Bioavailability of inorganic nanoparticles to planktonic bacteria and aquatic microalgae in freshwater

Abstract

Over the past few years, engineered nanomaterials (ENMs) have penetrated nearly every sector of modern life and their broad-scale use is steadily and rapidly increasing. The (expected) elevated levels of ENMs in the environment raise concerns with regard to their potential environmental impact, but environmental risk assessment of released ENMs lags behind invention and today's global consumption volumes. Although considerable progress has been achieved in understanding particle behavior in complex systems and numerous studies have investigated the environmental hazards of ENMs in recent years, the link between these two aspects is less developed. This review provides an overview of what is known about ENMs in freshwater systems and explores the applicability of the bioavailability concept known from aquatic trace metal toxicology. The concept of bioavailability may provide a useful framework to link the “chemical and physical speciation” of ENMs with their possible biological effects but likely requires some ENM specific adaptations. However, there are still considerable knowledge gaps with respect to ENM “speciation” in natural aquatic systems and it remains unclear if it is realistic (by analogy to free metal ions) to search for a specific ENM form that could be used as a measure of biological reactivity. Major knowledge gaps concern the effects of agglomeration on bioavailability, cellular internalization routes, intracellular compartmentalization as well as dissolved organic matter–protein competition on the surface of internalized ENPs.

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Nadia von Moos

Nadia von Moos obtained her Bachelor's degree in biology from the University of Basel in 2007, where she also graduated from the Master's program in Sustainable Development, with a major in ecotoxicology in 2010. She conducted her Master's thesis on the cellular effects of microplastics in the marine environment at the Alfred Wegener Institute for Polar and Marine Research in Bremerhaven, Germany. She is presently doing a PhD thesis on nanoparticles and their effects on aquatic microorganisms at the University of Geneva.

Paul Bowen

Dr. Paul Bowen has a BSc in Physics from Imperial College (UK) and obtained his Ph.D. in Physical Chemistry from the University of Cambridge (UK) in 1982. He worked at the BP Research Centre (Sunbury, UK) for 4 years in applied surface sciences before moving to Switzerland and EPFL in 1987. He has been at the Powder Technology Laboratory in the Materials Institute since its conception in 1988. He has over 160 publications and has co-authored an undergraduate textbook on ceramics. His research interests are powder synthesis, powder characterisation, colloidal processing and atomistic modelling of surfaces and interfaces.

Vera I. Slaveykova

Dr. Vera I. Slaveykova is a professor of environmental biogeochemistry and ecotoxicology at the University of Geneva and the director of the Institute F.A. Forel. Her primary research interests are in the development of new tools and concepts to study the basic processes governing the behavior of trace elements and nanoparticles and their interactions with various biotic and abiotic constituents of the aquatic systems that are highly relevant to water quality and environmental risk assessment. She is an editor of the Environmental Science and Pollution Research journal.

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Nano impact

This review provides a thorough overview of what is currently known about environmental transformations of nanomaterials as well as their interactions with and their toxicity towards bacteria and microalgae, two principal model microorganisms of aquatic risk assessment. In this way, we hope to provide a comprehensive synopsis that can serve as a practical reference study to help guide future research efforts in nanomaterial hazard and risk assessment.

Introduction

Biological availability (bioavailability) is defined as “the extent to which a contaminant in a source is free for uptake by an organism and to which it can cause an effect at the site of action”^1,2 (Box 1). It is considered a key concept in aquatic ecotoxicology and emerged from the recognition that not all metal forms present in natural waters are available to aquatic biota – it was, for example, found that total or dissolved metal concentrations are usually not good predictors of the acute toxicity of metals to aquatic biota.³ Indeed, a basic premise in aquatic ecotoxicology is that toxicants must first be accumulated above normally regulated levels by the organism in which an adverse effect is elicited. In other words, biota only respond to metals which become associated either in or on them. This association between metals and organisms is fundamental to the bioavailability concept and understanding this process is critical to the evaluation of potential adverse effects. The bioavailability concept has been adopted in the regulatory context for trace metals^4–8 and enables the quantitative link between chemical speciation and biological effects, as well as the integration of the modifying role of basic water quality parameters, such as water hardness ions, pH, dissolved organic matter and salinity.

Conversely, though there is an ever growing understanding of the physical and chemical transformations and toxic effects of engineered nanomaterials (ENMs, definition in Box 1), a general conceptual framework linking ENM behavior and effects as well as the influence of environmental parameters thereon is still lacking. As for trace metals, the bioavailability of ENMs could provide a conceptual framework for the integration of their environmental transformations, the influence of modifying factors and possible toxic effects and thus provide a means to understand and predict their potential (eco-)toxicity.^9–13 However, it is currently unclear to what extent this concept could apply to inorganic ENMs and what specific adaptations are likely required.

This review provides an overview of the main processes underlying the availability of inorganic ENMs to unicellular microorganisms, explores how physical and chemical transformations affect their toxicity towards bacteria and microalgae, two very common model aquatic microorganisms (AMOs) of (eco-)toxicology, and examines the applicability of the bioavailability concept in understanding ENM toxicity. The main topics covered are (i) direct interactions between inorganic metal and metal oxide ENMs and AMOs, adsorption to biosurfaces, penetration of cellular barriers and intracellular fate, (ii) environmental transformations of ENPs that affect physical and chemical “speciation” of ENMs in contact with biota and their bioavailability, such as agglomeration, dissolution, binding to dissolved organic matter, sulfidation and oxidation.

1. Bioavailability of engineered nanomaterials

In order to induce an effect in AMOs, ENMs suspended in the water column must first translate from the bulk medium to the immediate vicinity of a microorganism via diffusion (mass transport, Fig. 1). ENMs that agglomerate and reach a critical threshold size high enough for gravitational forces to outweigh buoyancy forces²² will settle and thus decrease the particle number that can encounter planktonic algal or bacterial cells. During mass transport, ENMs are not necessarily inert and undergo physical and chemical transformations.^23–27 Thus, ENMs will interact with AMOs in their environmentally transformed forms rather than in their original, as-manufactured or as-released forms.^21,28,29 ENMs interact with biological surfaces via adsorption and desorption processes either as released ions, single nanoparticles, agglomerates, or complexes (Fig. 1). At the nano–bio interface, target AMOs can modify interactions by their surface properties (e.g. by the presence of exopolymeric substances, surface molecules etc.) and by the release of small proteins and other exudates.³⁰ Dynamics at this interface stem from patchiness of the distribution and spatial localization of surface charge or embedded biomolecules (e.g. proteins, lipids, and glycosylated structures), which undergo constant flux due to normal cellular metabolism (e.g. protein turnover, ion transport, secretion) and reactions in response to environmental fluxes.¹³ The bioavailability of ENMs is thus ultimately shaped by the sum of the above processes (Fig. 1) occurring successively or simultaneously^31,32 in the inherently complex, heterogeneous and highly dynamic natural aquatic systems, which contribute yet another degree of complexity by their high spatial and temporal variability.^18,21

Fig. 1 Processes at the medium–bio-interface underlying the bioavailability of engineered nanomaterials to aquatic microorganisms. Engineered nanomaterials in the environment undergo chemical (e.g. dissolution, sulfidation, redox surface reactions, (de-)protonation, ligand exchange, photodegradation, complexation etc.) and physical (homo- and heteroaggregation) transformations as a function of their material properties and the abiotic factors of the ambient medium. They interact with biological surfaces by adsorption and desorption, either as released ions, transformed single nanoparticles, agglomerates or complexes. Cells can modify these interactions by their specific cell surface properties (e.g. by the presence of extracellular polymeric substances etc.) or by the release of small molecules/proteins. Following adsorption, engineered nanomaterials may, depending on their material properties and the biological target organism, actively or passively be internalized by cells and induce biological responses, which however can also be independent of direct interactions and internalization. All processes are highly dynamic. Adapted from ref. 33 and 13, not to scale.

The quantitative understanding of ENM bioavailability requires insights into their behavior during transport from the ambient medium to the AMO interface and of the processes underlying adsorption, internalization as well as intracellular fate, which will be discussed in the next sections. Although these processes depend on the ENMs' physical and chemical properties such as size, shape, surface properties, energy bandgap etc., the emphasis of this review lies on the environmental factors affecting the above processes, since the importance of ENMs' material characteristics has already been thoroughly reviewed.^16,17,34 For more detailed information concerning these processes in wastewater systems, the authors direct readers to recent reviews by Eduok et al.,³⁵ Liu et al.³⁶ or Brar et al.³⁷

2. ENM uptake by planktonic bacteria and microalgae

ENMs undergo environmental transformations that yield a complex mix of different ENM forms, comparable to the speciation of trace metals in the aquatic environment, which include homo- and hetero-agglomerates, free/complexed dissolved and partially dissolved ENMs. This section is devoted to the uptake of (undissolved) ENMs, which is defined as a process involving adsorption and internalization, as well as to the intracellular fate of internalized ENMs. The uptake of dissolved, inorganic species (ions) is well known from research on trace metals.³

2.1 Adsorption

Adsorption and direct contact of ENMs with AMOs are a prerequisite for toxicity^38–42 and the penetration of biological membranes. Adsorption of ENMs onto surfaces is governed by the combined effects of the nature and physicochemical properties of the ENMs and the ambient media.⁴³ Adsorption and internalization also strongly depend on the biological substrate, i.e. the cell type, cell membrane, differentiation stage and the pathways of cellular processing.¹³ The forces involved in adsorption are van der Waals forces, hydrophobic forces, electrostatic attraction and specific chemical interactions, including hydrogen bonding and receptor–ligand interactions.⁴⁴ Microorganisms in general,⁴⁵ particularly bacteria and microalgae, normally exhibit a net negative surface charge at freshwater pH values due to the dissociation of carboxylic, phosphate and other acidic functional groups of the membrane.^38,46,47 Electrostatic attraction is therefore likely to occur between AMOs and ENMs with a positive zeta potential, be it inherent or the result of previous modifying interaction processes at the particle surface.^38,47 Indeed, the importance of electrostatic attractions has, for example, been evidenced by interactions between positively charged nano-CeO₂ (at neutral pH) and outer membranes of Escherichia coli.⁴⁸ However, the simple electrostatic approach sometimes fails to explain direct interactions between ENPs and AMOs. This was, for instance, the case for nano-TiO₂ (−25.72 ± 0.5 mV) and the bacterial freshwater isolate Bacillus licheniformis (−28.47 ± 0.62 mV) whose negative zeta potentials in lake water seemed to preclude nano–bio interactions based on the electric charge but whose mechanism of toxicity under dark conditions was predominantly found to be due to the attachment of nano-TiO₂ to the bacterial cell membrane.⁴⁹ Electrostatic interactions were also deemed to be of secondary importance when no difference in the adhesion of more (carboxyl-PEG-QDs) or less (amine-PEG-QDs) negatively charged CdSe/ZnS quantum dots (QDs) onto the negatively charged bacterium Cupriavidus metallidurans CH34 was observed.²⁹ Negatively charged nano-Ag interacted with membranes of negatively charged E. coli cells⁵⁰ and un-capped nano-TiO₂ induced toxicity in the negatively charged bacteria, regardless of their own negative surface charge at physiological pH.³⁸ Hence, the possibility of other mechanisms than electrostatic forces governing the adhesion of ENPs to cell surfaces has been suggested.⁵¹ Nonetheless, electrostatic attraction cannot generally be ruled out.

Adsorption of ENMs onto unicellular freshwater microalgae (Table 1) has been shown for nano-Au and the microalga Scenedesmus subspicatus⁵² as well as for SiO₂ and CeO₂ and the microalga Pseudokirchneriella subcapitata.^39,53,54 According to Aruoja et al.,⁵⁵ the same microalga, P. subcapitata, adsorbed 2.3 times its own weight of nano-TiO₂ and the adsorption kinetics depended on pH, with the adsorption maximum at pH 5.5.⁵⁶ Similarly, nano-TiO₂ was shown to adsorb onto the cell walls of the microalga Chlamydomonas reinhardtii⁵⁷ and so did nano-Ag⁵⁸ and CdSe/ZnS QDs.⁵⁹ Furthermore, negatively charged carboxyl CdSe/ZnS QDs were adsorbed onto cells of the wall-less strain but not onto wild-type C. reinhardtii cells.⁶⁰

Table 1 Selected examples showing adsorption and internalization of ENPs by microalgae

ENP (size)	Organism	Location	Uptake mechanism	Analytical technique	Ref.
Ag (1–35 nm)	Ochromonas danica	Inside cells	Suggested: increased membrane permeability	TEM and STEM	61
Ag (29 nm)	Chlamydomonas reinhardtii	Inside cells	NA	HR-ICP-MS	58
Au (10 nm, amine-coated)	Scenedesmus subspicatus	In the cell wall	NA	TEM	52
FITC-mannose generation 0 (G0) glycodendrimer-coated Au (2 nm)	Wt and wall-less Chlamydomonas reinhardtii	On the cell wall of wt and inside wt and wall-less strains	NA	Fluorescence confocal microscopy and flow cytometry	62
CdTe/CdS quantum dots (core 3–4 nm, 5.7 ± 0.4 nm, 4.3 ± 0.4 nm, 4.2 ± 1.2 nm)	Chlamydomonas reinhardtii	Cd inside cells	NA	Graphite furnace atomic absorption spectrometry and ICP-MS	63
Carboxyl CdTe/ZnS quantum dots	Wall-less Chlamydomonas reinhardtii	Association with cell	NA	ICP-MS and flow cytometry	60
CuO, bare and polymer-coated (30–40 nm/148 nm, 65.4 nm)	Chlamydomonas reinhardtii	In cytoplasm	NA	TEM and ICP-AES	64
CuO (30–40 nm)	Chlamydomonas reinhardtii	Inside cells	Suggested: endocytosis, carrier proteins or ion channels	TEM	65
TiO2 (21 nm)	Chlamydomonas reinhardtii	In the cell wall, plasma membrane and cytoplasm	NA	SEM, TEM	57

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The adsorption of ENPs onto bacterial cells (Table 2) has, for instance, been shown for nano-CeO₂ and the bacteria E. coli and Synechocystis,^48,66 nano-Ag and laboratory-grown P. putida biofilm bacteria,⁶⁷ CeO₂ and the cyanobacterium Anabaena CPB4337,⁵⁴ Fe⁰ and the cyanobacterium Microcystis aeruginosa,⁶⁸ TiO₂ and B. licheniformis,⁴⁹ TiO₂ and Al₂O₃ and C. metallidurans and E. coli,⁶⁹ as well as for Al₂O₃, SiO₂, TiO₂, ZnO and the three bacteria B. subtilis, E. coli and Pseudomonas fluorescens.⁵¹ The same was observed for nano-Ag and the bacteria E. coli, V. cholerae, P. aeruginosa and S. typhus, which also entered the bacterial cells.⁴⁷ Based on the observation that internalized nano-Ag were approximately the same size as those adsorbed onto the bacterial membrane, it was suggested that uptake may occur via changes in membrane permeability induced by interactions with the ENPs.⁴⁷ As mentioned earlier, it is therefore very likely that interaction and adsorption are preconditions for the internalization of undissolved ENPs into bacterial cells. This presumption is supported by the hypothesis that adsorption may induce cell wall pitting⁵⁰ or membrane damage, e.g. perforation,^70,71 and thereby increase membrane permeability.⁴⁴

Table 2 Selected examples showing adsorption and internalization of ENPs by bacteria

ENP (size)	Organism	Location	Uptake mechanism	Analytical technique	Ref.
Ag (12 nm)	Escherichia coli	In the membrane structure and inside cells	Suggested: increased membrane permeability	SEM, TEM and EDAX	50
Ag (16 nm)	Escherichia coli	On the cell membrane and inside cells	Suggested: increased membrane permeability	High angle annular dark field STEM and TEM	47
Ag (10–15 nm)	Escherichia coli	On the cell wall and inside the cell	NA	TEM	70
Ag (30–50 nm)	Pseudomonas putida biofilms	On and inside cells	NA	TEM, graphite furnace atomic absorption and ICP-MS	67
Ag (104–118 nm)	Escherichia coli and Bacillus subtilis	Inside cells	NA	High resolution microscopy, TEM, TEM-EDS	72
CuO (<5 nm)	Cyanobacterium Microcystis aeruginosa	Inside cells	Cell wall pores, endocytosis	FAAS, HRTEM, SEM, TEM and EDS	73
TiO2, Al2O3 (12–707 and 13, respectively)	Cupriavidus metallidurans and Escherichia coli	In the periplasmic compartment	NA	TEM and STEM	69
TiO2 (192 nm)	Cyanobacterium Anabaena variabilis	Inside cells	NA	HRTEM and Raman imaging	71
TiO2 (20–100 nm)	Bacillus licheniformis	Outer and inner sides of the plasma membrane and inside cells	Increased membrane permeability	SEM, TEM, FTIR and ICP-OES	49
Carboxyl-CdSe/ZnS core/shell QDs (12.9 nm)	Cupriavidus metallidurans	In periplasmic space	Loss of membrane integrity	TEM, EDS, AFlFFF and ICP-MS	74
ZnO (ca. 7, 260 and 800)	Escherichia coli and Staphylococcus aureus	On/in membranes and in the cytoplasm	NA	TEM, HRTEM, EDS	75
ZnO (30 nm), TiO2 (50 nm)	Escherichia coli	On the cell wall and inside cells	NA	TEM, SEM	76
ZnO (30 nm), TiO2 (50 nm)	Salmonella typhimurium	Smaller ENPs inside cells, larger ones on the membrane	NA	TEM and flow cytometry	77

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Overall, the adsorption of ENPs onto biosurfaces has been shown for various planktonic bacteria and microalgae, although the underlying mechanisms and modifying factors are not yet well understood. Electrostatic attraction is the most obvious explanatory factor for surface adsorption but does not explain all direct interactions. Other factors may come into play. Molecular dynamic simulations have shown that the adsorption of polymers and dispersants is sensitive to the affinity of the surface for water.⁷⁸ This has recently been illustrated for the adsorption of polyacrylic acid (PAA) and polyaspartic acid (p-ASP) on calcite surfaces⁷⁸ and the attachment of small peptides onto an amorphous TiO₂ surface.⁷⁹ The different conformations of PAA and p-ASP lead to different disruptions of the ordered water layer at the calcite crystal surface and thus contributed to different adsorption free energies, which resulted in a more rapid adsorption and longer residence times for p-ASP at the calcite surface.⁷⁸ Also, the adsorption geometry of peptides was found to be directly influenced by local density changes in the water structure on the Ti or Si surface, which resulted in subtle differences in the adsorption configurations of different amino acid side chains (flat vs. upright). Calculations of adsorption free energy profiles revealed a correlation between the adhesion forces and the nanoscale features of the water structure at the solid/liquid interface. The authors concluded that electrostatic interactions are the major driving force behind approaching surfaces of opposite charge density but are of secondary importance for adhesion forces.⁷⁹

Furthermore, the kinetics of the adsorption process is also largely unknown. The kinetics of particle adhesion (and internalization) with respect to the diffusion flux from the bulk solution to the cell surface can indicate the limiting factor of the bioavailable fraction (and thus of the observed effects). In the case where mass transfer is slower than the adsorption process, bioavailability will primarily be determined by the diffusion flux, which depends on the concentration of the contaminant. In the inverse case, bioavailability will largely depend on cell properties governing adsorption (i.e. ligand binding etc.), in which case surface properties of ENPs (such as coatings, adsorbed DOM layer etc.) and cells (such as cell wall structure, extracellular polymeric substances etc.) will likely play an important modifying role. For example, it has been suggested that the reduced cytotoxicity (and thus bioavailability) of polyelectrolyte/NOM-coated nano-ZVI towards E. coli cells may be explained by DOM preventing direct physical contact with cells.⁸⁰ In bacteria, it has variously been shown that extracellular polymeric substances (EPS) can exert a protective function by decreasing ENP bioavailability by fending off ENP adsorption and toxicity. This was, for example, the case for Synechocystis cyanobacteria, which were able to prevent the adsorption of CeO₂ ENPs onto their outer membrane by trapping the ENPs with extracellular polysaccharides, much in contrast to E. coli that does not excrete EPS.⁸¹Synechocystis thereby protected itself from direct toxic effects mediated by ROS, which require a close contact between ENPs and cell membranes, a finding also supported by a more recent study on Synechocystis and TiO₂ ENPs.⁸² However, the extracellular polysaccharide layer of Synechocystis did not provide sufficient protection from indirect toxicity mediated by released toxic ions originating from CeO₂ ENPs.⁸¹ The protective function of EPS has also been demonstrated for the bacterium Pseudomonas chlororaphis exposed to nano-CuO.⁸³ Furthermore, engineered E. coli and S. meliloti overproduced EPS in reaction to nano-Ag exposure and thereby protected themselves from toxicity by trapping ENPs with EPS outside the cell.⁸⁴

However, to our knowledge, no studies have directly addressed the kinetics of adsorption onto AMOs with respect to bioavailability (and toxicity) so far. Further progress is probably limited by the available techniques, since most direct evidence comes from imaging techniques, which require important sample preparation and are not well suited for kinetic studies in liquid media.

2.2 Internalization

The internalization of ENPs by bacteria and microalgae has variously been described (Tables 1 and 2), disproving the common hypothesis that ENMs cannot penetrate the protective, relatively thick cell walls of bacteria and microalgae^12,71 without prior damage. For example, nano-TiO₂ localized in the cell wall and cytoplasm of the microalga C. reinhardtii⁵⁷ and inside cells of the nitrogen-fixing, prokaryotic blue-green alga Anabaena variabilis.⁷¹ Gold ENPs penetrated both the wild-type and the cell wall-deficient mutant C. reinhardtii strains⁶² and adsorbed onto the cellulosic layer of the cell walls of the microalga S. subspicatus but did not enter its cytoplasm.⁵² Bare and polymer-coated nano-CuO also penetrated cells of the microalga C. reinhardtii, and the higher toxicity of the polymer-coated nano-CuO was attributed to their greater cell-penetrating capacity.⁶⁴ Another study on C. reinhardtii suggested, based on the measured intracellular Ag, that the number of bioavailable nano-Ag was limited compared to cationic Ag.⁵⁸ However, nano-Ag was readily internalized and accumulated by the mixotrophic freshwater alga Ochromonas danica.⁶¹ Furthermore, Ag ENPs penetrated Gram-negative E. coli cells^47,72 and the Gram-positive B. subtilis.⁷² ZnO and TiO₂ ENPs were observed inside E. coli.⁷⁶ Similarly, CuO ENPs were internalized by the cyanobacterium M. aeruginosa and uptake was enhanced by the presence of SRFA.⁷³ The internalization of ZnO was observed in both the Gram-negative bacterium E. coli as well as the Gram-positive bacterium Staphylococcus aureus and was found to be size-dependent, i.e. smaller particles were internalized to a greater extent than larger ones.⁷⁵ Maybe less surprisingly (considering their smaller size range), CdTe/CdS QDs were also taken up by C. reinhardtii cells, but the observed bioaccumulation, in contrast, was greatest for the largest (orange) QDs.⁶³ However, above observations should be considered cautiously as the techniques for the investigation of ENM uptake have several limitations. While visual techniques provide qualitative information on subcellular localization, techniques based on mass spectrometry provide quantitative data on uptake but cannot differentiate between ionic and particulate forms. Flow cytometry can also furnish quantitative information on uptake but cannot necessarily distinguish between uptake and surface-associated ENMs. Although there is evidence for the internalization of ENMs in AMOs such as algae and bacteria, the underlying mechanisms are neither understood nor are the uptake rates known.

Nonetheless, the tough, more or less tensile and rigid, ca. 20 nm thick⁵⁹ cell walls of bacteria and microalgae form efficient first-tier barriers that screen the few nanometer thick cell membranes enwrapping the cytoplasm from the environment.⁴⁴ Cell walls are thus an important modulating factor of nano–bio interactions as demonstrated by a study on CdSe/ZnS QDs and the microalga C. reinhardtii, in which an association of QDs with the wall-less mutant strain was observed but not with the walled wild type.⁶⁰ Similarly, the cell wall-deficient marine microalga D. tertiolecta was more sensitive to nano-Ag induced ROS and lipid peroxidation than the freshwater microalga C. vulgaris.⁸⁵ Notwithstanding the generally suggested protective function of cell walls, it has also been shown that algal (C. reinhardtii) and bacterial (Vibrio fischeri) cells were more sensitive to nano-Au than two mammalian cell lines.⁸⁶ Cell walls contain pores with diameters ranging 5–20 nm.^87,88 These pores are potential entry ports for small ENMs, which may then become accessible for internalization across the lipid bilayer. The permeability of cell walls changes throughout cell cycling especially during the delicate phase of cell division in the course of which cell walls are newly synthesized.^73,89–91 Also, there is the possibility that interactions of cells with ENMs may even induce the formation of larger new pores permeable to bigger ENMs.⁸⁹

Once the cell wall has been penetrated, the cellular internalization of single particles or aggregates occurs through direct ingestion or across epithelial boundaries,⁹²i.e. the semi-permeable phospholipid bilayer or plasma membrane. Stable primary ENPs can theoretically passively diffuse through hydrophobic lipid bilayers, given they are nonpolar, i.e. uncharged and not too large. However, this has, to the best of our knowledge, not yet been shown to occur in bacteria and microalgae.

Theoretically, ENPs may possibly also appropriate membrane fusion as a non-invasive entry route, commonly observed to occur with biomacromolecules or alternatively, if cationic, may also be capable of creating transient membrane holes,⁹³ both of which remain to be shown. An uptake of ENPs via transport proteins for ions is deemed rather unlikely because they are much larger than ions (ranging from 30 pm to ≥200 pm) and therefore most likely do not fit the respective binding sites,⁴³ unless of course, if they dissolve.

Endocytic pathways via membrane-bound vesicles that emerge from inward folded plasma membrane buds, which are then pinched off, have been suggested to be a possible internalization route for ENMs into the cytoplasm. However, prokaryotes do not have endocytic mechanisms for the transport of “bulk” particles across their membranes. Much in contrast to eukaryotes that have evolved highly developed mechanisms for the active internalization of bulk materials via endocytosis⁹² (Fig. 2). Endocytosis has, in some cases, been suggested as the main ENM uptake route into bacteria and microalgae (see Tables 1 and 2), but it remains to be shown whether endocytosis is the predominant uptake pathway of ENMs into AMOs. In most cases, the actual uptake routes are unknown.

Fig. 2 Active and passive cellular uptake pathways for ENMs in eukaryotic cells. Passive uptake occurs via diffusion and facilitated diffusion via transport proteins, i.e. gated channel proteins and carrier proteins. Active uptake pathways involve transmembrane carrier proteins and endocytic pathways including receptor-mediated phagocytosis, clathrin-mediated endocytosis (120 nm, via clathrin-coated pits) and caveolae-mediated endocytosis (60 nm, via lipid rafts), non-specific endocytosis by macropinocytosis and non-clathrin, non-caveolae endocytosis (90 nm, fluid phase). All pathways except caveolae-mediated endocytosis and diffusion merge with the lysosomal degradation system comprising numerous vesicle maturation steps within the cell. A lysosome typically ranges from 0.2 to 0.5 μm in diameter.⁹⁸ Phagocytosis is mediated by specific membrane-receptors that are activated upon contact with a ligand to produce phagosomes (>250 nm). During their maturation process, phagosomes transform into late phagosomes, which fuse with lysosomes to form phagolysosomes. During macropinocytosis, internalization occurs via an unspecific invagination resulting in pinocytic vesicles (<150 nm), which eventually merge with lysosomes. Clathrin-mediated endocytosis and non-clathrin, non-caveolae-mediated endocytosis produce endosomes which evolve into early and late endosomes that mature into acidic lysosomes (pH = 5). Caveolae-mediated endocytosis produces caveosomes that either transfer their contents into the Golgi apparatus, endoplasmatic reticulum (ER) or into the cytosol or may also undergo transcytosis. Adapted from ref. 11, 92, 99 and 100, not to scale.

Ample evidence from mammalian cells suggests that endocytosis is size-dependent.⁹⁴ It has been put forth that the optimal sizes for endocytosis are diameters around 50 nm,⁹⁵ which corresponds well to theoretical model calculations yielding optimal radii between 27 and 30 nm for spherical particles.⁹⁴ Thus, the 1–100 nm scale can be considered the most critical size dimensions for biological interfaces,⁹⁶ especially since it is assumed that the two main endocytic entry routes for ENMs into eukaryotic cells are the clathrin- and caveolae-mediated endocytic pathways.⁹² Simulation results have furthermore suggested that cationic ENPs are more readily internalized than anionic ENPs, which must first overcome an energy barrier to reach the net negatively charged surface of the lipid bilayer.⁹⁷ However, this hypothesis remains to be corroborated by conclusive experimental data for AMOs.

Uptake of dissolved ENMs occurs by the same internalization routes as for trace metals, which principally depends on the metal form (i.e. charge, valency, size etc.) and the biological substrate (i.e. availability of pathways). These include active and passive transport mechanisms via (non-)specific transporter (carrier) proteins and channels, co-transport, endocytosis or simple diffusion.^1,3

Hence, ENPs of different chemical composition and various sizes (Tables 1 and 2) are capable of penetrating the cell walls and plasma membranes of Gram-negative and Gram-positive bacteria as well as microalgae. However, it seems that in at least some cases membrane damage occurs during the process and it remains to be shown if this is size or particle related. Furthermore, excretion of ENMs can theoretically occur by exocytosis,⁴³ but we have found no experimental accounts of this mechanism for microalgae and bacteria.

2.3 Intracellular fate

Understanding the intracellular fate of ENMs is critical to the quantitative link of internalized ENMs to their toxicity, since biological effects are a complex function of a contaminant's cellular concentrations, intracellular distribution and speciation. ENMs internalized by AMOs reach the cytoplasm where they potentially interact with cellular organelles. However, data concerning their intracellular distribution and localization in cells and organisms of ecotoxicological relevance are sparse, apart from a handful of isolated accounts. For example, nano-CuO in C. reinhardtii cells formed large agglomerates which mainly localized in the cytoplasm and were likely enclosed by vacuoles.⁶⁴ Nano-TiO₂ accumulated inside its cell wall and plasma membrane.⁵⁷ Similarly, gold ENPs localized in the cellulosic layer of the trilaminar membrane of S. subspicatus microalgal cells.⁵² Nano-Ag was found in irregularly shaped vacuoles of the microalga O. danica and it remained unclear whether Ag ENPs exerted toxicity directly by their presence or indirectly by the release of Ag⁺.⁶¹ Similar observations have been made in bacteria models. For instance, nano-Ag was found in the membrane and inside E. coli.^47,50,70 TiO₂ and Al₂O₃ ENPs adsorbed onto the surfaces of C. metallidurans and E. coli and accumulated in the periplasmic compartments of both species.⁶⁹ Similarly, CdSe/ZnS functionalized QDs adsorbed onto the bacterium C. metallidurans and accumulated in its periplasmic space.⁷⁴ In the Gram-positive bacterium B. licheniformis, nano-TiO₂ accumulated on the outer and inner sides of the plasma membrane and in the cytoplasm.⁴⁹ ZnO ENPs were detected in the membranes and cytoplasm of E. coli and S. aureus.⁷⁵ Another study described the uptake of ZnO and TiO₂ ENPs into E. coli cells and observed a uniform distribution of smaller ENPs in the cell and the adherence of agglomerated particles onto the cell wall⁷⁷ (Table 2).

Intracellular localization is largely determined by the chemistry of the biological fluids⁹² and the dynamic biopolymer (e.g. protein) corona on the particle surface.¹³ Unfried et al.⁹⁹ emphasize the importance of clarifying the relation between uptake and subsequent cellular distribution patterns in future research. ENMs in physiological fluids such as cytoplasm or interstitial fluids are immediately coated by biomolecules and proteins that form a dynamically changing biopolymer corona of partial (to, rather unlikely, full) coverage on a particle's surface that confers a unique biological identity.^101,102 The currently prevailing hypothesis proposes a highly dynamic corona consisting of a strongly associated hard phase and a more loosely bound soft phase. Polymer-coated ENPs may be more resistant to the formation of a protein corona, especially since coatings often are specifically designed to increase their biocompatibility. For ENPs which have previously acquired a dissolved organic matter (DOM) coating in the environment, the formation of a protein corona will be dominated by the competitive exchange between proteins, DOM and other present coatings for the core surface. However, we are not aware of any studies investigating these interactions.

What we can derive so far is that protein adsorption kinetics (association and dissociation) is governed by competitive binding mechanisms and furthermore by time, particle size (i.e. surface curvature), initial material surface properties, previous surface interactions and by the proteins present in the surrounding media and their respective equilibrium constants.^13,103,104 Protein concentrations and equilibrium binding constants determine the composition of a corona at any given time¹⁰¹ and particle–protein complexes can last microseconds to days.¹³ But it is still unknown how the epitopes of proteins embedded in the corona are affected in their ability to perform their usual biological functions.^103,105 It is however recognized that the protein corona strongly influences all interactions at the particle–bio interface as well as the biological fate and effect of ENMs and that biological responses may either reflect the adsorbed protein corona, the material itself or both.¹⁰¹ Yet, the ramifications for nanoecotoxicology have, to our knowledge, hardly been investigated and research in this area has almost entirely focussed on mammalian model cells.

While the bioavailability (and toxicity) of trace metals is thought to be directly linked to their ability of crossing membranes and is commonly predicted by internal metal concentrations or uptake rates,¹⁰⁶ this relation does not necessarily apply to ENMs in the same way. Firstly, specific uptake rates and bioconcentration factors of ENPs are currently not available for AMOs. Secondly, the current body of nanotoxicological evidence suggests that the bioavailability (and toxicity) of ENMs is possibly independent of internalization and thus more complex. It has, for instance, been shown that ENPs can indirectly trigger adverse biological responses without direct contact or uptake (via ions and ROS) and it has also been shown that internalized ENPs do not cause harm per se.⁴¹ Thus, the actual toxicity mediating the ENM form has so far not yet been explicitly identified. Whether ENP toxicity is ion or particle mediated is still an avidly debated question in nanoecotoxicology and the literature contains copious contradictory findings, which are treated in the section on dissolution.

In the following sections, toxicity will thus be considered as a preliminary proxy for the bioavailability of ENPs to AMOs. Sections three and four are devoted to the effects of agglomeration and dissolution on ENP bioavailability as well as to the modifying environmental factors.

3. ENM agglomeration and bioavailability

ENP agglomeration behavior can largely be explained by fundamental colloid science and there is increasing knowledge of the physical transformations of ENPs in freshwater and how these are affected by water quality parameters.¹⁰⁷ However, there is little knowledge of how agglomeration influences ENP bioavailability (and toxicity); in fact, very few studies actually address the direct relation between ENM agglomeration state and bioavailability to AMOs. Since the role of undissolved, particulate ENMs is still uncertain with respect to ENM toxicity, the role of agglomeration can only be resolved once the bioavailable ENM form responsible for the observed toxic effects has clearly been identified and quantified (i.e. particles vs. dissolved fractions etc.).

Generally, agglomeration is expected to diminish ENM fluxes towards biointerfaces (by increasing hydrodynamic size and decreasing their diffusivity) as well as the number concentration and available surface area and interfacial free energy for adsorption and reaction.^{31,53,108,109} But even this basic premise is still controversial since very few existing studies on AMOs report contradicting findings. The above assumption was, for example, confirmed for the crustacean Daphnia magna (immobilization) exposed to nano-Ag in OECD medium¹⁰⁹ but was refuted by a recent study investigating the effect of agglomeration of nano-CeO₂ on growth of the microalga Pseudokirchneriella subcapitata, which was independent of agglomeration. Similar findings were reported for tungsten carbide, albeit for gill cells of the rainbow trout Oncorhynchus mykiss, in which rapid agglomeration in minimal exposure media did not prevent uptake and subsequent toxicity.¹¹⁰

In addition to particle-specific characteristics, medium pH, ionic strength and DOM concentrations cause changes in the particle surface charge of metal oxides,^22,27 thus affecting interparticle interaction forces and agglomeration behavior.^111,112 Hence, medium properties likely affect ENM bioavailability. For example, Guzman et al.²² pointed out that various materials have pH_IEP values far from the pH of natural systems, e.g. uranium oxides (pH_IEP ∼ 5), iron oxides (pH_IEP ∼ 8), zinc sulfides (pH_IEP ∼ 2) and aluminum oxides (pH_IEP ∼ 9), which, in light of the basic premise discussed above, favor stable suspensions and high fluxes towards biointerfaces. At any given pH, the aggregation of nano-TiO₂ suspensions increased with increasing ionic strength.¹¹² This was also observed for nano-Ag.¹¹³ The influence of the nature of electrolytes present in the aqueous suspensions under investigation was nicely illustrated in a study comparing the effects of increasing Na⁺ and Ca²⁺ concentrations by French et al.¹¹⁴ who observed the agglomeration of nano-TiO₂ in the presence of the divalent cation Ca²⁺ within 5 min, which is significantly faster than the 15 min agglomeration time observed in the presence of monovalent Na⁺ at the same ionic strength (0.0128 M) and pH (4.8). The above non-exhaustive examples illustrate that water hardness (e.g. Ca²⁺) and increased water salinity (e.g. high ionic strength) could favor ENM agglomeration. By the same token, water hardness and other major ionic species are thus expected to mitigate ENM bioavailability and thus protect planktonic bacteria and microalgae from ENM-induced stress, but further experimental evidence is necessary to support such as a general consideration.

The effects of DOM on ENM suspensions are complex and can be difficult to predict.³¹ Depending on DOM composition, pH and ionic strength conditions, DOM can either have a stabilizing effect on suspensions,^24,25,113 favor agglomeration via bridging or even cause de-agglomeration of a given ENM.^{10,23,25,31,53,89,112,115–117} The stabilizing effect of DOM is thought to be due to the combined effects of increased steric hindrance¹¹⁸ and increased electrostatic inter-particle repulsion due to increased particle surface charges resulting from the adsorption of the negatively charged DOM.²⁴ For example, the nano-ZnO agglomerates formed at pH_IEP (pH = 9.3) decreased in size as concentrations of Suwannee River humic acid (SRHA) were increased from 0.1 to 0.5 mg L⁻¹, which is explained by the formation of an adsorbed layer of negatively charged SRHA causing an increase in surface potentials and the resulting electric double layer repulsive energy and steric repulsion.¹¹⁹ Stabilizing effects of DOM have, for example, been also shown for nano-TiO₂ (<25 nm) in lake water,⁴⁹ 15–30 nm sized nano-TiO₂, nano-ZnO and nano-CeO₂ suspensions under many different natural conditions¹²⁰ as well as for TiO₂ ENPs (5 nm) in the presence of SRFA.²⁶ Similarly, bacterial extracellular proteins were found to stabilize nano-Ag.¹²¹ In yet another study, DOM was even found to counteract the agglomeration-inducing effect of increasing ionic strength towards nano-Ag suspensions, producing more stable suspensions.¹²² Chowdhury et al.¹²³ found that both the bacterium E. coli and SRHA significantly stabilized nano-TiO₂ suspensions but that the stabilizing effect of DOM was greater. They also observed less nano-TiO₂ deposition in the combined presence of DOM and E. coli compared to each of the compounds alone, implying highly complex interactions. DOM-induced destabilization has also been variously observed. For example, nano-TiO₂ suspensions were destabilized by the two organic acids, oxalic and adipic acids.¹⁰⁸ DOM-induced de-agglomeration has also been described for iron oxide ENPs at pH 7. It was found that high concentrations of SRHA – as opposed to low SRHA concentrations or none – induced the de-agglomeration of the iron oxide ENPs over time.¹²⁴ In the same way, SRHA caused a partial disaggregation of nano-Ag agglomerates by coating individual ENPs with a thin film, which stabilized the particles by charge and steric effects yielding nicely dispersed, stable suspensions of primary Ag ENPs.⁶⁷ Another study investigating the effects of environmental concentrations of SRHA (0.02 to 0.5 mg L⁻¹) on nano-ZnO suspensions concluded that the SRHA surface coatings on the ZnO ENPs induced the disaggregation of ZnO aggregates and a reduction in their size with increasing SRHA concentrations.¹¹⁹ Environmental concentrations of HAs and alginates were found to not only stabilize TiO₂ suspensions (even at high concentrations) but also cause the disaggregation and dispersion of already formed TiO₂ aggregates.¹¹⁵ Domingos et al.¹¹² observed stable dispersions of TiO₂ ENPs under environmentally relevant conditions (SRFA, pH, I) and inferred that stable TiO₂ dispersions may occur more frequently in natural aquatic environments than predicted. A similar conclusion was also drawn by Cumberland and Lead¹¹³ and Khan et al.¹²¹ from their observations on the behavior of nano-Ag under environmentally relevant conditions, which implied longer residence times and thus higher bioavailability than initially expected. A recent study investigating different divalent cations (Ca²⁺ or Mg²⁺) and pH values with DOM concentrations of European streams demonstrated that the agglomeration behavior of nano-Au was, to a large extent, determined by DOM.¹⁰⁷ Therefore, depending on ENP properties and environmental conditions, certain ENMs form well-dispersed, stable suspensions in the water column. Others will form large particle clusters/agglomerates, which at a critical threshold size, sediment and decrease the particle concentration in the water column, at least temporarily.^125,126 According to the above premise, the former case thus favors high mobility¹²⁵ and fluxes towards microorganisms, while in the latter case, sedimentation leads to a decrease of the number of suspended ENMs available for interaction with planktonic AMOs.

Another important assumption, especially with respect to ENPs whose main bioavailable forms are ionic (see next section), is that agglomeration can minimize dissolution by kinetically interfering with the diffusion process.¹²⁷ This has, for example, been confirmed for the dissolution rate of nano-Ag, which tended to be slower for agglomerated ENPs.^128,129

Apart from these single accounts, the question of how agglomeration relates to the bioavailability of ENMs to AMOs has, to the best of our knowledge, not yet been systematically addressed. The scarcity of articles may possibly also be due to a publication bias in favour of positive (toxic) effects. Bridging the gap between ENP behavior and effects as well as the identification of the critical bioavailable ENM forms thus remains a major research priority.

4. ENM dissolution and bioavailability

On the one hand, dissolution leads to reduced environmental persistence of ENMs and their long-term bioavailability.³¹ On the other hand, the released toxic ions are immediately and easily bioavailable to AMOs and may trigger acute toxic effects.¹³⁰ Thus, dissolution decisively influences ENM bioavailability (and toxicity)^60,89,131 to algae and bacteria, which are thus exposed to complex mixtures of ENMs, partially dissolved ENMs, free/complexed dissolved species and ions adsorbed to ENM surfaces.¹³²

Ion-mediated toxicity of ENPs has been demonstrated by numerous studies, most prominently for nano-ZnO, nano-CuO and nano-Ag.¹³¹ A recent literature review on the ecotoxicity of ZnO ENPs has, for example, concluded that ionic Zn was the main mediator of detrimental effects towards bacteria, algae and plants, aquatic and terrestrial invertebrates and vertebrates.¹² Examples of Zn²⁺ as mediators of nano-ZnO toxicity towards AMOs include the freshwater microalga P. subcapitata,^55,111 the bacteria E. coli, B. subtilis, V. fischeri and P. putida (measured as inhibition of bacterial luminescence (INH), minimal inhibitory concentration (MIC), mortality, growth inhibition, or cell viability).^133–135 Likewise, the solubilized bioavailable fraction of nano-CuO ENPs was linked to the growth inhibition of microalga P. subcapitata.⁵⁵ There is also a wealth of articles describing Ag⁺ ion-mediated antimicrobial effects of nano-Ag.^{47,72,136–143} Recently, Xiu et al.¹⁴⁴ decidedly ruled out direct particle-specific biological effects by showing no toxicity of nano-Ag towards the bacterium E. coli under anaerobic conditions, precluding the oxidation of Ag and thus the release of Ag⁺. The hypothesis of ion-mediated ENM toxicity is further supported by findings showing how organic molecules can decrease the bioavailability of the released ions and thereby mitigate acute, short-term toxicity. For example, the complexing agent tannic acid significantly decreased the toxicity of nano-ZnO towards the bacterium P. putida by decreasing the bioavailable fraction of dissolved Zn²⁺ and was more efficient in doing so than humic, fulvic and alginic acids.¹³⁴ On the other hand, the effects of nano-Ag towards the planktonic bacterium P. fluorescens were independent of the presence of standard SRHA.⁶⁷ Furthermore, the toxicity of ZnO ENPs towards E. coli, mediated by dissolved Zn²⁺, was mitigated by the presence of cations (Ca²⁺, Mg²⁺) and increasing pH, HPO₄⁻ and DOM.¹⁴⁵ Thus, in the above cases where the main bioavailable fraction responsible for ENM toxicity is unequivocally found to be mediated by dissolved species, the conventional bioavailability models described at the end of this section can be directly applied.

However, despite the numerous studies supporting the role of dissolution in toxicity, the current body of evidence suggests that the toxicity of soluble ENMs is more complex and cannot generally be simply reduced to dissolution, i.e. that the major bioavailable forms are not ionic but may be particulate or a combination of different forms (both particulate and ionic). As such, the growth inhibition of CuO, NiO, ZnO, and Sb₂O₃ ENPs towards the three model bacterial species E. coli, B. subtilis and S. aureus was found to be principally due to particle rather than to ion “species”¹⁴⁶ and the MIC of ZnO ENPs towards the bacterium S. meliloti was higher than that of ionic Zn.¹⁴⁷ Also, it was concluded that dissolved Ce⁴⁺ was not responsible for the observed growth inhibition of CeO₂ ENPs towards the freshwater microalga P. subcapitata.¹⁴⁸ A specific ENP intrinsic effect has also been shown for Ag ENPs towards the bacterium P. fluorescens⁶⁷ and towards activated sludge communities composed of various bacteria.¹⁴⁹ Results obtained by Burchardt et al.¹⁵⁰ suggest a shared effect of Ag ENPs and released Ag⁺ towards the diatom Thalassiosira pseudonana and cyanobacterium Synechococcus sp., which has also been observed for Ag ENPs and E. coli (via ROS, measured by a recombinant luminescent bacterium)¹⁵¹ and Ag ENPs and the microalga C. reinhardtii (measured as photosynthetic yield).¹⁵² Similarly, Dimkpa et al.⁸³ concluded that released Cu ions only partially account for the observed effects of CuO ENPs towards the soil bacterium Pseudomonas chlororaphis O6 and neither did released Zn²⁺ ions fully explain the Chlorella sp. growth inhibition by ZnO ENPs.¹⁵³ Other results suggest that the observed effects of the metal oxide ENPs Al₂O₃, SiO₂ and ZnO towards the bacteria B. subtilis, E. coli and P. fluorescens were not only due to released ions but also due to their tendency to attach to the cell walls of the exposed organisms.⁵¹ Stable colloidal suspensions of Ag ENPs were more toxic towards P. fluorescens biofilms than purely ionic Ag and highly aggregated Ag ENP dispersions, but Ag⁺ significantly contributed to the reduction of viability.¹⁵⁴ Hence, there are a considerable number of cases in which ionic species do not seem to be the primary bioavailable ENM form responsible for toxic effects. In these cases, a clearer picture of ENM toxicity mechanisms is necessary to enable appropriate modifications to the conventional bioavailability models.

Dissolution of ENMs, especially under natural conditions, remains difficult to predict, as discussed elsewhere.¹³² The most relevant medium properties include temperature, pressure, pH, metal solubility, ionic species and the presence of complexing agents (e.g. biological metal ion chelators, DOM),^17,34,131 which affect both ENP solubility and the chemical speciation in the medium.

Overall, a thorough understanding of dissolution processes and the chemical speciation of released metal ions is required for the proper description of ENM bioavailability and ion-mediated toxicity. In this regard, the existing general models applied to trace metals, which incorporate chemical speciation (such as the free ion activity model¹⁵⁵), the competition between metal species and biotic ligands (i.e., biotic ligand model (BLM)),⁸ biodynamics¹⁵⁶ or more general dynamics linking chemo- and biodynamic models¹³⁰ may be useful starting points for the development of similar models for ENPs, especially for those cases in which ionic species are not the sole mediators of biological effects, for which particles or a mix of the two likely play crucial roles. We have seen that the interactions (and thus the resulting effects) of both dissolved ions and ENPs with biota are affected by different environmental factors including pH, major cations and anions, trace elements, ligands of different origins, dissolved organic matter and natural colloids. In general, hardness cations such as Ca²⁺ and Mg²⁺ are known to play a protective role with respect to biota by competing with toxic metal ions for the binding sites on biological membranes. Therefore, major cations could be expected to mitigate the toxicity of the ENPs by competing with released Ag⁺, Cu²⁺ and Zn²⁺ ions for the uptake sites on algal and bacterial surfaces³³ and thereby modifying ENM surface properties and charge, favoring aggregation of the ENPs and decreasing their persistence and the probability of contact with biota. pH affects ENP dissolution, surface charge, aggregation and thus ENP reactivity. DOM could promote the dissolution of ENMs, but on the other hand DOM decreases the bioavailability of the dissolved ions by complexation. Further studies, particularly for low, environmentally relevant concentrations of ENMs, are necessary to better understand which conditions favor the action of which specific ENM species (i.e. dissolved ions or particles).

5. ENM surface transformations and bioavailability

In addition to dissolution and aggregation processes, ENMs can bind different natural and anthropogenic substances in the ambient medium or undergo other chemical transformations such as sulfidation and oxidation. Surface coatings of ENMs can play crucial roles in nano–bio interactions. Engineered or natural (DOM, proteins, lipids etc.) surface coatings transform ENP surface properties and thus play a very important role in the bioavailability of ENMs.¹⁵⁷ For example, it has been shown that citrate-functionalized nano-Ag exhibited more ready dissolution and higher toxicity towards the nitrifying bacterium Nitrosomonas europaea than the less well stabilized gum arabic and polyvinylpyrrolidone-stabilized nano-Ag suspensions.¹⁵⁸ Surface functionalization also strongly influenced the toxicity of three differently organo-coated nano-Ag ENPs towards the bacterium E. coli¹⁵⁹ and increased the bactericidal properties of glucosamine-functionalized nano-Cu by UV irradiation.¹⁶⁰ A comparison of the toxicities of bare and polymer-coated nano-CuO towards the microalga C. reinhardtii showed that coated nano-CuO was more toxic than uncoated nano-CuO whose cell-penetrating capacity was lower.⁶⁴ Another study demonstrated how G0 polyamidoamine dendrimer-coated nano-Au induced very low toxicity in two mammalian cells (N2A and Vero) but caused adverse effects on the bacterium V. fischeri and the microalga C. reinhardtii, which was explained by the different cellular structures and by the different modification of the ENP surface properties in the respective cell culture media.⁸⁶ Hence, exposure media play a crucial role and acquired surface coatings of ENPs in the environment (e.g. protein/DOM corona) can decisively influence internalization processes.⁴⁴ This was nicely illustrated by findings showing enhanced cellular uptake and subsequent toxicity of nano-CuO in the prokaryotic microalga M. aeruginosa in the presence of SRFA.⁷³

ENM transformation by sulfidation was first detected for nano-Ag¹⁶¹ but is especially relevant for class B soft metal cations (e.g. Ag, Zn and Cu) due to their high affinities for electron-dense S.³¹ Sulfidation has been shown to decrease the dissolution of nano-Ag and effectively reduce its bactericidal properties.^{143,162–164} Other findings in environmentally realistic exposure scenarios have recently observed adverse effects of nano-Ag towards microbes in spite of prior transformation in biosolids through oxidation and sulfidation.¹⁶⁵ The oxidation of nano-Ag produced Ag(NH₃)⁺ in solution and led to an increase in dissolution rates and ion-mediated toxicity towards the ammonia-oxidizing bacterium Nitrosomonas europaea.¹⁶⁶

ENPs in the environment are also known to adsorb anthropogenic substances in aquatic systems, such as surfactants,¹⁶⁷ trace metals or persistent organic pollutants,¹⁶⁸ which could considerably modify their bioavailability. This is supported by findings showing an increased adsorption of Cd(II) on HA-coated TiO₂ ENPs as compared to bare, uncoated TiO₂ ENPs⁵⁷ and increased phenanthrene sorption by HA-coated TiO₂ and ZnO ENPs,¹⁶⁹ which may lead to either a decrease or an increase in the pollutants' bioavailabilities. This is evidenced by a study showing decreased bioavailabilities of Cu and Pb in solution to the algal strains Chlorella kessleri and C. reinhardtii in the presence of carboxyl-CdSe/ZnS QDs by a factor of 2.5 and 2, respectively.⁶⁰ In fact, ENMs such as nano-Fe, nano-TiO₂ or silica monoliths are also being applied in the targeted decontamination of wastewater from various organic environmental contaminants via adsorption.^37,170 On the other hand, the same QDs increased the bioavailability of Cu and Pb to the wall-less C. reinhardtii strain by a factor of 4 and 3.5, respectively.⁶⁰ In the case of increased bioavailability, ENPs become carriers that facilitate the internalization of previously unavailable pollutants by the so-called “Trojan Horse” mechanism. These kinds of mixed effects have received very little attention so far and remain to be investigated systematically. What is more, for nano-Ag, it was found that laundry surfactants (anionic sodium dodecylbenzenesulfonate (LAS), cationic dodecyl trimethylammonium chloride (DTAC), and nonionic Berol 266) modified ENP surface and significantly influenced speciation and stability.¹⁶⁷ Depending on charge and concentration, surfactants either reduced or enhanced the degree of particle agglomeration. These interactions affected the mobility of Ag ENPs, which, depending on the reversibility of the surface modifications and exposure settings, can positively or negatively affect ENP bioavailability.⁸⁹ Apart from this rather singular account, we virtually know nothing about mixed effects of this type, which potentially complicate the prediction of ENM behavior and bioavailability in natural aquatic systems. Their possible impact on ENP bioavailability remains to be examined.

6. Conclusions and outlook

Though an ever more accurate understanding of particle behavior in complex systems has been attained and a wealth of ecotoxicological information has been gathered on the environmental hazards of ENMs in recent years, the link between the two aspects is comparably less developed. There are still considerable knowledge gaps with respect to ENM bioavailability in the environment. Most importantly, ENM “speciation” in natural aquatic systems is not well understood or quantified. The major bioavailable fractions responsible for toxic effects have not been concludingly identified either. Considerable knowledge gaps remain with respect to the effects of agglomeration on bioavailability as well as on the exact uptake routes, intracellular compartmentalization as well as DOM–protein competition on the surface of internalized ENPs. Hence, the concept of bioavailability may provide a useful framework to link “chemical and physical speciation” of ENMs with their possible biological effects but require some ENM specific adaptations. In the environment, ENMs are present as complex mixtures of different chemical and physical forms including particles, particles complexed with DOM, sulfidized/oxidized forms, agglomerated and (partially) dissolved forms etc. For a given ENM, laboratory studies revealed that the relative proportion of these forms depends on the environmental parameters such as water hardness ions, pH and DOM.

In the case where particles as such are the major bioavailable fraction mediating toxicity, water hardness cations are expected to reduce the bioavailability of ENMs by favoring aggregation and thus protecting algae and bacteria from ENM stress. In the case where solubilized ENM forms are the main bioavailable form, hardness cations will reduce their bioavailability by competing with the released ions for sensitive sites on biological membranes. Similarly, ambient pH values are anticipated to affect aggregation, dissolution and interactions with bio-interfaces. Circumneutral pH values typically found in surface waters will favor the agglomeration of the ENMs with close pH_IEP. However, it has already been shown that the aggregation behavior of ENPs in surface waters is largely determined by DOM. Indeed, DOM is anticipated to play a key but dual role in ENM bioavailability to algae and bacteria. On the one hand, DOM is expected to affect chemical and physical “speciation”, e.g. by complexing ENPs and favoring the formation of stable suspensions, which would favor their contact with biointerfaces. On the other hand, DOM also binds a solubilized metal, which will decrease the bioavailability and uptake of dissolved ENM species. Therefore, for the same ENM, different bioavailabilities must be expected under different environmental conditions.

It is currently unclear if it is realistic (by analogy to free metal ions) to search for a specific form that could be used as a measure of the biological reactivity of the medium. Accordingly, the development of analytical tools and models that enable the characterization of ENM “speciation” or the measurement of specific dissolved and/or nanoparticulate forms would be highly useful. Also, it is still unclear if specific properties can universally be considered important for the uptake and toxicity of inorganic ENMs. In the case of soluble ENMs, it seems important to quantify and distinguish the contribution of soluble and particulate fractions to toxicity, as well as to define the “limiting” condition corresponding to the prevailing contribution of one or the other form.

To verify if basic ecotoxicological paradigms are applicable to ENM toxicity, it is crucial to improve our current understanding of possible particle-specific uptake routes in planktonic bacteria and microalgae, the kinetics of the uptake and excretion processes, evaluation of the importance of internalization with respect to adsorption to surfaces, as well as the intracellular fate of ENMs in algae and bacteria. Furthermore, it is still unsure whether endocytic pathways are common in microorganisms such as planktonic bacteria and microalgae or rather are an exception. It is also still unclear if cellular uptake is a precondition for biological effects and if such a general principle, which is proven for conventional micropollutants, is also valid for ENMs. We have seen that biological effects can, but not necessarily, be triggered by ENMs directly reacting with sensitive sites on the biological membrane and/or by the penetration of biological barriers (internalization), i.e. cell walls or epithelial boundaries and membranes. However, there are processes and indirect mechanisms of toxicity for which direct contact of ENMs with biota is not necessarily required. These include (i) the dissolution of ions, (ii) the production of ROS (mediated by UV, discontinuous crystal planes or material defects) and (iii) redox cycling and catalytic chemistry (e.g. Fenton and Quinone reactions), which can lead to oxidative damage.^41,171,172 In the case in which these processes dominate, the applicability of the bioavailability concept to ENMs may be limited due to the complexity of these interactions. The Trojan Horse mechanism is another important possible mechanism of indirect ENM toxicity, which is not considered in the bioavailability framework. Despite these limitations, the integration of key modifying parameters is critical to develop a deeper understanding of the bioavailability of ENMs and to establish general principles that allow us to link environmental transformations of ENMs with their potential effects. Furthermore, a well-established bioavailability concept would allow the development of descriptive and predictive models capable of taking into account specificities of different environmental systems and enable extrapolations of laboratory results to the natural environment. Important advances in understanding environmental hazards of ENMs to different organisms and their interactions at nano–biointerfaces were mainly achieved by bioassays in well-controlled media with relatively high ENP concentrations. However, in the environment, ENPs are subjected to complex and highly dynamic physical, chemical and biological transformations, which will alter their interactions with biota and toxicity.^31,162 A first attempt has already been done to predict the speciation and prevailing forms of Ag based on thermodynamic principles in various environmental scenarios relevant for natural and constructed environments (wastewater treatment plants).¹⁶²

Box 1. Definitions

Nanomaterials	The European Commission has recently recommended the following definition of nanomaterials:^14 “‘Nanomaterial’ means a natural, incidental or manufactured material containing particles, in an unbound state or as an aggregate or as an agglomerate and where, for 50% or more of the particles in the number size distribution, one or more external dimensions is in the size range 1–100 nm. In specific cases and where warranted by concerns for the environment, health, safety or competitiveness the number size distribution threshold of 50% may be replaced by a threshold between 1 and 50%.” There is currently no universally accepted definition which is equally practical and unambiguous to industry, consumers, legislators and scientists alike. There are good scientific grounds to argue that the 100 nm limit is scientifically somewhat imprecise, if not arbitrary.^15–17 Some argue that the nano-specific, non-bulk properties only emerge below a critical size of about 20–30 nm or below.^17,18 Kreyling, Semmler-Behnke and Chaudhry^15 provided a good overview of the different definitions currently in use. Furthermore, it is also important to clearly distinguish between engineered and naturally occurring nanoparticles,^19 which are ubiquitous in all water bodies and crucially influence water chemistry.^20 For the purpose of this review, we will adhere to the above definition and designate manufactured nanomaterials and nanoparticles “engineered” nanomaterials (ENMs) and nanoparticles (ENPs), respectively.
Bioavailability	According to the International Union of Pure and Applied Chemistry, biological availability is defined as “the extent of absorption of a substance by a living organism compared to a standard system”.^2
Agglomeration and aggregation	The term agglomeration is used for loosely adhering clusters of particles without any chemical bonds. Aggregates are clusters of particles irreversibly held together by chemical bonds^21 (the two terms are, however, commonly used ambiguously and interchangeably).

Acknowledgements

Many thanks to the anonymous reviewers who have helped improve this article. This work was financed by the Swiss National Science Foundation in the framework of the Swiss National Research Program 64 on the Opportunities and Risk of Nanomaterials.

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