Interactions of Nanoplastics with Freshwater Phytoplankton Species: History
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Subjects: Environmental Sciences
Contributor:
Vera I. Slaveykova
,
Matea Marelja

Nano-sized plastics (NPLs, size < 100 nm) are characterized by a very small size and high reactivity, allowing them to interact with diverse phytoplankton species. The processes governing the interactions of NPLs with phytoplankton cells include absorption onto cell, penetration into cells via endocytosis or physical damage, and the obstruction of substance and energy exchange with the surrounding medium. Upon association with the cells, elevated concentrations of NPLs can reduce phytoplankton growth and photosynthesis,  trigger overproduction of reactive oxygen species and damages, as well as alter cellular metabolic activity. NPLs can influence toxin production by cyanobacteria and  release of extracellular polymeric substances by different phytoplankton species. Conversely, phytoplankton species can modulate NPL fate by secreting biomolecules that form an eco-corona around the NPLs, as well as taking part in the NPLs biotransformation.

  • nanoplastics
  • uptake
  • toxicity
  • trophic transfer
  • algae
  • cyanobacteria

1. Introduction

Plastic particles are considered as nanomaterials if at least half of them, in an unbound state or as an aggregate or  an agglomerate state, have at least one dimension ranging from 1 to 100 nm [1]. As the usage of diverse plastic materials continues to expand, the accumulation of plastic waste is on the rise. Consequently, the presence of nanoplastics (NPLs) in the environment is increasing [2][3][4] together with the concerns regarding the possible environmental implications of NPLs   [5][6]. Indeed, NPLs, originating from various primary or secondary sources, represent the least explored facet of plastic pollution [5][6]. In environmental settings, diverse plastic debris can release a large amount of NPLs through various physical and chemical degradation pathways, as previously reviewed [7]. The concentrations of NPLs in the environment have not yet been measured due to the constraints of current analytical methodologies. Multi-media model estimations provide average concentrations of NPLs in surface water of 280 µg L−1 [8]. More recently, NPLs’ abundance has been estimated to be in the range 0.3–488 μg L−1 in freshwaters, which are higher than those for marine environments (2.7–67 μg L−1) [9]. These concentrations are generally lower as compared to the predicted no-effect concentrations (PNECs) derived from probabilistic species sensitivity distributions, resulting in values of 99 μg L−1 and 72 μg L−1 for freshwater and marine datasets [10]. They are also below the estimated hazard concentration affecting 5% of the species (HC5) of 410 μg L−1 for marine plankton for two types of materials, polymethylmethacrylate (PMMA) and polystyrene (PS) [11], as well as HC5 for NPLs with a size of 50 nm in freshwater, 187.9 (8.0–2978.3) μg L−1 [12]. Nevertheless, some hotspots of NPLs pollution can be of risk for aquatic organisms,  including phytoplankton. Consequently, considerable attention has been paid to the bioaccumulation of NPLs and their adverse effects within higher trophic levels [9]. Recent insigths into their accumulation within aquatic organisms  have highlighted numerous challenges [13]. The accumulation of NPLs in aquatic organisms has been shown to lead to detrimental effects on various freshwater organisms [4][5][6][7][9][1][14][15][16][17][18][19][20]. These effects include oxidative stress and damage, inflammation, altered development, reduced growth, energy and movement, genotoxicity, etc., as recently reviewed in references [9][16][19][21][22][23][24].  Modifying factors, such as particle characteristics, concentration, size, exposure duration, and co-factors like presence of other organic or inorganic contaminants, food availability, species, development stage, and environmental conditions were extensively discussed in references [16][22][24][25]. However, there is a scarcity of scientific evidences regarding the interactions and possible impacts of NPLs on phytoplankton species.

NPLs are characterized by a very small size and high reactivity [26]. NPLs interact with different abiotic and biotic components in aquatic environments. Consequently, they undergo various transformation, such as aggregation, sedimentation, and chemical or physical transformations, profoundly influencing their  reactivity and environmental impact [4][6][7][1][27]. Similarly to the engineered nanoparticles, the interactions of NPLs with phytoplankton species, such as algae and cyanobacteria, include  adsorption onto cell, penetration into the cell via endocytosis or physical damage, and the shading and blocking of substance and energy exchange with the surrounding medium (Figure 1) [21].

Figure 1. Key processes governing the interactions of the NPLs and phytoplankton cells.

Indeed, both amidine and carboxyl - functionalized polystyrene NPLs (PS-COOH) were found to adsorb onto the marine diatom Dunaliella tertiolecta. However, only amidine PS NPLs triggered the inhibition of algal growth, displaying an effective concentration for 50% of the algal population (EC50)  of 12.97 μg mL−1 [28]. In another study, fluorescent-blue 50 nm amino - modified polystyrene NPLs (PS-NH2) adhered to the diatom Chaetoceros neogracile, leading to impairment of the photosynthetic machinery and an overproduction of reactive oxygen species (ROS) at both low (0.05 μg mL−1) and high (5 μg mL−1) exposure concentrations [29]. The adsorption onto the cells of green alga Pseudokirchneriella subcapitata was significantly higher for neutral and positively charged PS-NH2 NPLs at a concentration of 100 mg L−1. Conversely, negatively charged PS-COOH NPLs displayed minimal adsorption onto the algal cell wall [30]. These examples highlight the important role of the surface charge of NPL particles and the specificity of interactions with different algal species. The rapid adsorption and/or absorption of PS NPLs onto/in diatom Phaeodactylum tricornutum was evidenced through an observed increase in cell complexity, size and microalgae fluorescence induced by 100 nm fluoresbrite [31]. Fluorescent 51 nm PS NPLs attached to and penetrated the outer layer of green alga Chlamydomonas reinhardtii during cell division [32]. In a recent study involving metal - doped PS NPLs, it was demonstrated that more than 60% of Fe-PS or Eu-PS NPLs remained associated with algal cells of P. subcapitata after 72 hours [33]. A recent study uncovered that fluorescent aggregation - induced emission fluorogens-incorporated nanoparticles (AIE-NPs) of sizes 40, 70, and 85 nm were internally taken up via clathrin - dependent endocytosis in C. reinhardtii, while the 140 nm AIE-NPs remained surface-bound [34]. Notably, the authors highlighted the importance  of endocytosis, algal cell membrane permeability, and the thickness of extracellular polymeric substances (EPS) and their cell cycle dependence in the uptake of AIE-NPs [34].

2. Effect of Nanoplastics on Phytoplankton Species

A recent review paper has examined the detrimental effect of different micro- and nanoplastics on phytoplankton [35]. The behavior and adverse effects of PS NPLs with positive and negative surface charges to marine plankton have been reviewed [36]. For instance, exposure of the marine green microalga Platymonas helgolandica to 70 nm PS NPLs at concentrations of 20, 200, and 2000 μg L−1 resulted in inhibited algal growth after 3 days, followed by stimulation after 5 days of exposure. Higher concentrations of PS NPLs (200 and 2000 μg L−1) increased membrane permeability and mitochondrial membrane potential, decreased light energy utilization in photochemical processes of microalgae, and affected cell morphology and organelle function [37]. Exposure to 100 nm PS NPLs at concentrations ranging from 10 to 200 mg L−1 resulted in the stimulation of growth in the green alga Scenedesmus quadricauda. At the higher concentration of 200 mg L−1, PS NPLs induced an increase in chlorophyll content, soluble proteins, and polysaccharides and an enhancement in the activities of antioxidant enzymes [38]. Exposure to 10–100 mg L−1 of 100 nm PS NPLs led to dose-dependent adverse effects on another green alga Chlorella pyrenoidosa growth, which was observed from the lag to the earlier logarithmic phases. However, during the transition from the end of the logarithmic to the stationary phase, C. pyrenoidosa demonstrated resilience to the adverse effects of NPLs. This resilience was achieved through mechanisms such as thickening of cell walls, algae homo-aggregation, and algae - NPLs hetero-aggregation, which triggered an increase in algal photosynthetic activity and growth promotion [39]. Further studies have documented that the exposure of various algae and cyanobacteria to different NPLs led to oxidative stress, membrane damages, and alterations in photosynthesis. Specifically, 50 and 100 nm PS NPLs induced increased oxidative stress biomarkers, damage to the photosynthetic apparatus, DNA damage, and the depolarization of mitochondrial and cell membranes in the marine diatom P. tricornutum following 24-hour exposure to concentrations starting from 5 mg L−1 [31]. PS NPLs sized at 80 nm within concentrations range from 5 to 50 mg L−1 inhibited algal growth, chlorophyll a level, and Fv/Fm by 27.73%, 29.64%, and 11.76%, respectively in C. pyrenoidosa upon 24 - 48 - hour exposure. However, these effects diminished following an exposure duration of 96 hours. Transcriptomic analysis has demonstrated that PS NPLs inhibited the gene expression of aminoacyl tRNA synthetase and the synthesis of related enzymes and proteins at low concentration (10 mg L−1), while at high concentration (50 mg L−1), they altered DNA damage repair and hindered photosynthesis [40]. A 28 - day exposure of green alga Chlorella vulgaris to carboxyl - functionalized and non - functionalized PS NPLs sized at 20 and 50 nm, at a concentration of 250 mg L−1, exhibited reduced algal cell viability and chlorophyll a concentration, increased cell size, deformed  cell wall, and increased  volume of starch grains [41]. When Euglena gracilis was exposed to 100 nm PS NPLs at 50 mg L−1, the inhibition of its growth reached 35.5% over a 96-hour period. This effect surpassed the impact of 5 μm PS NPLs (27.9%) within the same exposure duration. Both sizes of PS NPLs significantly decreased pigment contents, altered superoxide dismutase (SOD) and peroxidase (POD) activities, and dysregulated the expression of genes related to cellular processes, genetic information processing, organismal systems, and metabolisms [42]. On the contrary, 5 μm PS MPLs at 1 mg L−1 showed more pronounced growth inhibition of E. gracicis compared to 100 nm PS NPLs during 96-hour exposure  [43]. Positively charged PS-NH2 NPLs sized at 50 nm induced growth inhibition in the cyanobacterium Synechococcus elongatus with a 48 h EC50 of 3.81 mg L−1. The primary mechanism of toxicity  included oxidative stress, disruption of glutathione metabolism, and impairment of membrane integrity [44]. Conversely, negatively charged PS-NH2 NPLs at a size of 500 nm and a concentration of 2.5 μg mL−1 significantly decreased cellular esterase activity and neutral lipid content, suggesting an adaptive response in cellular energy metabolism to stress [45]. Exposure of C. reinhardtii to increasing concentrations ranging from 5 to 100 mg L−1 of 300–600 nm PS NPLs resulted in adhesion of the NPLs to the surface of microalgae, causing decrease of  chlorophyll a fluorescence yields and photosynthetic activities, and  membrane damage [46]. Exposure to Pd - doped PS NPLs reduced the growth of both the filamentous cyanobacterium Anabaena sp. (72 h EC50 of 151.3 ± 22.5 mg L−1) and the green alga C. reinhardtii (72 h EC50 of 247.8 ± 32.7 mg L−1), demonstrating the higher sensitivity of the cyanobacterium to PS NPLs compared to the green alga. Both algae exhibited a dose-dependent overproduction of ROS, membrane damage, and metabolic alterations [47]. A 96 - hour exposure to PMMA led to a species - specific reduction in the growth rate of four marine microalgae: Tetraselmis chuii (EC50 of 132.5 mg L−1), Nannochloropsis gaditana (EC50 of 116.5 mg L−1), Isochrysis galbana (EC50 of 123.8 mg L−1) and Thalassiosira weissflogii (EC50 of 83.4 mg L−1) [11]. Moreover, both PMMA and PMMA-COOH NPLs triggered an excess production of pigments, compromised membrane integrity, hyperpolarization of the mitochondrial membrane, enhanced ROS production, increased lipid peroxidation, reduced DNA content, and diminished photosynthetic capacity in the red marine alga Rhodomonas baltica. The interaction with cell walls suggested that PMMA NPLs might lead to the creation of small holes in the lipid layer, potentially resulting in the permeabilization and internalization of minute PMMA aggregates [11]. Exposure to PMMA-COOH NPLs  led to reduced algal growth, attributed to alterations of the cell cycle, resulting in decreased cell viability, metabolic activity, and photosynthetic performance [48]. PS-NPLs were internalized in the Anabaena sp., triggering an excessive generation of ROS, lipid peroxidation, membrane disruptions, intracellular acidification, and a decline in photosynthetic activity. When exposed in combination with poly(amidoamine) dendrimers of generation 7 (G7), the cellular internalization of PS decreased, subsequently mitigating their adverse effects [49]. A significant rise in the teratological frequency was observed in the diatom Cocconeis placentula upon exposure to 0.1 μg L−1 of poly(styrene-co-methyl methacrylate) (P(S-co-MMA)) with an average size of 425.70 ± 175.02 nm. However, no discernable impact on diatom growth was observed across the concentration ranges of 0.1, 1, 100, or 10,000 μg L−1 for 28 day  period  [50]. Overall, these examples highlight the importance of the composition, size, and surface functionalization of NPLs in their interactions with phytoplankton species, underscoring their species specificity. PS NPLs often demonstrate higher toxicity, whereas PE and PMMA NPLs show lower effects [22]. Previous research has stressed the central role of the particle size in biological interactions and the physicochemical behavior of plastics in the environment [57]. 
Exposure to NPLs has been found to influence the production of specific toxins by algae. For instance, both 100 nm and 1 μm-sized PS inhibited growth, photosynthetic parameters, nutrient uptake, and extracellular carbonic anhydrase activities (CAext) in harmful algal blooms, specifically affecting the dinoflagellate Alexandrium tamarense. Notably, the inhibitory effects were more pronounced with exposure to 1 μm - sized NPLs compared to 100 nm PS NPLs exposure [54]. While 100 nm PS NPLs increased the concentrations of intracellular paralytic shellfish toxins in this alga, 1 μm PS NPLs exposure reduced these toxin levels [54].  As PS NPLs concentration increased from 25 to 100 mg L−1 , it promoted the production and release of microcystine in M. aeruginosa, resulting in concentrations of 199.1 for intracellular microcystine and 166. 5 μg L−1 for extracellular microcystine concentrations at 100 mg L−1 PS NPLs [55].  Moreover, exposure to NPLs affected EPS production, leading to increased EPS release and changes in the protein-to-carbohydrate ratio. These effects were observed in marine phytoplankton species like T. pseudonana, Skeletonema grethae, P. tricornutum, and D. tertiolecta when exposed to 1 mg L−1 PS NPLs [56]. Carboxyl - functionalized and non-functionalized PS NPLs sized at 20 and 50 nm adhered to the cell wall of C. vulgaris, triggering the generation of a significant amount of cellular EPS  [41]. 
New evidence has emerged, demonstrating the impact of secondary NPLs to phytoplankton species. In a 48-hour exposure to 1 μg L−1–10 mg L−1 of reference polyethylene (PE) or NPLs derived from PE collected in the North Atlantic gyre (PEN), two microalgae, the green alga Scenedesmus subspicatus and the diatom Thalassiosira weissflogii were studied. Interestingly, while this exposure had no noticeable effect on the cell growth of T. weissflogii, it led to the growth inhibition of S. subspicatus upon PEN exposure [51]. The nano-sized fraction (< 100 nm) resulting from the degradation of polycaprolactone (PCL-plastics + PCL oligomers) prompted an overproduction of ROS, altered intracellular pH and impacted metabolic activity in the filamentous nitrogen-fixing cyanobacterium Anabaena sp., and the unicellular cyanobacterium Synechococcus sp. Moreover, it hindered nitrogen fixation in Anabaena sp. [52]. Secondary NPLs originating from the degradation of polyhydroxybutyrate (PHB) significantly reduced the growth of both Anabaena sp. and C. reinhardtii by 90 and 95%, respectively. These secondary NPLs increased intracellular ROS levels and induced membrane damage with more pronounced effects observed in Anabaena sp. [53].

3. Effect of Phytoplankton Species on Nanoplastics 

Phytoplankton species could influence the NPLs fate in the aquatic environment and thus their impact via release of biomolecules such as extracellular polymeric substance and via cellular transformations. Recent reviews have shed light on how phytoplankton can influence the fate and biological availability of NPLs by excreting diverse EPS, leading to the formation of the eco-corona [58][59]. Studies indicate that  EPS  produced by the marine phytoplankton play a role in forming an eco-corona arround various NPLs, thereby influencing their reactivity [58]. For example, the EPS derived from the diatom P. tricornutum, containing proteins with molecular weight ranging from 30 to 100 kDa along with high molecular weight carbohydrates, formed an eco-corona on 60 nm-sized PS-COOH NPLs, effectively reducing NPLs’ aggregation [60]. However, when EPS from P. tricornutum, Ankistrodesmus angustus, and Amphora sp. interacted with 23 nm PS NPLs, it led to the formation of gel-like micrometer aggregates, which was presumably driven by hydrophobic interactions [61]. The formation of the eco-corona has been found to depend on NPLs size, charges, and incubation duration [58]. Alginate, used as a model polysaccharide, formed an eco-corona on amidine functionalized PS NPLs, altering the surface charge, although aggregation was minimal [62][63]. Furthermore, aminated, carboxylated and plain NPLs aged in EPS reduced the oxidative stress and mitigated toxic effects in the marine alga Chlorella sp. [64].

The capacity of phytoplankton species, such as diatoms and cyanobacteria, to transform  NPLs has also been documented. Observations include the biodegradation of plastic materials through processes like fouling, corrosion, hydrolysis, and penetration, as along with the degradation of leaching components and the diffusion of pigment coloration into the polymers [65][66]. Furthermore, the activity of the algal PETaze enzyme in C. reinhardtii when acting on PET plastic demonstrated the potential for biological degradation with a high conversion rate [67]. However, to further understand the role of phytoplankton on NPLs degradation and transformations, as well as its environmental significance  in-depth research in the future is highly sought.

Overall, the accumulating evidences obtained from model NPLs demonstrated that these particles interact with phytoplanktonic organisms, potentially causing harm when present in concentrations significantly higher than those typically found in aquatic environments. However, further investigations are indispensible to understand the intricate interplays between phytoplankton species and more realistic nanoplastic materials, such as secondary NPLs and aged NPLs, at concentrations closer to those anticipated in aquatic settings. Enhanced and quantitative understanding of the fundamental processes governing the interactions between NPLs and phytoplankton species is pivotal for a comprehensive grasp of their potential impacts on aquatic ecosystems. Notably, phytoplankton play a critical role in global elemental cycling, contributing to nearly half of the global primary production, and occupies a foundational position at the base of aquatic food chains. 

This entry is adapted from the peer-reviewed paper 10.3390/microplastics2040029

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