Primary microplastics are those that are manufactured in the microplastic size range and include scrubbers or microbeads used in personal care products, abrasive materials used in cleaning and air-blasting products, and feedstocks used in plastic manufacturing (resin powder or nurdle pellets, see sidebar 1). Secondary microplastics are fragments from the breakdown of larger plastics through physical, chemical, or biological processes. Sources of secondary microplastics include synthetic fibers shed from textiles (clothing, carpets, upholstery) during general use, washing, and drying (Hartmann and others, 2019; Kapp and Miller, 2020; Tao and others, 2022), and breakdown of larger plastics including fishing equipment, car tires (through road wear and reuse applications in artificial turf and rubberized pavements), agricultural plastics and mulches, and single-use items such as bags, bottles, cups, takeout containers, straws, cigarette butts, and medical waste, among many others (Hidalgo-Ruz and others, 2012; Eerkes-Medrano and others, 2015; Duis and Coors, 2016; Alimi and others, 2018; Zangmeister and others, 2022). Further breakdown of plastics from these and other sources yields nanoplastics. For example, research has documented that single-use consumer plastic products (such as single-use bags and beverage cups) release trillions of nanoplastic particles per liter into water during normal use (Zangmeister and others, 2022). Sidebar 4 lists plastic types that are common sources of microplastics and example of items using each type.

Microplastics reach aquatic and terrestrial environments through pathways such as wastewater, stormwater runoff and sewage overflow, and the atmosphere. Wastewater represents two important pathways: wastewater treatment facility effluent (that is, microplastics in treated liquid waste) and biosolids (that is, microplastics in treated solid waste) (Edo and others, 2022). Many studies have shown that wastewater treatment facilities remove almost all microplastics from influent water (95–99 percent removal; Carr and others, 2016; Talvitie and others, 2017), yet wastewater effluent still represents a significant pathway for microplastics to receiving waters because of the high concentrations of microplastics in the influent, and the large volume and continuous flow of effluent (Rochman and others, 2015; Mason and others, 2016; Talvitie and others, 2017). The wastewater biosolids, or sludge, captured by treatment facilities are an additional and distinct pathway for microplastics to the environment (Edo and others, 2022). Biosolids contain thousands to tens of thousands of microplastic particles per kilogram (kg) dry weight (Mahon and others, 2017), and roughly 50 percent of the 730 million kg of biosolids collected annually in the United States are land-applied on agricultural fields, forests, reclamation sites, and lawns and gardens (U.S. Environmental Protection Agency, 2016). Microplastics in land-applied biosolids may be lofted to the air by wind, incorporated into soils, or enter waterbodies through runoff. Estimates of the microplastic burden on soils in North America and Europe due to biosolids application exceed estimates of the burden in surface waters of the global oceans (Nizzetto and others, 2016). Research from Australia estimates that 4,700 metric tons (t) of microplastics are released into the Australian environment annually through land application of biosolids (Okoffo and others, 2020). In the United States, irrigation diversions from streams represent an additional pathway for microplastics to the terrestrial environment (Kukkola and others, 2023).

Stormwater runoff and sewage overflow systems represent another important pathway for microplastics to waterways (Werbowski and others, 2021; Zhu and others, 2021). Concentrations of microplastics in stormwater exceed those in wastewater effluent by multiple orders of magnitude (Sutton and others, 2019; Werbowski and others, 2021; Zhu and others, 2021). Watershed-scale comparisons of microplastic loading from different pathways are limited, but a study of microplastic loading to San Francisco Bay estimated stormwater loads to be 300 times greater than those in wastewater effluent (Sutton and others, 2019). These estimates are almost certainly affected by local land cover, terrain, stormwater treatment practices, and climate characteristics, and thus, may not reflect relative load contributions in other locations but serve to highlight the importance of stormwater as a pathway for microplastics to reach aquatic environments. Preliminary studies indicate that bioretention cells, rain gardens, and other stormwater retention structures can be effective tools for reducing microplastic loads in stormwater, capturing 83–96 percent of particles (Gilbreath and others, 2019; Smyth and others, 2021; Werbowski and others, 2021).

The atmosphere is an additional pathway through which microplastics reach aquatic and terrestrial environments. Microplastic fibers and fragments are lofted into the atmosphere from roads, the ocean (through wind and wave action), agricultural fields, and population centers (Brahney and others, 2021), and once aloft, may be transported long distances (Allen and others, 2019; Bergmann and others, 2019). Atmospheric deposition of microplastics occurs in all landscapes, even remote areas such as the Alps (south-central Europe), the Arctic (above 66.5 degrees north latitude), and protected areas of the western United States (including hard-to-reach areas in Yellowstone, Alaska, and the Rocky Mountain region), at rates ranging from 50 to 700 particles per square meter per day (m²/d) (Cai and others, 2017; Dris and others, 2017; Allen and others, 2019; Bergmann and others, 2019; Wetherbee and others, 2019; Brahney and others, 2020a). An estimated 22,000 t of microplastics are deposited on the contiguous United States each year (Brahney and others, 2021).

Microplastic transport in aquatic systems is affected by many factors, including buoyancy (specific gravity), size, and shape, which determine whether a microplastic particle will settle or remain in the water column (Horton and Dixon, 2018). Environmental considerations such as flow velocity, turbulence, wind, currents, and wave action also affect microplastic movement and transport (Horton and Dixon, 2018). The chemical characteristics of the water further affect whether a microplastic particle or fiber will float on the surface, suspend in the water column, or deposit in bed sediment. The formation of biofilms on microplastic surfaces, aggregation with other microplastics and sediment particles, and interaction with organic matter also play a role in transport properties, deposition, and potential for resuspension (Semcesen and Wells, 2021).

Weathering processes (photodegradation, thermal degradation, biodegradation, and mechanical fragmentation) can change the physical shape, size, and chemical properties of microplastics, ultimately affecting microplastic fate (Duan and others, 2021). Some microplastics may be slowly biodegraded by micro-organisms (Eerkes-Medrano and others, 2015; Rocha-Santos and Duarte, 2015), although even degradable plastics remain in the environment for decades to centuries (Barnes and others, 2009; Roy and others, 2011). Some plastics are more susceptible to oxidative weathering by ultraviolet light (photodegradation), although rates vary based on water column or soil mobility considerations that affect intensity (light penetration at depth; Hebner and Maurer-Jones, 2020; Masry and others, 2021). Microplastics can be mistaken for edible material by wildlife, or simply swallowed with surrounding water. While some microplastics may become trapped in the guts and gills of organisms, others may fragment or undergo surface degradation by digestive enzymes (Dawson and others, 2018). These different weathering processes affect microplastic surface properties (creating cracks, for example), which can lead to increased adsorption of heavy metals and organic contaminants, and fragmentation into smaller particles (Duan and others, 2021; Bhagat and others, 2022; Hadiuzzaman and others, 2022). With fragmentation, microplastics become more prevalent, surface area increases relative to volume, and particles become more bioavailable to organisms at the base of the food web (Botterell and others, 2019).

In aqueous environments, microplastics float or settle depending on particle density, size, shape, degree of surface biofouling, and waterbody hydraulics (Cowger and others, 2021; Semcesen and Wells, 2021). However, even low-density microplastics will eventually sink with increasing fragmentation and surface biofouling (Peng and others, 2020; Semcesen and Wells, 2021). For this reason, and based on microplastic concentrations in these environments, streambeds, lakebeds, and sea floors are often considered the ultimate sinks for microplastics (Woodall and others, 2014; Peng and others, 2020; Lenaker and others, 2021; Drummond and others, 2022).

In terrestrial environments, soils are a sink for microplastics accumulated through atmospheric deposition, land application of reuse materials (for example, biosolids, livestock manure), and mismanagement and breakdown of trash (He and others, 2018; Brahney and others, 2020a; Jin and others, 2022; Tagg and others, 2022). However, soils may only be a temporary sink, as wind may resuspend microplastics into the atmosphere, runoff may wash microplastics into a nearby waterway, or microplastics may infiltrate into groundwater (Grbić and others, 2020; Brahney and others, 2021; Bigalke and others, 2022; Rezaei and others, 2022).

Throughout the microplastic cycle, microplastics can act as both a source and a sink for chemical pollutants. Compounds added to plastics during the manufacturing process (for example, flame retardants, plasticizers, antibacterial agents, ultraviolet inhibitors, reaction byproducts) can leach out of plastics (Schrank and others, 2019; McIntyre and others, 2021). Conversely, microplastics may absorb pollutants in aquatic and terrestrial ecosystems, potentially concentrating persistent, bioaccumulative and (or) toxic organic and microbiologic pollutants and metals (Imran and others, 2019; Wang and others, 2021a; Hu and others, 2022). The absorption and desorption of chemical contaminants from microplastics may create regional hotspots of chemical pollutants due to microplastic transport/deposition (Eerkes-Medrano and others, 2015) and (or) contribute to the bioaccumulation of hazardous compounds in aquatic and terrestrial organisms (for example, polybrominated diphenyl ether added to plastics as flame retardants; Rochman and others, 2014).

The forensics of source apportionment of microplastics is complicated by the pervasive and overlapping use of plastics for different applications. For example, polypropylene is used in packaging, construction, furniture, transportation, electronics, textiles, and many other application sectors (Di and others, 2021). Hence, identifying the source of a polypropylene particle found in the environment is a challenge. Microplastic morphology, additives, and surface contaminants including metals, organic contaminants, and biofilms may help identify microplastic sources and (or) pathways (Fahrenfeld and others, 2019). Microplastic assemblages may also provide clues: studies have found that waters from different pathways (stormwater runoff, agricultural runoff, and wastewater effluent) may have distinct signatures (Grbić and others, 2020; Zhu and others, 2021). Stormwater, for example, had a high proportion of black rubbery particles, presumably from car tires, whereas wastewater was dominated by fibers, a large proportion of which were cellulose, presumably from washing of clothes (Grbić and others, 2020; Zhu and others, 2021).

Most of the science opportunities related to microplastic sources, pathways, and fate require USGS access to a microplastics laboratory, either internal or external, with the capability of providing microplastic quantification and polymer identification for complex environmental matrices; for example, stormwater, biosolids, sediment, soil, and biological tissue (table 1). The laboratory needs to handle large numbers of samples and meet USGS quality-assurance standards. The USGS has the expertise and experience to develop these capabilities. Thus, in the short term (1–2 years), the USGS could begin to address several of the microplastic science gaps identified in this section by designing studies aimed at addressing science gaps, beginning sample collections, purchasing laboratory equipment, and developing analytical capabilities, but study completion would likely be a longer term effort requiring more than 2 years.

There are several long-term opportunities for the USGS to help fill science gaps related to microplastic sources, pathways, and fate (table 1). For example, the USGS could leverage its expertise and national monitoring networks for holistic, multimatrix watershed-scale evaluations of microplastic sources and pathways in different regions to determine the relative importance of different sources and pathways and inform source and pathway apportionment. Additionally, the USGS could assess microplastic mitigation and sequestration strategies across a variety of natural and built environments (for example, green infrastructures); evaluate natural and anthropogenic drivers that deliver microplastics to groundwater systems; investigate chemical sorption and desorption of contaminants in environmental matrices and leaching of chemicals to determine the relative effect compared to contaminants in other environmental compartments; characterize the fragmentation/breakdown properties of different polymers into smaller size fractions under varied environmental conditions; and determine how the fate of microplastics and microplastic-associated contaminants is affected by natural organic matter, micro-organisms, and growth of biofilms on microplastic surfaces.

Early research on the presence of microplastics in the environment was completed in aquatic ecosystems, with marine studies (77 percent) dominating (de Sá and others, 2018; Miller and others, 2020). In recent years, studies have expanded greatly into freshwater and terrestrial ecosystems (Eerkes-Medrano and others, 2015; Sarijan and others, 2021; Wong and others, 2020). Ingestion of food and water are considered the most important pathways of microplastic uptake for aquatic and terrestrial wildlife and humans (Prata and others, 2020; Roch and others, 2020; Zantis and others, 2021). However, a recent study reviewing ingestion of table salt, drinking water, and inhalation of air as exposure routes indicated inhalation (indoor air) to be the most significant pathway for human exposure (Zhang and others, 2020). In addition, fish (Kiryu and others, 2000; Lu and others, 2016; Bhagat and others, 2020) and aquatic invertebrates (Watts and others, 2014; Kolandhasamy and others, 2018) can also absorb microplastics through the skin, gills, and other exposed tissue. Microplastics in aqueous exposures attached to the chorion of zebrafish (Danio rerio) eggs but were not observed in the embryos (Batel and others, 2018). To date, no clear evidence of maternal transfer has been demonstrated in higher vertebrates, although microplastics were observed in human placentas (Ragusa and others, 2021).

Accidental ingestion may be common for passive filter feeders and for active foragers such as fish and some invertebrates, which may mistake the particles for food (Roch and others, 2020). In a laboratory study, fish species relying on visual foraging cues ingested more microplastics than those relying on chemosensory cues (Roch and others, 2020), which may explain why many pelagic fish species have been reported with higher concentrations than benthic species in the marine environment. However, in the riverine environment, benthivores may have an even higher risk than plankivores due to accumulation of microplastics in sediments (McNeish and others, 2018). Diet breadth may also affect microplastic uptake. In a study of intertidal fishes, omnivorous species had higher amounts of microplastics in their guts than herbivores or carnivores (Mizraji and others, 2017).

The potential for bioaccumulation (accumulation within the tissues of an organism at a level greater than the surrounding environment) of microplastics seems to be variable according to current (2023) literature. Gouin (2020) reviewed data from more than 800 organisms at all levels of biological organization in which microplastics have been observed. Approximately 90 percent of the studies enumerated microplastics in the stomach and (or) gastrointestinal tract. It is documented that the presence of microplastics in the gut is often transient as both ingestion and egestion or excretion of microplastics occurs (Woods and others, 2018; Xiong and others, 2019). Assessment of the data did not support bioaccumulation or biomagnification. Conversely, McIlwraith and others (2021) found that seven species of sportfish showed widespread presence of microplastics in the guts, livers, and fillets. Although microplastics in the gut may be excreted, in 2023, it is unknown if particles in other tissues will be as easily removed. In a review and meta-analysis of studies in the marine environment, Miller and others (2020) indicated that bioaccumulation occurs within trophic levels, but biomagnification (increasing concentrations with increasing trophic level) was not observed. The bioaccumulation appeared to be more strongly related to feeding strategies than trophic levels. More information is needed about how microplastics enter specific tissues and if they remain there.

Trophic transfer (the transfer of contaminants from one trophic level to another) of microplastics in simple food chains has been documented. For instance, consumption of contaminated zooplankton resulted in the translocation of microplastics to benthic filter feeders (Van Colen and others, 2020). Predatory midges were also found to accumulate microplastics through consumption of larval (Culex) mosquito prey (Cuthbert and others, 2019). Microplastics can also become entrapped in biofilms and then become available to biofilm grazers (McCormick and others, 2016). Microplastics adhered to or within biofilms on aquatic plants can be a route of exposure when the plants are eaten by herbivorous invertebrates and fishes (Gutow and others, 2016; Goss and others, 2018). Zhang and others (2019) measured microplastic abundance in 11 wild fish species and 8 wild crustacean prey species and found that abundance was significantly lower in crustaceans than fish species, leading to speculation that trophic transfer occurred.  

 The mechanisms driving microplastic trophic transfer in more complex food webs is relatively unknown (Provencher and others, 2019; Krause and others, 2021). A study evaluated microplastic transfer from water to tadpoles to fish to mice. Tadpoles ingested and accumulated microplastics from the water, and fish that ate those tadpoles had microplastics not only in the gut but also the liver. Microplastics were also observed in the livers of mice fed ground contaminated fish (da Costa Araújo and Malafaia, 2021).

Shifts in habitat caused by ontogenetic changes in animals (that is, changes in appearance, habitat, or diet, during its development) could also be important in mediating movements of microplastics within and across ecosystem boundaries. Foraging movements, migrations, and other ontogenetic shifts in habitat such as caused by insect metamorphosis can alter distribution and exposure to microplastics. Fish and bird migrations can move particles long distances from the source and between aquatic and terrestrial ecosystems (Bourdages and others, 2021). Microplastic beads in larval aquatic insects, such as mosquitos, can be transferred from larvae to pupae to adult, although the number of particles in adults are greatly reduced (10–100 times lower) compared to numbers in pupae or larvae (Al-Jaibachi and others, 2018). Because Culex oviposition, habitat selection, growth, and survival through metamorphosis were not affected by microplastic accumulation, even a small per individual transfer of microplastics could export a large quantity of microplastics from aquatic ecosystems to the terrestrial environment (Al-Jaibachi and others, 2019). Microplastics adhered to the periphyton can be ingested by tadpoles which, through metamorphosis, could be another transfer path from the aquatic to terrestrial environment (Boyero and others, 2020). Finally, feces and decomposition of plant and animal tissue can be another pathway of microplastic movement within and among ecosystems. For instance, filter-feeding benthic invertebrates can ingest microplastics from the water column and transfer them to benthic organisms through feces or pseudo feces (Krause and others, 2021).

The USGS is working with Federal agencies, State agencies, Tribes, nongovernmental organizations, and academic institutions to expand its research on human and wildlife exposure routes from microplastics (table 2). One short-term opportunity would be the derivation of a predictive framework for understanding the routes of microplastic exposure to wildlife based on previous knowledge of environmental transport and bioaccumulation of contaminants known to adsorb to particles. In addition, the USGS and its partners could study trophic and ontogenetic transfer, which would include the transfer of microplastics and degraded compounds among organisms, as well as determining the effects of migratory and ontogenetic movements on the transport of microplastics from place to place.

The USGS could contribute to and (or) start collaborative interagency or multi-institution research studies that would identify exposure routes of microplastics that have not been adequately researched. Paired laboratory and field studies of the patterns of uptake and transfer of microplastics and their associated chemicals would reveal causal mechanisms and landscape patterns of exposure to microplastics (table 2).

The toxicity of microplastics is concentration dependent (Gutow and others, 2016; Lu and others, 2016; Everaert and others, 2018) and varies based on the type, size, and shape of the particles (Huang and others, 2020; Zimmermann and others, 2020). Generally, smaller particles are more toxic than larger ones because they have greater surface-area-to-volume ratios and can translocate through cellular membranes (Jeong and others, 2017). However, this pattern is highly reliant on the mechanism of effect (for example, false satiation leading to reduction of food ingestion) versus physiological response (damage to gut walls and hypoxia), species morphology, and life stage (Martins and Guilhermino, 2018; Wu and others, 2019).

Microplastics are toxic to primary producers such as phytoplankton, disrupting feeding and photosynthesis, and negatively affecting growth, development, and reproduction. Additionally, polysaccharides secreted by phytoplankton may agglomerate (form into a mass) microplastics (Long and others, 2017), creating an exposure pathway for higher trophic species. Copepods (Calanus helgolandicus) feeding in the presence of microplastics consumed 40 percent less carbon biomass, resulting in lower growth, fecundity, and survival (Cole and others, 2015). Corals may be affected both through ingestion and passive adherence, leading to adverse effects on coral cleaning and feeding mechanisms. Soares and others (2020) documented many effects, including reduced growth, decreased activity of detoxifying and immunity enzymes, and negative effects on coral-Symbiodiniaceae relationships.

Exposure of insects such as honeybees (Apis spp.) to microplastics in food affected their gut microbiome, leading to decreased gut microbiota α-diversity (Wang and others, 2021b). These changes altered the expression of certain genes, including some related to the immune system, although bee growth was not affected. Dosing the bees with antibiotics greatly increased the lethality of the microplastics to the bees, leading the authors to indicate that the bee gut microbiome protects the bees from the toxic effects of microplastics (Wang and others, 2021b). Alterations of gut microbiomes of other invertebrates such as shrimp (Duan and others, 2021) have also been reported. Studies have also documented aquatic vertebrates, such as zebrafish, with intestinal inflammation and disorders of the metabolome (Qiao and others, 2019; Roch and others, 2020).

Microplastics in organisms such as fish and aquatic invertebrates affect feeding, growth, development, immune response, reproduction, and survival, and they can cause tissue inflammation (Foley and others, 2018; Ahrendt and others, 2020; Horn and others, 2020). The presence of microplastics is often evaluated in the stomach and gastrointestinal tracts, and correlations are made between presence of microplastic particles and observed effects. While it has been documented that microplastic particles can be absorbed from gut, gill, and body surfaces and transported to other tissues, many species quickly eliminate certain types and sizes of microplastics (Graham and others, 2019; Spanjer and others, 2020). The uptake of particles from the gut and translocation are related to particle size. Laboratory studies with zebrafish indicated that microplastic particles (<5 μm) were translocated to the liver (Lu and others, 2016). More toxicity data are needed at environmentally relevant concentrations and for environmentally relevant particle types.

During manufacturing, chemical additives such as metals and organic chemicals are often added to plastics to improve performance, elasticity, slow degradation, or reduce flammability, among other purposes. Chemical additives include, but are not limited to, flame retardants, phthalates, phenols, bisphenol A, antimicrobials, and synthetic antioxidants. Some of these additives may be endocrine disrupting and (or) toxic to organisms if they later leach from plastics into the environment or directly into their digestive tracts (Chen and others, 2022). Understanding the role of additives on the environment and wildlife is critical.

Multiple studies on aquatic organisms, including invertebrates, fish, and waterbirds, have shown the synergistic negative effects of microplastics and adsorbed contaminants, including changes in blood chemistry, foraging and swimming behavior, genotoxicity, oxidative damage to cells, metabolic activity, immunotoxicity, neurological responses, alterations in gene expression, liver toxicity, and pathology after chronic exposure (Huang and others, 2021). Leaching of these metals and organic compounds into the environment depends on the physicochemical properties of the chemicals and can change due to a range of factors, including salinity, hydrophobicity and (or) lipophilicity, molecular weight, surface charge, water or air velocity, and exposure to sunlight (Chen and others, 2019a). For example, ultraviolet radiation accelerates the release of polybrominated diphenyl ethers (Khaled and others, 2018), while salinity slows the leaching behavior of phthalates (Paluselli and others, 2019). Microplastics collected from the marine environment continued to leach endocrine-disrupting chemicals (for example, multiple phenols) when exposed to solar irradiation for 1 month (Chen and others, 2019b). Other researchers have demonstrated the leaching of toxic chemicals from tire rubber particles, including 6PPD-quinone, polyaromatic hydrocarbon congeners, benzothiazole, and zinc (Panko and others, 2013; Redondo-Hasselerharm and others, 2018; Capolupo and others, 2020; Kolomijeca and others, 2020; Tian and others, 2021).

Ingested plastics can leach chemicals directly into the gastrointestinal tract of consumers through the desorption of intrinsic contaminants. Estrogenic and polybrominated diphenyl ethers chemicals were shown to leach from multiple plastic types, including polypropylene, polystyrene, and polyethylene (see sidebar 4), when placed in laboratory mimics of gastrointestinal conditions of both fish and sea birds (Coffin and others, 2019; Guo and others, 2020), but co-ingestion of sediment and food items can confound the amount of leachate that makes it into the organism through the adsorption of these leached chemicals. Additional work has shown that multiple metals leach from ingested plastics, including bromine, cadmium, chromium, mercury, lead, and antimony, and that their mobilization in the gut fits standard diffusion models (Smith and Turner, 2020). Although many chemicals leach into the gut after ingestion, little work has demonstrated acute or chronic toxicity or determined if microplastics represent a pathway for intrinsic plastic chemical additives that bioaccumulate and biomagnify at environmentally relevant microplastic concentrations.

The exposure to and effects of leached microplastic chemicals on organisms can vary. Microplastics can act as both a vector (delivering contaminants to an organism) and a scavenger (restricting the uptake of contaminants to an organism) of organic contaminants, dependent on the concentration of both the additives and the plastic (Liu and others, 2020). Field studies have correlated microplastic uptake with chemical concentrations of leachate in tissues (Jang and others, 2016; Khoshmanesh and others, 2023). Laboratory studies have documented the potentially acute toxic effects of leached plastic additives (both organic and metal) on aquatic organisms through leaching experiments (Huang and others, 2020). However, the presence of microplastics in the gut along with exposure to different chemical additives can have confounding effects, with plastics acting both to promote (Wardrop and others, 2016) and prevent (Chua and others, 2014) the uptake of organic contaminants.

In addition to acting as a vector of leached plastic additives, microplastics are shown to effectively concentrate contaminants, including organic contaminants and metals, from their surrounding medium (that is, air or water) (Brennecke and others, 2016; Liu and others, 2019). The concern about microplastics acting as a vector for exogenous compounds is twofold: microplastics can act as vectors for these adsorbed contaminants and absorbed substances may increase microplastics uptake. For example, Athey and others (2020) found that larval fish more readily consumed zooplankton prey containing microplastics that had been treated with DDT (dichloro-diphenyl-trichloroethane) than they consumed zooplankton that contained untreated plastics.

Factors affecting the toxicity of adsorbed contaminants are similar to factors affecting contaminants that leach out from virgin (that is, newly manufactured) plastics. There are potentially thousands of contaminants that can be adsorbed to microplastic particles. The type of contaminants adsorbed by microplastics is dictated mainly by a chemical’s hydrophobicity (though, pi interaction and electrostatic interaction are also important); more hydrophobic chemicals are more likely to adsorb to microplastic polymers (Tourinho and others, 2019; Luo and others, 2022). Multiple studies, on invertebrates, fish, and waterbirds, have shown the synergistic negative effects of adsorbed contaminants and microplastics, including changes in blood chemistry, changes in foraging and swimming behavior, genotoxicity, oxidative damage to cells, metabolic activity, immunotoxicity, neurological responses, alterations in gene expression, liver toxicity, and pathology after chronic exposure (Huang and others, 2020). Microplastics can also act as antagonists by adsorbing aqueous phase contaminants, which reduces organism exposure, even when those contaminant-laden plastics are ingested (that is, desorption in the organism does not occur) (Rehse and others, 2018). Results are highly chemical, plastic, and species specific, making it difficult to generalize the effect of exogenously absorbed contaminants on aquatic organisms. A critical review and re-interpretation of empirical studies by Koelmans and others (2016) indicates that in most marine habitats, consumers assimilated hydrophobic organic chemicals from natural prey tissues in much greater amounts than they assimilated from ingested microplastics; thus, microplastic ingestion is not likely to increase exposure to those chemicals.

By providing a novel substrate and potential source of energy and toxicants to micro-organisms, microplastics can change the assemblages and functions of microbe communities in the environment and guts of living organisms (Lu and others, 2019; Helmberger and others, 2020; Ma and others, 2020). These changes have implications for human and environmental health. For example, microbial colonization of microplastics in wastewater can possibly lead to the spread of antimicrobial resistance genes in biofilms in wastewater treatment plants (Eckert and others, 2018; Lu and others, 2019). Bacteria, fungi, and other micro-organisms can colonize microplastics, forming biofilms called the plastisphere (Zettler and others, 2013). A study indicated that plastics were home to taxonomically enriched and distinct fungal communities and accumulated pathogenic fungi within a riverine system (Xue and others, 2021). Prata and others (2019b) reviewed the effects of microplastics on microalgal populations and concluded that current (2023) environmental concentrations would not cause toxicity. However, disruption of populations could occur by reducing the availability and (or) absorption of nutrients or an increase in populations by decreasing predator populations such as zooplankton that may be more susceptible to microplastic toxicity (Foley and others, 2018).

Population effects of microplastics have also been documented in invertebrate species. Mueller and others (2020) documented changes in nematode population density in sediments exposed to microplastic beads. Martins and Guilhermino (2018) documented that exposure of Daphnia (small planktonic crustaceans) to ecologically relevant microplastic concentrations (polluted sites) led to mortality even beyond the exposed first generation. After a single exposure, succeeding generations of Daphnia exhibited lower fecundity and did not recover to preexposure condition until the third generation (Martins and Guilhermino, 2018).

Finally, predator-prey interactions can be affected by the presence of microplastics and their leachates, although the mechanisms and magnitude of effects (if any) are variable. A meta-analysis of reported effects of microplastics on fish and aquatic invertebrates found that a reduction in prey consumption was the most consistent effect detected across studies (Foley and others, 2018). In an experimental trial, predation rates of zooplankton larvae by benthic filter feeders (bivalve mollusks) were 30 percent lower when zooplankton were contaminated with microplastics than when clean prey were available (Van Colen and others, 2020). The microplastic uptake by zooplankton altered their swimming behavior and made them less likely to be filtered from the water column and consumed (Van Colen and others, 2020). In another experiment, microplastic leachates from virgin and bleached microplastic pellets impaired the vigilance and predator-avoidance behaviors of an intertidal snail (Littorina littorea) by approximately 15–64 percent when exposed to chemical cues of a common predatory crab (Carcinus maenas) (Seuront, 2018). These findings imply that consumption rates of these larval fish could potentially increase in the presence of microplastics. Consumption of mosquitoes by Chaoborus (midge) larvae (attack rate and handling time) was not affected by uptake of microplastic particles by mosquito prey (Cuthbert and others, 2019). Thus, different studies indicated that microplastics could make prey better or worse at avoiding predation or evoke no change. In terms of how these effects could play out at the population level, a modeling study extrapolated that reductions in feeding due to presence of microplastics would have negligible effects on the population dynamics of the predator-prey system (Huang and others, 2020).

Ultimately, the effects of microplastics on microbes and other organisms may lead to changes in ecosystem functions and services, such as productivity, nutrient cycling, carbon sequestration, and pollination (Helmberger and others, 2020; Ma and others, 2020). As a result, scientists have used multiple lines of research to investigate the effect of microplastics on ecosystems. For productivity, a large-scale model of the North Sea (lat 56.5110° N., long 3.5156° E.) predicted that microplastics would not affect the total primary or secondary production, but spatial patterns of secondary production would be altered (Troost and others, 2018). Nutrient cycling could be affected by microplastics by providing a novel substrate for colonization by a plastic-specific assemblage of microbes (Eckert and others, 2018; Oberbeckmann and others, 2018; Chen and others, 2020) and by altering filtration rates by benthic bivalves (Cluzard and others, 2015; Green and others, 2017). In aquatic ecosystems, microplastic biofilms accelerate ammonia and nitrite oxidation as well as denitrification (Chen and others, 2020). For bivalves, filtering rates of Ostrea (oysters) increased in the presence of microplastics, leading to reduced porewater ammonium concentrations and biomass of benthic cyanobacteria (Green and others, 2017). On the other hand, Mytilus (mussel) filtering rates did not change in the presence of microplastics (Green and others, 2017), while ammonium concentrations increased in the water surrounding commercial clams contaminated with microplastic beads (Green and others, 2017). Green and others, (2017) implied that their results showing increased ammonium concentrations in the water around clams exposed to microplastics could potentially lead to increased algal blooms around commercial shellfish operations. In terrestrial ecosystems, microplastics increased nutrient contents of organic matter dissolved from soil (Liu and others, 2017). Finally, a review of the effects of microplastics on insects indicated that multiple ecosystem services provided by insects could be affected including pollination, mostly because microplastic beads could mimic pollen (Oliveira and others, 2019). Interestingly, Oliveira and others (2019) also indicated that certain insect larvae might be able to help remediate the effects of plastics by consuming and digesting them.

The USGS could prioritize toxicological studies with different plastic polymers by (1) using data from new studies, (2) identifying the relative likelihood of toxicity of different polymers and associated chemicals by consulting existing databases such as CompTox (Williams and others, 2017) and ToMEx (Thornton Hampton and others, 2022), and (3) identifying sensitive and economically important species. Examples of species to study include Atlantic and Pacific herring (Clupea spp.) or river herring (Alosa spp.), anchovies (family Engrulidae), American eel (Anguilla rostrata), shrimp (family Caridae), and red snapper (Lutjanus campechanus). Integrative studies could prioritize toxicological studies within the USGS from multiple disciplines and within multiple mission areas, including ecosystems, water resources, and energy and minerals. This prioritization would start the transition from microplastics being generally considered as a single contaminant to categorizing and defining microplastics by toxicologically relevant groups. After prioritization, studies could focus on acute and chronic toxicity tests measuring sublethal endpoints in sensitive species across trophic levels (microbes to secondary consumers). Toxicity studies should consider different exposure routes and potential mixtures while being designed for use in environmental risk assessment applications (well replicated over a range of concentrations for the development of dose-response curves). Polymer-specific toxicity testing methods should be developed for many polymer morphologies (types) and sizes and should reflect what is most prevalent in environmental occurrence studies (fibers and nanoplastics). For toxicity modeling and prediction, sorption constants of exogenous and intrinsic chemicals could be determined for chemicals on a polymer-specific basis (table 3).

Experimental approaches for understanding microplastic toxicity need to be identified to better characterize microplastic toxicity with both physical and chemical modes of action. Studies are needed to understand the toxicity of both virgin and weathered microplastics. Once individual polymer, exogenous chemical, and polymer with intrinsic chemical studies are completed, more work will be needed on polymer mixture toxicity and sorbed chemical mixtures. Systems toxicology studies could be designed to elucidate the mechanism of toxicity of different polymers by integrating in vitro and in vivo toxicity data with computational modeling to inform and predict toxicity of microplastics to species not tested. Lastly, microcosm and mesocosm studies, completed alongside field studies, could be used to confirm laboratory effects in natural systems and to test microplastics' effects on populations and at the ecosystem level. Field studies, in particular, could be used to reveal the ecological and ecosystem effects of microplastics on patterns and processes such as decomposition, insect emergence, nutrient cycling, predator-prey interactions, and biomass and distribution of species of interest.

The need for standard protocols for collection of environmental microplastic samples is widely recognized as a critical obstacle in the science of microplastics (Koelmans and others, 2019; Brander and others, 2020; Watkins and others, 2021; Gimiliani and Izar, 2022). This is especially true for collection of microplastics in surface water. As of 2023, meaningful comparison of data across studies of microplastics in water is hindered by inconsistent sample collection methods and analyses (Hermsen and others, 2018), which have evolved along with the understanding of environmental occurrence, the potential for sample contamination, and improved analytical techniques, among other factors. Thus, there is a need for a comprehensive assessment of existing sampling protocols, especially in water, to develop a set of standardized methods that can be used to collect comparable data by personnel across the United States and in a range of conditions. The recent development of a sampling method by ASTM International (ASTM International, 2020) is a significant step toward standardization and could improve interstudy comparability among USGS studies. Before using any method, USGS laboratories are required to verify they can perform outside methods while meeting established acceptance criteria or will have to validate the suitability and performance of any newly developed methods (U.S. Geological Survey, 2022). Practical considerations, such as the amount of time and the equipment and supplies required to collect a sample, should also be considered.

Quality-control considerations unique to microplastics need to be included in the development of standardized microplastic sampling protocols. Perhaps the most important consideration is the potential for contamination. Because microplastics come from a wide variety of sources and are ubiquitous in both indoor and outdoor environments, there is a high potential for contamination during all stages of sample collection, processing, and analysis (Brander and others, 2020). Equipment, field, and laboratory blanks are therefore critical to assess potential contamination throughout the sample collection and analysis process. Replicate samples are also necessary to understand microplastic heterogeneity and sample representativeness. A quality-control consideration unique to microplastics is that microplastic densities vary by polymer, and therefore particles of different polymer types will distribute differently in the water column and sediment (Lenaker and others, 2019). Although not considered in many microplastic studies, this vertical heterogeneity is an important consideration in minimizing bias and collecting vertically representative samples in water. In studies to date, quality-control efforts related to sample collection have been inconsistent (Koelmans and others, 2019). Thus, to help guide the development of robust sample-collection methods, there is a need for rigorous assessment of quality assurance related to (1) maximizing sample representativeness and reproducibility in different media, (2) minimizing bias, and (3) minimizing contamination.

In the short term (1–2 years), the USGS could use its expertise and existing laboratory capabilities (see section 2.5.3., “USGS Capabilities and Expertise”) to evaluate existing sampling protocols, with the longer term goal of developing a nationally consistent protocol for the collection of microplastics samples in water (like surface water, stormwater, groundwater, drinking water), bed and suspended sediment, soils, and biota (table 4). As an example, the USGS could use the flumes at its hydraulic laboratory (Hydrologic Instrumentation Facility) to test different water column microplastics sampling methods under varying flow conditions and identify the strengths, weaknesses, and biases of each method. These and other activities should be done in cooperation with researchers at other Federal agencies and universities, who have together made significant progress toward these goals.

In the long term (3 or more years), the USGS could refine and publish sampling protocols for microplastics in different media, based on the evaluations described in section 2.4.4.1., “Short-Term Opportunities,” to meet the data-quality standards of the USGS and other agencies. Publication of standardized, nationally consistent sampling protocols would enable meaningful comparison of microplastic data across studies and agencies. Additionally in the long term, the USGS could partner with researchers who are developing sensors and machine learning techniques to provide real-time measurements or surrogate estimates of microplastics in surface waters.

In 2023, the USGS can use existing mass spectrometry expertise, equipment, and techniques, such as gas chromatography-mass spectrometry, LC/MS, and inductively coupled plasma-mass spectrometry, for analyzing microplastic matrices (table 5). These techniques are currently used for contaminant analyses (for example, with mercury, per- and polyfluoroalkyl substances, and polychlorinated biphenyls). This equipment could be immediately put to use in the characterization of sorbed intrinsic and extrinsic compounds (organic and inorganic contaminants) and the role of microplastics in fate, exposure, and adverse effects of these sorbed chemicals. Investment in new mass spectrometry techniques like Pyr-GC/MS or thermal desorption gas chromatography-mass spectrometry (TD-GC/MS) would provide complementary chemical composition information, and quantification of mass of different microplastic polymer types, with better reproducibility than optical methods. The mass spectrometric determination allows microplastic quantities to be presented in traditional concentration units (mass per unit volume), which facilitates more accurate comparisons of pollution loading among polymer types. These data help with assessing source apportionment trends and allow better modeling of chemical sorption using known polymer chemical sorption kinetics.

The USGS could develop, validate, and standardize methods for studying microplastics, and it could devise best practices for analyses for freshwater, wastewater, sludge, sediment, biological samples, and soil media (table 5). One capability for the USGS is the monitoring of microplastics (<300 µm) contributing to the aquatic load and potential effects on freshwater systems, which requires tools that are fast, easily implemented, and capable of detecting small microplastics. Plastic particles are known to undergo continuous fragmentation when exposed to environmental stressors (photodegradation, thermal degradation, and biodegradation), so the number of particles is expected to increase steeply for smaller sizes (Cózar and others, 2014; Andrady, 2017). Despite smaller microplastics making up the largest relative percentage of the aquatic load, these smaller fractions of <300 µm, and especially those <20 µm, they are consistently neglected in most quantification studies because large mesh neuston nets and low resolution FTIR spectroscopy (or optical microscopes) are commonly used for particle counting (Araújo and others, 2018; Conkle and others, 2018; Enders and others, 2015). Raman microscopy capabilities would offer the best opportunity for the USGS to provide accurate quantification of aquatic load in the full microplastics size range from 1 to 5 mm, along with the identification of the polymer type and associated source materials and locations (Käppler and others, 2016; Elert and others, 2017; Ivleva and others, 2017). This choice primarily stems from micro-Raman being able to identify smaller microplastics (<20 µm) that are undetectable by micro-FTIR techniques, along with its known capabilities for improving sample analysis times through automated particle counting and machine learning techniques.

Additional opportunities exist for development of in situ and real-time quantitation techniques for microplastics and (or) their accumulated toxic chemicals in water and sediment using surrogate (proxy) measurements or advancing portable technologies. Important opportunities exist due to recent technological advances in portable measuring devices with X-ray fluorescence, optical laser sensors, and microparticle-electrode sensors that can be coupled with machine learning automation in the future (Shimizu and others, 2017; Turner, 2017; Asamoah and others, 2019; Peiponen and others, 2020). The USGS is uniquely positioned with expertise in these fields to develop and implement advancements to improve microplastic field capabilities.