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Transformations that affect fate, form and bioavailability of inorganic nanoparticles in aquatic sediments

Article in Environmental Chemistry · August 2015 DOI: 10.1071/EN14273

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Transformations that affect fate, form and bioavailability of inorganic nanoparticles in aquatic sediments

Richard Kynaston Cross,A,B Charles TylerA and Tamara S GallowayA

AGeoffrey Pope Building, Biosciences, College of Life and Environmental Sciences, University of Exeter, Stocker Road, Exeter, EX4 4QD, UK. BCorresponding author. Email: [email protected]

Environmental context. Engineered nanomaterials are increasingly being used and their release to the aquatic environment poses potential risk. We review the research on transformations of engineered nanomaterial in the aquatic sediment environments, and consider the implications of their release. The key factors defining the fate of engineered nanomaterials in aqueous and sediment systems are identified.

Abstract. Inorganic nanoparticles are at risk of release into the aquatic environment owing to their function, use and methods of disposal. Aquatic sediments are predicted to be a large potential sink for such engineered nanomaterial (ENM) emissions. On entering water bodies, ENMs undergo a range of transformations dependent on the physicochemical nature of the immediate environment, as they pass from the surface waters to sediments and into sediment-dwelling organisms. This review assesses the current state of research on transformations of metal-based ENMs in the aquatic environment, and considers the implications of these transformations for the fate and persistence of ENMs and their bioavailability to organisms within the benthos. We identify the following factors of key importance in the fate pathways of ENMs in aqueous systems: (1) extracellular polymeric substances, prevalent in many aquatic systems, create the potential for temporal fluxes of ENMs to the benthos, currently unaccounted for in predictive models. (2) Weak secondary deposition onto sediment grains may dominate sediment–ENM interactions for larger aggregates .500 nm, potentially granting dynamic long-term mobility of ENMs within sediments. (3) Sulfurisation, aggregation and reduction in the presence of humic acid is likely to limit the presence of dissolved ions from soluble ENMs within sediments. (4) Key benthic species are identified based on their ecosystem functionality and potential for ENM exposure. On the basis of these findings, we recommend future research areas which will support prospective risk assessment by enhancing our knowledge of the transformations ENMs undergo and the likely effects these will have.

Received 18 December 2014, accepted 6 May 2015, published online 14 August 2015

Introduction Attempts have been made to model the release of ENMs The use and application of engineered nanomaterials (ENM) is into the environment at both national and global scales[13–15]; rapidly expanding, with a global industry estimated to be worth however, a lack of data on production[16] and emission rates US$3 trillion by 2020.[1] This growth is reflective of their wide- inhibits accurate prediction of the release potential of ENMs.[17] ranging applications, spanning environmental remediation,[2,3] Recent estimates predicted 1100–29 200 Mg of the global ENM commercial products[4–7] and medicine.[8] The expansion of production in 2010 was released directly into water bodies[18]. commercial uses for ENMs is fuelled by the unique properties Most of these ENMs will pass through some form of water- materials display at the nanoscale compared with the larger treatment process before entering the aquatic environment, (bulk) form of the material, principally due to their high surface which can be highly effective at removing ENMs such as silver area-to-volume ratio.[9] It is this high reactivity at the nanoscale (AgNPs).[19] In one study, silver in the effluent from a model that has established ENMs as a contaminant separate from their waste-water treatment plant accounted for only 2.5 % of the bulk form. The commonly used definition of a nanomaterial is a original dosed concentration.[20] However, some ENMs may material with at least one dimension ,100 nm.[10,11] Nano- also be released directly to aquatic systems. Coatings and particles (NP) have all three dimensions ,100 nm, whereas cosmetics comprise the largest share of commercial products nanosheets and nanorods exist with one or two dimensions containing ENMs[21] and are the two largest potential sources of ,100 nm. In terms of regulation, the European Commission has ENM release into the environment.[18] Most ENMs utilised in a working definition of engineered nanomaterials as ‘containing coatings and cosmetics are inorganic metal or metal oxide NPs [22,23] particles, in an unbound state or as an aggregate or as an (MeO) such as titanium dioxide (TiO2), zinc oxide (ZnO), agglomerate and where, for 50 % or more of the particles in the silicon dioxide (SiO2) and cerium oxide (CeO2), which have number size distribution, one or more external dimensions is in high production and release volumes.[14,16,18] These may be the size range 1–100 nm’.[12] released from painted exterior facades and other consumer

Journal compilation Ó CSIRO 2015A www.publish.csiro.au/journals/env R. K. Cross et al. products directly into the aquatic environment[23,24] with uncer- cellular internalisation may be physically or temporally con- tain ecological consequences.[25] There is little regulation in strained, whereas bioavailable ENMs may be readily interna- place for their production or disposal, making the area of nano- lised by cells and tissues. As such, transformations of ENMs ecotoxicology one of vital importance for prospective risk and their effect on bioavailability in sediments are an important assessment.[25–27] component of ENM hazard prediction. There has been recent Water bodies receiving ENMs present a diverse range of progress in standardising approaches to measuring the toxicity environments, from fast-flowing rivers to lakes high in natural of chemicals in sediment environments,[36] but the existing organic matter (NOM) or oceans of high ionic strength. Differ- framework for chemical contaminants may not be suitable for ences in both physical and chemical conditions of the immediate ENMs owing to the fundamental differences between solute environment may drastically alter the fate and behaviour of and colloidal chemistry. International efforts are working ENMs. Both naturally occurring NPs[28,29] and ENMs[30,31] towards achieving the same quality of standardisation and have been shown to aggregate and sediment on transport into environmental realism for nano as for traditional chemical saline environments as occurs when entering estuarine waters. ecotoxicology.[25,37,38] To do this, a firm understanding of the In the Gironde estuary, France, silver from anthropogenic transformations that occur through the products’ lifecycle is sources contributes 24–90 % of particulate Ag fluxes into of utmost importance.[39] [32,33] sediments. In some instances, for example TiO2 in the The present review critically analyses knowledge concerning Rhine river, modelled sediment concentrations exceed that in the hazard potential of inorganic ENMs in the benthic environ- the overlying water by up to six orders of magnitude.[34] ENMs ment. We distinguish between organic and inorganic ENMs have also been observed to move in run-off from terrestrial soil for the purposes of the current review owing to unique fate into aquatic sediments.[35] Therefore, sediments are predicted as processes of inorganic ENMs concerning dissolution and the a major sink for ENM emissions.[11] presence of ions, and the differences in the specific challenges The transformations that an ENM may experience on enter- for investigators that organic and inorganic ENMs present, ing the aquatic and benthic environment can be subdivided into which must be dealt with separately. The reviews focus is on four main categories: the interactions that occur between inorganic ENMs with both biotic and abiotic components of the benthic environment and Physical transformations, including aggregation processes the implications of these transformations for the fate and and deposition onto sediment grains. bioavailability of ENMs to benthic species. Investigating these Chemical transformations, including dissolution, exchange of processes and transformations will help identify the form of surfactants, influence of other inorganic chemicals and redox inorganic ENMs likely to be present within the benthos. In doing reactions. so, this information can be used in support of standardised Interactions with macromolecules external to organisms such testing for predictive risk assessment of ENMs in sediments. as NOM and extra-cellular polymeric substances produced in the water column or by biofilms at sediment surfaces. Biologically mediated transformations involving partitioning Transformations in the water column among benthic communities, bioturbation and transforma- Many transformations occur in the water column on release of tions associated with ingestion–egestion dynamics. ENMs into the aquatic environment and as they pass from sus- Each of these transformations may influence the toxicity, pension in the water column to their incorporation into sedi- bioaccessibility and bioavailability of ENMs. Bioaccessible ments (Fig. 1). These transformations facilitate the flow of ENMs can be taken up by an organism, but their potential for ENMs from the water column into sediments and may persist

Richard Cross is a Ph.D. scholar whose research is focussed on the factors that determine bioaccumulation of engineered nanomaterials in sediment-dwelling organisms. His work considers the transformations that engineered nanoparticles undergo within aquatic ecosystems, in order to investigate the link between the physicochemical properties of engineered nanoparticles and biological factors that determine the bioavailability of these anthropogenic pollutants. The focus on sediment-dwelling species is in recognition of the important ecological role of these organisms and the potential risk they face from engineered nanoparticles in the future.

Charles Tyler is a reproductive physiologist and ecotoxicologist. He is Deputy Head of Biosciences and Academic Lead in the College of Life and Environmental Sciences at the University of Exeter. His research spans investigations into the effect mechanisms of endocrine-disrupting chemicals, pharmaceuticals and nanoparticles, to assessing population-level effects of these environmental contaminants in wildlife, principally fish. He has published more than 220 full research papers and peer- reviewed book chapters and reviews with ,13 000 citations. In 2012, Tyler was awarded The Fisheries Society of the British Isles Beverton Medal for ground-breaking research in fish biology.

Tamara Galloway is Professor of Ecotoxicology at the University of Exeter and also holds an honorary Chair at University of Exeter Medical School. Tamara’s research focus is in understanding how organisms adapt and survive in polluted environments, what makes some organisms more susceptible than others, and how we can use this information to protect the environment. She studies the health effects of some of the most pressing priority and emerging pollutants, including complex organics, plastics additives, metals and nanoparticles.

B Transformations of nanoparticles in sediments

NP coated with NOM may prevent homoaggregation, which can lead to enhanced stability. Homoaggregation occurs in waters of low NOM and high IS. Results in sedimentation. Large heteroaggregates formed with biologically derived organic matter, particularly under high IS conditions. Increases sedimentation.

Ingestion of suspended NPs by pelagic species or the pelagic phase of some sediment dwellers.

Dynamic cycling of NP Larger aggregates Contact with mucus on within the sediments targetted by filter external membranes with bioturbation and feeders, increasing resulting in flocculation egestion burying and uptake of NP in of these and other re-suspending NPs. epibenthos. excretory products and subsequent sedimentation.

Epibenthic organisms Sedimentation and association with sediment grains gives NPs suspended in the pore water can Endobenthic organisms access to sediment ingesting benthic be transported dwellers. through saturated sediments.

Fig. 1. Conceptualisation of the aggregation processes resulting in the accumulation of engineered nanomaterials (ENMs) in sediments and their exposure to benthic-dwelling organisms (NPs, nanoparticles; NOM, natural organic matter, IS, ionic strength). within the benthos. Transformations include physicochemical ENM.[43,49,50] The capacity of ionic strength to influence transformations such as the dynamic interplay between aggre- aggregation can be limited by particle shape. Changing ionic gation and dissolution, as well as interactions with NOM and strength had minimal influence on aggregation of rod-shaped aggregates of biogenic exudates known as ‘marine snow’, all of ZnO (aspect ratio 10 : 1) owing to the lower interaction energies which may have lasting implications for ENM fate within the involved between rod-shaped NPs, while dramatically increas- benthos. ing aggregation for 20-nm spherical ZnO until aggregation was no longer limited by ionic strength.[42] Increasing ionic strength Aggregation not only facilitates homoaggregation, but also heteroaggrega- On release into water bodies, metal ENMs exhibit behaviour tion with large natural colloids (.0.2 mm in diameter). This similar to other colloids, with their fate dominated by aggre- heteroaggregation has been demonstrated as the main cause gation and dissolution.[39–41] In principle, this follows behind sedimentation of ENMs in river water,[51] a process Derjaguin–Landau–Verwey–Overbeek (DLVO) theory, where exacerbated in waters of high ionic strength. CeO2, and AgNPs, aggregation behaviour is the product of attractive van der Waals coated with either polyvinylpyrrolidone (PVP) or SiO2 experi- forces and repulsive electrostatic forces. The balance of these enced higher rates of heteroaggregation with larger natural opposing forces is highly dependent on the physical and colloids in saline water and subsequently faster sedimenta- chemical properties of the ENM in question and the properties of tion.[52] Therefore, sediments are a likely reservoir of these the media in which the ENM is suspended. Most materials particles on release into surface waters, particularly if they move suspended in water have a negative surface charge, attracting along a salinity gradient such as that in an estuarine environ- counter-ions from the solution and inducing the repulsive ment, into waters of higher ionic strength.[53] electric double layer (EDL), which on close interaction with The aggregation behaviour of ENMs depends not only on the other negatively charged colloids results in electrostatic repul- physicochemical characteristics of the water but also the physi- sion. This prevents aggregation and results in a material cal properties of the ENMs themselves. Both concentration and becoming stabilised in suspension. Ionic strength of the sur- initial particle size have a pronounced effect on aggregation rounding water influences aggregation behaviour, with high behaviour. At higher concentrations, MeOs such as TiO2, ZnO ionic strength typical of saline marine waters, resulting in a and CeO2 sediment at a faster rate than when dosed at lower compression of the EDL, reducing electrostatic repulsion and concentrations.[50,51] Higher concentrations result in a greater therefore increasing homoaggregation.[42–46] A similar effect number of particle–particle interactions and so a greater poten- has been seen as the pH approaches the pH point of zero charge tial for homoaggregation. This relationship is dynamic. Evi- [47,48] (pHzpc) for the particle. This reduced stability of the ENM dence of aged TiO2 from commercially available sunscreens and subsequent increase in homoaggregation results in faster showed two distinct size distributions formed after 48 h, one rates of sedimentation under conditions of high IS or where the ,700 nm in diameter and the other group of larger low-density pH of the surrounding water is close to the pHzpc of the heteroaggregates .10 mm of TiO2 with other components of the

C R. K. Cross et al. sun creams.[53] Such bimodal size distributions may arise NOM with ENM may grant either steric[64–68] or electrostatic because rapid aggregation reduces the overall number of parti- stability[44,65,69,70] to the ENM, reducing sedimentation.[50,71] cles in suspension, reducing the frequency of particle interac- Keller et al.[50] found that in the presence of ,5 mM humic acid, tions; thus, a second group of smaller or individual ENMs TiO2 (27 nm), ZnO (24 nm) and CeO2 (67 8 nm) formed stable remain in suspension.[51,53] Such aggregation during aging can heteroaggregates of ,300 nm in size, which then remained in be irreversible. Dry powders of 85-nm hematite ENMs did not suspension. This can be compared with the same ENMs in the disaggregate below 100 nm when stored for 1 month.[46] absence of NOM, where aggregates formed across the 400-min Additionally, ENMs of a smaller initial size have been observed time period ranged from ,400 to 1200 nm[50] and settled out of in several cases to result in larger aggregates[45,46,54]; for suspension. At low concentrations of NOM, typical of ground- example, 10-nm citrate-capped AgNPs aggregated to 300 nm water (between 0.1 and 2 mg L1), the stabilising effect of NOM after 3 days, compared with the 100-nm particles, which formed becomes concentration-dependent. An exposure ratio of 2 % 200-nm aggregates in saline water.[55] Overall, it is likely that a NOM to ENM concentrations caused bridging and flocculation significant proportion of ENMs released into the aquatic envi- of TiO2 NPs, whereas NOM concentrations of ,20 % stabilised ronment will persist as larger aggregates and agglomerates. 25 % of the ENM as suspended colloids ,300 nm,[72] in The dynamic nature of ENM aggregation behaviour makes agreement with Keller et al.[50] These suspended ENMs can still characterisation of ENMs in natural matrices difficult, owing to have an effect on benthic communities through disrupting the the propensity for NPs to change during the process of charac- life cycle of the planktotrophic larvae produced by many sedi- terisation. Much of the past literature quotes the primary particle ment dwellers that feed in the water column. To the authors’ size when presenting results. However, as can be seen, these knowledge, no experimental data have been published investi- reported sizes may bear little resemblance to the actual size gating the effects of ENMs on planktotrophic stages of other- distributions of ENMs in the laboratory exposures. As such, the wise benthic-dwelling species. Such tests would be of particular current research efforts to improve separation and sizing tech- use because there is little acknowledgment of this benthic– niques such as field flow fractionation and single-particle pelagic coupling in current regulatory frameworks. inductively coupled plasma mass spectrometry are vital in order for progress to be made in our understanding of ENM aggrega- tion and its implications for environmental fate. Interaction with extracellular polymeric substances Clusters and films of aggregated organic and inorganic matter Dissolution are prevalent in both marine and freshwaters. Much of this forms from exudates of extracellular polymeric substances (EPS) from Dissolution also plays a significant role in the fate of MeO NPs a diverse range of microbes[73] that attach to these agglomer- on release into water bodies. Evidence suggests that there is a ates.[73,74] These microgels provide an efficient transport relationship between particle size and dissolution rate at the mechanism for the movement of organic and inorganic matter nanoscale,[56] though it is not the exponential predicted by solely from surface waters to the benthos known as ‘marine using thermodynamic theory.[44,57] This may be due to several snow’,[75,76] with sedimentation rates from 1 to 368 m day–1[77]. factors, including the larger surface area relative to volume MeOs are rapidly incorporated into EPS, with a 95 % incorpo- increasing the number of ‘hotspots’ for dissolution,[57] surface ration efficiency for TiO after 168 h.[78] Such marine snow morphology[42] and NP concentration.[58] In general, smaller 2 events are likely to occur after phytoplankton blooms and are not NPs have a greater rate of dissolution; however, other factors restricted solely to marine environments, with fluxes of partic- can reverse this general rule. For example, 78-nm CeO NPs had 2 ulate organic matter in sediments coinciding with the disap- a greater concentration of CeIII impurities than smaller, 33-nm pearance of EPS in the surface waters of Lake Constance, NPs, and so experienced greater dissolution at low pH , 5.[59] Germany.[79] Waste-water treatment plants also act as an Other factors that may suppress the rate of dissolution include important source of EPS. ENMs come into contact with an array ligand exchange in the presence of phosphate[59] and aggrega- of microbes and EPS associated with sewage sludge and so may tion.[44,56,60] In fact, aggregation has been observed to reduce become complexed with such EPS on release in the effluent into dissolution rates of ZnO,[61] in one case to values similar to the water bodies.[80] The presence of ENMs has been suggested to bulk form,[62] a finding that has been attributed to an increase in enhance aggregation and assembly of EPS microgels in marine the thickness of the diffusion layer with size.[63] Experimentally, waters. In the absence of ENMs, EPS aggregated to ,0.5 mm, a slower rate of aggregation in lower-ionic-strength media has whereas the presence of a high concentration (100 mgL 1)of been shown to increase ZnO aggregate density,[49] which could 23-nm polystyrene NPs resulted in EPS-NP aggregates of also reduce diffusion efficiency of solutes from the interior of 4–6 mm.[81] This could mean that in the presence of ENMs, the aggregate to the bulk solution.[63] Both of the transforma- marine snow may flocculate and sediment at a faster rate than tions involved in dissolution and aggregation are likely to con- usual, which may have wider ecosystem implications, through tinue in tandem with more sediment-specific processes such as altering the local dynamics of this process. interactions with sediment grains and biofilms on reaching the When EPS flocculate and sediment, not only is there an benthos. It is this dynamic interplay of dissolution and aggre- increased exposure of ENMs to the benthos, but there can also be gation that is a fundamental difference between traditional other transformations of ENMs, impacting on aggregation and solute chemistry and ENMs’ colloidal behaviour. dissolution.[82] EPS was shown to increase heteroaggregation for both positively and negatively charged cadmium selenide Interaction with natural organic matter (CdSe) quantum dots (used as a representative inorganic In waters containing high NOM, such as lakes, rivers and ENM). The size of aggregates differed under illumination, with estuaries, NP-NOM complexes dominate particle behaviour ,800–3800-mm aggregates formed under dark conditions and because NOM concentrations are likely to be orders of magni- ,700–1300 mm under a light : dark regime of 14 : 10 h. Interest- tude higher than ENM concentrations. This co-aggregation of ingly, aggregation should suppress dissolution through reducing

D Transformations of nanoparticles in sediments the surface area from which ions can dissolve. However, in this acid-capped AuNPs when in the presence of D. magna and the study, particularly under lit conditions, dissolution of the CdSe freshwater shrimp Gammarus pulex.[90] Such increases in was actually increased. Proteins in the EPS acted as a source for smaller inorganic colloids in the presence of the filter feeders CdSe-photocatalysed production of reactive oxygen species, have been suggested as a combination of selective ingestion of thus destabilising the CdSe. This effect increased dissolution larger aggregates and break-down of these aggregates during and degradation of the quantum dots. Although UV is attenuated digestion and egestion.[91] Although these studies concede that by waters and so does not reach the benthos in many cases, in population densities in the investigations were too low to have a natural lake waters with low dissolved organic carbon, ultravio- significant effect on the natural systems they represented, in let penetration can reach a depth of 10 m[83] and so photoactive areas with a high density of filter-feeding organisms, significant properties of some MeOs may still play an important role in change to the size distribution of ENMs reaching the benthos ENM transformations. Flocs of EPS may remain in surface may be possible. waters for 5–16 days,[84] allowing time for significant degrada- Selective feeding by epibenthic species will expose some tion of photocatalytic ENMs such as CdSe quantum dots or TiO2 NPs to biogenic mucus in both the gut and gills, which will close to the water’s surface. This potentially increases exposure dominate particle interactions in a similar way to EPS. Surface of pelagic species to dissolved ions from the ENM, while mucus represents a significant barrier to uptake of ENMs across reducing the amount of particles that could sediment out to the dermal pathways.[92] Studies on the mussel Mytilus gallopro- [84] benthos. However, during windy conditions and when con- vincialis showed most CeO2 and ZnO NPs dispersed in suspen- centrations of EPS are great (for example after algal blooms),[79] sion were rejected in pseudofaeces before ingestion could take rapid sedimentation can occur. This could result in seasonal place.[93] These pseudofaeces are produced in the gills at a much fluxes of high ENM concentrations in the benthos. This is a higher rate compared with normal faeces and so act as an hazard that is not recognised in current predictions of sediment efficient mechanism for sorting unwanted NPs during filtra- concentrations of ENMs. Marine snow is not acknowledged in tion.[93,94] The result of this is the deposition of mucus–NP predictive models and sedimentation rates are assumed to be complexes onto sediment surfaces, similar in composition to the constant across the year.[14,85] As such, ENM interactions with EPS-ENM complexes mentioned earlier. EPS need to be better understood and are expected to be an On ingestion, the gastrointestinal mucus is the first barrier for important factor in predictive modelling of the environmental uptake of ENMs into the organisms’ tissues and acts as another fate of ENMs. biological compartment for mucus–ENM interactions. Nega- tively charged ENMs may cross this barrier faster than their [95,96] Transformations at the surface of sediments cationic counterparts. Hydrophobic ENMs are also expected to experience far greater uptake, bioaccumulation At the sediment–water interface, deposition onto natural col- and cellular internalisation than their hydrophilic counter- loids appears to be the main mode of sedimentation and incor- parts.[96,97] However, even those ENMs that reach intestinal poration into sediments.[51] This is of particular importance in epithelial cells and are internalised may still be ejected back into turbulent, colloid-rich systems, e.g. rivers, estuaries and coastal the gastrointestinal mucus to be eliminated from the organisms. systems. Heteroaggregation of CeO , PVP-Ag and silica-coated 2 Goblet cells have been recorded to internalise activated carbon silver (SiO -Ag) NPs with suspended sediment flocks has been 2 NPs into intracellular vacuoles, which are then secreted at the observed 12 orders of magnitude higher in turbulent waters than gut lumen through exocytosis.[98] These excreted ENMs will in quiescent conditions. This resulted in sedimentation rates one have undergone mechanical and chemical transformations due to two orders of magnitude higher than in still or stagnant to changes in pH throughout the digestive process, for example, waters.[86] In lakes, this scavenging of ENMs out of the water 45-nm gold (Au) NPs excreted by the clam Corbicula flumine column by suspended sediment particles dominates ENM sed- were far less spherical and uniform in size than pre-exposure.[99] imentation in shallow waters, in particular in the diffusive This transformation of NPs in faeces and mucus exudates benthic boundary layer, or ‘fluff layer’ just above the sediment represents an important process in benthic–pelagic coupling. surface.[87] Therefore, for many MeOs, persistence in the ben- In areas with high population densities of burrowing and thos will involve aggregated ENMs deposited onto sediment sediment-ingesting organisms, bioresuspension of nanomater- grains or other natural colloids. Further transformations will ials in faecal pellets brings a significant volume of sediments then alter the composition and form that ENMs take at the sur- and their associated contaminants to the sediment surface.[100] face of sediments. These include interactions with epibenthic For example, bioresuspension through release of faecal pellets species, bioresuspension of NPs and EPS derived from plants of the burrowing mud shrimp Callinassa subterranea has been and microbes inhabiting the surface of aquatic sediments. estimated to equal an annual sediment turnover budget of 11 kg dry sediment m2 year1.[101] Although to date there has been Effects of animal species on composition and distribution no published direct investigation into the effect this bioresus- of ENMs at the sediment–water interface pension has on ENMs, the importance of bioturbation and The species composition of epibenthic communities will have a resuspension in many other nutrient and chemical cycles[102] profound effect on the size distribution and form of ENMs suggests that these processes are likely to play an important role before settling to the sediment surface. Daphnids (water fleas) in benthic–pelagic coupling of ENM cycles. have an average filter mesh size of 400–700 nm and may only feed on suspended particles greater than this size.[88] As a con- sequence, the species Daphnia hyalina alters the size distribu- Interactions with EPS at the sediment–water interface tion of natural inorganic colloids, simultaneously reducing the EPS exuded from bacteria will also affect the distribution of number of aggregates larger than their filter mesh size and ENMs at the sediment surface. Induced aggregation of biogenic increasing the number of smaller aggregates ,400 nm.[89] A ZnS NPs to sizes of 1–10 mm by proteins exuded from microbes similar effect was observed for 30-nm mercaptoundecanoic has been demonstrated to limit their transport in subsurface

E R. K. Cross et al. waters, and in turn incorporate excess Zn ions from pore attraction occurs and beyond which an energy barrier exists to waters.[103] This demonstrates how EPS from microbial com- deposition.[107] The strength of this barrier to deposition munities can alter the form of ENM at the sediment surface, depends on the ionic strength of the external media, with greater reducing the concentration of Zn ions in interstitial pore waters ionic strength resulting in a smaller EDL and so smaller energy within sediments, while increasing the nanoparticulate fraction. barrier.[108] Under cases of high ionic strength or low pH, this Such EPS may not only originate from bacteria. Bone et al. energy barrier can be entirely overcome, allowing deposition in prepared different exposure stocks of 50-nm PVP-AgNPs and this primary energy minimum[109] Surface charge of different 12-nm gum arabic (GA)-capped AgNPs by preparing them in minerals within sediments may also alter deposition, with more the presence of sediment or plants or both sediments and negatively charged silica and iron oxide collectors experiencing [104] [110] plants. EPS produced by the plants as a detoxification greater deposition of CeO2 NPs than alumina. Alternatively, response to the presence of the NPs caused the ENMs to as electrostatic repulsion decreases at an exponential rate with aggregate and sediment, reducing the ENM concentration sus- distance, a zone of weak attraction can exist further from the NP pended in the water, and so reducing toxicity to these organisms surface, known as the secondary energy minimum[111] Under feeding in the water column. Therefore, in shallow waters where less-favourable conditions for primary deposition, secondary there are likely to be aquatic flora, EPS production will be higher deposition in this weaker zone of attraction can still occur. The and significantly increase sedimentation of ENMs. Whether this factors that determine deposition of an ENM can thus be con- decrease in exposure and toxicity to feeders in the water column ceptualised as those that increase interactions between ENMs would be mirrored by an increase in toxicity to sediment- and sediment grains and those that increase the attachment dwelling species was not assessed, because no benthic species efficiencies of these interactions. was included in the exposure. This highlights the importance of Ionic strength has an important role in determining the including a sediment-ingesting benthic organism as part of the attachment efficiency of interactions between NPs and sediment base set of organisms for the standardised testing of NP aquatic grains in the benthos. An increase in ionic strength has been [112,113] [114] hazards, because reduced toxicity to pelagic species due to demonstrated to increase deposition of TiO2, CeO2, [115] [116] sedimentation across the exposure period could be negated by an copper oxide (CuO), ZnO and aluminium oxide (Al2O3) increase in toxicity to benthic species. in saturated porous media. The form this deposition takes Incorporation of MeOs into larger aggregates may also depends on the size and surface charge of the NP in question. increase exposure to some epibenthic organisms. For example, As negatively charged aggregates become larger, their surface the mussel Mytilus edulis and oyster Crassostrea virginica both potential becomes more negative.[47,69] As such, secondary experienced rapid uptake and accumulation of 100-nm fluores- deposition would dominate the fate of larger NPs .500 nm cently labelled polystyrene NPs when they were embedded into because they induce a stronger interaction barrier and deeper laboratory-generated aggregates (.100 mm), whereas the freely secondary energy minimum, preventing deposition in primary suspended particles (,1 mm) experienced no uptake over a energy minima but encouraging secondary deposition.[111,117] 45-min exposure.[105] The aggregated ENMs were bioaccessible However, pH and ionic strength can significantly reduce these whereas the untransformed ENMs were not owing to these energy barriers; for example, at pH 7, a value close to the pHzpc species’ greater capture efficiency of particles .1 mm in size. of both ZnO and iron oxide (Fe3O4), these particles aggregate to Research now needs to address whether this transformed, .1 mm and no energy barrier is present, so conditions are bioaccessible fraction of ENMs is also bioavailable to benthic- favourable for primary deposition.[115] This has wide implica- dwelling species. tions for the longer-term fate of ENMs in sediments because primary deposition is generally irreversible whereas deposition Subsurface transformations in the benthos in the secondary energy minima can be reversed. This can occur through Brownian motion[117,118] or if there is a change in flow On incorporation into sediments, the ENM will interact with two rate,[119] or ionic strength,[107,111] which would be periodical in distinct environments: the pore waters and the sediment grain estuarine systems. The literature suggests that under some surfaces. These interactions are highly dependent on the phys- conditions the secondary energy minimum will play a dominant icochemical transformations that the ENM has undergone and role in NP deposition. This is primarily for aggregated particles will determine the particles’ bioavailability to benthic-dwelling approximately .500 nm where an increase in size may result in organisms. Geochemical heterogeneities in the sediment such as greater deposition in the secondary energy minimum,[111] but mineral form, distribution and concentration also add com- also in some cases for smaller NPs ,30 nm,[118,120]. However, plexity to the task of understanding subsurface transformations for the majority of NPs ,100 nm[30] or for rod-shaped NPs and greatly influence deposition and transport behaviour of [106] where the interaction energies are reduced, the secondary ENMs in saturated porous media. There is a lack of infor- minima are expected to play a far smaller role in ENM deposi- mation on the fate and processes acting on ENMs in the aquatic tion.[121] Towards the larger end of the scale where ENM sediments; thus, many processes are inferred through the use of aggregates are .1 mm, physical straining dominates ENM our knowledge on ENM behaviour in similar matrices such as removal from the interstitial water.[108,109] This is where the saturated soils, for which there has been greater research effort. pores become too small for the ENM to pass through (Fig. 2) and so the particles are trapped and no longer mobile. Physical transformations will determine deposition and Over time, deposition and straining can result in a ‘ripening’ transport of ENMs within sediments effect where deposited and strained aggregates act as collecting The balance between attractive van der Waals forces and surfaces themselves, allowing further deposition of the ENM, as [112] repulsive electrostatic forces are the main interactions that has been shown to occur for TiO2. Such deposition can be determine the form of deposition of ENMs onto sediment grain enhanced by particle shape; for example, the higher deposition surfaces. These two forces create a primary ‘energy minimum’ and ripening effect observed of rod-shaped latex NPs (aspect close to the sediment grain surface (,,1 nm) where a strong net ratio 4 : 1) that was not seen with spherical latex NPs under

F Transformations of nanoparticles in sediments

Larger aggregates (~1000 nm) are ENMs deposited strained at joins strongly in primary energy minima

ENMs may sheer “Ripening” effect as off from trapped deposited or trapped aggregates due ENMs become Blocking of ENM owing to to hydrolic collectors of other hydrodynamic radius of pressure ENMs attached NPs preventing further binding to grains at high NP concentrations

ENMs suspended in the ENMs weakly pore waters available deposited in across organism–water secondary interface energy minima

Interstitial water

Sediment grain

Fig. 2. Physical interactions of engineered nanomaterials (ENMs) with sediment grains and potential for ENM transport within pore waters. unfavourable conditions.[121] This ripening effect has now been increased transport and suspension of ENMs in the interstitial found to occur also for mixtures of ENMs, with fullerene (C60) waters in freshwater sediments. However, in pore waters of high NPs showing reduced transport in the presence of TiO2 NPs as ionic strength, such as marine or estuarine sediments, penetra- [122] the C60 becomes retained by pre-deposited TiO2. Although tion of ENMs deep into the sediment is still likely to be [125] straining initially reduces transport of ENMs through sediments, limited, and retention profiles of TiO2 have shown most of these strained aggregates may act as a longer-term source of NPs the retained ENMs are concentrated in the top 4 cm of the into the pore waters as hydraulic forces shear particles from the saturated sand columns.[122,126,127] This reduced transport may surface of these aggregates.[109] Interestingly, favourable con- be exacerbated, as has been found for AgNPs, by heteroaggre- ditions for aggregation and deposition can actually result in gation with clay particles[124,128] and other natural colloids such greater transport of NPs in the pore waters. If, for instance, larger as maghemite or montmorillonite.[106] However, bioturbation in aggregates avoid straining, their size would allow movement areas with high densities of burrowing organisms may overcome only in wider pores with a greater flow rate and so they would these factors, which limit ENM mobility. For example, biotur- travel greater distances.[113] This is known as size exclusion. bation may lower the bulk density and rigidity of the surficial Another scenario in which conditions favourable for deposition sediment layer,[129] thus making it more prone to erosion can result in greater concentrations of ENM suspended in pore processes and resuspension of any sediment and associated waters is where there are high concentrations of the ENM. In NPs into the overlying waters. these instances, the surface of a sediment grain may become saturated with deposited negatively charged NPs, creating an electrostatic barrier preventing further deposition.[116,121] This Bioavailability of ENMs within sediments process of ‘blocking’ is particularly effective under lower ionic The different forms of deposition in sediments and the relative strength as the EDL of the ENMs expands, thus ‘blocking’ an distribution of ENMs between sediment grains and pore waters area of the grain surface much larger than the NP’s actual size. will affect the bioavailability of the ENMs to benthic-dwelling As such, point sources of high ENM concentrations could result organisms and in turn their toxicity. Ingestion has been estab- in high deposited and strained concentrations close to the source, lished as a dominant pathway for ENM uptake in aquatic which might then act as a constant release of NPs over time organisms. Association to food material increased both uptake through shearing from aggregates, size exclusion and blocking and toxicity of CdSe–ZnS quantum dots and AgNPs to Daphnia (Fig. 2). magna compared with those simply exposed in a water sus- [130,131] ENMs such as Ag and CeO2 are likely to be more mobile than pension. Deposition into the primary or secondary their ionic counterparts,[123] which are unlikely to persist as free energy minima of sediment grains therefore could increase ions in sediments owing to strong partitioning to organic exposure through ingestion for Oligochaetes and other matter.[124] In areas of high ENM concentrations, shearing of sediment-ingesting organisms. The limited penetration of MeOs strained aggregates, blocking and size exclusion all result in into surface sediments also makes them bioaccessible to those

G R. K. Cross et al. organisms living within the surface layer of sediments. This was toxicity is simply due to the release of free ions.[62,144–147] in evidence in a study where TiO2 spiked into the surface sed- Alternatively, both of these modes of toxicity can be considered iment at a depth of 1 mm resulted in greater uptake and toxicity under the toxic capacity of the NP, because the NP may act as a to Hyalella azteca and Chironomus dilutus than when spiked source or store for the later release of toxic ions. This question of deeper into the sediment (7 mm)[132] because these organisms persistence needs to be addressed using long-term studies feed primarily at the surface sediments. Coutris et al. found that investigating aging and its effect on toxicity. uncoated 19-nm AgNPs, although rapidly retained in soils Phosphate and nitrate also influence the chemical transfor- through deposition, were increasingly re-entrained in the pore mations of MeOs in the benthos. In the presence of phosphate, water over time.[133] These uncoated NPs reached concentra- there is a rapid reduction in the rate of dissolution of ZnO.[148] tions in the pore waters 8–9 times higher than for citrate- The presence of phosphate also altered the morphology of ZnO stabilised AgNPs or ionic AgNO3. This re-entrainment of NPs from structurally uniform spheres to a mixture of amorphous and into pore waters means that ENMs deposited onto sediment crystalline phases of ZnO and zinc phosphate (Zn3(PO4)2), grains may still act as a reasonably stable and continuous source which in turn could increase its persistence and alter reactivity of ENMs suspended in the pore waters, which are available for or bioavailability. Nitrate is also observed to increase persis- dermal uptake into sediment dwellers. tence of some ENMs; for example, it increased the stability of Fe0 to such an extent that after 6 months aging in sediments, the Fe0 and mineral content of aged samples differed little from Chemical transformations and the persistence of ENMs fresh samples of Fe0[149]. A major route of ENM discharge into in sediments waters is via sewage treatment plants and the high levels of Surfactants can dramatically increase the persistence and sulphur, nitrate and phosphate in these effluents are likely to mobility[134] of MeOs in sediments. ENMs coated with surfac- increase the stability and persistence of ENMs in the benthos. tants that confer steric stability are more resistant to aggregation and deposition. PVP-capped AgNPs had an attachment effi- ciency to quartz sand grains two orders of magnitude lower than NOM affects ENM mobility and the prevalence of dissolved that of pristine AgNPs.[135] Likewise, polyelectrolyte-coated ions within sediments zero-valent iron (Fe0) remained mobile in sediments even after Dissolution may occur in the water column or on uptake within 8 months aging, demonstrating the slow rate of degradation of organisms,[150,151] but within sediments, the rate of ENM dis- some surfactants.[136] Although many studies are investigating solution is reduced. Ions such as Agþ are transformed into NPs fate and toxicity of coated or stabilised NPs,[137–139] limited data within 4–14 days through reduction in the presence of humic on the production volumes or nature of the surfactants used by acid.[152] The presence of both humic and fulvic acids rapidly manufacturers mean it is not yet possible to accurately predict reduces ion release from AgNPs, even at low concentrations of which coatings of NPs will be most prevalent in the environment. 5mgL1, resulting in negligible dissolution at concentrations Several studies have looked at the longer-term chemical of 50 mg L1.[153] This strongly suggests a low prevalence of transformations of MeO NPs within sediments. Inorganic che- dissolved ions derived from metal ENMs within the benthos. micals, in particular, appear to have a significant and long-term NOM will not only reduce the prevalence of dissolved effect on the transformation products of ENMs and their ions from AgNPs in pore waters, but will also increase bioavailability and toxicity. Manufactured surfactants have transport distances for many ENMs through saturated been demonstrated to make ENMs highly persistent in the porous media. Humic acids in particular have been shown to [115] benthos and inorganic constituents of the sediment can act in a increase the mobility of nanoparticulate Fe3O4 and ZnO, [126] 0[154] [106] similar way. Such chemical transformations can dominate TiO2, Cu and Ag within pore waters. At low surface properties to the point where different ENMs behave concentrations, NOM-ENM complexation may provide elec- uniformly, irrespective of ENMs’ original surface properties. trostatic stability, reducing the rate of deposition by making the Dale et al. highlight that sulfurisation plays a dominant role in NP surface charge more negative.[126] At higher concentrations Ag fate within sediments,[140] overriding any difference of between 1 and 20 mg L1, NOM can also provide steric between surfactants, with ,70 % of citrate-capped, PVP- stabilisation,[134,155,156] preventing close proximity between capped and pristine AgNPs persisting as Ag2S in sediment pore the NP and the grain surface and making deposition in the waters when amended with sewage sludge.[141] This sulfidation primary energy minima impossible.[157] This steric repulsion is also limits the prevalence of dissolved ions in the pore waters. not affected by changing ionic strength, pH or electrolyte AgNPs dosed into saturated soils and bound to sediment grains composition.[158] Fulvic acids, however, have a smaller molec- were shown to release a constant supply of AgNPs into the ular size and so when coating ENMs, confer less steric stability interstitial waters over a 4-week exposure period. Particulate Ag and thus less resistance to deposition.[156] underwent some degree of dissolution before sulfidation Other components of NOM can reduce mobility of ENMs in occurred and the resulting ionic Ag was rapidly complexed with sediments. Examples of this include oxalic and adipic acids, pore-water chloride. This formed NPs smaller than the original which have been shown to reduce the stability of TiO2 when [142] [159] particles, 50–70 nm in diameter with an Ag2S shell. adsorbed to the NPs, promoting aggregation and straining of [143] Although the formation of a highly insoluble Ag2S shell these aggregates. The process of adsorption of organic acids to reduces toxicity caused by the release of Agþ ions, it does grant the NP was unchanged by either initial or aggregated particle it a half-life in sediments between 10 and .100 years depending size, suggesting attachment was uniform, irrespective of the size on the organic carbon content.[140] Where we draw the boundary of NP coming into contact with these organic acids. Another for what can still be described as the primary contaminant is example of where NOM can reduce ENM mobility is for uncertain. From an ecotoxicologist’s perspective, discerning polysaccharide NOM present in waste-water sludge and EPS, between dissolved free ions and the primary NP is key to which increased the retention of C60 NPs in saturated porous understanding if there is a nano-specific toxic effect, or if NP media.[160] Whether such a reduction in mobility also occurs for

H Transformations of nanoparticles in sediments

MeO ENMs has not yet been established. Many interactions total TiO2 available through ingestion should have been lower. with NOM initially reduce the mobility of ENMs in sediments, However, in this mesocosm, the P. acuta ingested .50 % of the through formation of aggregates leading to physical straining. total S. ulna biofilm, far more than any other biofilm and so Although most ENMs will experience reduced transport in ingestion and accumulation of TiO2 were greater. Patterns of sediments, the potential for re-entrainment of ENMs into pore behaviour can also reduce exposure of ENMs to organisms. The waters under some conditions from weak secondary deposition larva of the nematoceran fly Chironomus dilutes creates a tube onto sediment grain surfaces means a significant percentage of sand around itself in sediments, totally preventing surface [132] may still experience longer-term mobility. attachment of TiO2NPs. These studies highlight how the toxicity of an ENM is dependent not only on the inherent toxicity of the NP itself, but also on biological parameters Uptake of ENMs into benthic organisms defined by organism behaviour and how it partitions among Sediment ecosystems have historically acted as a buffer against benthic communities. extinction events. As an example, in the late Permian mass The characteristics of some exemplar benthic species of great extinction event, globally only one functional group (based on importance for hazard prediction for ENMs in the benthos are mode of life and functional morphology) was lost from the identified in Table 2, by defining their functional role within benthos.[161] However, at the local and regional scale, pertur- ecosystems and the route through which they will be exposed to bations to the functional diversity of benthic ecosystems can ENMs. The five groups outlined include a range of organisms have a profound effect on processes such as bioturbation[162] that dominate ecosystem engineering within the benthos and that are essential for global nutrient cycling and the productivity include groups that represent the benthic–pelagic coupling. of aquatic ecosystems.[163,164] The importance of this linking Routes of exposure are suggested, based on the evidence between the benthos and the overlying waters, known as the collected in the present review. Examples of species used to benthic–pelagic coupling is rarely referred to explicitly within investigate the question of NP hazard in the benthos are included the nano-ecotoxicology literature. Freshwater sediments act as a in the table along with the corresponding standardised chemical zone for diapause for zooplankton, an essential survival mech- ecotoxicity testing methods that are currently in use. Owing to anism that increases the resilience of zooplankton throughout the differences between solute and colloidal chemistry, thorough periods of unfavourable environmental conditions.[165] Benthic characterisation of NPs in the test media and their transforma- dwellers themselves act as ecosystem engineers and as such, tions is necessary to determine the suitability of these tests when their composition and diversity determine the functioning of using NPs, and in order to accurately interpret any results. entire aquatic ecosystems.[166–168] Given this, understanding the risk that ENMs pose to the benthos is important, not only to Conclusions and future recommendations predict the localised effects of ENMs, but also to understand the The various processes and transformations ENMs undergo on effects on the wider aquatic ecosystem. their release into the environment have the potential to drasti- Organisms exposed to ENMs within the benthos vary from cally alter their fate, form and toxicity. Realism in ecotoxico- sedentary organisms to endobenthic, sediment-ingesting Oligo- logical testing of ENMs requires a better understanding of these chaetes to the juvenile stages of some pelagic species.[169] processes to inform us on their fate and nature and in turn to Biofilms are the major biological compartment for partitioning better understand which compartments and organisms are most of ENMs when sedimenting to the benthos.[24] In a study on at risk of exposure. From our analysis of the available literature, positively charged cetyltrimethylammonium bromide (CTAB)- we identify that the following factors are critical in under- stabilised Au nanorods, 61 % was shown to partition into standing the fate and form of transformed ENMs within the biofilms,[170] reducing ENM mobility within sediments.[171] benthos. ENM toxicity towards these biofilms is of importance because they act as sediment biostabilisers, performing an essential role The presence of NOM and other biogenic exudates such as in the structure and functioning of the benthos.[172] Biofilms and EPS will dominate inorganic ENM behaviour in the aquatic other microbes also act as an important food source for many environment through heteroaggregation or by conferring benthic grazers. Association with algae may increase the assim- steric stability. EPS can derive from both biofilms and free- ilation efficiency of MeOs and so bioavailability. In Daphnia floating bacteria in the water column but can also arise as a magna, ,70 % of accumulated Ag was attributable to AgNPs response to stress in aquatic plants, and so ENMs will interact associated with algal food.[131] Some particles appear to become with EPS both in the water column and at the sediment toxic only when associated with phytoplankton or biofilms. For surface. example, SiO2 and CeO2 NPs associated with algal food reduced The formation of ‘marine snow’ from EPS and its potential to sea urchin larvae survival.[173] This was also demonstrated with act as a vehicle for sedimentation of ENMs out of the water CdSe–ZnS quantum dots, which were toxic when ingested in column may present a seasonal flux of ENMs to the benthos in association with algal food in D. magna but not when freely both fresh and marine waters. This is not currently accounted suspended, even though bioaccumulation occurred for both for in attempts to model environmental concentrations of exposure routes.[130] ENMs in the benthos. Feeding patterns of organisms will also determine the bio- Aggregation, complexation with NOM and EPS, and chemi- availability of ENMs to benthic dwellers via dietary uptake. The cal transformations such as sulfidation all reduce rates of grazing aquatic snail Physa acuta selectively grazes on biofilms dissolution and the release of toxic metal ions from ENMs. based on palatability. In stream mesocosm studies, various Therefore, ENMs within the sediments are likely to be in the biofilms were exposed to a range of concentrations of TiO2 particulate form, prevailing as large aggregates or complexed and the greatest accumulation of the NP in P. acuta was when with NOM and EPS. feeding on biofilms of the diatom Synedra ulna.[174] S. ulna Complexation with NOM and EPS results in deposition produced the lowest biomass of all the mesocosms tested and so into the weaker secondary energy minimum dominating NP

I Table 1. Important surface and subsurface transformations of engineered nanomaterials (ENMs) and our understanding of their effect on fate of ENMs in aquatic sediments NPs, nanoparticles; PVP, polyvinylpyrrolidone; OC, organic carbon

Transformation Environmental ENM composition (pri- Outcome Reference compartment mary particle size) Natural organic matter Surface waters ZnO (4 nm) Dissolved Suwannee River humic acid at 7.3 mg L 1 resulted in zeta potential below 30 mV and reduced Bian et al. 2011[44 ] forms heteroaggregates sedimentation over 120 min. Lower levels of HA (1.7 mg L 1) caused greater sedimentation through charge that can stabilise ENMs neutralisation by the humic acid of the NPs’ positive charge, resulting in greater aggregation. ZnO (24 nm), TiO2 All NPs formed aggregates ,300 nm in diameter that were stable and experienced no sedimentation over 7 h in Keller et al. [50] (27 nm) and CeO2 (8 nm) mesocosm freshwater with total organic carbon of 5.3 mM. 2010 1 [71 ] TiO2 (30 nm) Suwannee River humic acid at 10 mg L stabilised TiO2 NPs through both electrostatic and steric repulsion, Thio et al. 2011 þ particle zeta potential being between 30 and 40 mV at 0.1–100 mM NaCl (Na cation dominates marine þ waters). Ca , the cation that is prevalent in freshwaters, was found to be more effective at causing aggregation at lower concentrations than NaCl; this effect was negated by the presence of the humic acid. TiO2 nanorod core, coated A TiO2 nanocomposite used in sunscreens; 35 % of the particles remained in suspension as aggregates ,300 nm Labille et al. 1 [72] with Al(OH)3 and poly- after 48 h. Humic acid at 2 % w/w (,1mgL ) caused flocculation through bridging between NPs, whereas at 2010 dimethylsiloxane layers 20 % w/w, the NP was stabilised, experiencing no sedimentation over 48 h. (10 50-nm TiO2 core) Filter feeders may alter the Surface waters and Mercaptoundecanoic Particles aggregated to 10 primary particle size, averaging ,300 nm. In presence of Daphnia magna, this Park et al. 2014[90 ] size distribution of sediment–water acid-capped Au NPs decreased to ,200 nm within 10 h. In the presence of Gammarus pulexa, size stabilised at ,150 nm over 100-h ENMs reaching the boundary (30 nm) exposure. benthos Natural inorganic colloids After 310 min in the presence of Daphnia magna, inorganic colloids ,400 nm increased to between 100 and 130 % Filella et al. in lake water (50– of their original concentration before the organisms were added. Colloids .400 nm decreased to a minimum of 2008[89]

J 2000 nm) ,40 % of their original concentration.

1 Natural organic matter can Sediment AgNO3 (ionic) Silver nitrate solutions were exposed to 1–100 mg L Suwannee River humic acid and three sedimentary humic Akiaghe et al. þ reduce presence of dis- acids, all of which reduced the Ag ions, forming AgNPs within 2 to 4 days at 22 8C. NPs had a wide size 2011[152 ] solved ions distribution but with similar means between 70 and 90 nm found by dynamic light scattering. Citrate-stabilised Ag NPs 5mgL 1 of both Suwannee River humic and fulvic acid reduced dissolved Ag by 1 order of magnitude, from Liu et al. 2010[153 ] (2–8 nm) ,15 % of total Ag to ,1%. 0 Sulfurisation increases Sediment Modelled Ag NPs Modelled half-life of Ag NPs (both Ag and AgS2) varied with OC, with low OC half-life of 6.6 years, mid OC 77 Dale et al. persistence but reduces years and high OC 280 years. 2013[140 ] dissolution PVP-Ag NPs (60 nm) When aged in the presence of sewage sludge, 70 and 78 % of the PVP and citrate-coated NPs respectively formed Whitley et al. [141 ] Citrate Ag NPs (60 nm) AgS2. Aging reduced Ag ions in the pore water to ,2 % of total Ag. 2013 Natural organic matter Sediment Fe3O4 (,50 nm), TiO2 All ENM were aggregated in test water .100 nm. Mobility decreased in order of TiO2 . CuO . ZnO . Fe3O4. Ben-Moshe et al. 1 [115 ] increases ENM transport (,100 nm), CuO 62 % of TiO2 was initially eluted from the bead column, but in the presence of 60 mg L humic acid, this 2010 in saturated porous (,100 nm) and ZnO increased to 98 % for both TiO2 and CuO. media (,50 nm) 1 TiO2 (10 40 nm) At pH 5.7, humic acid (1 mg L ) reduced deposition onto sediment grains, increasing transport through elec- Chen et al. ,90 nm in test trostatic and steric effects. At pH 9, the effect of NOM was limited because humic acid did not adsorb well to the 2012[126 ] conditions TiO2 NPs. [122 ] Co-transport of suspended Sediment TiO2 (25 nm) 200 nm in At pH 7, transport of TiO2 increased with increasing C60 concentration, as the C60 competed for deposition sites on Cai et al. 2013

colloids and ENMs test conditions and C60 quartz sand grains.TiO2 NPs that were not deposited acted as additional sites for deposition, reducing transport of al. et Cross K. R. (135–505 nm in test C60 when in the presence of TiO2. conditions) PVP-capped AgNPs Ag NPs experienced rapid heteroaggregation with maghemite or montmorillonite. In soils with a higher porosity, Cornelis et al. (10 nm) 40 nm in test this leads to size exclusion and greater heteroaggregate transport. In most soils, this reduced mobility owing to 2013[106 ] conditions primary deposition, resulting in a maximum transport depth of 12 cm. Transformations of nanoparticles in sediments

interactions with sediment grains, increasing the mobility of many larger inorganic ENM aggregates within saturated porous media. Therefore, sediment-dwelling organisms will be at risk from a variety of routes to exposure, including ENMs in suspension in the interstitial waters, bound to sediment grains and at the sediment surface where they may be recirculated into the water column through bioturbation. Biofilms at the sediment surface have the capacity to adsorb ENMs effectively and play an important role both in the cycling of nutrients within aquatic ecosystems and as primary producers. As such, the high incorporation of ENMs into 06(2013) 20745761 20745761, OECD test number 31510.1787/2074577x, doi: EPA 600-R-99–064 available None available to date Recommended by OECD, ASTM E2455– OECD test number 233 doi:. 10.1787/ OECD test number 225 doi:. 10.1787/ Standardised chemical testing protocols biofilms could both disrupt local nutrient cycling and act as a source for trophic transfer of ENMs between benthic- dwelling organisms, because many benthic organisms graze on these microbial biofilms. We would argue that better understanding of the complex nature of the transformations of ENMs within the aquatic environment is essential for developing standardised (and appropriate) testing methods for the ecotoxicity of ENMs. Transformations within sediments represent a significant gap in our knowledge and our recommendations for the immediate future of research needs in this field are: Studies into bioaccumulation and toxicity of ENMs to benthic species should include sufficient characterisation of the ENMs in the test water and where possible in the sediment the sediment surface layer’ or incorporation into pseudofaeces planktotrophic benthic species larvae orsediments in for pelagic species larvae exposure possible through pore waters Adsorption in the ‘fluff layer’ and settling on Filter feeding out of suspension in the ‘fluff Potential exposure through root systemsEither ingestion from the water column for OECD test number 239 Ingestion of sediment-deposited ENMs, some Expected route to exposure? matrix to examine the processes and transformations that dictate patterns of ENM fate and bioavailability.

, Studies using manufactured ‘transformed’ ENMs such as AgNPs after sulfidation are needed to understand how such transformations affect ENM fate, behaviour and toxicity to Chron- benthic organisms compared with ‘pristine’ particles. Egeria densa

, Assessment of the potential for fluxes of ENMs carried in Tubifex tubifex , ‘marine snow’ to act as a vector for NP sedimentation to the Crassostrea gigas , benthos and to include such transport systems in predictive modelling of environmental concentrations of ENMs. Toxicity and bioaccumulation tests on a range of the benthic-

larvae dwelling species identified in the present review (Table 2)as being of particular risk of exposure, including aquatic plants

sediments; epipelic in clay-based sediments Myriophyllum spicatum omidae and biofilms. Epipsammic diatom species in sandy Lumbriculus variegatus Scrobicularia plana Potamogeton diversifolius Dormancy-phase zooplankton, Species examples References [1] M. C. Roco, C. A. Mirkin, M. C. Hersam, Nanotechnology Research Directions for Societal Needs in 2020: Retrospective and Outlook ] 2011 (Springer Netherlands). doi:10.1007/978-94-007-1168-6 165 [

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