Science of the Total Environment 850 (2022) 157912 Contents lists available at ScienceDirect Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv Metal transfer to sediments, invertebrates and fish following waterborne exposure to silver nitrate or silver sulfide nanoparticles in an indoor stream mesocosm Nathaniel Clark a, Joanne Vassallo a, Patrícia V. Silva b, Ana Rita R. Silva b, Marta Baccaro c, Neja Medvešček d, Magdalena Grgić e, Abel Ferreira b, Martí Busquets-Fité f, Kerstin Jurkschat g, Anastasios G. Papadiamantis h,i, Victor Puntes j,k,l, Iseult Lynch h, Claus Svendsen m, Nico W. van den Brink c, Cornelis A.M. van Gestel n, ⁎ Susana Loureiro b, Richard D. Handy a, a School of Biological and Marine Sciences, University of Plymouth, Plymouth, UK b Department of Biology and CESAM, University of Aveiro, Campus Universitário de Santiago, 3810-193 Aveiro, Portugal c Department of Toxicology, Wageningen University, Wageningen, the Netherlands d Department of Biology, Biotechnical Faculty, University of Ljubljana, Večna pot 111, 1000 Ljubljana, Slovenia e Department of Biology, Josip Juraj Strossmayer University of Osijek, Cara Hadrijana 8/A, 31000 Osijek, Croatia f Applied Nanoparticles SL, C Alaba 88, 08018 Barcelona, Spain g Department of Materials, Oxford University Begbroke Science Park, Begbroke, United Kingdom h School of Geography, Earth and Environmental Sciences, University of Birmingham, Edgbaston, B15 2TT Birmingham, UK i NovaMechanics Ltd., 1065 Nicosia, Cyprus j Institut Català de Nanociència i Nanotecnologia (ICN2), CSIC, The Barcelona Institute of Science and Technology (BIST), Campus UAB, Bellaterra, Barcelona, Spain k Institució Catalana de Recerca i Estudis Avançats (ICREA), 08010 Barcelona, Spain l Vall d'Hebron Institut de Recerca (VHIR), 08035 Barcelona, Spain m Centre of Ecology and Hydrology (CEH-NERC), Wallingford, UK n Amsterdam Institute for Life and Environment (A-LIFE), Faculty of Science, Vrije Universiteit Amsterdam, the Netherlands H I G H L I G H T S G R A P H I C A L A B S T R A C T • Mesocosm experiments allow a more real- istic understanding of nanomaterial fate. • Silver as AgNO3 or Ag2S NPs caused simi- lar water and sediment contamination. • Ag-containing particles were found in the water of both Ag-treatments. • At day 14, Ag-treatment had no difference in Ag uptake in the snails or worms. • Silver as AgNO3 was more bioavailable to planarians and fish than Ag2S NPs. ⁎ Corresponding author. E-mail address: r.handy@plymouth.ac.uk (R.D. Handy). http://dx.doi.org/10.1016/j.scitotenv.2022.157912 Received 9 June 2022; Received in revised form 25 July 2022; Accepted 4 August 2022 Available online 8 August 2022 0048-9697/© 2022 The Authors. Published by Elsevier B.V. This is an open access article under the CC BY-NC license (http://creativecommons.org/licenses/by-nc/4.0/). N. Clark et al. Science of the Total Environment 850 (2022) 157912 A R T I C L E I N F O A B S T R A C T Editor: Julian Blasco The fate of engineered nanomaterials in ecosystems is unclear. An aquatic stream mesocosm explored the fate and bio- accumulation of silver sulfide nanoparticles (Ag2S NPs) compared to silver nitrate (AgNO3). The aims were to deter- Keywords: mine the total Ag in water, sediment and biota, and to evaluate the bioavailable fractions of silver in the sediment Freshwater sediments using a serial extraction method. The total Ag in the water column from a nominal daily dose of 10 μg L−1 of Ag for Engineered nanomaterials the AgNO3 or Ag2S NP treatments reached a plateau of around 13 and 12 μg L−1, respectively, by the end of the Benthic and planktonic invertebrates Trout study. Similarly, the sediment of both Ag-treatments reached ~380 μg Ag kg−1, and with most of it being acid- Silver uptake extractable/labile. The biota accumulated 4–59 μg Ag g−1 dw, depending on the type of Ag-treatment and organism. Metal partitioning The oligochaete worm, Lumbriculus variegatus, accumulated Ag from the Ag2S exposure over time, which was similar to the AgNO3 treatment by the end of the experiment. The planarian, Girardia tigrina, and the chironomid larva, Chironomus riparius, showed much higher Ag concentrations than the oligochaete worms; and with a clearer time- dependent statistically significant Ag accumulation relative to the untreated controls. For the pulmonate snail, Physa acuta, bioaccumulation of Ag from AgNO3 and Ag2S NP exposures was observed, but was lower from the nano treat- ment. The AgNO3 exposure caused appreciable Ag accumulation in the water flea, Daphnia magna, but accumulation was higher in the Ag2S NP treatment (reaching 59 μg g−1 dw). In the rainbow trout, Oncorhynchus mykiss, AgNO3, but not Ag2S NPs, caused total Ag concentrations to increase in the tissues. Overall, the study showed transfer of total Ag from the water column to the sediment, and Ag bioaccumulation in the biota, with Ag from Ag2S NP exposure gener- ally being less bioavailable than that from AgNO3. 1. Introduction In general, Ag NPs are less toxic than silver salts such as AgNO3, with le- thal concentrations for the former in the range ~360, ~10, and ~1360 μg Silver is well known for its antimicrobial properties, and colloidal solu- L−1 for algae, crustaceans and fish, respectively (Bondarenko et al., 2013). tions containing silver have been used for many years in photography and Johari et al. (2013) reported a 96 h LC50 for Ag NPs in juvenile rainbow microbiology (Silver, 2003). Silver-containing engineered nanomaterials trout of around 2160 μg L−1. In freshwater invertebrates, such as Hydra (ENMs) are also one of the most frequently reported ENMs in consumer vulgaris, and the shrimp, Paratya australiensis, the 96 h LC50 values are products (Vance et al., 2015). Around one thousand tons or more of around 1941 and 55 μg L−1, respectively (Lekamge et al., 2018). However, Ag-containing ENMs are produced annually, with applications in food pack- the extrapolation of toxicity data from laboratory studies to natural waters aging, food supplements, medical devices and biocides; contributing to the or field situations is challenging. Ag NPs may be adsorbed to organic matter estimated 150–200 billion Euro economic value of nanotechnology (Laux in the water column, show aggregation depending on the factors from col- et al., 2018). Inevitably, the manufacture and use of ENMs is contributing loid chemistry (pH, ionic strength, presence of divalent ions, organic mat- to their release into the environment. Early exposure models gave predicted ter, etc.), or be chemically transformed (Dale et al., 2013; Lead et al., environmental concentrations (PECs) for Ag nanoparticles (NPs) ranging 2018; Levard et al., 2012; Zhang et al., 2019). Notably, pristine Ag NPs from ng L−1 in surface waters to mg kg−1 in sediments (Fabrega et al., are transformed into much less toxic, but very persistent, Ag2S NPs when 2011). In Europe, the predicted mean concentration of Ag from ENMs sulfide is present (e.g., in sediments), or in the reducing conditions of was 4.15 ng L−1 and 88.2 μg kg−1 in surface waters and sediments respec- poorly oxygenated water (Dale et al., 2013; Lead et al., 2018). Indeed, pris- tively, by 2020 (Sun et al., 2016). Similar predictions were made for tine Ag NPs are rapidly transformed into Ag2S NPs during wastewater treat- German watercourses, with an anticipated six-fold increase by 2050 (Giese ment processes (Kaegi et al., 2013) and Ag2S NPs are the likely form et al., 2018). Field measurements are beginning to emerge, and in the Seine discharged to the environment. While the Ag2S form is less toxic than pris- River watershed of France, measured concentrations ranged from 0.4 to tine Ag NPs (e.g., to Chironomus riparius (Lee et al., 2016); adult zebra fish, 28.3 ng Ag L−1 (Wang et al., 2020). Clearly, there are concerns for Ag- Danio rerio (Devi et al., 2015)), it is not clear if the bioaccumulation hazard containing ENMs in surface waters, and with sediment as a possible sink, remains for Ag2S NPs. In at least one study with rainbow trout, Ag2S NPs the bioaccumulation hazard and effects on biota in real ecosystems is were much less bioavailable than either the metal salt or Ag NPs (Clark not clear. et al., 2019a), with similar observations made in snails and planarians Historically, the main environmental releases of dissolved silver were (Silva et al., 2022). However, in fish at least, nanoparticles arising from from mining and smelting, biocides and the photographic industry, with the Ag2S NP exposure were also detected in the internal organs (Clark standards for total Ag in surface waters set around 0.1 μg L−1 et al., 2021), indicating that some bioaccumulation of particulates oc- (e.g., Canada, Purcell and Peters, 1999). Unfortunately, silver is the second curred. most toxic metal to aquatic organisms in the periodic table (Ratte, 1999). Clearly, there are a myriad of abiotic and biotic factors in natural ecosys- For freshwater fish, the 96 h 50 % lethal concentration (LC50) for AgNO3 tems that may influence the fate and behaviour of ENMs (Lead et al., 2018). is 5–70 μg Ag L−1 (0.05–0.65 μM) with the toxicity resulting from One useful approach has been to conduct mesocosm studies that offer some the ionic form of silver (Hogstrand and Wood, 1998). For freshwater inver- environmental realism, while also retaining the benefits of a controlled tebrates such as Daphnids, the acute toxicity of Ag+ is around 0.3–4 μg L−1, ‘laboratory’ experiment (Holden et al., 2016). Early mesocosm exposures depending on the exposure conditions (Naddy et al., 2007). In freshwaters, with inert gold nanorods showed transfer of the ENMs from the water col- dissolved silver may preferentially bind to natural organic matter and/or umn to mainly the estuarine sediment (25 % of the dose) and biofilms form less soluble chloride complexes, such that the free ion activity of (61 %), with limited accumulation in the plants and animals present Ag+ is low. Consequently, there is a tendency for total Ag to deposit into (Ferry et al., 2009). Similarly, Furtado et al. (2015) found that total Ag sediments with a risk of bioaccumulation for the local biota (Hogstrand from Ag NPs added to freshwater boreal lake mesocosms was transferred and Wood, 1998; Ratte, 1999). In freshwater, algae and invertebrates can from the water column mainly to the sediments and periphyton biofilm. have high bioconcentration factors (BCFs) for silver salts (from 4.8 to In wetland mesocosms, both Ag NPs and Ag2S NPs accumulated mainly >105 for algae, 9 to >2200 for invertebrates, (Ratte, 1999)), with conse- in/on the surface of sediments and aquatic plant materials (Stegemeier quent concerns for dietary bioaccumulation in fish. et al., 2017). The fate of Ag ENMs in mesocosms may also depend on the 2 N. Clark et al. Science of the Total Environment 850 (2022) 157912 shape of the materials, and with some evidence of sulfidation of the materials biting midge) (150 per river), Oncorhynchus mykiss (rainbow trout) (3 per in the benthic compartment (Auffan et al., 2020). However, mesocosm stud- river), Daphnia magna (water flea) (0.25 % ration for the fish equating to ies with Ag-containing ENMs have mostly been performed in static systems to approximately 13 Daphnia per river per day), Lumbriculus variegatus (oligo- simulate a lake or a pond. River mesocosms are less well understood. chaete worm) (approx. 900 mg fresh weight per river). The life stages In our previous work, Ag2S NPs, as the most relevant environmental were as follows: snails, 2 months old; chironomid larvae, 6 to 7 days old, transformation (form) of Ag NPs, was explored for its bioaccumulation adult planarians (length around 1.5–2 cm), adult oligochaete worms potential in single species studies on: dicotyledonous crop, Khodaparast (~6–10 mg fresh weight), Daphnia (adults and juveniles from the culture et al. (2022); wheat crop, Lahive et al. (2021); earthworms, Baccaro et al. to ensure enough food for the fish), fingerlings of rainbow trout (<5 g). (2019); meal worms, Khodaparast et al. (2021); planarians, Silva et al. Rainbow trout were kept in rectangular plastic chambers (dimensions: (2022); aquatic snails, Silva et al. (2020) and trout, Clark et al. (2019a). height 11.5 cm, length 16.5 cm, width 14 cm), weighted to keep them sub- In the present study, the same Ag2S NPs were used in indoor stream merged, and with plastic mesh-covered apertures (1 mm mesh) on both mesocosms. The fate and bioaccumulation of total Ag from the Ag2S NP ex- sides of each chamber to allow free water movement, but also to prevent posure compared to that of AgNO3 was measured. The specific objectives the invertebrates from entering the box. This was done so that the food were to analyse the water and sediment samples to confirm and understand items eaten by the fish in each river could be managed and to restrict the the dynamics of the exposure, and to evaluate the bioavailable fractions of fish to the water column. Similarly, for the unexposed control and Ag2S silver in the sediment using a serial extraction method. The mesocosms NP treatments, the daphnids were kept in cylindrical plastic chambers contained a range of invertebrates that occupied different niches and tro- (dimensions: height 16 cm and diameter 11.5 cm, apertures covered by phic levels (planarians, oligochaete worms, chironomids, snails, daphnids) 1 mm mesh). These Daphnia were restricted to their chamber in the water and with rainbow trout as a ‘top’ predator in the food chain. The total Ag column and were collected to feed the fish each day. Two weeks prior to concentrations in the biota were determined in order to understand the bio- the start of the experiment, unglazed ceramic tiles (20 cm2) were incubated accumulation in the different organisms, and related changes in trace elec- in the Mau river (an unpolluted area in Sever do Vouga, Portugal) to allow a trolytes in these organisms. Furthermore, where possible the study natural colonization of biofilm to serve as a food source for the snails. determined the particle number concentration in the water using single Daphnids are especially sensitive to dissolved silver compared to the particle inductively coupled plasma mass spectrometry (spICP-MS). other invertebrates used in the mesocosms, and in order to create an Ag- contaminated food source for the fish in the mesocosms with 10 μg L−1 2. Methodology AgNO3, the Daphnia were exposed in a separate aquarium of APW (not di- rectly exposed within the mesocosm) at a nominal concentration of 1 μg Ag 2.1. Experimental design for the mesocosm studies L−1 (i.e., 10-fold less to give a sub-lethal dose to the Daphnia). This was lower than the nominal 10 μg L−1 of total Ag used in the main mesocosm, The experiment was conducted in an indoor modular mesocosm system but was necessary to avoid mortality, considering the reported LC50 at 48 h at the University of Aveiro, Portugal, in April 2018. The study design of exposure to AgNO3 of 3.38 μg L−1 with food and 1.04 μg L−1 without consisted of 36 artificial rivers made of glass (length: 2 m, width: 0.200 m food for Daphnia (Ribeiro et al., 2014). Daphnids were fed every other and depth: 0.225 m). The rivers were arranged in triplicates, where each day with the microalgae Raphidocelis subcapitata at concentration of 3 × triplicate set was fed by water from the same sump (102 L) at a flow rate 105 cells mL−1. The daphnids were then used as ‘contaminated food’ for of 4 L min−1. A water reservoir tank with 800 L capacity supplied the the rainbow trout in the mesocosms. Trout were fed daily on a 0.25 % ra- water to the system. The plumbing was designed such that each set of tion of daphnids (equating to about 13 Daphnia per river). The other food three rivers could operate on flow-through (i.e., single pass for flushing chain in the mesocosm involved the snails in the mesocosm that fed on the system) or be recirculated. The latter mode was used for the experi- the biofilm on the ceramic tiles or growing on the sediment/glass walls of ments. Each river was filled with 7 kg of sediment (99 % of sand with each river. The planarians may have predated on the snails, chironomids <2 mm grain size, previously burned at 500 °C for 4 h) mixed with or Lumbriculus and the chironomids grazed on the organic matter (ground grounded alder leaves (Alnus glutinosa) (1 % w/w), to provide organic mat- Alder leaves) in the sediment. ter and serve as the food source for chironomid larvae. The sediment was The unit of replication in the study design was a group of three rivers fed carefully spread over the bottom of each river to a thickness of approx. by a shared sump. There were 12 sets of three rivers in all (4 sets per treat- 2 cm. A sandy sediment was chosen for several reasons. Firstly, ecological ment). Groups of three rivers were randomly designated either as controls relevance since river ecosystems especially in Portugal do have areas of (no added silver), or exposed to 10 μg L−1 of total Ag via the water as either 100 % sand (Balsinha et al., 2009), but mainly to minimise the risk of silver nitrate (AgNO3, analytical grade from Sigma Aldrich) or as silver sul- Ag2S NPs being further transformed by geochemical reactions with clays, fide nanoparticles (Ag2S NPs, batch number 6, from Applied Nanoparticles, silt and complex NOM (Dale et al., 2015), and pragmatically to enable spe- Barcelona). We have characterised Ag2S from Applied Nanoparticles many ciation calculations of the Ag species present in the mesocosm. The organic times before (e.g., Clark et al., 2019a; Clark et al., 2019b), with details in matter was kept to 1 % to minimise the risk of fungal growth in the Silva et al. (2020) for the aquatic media used here, and the specific batch mesocosm and to be consistent with our previous work. Each river was used in the current experiments in Peixoto et al. (2020). The Ag2S NPs then filled with 35 L of ‘artificial pond water’ (APW, Naylor et al., 1989; (Manufacturer's information) had a nominal size and concentration of Vidal et al., 2014). The APW medium was prepared by filling the reservoir 20 nm and 1.32 g Ag2S L−1, respectively, with the latter confirmed directly tank with deionised water and adding the following solutions: 1.99 mmol before use in the experiment. The Ag2S NPs were supplied in a colloidal dis- L−1 CaCl2·2H2O, 0.499 mmol L−1 MgSO4·7H2O, 0.77 mmol L−1 persion stabilized by 1 mg mL−1 of 55 kDa polyvinylpyrrolidone (PVP). NaHCO3, 0.08 mmol L−1 KCl, 0.046 mmol L−1 K2HPO4, 1 mmol L−1 Transmission electron microscopy was conducted to confirm the primary NaNO3, 0.107 mmol L−1 NaSiO3·9H2O, at a pH ranging from 7.81 to particle diameters of the Ag2S NPs to be 20.4 ± 11.9 nm (mean ± S.D., 7.90, at day zero. The final water volume in each unit (sump + 3 rivers) n = 613, Peixoto et al. (2020)). Dynamic light scattering was performed was approximately 207 L. The APW represents a moderately hard freshwa- in artificial pond water with 1 mg L−1 dispersions of the Ag2S NPs and ter with measured concentrations (Table S1) of Na+ (~52 mg L−1 or gave a mean (±S.D.) hydrodynamic diameter of 336 ± 26.6 nm (by inten- 2.26 mmol L−1), K+ (~7.8 mg L−1 or 0.2 mmol L−1), Mg2+ (~14 mg sity at pH 7.8, n = 3). The measured total Ag concentrations and particle L−1 or 0.57 mmol L−1), and Ca2+ (~83 mg L−1 or 2.07 mmol L−1). number concentrations in the rivers are reported below. The rivers were initially filled with freshly prepared APW and left to ac- The exposure to AgNO3 was achieved by preparing a stock solution of climate for 2 days for equilibration of the water chemistry. After that, the 13.04 mg L−1 of total Ag in ultrapure water and adding 250 mL of the following invertebrates were introduced in each river: Physa acuta (snail) stock solution to the sump that fed each of the three rivers, allowing it to (60 per river), Girardia tigrina (planarian) (50 per river), C. riparius (non- mix to achieve a final nominal concentration of 10 μg L−1 of total Ag. For 3 N. Clark et al. Science of the Total Environment 850 (2022) 157912 the exposure to Ag2S NPs, the required volume of total Ag was added di- 2.3. Silver and electrolyte determination in biota rectly from the liquid stock as supplied by Applied Nanoparticles (with stock concentration being of 1.15 g Ag L−1) to the sump, then also allowed Elemental analysis in biota followed the method of Clark et al. (2019a). to mix to achieve a final nominal concentration of 10 μg Ag L−1 of total Ag. For each replicate stream, typically three animals were pooled from Each unit of three rivers was dosed daily at 10 a.m. to ensure the exposure that stream to get enough biomass to analyse (in mg dry weight): snails, of 10 μg Ag L−1 was maintained. This concentration was selected to be sub- 2–9 mg; planarians, 2–10 mg; Lumbriculus 30–80 mg; chironomid larvae, lethal to all of the test species (except for daphnids and AgNO3, as explained 6–18 animals to obtain a dry mass of 2–6 mg; Daphnia, 15–30 animals to above) and to enable a physiological accumulation of Ag in the organisms, obtain a dry mass of 3–5 mg. Tissues were freeze-dried (Lablyo freeze but also to facilitate measurement of Ag in the system. Each triplicate set of drier), then weighed and digested in Eppendorf tubes (n = 2 to 4 samples) rivers was exposed for up to 14 days to one treatment (AgNO3 or Ag2S NPs), with 0.5 mL of neat nitric acid for 2 h in a water bath set at 65 °C. After and the rivers were subjected to destructive sampling on either day 2, 7, or cooling, the digests were diluted to 1.5 mL using ultrapure deionised water 14. The temperature was set at 15 ± 1 °C and a photoperiod of 16 h light: and stored in the dark until subsequent analysis by ICP-MS for total Ag. 8 h dark was applied. Following metal analysis by ICP-MS, the digest samples were diluted 1:2 The rivers were measured daily for routine water chemistry (dissolved with 2 % (v/v) nitric acid and analysed by ICP-OES for Na, K, Ca, Mg, Fe, oxygen (DO), pH, temperature, total ammonia, electrolytes) and for total Cu and Zn. The certified reference tissue DORM-4 (National Research Council Ag determination. In addition to the daily water quality measurements on Canada) was also analysed as above for total silver (80.9 ± 15.2 as mean per- days 1, 6, and 13, water samples were collected hourly over the first four centage recovery ± S.D., n = 20, with percentage coefficient of variation = hours after dosing and after 24 h before the next dosing, in order to under- 18.8). stand the exposure profile within a single dosing period. Water and sedi- ment samples were also collected along the length of each river for Ag 2.4. Single particle ICP-MS for particle size and particle number concentration determination, near the inflow, middle, and outflow of each river on the day of destructive sampling. Biota were sampled for total Ag analysis on Water was sampled on days 1, 2, 6, 7, 8, 10, 13 or 14 to determine if the days 2, 7, and 14 (end of the experiment), except the fish which were with- total Ag measured in the water column included particulate forms of Ag. A drawn from the study early for ethical reasons and sampled on days 2 and 6 10 mL volume of each river water was taken and stored at 4 °C until analy- only. In each sampling time and for each treatment, invertebrates were col- sis. Analysis was performed according to Clark et al. (2021), with minor lected from 4 rivers (one river per replicate) with the help of Pasteur pi- modifications. The particulate content of the water was measured using pettes and placed in containers with clean APW medium for a 24 h an iCAP RQ ICP-MS (Thermo Fisher) with the same settings as previously depuration period. Afterwards, the organisms were rinsed in ultrapure reported (Clark et al., 2021) in time resolved analysis mode. Each sample deionised water, pooled in Eppendorf tubes and frozen at −80 °C until re- was analysed for 60 s by measuring the Ag107 m/z ratio with a 3 ms dwell quired for analysis. time. The sample wash out time was 60 s, and between samples a wash so- lution of 4 % HCl and 2 % HNO3 was used. The sample flow rate was calcu- 2.2. Metal analysis of water and sediment samples lated by aspirating deionised ultrapure water and determining the difference gravimetrically over 2 min (n = 5). The transport efficiency A volume of 10 mL of mesocosm water was taken from each river for was calculated using a well characterised 60 nm Au NP standard (BBI solu- total Ag determination (n = 4 per treatment at days 2, 7 and 14) and acid- tion, UK). The instrument was calibrated using dissolved standards ranging ified with 0.5 mL of neat nitric acid (Fisher, Primar Plus Trace Metals Anal- from 0.5 to 4 μg L−1. All solution and suspension preparation, and ICP-MS ysis Grade). The samples were then analysed for Ag, Cu, Zn and Fe using analyses, were conducted in a laboratory managed under an ISO 9001 cer- ICP-MS and Ca, Mg, K and Na using inductively coupled plasma optical tified Quality Management system. Each sample (standards and unknowns) emission spectrophotometry (ICP-OES). It was also expected that the produced 20,000 data points that were used to calculate the particle num- daily dosing with Ag would result in a peak concentration of Ag that then ber concentration using a bespoke Excel sheet that has been previously equilibrated in the system in the subsequent 24 h before the next dosing. assessed (Clark et al., 2019c), and were standardised to the number of Therefore, water samples were collected more frequently (at 10 min, 1 h, particles L−1. 2 h, 4 h and 24 h after dosing) to understand the 24 h profiles of Ag concen- trations in the water column at the start, middle and end of the experiment. 2.5. Statistics These were analysed as above. The total concentration of Ag, along with the elements Ca, Mg, Na, K, Unless stated otherwise, all data are shown as mean ± standard error of Zn, Cu and Fe, present in the sediments was determined at days zero, 2, 7 the mean (S.E.M). Following descriptive statistics, the Kolmogorov- and 14. Approximately 30 g of the moist sediment (n = 12 samples per Smirnov test was used to assess the normality of the distribution of the treatment) was oven dried to constant mass at 85 °C. Sub-samples of data. Independent Student's t-tests and one-way analysis of variance about 250 mg of dry sediment were digested using 10 mL of neat nitric (ANOVA, Tukey post hoc test) were used to check for significant differences acid in covered 50 mL glass beakers. The acid mixture was allowed to amongst responses from the treatments. Data for fish tissues were statisti- gently simmer at 110 °C on a hotplate for at least 2 h in the fume hood cally analysed as described in Clark et al. (2019a). Briefly, after checking (adapted from Chen and Ma, 2001). The digests were then cooled and di- for normality (Shapiro-Wilk test) and equal variance (Brown Forsythe), luted with 2 % (v/v) nitric acid into 25 mL volumetric flasks. They were data were analysed by one-way ANOVA (e.g., treatment effects), or two- transferred to propylene tubes, tightly stoppered and stored in a dark way ANOVA (e.g., treatment by time effects). Where data were not para- place until required for ICP-OES or ICP-MS analysis, as appropriate. metric, the Kruskal-Wallis test was applied. Figures were prepared using The extractable Ag fractions in all dried sediment samples were deter- SigmaPlot 14.5 and statistical analyses were carried out using IBM SPSS mined by a two-step sequential extraction method at days zero, 2, 7 and Statistics 22. 14, following Tatsi et al. (2018). The first extraction was with ultrapure deionised water (Milli-Q water, Elga, 18.2 Ω) to release the water extract- 3. Results able fraction of Ag and followed by the dilute acid extractable fraction of Ag extracted with 0.1 M nitric acid; both at ratios 1:10 (sediment:solution) 3.1. Confirming the exposure – profiles of total Ag in the water column in 15 mL polypropylene centrifuge tubes. The tubes were shaken (IKA Labortechnik KS250) for 1 h, followed by centrifugation for 10 min at The total silver concentration was measured in the water column during 4500 ×g (Harrier 18/80). Each solution was decanted, then acidified the experiment (Fig. 1). In the control rivers, the total Ag concentration re- with 2 % nitric acid and stored in the dark until analysis by ICP-MS. mained below the procedural detection limit (<0.155 μg L−1). The total Ag 4 N. Clark et al. Science of the Total Environment 850 (2022) 157912 from the AgNO3 treatment showed a progressive increase in the water col- umn over the first seven days of dosing and reached a steady plateau of around 13 μg L−1 for the remainder of the experiment. In contrast, the total Ag from the Ag2S NPs treatment continued to rise until day 11, peaking at around 28 μg L−1, and then decreasing to around 12 μg L−1 by the end of the study. However, the grand mean values for the total Ag concentrations measured in the water column (after extraction in nitric acid) were close to the nominal exposure concentration (10 μg L−1), with values of around 12–13 μg L−1 in both silver treatments at days 7 and 14, suggesting that the water was reaching a steady-state concentration (Fig. 1C). This was also confirmed by the daily profiles of total Ag in the water col- umn after dosing (Fig. S1). On the first day of the experiment, the silver dose, regardless of treatment, was quickly removed from the water column. The latter observation is presumably due to adsorption of dissolved Ag and/ or the Ag-containing particles to the glass of the mesocosm and to the sed- iments. The nominal dose was not achieved on the first day with total Ag concentrations peaking at around 5 μg L−1 (Fig. S1A), and then gradually dissipating over the remainder of the 24 h so that only around 1 μg L−1 re- mained by the next dosing period. However, by day 6 the dosing profile for total Ag achieved much steadier concentrations, with the AgNO3 treatment being close to nominal, but the Ag2S NPs treatment showing a spike in con- centration in the first five hours after dosing, but then declining to around 12 μg L−1 over the next 24 h (Fig. S1B). Ultimately, by day 13, the daily profile after dosing was steadier and maintaining close to the nominal con- centrations, suggesting the water was reaching an equilibrium with the other compartments in the mesocosm (Fig. S1C). The presence of silver-containing particles in the water column was measured using spICP-MS (Fig. 1B). There were no silver particles detected in the controls. The silver nitrate reported particulate Ag that was most likely derived from the complexation of Ag with the chloride in the APW to form sparingly soluble AgCl particles. For the Ag2S NPs treatment, parti- cles were detected, but initially at a lower particle number concentration than for the AgNO3 treatment. For both Ag treatments, the particle number concentration increased over time and this observation is consistent with the repeated dosing procedure used to maintain the exposure. The electro- lyte composition of the APW in the rivers was monitored (Table S1), but there were no statistically significant differences between treatments and a consistent water chemistry was observed throughout. 3.2. Total Ag concentrations in the sediment The sediments were analysed for total Ag concentrations following a strong acid digestion (Fig. 2A). The control sediments at the beginning of the experiment had a natural Ag background of around 50 μg Ag kg−1 dry weight (dw) and this showed a decreasing trend over time (not statisti- cally significant), suggesting some redistribution of the background Ag to other compartments in the mesocosm. In contrast to the control, both the Ag treatments showed a progressive increase in total Ag concentrations within the sediment, which were significantly higher than at day 2 (Fig. 2A). One concern for a flowing river where the dosing is arriving at the inflow is that the Ag might tend to accumulate in the sediments at the top of each river (i.e., immediately below the inflow), rather than through- out the sediment along its length. Sediment was collected from the upper, middle and lower parts of each river. Overall, there was no evidence of a position effect with the sediment from all regions of each river showing similar amounts of silver accumulation at each time point (Fig. 2A). Fig. 1. Total measured silver concentrations (μg L−1) and particle number The available fractions of Ag in the sediments were assessed using a se- concentrations (×106 L−1) found in the water column of the rivers from the rial extraction method using water, dilute (mild) acid and strong acid for AgNO3 and Ag2S NP treatments, respectively. Data are mean ± S.E.M., n = 4–12 the total metal (Fig. 2B, C and D). Only a small fraction of the silver mea- water samples. Note, control values were all below the procedural limit of sured, regardless of the exposure, was readily extractable in water, indicat- detection (<0.155 μg L−1). In Panel (A) and (C), there was no statistically significant difference between treatments (Kruskal-Wallis or t-test, p > 0.05) at ing that most of the silver was adsorbed onto or bound to the grains of the day 2, 7 and 14, respectively. Also, in Panel (A) there were no significant time- sediment and not exchangeable into the ‘pore water’ of the sediment layer. dependent changes in the AgNO3 concentrations. In Panel (C) “*” refer to statisti- However, around 28 % in the AgNO3 exposure and 12 % in the Ag2S NP ex- cally significant differences (t-test, p < 0.05) in measured silver concentrations by posure was extractable with dilute acid, suggesting that there was a labile treatment on that day of sampling, relative to sampling on day 2. pool of Ag in the sediment (Fig. 2C and D). 5 N. Clark et al. Science of the Total Environment 850 (2022) 157912 Fig. 2. Total measured silver concentration (μg Ag kg−1) in dry sediment, from the different exposure positions (upper, middle, lower) at days 2, 7 and 14, respectively, in the mesocosm experiment. Panel (A) is the combined total Ag in each fraction. Subsequent panels are broken down into the water extracted, mild acid extracted and concentrated acid fractions, in the control (B), AgNO3 (C) and Ag2S NP (D) treatment, respectively. Data are mean ± S.E.M (n = 4). In panel (A), different letters indicate statistically significant differences amongst the treatments, within the respective day of sampling and exposure position (ANOVA, p < 0.05). Asterisk “*” refers to statistically significant difference in measured silver concentrations by treatment, day of sampling and exposure position, relative to treatment and exposure position in sampling day 2. Absence of letters or symbols means no significant difference amongst the treatments, sampling days and exposure positions investigated. In panels (B), (C) and (D) for each exposure position and day of sampling, “*” indicates a statistically significant difference in measured silver concentration by that extraction method, relative to the other extraction methods (ANOVA, p < 0.05). Absence of symbols means no significant difference amongst the silver extraction methods investigated. 3.3. Total silver concentrations in biota showed a much clearer response, where both organisms showed a statisti- cally significant time-dependent uptake of Ag from the AgNO3 and Ag2S Some of the organisms were very small, or because of matrix effects in NP treatment, respectively. For the planarians, it made no difference if the tissue, the samples were sometimes pooled within stream replicate to the animals were depurated before determining their bodily Ag concentra- get enough biomass for a reliable metal determination. In general, for the tion (data not shown), suggesting that the silver was incorporated into the AgNO3 treatment, all the organisms accumulated total Ag relative to the internal tissues (i.e., not incidental Ag in the gut lumen) and/or irreversibly equivalent control. In most cases, the Ag accumulation from the AgNO3 bound to their epithelial surfaces. For the freshwater snail, P. acuta, the treatment was greater than from the Ag2S NP treatment (Table 1). shell was separated from the soft body in order to understand if there was Amongst the sediment dwelling organisms, L. variegatus, showed a any adsorption of Ag to the former. The shell did show some apparent Ag steady accumulation of Ag from the Ag2S NP exposure over time, which accumulation in both Ag-treatments (Fig. S2). However, the concentration was not different from the AgNO3 treatment by the end of the experiment of Ag in/on the shell was trivial compared to the concentration in the tissue (Table 1). However, the planarians and the chironomid larvae (Table 1) itself, with the latter being an order of magnitude higher than the Ag 6 N. Clark et al. Science of the Total Environment 850 (2022) 157912 Table 1 Total measured silver concentration (μg Ag g−1) in dry invertebrate tissue at sampling day 2, 7 and 14, respectively of the mesocosm experiment with AgNO3 and Ag2S NPs. Test organism Time Treatment (days) Control AgNO3 Ag2S NPs (μg Ag g−1) (μg Ag g−1) (μg Ag g−1) Non-biting midge (Chironomus riparius) 2 0.01 ± 0.002a 12.5 ± 3.72b 3.42 ± 0.90c 7 0.11 ± 0.02a⁎ 33.7 ± 8.11b⁎ 11.7 ± 2.66c⁎ 14 ND ND ND Water flea (Daphnia magna) 2 0.88 ± 0.74a 9.99 ± 1.07b 51.8 ± 10.9b 7 0.05 ± 0.02a ND 40.3 ± 23.9b 14 0.39 ± 0.13a ND 52.0 ± 7.97b Planarian (Girardia tigrina) 2 0.03 ± 0.02a 1.35 ± 0.74b⁎ 0.01 ± 0.001a⁎ 7 0.03 ± 0.01a 16.4 ± 7.49b⁎⁎ 1.17 ± 0.41c⁎⁎ 14 0.08 ± 0.02a 62.0 ± 10.3b⁎⁎⁎ 5.65 ± 2.80c⁎⁎⁎ Oligochaete worm (Lumbriculus variegatus) 2 0.16 ± 0.02a 2.37 ± 0.89b 1.43 ± 0.19b 7 0.21 ± 0.02a 5.24 ± 1.63b 1.91 ± 0.40c 14 0.28 ± 0.02a 4.45 ± 1.97b 4.00 ± 1.14b⁎# Snail (Physa acuta) 2 0.34 ± 0.05a 109 ± 6.72b 28.4 ± 7.24c 7 0.54 ± 0.03a 120 ± 13.7b 53.0 ± 13.7c 14 1.21 ± 0.23a⁎ 59.2 ± 5.82b⁎ 41.3 ± 1.13b Data are mean ± S.E.M (n = 2–4). Different letters indicate significant differences amongst the treatments by organism type, within the respective day of sampling (rows). For each organism type, asterisk (*) symbols refer to statistical difference in measured silver for each treatment type with time. Absence of asterisk (*) symbols refer to no differences in measured silver, for each treatment type, with time. #, for L. variegatus exposure to Ag2S NPs, day 14 is significantly different (p < 0.05) only with respect to day 2. All statistics were determined using t-test, 1-way or 2-way ANOVA with p = 0.05 level of significance. ND - not determined as no animals were found at the sampling time point. deposition on the shell. Importantly, the time course of the changes in the showed a clearer trend wherein the carcass of animals from the AgNO3 soft tissue concentrations of total Ag was different to those in the shell treatment accumulated more Ag than those of the Ag2S NP treatment, (Fig. S2), and not consistent with the notion of a biogenic redistribution and with the least in the unexposed controls (Fig. 3E). The presence of from the tissue to the shell. For the shell, ‘depuration’ had no effect more Ag in the carcass than the skeletal muscle also suggests that the ner- (Fig. S2). However, for the soft tissue, the non-depurated animals showed vous tissue, bone, spleen, kidney (i.e., carcass tissues other than the skeletal more total Ag than the depurated ones by day 14, suggesting that around muscle) were also accumulating silver, with Ag from AgNO3 being more half or more of the apparent accumulation was most likely in the gut con- bioavailable than Ag from Ag2S NPs. tents as rapidly excreted compartment, although some desorption from the mucus on the surface of the foot is also possible. 3.4. Trace metal and electrolyte composition of biota For the organisms that live primarily in the water column (Daphnia and fish) the Ag accumulation was somewhat different to the other inverte- Silver is well-known for its effects on sodium homeostasis in animals brates. For Daphnia (Table 1), the animals survived well in the mesocosm and for its interactions with other trace metals in biota. Consequently, a for the control and Ag2S NP treatments and showed steady clear Ag accu- suite of trace elements and electrolytes were measured in the biota includ- mulation from the Ag2S NP treatment. However, how much of this apparent ing Ca, Mg, Na, K, Cu, Zn and Fe (Tables S2 and S3). There were some time- accumulation in Daphnia was inside the tissue is unclear, as the animals dependent increases in some of the electrolytes in the organisms, perhaps tend to ingest particulates from the water column. These Daphnia, used to reflecting growth and nutritional health in the organisms. However, there feed the fish, were not depurated before the whole body Ag determination. were few treatment-dependent effects and where these occurred, the For the AgNO3 treatments, the Daphnia did not survive in the mesocosms changes were within the physiological scope of the animals (Tables S2 beyond day 2, and so there was no data for days 7 or 14. However, a sepa- and S3). Crucially, overt electrolyte depletion and ionoregulatory stress rate trial was conducted in APW alone to explore AgNO3 toxicity to Daphnia were not observed in any of the organisms. in this medium, and these were used to feed to the fish. The animals did ac- cumulate total Ag from the AgNO3 exposure and depurating the animals for 4. Discussion 24 h made no statistically significant difference to the total body burden. The total measured Ag concentrations (mean ± S.E.M, n = 4) for 4.1. Total Ag in the water column and partitioning to the sediment depurated Daphnia were 15.8 ± 1.39 μg Ag g−1 dry tissue, and for non- depurated Daphnia were 18.9 ± 0.77 μg Ag g−1 dry tissue, respectively, The total silver concentration in the water column in the control with t-test, p < 0.05. mesocosm remained below the procedural detection limit (<0.155 μg The rainbow trout were inevitably exposed to Ag via the water in the L−1, Fig. 1) and with only a trace amount of around 50 μg Ag kg−1 dw or mesocosms, but the fish were also fed contaminated Daphnia to simulate di- less in the sediments (Fig. 2A). Some background of Ag is unavoidable in etary intake as would also occur in a natural river. The animals were ex- aquatic sediments, even from pristine river systems, and the values here posed for a total of six days. The AgNO3 exposure, but generally not Ag2S are in the microgram range similar to previous reports (e.g., sediment NPs, caused total Ag concentrations to increase in the tissues of trout cores from the upper reaches of the River Lot in France, around 310 μg (Fig. 3). The total Ag concentrations in the intestines were variable, with kg−1 dw, (Lanceleur et al., 2011)). For both Ag treatments, the exposure no apparent trend, perhaps suggesting the Daphnia that the fish were fed was progressive with daily dosing of Ag into the water column (Fig. 1) were not a large enough ration or dose to produce a clear time-dependent that inevitably also contaminated the sediments (Fig. 2). For AgNO3, this contamination of the intestinal epithelium (Fig. 3A). Nonetheless, the is expected to be a dynamic process where dissolved Ag species are formed AgNO3 treatment did elevate Ag concentrations in the fish gills (Fig. 3C) in the water column, which then adsorb to surfaces of the particulate min- and livers (Fig. 3B) compared to either the unexposed controls or the erals (e.g., sand) in the sediment, and/or bind to –SH groups and other li- Ag2S NP treatment. The skeletal muscle showed a time-dependent increase gands in any organic matter present, such as humic acids (Adams and in Ag for all treatments (Fig. 3D), but this was at a background concentra- Kramer, 1999; Chen et al., 2012; Herrin et al., 2001). This is partly reflected tion and much lower than that of the other organs. The carcass, which in- in the daily dosing where peak concentration in the water column (Fig. S1; cluded most of the skeletal muscle, bone, nervous tissue and kidney, Fig. 1A) dissipated, presumably through mixing, and by transfer into the 7 N. Clark et al. 8 Fig. 3. Total measured silver concentration (μg Ag g−1) in dry tissue of rainbow trout (Oncorhynchus mykiss) within (A) intestine, (B) liver, (C) gill, (D) muscle and (E) the remaining carcass, at sampling days 0, 2 and 6, in the mesocosm experiment with AgNO3 and Ag2S NPs. Data are mean ± S.E.M (n = 4–5). Different letters indicate statistically significant differences amongst the treatments, within the respective day of sampling (ANOVA, p < 0.05). Absence of letters indicates no statistically significant difference in measured silver concentrations amongst the treatments, within the respective day of sampling. Asterisk “*” refers to statistically significant difference in measured silver concentrations by treatment type, with time (day 2 and day 6). Absence of asterisk symbols indicates no statistically significant differences in measured silver concentrations, for each treatment type, with time (t-test, p > 0.05). Science of the Total Environment 850 (2022) 157912 N. Clark et al. Science of the Total Environment 850 (2022) 157912 sediment compartment of the mesocosm (Fig. 2). Speciation calculations accumulation with the induction of metallothionein inferring metal chela- for AgNO3 in APW (Visual MINTEQ version 3.1 https://vminteq.lwr.kth. tion inside the tissue (Wu et al., 2011). For AgNO3, a recent dietary expo- se/) indicated that the following species were present: 7.19 % of Ag+, sure using contaminated prey items (freshwater snails) showed that the 64.96 % of AgCl (aq), 27.62 % of AgCl− 2− 2 , 0.19 % of AgCl3 , and 0.04 % planarian, G. tigrina, could accumulate total Ag from food (Silva et al., AgSO− 4 . At a nominal total Ag concentration of 10 μg L −1 in the water col- 2022). Crucially, the total Ag accumulation from eating AgNO3-exposed umn, the free Ag+ ion concentration would be <1 μg L−1 and most of the prey items was twice that of Ag2S NP treatment (Silva et al., 2022), indicat- Ag would be readily adsorbed to the sediment or glass of the mesocosm ing that the former was more bioavailable. Similar observations were made each day. Of course, the effect is cumulative, and after the first seven days here with much higher total Ag concentrations in the planarians from the of dosing the total Ag concentration in the water column reached a steady AgNO3 treatment than from the Ag2S NP treatment. Notably, in the present plateau of around 12–14 μg L−1 which remained for the rest of the experi- study, planarians that were depurated had the same Ag concentrations as ment for the AgNO3 (and Ag2S NP) treatment (Fig. 1). animals that were not depurated (data not shown, regardless of type of The speciation calculations indicated that two thirds of the total Ag was Ag exposure), suggesting biological incorporation of the Ag into the tissue, likely present as the AgCl species. Notably, the AgNO3 exposure resulted in as observed for Cd by Wu et al. (2011). Silva et al. (2022) also reported min- Ag-containing particles in the water column (Fig. 1B) and these are likely imal differences in Ag body concentrations between depurated and non- sparingly soluble AgCl particles which are known to form in the presence depurated planarians exposed to different forms of Ag in the food. of chloride (see discussion in Bradford et al., 2009). The APW had a reason- Chironomids are known to accumulate metal from metal salts, depend- able ionic strength (millimolar) and is not unlike physiological saline in that ing on the type of metal, and age/body size of the animals (Krantzberg, regard; where the formation of nanoscale AgCl particles from AgNO3 is also 1989). However, while there have been studies on the toxicity of AgNO3 routinely observed (Besinis et al., 2014; Clark et al., 2019c). Crucially, total and Ag NPs to C. riparius (e.g., Park et al., 2015), the small size of the ani- Ag concentrations in the water or sediment from the AgNO3 compared to mals has hampered reports on Ag accumulation. One food chain study inci- Ag2S NP exposures were similar (Figs. 1 and 2). This suggests that their dentally reported Ag concentrations in adult Chironomus spp., between 38 fate in the abiotic compartments of the mesocosms were the same, and ul- and 45 μg g−1 for AgNO3 and Ag NP exposures respectively, and although timately were likely driven by particle settling from the water column to no unexposed controls were mentioned (Yoo-iam et al., 2014), the Ag accu- the sediment, regardless of the original form of the test substance (Dale mulation reported is of the same magnitude as in the present study et al., 2015; Zhao et al., 2021). (Table 1). Notwithstanding particle formation, dynamic changes in total Ag con- For the freshwater snail, P. acuta, bioaccumulation of total Ag from both centrations in the water and sediment with repeated dosing of the the AgNO3 and Ag2S NP exposures was observed (Table 1), but with less Ag mesocosm are expected, and similar observations were made by Bradford accumulated from the NP treatment. There was also some evidence of et al. (2009) for marine mesocosms with AgNO3 and Ag NPs, where peak depuration (data not shown) suggesting a fraction of the apparent total concentrations in the water eventually plateaued and Ag was adsorbed to Ag was likely in the gut lumen rather than in the tissues. Both AgNO3 and the sediments. In the present study, the partitioning of total Ag in the sedi- Ag NPs are not toxic to P. acuta at the concentrations tested in the APW ment was also measured (Fig. 2A). For both forms of Ag added to the used here, although it is known that the toxicity extent is also life stage mesocosm, the total Ag concentration in the sediment remained within and/or size dependent (Gonçalves et al., 2017). However, there are few de- the sediment, as relatively small amounts of total Ag were being readily ex- tailed reports of Ag accumulation in this species of aquatic snail. At the tracted by water or dilute acid (Fig. 2B, C and D). The sediment was very same concentrations used in the present study, Silva et al. (2022) demon- sandy with negligible organic matter to bind the Ag, and so a small labile strated that P. acuta could accumulate around 40 μg g−1 dw of total Ag pool of total Ag may be expected. from AgNO3 exposure over seven days, compared to 59 μg g−1 dw over 14 days in the present study (Table 1). Crucially, Silva et al. (2022) showed 4.2. Bioaccumulation of total Ag from AgNO3 versus Ag2S NP exposure in the or- that there was more total Ag accumulation from the metal salt (AgNO3) ganisms than from the Ag2S NP treatment in laboratory studies, partly because the snails were not able to excrete Ag (form unknown inside the tissue) after The extractable Ag in the sediments and that in the water column is the AgNO3 exposure (Silva et al., 2022). P. acuta also showed higher accu- likely to be bioavailable to the organisms in the mesocosm. The oligochaete mulation of Ag from AgNO3 compared to Ag2S NP waterborne exposures worm, L. variegatus (Table 1), showed a steady accumulation of Ag from (Silva et al., 2020). The metal salt was therefore more bioaccumulative both the Ag2S NP and AgNO3 exposures, and with no difference between than the nano form by P. acuta, and this was also the case in the mesocosm the Ag treatments by the end of the experiment. Bioaccumulation studies (Table 1). There could be species differences in the handling of Ag amongst on Ag metal salts with L. variegatus are sparse, and often do not include snails. In the freshwater snail, Lymnaea stagnalis, the uptake of total Ag from AgNO3. One study on natural sediments spiked with silver sulfide indicated either AgNO3 or Ag NPs was measured from the water, with uptake rates a trend (no statistical analysis) of increased total Ag accumulation in the greater than the elimination rates, leading to a net bioaccumulation for worms compared to unexposed controls (Hirsch, 1998). The apparent up- both substances (Croteau et al., 2011). Notably, L. stagnalis accumulated take of Ag from Ag2S NPs has not been previously reported for Ag from both food or water exposures (Croteau et al., 2011), and a combi- L. variegatus. However, Ag (form unknown in the tissues) from exposure nation of uptake routes is also likely in the mesocosm study here for P acuta. to AgNO3 or Ag NPs with different coatings, has been reported to be Similar to the snails, AgNO3 exposure caused appreciable Ag accumula- taken up by the oligochaete worm, with the uptake rates being faster than tion in D. magna (Table 1). At day 2 there was a five-fold greater total Ag in the elimination rates, implying that the Ag was bioaccumulative in the or- animals from the Ag2S NP treatment than from AgNO3 (Table 1) and this ganism (Khan et al., 2015), as also observed here (Table 1). Similar obser- could simply reflect that Daphnia are filter feeders and would be removing vations were made by Coleman et al. (2013), who also confirmed the particulates from the water column. There have been a number of previous presence of Ag-containing particles in the oligochaete worm by spICP-MS studies on the bioaccumulation of either dissolved silver or Ag NPs on and spectral imaging approaches. D. magna (Khan et al., 2015; Kim et al., 2016; Ribeiro et al., 2017), but The planarians and the chironomid larvae (Table 1) showed much not on Ag2S particles. Daphnia can accumulate around 10–30 μg g−1 dw higher Ag concentrations than the oligochaete worms, with a clearer of Ag from sublethal exposure via the water to AgNO3 or Ag NPs over time-dependent Ag accumulation from both the AgNO3 and Ag2S NPs treat- 48 h (Ribeiro et al., 2017), a similar magnitude to the present study ment, respectively. There are only a few reports on the bioaccumulation of (Table 1). Ribeiro et al. (2017) also showed more total Ag accumulation metals in planarians, and only one recent study on Ag from the same labo- from the Ag NP treatment than that for AgNO3, with the food being an im- ratory that conducted the mesocosm experiment reported here (Silva et al., portant aspect. Unfortunately, not enough Daphnia remained in the 2022). Nonetheless, for Cd, the planarian, Dugesia japonica showed metal mesocosm to collect for total metal determination at days 7 and 14 for 9 N. Clark et al. Science of the Total Environment 850 (2022) 157912 the AgNO3 treatment, mainly because these are a favourite food of trout. In into account, the animals in the Ag2S NP treatments were probably not the current study, the Ag accumulated by D. magna also represented a die- showing biomagnification relative to the sediments, beyond that already tary hazard to the trout in the mesocosm, because the fish were fed on con- present in the controls. The molecular mechanisms relating to the apparent taminated Daphnia from the appropriate treatments. lack of biomagnification of Ag2S NPs requires further investigation, and in The target organs for freshwater-adapted rainbow trout exposed to part, might be due to the low bioavailability of the ENM at epithelia (Clark AgNO3 via the food (Clark et al., 2019b), or the water (Hogstrand et al., et al., 2019b). In terms of differences in Ag accumulation by habitat or 1996), are relatively well-known and include the gills, gut and liver. In ecological niche, only the fish and Daphnia were mainly in the water col- the mesocosm study, AgNO3, but not Ag2S NPs, caused total Ag concentra- umn, and the fish generally had lower tissue concentrations (Fig. 3) than tions to increase in the tissues (Fig. 3). This is consistent with our previous the benthic invertebrates (Table 1), perhaps reflecting that the fish were findings on in vivo (dietary) exposures in trout where total Ag (form un- denied access to the sediment in the study design. known in the tissue) from the AgNO3 exposure was more bioavailable The Daphnia fed to the fish, and the fish themselves, had negligible con- than that of Ag2S NPs (Clark et al., 2019b). Subsequently, Clark et al. tact with the sediment and so bioconcentration factors relative to the water (2021) identified that dietary exposure to either the metal salt or Ag2S are of more interest here. For the AgNO3 treatment, the Daphnia achieved NPs resulted in Ag-containing particles being present in the internal organs, around 50 μg g−1 dw and the trout maximally around 7.5 μg g−1 dw in as measured by spICP-MS. The total Ag concentrations in the target organs the liver, with the latter consistent with reports in other fish species of around 1 μg g−1 dw or a few hundred ng g−1 dw (Fig. 3) are typical of (Carassius auratus; Ribeiro et al., 2022). At the nominal water concentration previous reports (e.g., AgNO3, Hogstrand et al., 2003); AgNO3 and Ag2S of 10 μg L−1 (or 10 ng mL−1), this represents a ‘transfer factor’ of around NPs, (Clark et al., 2019b; Clark et al., 2021). Interestingly, when silver is x5000 for Daphnia and x750 for the fish. A precise bioconcentration factor presented as a silver sulfide metal salt in the diet of trout, there can be ap- (BCF) would require a demonstration that both water and biota concentra- preciable accumulation of total Ag in the liver, albeit after several months tions were in steady-state, and this is difficult to prove with only a 2-week of exposure (Galvez and Wood, 1999). This at least demonstrates that exposure for the Daphnia, and 6 days for the fish. Nonetheless, the calcula- Ag2S as a chemical substance can be bioavailable to fish, but our previous tion shows, that for AgNO3, there is bioaccumulation of total Ag in the prey findings show this only to a very limited extent when the Ag2S is presented items, but this does not lead to biomagnification at the next trophic level in the nano form (Clark et al., 2019b). This was also the case in the (i.e., in predatory fish). For the Ag2S NP treatment, the Daphnia achieved mesocosm where there were very limited increases of total Ag concentra- around 59 μg g−1 dw, but the trout liver showed <1 μg g−1 dw and was tions in the tissues from Ag2S NP treatment, and with the AgNO3 exposure similar to controls. At the nominal water concentration of 10 μg L−1, this resulting in much more Ag accumulation than that from the nano form represents a ‘transfer factor’ of around x5900 for Daphnia, but with no evi- (Fig. 3). dence of biomagnification of Ag from the Ag2S NP exposure to the next tro- phic level (fish). Silva et al. (2022) reached an identical conclusion for 4.3. Evidence of Ag transfer in the mesocosm and/or biomagnification planarians predating on snails with the same substances used here. A recent study by Xiao et al. (2022) showed the biomagnification factors for dietary A central question for the environmental risk assessment of silver is Ag transfer from Daphnia (prey items) to zebrafish were all <0.2 for AgNO3 whether or not Ag is transferred from the water to sediments and then to and Ag2S NP exposures, indicating negligible biomagnification in the food the biota to cause a bioaccumulation potential, or even a biomagnification web, that was partly explained by lower bioavailability of Ag in all its risk to the food web. It is clear that both AgNO3 and Ag2S NPs in the water forms, possibly due to sulfidation in the intestine of the fish. column resulted in contamination of the sediments due to dissolved metal partitioning into the abiotic compartments and/or particle settling (dis- 4.4. Conclusions and regulatory perspective cussed above). This occurred at a similar magnitude for both the AgNO3 and Ag2S NP exposures, with at least a 100-fold enrichment factor for trans- Additions of either AgNO3 or Ag2S NPs resulted in a progressive in- fer of total Ag from water to sediment (compare Figs. 1 and 2). This high crease in the particle number concentrations in the water column, with level of enrichment is to be expected. For example, dissolved Ag releases the formation of Ag-containing particles from the metal salt. Thus, both ex- into the river Seine resulted in sediment enrichment factors of 100–300 posures ultimately presented a particulate hazard in the water. Irrespective above the natural background level (Ayrault et al., 2010). Similar argu- of the form, the total Ag dosed into the water column rapidly became asso- ments have been presented for the fate and behaviour of Ag NPs (Lead ciated with sediment. However, this sink of total Ag was labile, being ex- et al., 2018) and these processes have been modelled for mesocosms con- tractable in nitric acid, and is therefore likely to be bioavailable. In terms taining Ag2S NPs where nearly all of the microgram amounts of silver of bioaccumulation potential for the AgNO3 treatment, the total Ag concen- dosed into the water would incorporate into the sediments over about ten tration in the benthic invertebrates was three orders of magnitude higher days (Dale et al., 2013). than that in the sediment, and also high for the Daphnia and fish living in Crucially, a fraction of the Ag present in the sediment was water or di- the water column. From a regulatory perspective, this is expected since dis- lute acid extractable (Fig. 2B, C and D), and there were some modest in- solved Ag salts are known to bioaccumulate. Notably, the exposure to Ag2S creases in Ag concentrations in the biota (Table 1). First consider the NPs caused much less total Ag accumulation in the organisms than the AgNO3 treatment, where, with some 200 μg kg−1 of total Ag in the sedi- AgNO3 exposure, with Ag values from Ag2S NP exposures often similar to ment, the values in the biota were around 4, 60, 35, and 60 μg g−1 dw controls. The Ag2S NPs therefore showed little bioaccumulation and were (or mg kg−1 dw) for the oligochaete worms, planarians, chironomid larvae not biomagnified to higher trophic levels. and snails, respectively. On a concentration basis, the organisms in direct In the United States the national Criteria Maximum Concentration contact with sediment therefore contained around ×1000 higher concen- (CMC) for dissolved Ag in freshwater is 3.2 μg L−1 (EPA U, 1980). In trations of silver compared to the control organisms. This apparent Europe a recent review of the regulations for the Water Framework Direc- biomagnification from the sediment to the benthic organisms is typical tive suggests an Annual Average Quality Standard (AA-QS) of around for dissolved Ag and can even be much higher. Lanceleur et al. (2011) re- 0.1 μg L−1 for freshwater and 4.78 μg kg−1 for freshwater sediments ported a biomagnification estimate of 10 × 106 for silver into shellfish. In- (SCHEER, 2021). The latter more stringent regulations are clearly needed door microcosm studies estimated bioconcentration factors from the water since the 10 μg L−1 exposure to Ag as AgNO3 in the mesocosms used here to invertebrate biota for dissolved silver of around ×1000, but estimates caused considerable contamination of the sediment and bioaccumulation were much lower from field measurements (Ratte, 1999). The situation in the biota. While these environmental standards would also protect eco- for Ag2S NPs was different. The measured concentrations for total Ag systems from exposure to Ag2S NPs, one might argue that they would also from the Ag2S NP exposures in the benthic invertebrates were very modest, ‘over regulate’ the nano form as the latter is much less bioaccumulative. or broadly similar to the control animals in most cases (Table 1). Taking this Further research is needed to set any adjustment of the environmental 10 N. Clark et al. Science of the Total Environment 850 (2022) 157912 standards for nano forms of silver, and to define the relative importance of Clark, N.J., Boyle, D., Handy, R.D., 2019b. An assessment of the dietary bioavailability of sil- ver nanomaterials in rainbow trout using an ex vivo gut sac technique. Environ. Sci.: Nano waterborne and dietary sources of Ag2S NPs in real ecosystems. 6, 646–660. Clark, N.J., Clough, R., Boyle, D., Handy, R.D., 2019c. Development of a suitable detection method for silver nanoparticles in fish tissue using single particle ICP-MS. Environ. Sci.: CRediT authorship contribution statement Nano 6, 3388–3400. Clark, N.J., Clough, R., Boyle, D., Handy, R.D., 2021. Quantification of particulate Ag in rain- All of the authors actively contributed to this study, either by directly bow trout organs following dietary exposure to silver nitrate, or two forms of engineered silver nanoparticles. Environ. Sci.: Nano 8, 1642–1653. preparing and characterising the nanomaterials used in the work, con- Coleman, J.G., Kennedy, A.J., Bednar, A.J., Ranville, J.F., Laird, J.G., Harmon, A.R., et al., ducting the freshwater mesocosm experiment at the University of Aveiro 2013. Comparing the effects of nanosilver size and coating variations on bioavailability, led by the team headed by Professor Susana Loureiro, and/or the subse- internalization, and elimination, using Lumbriculus variegatus. Environ. Toxicol. Chem. 32, 2069–2077. quent chemical analysis of biota, sediment or water samples from the exper- Croteau, M.-N., Dybowska, A.D., Luoma, S.N., Valsami-Jones, E., 2011. A novel approach re- iments. All of the authors contributed to the writing of the manuscript, with veals that zinc oxide nanoparticles are bioavailable and toxic after dietary exposures. the editing and data analysis of the submitted paper led by Professor Nanotoxicology 5, 79–90. Dale, A.L., Lowry, G.V., Casman, E.A., 2013. Modeling nanosilver transformations in freshwa- Richard Handy and his team at Plymouth University. ter sediments. Environ. Sci. Technol. 47, 12920–12928. Dale, A.L., Lowry, G.V., Casman, E.A., 2015. Stream dynamics and chemical transformations Data availability control the environmental fate of silver and zinc oxide nanoparticles in a watershed-scale model. Environ. Sci. Technol. 49, 7285–7293. 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