Advertisement

Microplastics in Freshwater Ecosystems

  • Shaun A. Forrest
  • Madelaine P. T. Bourdages
  • Jesse C. VermaireEmail author
Living reference work entry

Abstract

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.

Introduction

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

Rivers have been identified as a major source of plastic pollution reaching the world’s oceans (Lebreton et al. 2017); however, research on microplastic contamination in rivers is still gaining momentum with only a limited understanding of the mechanisms of what controls transport, dispersion, and fate of microplastics in river systems (Vermaire et al. 2017; Wagner and Lambert 2018). What is clear is microplastic contamination is prevalent in many river systems around the world, including in remote or rural regions of river systems (Forrest et al. 2019) (Fig. 1). Plastics and microplastics originating from terrestrial or atmospheric sources released into freshwater environments may be transported to the oceans via rivers but could also be retained in sinks within freshwater environments, including river sediments or isolated lakes (Free et al. 2014; Jiang et al. 2019). Microplastic entrainment and deposition in rivers can be attributed to factors including (but not limited to) upstream land use (Mani et al. 2016), urban runoff (Nizzetto et al. 2016), relative effluent discharge volumes (Moore et al. 2011), river hydraulics and river morphology (Besseling et al. 2017), and population density and the amount of urbanization and industrialization surrounding a river (Yonkos et al. 2014). In addition, barriers such as dams and weirs also influence the concentration, composition, and transport of microplastics in freshwater ecosystems (Zhang et al. 2015). Other important factors to consider are related to the polymer itself, size, shape, density, and if it has any organic and/or inorganic materials, including biofouling attached to the plastics (Fig. 2). Environmental factors that could affect microplastic distribution in rivers include winds, currents, waves, and water density (Prata et al. 2019). In short, microplastic concentration in river systems can be highly variable both spatially and temporally (e.g., high flows vs low flows), and researchers are at the early stages of understanding these mechanisms.
Fig. 1

Example of microplastics, microfibers, and fragments, extracted from the Trent-Severn waterway in Ontario, Canada

Fig. 2

Microplastics from the Ottawa River often contain biofilms with diatoms and other algae growing on them. Scanning electron microscope image of a microplastic fragment with the diatom Cocconeis placentula Ehrenberg growing attached to the plastic fragment

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

To further understand microplastic contamination in freshwater ecosystems, it is imperative to understand potential inputs of microplastics to freshwater environments. There are multiple pathways for microplastics to enter freshwater ecosystems including wastewater and stormwater, atmospheric deposition, and the breakup of larger plastics within rivers, lakes, and shorelines (Fig. 3). In particular, previous research has identified wastewater treatment plants and stormwater flow (including overland runoff) as a potentially important source of microplastics to freshwater environments (Horton et al. 2017; Vermaire et al. 2017; Windsor et al. 2019). This is not to disregard the potential of other inputs, for example, atmospheric fallout; however, to date, very little research exists on the influence of atmospheric deposition on microplastic concentration or composition in freshwater ecosystems.
Fig. 3

The movement of microplastics (primary and secondary) throughout freshwater environments. Arrows demonstrate the fate of microplastics within lakes and rivers and the input of microplastics from atmospheric and terrestrial sources including urban and agricultural land use. The diagram highlights the numerous interactions throughout freshwater environments, including ingestion by freshwater organisms (represented by the dotted line arrows) with a final output of the majority of these microplastics (microplastic sinks being the exception) to the marine environment, whereby rivers are one of the major sources to these environments. Further research is required to provide estimates of the magnitude of these pathways and to better understand the global plastic cycle. Our goal as scientists should be to better quantify these pathways so that we can provide an understanding of the global plastic cycle

Wastewater Treatment Plants

One of the main conduits of microplastics to freshwater environments has been identified as wastewater treatment plants (WWTP; Fig. 4). This is especially true in urban rivers, where effluent disposal can expel primary and secondary microplastics into river environments (Leslie et al. 2017; Talvitie et al. 2017). Additionally, wastewater sludge captures a substantial amount of microplastics and in many cases is returned to the land as soil conditioner (Nizzetto et al. 2016), creating the potential for the microplastics within the sludge to enter freshwater environments through runoff. For example, the European Union has estimated between 63,000 and 430,000 metric tons of microplastics within sewage sludge are deposited annually on European agricultural lands (Nizzetto et al. 2016).
Fig. 4

Microplastics extracted from the Gatineau wastewater treatment plant in Quebec, Canada, showing a high concentration of microfibers

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

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.

Atmospheric Deposition

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.

Conclusions

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.

References

  1. Anderson PJ, Warrack S, Langen V, Challis JK, Hanson ML, Rennie MD (2017) Microplastic contamination in Lake Winnipeg, Canada. Environ Pollut 225:223–231.  https://doi.org/10.1016/j.envpol.2017.02.072CrossRefGoogle Scholar
  2. Auta HS, Emenike CU, Fauziah SH (2017) Distribution and importance of microplastics in the marine environment: a review of the sources, fate, effects, and potential solutions. Environ Int 102:165–176.  https://doi.org/10.1016/j.envint.2017.02.013CrossRefGoogle Scholar
  3. Ballent A, Corcoran PL, Madden O, Helm PA, Longstaffe FJ (2016) Sources and sinks of microplastics in Canadian Lake Ontario nearshore, tributary and beach sediments. Mar Pollut Bull 110:383–395.  https://doi.org/10.1016/j.marpolbul.2016.06.037CrossRefGoogle Scholar
  4. Barrows APW, Neumann CA, Pieper C, Berger ML, Shaw SD (2017) Guide to microplastic identification, a comprehensive methods guide for microplastics identification and quantification in the laboratory. Marine & Environmental Research Institute, Blue HillGoogle Scholar
  5. Besseling E, Wegner A, Foekema EM, van den Heuvel-Greve MJ, Koelmans AA (2013) Effects of microplastic on fitness and PCB bioaccumulation by the Lugworm Arenicola marina (L.). Environ Sci Technol 47:593–600.  https://doi.org/10.1021/es302763xCrossRefGoogle Scholar
  6. Besseling E, Quik JTK, Sun M, Koelmans AA (2017) Fate of nano- and microplastic in freshwater systems: a modeling study. Environ Pollut 220:540–548.  https://doi.org/10.1016/j.envpol.2016.10.001CrossRefGoogle Scholar
  7. Biesinger MC, Corcoran PL, Walzak MJ (2011) Developing ToF-SIMS methods for investigating the degradation of plastic debris on beaches. Surf Interface Anal 43:443–445.  https://doi.org/10.1002/sia.3397CrossRefGoogle Scholar
  8. Boehrer B, Schultze M (2008) Stratification of lakes. Rev Geophys 46:RG2005.  https://doi.org/10.1029/2006RG000210CrossRefGoogle Scholar
  9. Cai L, Wang J, Peng J, Tan Z, Zhan Z, Tan X, Chen Q (2017) Characteristic of microplastics in the atmospheric fallout from Dongguan city, China: preliminary research and first evidence. Environ Sci Pollut Res 24:24928–24935.  https://doi.org/10.1007/s11356-017-0116-xCrossRefGoogle Scholar
  10. Carpenter EJ, Smith KL (1972) Plastics on the Sargasso Sea surface. Science 175:1240–1241.  https://doi.org/10.1126/science.175.4027.1240CrossRefGoogle Scholar
  11. Chua EM, Shimeta J, Nugegoda D, Morrison PD, Clarke BO (2014) Assimilation of polybrominated diphenyl ethers from microplastics by the marine amphipod, Allorchestes Compressa. Environ Sci Technol 48:8127–8134.  https://doi.org/10.1021/es405717zCrossRefGoogle Scholar
  12. Cole M, Lindeque P, Halsband C, Galloway TS (2011) Microplastics as contaminants in the marine environment: a review. Mar Pollut Bull 62:2588–2597.  https://doi.org/10.1016/j.marpolbul.2011.09.025CrossRefGoogle Scholar
  13. Cole M, Lindeque P, Fileman E, Halsband C, Goodhead R, Moger J, Galloway TS (2013) Microplastic ingestion by zooplankton. Environ Sci Technol 47:6646–6655.  https://doi.org/10.1021/es400663fCrossRefGoogle Scholar
  14. Corcoran PL, Norris T, Ceccanese T, Walzak MJ, Helm PA, Marvin CH (2015) Hidden plastics of Lake Ontario, Canada and their potential preservation in the sediment record. Environ Pollut 204:17–25.  https://doi.org/10.1016/j.envpol.2015.04.009CrossRefGoogle Scholar
  15. Crawford CB, Quinn B (2017) Microplastic pollutants. Elsevier, AmsterdamGoogle Scholar
  16. Dehghani S, Moore F, Akhbarizadeh R (2017) Microplastic pollution in deposited urban dust, Tehran metropolis, Iran. Environ Sci Pollut Res 24:20360–20371.  https://doi.org/10.1007/s11356-017-9674-1CrossRefGoogle Scholar
  17. Driedger AGJ, Dürr HH, Mitchell K, Van Cappellen P (2015) Plastic debris in the Laurentian Great Lakes: a review. J Great Lakes Res 41:9–19.  https://doi.org/10.1016/j.jglr.2014.12.020CrossRefGoogle Scholar
  18. Dris R, Gasperi J, Mirande C, Mandin C, Guerrouache M, Langlois V, Tassin B (2017) A first overview of textile fibers, including microplastics, in indoor and outdoor environments. Environ Pollut 221:453–458.  https://doi.org/10.1016/j.envpol.2016.12.013CrossRefGoogle Scholar
  19. Eriksen M, Mason S, Wilson S, Box C, Zellers A, Edwards W, Farley H, Amato S (2013) Microplastic pollution in the surface waters of the Laurentian Great Lakes. Mar Pollut Bull 77:177–182.  https://doi.org/10.1016/j.marpolbul.2013.10.007CrossRefGoogle Scholar
  20. Eriksen M, Lebreton LCM, Carson HS, Thiel M, Moore CJ, Borerro JC, Galgani F, Ryan PG, Reisser J, Dam HG (2014) Plastic Pollution in the World’s Oceans: More than 5 Trillion Plastic Pieces Weighing over 250,000 Tons Afloat at Sea. PLoS ONE 9(12):e111913Google Scholar
  21. Fischer EK, Paglialonga L, Czech E, Tamminga M (2016) Microplastic pollution in lakes and lake shoreline sediments – A case study on Lake Bolsena and Lake Chiusi (central Italy). Environ Pollut 213:648–657Google Scholar
  22. Forrest SA, Holman L, Murphy M, Vermaire JC (2019) Citizen science sampling programs as a technique for monitoring microplastic pollution: results, lessons learned and recommendations for working with volunteers for monitoring plastic pollution in freshwater ecosystems. Environ Monit Assess 191:172.  https://doi.org/10.1007/s10661-019-7297-3CrossRefGoogle Scholar
  23. Fossi MC, Coppola D, Baini M, Giannetti M, Guerranti C, Marsili L, Panti C, de Sabata E, Clò S (2014) Large filter feeding marine organisms as indicators of microplastic in the pelagic environment: the case studies of the Mediterranean basking shark (Cetorhinus maximus) and fin whale (Balaenoptera physalus). Mar Environ Res 100:17–24.  https://doi.org/10.1016/j.marenvres.2014.02.002CrossRefGoogle Scholar
  24. Free CM, Jensen OP, Mason SA, Eriksen M, Williamson NJ, Boldgiv B (2014) High-levels of microplastic pollution in a large, remote, mountain lake. Mar Pollut Bull 85:156–163.  https://doi.org/10.1016/j.marpolbul.2014.06.001CrossRefGoogle Scholar
  25. Galgani F, Fleet D, Van Franeker JA, Katsanevakis S, Maes T, Mouat J, Oosterbaan L, Poitou I, Hanke G, Thompson R, Amato E, Birkun A, Janssen C (2010) Marine strategy framework directive: task group 10 report marine litter. Office for Official Publications of the European Communities, LuxembourgGoogle Scholar
  26. Hengstmann E, Fischer EK (2019) Nile red staining in microplastic analysis—proposal for a reliable and fast identification approach for large microplastics. Environ Monit Assess 191.  https://doi.org/10.1007/s10661-019-7786-4
  27. Hirai H, Takada H, Ogata Y, Yamashita R, Mizukawa K, Saha M, Kwan C, Moore C, Gray H, Laursen D, Zettler ER, Farrington JW, Reddy CM, Peacock EE, Ward MW (2011) Organic micropollutants in marine plastics debris from the open ocean and remote and urban beaches. Mar Pollut Bull 62(8):1683–1692Google Scholar
  28. Horton AA, Svendsen C, Williams RJ, Spurgeon DJ, Lahive E (2017) Large microplastic particles in sediments of tributaries of the River Thames, UK – abundance, sources and methods for effective quantification. Mar Pollut Bull 114:218–226.  https://doi.org/10.1016/j.marpolbul.2016.09.004CrossRefGoogle Scholar
  29. Hurley RR, Nizzetto L (2018) Fate and occurrence of micro(nano)plastics in soils: knowledge gaps and possible risks. Current Opinion in Environmental Science & Health 1:6–11.  https://doi.org/10.1016/j.coesh.2017.10.006CrossRefGoogle Scholar
  30. Hurley RR, Woodward JC, Rothwell JJ (2017) Ingestion of microplastics by freshwater Tubifex worms. Environ Sci Technol 51:12844–12851.  https://doi.org/10.1021/acs.est.7b03567CrossRefGoogle Scholar
  31. Jiang C, Yin L, Li Z, Wen X, Luo X, Hu S, Yang H, Long Y, Deng B, Huang L, Liu Y (2019) Microplastic pollution in the rivers of the Tibet Plateau. Environ Pollut 249:91–98.  https://doi.org/10.1016/j.envpol.2019.03.022CrossRefGoogle Scholar
  32. Koelmans AA, Gouin T, Thompson R, Wallace N, Arthur C (2014) Plastics in the marine environment: ET&C perspectives. Environ Toxicol Chem 33:5–10.  https://doi.org/10.1002/etc.2426CrossRefGoogle Scholar
  33. Koelmans AA, Mohamed Nor NH, Hermsen E, Kooi M, Mintenig SM, De France J (2019) Microplastics in freshwaters and drinking water: critical review and assessment of data quality. Water Res 155:410–422.  https://doi.org/10.1016/j.watres.2019.02.054CrossRefGoogle Scholar
  34. Kole PJ, Löhr AJ, Van Belleghem F, Ragas A (2017) Wear and tear of tyres: a stealthy source of microplastics in the environment. IJERPH 14:1265.  https://doi.org/10.3390/ijerph14101265CrossRefGoogle Scholar
  35. Lebreton LCM, van der Zwet J, Damsteeg J-W, Slat B, Andrady A, Reisser J (2017) River plastic emissions to the world’s oceans. Nat Commun 8:15611.  https://doi.org/10.1038/ncomms15611CrossRefGoogle Scholar
  36. Leslie HA, Brandsma SH, van Velzen MJM, Vethaak AD (2017) Microplastics en route: field measurements in the Dutch river delta and Amsterdam canals, wastewater treatment plants, North Sea sediments and biota. Environ Int 101:133–142.  https://doi.org/10.1016/j.envint.2017.01.018CrossRefGoogle Scholar
  37. Lithner D, Larsson Å, Dave G (2011) Environmental and health hazard ranking and assessment of plastic polymers based on chemical composition. Sci Total Environ 409:3309–3324.  https://doi.org/10.1016/j.scitotenv.2011.04.038CrossRefGoogle Scholar
  38. Lohmann R (2017) Microplastics are not important for the cycling and bioaccumulation of organic pollutants in the oceans-but should microplastics be considered POPs themselves?: Should microplastics be considered POPs. Integr Environ Assess Manag 13:460–465.  https://doi.org/10.1002/ieam.1914CrossRefGoogle Scholar
  39. Mani T, Hauk A, Walter U, Burkhardt-Holm P (2016) Microplastics profile along the Rhine River. Sci Rep 5:17988.  https://doi.org/10.1038/srep17988CrossRefGoogle Scholar
  40. Mason SA, Garneau D, Sutton R, Chu Y, Ehmann K, Barnes J, Fink P, Papazissimos D, Rogers DL (2016) Microplastic pollution is widely detected in US municipal wastewater treatment plant effluent. Environ Pollut 218:1045–1054.  https://doi.org/10.1016/j.envpol.2016.08.056CrossRefGoogle Scholar
  41. Masura J, Baker J, Foster G, Arthur C (2015) Laboratory methods for the analysis of microplastics in the marine environment: recommendations for quantifying synthetic particles in waters and sediments.  https://doi.org/10.25607/OBP-604
  42. Mattsson K, Hansson L-A, Cedervall T (2015) Nano-plastics in the aquatic environment. Environ Sci Process Impacts 17:1712–1721.  https://doi.org/10.1039/C5EM00227CCrossRefGoogle Scholar
  43. Mintenig SM, Int-Veen I, Löder MGJ, Primpke S, Gerdts G (2017) Identification of microplastic in effluents of waste water treatment plants using focal plane array-based micro-Fourier-transform infrared imaging. Water Res 108:365–372.  https://doi.org/10.1016/j.watres.2016.11.015CrossRefGoogle Scholar
  44. Moore CJ, Lattin GL, Zellers AF (2011) Quantity and type of plastic debris flowing from two urban rivers to coastal waters and beaches of Southern California. RGCI 11:65–73.  https://doi.org/10.5894/rgci194CrossRefGoogle Scholar
  45. Munno K, De Frond H, O’Donnell B, Rochman CM (2020) Increasing the accessibility for characterizing microplastics: introducing new application-based and Spectral Libraries of Plastic Particles (SLoPP and SLoPP-E). Anal Chem 92:2443–2451.  https://doi.org/10.1021/acs.analchem.9b03626CrossRefGoogle Scholar
  46. Nalbone J (2014) Unseen threat: how microbeads harm New York waters, wildlife, health and environment, environmental protection Bureau of the New York State Attorney General’s Office. Albany (NY). http://ag.ny.gov/pdfs/Microbeads_Report_5_14_14.pdf
  47. Nelms SE, Barnett J, Brownlow A, Davison NJ, Deaville R, Galloway TS, Lindeque PK, Santillo D, Godley BJ (2019) Microplastics in marine mammals stranded around the British coast: ubiquitous but transitory? Sci Rep 9:1075.  https://doi.org/10.1038/s41598-018-37428-3CrossRefGoogle Scholar
  48. Nizzetto L, Futter M, Langaas S (2016) Are agricultural soils dumps for microplastics of urban origin? Environ Sci Technol 50:10777–10779.  https://doi.org/10.1021/acs.est.6b04140CrossRefGoogle Scholar
  49. Obbard RW (2018) Microplastics in polar regions: the role of long range transport. Current Opinion in Environmental Science & Health 1:24–29.  https://doi.org/10.1016/j.coesh.2017.10.004CrossRefGoogle Scholar
  50. Oliveira P, Barboza LGA, Branco V, Figueiredo N, Carvalho C, Guilhermino L (2018) Effects of microplastics and mercury in the freshwater bivalve Corbicula fluminea (Müller, 1774): filtration rate, biochemical biomarkers and mercury bioconcentration. Ecotoxicol Environ Saf 164:155–163.  https://doi.org/10.1016/j.ecoenv.2018.07.062CrossRefGoogle Scholar
  51. Prata JC, da Costa JP, Duarte AC, Rocha-Santos T (2019) Methods for sampling and detection of microplastics in water and sediment: a critical review. TrAC Trends Anal Chem 110:150–159.  https://doi.org/10.1016/j.trac.2018.10.029CrossRefGoogle Scholar
  52. Provencher JF, Vermaire JC, Avery-Gomm S, Braune BM, Mallory ML (2018) Garbage in guano? Microplastic debris found in faecal precursors of seabirds known to ingest plastics. Sci Total Environ 644:1477–1484.  https://doi.org/10.1016/j.scitotenv.2018.07.101CrossRefGoogle Scholar
  53. Quinn B, Murphy F, Ewins C (2017) Validation of density separation for the rapid recovery of microplastics from sediment. Anal Methods 9:1491–1498.  https://doi.org/10.1039/C6AY02542KCrossRefGoogle Scholar
  54. Rochman CM, Browne MA, Halpern BS, Hentschel BT, Hoh E, Karapanagioti HK, Rios-Mendoza LM, Takada H, Teh S, Thompson RC (2013) Classify plastic waste as hazardous. Nature 494:169–171.  https://doi.org/10.1038/494169aCrossRefGoogle Scholar
  55. Rusina TP, Smedes F, Klanova J, Booij K, Holoubek I (2007) Polymer selection for passive sampling: a comparison of critical properties. Chemosphere 68:1344–1351.  https://doi.org/10.1016/j.chemosphere.2007.01.025CrossRefGoogle Scholar
  56. Stanton T, Johnson M, Nathanail P, MacNaughtan W, Gomes RL (2019) Freshwater and airborne textile fibre populations are dominated by ‘natural’, not microplastic, fibres. Sci Total Environ 666:377–389CrossRefGoogle Scholar
  57. Sutton R, Mason SA, Stanek SK, Willis-Norton E, Wren IF, Box C (2016) Microplastic contamination in the San Francisco Bay, California, USA. Mar Pollut Bull 109:230–235.  https://doi.org/10.1016/j.marpolbul.2016.05.077CrossRefGoogle Scholar
  58. Talvitie J, Mikola A, Setälä O, Heinonen M, Koistinen A (2017) How well is microlitter purified from wastewater? – A detailed study on the stepwise removal of microlitter in a tertiary level wastewater treatment plant. Water Res 109:164–172.  https://doi.org/10.1016/j.watres.2016.11.046CrossRefGoogle Scholar
  59. Teuten EL, Saquing JM, Knappe DRU, Barlaz MA, Jonsson S, Björn A, Rowland SJ, Thompson RC, Galloway TS, Yamashita R, Ochi D, Watanuki Y, Moore C, Viet PH, Tana TS, Prudente M, Boonyatumanond R, Zakaria MP, Akkhavong K, Ogata Y, Hirai H, Iwasa S, Mizukawa K, Hagino Y, Imamura A, Saha M, Takada H (2009) Transport and release of chemicals from plastics to the environment and to wildlife. Philos Trans R Soc B 364:2027–2045.  https://doi.org/10.1098/rstb.2008.0284CrossRefGoogle Scholar
  60. Thompson R (2005) New directions in plastic debris. Science 310:1117b–1117b.  https://doi.org/10.1126/science.310.5751.1117bCrossRefGoogle Scholar
  61. Van Cauwenberghe L, Claessens M, Vandegehuchte MB, Janssen CR (2015) Microplastics are taken up by mussels (Mytilus edulis) and lugworms (Arenicola marina) living in natural habitats. Environ Pollut 199:10–17.  https://doi.org/10.1016/j.envpol.2015.01.008CrossRefGoogle Scholar
  62. Vermaire JC, Pomeroy C, Herczegh SM, Haggart O, Murphy M (2017) Microplastic abundance and distribution in the open water and sediment of the Ottawa River, Canada, and its tributaries. FACETS 2:301–314.  https://doi.org/10.1139/facets-2016-0070CrossRefGoogle Scholar
  63. von Moos N, Burkhardt-Holm P, Köhler A (2012) Uptake and effects of microplastics on cells and tissue of the blue mussel Mytilus edulis L. after an experimental exposure. Environ Sci Technol 46:11327–11335.  https://doi.org/10.1021/es302332wCrossRefGoogle Scholar
  64. Wagner M, Lambert S (2018) Freshwater microplastics: emerging environmental contaminants? Springer Berlin Heidelberg, New YorkCrossRefGoogle Scholar
  65. Windsor FM, Tilley RM, Tyler CR, Ormerod SJ (2019) Microplastic ingestion by riverine macroinvertebrates. Sci Total Environ 646:68–74.  https://doi.org/10.1016/j.scitotenv.2018.07.271CrossRefGoogle Scholar
  66. Yonkos LT, Friedel EA, Perez-Reyes AC, Ghosal S, Arthur CD (2014) Microplastics in Four Estuarine Rivers in the Chesapeake Bay, U.S.A. Environ Sci Technol 48:14195–14202.  https://doi.org/10.1021/es5036317CrossRefGoogle Scholar
  67. Zbyszewski M, Corcoran PL (2011) Distribution and degradation of fresh water plastic particles along the beaches of Lake Huron, Canada. Water Air Soil Pollut 220:365–372.  https://doi.org/10.1007/s11270-011-0760-6CrossRefGoogle Scholar
  68. Zhang K, Gong W, Lv J, Xiong X, Wu C (2015) Accumulation of floating microplastics behind the Three Gorges Dam. Environ Pollut 204:117–123.  https://doi.org/10.1016/j.envpol.2015.04.023CrossRefGoogle Scholar

Copyright information

© Springer Nature Switzerland AG 2020

Authors and Affiliations

  • Shaun A. Forrest
    • 1
  • Madelaine P. T. Bourdages
    • 1
  • Jesse C. Vermaire
    • 1
    • 2
    Email author
  1. 1.Geography and Environmental StudiesCarleton UniversityOttawaCanada
  2. 2.Institute for Environmental and Interdisciplinary SciencesCarleton UniversityOttawaCanada

Section editors and affiliations

  • João Pinto da Costa
    • 1
  • Armando da Costa Duarte
    • 2
  1. 1.Department of Chemistry and Centre for Environmental and Marine StudiesUniversity of AveiroAveiroPortugal
  2. 2.Department of Chemistry & CESAMUniversity of AveiroAveiroPortugal

Personalised recommendations