Microplastics in Freshwater Ecosystems
Microplastics have been found in nearly all types of freshwater environments, including remote lakes and rivers. Although all types of microplastics have been reported in freshwater ecosystems, microfibers are typically the most common microplastic type, often accounting for more than 80% of all the plastic fragments recovered. Understanding of the sources, movement, and fate of microplastics in freshwater ecosystems is still an active area of research; however, wastewater treatment plants and stormwater runoff appear to be important conduits of microplastics to lakes and rivers. More research is required to determine the role of atmospheric fallout in loading microplastics to freshwater ecosystems. Field and laboratory techniques for sampling microplastics in freshwater environments closely follow protocols for marine systems, although the lower density of freshwater compared to salt water can alter results if certain plastic polymers sink in freshwater compared to salt water. Further research is required to increase our understanding of the sources, movement, and fate of microplastic in aquatic ecosystems and the potential impacts of microplastics on freshwater organisms. This research will greatly increase our understanding of the role of freshwaters in the global plastic cycle.
Plastic waste has become a worldwide problem, with discarded plastics evident throughout global aquatic environments. Global production of plastics has exceeded 280 million tons a year, with the vast majority of this production disposed of after use (Koelmans et al. 2014). Plastic waste is consistently found in aquatic environments throughout the globe including the ocean surface, deep-sea sediments, beaches, and freshwater lakes and rivers (Van Cauwenberghe et al. 2015; Vermaire et al. 2017). As most plastics are not biodegradable, they can continue to break up into smaller and smaller fragments without completely disappearing; in fact it is estimated that most plastics will take thousands of years to completely degrade (Crawford and Quinn 2017). As a consequence, tiny plastics, or microplastics (plastic pieces <5 mm in size; Masura et al. 2015), are now present in soils (Nizzetto et al. 2016), remote lakes (Free et al. 2014), remote rivers (Jiang et al. 2019), polar regions (Obbard 2018), and all environmental compartments in all continents, including aquatic environments (Horton et al. 2017). Within the marine environment, microplastics are now considered the primary constituents of marine debris (Cole et al. 2011).
The term “microplastics” was not formally adopted until after 2004 (Thompson 2005) even though small plastics have been observed in the marine environment since the 1970s (Carpenter and Smith 1972). Microplastics can be classed as either primary or secondary microplastics. Primary microplastics are specifically manufactured to be plastics of microscopic size. Secondary microplastics are the result of a breakup or fragmentation of larger plastics. The plastic breakup can be due to a combination of physical, chemical, or biological processes reducing the integrity of the plastic, leading it to fragment (Cole et al. 2011). Another example of secondary microplastics is microfibers that are shed from synthetic clothing. Currently, microfibers are one of the most common forms of microplastic pollution encountered in freshwater systems, including lakes and river systems (Vermaire et al. 2017).
Freshwater microplastic research has only recently gained momentum, with only about 4% of microplastic research related to freshwater (Wagner and Lambert 2018). Freshwater environments can act as a conduit of microplastics to global marine areas through rivers or can also act as a sink for microplastics, for example, in isolated freshwater lakes or lake and river sediments (Wagner and Lambert 2018). Additionally, freshwaters can play a role as transformers of plastic pollution modifying the shape, size, and texture of plastic pollution through physical, chemical, and biological processes. Once microplastics are released into freshwater systems, their fate is poorly understood; however, a portion would be transported to ocean environments by rivers, with some retained in sinks such as sediments (Free et al. 2014).
The objective of this chapter is to review the current state of knowledge on microplastic pollution of freshwater ecosystems including sources, transport, and fate of plastic pollution in addition to reviewing current field and laboratory sampling methodologies utilized to sample microplastics in freshwater environments.
Microplastics in River Systems
Research has suggested that microfibers do not tend to settle in river sediments as often as fragments (Leslie et al. 2017), but are rather entrained in the water column and more likely to be transported downstream. Nonetheless, some research has indicated the presence of microfibers in river sediments (Vermaire et al. 2017; Hurley and Nizzetto 2018) that have potentially precipitated out of the water column in lower-energy environments and/or as part of a conglomerate of material that has become more dense than water. River sediments also have been shown to contain a substantial amount of primary microplastics in more industrialized river systems (Vermaire et al. 2017). However, floods or high-flow events such as spring freshets can dislodge microplastics from sediments (Hurley and Nizzetto 2018) and entrain them in the water column, sending microplastics further downstream to settle again or to be transported to marine environments. In order to understand the occurrence, transport, and fate of microplastics in river ecosystems, there is a growing need to increase both the spatial and temporal resolution of sampling in these systems.
Microplastics in Lakes
Microplastics were first reported in freshwater lakes only in 2013 (Eriksen et al. 2013) highlighting the recent concern and focus of microplastic contamination in these systems. Research has focused on microplastic contamination in lakes of different sizes, dimensions, and proximities to urban areas, with all demonstrating striking amounts of microplastic pollution (Eriksen et al. 2013; Free et al. 2014; Driedger et al. 2015). Lakes have also been described as a “semi-closed” system for microplastics (Fischer et al. 2016) whereby microplastics may be contained in sinks in lakes (i.e., sediments) and/or circulated within the open water system or enter tributaries and exit the lake environment. The movement of microplastics within lake systems will depend on wind strength, wind direction, lake morphology, prevailing currents, and storm events. Additionally, a lake’s geographic position (latitude and longitude) can determine microplastic transport and deposition within a lake system, as seasonal temperatures and density stratification in the water can cause varying mixing and stagnation phases of microplastics in the water column (Boehrer and Schultze 2008).
The fate of microplastics in the surface waters of lakes depends on the polymer type, size, shape, and additional factors such as biofouling which could affect the buoyancy of the plastics. Additionally, as freshwater is not as dense as salt water, the critical density of a plastic to sink is slightly lower than in marine systems. For example, plastics prone to float in marine environments, namely, polyethylene, polypropylene, and polystyrene, have an increased chance to sink in freshwater environments (Ballent et al. 2016), especially if there are any biofouling mechanisms.
Similar to river systems, lakes can present large shorelines and thus exhibit some similarities to marine environments, as lake and river shores can provide the opportunity for the breakdown of larger plastics through mechanical weathering and photodegradation. In this way freshwater systems can also act as transformers of plastic pollution before it reaches marine systems. There are numerous studies examining the integration of plastics and microplastics in marine beaches, but little research has focused on microplastics on lake or river shores (Zbyszewski and Corcoran 2011). Though similarities have been drawn to marine beaches, weathering on lake or river shores can still differ due to factors such as the varying weathering rates in different water chemistries, i.e., salt versus freshwater (Biesinger et al. 2011). Microplastics have also been identified in bottom sediments of lakes (Corcoran et al. 2015), further indicating that lakes are sinks for microplastic pollution.
Similar to river systems, microfibers are also the dominant microplastic in lake environments (Anderson et al. 2017), highlighting the potential influence of urban areas and wastewater treatment effluent. However, research has suggested even remote lakes away from urban areas and anthropogenic influence exhibit microplastic contamination (Free et al. 2014). These findings suggest the role of atmospheric transport of microplastics (Cai et al. 2017; Dehghani et al. 2017) to remote aquatic systems or potentially trophic transfer by animals (Hurley and Nizzetto 2018; Provencher et al. 2018). Though research into microplastics in lakes and freshwater environments is somewhat limited, there is growing evidence of interaction (e.g., microplastic ingestion and adhesion to organisms) with benthic environments in lake (Driedger et al. 2015) and river systems (Windsor et al. 2019), suggesting bioaccumulation is also a potential mechanism of microplastic pollution in both marine and freshwater systems.
Sources of Microplastics to Freshwater Ecosystems
Wastewater Treatment Plants
The wastewater treatment process is very efficient in removing microplastics received from industrial and domestic wastewater. Removal efficiency is as high as 99% in tertiary-treated effluent (Talvitie et al. 2017); however, globally many municipalities only have primary or secondary wastewater treatment (or none at all), in which case microplastic removal efficiency will be greatly reduced (Mintenig et al. 2017). However, given the amount of water, WWTP processes, and the huge amounts of plastic pollution entering our wastewater systems, even high efficiency rates can still amount to a substantial deposition of microplastics to receiving waters (Leslie et al. 2017; Talvitie et al. 2017). For example, in the USA, 17 WWTP effluent streams were tested (Mason et al. 2016) noting 0.05 (+/− 0.024) microplastics per liter of effluent. Another study in San Francisco noted a slightly higher 0.086 microplastics per liter of effluent (Sutton et al. 2016). These figures may appear low; however, these concentrations are orders of magnitude above the concentration of microplastics in most lakes and rivers, and most treatment facilities process millions of liters of wastewater a day resulting in estimated microplastic discharges from WWTPs in the order of millions of microplastic particles per day (Mason et al. 2016).
It is also noted the level of treatment can affect the amount of microplastics released to the aquatic environment, for example, tertiary or advanced wastewater treatments are typically more efficient in microplastic removal. This can be demonstrated by research in Lake Ontario and Lake Erie, whereby it has been established only 66% of New York State WWTPs utilize advanced treatment and microbeads or (micro)pellets were identified in six out of seven WWTP effluent streams discharging to the two great lakes (Nalbone 2014). Additionally, the presence of combined sewer overflows can significantly load lake environments of microplastics during storm or overflow events, similar to river systems.
Microplastic conveyance through WWTP can also depend on the adoption of a combined or separate sewer system. A combined sewer system conveys both sewage and stormwater to a WWTP. During normal conditions, all of this influent gets treated to a set standard, and the effluent is discharged to a receiving water body. However, during high flow or storm events, a combined sewage overflow may redirect excess flow away from the WWTP directly to a receiving body of water. The typical 99% removal rate (in the case of most tertiary treatments) is now bypassed potentially resulting to a substantial microplastic load into a receiving body of water. The other construction option adopted by urban municipalities is the implementation of a separate system, where sewage and stormwater are separated and stormwater flows directly to an aquatic environment. From a microplastic pollution perspective, the separate sewer system may also not be ideal, as any stormwater flow is not processed for microplastics; thus any microplastic load that may have been removed by a WWTP within a combined system is now conveyed directly into a river, lake, or ocean. Stormwater runoff entering aquatic environments directly can contain microplastics from tires (Dehghani et al. 2017) and can collect plastic debris where these macroplastics can then begin to break down, depending on the environment it ends in. Further research on microplastic loading to lakes and rivers following large rainfall events will help quantify the role of storm events in loading microplastic pollution to freshwater ecosystems.
Stormwater is an additional identified conduit for microplastics into rivers. The types of microplastics in stormwater can vary to a greater extent than WWTPs, due to an increase in potential microplastic sources, for example, vehicle abrasion from tire treads (Dehghani et al. 2017; Horton et al. 2017). Stormwater does add an important temporal analysis component to research with outflow occurring in rain and/or melt events. For example, a Los Angeles study suggested that up to three times the plastic concentrations in rivers were measured after wet events, leading to runoff as a major contributor in transporting plastics to river systems (Moore et al. 2011). Additionally, urban creeks or waterways can be a collection point for stormwater runoff during precipitation events.
Remote locations, away from any anthropogenic influences, are also prone to microplastic contamination (Free et al. 2014; Jiang et al. 2019). There are suggestions that microplastics, more specifically microfibers, are capable of being deposited into freshwater environments through atmospheric deposition (Cai et al. 2017; Dehghani et al. 2017; Dris et al. 2017). Most of these microfibers are likely from natural textiles such as modified cotton or wool; however, some of them are microplastics (Stanton et al. 2019). Some research also suggests other microplastic types besides microfibers can also be transported by the atmosphere including fragments (Cai et al. 2017) and even fragments from car tires (Kole et al. 2017). Recent research suggests microplastic concentrations of 88 to 605 microplastics per 30 g of urban dust (Dehghani et al. 2017) with potential atmospheric fallout rates between 2 and 355 fibers per m3 per day.
Sampling for Microplastics in Freshwater Environments
Since microplastic research is still relatively new field, there is a general lack of harmonization and standardization of methodologies for sample collection and analysis. For years, many researchers have been urging for a harmonization and/or standardization of methods (Van Cauwenberghe et al. 2015; Koelmans et al. 2019), particularly since without harmonization and/or standardization, comparison of data between studies is challenging. Formalizing field and laboratory techniques in both marine and freshwater environments will greatly aid the comparison of results among studies.
Many methodologies used in freshwater studies follow those used in the marine environments, and just like research conducted on microplastics in the marine environment, the ways in which samples are collected and analyzed in freshwater environments vary between studies. Although there are several different sampling methodologies for collecting water and sediment samples from freshwater environments, some methods appear to be more commonly employed than others.
The most common technique used to collect water samples in rivers or lakes is net tows (either Manta or Neuston) (Eriksen et al. 2013; Free et al. 2014; Fischer et al. 2016; Anderson et al. 2017; Vermaire et al. 2017), which are also commonly used to sample marine waters (Eriksen et al. 2014) and allow for the filtration of large volumes of water. Collecting grab samples either for in situ filtration or for laboratory filtration has also been used (Leslie et al. 2017; Forrest et al. 2019; Jiang et al. 2019); however, an important limitation of this method is the low volume of water that can be filtered compared to using a net, which may increase variability between samples (Barrows et al. 2017). Although most studies use one of these two methods to collect water samples, inconsistencies are still present within both methods. Specifically, the lower size limit being examined can vary greatly (i.e., the filter or sieve size), and the volume of water that is filtered varies between studies or in some cases is not indicated in the literature.
Similar to water samples, methods used to collect sediment samples in freshwater environments follow the methods also used in marine environments. The two main methods used to collect sediments in lakes and rivers are through collecting surface sediment samples along shorelines (Ballent et al. 2016; Fischer et al. 2016; Jiang et al. 2019) or by collecting benthic sediment samples offshore using grab samplers or corers (Ballent et al. 2016; Leslie et al. 2017; Vermaire et al. 2017).
Laboratory processing of freshwater water and sediment samples also tends to follow similar processing methods used in the marine environment (Masura et al. 2015), and just like sampling methods, there are many ways in which samples are processed. Water samples are often filtered and rinsed through sieves or filters before undergoing chemical digestion to get rid of any organic material from the sample. Next, particles are usually picked, counted, and categorized from the samples using stereomicroscopy.
Sediment samples are typically first dried and weighed and then undergo steps for size fractioning and/or density separation. In some cases, chemical digestion is used to remove organic material. Although most studies perform a density separation, the solutions used for this step vary greatly in the literature (Quinn et al. 2017). One of the most common solutions for density separation in the literature is sodium chloride (NaCl) (Leslie et al. 2017; Vermaire et al. 2017), while others use various different heavy liquids to obtain higher solution density than can be achieved using NaCl, such as sodium polytungstate (Ballent et al. 2016) or zinc chloride (Jiang et al. 2019).
Initial identification of microplastic particles typically involves visual classification of plastic particles using a stereomicroscope (Barrows et al. 2017; Vermaire et al. 2017), followed by polymer identification for a subset of particles. The step of identifying and chemically characterizing particles is increasingly thought to be crucial to make sure that particles are in fact microplastics, especially as more studies are now including smaller-sized particles. The most common methods for particle identification and chemical characterization are Fourier-transform infrared spectroscopy (FTIR) (Zbyszewski and Corcoran 2011; Mintenig et al. 2017) and Raman spectroscopy (Free et al. 2014). However, other methods of identification have been published, such as the use of scanning electron microscopes (SEM) in the case of polyvinyl chloride (Anderson et al. 2017) and Nile red dye (Hengstmann and Fischer 2019). To improve the accessibility of identifying microplastics, spectral libraries specifically geared toward plastic particles have recently been developed (Munno et al. 2020).
Contamination and QA/QC
Microfiber contamination in both the field and the laboratory should also be considered while collecting and processing samples, with necessary steps taken to reduce contamination potential. Atmospheric transport and deposition of microplastics are possible throughout even in the most remote locations (Free et al. 2014; Obbard 2018; Jiang et al. 2019). This highlights the importance of conducting field controls to allow for potential contamination while sampling and also collecting sufficiently large samples to distinguish an environmental signal from potential contamination, especially as one or two microfibers from contamination can greatly skew microplastic estimates in lower-volume testing.
Indoor microfiber contamination is also important to consider, with suggested indoor fallout rates of 0.4 to 59.4 microfibers per cubic meter (Dris et al. 2017). This emphasizes the importance of reducing contamination potential wile processing samples in the laboratory. Laboratory coats, made of natural microfibers if possible, and/or bright distinguishable colors should be worn at all times. If laboratory coats shed, the bright colors can be identified as contamination in samples, and the natural microfibers can be removed during digestion protocols or can be discounted under Raman or FTIR analysis. All petri dishes, mesh, and containers (e.g., beakers) should be triple rinsed with deionized water to remove any settled fiber contamination throughout sample processing. If samples are idle at any time during processing, they should be covered to reduce and placed in fume hoods or environments where the air is extracted, thus removing airborne microfibers away from samples. Furthermore, all sample processing where possible should be conducted under a laminar flow hood or even better a clean room to reduce the settling potential of airborne microfibers. Microfiber contamination highlights the importance of conducting controls and blank samples throughout sampling and processing, as it establishes the detection limit and potential contamination limit of the samples.
Finally, while there should be an acknowledgment of contamination potential through controls, consideration should also be given to positive controls. During collection and processing of samples, loss of microplastics is possible. For example, various digestion protocols have been suggested to consume some polymers; thus, consideration should be given to which digestion protocol is selected, or a positive control should be applied to establish potential microplastic loss. Additionally, during visual identification stages, especially in samples with numerous microplastics, there is potential to miss microplastics in samples; therefore, utilizing positive controls is a useful method to establish the potential loss of microplastics during sample processing. To date, very few studies of microplastics have employed positive controls to assess the potential loss of microplastics through processing and sorting (Koelmans et al. 2019).
Impacts of Microplastics on Freshwater Ecosystems
Ingestion of microplastics can potentially cause chemical and physical harm to freshwater organisms (Auta et al. 2017). However, research to date on the potential impacts of microplastic pollution on organisms has largely focused on marine biota. In many cases, the potential physical and chemical harm due to microplastic ingestion (or attachment of microplastics to biota surfaces) on marine organisms is assumed to apply to similar freshwater organisms. Nonetheless, some emerging research is demonstrating similar physical effects of microplastics between marine and freshwater species (Fossi et al. 2014; Oliveira et al. 2018).
Bioaccumulation of Microplastics in Freshwater
Microplastic interaction with biota can be somewhat dependent on microplastic shape, size, color, aggregation, and abundance and will affect their potential bioavailability (Van Cauwenberghe et al. 2015). The smaller sizes of microplastics increase their bioavailability to a wider array of organisms, especially indiscriminate feeders. Microplastics have the potential to sorb (and/or release) toxic chemicals, with the potential transfer through the food chain (Hurley et al. 2017). Previous research has highlighted microplastics as a vector due to the sorption of waterborne pollutants, from invertebrates to higher trophic levels (Teuten et al. 2009). The accumulation of these small plastics in organisms can potentially cause adverse physical effects in addition to potential adverse chemical effects. Organisms can accumulate microplastics in their tissues serving as a vehicle of pathogens, absorbing and accumulating toxins (Auta et al. 2017). Organisms have also been found to metabolize persistent organic pollutants (POPs) from trophic surfaces of microplastics (Chua et al. 2014).
Microplastics can contain organic pollutants added during plastic production, or their relatively large surface area-to-volume ratio makes microplastics liable to contamination from waterborne contaminates such as POPs and metals (Cole et al. 2011). POPs prone to sorption by microplastics include dichlorodiphenyl (DDT), polycyclic aromatic hydrocarbons (PAH), and polychlorinated biphenyls (PCB) with potential metal sorption including copper, silver, zinc, lead iron, manganese, and mercury (Hirai et al. 2011; Chua et al. 2014; Auta et al. 2017; Oliveira et al. 2018). Some of these chemicals can be found in high quantities in aquatic environments, especially on the surface, where low-density microplastics can be present in large numbers (Teuten et al. 2009). Furthermore, plastics can sorb contaminants from the surrounding environment up to 100 times more than sediments, and this sorption can include organic chemicals that are persistent, bioaccumulative, and toxic (Rochman et al. 2013).
Moreover, chemical ingredients present in plastics sorbed from the environment can include 78% of chemicals listed as priority pollutants as they have been identified to be bioaccumulative, persistent, and/or toxic (Rochman et al. 2013), with more than 50% of the plastic polymers produced including chemical ingredients considered hazardous by the UN’s Globally Harmonized System (Lithner et al. 2011). Nonetheless, it is important to also consider the plastic or polymer itself as the type of polymer can influence the amount of sorption of organic pollutants (Rusina et al. 2007). For example, it has been suggested polyethylene can accumulate more organic pollutants than polyvinylchloride (Teuten et al. 2009).
The ingestion of these small plastics with toxins by organisms at the base of the food chain highlights the potential for bioaccumulations (Teuten et al. 2009). However, the research on the joint toxicology of microplastics and POPs is still limited (Mattsson et al. 2015), and less attention has been given to the chemical effects associated with the ingestion or bioconcentration of plastic debris, even with evidence growing of a wide range of ingestion by numerous species (Rochman et al. 2013).
Nonetheless, some researchers suggest there is no significant connection and thus importance of the cycling and bioaccumulation of organic pollutants or hydrophobic organic contaminants (HOCs) with microplastics (Lohmann 2017). POPs are a subset of persistent HOCs, and there is evidence that suggests that microplastics can and do absorb high concentrations of organic pollutants, though the significance of the sorption and transfer of organic contaminants has been suggested to be relatively low (Koelmans et al. 2014). One suggestion is that organisms can uptake these contaminants already from the water, sediment, and food and the added uptake from plastics does not cause any substantial increase in these chemicals for the organism (Koelmans et al. 2014). Furthermore, it is suggested that there are simply not enough microplastics in the (marine and likely freshwater) environment to outcompete the partitioning of POPs to water and natural organic matter (Lohmann 2017). However, this is not to suggest that the indiscriminate dumping of plastic should continue; rather, a critical limit of microplastic pollution may not have been reached yet in most environments but could become problematic in the future or at high-density locations.
Another chemical factor of microplastics to consider are the additives used in the manufacturing of polymers and plastics. As plastics have been suggested to absorb POPs and potentially released after ingestion, are plastic additives also released after ingestion? Again, this is a sparsely researched topic in freshwater ecosystems, and there needs to be more attention given to phenolic additive-derived chemicals from microplastics in the food web (Teuten et al. 2009). That said, some marine studies indicate there is no relevance of these chemicals in the uptake of organisms (Koelmans et al. 2014), further suggesting that more research is needed on the interaction of microplastics, chemicals, pollutants, and biota, particularly in freshwater environments where data is lacking.
Physical Effect of Microplastic Ingestion by Organisms in Freshwater
The chemical effects of microplastics are still relatively unknown, particularly in freshwater environments. There is, however, more research into the potential physical damage of microplastics in aquatic organisms although again the focus is primarily on marine organisms and researchers are left to draw parallels to freshwater organisms.
Adverse physical reactions in organisms to microplastic ingestion include oxidative stress (Fossi et al. 2014; Oliveira et al. 2018), decreased feeding rates (Cole et al. 2013), increased mortality rates (Nelms et al. 2019), weight loss (Besseling et al. 2013), fitness reduction (Besseling et al. 2013), digestive tract blockages (Galgani et al. 2010), inflammatory responses (von Moos et al. 2012), and neurotoxicity (Oliveira et al. 2018). There is however a serious research needed to better understand how microplastics may be impacting freshwater organisms specifically as the physical, chemical, and biological conditions of marine environments are vastly different from freshwater ecosystems.
Microplastics are now considered widespread throughout all environmental compartments, including freshwater ecosystems. Society continues to use plastic as it is durable, lightweight, and relatively inexpensive. Furthermore, the lack of waste management and indiscriminate disposal have contributed to an exponential release of plastics into the environment where they have been identified as a global concern to aquatic life. Plastics start from anthropogenic sources with microplastics contaminating aquatic environments though various pathways, including wastewater treatment plants, storm runoff, atmospheric deposition, and littering near shorelines. Some of these microplastics may return to terrestrial environments in wastewater sludge, for example, but they can still be transported back to aquatic environments though land runoff and atmospheric deposition.
Though research on freshwater microplastic contamination is still somewhat limited, freshwater environments have still been identified as important conduits, sinks, and transformers of microplastic pollution. Lakes can accumulate microplastics on lakeshores, in bottom sediment, and in the water column where they can feed tributaries, transporting microplastics out of lake environments. Rives can act as a major collector and conduit of microplastics to the marine environment, with potential adverse effects to freshwater organisms in both systems. In many ways microplastics research is in the early stages in freshwater ecosystems as scientists try and move past simply identifying the presence of microplastics in different lakes and rivers to better understand the sources, movement, and fate of microplastic in freshwater systems to better model the global plastic cycle. In addition, research on the impact of microplastics on freshwater organisms is sorely needed to better understand what types of microplastics at what concentration and under what environmental conditions negatively affect what organisms. This research is essential if we are to set management objectives for microplastic contamination in freshwater systems. The overwhelming consensus is the necessity for additional research, especially in freshwater environments, to better comprehend microplastics and their interaction with freshwater ecosystems.
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