Nanoplastic toxicity towards freshwater organisms

INTRODUCTION

Global plastic pollution is a social, political, and scientific cause for concern due to the large amount of plastic litter currently ending up in the environment^[1]. Approximately, 9 × 10¹² tons of plastics have been marketed since 1950, with a current annual production of > 360 million tons^[2]. Despite the recent slight reduction in global plastic manufacturing, the increasing social awareness concerning this type of materials, and the political attempts to regulate single-use plastics, the global trends in plastic use by segments are preserved^[3,4]. A considerable amount of these plastics ends up in the environment through different dissemination pathways^[5-8]. For instance, the occurrence of microplastics (MPs) in the upper ocean layer has been estimated at 0.8-5.8 × 10⁵ tons, equivalent to > 10¹⁹ items^[9]. A major source of ocean plastic pollution comes from rivers, the contribution of which has been estimated to range 0.8-2.7 × 10⁶ tons/year^[10].

Plastic fragmentation proceeds due to environmental factors such as photodegradation, hydrolysis, or physical abrasion that ultimately result in small fragments, the smallest of which are termed nanoplastics (NPLs)^[11-13]. Despite the limited knowledge of the actual role of different aging processes, the potential release of NPLs from MPs raises the possibility of increasing by several orders of magnitude the number of plastic fragments in the environment^[14]. The main property defining NPLs is their size, specifically the length of their largest dimension. There is no agreement within the scientific community regarding the upper limit of this size range. Some authors use the limit of 100 nm^[15,16], while others prefer 1000 nm^[17,18]. There are reasons to support any of these definitions based on analytical limitations or colloidal behavior in water suspension, but a detailed discussion is outside the scope of this review. Overall, NPLs must be considered emerging pollutants with specific properties, different from both larger plastic items, such as MPs, and engineered nanomaterials^[19].

Plastic pollutants can be divided into two categories: “primary” refers to plastic items intentionally produced in that specific size and shape that end up in the environment as a consequence of their use or due to waste mismanagement^[20] and “secondary” denotes plastic items that are caused by the environmental fragmentation of larger particles^[21,22]. This criterion allows policymakers to establish regulations based on their different environmental risks^[23]. It is also important to consider the aging processes they undergo because of their influence on reactivity, release of additives, pollutants adsorption behavior, and integrity of plastic particles, among others^[24,25]. Furthermore, it is important to include the weathering degree of plastic as an additional criterion for particle characterization. For this propose, standardized methods are needed. The literature provides some approaches that may be useful, such as those based on the oxygen-containing surface groups^[26-28]. However, precise characterization criteria for plastic particles are still needed to evaluate the environmental risk of plastic pollution, especially concerning the lower size ranges.

The assessment of NPLs in complex matrices has been hindered by the limited availability of adequate analytical techniques, although new recent tools and methodologies, particularly those based on mass spectrometry, have allowed significant progress in that direction^[28,29]. Recent reports concerning NPLs occurrence in the environment have shown their presence in different aqueous and terrestrial compartments, and their widespread presence is generally assumed^[30-32]. In parallel, investigation on the potential effects of NPLs to the biota is receiving increasing attention driven by data showing that they are potentially more harmful than larger fragments. NPLs can be internalized by cells through either passively crossing the cellular membrane (promoted by their hydrophobicity and small size) or endocytic processes^[33,34]. Furthermore, their large surface area to volume ratio makes them more prone to interact with environmental contaminants^[35]. The capacity to act as a vector for the transfer of pollutants to aquatic organisms has been termed the “Trojan Horse” effect and is the subject of active research^[36].

The goal of this review is to discuss the recently reported studies (since 2019) on the effects of NPLs to freshwater organisms. The articles were selected from a first thorough search using the Web of Science citation database with the keywords defining this review (nanoplastics, environmental fate, toxicity, and freshwater organisms) followed by a cross-referencing search in an attempt to identify all relevant articles covering the nanoplastic toxicity towards freshwater organisms. A screening of similar articles from the same groups led to the set of references cited herein. Although most published results refer to polystyrene (PS) NPLs, we also review those obtained with other polymers, with an emphasis on secondary NPLs rather than those specifically produced in that size. Furthermore, as potential replacing material for the traditional oil-based plastics, we include a section focused on the impact of biodegradable plastics in the environment. As the environmental fate of NPLs is widely determined by the stability of their colloidal properties, we reserve one specific section to review the existing body of knowledge on this specific topic. In what follows, studies are classified based on the trophic level of the organisms: primary producers, primary consumers, and secondary producers. Studies concerning the combined toxicity of NPLs and other emerging pollutants are also reviewed. Finally, this review identifies research needs and gives recommendations aimed at minimizing NPLs pollution in the environment.

SOURCE AND OCCURRENCE OF NPLS IN FRESHWATER ENVIRONMENTS

The emission sources as well as the impact of NPLs in the environment remain largely unknown mainly due to the limitations concerning the characterization and identification of small carbon-based particles in complex matrices^[15]. In this context, studies are increasingly addressing the potential release of primary and secondary NPLs under relevant conditions. Regarding primary NPLs, it has been shown that the use of facial scrub may release over 10¹³ sub-micron particles per gram of product, mostly discarded with household wastewater^[29]. Several studies have addressed the continuous release of NPLs from larger plastic items subject to environmental degradation. Nylon and polyethylene terephthalate (PET) teabags have been shown to release > 10¹² NPLs (< 100 nm) along with a similar number of MPs into a single cup of tea^[37]. Morgana et al. determined that a single face mask could release up to 10⁸ NPLs under mechanical stress forces mimicking those encountered in the environment^[38]. Zhang et al. demonstrated the release of NPLs from the surface of recycled PVC^[39]. Sorasan et al. reported the generation of up to 10¹⁰ NPLs per gram of low-density polyethylene (LDPE) after the exposure of MPs to mechanical agitation and the equivalent to one year of solar irradiation^[11]. Luo et al. estimated a release of up to 3000 polypropylene (PP) items (MPs and NPs) per mm² of plastic chopping boards^[40]. Munoz et al. used in vitro experiments to simulate vaginal conditions and estimated a release of up to 1.7 × 10¹³ NPLs per tampon after 2 h of use^[41]. The available results show the existence of a number of potential sources of NPLs that may eventually end up in the environment, posing a risk to biota and human health.

Obtaining reliable NPL concentrations in environmental samples remains highly challenging despite the considerable efforts paid and the advances currently ongoing. In this regard, Xu et al. combined a concentration pretreatment (< 1 µm followed by ultrafiltration through 100 kDa membranes) with pyrolysis gas chromatography-mass spectrometry (Py-GC/MS) to identify and quantify NPLs from surface water and groundwater and reported a total mass concentration reaching 100s ng/L for PP and polyethylene (PE), which were the main polymers identified^[42]. It is important to note that mass spectrometry techniques provide mass concentrations but cannot give particle concentrations. Materić et al. analyzed the presence of NPLs in freshwaters from the Siberian Arctic tundra and a forest location in southern Sweden using thermal desorption proton transfer-reaction mass spectrometry (TD-PTR-MS). They identified four polymers, PE, polyvinyl chloride (PVC), PP, and PET, in different lake and stream samples, with a mean concentration of total nanoplastics (< 0.2 μm) as high as 563 μg/L^[43]. A study from the same group reported a concentration of PET and PVC in the 5-23 μg/L range in Alpine snow^[44]. These experimental data are not only very different from each other but also several orders of magnitude higher than the 0.14-1.4 ng/L range calculated combining the total estimated mass of plastic debris with 3D fragmentation models^[45]. The differences may be attributed to the heterogenous distribution of plastics or to a low model accuracy, but they evidence the lack of reliable field data concerning actual environmental concentrations of NPLs. Further efforts should be done in this direction supported by the current development of more accurate and efficient techniques that allow NPL identification and quantification in environmental samples.

PHYSICOCHEMICAL BEHAVIOR OF NPLS IN FRESHWATER

The occurrence of NPLs in freshwater implies their interaction with biota as well as other compounds naturally present such as natural organic matter (NOM), extracellular polymeric substances (EPS), or inorganic compounds (ions, including metals, clay, and other minerals), among others. Such interactions modulate the different mobility, toxicity, bioavailability, distribution, and fate of NPLs^[20,46]. The stability/aggregation behavior of NPLs is addressed based on the Derjaguin-Landau-Verwey-Overbeek (DLVO) theory that combines the effects of van der Waals attraction forces and electrostatic repulsion between charged particles, although other non-DLVO interactions, such as bridging flocculation, patch-charge attraction, π-π interaction, or steric repulsion, may be also involved in the aggregation process of NPLs^[47,48]. The aggregation kinetics of NPLs is frequently given by the evolution of hydrodynamic size (d_H) with increasing concentration of a given ion, which allow determining the critical coagulation concentration (CCC, or the concentration at which the aggregation rate is maximum). CCC has been commonly used to assess the stability of NPLs under different environmental conditions^[49-51]. Table 1 shows CCC values reported in the literature for different NPLs.

The stability of PS-NPL suspensions decreases as the concentration of monovalent ions increases as result of the screening effects of the ions that reduce repulsion forces. The aggregation is higher in the presence of divalent ions (using CaCl₂), in agreement with the Schulze-Hardy rule stating that higher valence ions result in faster aggregation due to the compression of the electrical double layer^[50,52,60]. This tendency has also been reported for other mono- and divalent ions using K⁺, Mg²⁺, and Ba^2+[55]. Indeed, trivalent ions such as Al³⁺ have been proposed as strong coagulants for NPL removal^[62]. Similar results were obtained for heavy metal salts^[50]. Concerning surface modified NPLs, it has been reported that carboxyl modified PS-COOH displays similar stability as non-functionalized PS-NPLs, while, in the case of amino-modified PS-NH₂, the suspension remained stable even at concentrations as high as 1 and 0.15 M of NaCl and CaCl₂, respectively^[51]. Similar observations have been reported for irregularly shaped NPLs, which otherwise presented lower CCC values compared with spherical PS-NPLs under similar conditions, indicating less stability of irregularly shaped NPLs aqueous suspensions^[49,61]. Weathering is another effect that influences NPL fate in the environment. Recent reports assessed the stability of UV-aged NPL suspensions, showing higher stability of aged NPLs due to the increase in oxygen-containing functional groups that decrease particle hydrophobicity and increase the absolute value of the ζ-potential of negatively charged NPLs^[55,59]. The stabilization of irradiated and oxidized PS-NPLs has been attributed to stronger Lewis acid-base interactions that resulted in higher hydration forces. In contrast, UV-aged PS-NPL suspensions displayed less stability than their pristine counterparts when exposed to increasing concentration of divalent cations. This finding has also been attributed to the bridging of oxygen-containing functional groups with Ca²⁺, thereby promoting the aggregation of UV-irradiated PS-NPLs in CaCl₂ solutions^[60]. Interestingly, when the weathering is simulated by ozonation, the stability of PS-NPL suspensions increased in the presence of both monovalent and divalent cations attributed to the steric repulsion caused by the attachment of organic matter released from PS degradation^[60]. Temperature has also been reported as an important factor determining NPL behavior in aqueous suspensions. It has been shown that temperature increases reduce the CCC values of PS-NPLs, promoting NPL aggregation^[50].

NOM, ionic strength (IS), and pH are the most relevant parameters determining the fate of NPLs in real freshwater^[49]. The influence of NOM on the stabilization of NPL suspensions depends on the simultaneous presence of electrolytes as well as the concentration and type of NOM^[60]. The differences have been attributed to the thickness of the macromolecular layer adsorbed onto the surface of NPLs^[54]. Despite the limited number of studies addressing the stability of non-pristine NPL suspensions, it has been reported that mechanically degraded PS-NPL suspensions stabilize in the presence of humic acids (HA) through electrostatic and steric repulsions, as well as with sodium alginate (SA) via hydrogen bonds and van der Waals interactions^[49]. The stability of UV-aged PS-NPLs increased in the presence of monovalent electrolytes, although SA yielded less stable suspensions. However, ozonated PS-NPL suspensions displayed higher stability in the presence of HA and SA, in the presence of both mono- and divalent electrolytes^[60].

The concentration of mono- and divalent ions reported for freshwater bodies rarely exceeds ~50 mM for Na⁺ and ~2.5 mM for Ca²⁺. Therefore, it is likely that most NPL suspensions are stable in natural freshwater ecosystems since their CCC values are considerably higher^[63]. However, freshwater may contain other substances than those reviewed here that could co-occur with NPLs and modify the stability of their suspensions. The results summarized in the Table 2 indicate that most tested NPL suspensions displayed considerable stability. This implies that NPLs in freshwater ecosystems would be bioavailable within the water column and that their fate and distribution would be dominated by their mobility throughout the water column.

It is important to note that all the results listed in the preceding tables (and most of those in the following ones) correspond to spherical PS particles specifically produced in that size and not to incidental secondary NPLs, which would be expected to display a variety of shapes. This is a limitation found in most of the literature concerning NPLs and the reason we also include in this review article some other polymeric NPs such as poly(amidoamine) (PAMAM) dendrimers, which are not conceptually distant from PS latexes.

NPL TOXICITY TOWARDS FRESHWATER PRIMARY PRODUCERS

Freshwater primary producers such as benthic algae and cyanobacteria (periphyton), phytoplankton (suspended algae and cyanobacteria), and macrophytes are crucial for the preservation of freshwater trophic chains. Considering the large amount of plastic litter transported by rivers, NPLs are expected to influence primary producers^[65]. Table 3 summarizes the main recently published findings on the single and combined toxicity of NPLs towards freshwater primary producers (excluding macrophytes), highlighting the more sensitive toxicity endpoints as reported by the authors. Thus far, most of the in vitro studies have assessed NPL toxicity at high concentrations. This approach may disclose potential biological targets (such as ROS homeostasis alteration and photosynthesis impairment) of NPLs and establish dose-response curves to further understand the toxicological behavior of NPLs and their interaction with other pollutants^[91]. It has been shown that both micrometric and nanometric plastic particles may trigger clear effects at high concentrations. In addition, 100 nm PS-NPLs have been shown to cause higher growth inhibition, higher ROS and lipid peroxidation levels, and overproduction of antioxidant enzymes in the algae Chlamydomonas reinhardtii compared to 100 µm MPs at the same mass concentration^[66]. Furthermore, the internalization of NPLs in algae and cyanobacteria has been reported for sizes between 20 and 100 nm^[67,78], as well as in microalgae for sizes up to 2 µm^[68]. This process occurs through different potential pathways: (1) direct crossing through the porous structures of cell envelopes for NPLs < 20 nm; (2) direct passage through the cell wall owing to increased cell membrane permeability during cell cycling (up to 140% of the normal permeability); and (3) endocytosis for larger NPLs^[92]. These processes, together with NPL attachment onto the cell surface, may result in the ingestion of NPLs by grazers^[93]. Apart from their effects at high concentrations, NPL concentrations ≤ 1 mg/L have been reported to cause effects on primary producers. Xiao et al. observed a reduction in pigment content (chlorophyll b) and an increase in superoxide dismutase (SOD) activity in Euglena gracilis exposed to 1 mg/L of 100 nm PS-NPLs^[67]. Wang et al. reported a growth inhibition of 15.6% in Chlorella pyrenoidosa upon exposure to 1 mg/L of 600 nm PS-NPLs^[79]. Baudrimont et al. reported significant growth reduction of the green alga Scenedemus subspicatus exposed to 1 mg/L of PE-NPLs, which was greater when using PE-NPLs gathered from the North Atlantic gyre than reference PE-NPLs. The difference attributed to the accumulation of trace metals in plastics exposed to environmental pollutants^[69]. Similarly, it has been described that the photosynthetic activity of the alga E. gracilis was impaired by 100 nm PS-NPLs at concentrations as low as 0.5 mg/L^[80]. In contrast, low concentrations of NPLs have been reported to promote algal growth at concentrations < 1 mg/L, sometimes in long-term exposure assays^[81,94]. This discrepancy may be related to the physicochemical conditions that alter the colloidal status of NPLs suspensions and, consequently, their toxicological behavior. Finally, it is important to stress that, in combination with other pollutants, low concentrations of NPLs may induce considerable effects. For instance, co-exposure to PS-NPLs and Cd²⁺ at the same concentration (0.05 mg/L) resulted in significant and synergistic inhibition of algal growth^[80]. Further efforts should be made to reach a deeper understanding of the effects of NPLs on primary producers at realistic concentrations and environmental conditions in order to accurately assess their environmental fate and risk. The use of real secondary NPLs is highly recommended.

TOXICITY TOWARDS FRESHWATER PRIMARY CONSUMERS

Freshwater primary consumers, organisms that feed on primary producers, include invertebrates, some fishes, and a few amphibian larvae. Among the primary consumers, invertebrates are the dominant grazers in the freshwater ecosystems of temperate latitudes. In this regard, most of the NPLs toxicological studies have been carried out using different species of Daphnia, one of the preferred organisms for toxicity assessment^[95]. Table 4 summarizes the main findings reported since 2019 concerning single and combined toxicity of NPLs towards freshwater primary consumers, highlighting the most sensitive assessed toxicity endpoints. EC₅₀ for negatively charged NPLs (more common than positively charged NPLs) has been reported to range between 30 and 300 mg/L, depending on the plastic used and the organism assessed^[96,102]. However, harmful effects, such as ROS overproduction or the reduction in the number of neonates per brood, have been reported at lower concentrations (≤ 1 mg/L)^[108,109]. The advantage of behavioral endpoints is the possibility to observe sublethal effects, which are generally undetectable for tests based on global or lethal endpoints, although their findings may be difficult to interpret^[95]. D. magna swimming behavior has been reported to change by the exposure to PS-NPLs at concentrations > 16 mg/L^[98]. However, such alterations do not seem to appear at concentrations < 1 mg/L, despite a clear accumulation of NPLs in the gut and external body appendages^[110,111]. It is worth noting that some of the effects observed in daphnids at low concentrations only appeared during multigenerational or long-term assays, revealing that the interaction between these organisms and NPLs may occur through parental transfer, and their effects may span the entire lifetime of the organism^[97,98]. The effect of NPL ingestion by D. magna through feeding has been tracked using algae pre-exposed to 270 and 640 nm metal-doped NPLs (4.8 × 10¹⁰ particles/L, 1-7 mg/L). The results show that both types of NPLs became attached to the cell surface of algae and are ingested by the daphnids and excreted without any effects on daphnids growth were observed, but the smaller ones increased the reproduction time, reduced the number of neonates, and induced higher offspring mortality^[112]. These finding reveal that low concentrations of NPLs may damage primary consumers not only through direct exposure but also through feeding on NPL-polluted algae. Finally, it is important to note that the effects described above may be enhanced by the co-occurrence of NPLs with other chemicals. For instance, it has been described that glyphosate (the active compound of several herbicides) in combination with PS-NPLs exerts synergistic and multigenerational effects on D. magna^[103]. The interaction between NPLs and potential co-occurring contaminants should be studied to better understand their environmental risk under realistic conditions. Furthermore, given the relatively long lifetime of higher organisms and the data showing multigenerational effects, the long-term fate and toxicity of NPLs on freshwater biota should be addressed.

TOXICITY TOWARDS FRESHWATER SECONDARY CONSUMERS

Secondary consumers are also crucial for the equilibrium of freshwater ecosystems. Located at the top of the trophic chain, any alterations to them may cause a potential cascade of interactions through the food web. Furthermore, human health may be jeopardized due to the consumption of organisms such as fish or crustaceans that could constitute an important route for NPL transfer to humans^[25]. Table 5 summarizes the main findings reported since 2019 on the toxicity of NPLs towards freshwater secondary consumers. The exposure of Danio rerio (zebrafish) to 1 mg/L of 500 nm fluorescent PS-NPLs revealed particle translocation from the gut epithelium of the digestive tract to different tissues where they activated enzymatic responses against oxidative stress^[113]. Small NPLs (70 nm) have been reported to accumulate in zebrafish gonads, intestine, liver, and brain causing oxidative stress, metabolic alterations, and neurological impairments, including the decrease in acetylcholine esterase, acetylcholine, or gamma-aminobutyric acid, as well as neurobehavioral alterations, at concentrations as low as 0.5 mg/L^[114]. Important effects have also been found for 20 nm PS-NPLs, which resulted in increased fish mortality, occurrence of abnormalities, and excessive ROS formation and apoptosis, particularly in the brain^[115]. Other studies reported physical abnormalities found in different freshwater organisms such as Xenopus laevis or Hydra viridissima at concentrations as low as 1 mg/L, highlighting the importance of using approaches that overcome the limitations of traditional toxicity tests^[116,117]. Interestingly, several recent studies focused on the potential effects induced by NPLs in the intestinal microbiome of freshwater secondary consumers. It has been found that both MPs and NPLs may cause dysbiosis in the zebrafish gut at very low concentrations (1 µg/L), but NPLs can also increase the presence of pathogenic genera, such as Aeromonas^[131]. It has also been found that PS-NPLs at concentrations ≤ 0.1 mg/L affected the brain-intestine-microbiota axis of zebrafish, causing reduced growth, inflammatory responses, and altered intestinal permeability, even inducing transgenerational effects such as NPL accumulation in the gastrointestinal tract of the offspring^[132]. Microbiome alterations have been described in the freshwater crustacean Procambarus clarkia (crayfish) exposed for 48 h to 75 nm PS-NPLs, which resulted in a reduced abundance of Lactobacillus and an increase in the number of pathogenic bacteria, probably linked to a lower immunity^[133]. Concerning the co-occurrence of NPLs with other pollutants, the results are still limited. No clear toxicological interactions have been described in zebrafish exposed to polycyclic aromatic hydrocarbons or the herbicide phenmedipham (PHE) in the presence of NPLs^[18,128]. However, the combined exposure of PE-NPLs with bovine serum albumin (as a model of a naturally occurring protein) induced more toxic effects on zebrafish than their co-exposure with an artificial surfactant such as sodium dodecyl sulfate, which was attributed to the higher colloidal stability provided by the first^[129]. Considering the complexity of this type of organisms, the combination of in vitro studies with different cell lines together with in vivo studies using the whole organism are deemed necessary for understanding the potential adverse outcome pathways of NPLs to secondary consumers. Finally, attention should also be paid to artifacts when using labeled plastic particles due to the leaching of fluorochromes or metals.

BIODEGRADABLE NPLS

The materials produced to replace the traditional petroleum-based plastics are ambiguously referred to using several terms such as biodegradable or biobased plastics. A wide denomination for all these types of plastic material is “bioplastics”. Table 6 summarizes the information concerning the main types of bioplastics developed for replacing the traditional non-biodegradable petroleum-based ones. Among them, the category of biobased plastics refers to plastic materials manufactured using renewable resources. It is important to note that biobased plastics are not free from environmental issues. The life cycle assessment of biobased plastics shows that they may reduce carbon emissions, but other characteristics, such as their persistence, are not necessarily better than those of conventional plastics. Furthermore, there is an important problem concerning the occupation of agricultural land for their production. The materials based on conventional plastics supplemented with additives that allow their rapid degradation are not a realistic solution since this process enables only their fragmentation into smaller pieces but not their complete degradation. Accordingly, oxo-degradable plastics have been restricted in the EU and Switzerland. Regardless of the type of plastic, recycling is difficult due to the presence of additives in almost every finished plastic product. However, it is important to note that recycling is clearly the most environmentally friendly option for end-of-life plastic management, even better than composting. Accordingly, the preferred bioplastics would be those both biobased and biodegradable. This is the case of bioplastics such as polylactide or polylactic acid (PLA) and polyglycolide (PGA) along with those obtained from bacteria or algae that do not imply the use of lands for agriculture, such as polyhydroxyalkanoates (PHA).

The largest plastic demand by segment in 2015 was in packaging (36%), which is considered the greatest source of waste, globally accounting for 146 million tons that year, of which > 95% was not recycled^[140]. Bioplastics are mainly being developed for single-use products, such as packaging, in order to reduce the environmental burden of plastic wastes^[141]. The global production capacity for biodegradable plastics is still modest, 2.24 million tons in 2021, but it is expected to expand up to 7.5 million tons by 2026 (source: European Bioplastics). Thus, in the near future, bioplastics are expected to reach the aquatic ecosystem following similar routes as petroleum-based materials^[142]. However, their potential impact on organisms of freshwater environments has been shown to be similar to that from conventional plastics, or even larger due to their more rapid degradation^[143,144]. Furthermore, during their degradation, bioplastics may release millions of MPs and billions of NPLs per gram^[145]. Table 7 summarizes the main findings reported on the toxicity of biodegradable NPLs (including some carbon-based nanoparticles) towards freshwater organisms. Secondary biodegradable NPLs have been shown to consist of short polymeric chains (< 1600-3000 kDa) produced during the degradation of larger items that will continue to release as long as the source (any biodegradable plastic litter) remains in the environment.

There is a considerable lack of information concerning the colloidal stability of bioplastics, but the knowledge gathered during the last decade with other NPLs suggests that their higher degradability could lead to a more oxidized surface (probably along with a more negative surface net charge) and higher stability in aqueous suspension. This colloidal stability could be comparable to that observed in artificially aged petroleum-based NPLs (see Table 1) but in considerably less time and under softer weathering conditions. The information on the toxicity of biodegradable NPLs is also scarce but points towards a non-negligible biological impact on freshwater organisms. For instance, unlike the studies summarized in Table 7, Tong et al. did not find that PBAT- or PLA-NPLs affected the survival of the copepod Tigriopus japonicas^[150]. Likewise, Götz et al. did not report any adverse outcomes to the freshwater invertebrate Gammarus roeseli exposed to different particle sizes of PS- and PLA-NPLs at concentrations up to 430 ng/mg of food^[151]. Overall, the environmental fate and risk of biodegradable nanometric plastic remains poorly understood and needs further scientific efforts to be properly assessed. As the substitution of petroleum-based plastics by biodegradable materials is accelerating, a thorough risk assessment is urgently needed to ensure a sustainable replacement for petroleum-based plastics.

REMARKS AND FUTURE RESEARCH NEEDS

The current knowledge on the distribution of plastic litter and the information available on the effect of the weathering processes suffered by plastic debris suggest a widespread presence of NPLs in all freshwater compartments. The first attempts to measure the environmental concentration of NPLs in freshwater systems, along with the estimations obtained from mathematical models, point to probable environmental concentrations in the parts per billion (< 1 µg /L) range, similar to other anthropogenic pollutants. The development of techniques for the routine monitoring of NPLs in environmental samples is urgently needed.

Most environmental fate and toxicity studies have been performed using commercially available or synthetic NPLs, especially PS-NPLs (PS latexes). This approach allowed gaining a considerable body of knowledge on the colloidal stability of NPLs in water bodies and insight into their main toxicity drivers, but it does not represent the variety of shapes and chemical compositions of real secondary NPLs that can be found in the environment. Special attention should be paid to possible artifacts due to the leaching of the substances used to label plastic particles.

The available data show clear damage upon NPL exposure at concentrations as high as tens or even hundreds of milligrams per liter. The use of high concentrations clarifies potential biological targets, but efforts should be made to assess the possible effects upon exposure to realistic environmental concentrations. The effect of NPLs is expected to be enhanced at low concentrations due to higher colloidal stability and possibly triggered by the release of oligomeric fractions detached from larger particles. Long-term assays and mesocosm studies using low concentrations of secondary NPLs would be needed to perform realistic risk assessments for regulatory purposes.

Although biodegradable plastics are considered environmentally friendly substitutes for traditional petroleum-based polymers, their risk must be assessed in the same way it is being performed for their non-biodegradable counterparts. Thus far, there is very limited information regarding the physicochemical behavior of biodegradable plastics in relevant conditions and their impact on biological organisms and ecosystems. This is a particularly urgent need as bioplastics are already replacing conventional plastics in various segments of the global plastic market.