | 2007 | ||||
| Algarve, Portugal | 0.014±0.012 | 2006 | ||||
| Africa | ||||||
| South Africa South Eastern coastline | 257.9±53.36–1,215±276.7 | Dissecting microscope | 80–5,000 | na | 2014 | Nel and Froneman (2015) |
| Mdloti, Durban, KwaZulu-Natal, South Africa | 39.8±32.4 | FTIR–ATR | 300–5,000 | PS, PU, PES, nylon, PP | na | Naidoo et al. (2015) |
| uMgeni, Durban, KwaZulu-Natal, South Africa | 83.4±46.0 | |||||
| Durban Harbour, Durban, KwaZulu-Natal, South Africa | 319.8±542.4 | |||||
| Isipingo, Durban, KwaZulu-Natal, South Africa | 95.2±45.6 | |||||
| iLovu, Durban, KwaZulu-Natal, South Africa | 27.4±11.2 | |||||
| Richard’s Bay Harbour, South Africa | 413.3±77.53 | dissecting microscope | 63–5,000 | na | 2016 | Nel et al. (2017) |
| Durban Harbour, South Africa | 1,200±133.2 | |||||
| Denu, Ghana | 2.36±1.95 | FTIR–ATR | 333–5,000 | PP, PE, PS | 2021 | Nuamah et al. (2022) |
| Tema, Ghana | 2.79±2.79 | |||||
| Cape Coast, Ghana | 1.14±0.63 | |||||
| Sekondi, Ghana | 1.33±1.02 | |||||
| Middle East | ||||||
| Qatar’s Exclusive Economic Zone | 0–3 | FTIR | 125–5,000 | PP, PE, LDPE, PA, ABS, PMMA, cellophane | 2015 | Castillo et al. (2016) |
| Oceania | ||||||
| Coast of Kimberly, Australia | 0.01–0.41 | FTIR–ATR | 350–5,000 | PE, PMMA, PP, PES, PU | 2015 | Kroon et al. (2018) |
| Ceduna, Australia | 2.95±0.42 | FTIR | 50–5,000 | PA, PE, PET, PP | 2019 | Klein et al. (2022) |
| Semaphore Beach, Adelaide, Australia | 16.31±2.42 | |||||
| South Australian Coastline | 8.16±0.55 | |||||
| Vava’u archipelago, Tonga | 1.05±0.13 | FTIR | >100 | PES, PP, PE | 2017 | Markic et al. (2022) |
FTIR: Fourier transform infrared, ATR: Attenuated Total Reflection
PE: polyethylene, PP: polypropylene, PES: polyester, PET: polyethylene terephthalate, PMMA: polymethylmethacrulate, PU: polyurethane, PS: polystyrene, PA: polyamide, PVC: polyvinyl chloride, HDPE: high-density polyethylene, LDPE: low-density polyethylene, EVA: ethylene vinyl acetate, ABS: acrylonitrile butadiene styrene; na=not available.
| Location | MP Abundance (particles/m3) | Identification Method | MP Size Range (µm) | Polymer Types | Sampling Year | Reference |
|---|---|---|---|---|---|---|
| Open Oceans/Seas/Offshore | ||||||
| Western Tropical Atlantic Ocean, Abrolhos Archipelago | 0.04 | stereomicroscope | 300–5,000 | na | 2011–2013 | Ivar do Sul el al. (2014) |
| Western Tropical Atlantic Ocean, Fernando de Noronha | 0.015 | |||||
| Western Tropical Atlantic Ocean, Trindade Island | 0.025 | |||||
| Tropical Atlantic Ocean | 0.03 | |||||
| North East Atlantic Ocean | 0–1.5 | dissecting microscope | 333–5,000 | na | 2011 | Maes et al. (2017) |
| Atlantic Ocean | 0–2.5 | FTIR–ATR | 250–5,000 | PES, PA, PP. acrylic, PVC, PS, PU, PET | 2015 | Kanhai et al. (2017) |
| Atlantic Ocean, offshore of Namibia | 8.5 | |||||
| Atlantic Ocean, West coast of Morocco | 6–6.5 | |||||
| Atlantic Ocean, Bay of Biscay | 3.5 | |||||
| Atlantic Ocean, western coast of Portugal | 3.5 | |||||
| European Coast to North Atlantic Subtropical Gyre | 13–501 | μ-Raman spectrometer | 10–10,000 | PE, PP, PS, PA, PMMA, PU, PVC, PES | 2014 | Enders et al. (2015) |
| Monterey Bay, California, USA | 8–9 | Raman spectrometry | >100 | PET, PA, PC, PVC | 2017 | Choy et al. (2019) |
| North East Pacific Ocean | 279±178 | optical microscopy | 62–5,000 | na | 2012 | Desforges et al. (2014) |
| Arctic Ocean | 0–18 | FTIR | 100–5,000 | PES, PA, PVC | 2016 | Kanhai et al. (2020) |
| Arctic Ocean, Svalbard, Norway | 0–11.5 | FTIR | >250 | PES, PA, PE, acrylic, PVC | 2014 | Lusher et al. (2015) |
| Barent Sea | 0.005 | FTIR | 200–5,000 | PE, PU, PVC, PES, PA, PS, PP | 2019 | Yakushev et al. (2021) |
| Kara Sea | 0.003 | |||||
| Laptev Sea | 0.002 | |||||
| East-Siberian Sea | 0.01 | |||||
| Southern Ocean, Antarctica | 6.2×10−4 | μ-FTIR | 200–5,000 | PE, PP, PS, PVC, nylon, PMMA | 2016–2017 | Suaria et al. (2020) |
| Sub-Antarctic Zone | 8.4×10−4 | |||||
| Antarctic Polar Front | 1.9×10−4 | |||||
| Antarctic Zone | 1.18×10−3 | |||||
| Southern Ocean, Antarctica | 3.1×10−2 | FTIR | 350–5,000 | PE, PP, PS, PVC | 2016 | Isobe et al. (2017) |
| East Asian Seas | 0.03–491 | FTIR | 350–5,000 | na | 2014 | Isobe et al. (2015) |
| Eastern Sea, China | 252.0±58.90 | FTIR | 330–5,000 | PE, PP, PS, PES, PU | 2017 | Tang et al. (2018) |
| Tokai region offshore, Japan | 0.04–0.17 | FTIR-ATR | 350–5,000 | PE, PP, PS, PA, PET | 2020 | Xu et al. (2022) |
| Tokai region offshore, Japan | 1,028–5,091 | μ-FTIR | 50–350 | PE, PP, PA, PS, PET, PVC | ||
| Terengganu offshore, Malaysia | 211.2±104 | μ-FTIR | >20 | PA, PE, PP | 2018 | Taha et al. (2021) |
| Southern Indian Ocean | 0–12 | μ-FTIR | 110–5,000 | PET, PE, Rayon, PA, PVDC | na | Li et al. (2022) |
FTIR: Fourier transform infrared, ATR: Attenuated Total Reflection
PE: polyethylene, PP: polypropylene, PES: polyester, PET: polyethylene terephthalate, PMMA: polymethylmethacrulate, PU: polyurethane, PS: polystyrene, PA: polyamide, PVC: polyvinyl chloride, PVDC: polyvinylidene chloride; na=not available.
Coastal areas include mangroves, estuaries, coral reefs, sea grasses, coastal shelves, and salt marshes (Burke et al., 2001). These coastal areas are important because they provide highly valuable goods and services to their surrounding communities and societies. Hence, coastal areas are usually highly populated with large settlements (Martínez et al., 2007). Because of the overwhelming amount of anthropogenic activities along global coastlines, these coastal zones can be considered a significant source of plastic pollution and are also the input points at which inland plastic waste flows into the oceans from rivers (Chenillat et al., 2021). Plastic wastes that enter the marine environment then spread from coasts into open oceans via currents, winds, and waves.
Asia has been an area of focus for microplastic monitoring, with most articles conducting case studies in China, Korea, and Japan. Kwon et al. (2020) and Song et al. (2018) reported on several urban and rural coastal locations along South Korea’s coastline focusing on particle sizes of 330–5,000 μm and >20 μm respectively. Overall, urban coastal locations were reported to have higher microplastic abundances, and they mostly overlapped. Kwon et al. (2020) reported abundances of 1.35–4.74 particles/m3 in urban coasts and 1.12–2.79 particles/m3 in rural coasts. Meanwhile, Song et al. (2018) reported abundances of two to three orders of magnitude higher with average microplastic abundance for urban and rural coasts 1,051 and 560 particles/m3 respectively, demonstrating that particle size greatly affected the results. On top of that, sample collection method also affected the accuracy of quantifying microplastic abundance in water columns. Sampling via large mesh plankton nets could underestimate microplastic abundance by up to four orders of magnitude compared with using smaller mesh size nets or bulk water sampling that targets smaller size particles (Covernton et al., 2019; Schönlau et al., 2020). Similarly, in a survey conducted in Japanese waters, H. Xu et al. (2022) reported several orders of magnitude higher abundances for samples with smaller particle sizes. For instance, Tokyo Bay recorded 3,060–3,220 particles/m3 and 2.47–20.5 particles/m3 for the size ranges of 50–350 μm and 350–5,000 μm, respectively. Meanwhile, offshore of the Tokai region recorded 1,028–5,091 particles/m3 and 0.04–0.17 particles/m3, respectively.
Very high microplastic abundance was found in China, particularly along the Guangdong coast, with a range of 3,000–19,000 particles/m3. Relatively high abundance was also found in Beibu Gulf, with 399–5,531 particles/m3. It has been suggested that the source of plastics into the Guangdong coast could be inland waste carried by the Pearl River, which is a socioeconomic hub with rapid development and urbanization (Chan et al., 2021). China has also been known to import up to 45% of the global plastic waste for processing since the early 1990s (Brooks et al., 2018). However, this brought about a myriad of environmental problems causing China to introduced the Prohibition of Foreign Garbage Imports: the Reform Plan on Solid Waste Import Management, banning the import of solid waste, including plastic waste (Wen et al., 2021). However, previously imported plastic waste could still persist and is a possible source of microplastics in China’s waters. India also has a high population and rapid development in recent years, especially in coastal areas (Takar et al., 2020). India has been struggling with poor waste management, such as unregulated landfills and open littering (Sharholy et al., 2008), resulting in large inputs of waste into river systems and water bodies. Mumbai, India’s largest and most populous city, reported 165–547 particles/m3 of microplastics in its coastal waters. This is noticeably lower than the number observed in China, considering that Mumbai receives a large amount of plastic waste through industrial and sewage discharge, untreated wastewater, and other anthropogenic activities (Takar et al., 2020).
Southeast Asia is also plagued with plastic pollution, mostly due to poor waste management (Omeyer et al., 2022). The main source of plastic debris in the ocean is inland sources carried by rivers from runoff, direct littering, and leakage from open landfill sites (Jambeck et al., 2015; Marks et al., 2020). The abundances of microplastics was found to be relatively high, ranging from hundreds to thousands. The Lukut Estuary and Port Dickson Coast, situated on the highly urbanized west coast of the Malaysian Peninsula, have high microplastic abundances of 4,170 and 4,990 particles/m3, respectively (Shahbazi et al., 2010). Meanwhile, Terengganu Estuary, situated on the east coast reported only 421.8±110 particles/m3. Lamong Bay, Kenjeran Beach, and Wonorejo Beach in Indonesia all had similar concentrations to Terengganu Estuary with 380–610, 460–550, and 440–530 particles/m3 respectively. However, the quantified particle sizes were larger at >450 μm than those in Malaysia, which were >20 μm. The three surveys obtained from Thailand were conducted based on different particle size ranges, making it difficult to compare them. Nevertheless, compared to the other South East Asian samples, the concentrations in Thailand’s coast were lower, such as Gulf of Thailand and Bandon Bay with 21.29±36.21 and 0.28±0.07–0.63±0.13 particles/m3, respectively. Meanwhile, Rayong Coast had the highest microplastic abundance in this region, at 200–8,500 particles/m3. It is situated within the Gulf of Thailand, where inputs of plastic debris come from four main runoff sources. Moreover, various agricultural, industrial, tourism, and urbanization activities have contributed to plastic pollution in the Gulf (Vibhatabandhu and Srithongouthai, 2022).
In North America, higher microplastic abundances were reported in Canada compared to the United States of America (USA). Although the particle size ranges differed between the studies, they were not too far apart (62–5,000 and >100 μm). The abundances were mostly in the thousands for Canada, with the highest being in Queen Charlotte Sound with 7,630±1,410 particles/m3. Meanwhile, the abundances in the USA were mainly around 10 particles/m3 and under, such as Moss Landing Harbour in California with 4–11 particles/m3 and Gulf of Maine with 6.03±1.03 particles/m3. The identification process used in studies in Canada was visual identification via optical microscopy; thus, overestimation via misidentification is a possibility. In South America, extremely high microplastic abundance can be observed, with the highest being the Bahía Blanca Estuary in Argentina with 5,900–782,000 particles/m3, followed by Saigon Bay and Bocas del Toro in Panama with 107,000±25,000 particles/m3. These concentrations were significantly higher than any other global results. Both these locations are considered high-impact areas with strong human influence and are highly inclined to anthropogenic pollution (Fernández Severini et al., 2019; Fallon and Freeman, 2021). The Bahía Blanca Estuary in Argentina is also considered to be one of the largest petrochemical centers in South America, with factories that produce polyvinyl chloride and various types of polyethylene (PE), suggesting that these could be a major source in the estuary (Fernández Severini et al., 2019). High input sources coupled with underdeveloped waste management systems could be the main factor for high microplastic pollution levels (Fallon and Freeman, 2021). In both of these studies, the particle size ranges were also large, ranging from 0.22–5,000 and 10–5,000 μm, respectively. In the same study, a sample with a particle size range of 60–5,000 μm was also studied and recorded a much lower abundance of 42.6–113.6 particles/m3. Other surveys within the South American regions reported lower abundances despite the expectations of poor waste management, such as Esmeraldas in Ecuador with 6–340 particles/m3 and Guanabara Bay, Rio de Janeiro in Brazil with 1.40–21.3 particles/m3. The lower abundance found in Esmeraldas in Ecuador could be due to the low population and lack of industrial activities in the area (Capparelli et al., 2021). As for Guanabara Bay in Brazil, the lower abundance in comparison to the other locations could be due to the analyzed particle size range, which only looked at larger microplastic particles between 300–5,000 μm. The Galapagos Islands are remote areas with minimal urbanization and industrialization. However, microplastics were still detected in the surrounding waters, at approximately less than 0.5 particles/m3.
As for surveys conducted in Europe, most were sampled using neuston nets, except for Matosinhos Coast and Porto of Leixões in Portugal, which were collected as bulk water samples. The two samples collected via bulk water samples had much higher abundances than most of the other European surveys with 2,748±2,617 and 4,028±1,878 particles/m3, respectively. Only the Black Sea has a high microplastic abundance of 600–1,200 particles/m3, despite the neuston net collection method and a particle size range of 200–5,000 μm. This could be due to the high river discharge into a fairly small semienclosed sea (Aytan et al., 2016), which prevents the dilution and outflow of pollutants. In Swedish waters, which encompassed semienclosed seas and gulfs, low microplastic abundances of 0–0.46 particles/m3 over eight locations have been reported. Meanwhile, Frias et al. (2014), conducted a case study in Portugal using smaller 180 mesh size neuston nets. Microplastic abundance across the four locations was low with higher abundances detected in Lisboa and Costa Vicentina with 0.033±0.021 and 0.036±0.027 particles/m3 respectively. Costa Vicentina had the highest abundance in this study because of its close proximity to Sines, which is filled with industries and port facilities (Frias et al., 2014).
Unfortunately, there is still a scarcity of microplastic surveys in the African and Middle Eastern regions. Although there are overlapping sampling locations among the studies conducted in South Africa, mostly centered around Durban, they differ in sampling methods; Naidoo et al. (2015) used a neuston net with a mesh size of 300 μm, while the other two studies used bulk water sampling focusing on particle sizes of 80–5,000 and 63–5,000 μm. Again, bulk water samples had higher abundances than neuston net samples by approximately an order of magnitude. South African waters along the south eastern coastline had moderate to high microplastic abundances of 257.9±53.36–1,215±276.7 particles/m3. Durban Harbour was observed to have very high concentrations of microplastic with 1,200±133.2 and 319.8±542.4 particles/m3 for both bulk water and neuston net sampling respectively. Richard’s Bay Harbour also reported high concentrations with 413.3±77.53 particles/m3. Harbours, particularly Durban Harbour, have large populations, are prone to industrial waste, and are considered to be a source of microplastic pollution in the surrounding areas (Naidoo et al., 2015; Nel et al., 2017). Besides that, locations located away from the central city of Durban, such as the Mdloti and iLovu estuaries, still reported moderate abundances of 39.8±32.4 and 27.4±11.2 particles/m3. These abundances are higher than those in most other locations globally within the same particle size range, possibly due to mismanaged plastic waste (Jambeck et al., 2015). On top of that, wastewater infrastructure in South Africa is extremely limited, poorly operated, and unable to keep up with the growing demands and population (Mema, 2010). Similarly, Ghana is plagued by poor solid waste and wastewater management, with waste freely entering drainage channels and water bodies (Owusu Boadi and Kuitunen, 2002; Owusu, 2010). However, the Gulf of Guinea bordering the coast of Ghana were observed to have significantly lower microplastic abundances ranging from 1.14–2.79 particles/m3 over four locations. This could be due to the self-constructed Low-Tech Aquatic Debris Instrument trawl, which is suitable for small motorized boats, but may under sample microplastic quantities (Nuamah et al., 2022). Despite strong economic growth, increasing population, and the development of petrochemical industries, the microplastic abundance in Qatar’s Exclusive Economic Zone was surprisingly low, with 0–3 particles/m3, which is equivalent to some urban areas.
Most locations in the Oceania region consist of rural and remote areas; thus, can be considered as background levels of microplastic pollution. An accurate comparison between Australian studies would be difficult due to inconsistent sampling methodologies, but a general idea of the pollution status could be obtained. The coast of Kimberly is an extremely rural area with a very small and spread-out population; thus, it has a low abundance of 0.01–0.41 particles/m3. Although Ceduna and the South Australian Coastline are also considered to be secluded areas with minimum anthropogenic activities, their abundances are higher than that of Kimberly due to bulk water sampling and filtration through a smaller mesh size (50 μm) at 2.95±0.42 and 8.16±0.55 particles/m3 respectively. Semophore Beach is the only location in proximity to a moderately urbanized city with a considerable population size, that is, Adelaide. With moderate abundance of 16.31±2.42 particles/m3, it is almost double that of the South Australian Coastline. The Vava’u archipelago of Tonga is also a remote location. Using a tow net with a mesh size of 100 μm, the abundance in Tonga was slightly higher than that found in the Coast of Kimberly with 1.05±0.13 particles/m3. This is within the range of other remote islands and offshore samples (Markic et al., 2022).
MICROPLASTICS IN OPEN OCEANS/OPEN SEAS/OFFSHOREApart from coastal environments with close proximity to inland sources, plastic debris can also be found far offshore in the open seas and oceans. Plastic debris has been shown to converge at all five subtropical gyres, attributable to current and wind patterns (Moore et al., 2001; Eriksen et al., 2013, 2014; Law et al., 2014). Apart from these convergent zones, microplastics can be found even in most remote areas, such as close to the Arctic or Antarctica (Isobe et al., 2017; Kanhai et al., 2020; Suaria et al., 2020), demonstrating the omnipresence of plastic debris in the marine environment. Possibly due to long-range transport, long exposure to elements, and fragmentation, most plastic debris found in offshore ocean samples were small, less than 10 mm in size, with higher percentages of smaller microplastic particles (Rios et al., 2007; Barnes et al., 2009; Morét-Ferguson et al., 2010; Eriksen et al., 2014; Isobe et al., 2017).
The East Asian Sea, surrounded by urbanized and industrialized countries, is a hotspot for microplastic pollution. Hence, the microplastic abundances found in this region are higher than any other nonconvergent zone in the open sea/ocean, with 0.03–491 particles/m3. The source is suspected to be degraded mesoplastics, as they are slowly transported by northeastward ocean currents (Isobe et al., 2015). The Southern Indian Ocean, expressed lower abundances with 0–12 particles/m3. Plastic particles were suggested to have originated from continent-based sources and are expected to accumulate in the Indian Ocean Gyre (Li et al., 2022).
The Western Tropical Atlantic Ocean off the shores of Brazil has reasonably low microplastic abundances with 0.015–0.04 particles/m3. It is within the abundance range reported in the North East Atlantic Ocean (0–1.5 particles/m3), but significantly lower than the abundances that can be found in the North Atlantic Gyre (>20,000 particles/km2) (Law et al., 2010). Enders et al. (2015) conducted a transect survey from the European Coast to the North Atlantic Subtropical Gyre using bulk water sampling with a filtration mesh size of 10 μm. The range of abundance was 13–501 particles/m3, which is significantly higher than that of the Western Tropical Atlantic Ocean, possibly due to the difference in sampling technique and particle size range. To better understand the distribution of microplastic in the Atlantic Ocean, a more reliable comparison is necessary. Kanhai et al. (2017) also did a transect survey from the North Atlantic, starting in the Bay of Biscay in United Kingdom to the South Atlantic Ocean, ending at Cape Town in South Africa, using bulk water sampling with filtration mesh size of 250 μm. Most locations throughout the transect were reported to have abundances ranging from 0–2.5 particles/m3. Several outlining hotspots exceeded that range, notably, the Bay of Biscay and Western Coast of Portugal with 3.5 particles/m3, the west coast of Morocco with 6–6.5 particles/m3, and offshore of Namibia with 8.5 particles/m3. Because the sampling locations did not traverse through any locations in which microplastics have been known to accumulate, such as gyres, their abundances were lower than those reported from the North Atlantic Gyre (Kanhai et al., 2017). Meanwhile, the microplastic abundance of the North East Pacific Ocean was within the range of the North Atlantic Subtropical Gyre with 279±178 particles/m3.
Microplastics can even be found in pristine Arctic and Antarctica waters with early reports on microplastic particles in sea ice (Peeken et al., 2018; Kelly et al., 2020). Microplastic abundances in the Arctic reported by both Lusher et al. (2015) and Kanhai et al. (2020) were within the same order of magnitude as those found in the North Atlantic and Pacific Oceans with 0–11.5 and 0–18 particles/m3, respectively. The Eurasian Arctic, covering the Barent, Kara, Laptev, and East-Siberian Sea recorded very low abundances with 0.005, 0.003, 0.002, and 0.01 particles/m3 respectively. Considering the lack of anthropogenic sources in the Arctic, the presence of microplastics in this region are suspected to have been transported from more urbanized areas along the North Atlantic to the Arctic, possibly through the Barent Sea (Lusher et al., 2015; Yakushev et al., 2021). In addition, the great Siberian Rivers (Yakushev et al., 2021), as well as fishing and shipping activities, have also been suggested as direct sources of microplastics (Lusher et al., 2015). Sea ice, which could act as a sink for microplastic particles, has also been suggested to be able to transport microplastics within the region (Peeken et al., 2018; Kanhai et al., 2020). In Antarctica, the microplastic abundance reported by Isobe et al. (2017) was similar to those in the Arctic with 3.1×10−2 particles/m3. However, Suaria et al. (2020) revealed microplastic abundances of approximately one order of magnitude lower with 1.9×10−4–1.18×10−3 particles/m3. This could be due to the sampling conditions. During the sampling by Isobe et al. (2017), it was stormy with strong winds and high waves. Thus, this turbulence could have caused the higher abundance of microplastics. Akin to the Arctics, microplastic sources in Antarctica were most likely carried by ocean currents and winds from more polluted waters. Hence, the lower abundances could be due to the larger distance between the Antarctic region to populated areas and the Antarctic Circumpolar Currents that helps prevents matter from entering the Southern Ocean.(Suaria et al., 2020).
MICROPLASTIC ABUNDANCES IN COASTAL vs. OPEN WATERSOverall, microplastic abundance in coastal areas were substantially higher (up to three orders of magnitude) than that in open waters. Coastal areas reported abundances up to hundred thousand, while in the open oceans, the highest abundances were only in the hundreds. Generally, microplastic abundance is influenced by a combination of factors, including location, atmospheric parameters, and oceanographic conditions (Kanhai et al., 2017). Previous studies have also observed that microplastic abundance tends to decrease with increasing distance from the shore (Desforges et al., 2014; Steer et al., 2017; Lindeque et al., 2020).
Coastal areas tend to have higher microplastic abundance as they receive both terrestrial and marine sources and are close to human activities, such as urbanization, industrialization, shipping, recreational activities, and river outflows (Derraik, 2002; Ryan et al., 2009; Jambeck et al., 2015; Lusher et al., 2015; Takar et al., 2020; Chan et al., 2021). In addition, coastal regions often have less water circulation, making them reservoirs for marine plastic debris. Additionally, with little to no transport from coastlines to the open ocean, plastic debris entering the coast has a higher chance of remaining in the area (Onink et al., 2021). In contrast, open waters tend to have lower microplastic abundance because of their vastness and water circulation, leading to the dilution of plastic particles in surface waters (Desforges et al., 2014; Kanhai et al., 2017). However, plastic waste can still accumulate in certain areas, such as in the centers of oceanic gyres, where ocean currents converge and create stagnant zones known as “garbage patches” with six identified so far (Eriksen et al., 2013; Lebreton et al., 2018; Leal Filho et al., 2021). Overall, microplastic abundances can vary significantly depending on the specific location, global weather patterns, ocean currents and local conditions. However, it is important to note that microplastic pollution is a global problem affecting oceans, regardless of their location.
In terms of polymer type, the most common polymers found on the water surfaces, both in coastal and open ocean areas, were polyethylene (PE) and polypropylene (PP), followed by polystyrene (PS). This could be because PE and PP are the most commonly used polymer types, especially in single-used products. In addition, their low density allows them to float on the surface water where most sampling takes place (Enders et al., 2015; Erni-Cassola et al., 2019; Marrone et al., 2021). Besides, since an extended amount of time is required for microplastics to arrive at the center of open oceans or remote waters, most high-density polymers would have begun to sink into the water column (Enders et al., 2015), leaving mostly low-density ploymers on the surface of open oceans. Polyester (PES) and nylon fibers were also detected, with PES more commonly found in open and remote oceans than in coastal areas, as reported by Cui et al. (2022) and Lusher et al. (2015). PES is the most produced synthetic fiber (Enders et al., 2015) and can be easily lost from wastewater treatment plants globally; thus, it has become ubiquitous. However, coastal areas have a large variety of polymer types due to their close proximity to various sources, as most plastic debris tends to come from land-based sources (Enders et al., 2015; Jambeck et al., 2015).
The summarized data on microplastics concentrations in biota are shown in Table 3. The units were standardized for sufficient comparison, reporting mostly in particles per individual (N/ind.) and particles per gram (N/g). However, there are still differences in analytical, identification, and quantification methodologies. Selected studies on a variety of marine organisms, such as shellfish, zooplankton, various fish species, sponges, and others, were noted. They were then categorized into selective and nonselective feeders for further discussion. Selective feeders are usually predators with the ability to choose their diet. Nonselective feeders are usually filter feeders that simply filter the water surrounding them for nutrition.
| Location | Species | Conc. (N/ind.) | Conc. (N/g) | Feeding Type/Habitat | Identification Method | Polymer Types | Plastic Types | Tissue | Reference |
|---|---|---|---|---|---|---|---|---|---|
| Asia | |||||||||
| Bandon Bay, Thailand | Green mussels (P. viridis) | 0–13 | 0–1.85 | Filter | FTIR | Rayon, PP, PE, PET, nylon | Fibers (100%) | Soft tissue | Chinfak et al. (2021) |
| Clams (M. lyrata) | 0–3 | 0–1.06 | Filter | Rayon, PP, PE, PET | Fibers (95%) | ||||
| Coasts of Korea | Oyster/Mussel | 1.21±0.68 | 0.33±0.23 wet | Filter | μ-FTIR | PP, PE, PES | Fragment (69%), fiber (31%) | Soft tissue | Cho et al. (2021) |
| Manila clam | 2.19±1.20 | 0.43±0.32 | Filter | PP, PE, PES | Fragment (72%), fiber (23%) | ||||
| Tokyo Bay, Japan | Japanese Anchovy (Engraulis japonicus) | 2.3 | Selective | FTIR | PE, PP | Fragments (86.0%), beads (7.3%) | Digestive tract | Tanaka and Takada. (2016) | |
| Osaka Bay, Japan | Bivalve (Corbicula japonica) | 0–1 | Filter | FTIR–ATR | PE, PET | Fragments (100%) | Soft tissue | Nakao et al. (2020) | |
| Crabs (Chiromantes dehaani) | 0–2 | Selective | PET, PP | Fibers (100%) | |||||
| Mumbai Coast, India | White sardine (Escualosa thoracata) | 6.74±2.74 | Pelagic selective | Raman spectroscopy | PP, PE, PS, PA, nylon, polycarbonate, PMMA, PES, PET | Fiber, fragment, film, pellet/ beads | Gastrointestinal tracts | Gurjar et al. (2022) | |
| Belanger croaker (Johnius belangerii) | 6.42±2.65 | Demersal selective | |||||||
| Kadal shrimp (Metapenaeus dobsoni) | 6.60±2.98 | Demersal selective | |||||||
| Malabar sole (Cynoglossus macrostomus) | 5.62±2.27 | Demersal selective | |||||||
| Qingdao, Shandong, China | 7 types of shellfish | 1.2–4.1 | 0.8–4.4 | Filter | μ-FTIR | rayon, PET, CPE, PVC | Fiber (53.3%) | Digestive systems | Ding et al. (2020) |
| Xiamen, Fujian, China | 7 types of shellfish | 1.3–6.0 | 2.1–4.0 | Filter | μ-FTIR | rayon, PVDF, CPE, PVC, PET | Fiber (31.5%) | ||
| Qingdao, Shandong, China | Scallop (Chlamys farreri) | 0.5–2.9 | 0.4–3.4 | Filter | μ-FTIR | PVC, rayon, PES, PE, PET | Fiber (45%), fragment (23%), film (28%), granule (4%) | Digestive systems | Ding et al. (2021) |
| Mussel (Mytilus galloprovincialis) | 0.8–2.1 | 1.6–2.6 | Filter | ||||||
| Oyster (Crassostrea gigas) | 1.2–3.3 | 0.3–3.0 | Filter | ||||||
| Clam (Ruditapes philippinarum) | 1.2–3.2 | 4.5–20.1 | Filter | ||||||
| Coasts of Northern China | Sea Urchin (4 species) | 2.2±1.5–10.04±8.46 | 0.16±0.09–2.25±1.68 wet | Filter | μ-FTIR | PET, PE | Fiber, fragment, sheet, microbeads | Gut, coelomic fluid, gonads | Feng et al. (2020) |
| Yangtze Estuary, Shanghai, China | Collichthys lucidus | 6.2±2.4 | 17.2±9.7 | Demersal selective | μ-FTIR | Cellophane, PET, PES | Fiber, fragment, pellet | Gastrointestinal tracts | Jabeen et al. (2017) |
| Muraenesox cinereus | 2.4±0.6 | 0.4±0.2 | Demersal selective | ||||||
| Pampus cinereus | 3.0±0.8 | 0.5±0.2 | Benthopelagic selective | ||||||
| Harpodon nehereus | 3.8±2.0 | 1.9±0.1 | Benthopelagic selective | ||||||
| Coilia ectenes | 4.0±1.8 | 11.5±6.1 | Pelagic selective | ||||||
| Lateolabrax japonicus | 2.1±0.3 | 0.5±0.1 | Pelagic selective | ||||||
| Beibu Gulf, South China Sea, China | Upeneus sulphureus | 0.027 | Demersal selective | μ-FTIR | PES, nylon, PP, PE | Fiber (96%), fragment (2%), Film (2%) | Gastrointestinal tracts, gills | Koongolla et al. (2020) | |
| Caranx pectoralis | 0.125 | Pelagic selective | |||||||
| Gastrophysus spadiceus | 1 | Demersal selective | |||||||
| East Coast of China | Mussels (Mytilus spp.) | 1.5–7.6 | 0.9–4.6 | Filter | μ-FTIR | Cellophane, PET, PES | Fiber, fragment, sphere, flake | Soft tissue | Li et al. (2016) |
| Beibu Gulf, Guangxi province, China | Snail (Ellobium chinense) | 0.007±0.002–0.053±0.006 | Filter | μ-Raman spectrometer | PP | na | Whole organism | Li et al. (2020) | |
| Coast of China | Mussels (M. edulis, Perna viridis) | 0.77–8.22 | 1.52–5.36 | Filter | LUMOS (Bruker) microscope (ATR mode) | PES, rayon, PE, PVC, PP | Fiber (86%), Fragment (12%), Pellet (2%) | Soft tissue | Qu et al. (2018) |
| North America | |||||||||
| Central North Pacific | Longnosed lancetfish (Alepisaurus ferox) | 2.7±2.0 | Selective | Naked eye | na | na | Stomach | Choy and Drazen (2013) | |
| big-eye moonfish/opah (Lampris) | 2.3±1.6 | Selective | |||||||
| small-eye moonfish/opah (Lampris) | 5.8±3.9 | Selective | |||||||
| Northeast Pacific Ocean | zooplankton (N. cristatus) | 0.01–0.04 | Filter | Stereo discovery microscope | na | fiber (44%), fragment (56%) | Whole organism | Desforges et al. (2015) | |
| zooplankton (E. pacifica) | 0.03–0.07 | Filter | na | fiber (68%), fragment (32%) | |||||
| North Pacific Subtropical Gyre | Barnacles | 0–30 | Filter | Raman spectrometer | PE, PP, PS, PET | na | Stomach, intestinal tract | Goldstein and Goodwin (2013) | |
| McCormack’s Beach, Nova Scotia, Canada | Mussels (Mytilus edulis) | ~126 | Filter | Motic Dissecting Microscope | na | Fibers | Soft tissue | Mathalon and Hill (2014) | |
| Rainbow Haven Beach, Nova Scotia, Canada | Mussels (Mytilus edulis) | ~106 | Filter | ||||||
| Hudson–Raritan estuary, USA | Zooplankton (Acartia tonsa, Paracalanus crassirostris, Centropages typicus) | 0.30–0.82 | Filter and Selective | μ-FTIR, Raman spectroscopy | PE, PP | Fragment, Beads | Whole organism | Sipps et al. (2022) | |
| South America | |||||||||
| Bahía Blanca Estuary, Argentina | Whitemouth croaker (Micropogonias furnieri) | 12.1±6.2 | Selective | Stereomicroscope | na | Fibers (56.2%), pellets (15.6%), fragments (8%) | Gastrointestinal tracts | Arias et al. (2019) | |
| Saigon Bay, Bocas del Toro, Panama | Sponge (Callyspongia vaginalis) | 169±71 | Filter | Fluorescence microscope | na | Fibers, particles | Whole organism | Fallon and Freeman (2021) | |
| Sponge (Aplysina cauliformis) | 113±23 | Filter | |||||||
| Sponge (Niphates erecta) | 75±38 | Filter | |||||||
| Sponge (Ircinia campana) | 71±20 | Filter | |||||||
| Sponge (Amphimedon compressa) | 14±2 dry | Filter | |||||||
| Sponge (Mycale laevis) | 6±4 | Filter | |||||||
| Fernando de Noronha Archipelago, Rocos Atoll, Northeast Brazil | Vampire squid (Vampyroteuthis infernalis) | 9.58±8.25 | Fecal forager | Stereomicroscope | PE, PET, PVC, PA, PU, styrene–butadiene rubber | Fragment, Fibers, beads, foam, film | Whole organism | Ferreira et al. (2022) | |
| Midwater squid (Abralia veranyi) | 2.37±2.13 | Selective | Fragment, fibers, beads | ||||||
| Europe | |||||||||
| Baltic Sea | Herring (Clupea harengus) | 0–2 | Selective | Raman Microscope | PE, PP, PS, PC, ABS | na | Digestive tract, gills | Białowąs et al. (2022) | |
| Cod (Gadus morhua) | 0–2 | Selective | |||||||
| Southern North Sea/ Channel area | shrimp (Crangon crangon) | 1.23±0.99 | 0.68±0.55 | Filter | Stereo microscope | na | na | Whole organism | Devriese et al. (2015) |
| Groyne, Netherlands | Mussels | 0.26 | Filter | Stereo microscope | na | Fibers | Whole organism | De Witte et al. (2014) | |
| Quayside, Netherlands | Mussels | 0.51 | Filter | ||||||
| North Sea | Fish (5 species) | 0–4 | Selective | FTIR | PE, PP, PET, styrene acrylate | na | Esophagus, stomach, intestines | Foekema et al. (2013) | |
| Coast of Plymouth, United Kingdom (English Channel) | Fish (10 species) | 1–15 | Selective | FTIR | Rayon, PA, PES, PS, PE | Fibers (68.3%), Fragments (16.1%), beads (11.5%) | Gastrointestinal tract | Lusher et al. (2013) | |
| Adriatic Sea, Italy | Sardina pilchardus | 0.63±1.10 | 0.033±0.052 | Pelagic selective | μ-Raman spectrometer | PE, PP, PET, PS | na | Gastrointestinal tract | Mistri et al. (2022) |
| Engraulis encrasicolus | 0.47±0.86 | 0.041±0.077 | Pelagic selective | ||||||
| Merluccius merluccius | 1.37±1.56 | 0.011±0.012 | Demersal selective | ||||||
| Pegusa impar | 2.47±2.99 | 0.016±0.018 | Demersal selective | ||||||
| Mullus surmuletus | 1.90±1.81 | 0.048±0.044 | Demersal selective | ||||||
| Gobius paganellus | 0.93±1.23 | 0.037±0.065 | Demersal selective | ||||||
| North Sea Coast, France, Belgium, Netherlands | Blue Mussel (Mytilus edulis) | 0.2±0.3 | Filter | Raman Microscope | LDPE, HDPE, PS | na | Whole organism | Van Cauwenberghe et al. (2015) | |
| Lugworm (Arenicola marina) | 1.2±2.8 | Filter | |||||||
| Africa | |||||||||
| Dabaso, Kenya | Jellyfish (Crambionella orsini) | 0.05 | Passive | Hot needle test | na | Fibers | Whole organism | Awuor et al. (2021) | |
| Mikindani, Kenya | Jellyfish (Crambionella orsini) | 0.03±0.003 | Passive | ||||||
| Makupa, Kenya | Jellyfish (Crambionella orsini) | 0.03±0.01 | Passive | ||||||
| South African Coastline | European anchovy (Engraulis encrasicolus) | 1.13 | Pelagic selective | FTIR–ATR | PEPD, PE, PA, PES, PP | Fibers (80%), fragment (20%) | Gastrointestinal tract | Bakir et al. (2020) | |
| West Coast round herring (Etrumeus whiteheadi) | 1.38 | Pelagic selective | |||||||
| South African sardine (Sardinops sagax) | 1.58 | Pelagic Selective | |||||||
| Jamestown fish landing, Gulf of Guinea, Ghana | Pseudupeneus prayensis | 1.26±1.67 | Selective | μ-FTIR | PE | Fibers, Pellets | Gastrointestinal tract | Pappoe et al. (2022) | |
| Pagellus bellottii | 0.94±1.18 | Selective | PE, PVA, PA | Fibers | |||||
| Sardinella maderensis | 1.49±1.48 | Selective | PVA, PE, PA | Fibers | |||||
| Decapterus rhonchus | 0.96±1.055 | Selective | PE, PVA | Fibers, Pellets | |||||
| Eastern Central Atlantic Ocean, Off the Coast of Ghana | Sardinella maderensis | 40.0±3.8 | Selective | Stereo microscope | na | Pellets (31%), microbeads (29%), burned film (22%), fragment (9%), thread (2%), microfibers (2%) | Gastrointestinal tract | Adika et al. (2020) | |
| Dentex angolensis | 32.0±2.7 | Selective | na | ||||||
| Sardinella aurita | 25.7±1.6 | Selective | na | ||||||
| Off shore of Agulhas Bank, South Africa | Cape horse mackerel (Trachurus capensis) | 3.9±1.0 | 0.21±0.06 | Selective | Hot needle test | na | Mainly fibers | Digestive tract | Sparks and Immelman (2020) |
| Shallow-water Cape hake (Merluccius paradoxus) | 4.2±0.6 | 0.25±0.03 | Selective | ||||||
| Deep water Cape hake (Merluccius paradoxus) | 3.8±0.7 | 0.15±0.04 | Selective | ||||||
| Round herring (Etrumeus whiteheadi) | 3.3±0.5 | 0.05±0.01 | Selective | ||||||
| Chub mackerel (Scomber japonicus) | 4.6±0.8 | 0.28±0.05 | Selective | ||||||
| Cape Gunard (Chelidonichthys capensis) | 3.4±0.4 | 0.25±0.04 | Selective | ||||||
| Carpenter seabream (Argyrozona argyrozona) | 2.8±0.7 | 0.03±0.01 | Selective | ||||||
| Gulf of Guinea, Ghana | Mangrove oysters (Crassostrea tulipa) | 1.4±1.3–3.4±1.0 | 0.29±0.14–1.64±0.63 | Filter | FTIR | PE, PP, PS | Fiber (69%), Fragment (27%), Film (4%) | Soft tissue | Addo et al. (2022) |
| Tanzanian Coast | Cockle (Anadara antiquata) | ≤1 | Filter | FTIR–ATR | PP, PE | Fibers (75%), Fragment (25%) | Soft tissue | Mayoma et al. (2020) | |
| Mtoni Kijichi Creek, Tanzania | Cockle (Anadara antiquata) | 2.1±1.8 | Filter | ||||||
| Cape Town, South Africa | 3 species of mussels | 4.27±0.5 | 2.33±0.2 | Filter | Stereo microscope, hot needle test | na | Filaments (67%), Fragment (21%), Spheres (12%) | Soft tissue | Sparks (2020) |
| Oceania | |||||||||
| Brown’s Beach, Kangaroo Island, Australia | Wild mussels (Mytilus) | 2.83±0.6 | Filter | FTIR | PA, PP | Fiber (65.4%), fragments (34.6%) | Soft tissue | Klein et al. (2022) | |
| Victor Harbour, Australia | Wild mussels (Mytilus) | 10.17±4.1 | Filter | ||||||
| South Australia Coastline | Wild mussels (Mytilus) | 6.67±1.25 | Filter | ||||||
| Brisbane, Australia | 5 commercial species | 1.58±0.23 | Selective | μ-FTIR | PE, PEVA, PP, PS | Fiber (82.4%), fragment (10.2%), film (7.4%) | Gastrointestinal tract | Wootton et al. (2021) | |
| Suva, Fiji | 5 commercial species | 0.86±0.14 | Selective | PE, PEVA, PP, PS, PU | Film (50%), fragment (25.5%), fiber (24.5%) | ||||
| Seaports in New South Wales, Australia | Oysters (Saccostrea glomerata) | 0.15–0.83 | Filter | FTIR | Nylon, PET | Fibers (43% –80%), Fragment (14%–43%) | Soft tissue | Jahan et al. (2019) | |
| Coast of New Zealand | Green-lipped mussel (Perna canaliculus) | 0–1.5 | 0–0.48 | Filter | FTIR | PE | Fragments | Soft tissue | Webb et al. (2019) |
| Middle East | |||||||||
| Turkish territorial waters, Mediterranean Sea, Turkey | Argyrosomus regius | 1.84 | Benthopelagic selective | FTIR | PS isoprene, PE, PP | Fibers (70%), hard plastic (20.8%), nylon (2.7%), others (6.3%) | Gastrointestinal tract | Güven et al. (2017) | |
| Caranx crysos | 5 | Reef selective | |||||||
| Dentex dentex | 0 | Benthopelagic selective | |||||||
| Dentex gibbosus | 0.29 | Benthopelagic selective | |||||||
| Diplodus annularis | 1.96 | Benthopelagic selective | |||||||
| Lagocephalus spadiceus | 0 | Demersal selective | |||||||
| Lithognathus mormyrus | 0.63 | Demersal selective | |||||||
| Liza aurata | 3.26 | Pelagic selective | |||||||
| Mullus barbatus | 1.39 | Demersal selective | |||||||
| Mullus surmuletus | 1.18 | Demersal selective | |||||||
| Nemipterus randalli | 1.31 | Demersal selective | |||||||
| Pagellus acarne | 1.63 | Benthopelagic selective | |||||||
| Pagellus erythrinus | 0.63 | Benthopelagic selective | |||||||
| Pagrus pagrus | 1.44 | Benthopelagic selective | |||||||
| Pelates quadrilineatus | 1.48 | Reef selective | |||||||
| Pomadasys incisus | 0.79 | Demersal selective | |||||||
| Sardina pilchardus | 2.14 | Pelagic selective | |||||||
| Saurida undosquamis | 1.22 | Reef selective | |||||||
| Sciaena umbra | 3 | Demersal selective | |||||||
| Scomber japonicus | 6.71 | Pelagic selective | |||||||
| Serranus cabrilla | 1.5 | Demersal selective | |||||||
| Siganus luridus | 3.13 | Reef selective | |||||||
| Sparus aurata | 0.87 | Demersal selective | |||||||
| Trachurus mediterraneus | 1.77 | Pelagic selective | |||||||
| Trigla lucerna | 0.75 | Demersal selective | |||||||
| Umbrina cirrosa | 0 | Demersal selective | |||||||
| Upeneus moluccensis | 0.78 | Reef selective | |||||||
| Upeneus pori | 0.69 | Demersal selective | |||||||
| The Persian Gulf, Iran | 3 bivalve species | 3.9–6.9 | 0.2–2.2 | Filter | Stereomicroscope | na | Fiber, film, fragment, pellet | Soft tissues | Naji et al. (2018) |
| Predatory snail (Thais mutabilis) | 17.7±0.3 | Selective | |||||||
| Northern shores of the United Arab Emirates | Oyster | 0.10±0.09 | Filter | FTIR–ATR | PE, nylon | Fiber (93%), Fragment (7%) | Soft tissues | Hammadi et al. (2022) |
FTIR: Fourier transform infrared, ATR: Attenuated Total Reflection
PE: polyethylene, PP: polypropylene, PES: polyester, PET: polyethylene terephthalate, PMMA: polymethylmethacrulate, PU: polyurethane, PS: polystyrene, PA: polyamide, PVC: polyvinyl chloride, HDPE: high-density polyethylene, LDPE: low-density polyethylene, PEVA: polyethylene–vinyl acetate, PVDF: polyvinylidene fluoride, CPE: chlorinated polyethylene, PVA: polyvinyl acetate, PEPD: polyethylene polypropylene diene; na=not available.
Microplastics have been detected in various filter feeder species globally, with shellfish being the most commonly reported species. In the analysis of filter feeders, some have studied only the digestive tracts, but in most cases, the soft tissues were analyzed. Apart from shellfish, other filter feeders will also be explored such as shrimps, lugworms, sponges, barnacles, snails, zooplankton, and sea urchins.
Significant microplastic concentrations have been reported in Asia, and many case studies have been conducted across China. The Coast of China has high microplastic abundance, which is reflected in the biota. In Qingdao and Xiamen, microplastics were found in 70%–100% of all seven shellfish species analyzed with 1.2–4.1 N/ind. and 1.3–6.0 N/ind. (Ding et al., 2020). The microplastic size range reported for these shellfish was 10–4,377 μm, with microplastics smaller than 500 μm being the dominant fraction. Furthermore, Ding et al. (2021) specifically analyzed and compared four shellfish species, namely, scallops (Chlamys farreri), mussels (Mytilus galloprovincialis), oysters (Crassostrea gigas), and clams (Ruditapes philippinarum). Similar to a previous study, the size range of microplastics found in the four shellfish species was 7–5,000 μm, with a majority of abundances in the <500 μm range. The concentrations differed slightly between the species. Oysters and clams with 1.2–3.3 and 1.2–3.2 N/ind., respectively, were shown to have higher concentrations than scallops and mussels with 0.5–2.9 and 0.8–2.1 N/ind. respectively. In the East Coast of China, the microplastic concentrations in mussels had a higher range than those in Qingdao and Xiamen with 1.5–7.6 N/ind., but was comparable with the range reported by X. Qu et al. (2018), who sampled mussels throughout the Coast of China at 0.77–8.22 N/ind. These studies also have similar microplastic size ranges, with Li et al. (2016) detecting a microplastic size range of 5–5,000 μm with those <1,000 μm dominating. In addition, X. Qu et al. (2018) reported that most microplastics detected in mussels were within the size range of 250–1,000 μm, which coincided with the distribution found in the water column. The coast of Korea has also reported similar concentrations of 1.21±0.68 N/ind. for oysters/ mussels and 2.19±1.20 N/ind. for manila clams. The microplastic size on the longest axis ranged from 46–13,298 μm, with a size range of 100–200 μm being the most common. Bivalves in Osaka Bay had the lowest concentration range within Asia, with microplastics detected in only 10% of all samples, with a range of 0–1 N/ind. The detected microplastics had a mean size of 760±420 μm. Green mussels in Bandon Bay, Thailand reported a slightly higher range of 0–13 N/ind., with sizes ranging from 103–4,301 μm, although the microplastic abundance in the water column was relatively low (Chinfak et al., 2021). Meanwhile, microplastics were detected in four species of sea urchins in the Northern China Coast, with concentrations ranging 2.2±1.5–10.04±8.46 N/ind., which was slightly higher than that of shellfish. Sea urchins have been shown to contain microplastics in the size range of 27–4,742 μm, which is similar to the size range reported for shellfish.
Besides shellfish, certain zooplankton species are filter feeders and have been shown to ingest microplastics. This is particularly alarming because zooplanktons are an important foundation of the food chain. In North America, microplastics were reported in two zooplankton species sampled from the Northeast Pacific Ocean at 0.01–0.07 N/ind., which is considerably lower than those found in the Hudson-Raritan estuary 0.30–0.82 N/ind. reflecting the pollution status of the respective locations. Zooplankton have been reported to ingest microplastics with a smaller size range of <1,000 μm compared to shellfish, and depending on the species, most detected microplastics were below 100 μm (Desforges et al., 2015; Sipps et al., 2022). Meanwhile, barnacles from the North Pacific Subtropical Gyre exhibited higher microplastic concentrations ranging 0–30 N/ind., which reflects the higher microplastic abundance in the convergence zones. The microplastics detected in the barnacles in this study were larger than those in the other shellfish, with the smallest particle ingested having a maximum diameter of 609 μm and the largest particle diameter was 6,770 μm. Along the Atlantic Coast of Canada, mussels were sampled from two beaches in a large urban estuary. Microplastics, particularly fibers, were detected in all samples from both beaches, with average concentrations of 126 and 106 N/ind. This can be attributed to the densely urbanized and populated Halifax Harbour (Mathalon and Hill, 2014). In this study, only the median diameter of the detected microplastics was reported with a median range of 11–220 μm, which coincides with the most commonly found microplastic size range in mussels reported in other studies.
Sponges, have also been considered as bioindicators for monitoring microplastic pollution in marine environments. This is due to their widespread existence, ability to filter a large volume of seawater, and to intake and retain foreign particles, including microplastics (Fallon and Freeman, 2021; Girard et al., 2021). Six species of sponges were investigated in Saigon Bay, Panama, Aplysina cauliformis, Amphimedon compressa, Callyspongia vaginalis, Ircinia campana, Mycale laevis, and Niphates erecta. Significant pairwise differences were observed among the six species with Callyspongia vaginalis and Aplysina cauliformis, having the highest concentrations of 169±71 and 113±23 N/g-dry weight, respectively. This was followed by Niphates erecta (75±38 N/g-dry weight), Ircinia campana (71±20 N/g-dry weight), Amphimedon compressa (14±2 N/g-dry weight), and Mycale laevis (6±4 N/g-dry weight). Most microplastics found in the sponges were extremely small (approximately 10–20 μm), explaining the low concentrations in sponges despite the high microplastic abundance in the water column (Fallon and Freeman, 2021).
In Europe, the most popular filter feeder species in literature were mussels, but shrimp and lugworms have also been studied in the North Sea. Fibers were detected in 63% of all shrimp collected from the Southern North Sea with 1.23±0.99 N/ind. or 0.68±0.55 N/g-wet weight, respectively. The fibers ranged from 200–1,000 μm in size. However, a small percentage of shrimps were also found to contain plastic granules and films, which were found in a smaller size range compared to fibers of 20–100 μm. In terms of weight, shrimp showed slightly higher microplastic concentrations compared to blue mussels collected from the North Sea Coast with 0.2±0.3 N/g-dry weight. Similar concentrations were also observed in mussels from two coastal locations in the Netherlands, Groyne and Quayside with 0.26 and 0.51 N/g-wet weight respectively. In this study, only synthetic fibers were detected, ranging from 200–1,500 μm, with 1,000–1,500 μm being the most common size class. Lugworms live in sandy sediments and adult lugworms have been shown to remain in a single location throughout their lifetime (Van Cauwenberghe et al., 2015). As nonselective feeders, lugworms inevitably ingest microplastics within sediment and thus can be used to estimate microplastic abundance in the environment (Besseling et al., 2013; Van Cauwenberghe et al., 2015). Lugworms were shown to have the highest microplastic concentrations among the organisms within the North Sea Coast with 1.2±2.8 N/g-wet weight. Since both mussels and lugworms could reflect the microplastic abundance of their surroundings, the difference in microplastic abundance between the organisms could be due to differences in plastic retention, particle selection, and feeding habits. Lugworms are fully nonselective feeders that feed on organic matter by ingesting sediments (Retraubun et al., 1996). Mussels have the ability to reject other particles via pseudofaeces that are not appropriate in size or shape while ingesting water (Defossez and Hawkins, 1997; Ward and Shumway, 2004). Lugworms were also found to ingest larger particles (30–90 μm) than mussels (10–30 μm), based on the laboratory exposure experiments (Van Cauwenberghe et al., 2015).
Shellfish from the African continent were roughly within the same order as those found in Asia and Europe. Concentrations in mussels (three species) from Cape Town, South Africa were reported to be 4.27±0.5 N/ind. or 2.33±0.2 N/g-wet weight, respectively. The mussels were collected across 27 sites, and significant differences were found among the locations; however, no significant differences were found among the three mussel species, Aulyacoma ater (ribbed mussel), Choromytilus meridionalis (black mussel), and Mytilus galloprovincialis (blue mussel) (Sparks, 2020). Most microplastics particles detected were less than 100 μm in size, with only 22% of microplastics detected in the size range of 1,000–2,000 μm. Mangrove oysters in the Gulf of Guinea, Ghana had slightly lower concentrations than the Cape Town mussels with an average of 1.4±1.3–3.4±1.0 N/ind. or 0.29±0.14–1.64±0.63 N/g-wet weight. They also exhibited spatial variation according to the intensity of anthropogenic and commercial activities (Addo et al., 2022). The length of the microplastics detected in the oysters ranged from 33–4,870 μm, with more than 90% within the <1,000 μm size range. Along the Tanzanian Coast, microplastics were found in 48% of all cockles analyzed. Spatial differences were also observed, with the highest concentration found in Mtoni Kijichi Creek (2.1±1.8 N/ind.). Other locations reported lower concentrations of ≤1 N/ind. Although Mtoni Kijichi Creek was indeed the more industrialized location along the Tanzanian Coast, there was little correlation between environmental conditions and cockle concentrations in the other locations (Mayoma et al., 2020). The microplastic size range was not reported in this study.
Along the South Australian coastline, microplastic concentrations in mussels exhibited no significant differences among the study sites, with an average of 6.67±1.25 N/ind. (Klein et al., 2022). Only slight variations were observed, with the highest concentration being Victor Harbour with 10.17±4.1 N/ind. and the lowest, Brown’s Beach, Kangaroo Island with 2.83±0.6 N/ind. South Australian coastline mussels had higher microplastic concentrations than mussels from Africa, Europe, and most Asia, apart from Thailand. The specific size range for the detected microplastics was not clearly reported in the study, except for the 50 μm detection limit, and particles smaller than 90 μm were trapped in mussel tissues. Green-lipped mussels from the Coast of New Zealand also had lower concentrations of 0–1.5 N/ind. or 0–0.48 N/g-wet weight, respectively. Concentrations were low overall in New Zealand, with little to no difference between urban and rural sites (Webb et al., 2019). The size range of microplastics detected was 50–700 μm, with 52% ranging between 100 and 200 μm. Oysters from seaports in New South Wales have been reported to have microplastic concentrations ranging 0.15–0.83 N/g-wet weight. This is within the global range of microplastic concentrations in oysters, as listed in Table 3. In terms of weight, it was also within the range of mussels from the Coast of New Zealand. Apart from that, statistical differences in oyster microplastic concentrations have been observed among New South Wales seaport locations (Jahan et al., 2019). In terms of microplastic size, the reported range was 100–4,900 μm, with most found in the range of 100–1,000 μm. Oysters were also investigated on the northern shore of the United Arab Emirates (UAE). The reported size range of microplastics was comparable to that reported for oysters in other studies (146–3,172 μm). Although similar to global levels, concentrations of microplastics detected were on the lower side with 0.10±0.09 N/g-wet weight and only 51% of all oysters were found to contain microplastics (Hammadi et al., 2022). Meanwhile, in the Persian Gulf, bivalves (three species) were shown to have microplastic concentrations equivalent to the UAE with 0.2–2.2 N/g or 3.9–6.9 N/ind. Microplastics were detected in all three size groups 10–25, 25–250, and 250–5,000 μm, and the most detected size group was 10–25 μm. All five molluscan species examined in this study were shown to ingest and accumulate microplastics with higher concentrations of predatory species, suggesting that microplastics can be transferred through the food chain (Naji et al., 2018).
Overall, the microplastic size ranges found in filter feeders were similar, mostly ranging from 10–5,000 μm, with only two studies reporting results with larger microplastics. The larger size range could be due to measuring the longest axis, especially when it comes to measuring fibers, as in the case of Cho et al. (2021). Fibers were also reported to generally have a larger size range than microplastic particles, as reported by Devriese et al. (2015), suggesting that fibers can disguise their lengths by compacting and entangling themselves (Watts et al., 2015; Scott et al., 2019; Klein et al., 2022). The most frequently detected microplastic size range in the filter feeders was <500 μm. Studies that compared the microplastic size range between those found in the water column and bivalves have reported that microplastics found in bivalves were smaller than those found in the surrounding water (Klein et al., 2022; Hammadi et al., 2022).
MICROPLASTIC IN FISH, FORAGERS, PREDATORS, AND PASSIVE FEEDERS: SELECTIVE FEEDERSMicroplastic particles have also been detected in various marine species globally, from pelagic to deep-ocean biota (Lusher et al., 2013; Adika et al., 2020; Pereira et al., 2020). Predatory and selective feeder organisms have been shown to reject microplastics, suggesting that microplastic ingestion was mostly accidental through the act of breathing or passive feeding via drawing water (Li et al., 2021; J. Xu et al., 2022). Apart from that, there is evidence of possible bioaccumulation and biomagnification of microplastic through trophic transfer (Nelms et al., 2018; Setälä et al., 2018; Zhang et al., 2019, 2022; Hasegawa and Nakaoka, 2021). Since the selection of species in this subchapter is considerably larger than the previous one, spatial comparison might not be accurate due to the differences in feeding habits, capacity to retain microplastics, and depuration rate of each species (Au et al., 2017). Microplastic analysis of higher trophic level species were mostly conducted only on the gastrointestinal tract, digestive system, and stomach.
Microplastics within the size range of 68–6,890 μm were found in 77% of the Japanese anchovies in Tokyo Bay, with an average of 2.3 N/ind. and 0–15 N/ind., respectively. This reflects the pollution status of the area (Tanaka and Takada, 2016). Jabeen et al. (2017) investigated 21 species of sea fish in Yangtze Estuary, Shanghai, China for microplastics. A microplastic size range of 40–24,800 μm was detected, with plastic particles smaller than 5,000 μm being the most common. Only six were listed in Table 3, consisting of the highest and lowest concentrations of each habitat category (pelagic, benthopelagic, and demersal). Microplastics were 100% detected in all sea fish species, and overall, the average microplastic concentrations was 1.1–7.2 N/ind. or 0.1–3.9 N/g-wet weight, respectively. Demersal species were shown to have significantly higher concentrations compared to pelagic species with the range of 2.4±0.6–6.2±2.4 N/ind. and 2.1±0.3–4.0±1.8 N/ind., respectively (Jabeen et al., 2017). Similarly, Koongolla et al. (2020) studied 24 fish species in the Beibu Gulf, China; only half of the species were shown to contain microplastics, with Gastrophysus spadiceus and Siganus canaliculatus, the species with the highest concentrations. Table 3 lists the highest and lowest concentrations of both pelagic and demersal species, with concentrations ranging 0.125–0.222 and 0.027–1.000 N/ind. respectively. Concentrations observed in Beibu Gulf were lower than those in fish samples from the Yangtze Estuary. The microplastic size range reported here is also smaller (20–3,000 μm) with most being less than 1,000 μm.
Large fish species in the Central North Pacific have an ingestion frequency of 19% with the three highest ingestion frequencies species being the longnose lancetfish (Alepisaurus ferox) and two species of moonfish (Lampris spp.) (Choy and Drazen, 2013). Microplastic concentrations in these species were similar to those found in fish from Asia, with concentrations ranging 2.3±1.6–5.8±3.9 N/ind. Considering that these are large fish species, the microplastic size range is also much larger than that reported in Asia, with an average debris length of 56,800±75,500 μm. In South America, juvenile white-mouth croaker (Micropogonias furnieri), which are an important commercial species, were sampled from the Bahía Blanca Estuary in Argentina. Microplastic particles ranging in size of 200–5,000 μm were discovered in all individuals, with a majority being under 1,000 μm, and the average concentration was 12.1±6.2 N/ind. This result is comparable to those found mostly in demersal species from more contaminated areas in Asia, such as Shanghai and Mumbai. Ferreira et al. (2020) also reported similar concentrations of microplastics within the size range of 40±60 μm to 910±920 μm, in Vampire squid (Vampyroteuthis infernalis) from Fernando de Noronha Archipelago in Brazil with 9.58±8.25 N/ind. where the feeding habitat is in the deeper waters. However, Midwater squid (Abralia veranyi), which is also known to be a deep-sea cephalopod, was reported to have lower concentrations of microplastic with 2.37±2.13 N/ind., within the size range of 30±40 μm to 640±1,230 μm, probably because of its migration to the epipelagic zone at night to forage (Ferreira et al., 2020).
In the pristine Baltic Sea, microplastics were found in 12.7% of herrings, and 14.8% of cods with both fishes exhibited the same microplastic concentration range of 0–2 N/ind. (Białowąs et al., 2022). Gills have been demonstrated to be a significant part of microplastic transfer to cods, but this was not observed in herrings expressing differences between fish species in terms of water filtration rates and lifestyle (Białowąs et al., 2022). The specific size range of microplastics was not reported in this study, except that the majority of plastics were found to be less than 5,000 μm and only two items were larger than 25,000 μm. Fish species in the North Sea also expressed microplastic concentration ranges similar to those in the Baltic Sea, with a range of 0–4 N/ind., although most fish were reported to have either none or only one piece of microplastics. Of the seven species investigated, herring, gray gurnard, whiting, horse mackerel, haddock, Atlantic mackerel, and cod, microplastics were detected in only 2.6% of all surveyed fishes, and no microplastics were found in gray gurnards and mackerels (Foekema et al., 2013). All microplastics found in the fish were less than 5,000 μm, with a median size of 800 μm. Microplastics <1,000 μm were found in fish of all sizes. Meanwhile, microplastics >1,500 μm were only found in fish larger than 20 cm in length. Lusher et al. (2013) surveyed five pelagic and five demersal fishes in the English Channel (Coast of Plymouth) and found microplastic particles in all sampled fishes with a range of 1–15 N/ind. and an average of 1.90±0.10 N/ind. Microplastic sizes ranged 130–14,300 μm, with 1,000–2,000 μm being the most common size class. Overall, microplastic abundances in fish from the English Channel were higher than those from the Baltic Sea and North Sea. Similar concentrations were observed in six species of commercial fish from the Adriatic Sea consisting of two pelagic and four demersal species. The size range of microplastics reported in this study was smaller than that reported in other studies on fish, (54–765 μm). Out of the 180 individuals examined, 46.8% had microplastics in them with Adriatic soles (Pegusa impar) having the highest concentration with 2.47±2.99 N/ind., while anchovies (Engraulis encrasicolus) having the lowest with 0.47±0.86 N/ind.
Awuor et al. (2021) established that jellyfish ingest microplastic with abundances of 0.03±0.01–0.05 N/g-wet weight. The ingested microplastics were observed to be mainly fibers with sizes ranging from 300–3,000 μm. Although jellyfish were sampled from three locations along the Kenyan coast, there were no significant differences in concentrations among the locations. Microplastic concentrations in jellyfish were approximately an order of magnitude lower than those in most fish in African waters (Table 3). Fishes in Africa exhibited high concentrations, with the Eastern Central Atlantic Ocean, Coast of Ghana having the highest global concentrations. The Eastern Coast of Ghana is deemed to be highly polluted by marine litter, explaining the high concentrations found in all three fish species as follows: Sardinella maderensis, Dentex angolensis, and Sardinella aurita with 40.0±3.8, 32.0±2.7 and 25.7±1.6 N/ind., respectively (Adika et al., 2020). Specific size ranges or categories of ingested microplastics were not reported in this study. Microplastics were detected in all the seven species surveyed on the shore of the Agulhas Bank in South Africa with a detection rate of 86.67%. Most of the detected microplastics were within the size range of 500–1,000 μm, followed by 1,000–2,000 μm. The size range categories in this study were from <125 μm to >2,000 μm, and all size categories were found in the samples. Relatively high abundances were also reported ranging from 2.8±0.7–4.6±0.8 M/ind. or 0.03±0.01–0.28±0.05 N/g-wet weight (Sparks and Immelman, 2020). However, lower microplastic concentrations have been reported in fish from the South African Coastline and Gulf of Guinea, Ghana by Bakir et al. (2020) and Pappoe et al. (2022), respectively. The concentrations in these two studies were similar ranging 1.13–1.58 N/ind. for the South African Coastline and 0.94±1.18–1.49±1.48 N/ind. for the Gulf of Guinea, Ghana. Bakir et al. (2020) did not report on microplastic size ranges, meanwhile Pappoe et al. (2022) investigate three microplastic size range categories, which are: 100–500 μm, 500–1,000 μm, and 1,000–5,000 μm, with 500–1,000 μm being the most dominant size range.
In both Brisbane, Australia, and Suva, Fiji, Wootton et al. (2021) reported on five important commercial species. Microplastics were less frequently detected in Fiji’s fish (35.3%), whereas in Australia, 61.6% of fish had plastic particles. Average microplastic abundances also imitated the detection rate, as fishes from Australia were shown to have almost double the concentrations of that from Fiji at 1.58±0.23 and 0.86±0.14 N/ind., respectively. The differences in concentrations implied that microplastic pollution is higher in Australian marine waters, as Brisbane has a larger and denser population than Fiji (Wootton et al., 2021). On top of that, it has also been insinuated that Australia produces 11 times more plastic waste than Fiji (Jambeck et al., 2015). Most of the microplastics found were larger than 1,000 μm in size, but microplastic particles as small as 38 μm were also detected. Microplastics were also found in all 28 species of fishes collected from the Mediterranean Sea in Turkish territorial waters, and 58% of 1,337 individuals were confirmed to contain microplastics in either their stomach or intestinal tract. The particle size of microplastics found in the fishes were 9–12,074 μm, with a mean of 656±803 μm. Although the size range in this study was large, most particles were below 5,000 μm, with only five out of the total 1,822 particles being larger than 5,000 μm. The average concentration of microplastics in all samples was 1.36 N/ind., which is almost equivalent to the concentrations reported in Australia, Europe, and some parts of Asia and Africa. In the Persian Gulf, Naji et al. (2018) suggested the trophic transfer of microplastics through the food web by revealing that the predatory snail (Thais mutabilis) had significantly higher concentrations of microplastics with 17.7±0.3 N/ind. compared with other mollusks collected from the same location. The snails were observed to ingest microplastics ranging from 10–5,000 μm, with most falling within the size category of 10–25 μm.
Overall, the size range of microplastics found in selective feeders was larger than that of filter feeders, with particles as small as approximately 10 μm to mesoplastic sizes of >60,000 μm. This is because there is a large variety of selective feeder species that vary in a large variation of size, which could suggest the capability of ingesting larger plastic particles. However, it has been observed that most ingested microplastics were less than 5,000 μm in size and that the size of the organism does not affect the ingestion rate or the size of ingested microplastics (Foekema et al., 2013; Güven et al., 2017; Sparks and Immelman, 2020). On top of that, smaller sized microplastics may have more possibilities of being accidentally ingested by selective feeders (Koongolla et al., 2020).
BIOTAS AS SUITABLE BIOINDICATORS FOR MICROPLASTIC POLLUTION IN THE MARINE ENVIRONMENTBased on the overall global data on microplastic concentrations in biota, there are no clear trends between filter feeders and other species in the food chain. Microplastic concentrations in biota were shown to correspond more to geographical locations and the pollution status of the surrounding areas. Similarly, in terms of polymer and particle types, there were no clear trends among the species of biota globally. The polymer and particle types were more influenced by the availability at each sampling location and are listed in Table 3.
Filter feeders, particularly shellfish, have been considered popular choices as bioindicators of microplastic pollution in marine environments. This is mostly due to the previous success of shellfish, particularly mussels, acting as bioindicators for all types of monitoring purposes, such as the monitoring of persistent organic pollutants, heavy metals, and emerging pollutants via “Mussel Watch” (Goldberg, 1986; Cantillo, 1998; O’Connor, 1998; Nakata et al., 2012). Their suitability for monitoring is due to their widespread availability in global coastal waters, sessile lifestyle, resilience, and ability to concentrate pollutants from their surrounding waters, and they can be found in large populations (Goldberg, 1986; De Witte et al., 2014; Cho et al., 2021). Thus, bivalves have also been considered alongside other shellfish species as bioindicator species for microplastic pollution, because they can filter large amounts of seawater and retain microplastics from the water column (Cho et al., 2021).
Several studies have explored the potential of shellfish as a bioindicator of microplastic pollution by investigating the correlation between microplastic abundance in shellfish and their surrounding environment. While clams have been shown to have a closer association with microplastic abundance in sediment (Ding et al., 2021), mussels and oysters have been reported to exhibit a positive correlation to a certain extent with their surrounding waters (Li et al., 2019; X. Qu et al., 2018; Chinfak et al., 2021; Ding et al., 2021; Wang et al., 2021). However, Cho et al. (2021) were unable to obtain a positive correlation between microplastic abundance in mussels/oysters and the surrounding seawater. Both mussels and oysters showed a lower microplastic abundance than the calculated estimate based on seawater. This could be attributed to the intricate movement of microplastics in water columns and the rejection of some microplastic particles through pseudofaeces, and that a portion of the captured microplastics could be egested prior to ingestion (Li et al., 2019; Chinfak et al., 2021; Cho et al., 2021). Similar to clams, a particular species of barnacles (Amphibalanus Amphitrite) was reported to have a positive correlation with sediment (Xu et al., 2020), suggesting its potential as a bioindicator. Sea urchins have also demonstrated significant spatial variability and have been suggested to a have positive correlation with microplastic abundance in the water column (Feng et al., 2020). Fallon and Freeman (2021) revealed that sponges have lower microplastic concentrations than the surrounding seawater. This could be due to the inclination of sponges for very small particles and their resistance to ingesting or retaining microplastics. Generally, apart from sponges, most studies indicated that concentrations found in filter feeders were somewhat associated with the pollution status of an area, in which filter feeders with higher concentrations of microplastics were almost always found in more urbanized and industrialized areas.
For selective feeders such as predators and foragers, microplastic abundance, polymer type, and particle type seemed to be influenced by feeding habits, habitat, size, and microplastic bioavailability (Jabeen et al., 2017; Arias et al., 2019; Nakao et al., 2020; Sparks and Immelman, 2020; Białowąs et al., 2022). In terms of spatial variation in microplastic abundance, selective feeders were also shown to reflect, to a certain degree, the abundance of its surrounding feeding habitat. Unlike filter feeders, which are only able to ingest smaller microplastics (<1,500 μm), selective feeder might be able to indicate a more generally view of the microplastic pollution status of an area, due to its ability to ingest larger size microplastics (~5,000 μm or larger) as well (Teng et al., 2019; Białowąs et al., 2022). However, the wide variety of species, habitats, feeding habits, metabolic rates, and behaviors could make them difficult to be used as bioindicators. Most importantly, fishes have also been shown to actively reject microplastics by spitting them out after intake (Li et al., 2021). There were no clear trends among the species and habitats. Pelagic species mostly feed in shallow water near the surface, where lower density microplastics tend to float and contained more microplastics (Lusher et al., 2013; Tanaka and Takada, 2016; Ferreira et al., 2020; Gurjar et al., 2022). However, in some cases, demersal species, which are mostly carnivorous, feeding in deeper waters on benthic fish, mollusks, and crustaceans were shown to have higher concentrations than pelagic species. This could be due to biofouling, the presence of denser polymers, prey types and higher microplastic abundance near the seabed (Jabeen et al., 2017; Koongolla et al., 2020; Mistri et al., 2022). Lusher et al. (2013) reported no significant differences in microplastic abundance between pelagic and demersal species. Nevertheless, Lusher et al. (2013) revealed that polymers with lower densities, such as polystyrene, acrylics, and low-density polyethylene, were exclusively found in pelagic species. Meanwhile, low-density polyamide and denser fibers, such as polyester and rayon, have been found in both pelagic and demersal fishes (Lusher et al., 2013). In terms of particle type, Sparks and Immelman (2020) suggested that demersal species ingested more fragments, whereas pelagic species ingested more fibers because of the bioavailability of these microplastics in their respective habitats. Although fibers and fragments have been shown to be prevalent in fish (Tanaka and Takada, 2016; Jabeen et al., 2017; Arias et al., 2019; Bakir et al., 2020; Sparks and Immelman, 2020; Wootton et al., 2021), rare cases in which plastic pellets or film predominated were also observed (Adika et al., 2020; Wootton et al., 2021). In addition, some species were speculated to have little accumulation potential for microplastics in the gastrointestinal tract and would only reflect the recent ingestion of microplastics (Foekema et al., 2013; Güven et al., 2017; Jovanović et al., 2018).
The current main issue with microplastic monitoring in the environment, regardless of matrix, water, biota, sediment, air, etc., is a challenge when trying to compare and review results among different studies. This is due to the many different methodologies used by different laboratories globally to sample, extract, identify, and quantify microplastics from environmental samples. There have been many discussions regarding transparency and reproducibility in the methods described in the literature, as well as the necessity for standardization in the monitoring of microplastics in the environment (Müller et al., 2020; Lv et al., 2021; Ding et al., 2022). Both Ding et al. (2022) and Hermsen et al. (2018) have suggested various protocols and strategies that can be adopted for standardization and to increase the reliability of the studies, as well as to improve comparability among studies.
In favor of standardization, we suggest that mussels can be used as standard bioindicators for microplastic monitoring in marine environments. As previously discussed, the widespread distribution and accessibility of mussel sampling allows for global sampling, analysis, monitoring, and comparison. Their filter-feeding behavior also allows them to efficiently accumulate microplastic particles from the water column. Furthermore, mussels are sessile organisms, meaning that they remain in one place for an extended period, which enhances their ability to integrate and accumulate microplastics over time, providing valuable insights into long-term contamination trends (Goldberg, 1986; De Witte et al., 2014; Cho et al., 2021). Mussels have also been shown to positively reflect the pollution status of surrounding waters (X. Qu et al., 2018; Li et al., 2019; Chinfak et al., 2021; Ding et al., 2021; Wang et al., 2021). Owing to these characteristics of mussels, there will be fewer confounding variables to deal with; thus, monitoring using mussels can be more accurate, consistent, and reliable. There are still limitations in terms of representation, in which mussels can only represent the pollution status in certain water column depths, habitats, and microplastic size ranges. However, mussels may still be a better bioindicator than other marine organisms in terms of variability control. This has been proven by the success of “Mussel Watch” over the years for the monitoring of persistent organic pollutants, heavy metals, and emerging pollutants (Goldberg, 1986; Cantillo, 1998; O’Connor, 1998; Nakata et al., 2012).
Microplastic abundance tends to be higher in coastal areas that receive inputs from land and marine sources. Microplastics tend to accumulate and remain in coastal areas due to the lack of water circulation. Convergence zones exist in the world’s oceans, where microplastic abundances are denser, such as in gyres. Most marine biota demonstrated a correlation between microplastic abundance and surrounding waters. However, there are too many variables among selective feeders, making them less reliable as bioindicators. As such, filter feeders, especially mussels, are recommended as the most reliable bioindicators for microplastic monitoring in marine environments. However, it is still important to monitor microplastics in commercial species, as they are widely consumed as delicacies, which can affect public health.
We would like to express our sincere gratitude to all the members of our laboratory for their valuable feedback and assistance throughout the writing of this review paper. We appreciate their enthusiasm, dedication, and collaboration in advancing our research field. I would also like to thank my colleague, Dr. Jeffrey Robens, for his understanding and flexibility in accommodating my workload and deadlines. Finally, I would like to acknowledge Ruben Soares Luis for his constant advise and support throughout the writing of this paper. His support and encouragement were very helpful and motivating.
This review paper does not involve any original data or materials. All the data used in this paper are derived from published literature and are cited in the main text or main tables. No additional data or materials are available.
The authors declare that they have no conflicts of interest regarding the publication of this paper.
