Interactive Effects of Silver Nanoparticles and Phosphorus on

of Biology, Trent University, Peterborough, Ontario, Canada. Environ. Sci. Technol. , 2014, 48 (8), pp 4573–4580. DOI: 10.1021/es405039w. Public...
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Interactive Effects of Silver Nanoparticles and Phosphorus on Phytoplankton Growth in Natural Waters Pranab Das,† Chris D. Metcalfe,‡ and Marguerite A. Xenopoulos*,§ †

Environmental and Life Sciences Graduate Program, Trent University, Peterborough, Ontario, Canada Environmental and Resource Studies, Trent University, Peterborough, Ontario, Canada § Department of Biology, Trent University, Peterborough, Ontario, Canada ‡

S Supporting Information *

ABSTRACT: Increasing amounts of silver nanoparticles (AgNPs) are expected to enter the aquatic ecosystems where their effects on natural phytoplankton communities are poorly understood. We investigated the effects of AgNPs and its interactions with phosphorus (P) supply on the growth kinetics and stoichiometry of natural phytoplankton. Lake water was dosed with AgNPs (carboxy-functionalized capping agent; ∼10-nm particle size; ∼20% Ag w/w) at four different concentrations and five P concentrations and incubated in situ for 3 days. A treatment with ionic silver (AgNO3) was used as a positive control. We found that growth rates, calculated from changes in seston carbon and chlorophyll, responded significantly and interactively (p < 0.0001) to both AgNPs and P. AgNPs reduced the maximum phytoplankton growth rates by 11−85%. In the positive control, no or very little growth was observed. Inhibition of growth rates after exposure to Ag might be related to the reduction in chlorophyll and the inhibition of C and N acquisition rather than P uptake mechanisms. AgNPs, P supply and their interactions also significantly (p < 0.0001) reduced sestonic C:P and N:P ratios and increased C:N, C:Chl and cell-bound Ag stoichiometry. Our results indicate that fate and toxicity of AgNP will vary with phosphorus pollution level in aquatic ecosystems.

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INTRODUCTION Silver nanoparticles (AgNPs) have recently received much attention because of their antimicrobial, antiviral, and antifungal properties, making them the most commonly used nanomaterials in consumer products.1 Given these properties, it is not surprising that bacterioplankton communities and their metabolic rates are negatively affected by AgNPs, at least over shortterm exposures.2,3 AgNPs show potential for release into the aquatic environment4 and therefore, their environmental effects are currently under intense study. There is also potential that the risks posed by AgNP on lower trophic levels5 could cascade through the food web and alter ecosystem function. As the autotrophic microbial component of aquatic systems, phytoplankton can also be negatively affected by AgNP. Data on the toxicity of AgNPs to natural algae are limited to a few small scale laboratory studies.6−9 Past studies are heavily based on single cultured species, with questionable applications to natural ecosystems. Consequently, there is a need to study AgNP effects on natural algal communities that vary in species composition, physiological state, and experience widely diverse ambient physicochemical conditions. From an ecological standpoint, measuring the growth kinetics and ecological stoichiometry of phytoplankton can provide useful information on its physiological state as well as the ecological context of aquatic systems.10 Growth of phytoplank© 2014 American Chemical Society

ton is controlled by several environmental variables including nutrients,11 light12 and contaminants,13,14 which interact in complex ways in the natural environment. As such, an algal cell growing under the stress of nutrient limitation in the natural environment might respond differently to AgNP exposure than a cell growing under more favorable conditions, such as in laboratory-based experiments. The interaction between phosphorus, the most limiting nutrient in freshwater11 and growth is especially important not only for the algal state but for food web interactions. Indeed, P-limited algae increase the susceptibility of their grazers to contaminants,15 thus potentially triggering a food web response. In natural waters, the toxicity of AgNP can be dictated by its interactions with phosphorus. While phosphorus (particularly free phosphate) is a primary determinant of algal production, it can also act as a ligand for toxic Ag ions released from AgNP.16,17 Because of this, high concentrations of phosphate potentially reduces the Ag toxicity to exposed organisms. Ecological stoichiometry is a framework that explicitly links the physiology of an organism to food web interactions and Received: Revised: Accepted: Published: 4573

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ultimately ecosystem function.10 Algal stoichiometry (or C:N:P ratios) provide clues to the nutritional state of algae and how they may affect food webs. One way that AgNPs could affect ratios of algal carbon to nitrogen to phosphorus (C:N:P) is by reducing the C fixation rate due to the disruption of photosynthesis processes,18 and hence, decreasing the C acquisition. In turn, if N- and P-uptake mechanisms are disrupted by AgNP, then this will tend to increase C:N:P ratios. Alternatively, AgNP may enhance uptake rates of N and P, due to increased demand of these nutrients for protein and RNA synthesis, respectively, to repair damaged cells exposed to AgNP, thus also altering C:N:P ratios. If AgNPs alter algal growth kinetics, this can modify the cell’s elemental composition19 and hence, biogeochemical processes in algae and in the aquatic environment.20 These changes in the stoichiometry by AgNPs could alter energy flow and aquatic food web structure and the transfer efficiency of energy to higher trophic levels,10,21 representing a novel approach to understanding the toxicity of nanoparticles. In this study, we examined the effects of AgNPs on the growth and elemental composition of phytoplankton at different levels of P-supply. We showed that under natural conditions, AgNPs exert a strong influence on phytoplankton growth rates and physiological conditions over short (72 h) incubation periods.

a nominal stock concentration of 1.5 mg/mL, and the weight of silver (Ag) as 31%.26 The actual Ag concentration measured by inductively coupled plasma-mass spectrometry (ICP-MS) was 309 ± 9 mg/L,2 corresponding to 20% of the nominal stock concentration. The remainder of the weight of the AgNPs is likely comprised of the capping agent. The characterization of this AgNPs suspension determined by flow field flow fractionation, dynamic light scattering, transmission electron microscopy and ICP-MS have been described in detail elsewhere.2,27 A nominal working suspension of 15 mg/L was prepared by diluting (1:100) the AgNPs stock suspension in Milli-Q water. Based on the 20% proportion of Ag in the material, the nominal AgNPs concentrations of 5, 25, and 50 μg/L correspond to total Ag concentrations of 1, 5, and 10 μg/L. These concentrations were chosen from a trial experiment with algal communities of Chemung Lake using AgNPs concentrations of 1−20 μg-Ag/L, where increasing growth was observed up to ∼12 μg-Ag/L with the addition of P supply as compared to the initial biomass. A working solution of AgNO3 at a concentration of 10 μg/L was prepared by dissolving AgNO3 crystals (Fisher Scientific) in Milli-Q water. The mean concentrations of total Ag measured by ICP-MS after 72 h in the exposure bags ranged from approximately 14−16% of nominal concentrations for both experiments (SI Figure S1A and B), which is less than 20% proportion of the total Ag applied. Hence, the median effective concentrations (EC50) and the lowest observed effective concentrations (LOEC) calculated in the present study were based upon the 20% proportion of the nominal AgNP stock concentration. The concentration of dissolved Ag measured after 72 h by ultrafiltration of the water in the exposure bags was always less than 3.2% of the total Ag (SI Figure S1C and D) as described by Das et al.2 Briefly, for both total and dissolved Ag, aliquots of 2.5 mL of the extract were analyzed by ICP-MS (XSeries II; Thermo Scientific, Germany) to detect 107Ag, with Indium (115In) as the internal standard, as described previously.2 The instrument was optimized by Plasmalab software using 5 μg/ L tuning solution of In and Ag. The procedural limit of detection for 107Ag was 0.01 μg/L. Experimental Design. Algal growth bioassays were repeated twice in Chemung Lake following dilution bioassay procedures modified from Xenopoulos et al.25 Briefly, for each treatment, 1 L of mixed water (60:40) was added to a 1.2-L Whirlpak polythelene bag (Fisher Scientific). Water samples in all Whirlpak bags were then supplied with 800 μg-N/L (as NaNO3) and five different concentrations of P (0 − 100 μg-P/ L as NaH2PO4.H2O) (SI Table S2). P conditions varied depending on the experiment and time. In the first experiment, chlorophyll growth rates were likely to be saturated at 30−40 μg/ L P concentrations. The maximum P addition was reduced in the second experiment. Treatments with AgNPs for both experiments were at nominal concentrations of 0 (Control), 5, 25, and 50 μg/L which corresponds to 0, 1, 5, and 10 μg-Ag/L. There was one treatment with AgNO3 at a nominal concentration equivalent to 3.8 μg-Ag/L. Triplicate incubation bags were prepared for each treatment combination (n = 75 per experiment). After adding AgNPs and AgNO3, all bags were then placed inside floating wire baskets and incubated in the lake at a depth of about 50 cm for 72 h under natural light and ambient environmental condition. After 72 h, water samples in bags were processed for Chlorophyll a (Chl), cellular (seston) carbon, nitrogen, phosphorus and silver. Chl was collected from 250 mL volumes

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MATERIALS AND METHODS Site Selection and Environmental Sample Collection. Lake water was collected in June and July, 2011 from Chemung Lake (44°37′ N, 78°41′W), located in south-central Ontario, Canada. We used a Van Dorn water sampler to collect water samples from 45 to 50 cm below the surface at the deepest point of the lake and then passed all water through a 80 μm mesh Nitex filter in situ to remove metazoan grazers.22,23 All prefiltered water samples were then collected using acid-washed 25 L cubitainer plastic containers and kept in the dark until laboratory processing within 1−2 h of collection. For both time periods, physicochemical parameters were measured following standard field and laboratory procedures that have been described previously.2 The ambient water quality conditions for Chemung Lake were similar for both periods. Waters were slightly alkaline, with low conductivity and the water column was oxygenated at the time of sampling (see Supporting Information (SI) Table S1). Dissolved organic carbon ranged between 5.4 to 6.9 mg C/L and consisted of humic-like organic matter of a terrestrial nature.24 Chlorophyll-a, total dissolved phosphorus, and total dissolved nitrogen were consistent with oligo- to mesotrophic water bodies (SI Table S1). In the laboratory, a portion (40%) of the 80-μm filtered water samples was again filtered through 0.2 μm (142 mm) polycarbonate filters (Millipore). This filtered water was then added in 60:40 (80 μm filtered water:0.2-μm filtered water) ratios to the 80 μm screened whole lake water. Unless otherwise mentioned, phytoplankton in this 60:40 diluted water samples are hereafter defined as a natural algal community. Dilution of the water sample was designed to reduce the competition for resources and the effects of microzooplankton grazing25 and also to give the algal community more scope for growth during the incubation period. This is common practice for dilution bioassays22,25 seeking to measure growth rates. Exposure Materials. An aqueous suspension of carboxyfunctionalized polyacrylate capped AgNPs purchased from ViveNano (Toronto, CA; PN3010L) was used for AgNPs exposures in the present study. For this suspension, the manufacturer reported particles sizes of ∼10 nm diameter and 4574

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of water from each treatment bag onto ∼1 μm GF-C filter (Whatman), extracted in boiling 95% ethanol and quantified by fluorometry.2 Seston carbon and nitrogen were collected from 300 mL volumes of post-incubation water from each treatment bag onto ∼1 μm GF-C filter (Whatman) and estimated using an Elementar Analyzer (Vario EL, Elementer, Hanau, Germany). Initial concentrations of seston C and N were similarly estimated on lake water prior to incubation. Seston phosphorus was collected from 150 mL volumes of water from each treatment bag onto ∼1 μm GF-C filter (Whatman) and estimated using the molybdate-ascorbic acid method.28 Initial seston P was similarly estimated for lake water prior to incubation. For ambient total dissolved phosphorus (TDP) analysis, water samples were filtered through 0.2 μm polycarbonate membrane filters (Millipore) and then measured after potassium persulfate autoclave digestion using the molybdate-ascorbic acid method.28 Chlorophyll- and carbon-specific net P fluxes were calculated as described by Frost and Xenopoulos.29 Briefly, chlorophyllspecific net P flux (μg P. μg chlorophyll−1. d−1) was calculated as the difference between initial and final particulate P concentrations divided by final chlorophyll concentrations and the length of incubation (3 days). Similarly, carbon-specific net P flux (μg P. μg C1−. d−1) was calculated as the difference between initial and final particulate P concentrations divided by final seston C concentrations and the length of incubation (3 days). Cell-bound (or cellular) Ag was collected from 100 mL volumes of water from each treatment bag by filtration with ∼1-μm polycarbonate membrane filter (Millipore). The filter was placed immediately into a 15 mL screw cap conical base tube (Sarstedt), and then nitric acid (HNO3) was added to a final concentration of 2% before analysis. Aliquots of 2.5 mL of the extract were analyzed by ICP-MS as described above. For selected treatments and using only one sample, phytoplankton composition and cell numbers were measured by enumeration in Utermöhl chambers with an inverted microscope (400×) following methods previously described by Xenopoulos and Frost (2003).23 The total biovolume of individual phytoplankton genera was estimated from cell volume measurements and the geometric dimensions of an average of 15−20 cells of an individual genus using the formula that best described the shape, as described by Hillebrand et al. (1999).30 Statistical Analysis. Natural algal response variables in the dilution bioassay exposure experiments were log(x+1) transformed to better meet the assumptions of normality. Growth rates of C and Chl were calculated assuming exponential growth. Nonlinear regression analysis was then used to fit growth rates with P gradients following Monod growth kinetics:31

content), using observations after 72 h exposure to AgNPs and AgNO3 at different P conditions. The LOEC were also calculated as the square root of the lowest concentration observed with less than 50% negative effects multiplied by the lower bound of the EC50 95% confidence interval.

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RESULTS Variations in Total Ag, dissolved Ag, cell-bound Ag, and Chl:Ag, C:Ag, N:Ag, P:Ag ratios. Total Ag, dissolved Ag and cell-bound Ag measured after 72 h in the exposure bags showed a temporarily consistent concentration dependent increase across AgNP gradients, but the pattern remained relatively unchanged across phosphorus gradients (Figure 1; see

Figure 1. Changes in the cell-bound (cellular) Ag, sestonic Chl:Ag, C:Ag, N:Ag, and P:Ag ratios as a function of total dissolved phosphorus (TDP) at 1(filled circle), 5 (open circle) and 10 (filled down triangle) μg-Ag/L in AgNPs and 5 (open up triangle) μg/L AgNO3 treatments for experiments in June (A, C, E, G, I) and July (B, D, F, H, J) of 2011 in Chemung Lake of southern Ontario, Canada. Note that x- and y-axis scales vary in panels.

μ = (μmax × x)/(Ks + x)

where, μ is the specific growth rate (d−1), x is the P concentration (μg/L), μmax is the maximum specific growth rate (d−1) at the saturating phosphorus concentration, and Ks is the half saturation constant. Estimates of Ks and μmax, along with their respective 95% confidence intervals, were determined using the nonlinear fitting component of SPSS (SPSS Inc., Illinois). Effects of AgNPs, AgNO3 and phosphorus treatments on growth rates, C:P, C:N, N:P, C:Chl, C:Ag, P:Ag, N:Ag, Chl:Ag ratios and cellbound Ag were evaluated with two-way analysis of variance. Tukey’s HSD for honest significant difference posthoc test was used to evaluate whether differences between any two pairs of means are significant. SoftTOX Ver 1.1 (WindowChem) software was used to calculate by Probit transformation the EC50 for negative effects on Chemung Lake algal biomass (Chl

also SI Figure S1). The interaction between sestonic Ag (cellbound) and total dissolved phosphorus was statistically significant (p < 0.001; SI Table S3). In both experiments, P enrichment and AgNPs treatments also had significant effects on Chl:Ag, C:Ag, N:Ag and P:Ag ratios (p < 0.0001 for all ratios; Figure 1). At ambient P conditions, these ratios were the lowest in all Ag treatments and then gradually increased in a concentration dependent manner up to a certain level of P supplied (15−25 μg/L P supply) indicating their increasing growth, followed by a slight decrease in the highest P supply in 4575

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over the 72 h incubation period. The highest C:P ratios were observed in the nonexposure groups (no Ag). AgNPs reduced C:P ratios regardless of P-supply, and AgNO3 treatments further reduced C:P ratios (Figure 3A and SI Figure S3A). C:N ratios increased with P-supply and with exposure to increasing levels of AgNPs. At the highest level of P addition, the C:N ratio was the lowest in the nonexposed group (no Ag) and the highest in the high AgNPs or AgNO3 treatments (Figure 3B and SI Figure S3B). In contrast, N:P ratios decreased with Psupply, but N:P ratios were variable at the highest level of P addition (Figure 3D and SI Figure S3D), potentially highlighting differential responses of AgNP treatment on N and P dynamics. Ag treatments and P supply also had a significant influence on C:Chl ratios (p < 0.0001; Figure 3C and SI Figure S3C). We found a dose dependent increase in C:Chl ratios, whereas addition of P decreased C:Chl ratios in each treatment at each P level after 72 h of incubation. At all enrichment of P levels, C:Chl ratios in all Ag treatments were always significantly (p < 0.0001) higher than the nonexposed groups. In addition, AgNO3 showed the highest C:Chl ratios relative to other AgNPs treatments and the lowest was observed in the control group (no Ag) at each P level. Seston C:P ratios were negatively correlated with chlorophyll and seston C growth rates, and the strength of this relationship (R2) varied with AgNPs and AgNO3 treatments (Figure 4 and SI Figure S4). In addition, exposures of AgNPs shifted C:P relationships with algal growth rates toward the x-axis for both experiments. Algae grew more slowly in treatments with Ag exposures. Surprisingly these exposed algae always had lower C:P ratios than algae that grew at the same rate in control treatment (no Ag). Influence on Chlorophyll- and Carbon-Specific P Flux. In both experiments, chlorophyll- and carbon-specific net P fluxes showed a significant increase (p < 0.05) between the control and Ag treatments (Figure 5 and SI Figure S5). At higher P levels in both experiments, chlorophyll-specific net P flux was higher either in the highest AgNPs concentration or in the AgNO3 treatment compared to the control. At low levels of P addition, carbon-specific P flux was either not affected or slightly higher in Ag treatments compared to the control. At the highest level of P addition, carbon-specific P flux was the highest in the control and the lowest either in the highest AgNPs or the AgNO3 treatment. Influence on Phytoplankton Biovolume. At ambient P condition, the biovolume of phytoplankton was reduced by 70− 90% after exposure to the highest concentration of AgNPs (i.e., 10 μg-Ag/L) relative to the control after 72 h of incubation (SI, Table S8 and Figure S6). The biovolume was also decreased (22−94%) in other treatments with 10 μg/L AgNPs, but higher concentrations of P mitigated these responses (SI, Table S8 and Figure S6). Regardless, there was evidence of toxicity in all treatments with AgNPs for bacillariophytes, chlorophytes and chrysophytes. In contrast, cyanobacteria were almost eliminated by exposure to AgNPs, even in treatments with the addition of P.

most cases (Figure 1). In addition, Chl:Ag and nutrient:Ag ratios in the AgNO3 treatment were similar to the ratios in the highest AgNPs treatment (10 μg-Ag/L) at each P level. As the background levels of Ag in Chemung Lake waters at ambient conditions were below detection limit by ICP-MS analysis, these ratios (Chl:Ag, C:Ag, N:Ag and P:Ag) could not be determined in the control groups which did not receive any silver. Influence on Algal Growth. Chlorophyll and seston C growth rates were affected significantly by exposure to AgNPs and AgNO3, and also by their interactions with P in natural waters (p < 0.0001, SI Table S3). There was a clear concentration dependent response for chlorophyll and seston C growth after 72 h (Figure 2 and SI Figure S2). In general, no (July) or very little

Figure 2. The Monod phosphorus-dependent (A) chlorophyll, and (B) seston carbon growth curves as a function of total dissolved phosphorus (TDP) at 0(filled circle), 1 (open circle), 5 (open up triangle) and 10 (filled down triangle) μg-Ag/L in AgNPs and 5 (open square) μg/L AgNO3 treatments for experiments of June, 2011 in Chemung Lake of southern Ontario, Canada. Note that y-axis scales vary in panels.

(June) growth in chlorophyll and seston C was observed in treatments with AgNO3 (Figure 2 and SI Figure S2). The small difference in the growth response to AgNO3 exposed algae between June and July could be related to slight differences in the lake physicochemical parameters (e.g., water temperature; SI Table S1) or different algal communities. Chlorophyll and seston C growth were positively correlated (R2 = 0.81, 0.69, respectively; p < 0.0001). Results from the Tukey’s HSD posthoc multiple comparisons also showed differential effects of AgNPs gradients and AgNO3 on chlorophyll and seston C growth at the five P levels (SI Table S4). There was a concentration dependent decrease in the μmax of Chl and C growth kinetics after 72 h (SI Table S5). AgNPs reduced μmax of chlorophyll by 16−88%, while AgNO3 reduced this parameter by 90−99%, compared to the control. For seston C, AgNPs caused a reduction in μmax by 15−80%, whereas AgNO3 caused a 92−100% reduction compared to the control. On the contrary, Ks were significantly increased (p < 0.05) in Ag exposures in comparison to the control (SI Table S5). Influence on C:P, C:Chl, C:N, and N:P ratios. AgNPs and P additions significantly (p < 0.05 or p < 0.0001) reduced algal C:P ratios in both experiments (SI Table S6). Statistical analysis using two-way ANOVA revealed that AgNPs, P supply and their interactions had a significant influence on sestonic C:P ratios

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DISCUSSION Growth of a natural community of phytoplankton was significantly inhibited by exposure to AgNPs capped with carboxy-functionalized polyacrylate in treatments with P additions. We found reductions in growth rates and μmax with AgNPs and AgNO3, and a consistent significant interaction between P-supply and AgNP. The addition of P increased algal growth and may have also reduced toxicity. This interaction 4576

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Figure 3. Changes in sestonic molar (A) C:P, (B) C:N, (C) C:Chl, and (D) N:P ratios as a function of total dissolved phosphorus (TDP) after 72 h post incubation at 0(filled rectangle), 1 (open rectangle), 5 (diagonal rectangle), and 10 (dotted rectangle) μg-Ag/L in AgNPs and 5 (bricks rectangle) μg/L AgNO3 treatments for experiments in June, 2011 in Chemung Lake. Within each total dissolved phosphorus (TDP) level treatment identified with different lowercase letters, responses were significantly different from each other at p < 0.05, whereas, no letters indicate that there was no significant difference. Note that y axis scales vary in panels.

Figure 4. The regression analysis of (A) chlorophyll and (B) seston carbon growth rates as a function of sestonic molar C:P ratios at 0(filled circle), 1 (open circle), 5 (open up triangle), and 10 (filled down triangle) μg-Ag/L in AgNPs and 5 (open square) μg/L AgNO3 treatments for experiments of June, 2011 in Chemung Lake of southern Ontario, Canada. Note that y-axis scales vary in panels.

Figure 5. Chlorophyll-specific (A) and carbon-specific (B) P flux into the phytoplankton as a function of total dissolved phosphorus (TDP) after 72 h post incubation at 0(filled rectangle), 1 (open rectangle), 5 (diagonal rectangle), and 10 (dotted rectangle) μg-Ag/L in AgNPs and 5 (bricks rectangle) μg/L AgNO3 treatments for experiments in June, 2011 in Chemung Lake of southern Ontario, Canada. Within each total dissolved phosphorus (TDP) level treatment identified with different lowercase letters, responses were significantly different from each other at p < 0.05, whereas, no letters indicate that there was no significant difference. Note that y axis scales vary in panels.

between P and AgNP highlights the importance of considering nutrient pollution levels in studies of the fate and toxicity of AgNP. We also showed for the first time that AgNP exposure changes the phytoplankton elemental composition (C:N:P ratios), which may have implications for food web dynamics and biogeochemistry in aquatic systems. Our results indicate that exposure to AgNPs increases biomass-specific net P flux. This resulted in decreases in C:P and N:P ratios, which when coupled to increases in C:Chl, C:N and cell-bound Ag stoichiometry after exposure to AgNPs, indicate that other mechanisms (e.g., N

uptake, photosynthesis, respiration, etc.) rather than P uptake might be responsible for the reduction in algal growth. Finally, in the treatment with AgNO3, Chl- and C-growth were either fully 4577

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exposed phytoplankton were still able to acquire P from the water at rates similar to or higher than those observed in the nonexposed group. The higher chlorophyll- and carbon-specific P fluxes into AgNPs-exposed phytoplankton might be due to an increased demand for P to repair nanoparticle damaged cells or might reflect lower cellular chlorophyll concentrations. The calculated 72 h LOEC and EC50 values for inhibition of algal biomass increased with P supply (Table 1). This means that

inhibited or greatly reduced, thus indicating that the toxicological effects of Ag+ on natural algal communities are different from those of AgNPs. We found a substantial decrease in μmax coupled with a rise in Ks values with increasing AgNPs concentrations, indicating slower growth of phytoplankton compared to the control groups. A strong decrease in chlorophyll content was also observed by Oukarroum et al.7 in algae exposed to AgNPs. The significant reduction in chlorophyll observed in the present study potentially affected carbon acquisition, photosynthesis and respiration processes. Although not yet completely understood, several mechanisms have been proposed to explain the toxicity of nanoparticles on microalgae.6,7,32,33 One mechanism of algal toxicity may be linked with the release of silver ion (Ag+) from the nanoparticles,34−37 usually regulated by the pH and the ionic strength of the aquatic system.38,39 The physicochemical conditions of our exposure water (e.g., pH) did not change over the course of this short experiment thus not contributing to a large release of Ag+. Because the amount of Ag+ that was likely released in the present short-term exposure was small (0.01− 0.19 μg/L; calculated as 2.8% of total Ag2,3), it is unlikely that Ag + was an important mechanism of toxicity in this study. Furthermore, possible complexation of Ag+ with the organic material in the lake water35 (e.g., humic-like DOC of 5.4−6.9 mg C/L measured in this study) or with the added phosphate could have reduced the toxicity to algae.16,17 For example, release of Ag+ (equivalent to 0.1−1.7 nM) from AgNPs could precipitate as silver phosphate after potential complexation with the dissociated phosphate ion (equivalent to 0.06−3.24 μM PO43‑) from the added NaH2PO4 in the present study . Nevertheless, algal growth was completely inhibited or greatly reduced in the AgNO3 (5 μg/L) treatments, an indication that Ag+ is more toxic than AgNPs. Toxicity to phytoplankton could also stem from the accumulation of nanoparticles inside the algal cells,40,41 attachment onto the algal cell wall,42,43 or physical shading by the nanoparticles that reduces the light needed for algal growth.44 These toxicity mechanisms may in turn be influenced by particle size, formation of reactive oxygen species (ROS) and surface coating.7,34,42 For example, Yang et al.34 and Oukarroum et al.7 reported that the toxicity of AgNPs occurred through the effects of ROS. This interaction with ROS might have played a role in the present study, as we incubated our treatments under natural light conditions,35,35 although some ROS have not been found to be toxic to algae at ambient concentrations.43 As for cellular accumulation of AgNP as a potential factor to explain toxicity,32,33,41−43 we were able to measure total silver in the algal fraction collected after exposures, although this silver may be inside the cell or adsorbed onto the cell wall. In any case, there appears to be some potential for biochemical interactions within the algal cells that may cause the growth inhibition observed.40,41 Although cell-bound Ag increases with AgNP added, it does not increase with the P-supply indicating a nonlinear increase in Ag per unit algal biomass and highlighting an interesting interaction between cell growth and silver toxicity that needs to be considered in future studies. When measured as changes in seston cellular-P, P-uptake was increased in AgNPs- and AgNO3-stressed algae across P gradients in a manner that eventually decreased the C:P and N:P ratios. When adjusted for the algal biomass, Ag treatments significantly increased biomass-specific (chlorophyll- and carbon-specific) net P fluxes into phytoplankton, at all P-levels. An increase in net P-fluxes in the present study indicates that AgNPs-

Table 1. Median Effective Concentrations (EC50) and Lowest-Observed-Effective Concentrations (LOEC) for Production of Natural Algae from Chemung Lake After 72 h Exposure to AgNP P condition (μg/L)

EC50 (95% CI) μg-Ag/L

LOEC μg-Ag/L

8 (ambient) 18 33 58 108

2.21 (1.9−3) 3.78 (3.67−4.83) 5.12 (4.9−5.37) 6.83 (6.66−7.63) 6.08 (5.1−6.27)

0.92 1.43 1.81 2.28 2.4

toxicity to algae not only depends on the AgNP dose but also on the P-content of algal cells and that the nutritional state of an organism should be incorporated in future risk assessments of nanoparticles.46 Algal LOEC and EC50 values at different P conditions after 72 h exposure to AgNPs were calculated as 0.92−2.4 and 2.21−6.83 μg-Ag/L, respectively, which are 15−90 fold lower than the EC50 that we previously determined for natural aquatic bacterioplankton exposed to the same AgNPs.2,3 These toxicity thresholds indicate that algae may be more sensitive to the toxic effects of AgNPs than natural bacterial communities. Future studies should consider the interaction and competition for resources between algae and bacteria in the presence of nanoparticles. Despite enrichment of P, total cellular C and N were significantly lower in treatments with AgNPs compared to the control (see SI Figure S7). The reduction in particulate C indicates that exposure to AgNPs might affect natural algal photosynthesis and respiration.18,47 Thus, P-uptake may not be affected by AgNPs but photosynthesis, respiration and N-uptake might be. Total net N flux was indeed lower in the AgNPs treatments compared to the control (see SI Figure S7) resulting in an increased and decreased C:N and N:P ratios.48 It is perhaps not surprising that N-uptake mechanisms in phytoplankton might be more sensitive to AgNP than P-uptake mechanisms, given the recent finding that nitrogen cycling by bacteria respond negatively to AgNP.49 The lower uptake rates of N could also explain, in part, the AgNP-reductions in chlorophyll, a pigment containing nitrogen atoms. The C:Chl ratios in the present study decreased with enriched P levels in control groups over the incubation time, consistent with past studies.50,51 Our data revealed that, despite additional P supply and decreasing C:Chl ratios, exposure to AgNPs is responsible for the high C:Chl ratio in each P level over 72 h incubation period. One explanation is that AgNPs disproportionally reduced the Chl concentration compared to the cellular C, causing a substantial increase (17−126%) in the C:Chl ratios at all AgNPs exposures relative to the control groups over time. This is an indication that photosynthesis (requiring chlorophyll, producing glucose) might be especially sensitive to AgNP. This is complicated by algal respiration mechanisms, which might also be affected as a result of the impairment of photosynthesis and N assimilation52 after exposures to AgNPs. 4578

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Environmental Science & Technology We found that there was a negative correlation between concentrations of cell-bound Ag and phytoplankton growth rates (p < 0.0001). Algal biomass or elemental composition to Ag ratios (Chl:Ag, C:Ag, N:Ag, P:Ag) in both experiments decreased with Ag concentrations in a concentration dependent manner. Together, the increase in cell-bound Ag, coupled with a decrease in Chl, C, N, P to Ag ratios, decrease in biovolume and algal growth inhibition revealed that exposure to AgNPs is the main driver for inhibition of chlorophyll and sestonic C growth. However, algal sensitivity to AgNPs may vary due to many factors such as water chemistry and temperature, and the species composition of the algal community. For example, in the present study, cyanobacteria species seemed to be more sensitive to AgNPs compared to other algal genera, which indicates that there could be algal community composition shifts in natural waters impacted by AgNPs. Generally, a lower C:P ratio is considered improved food quality for herbivores, which enhances their growth.10 Since we found a reduction in C:P and N:P ratios (but not C:N ratios) in algae with exposure to AgNP, this might have an indirect beneficial effect for grazers through improved stoichiometric food quality.53 On the other hand, in laboratory exposures with controlled food quality (replete conditions), grazers such as Daphnia magna have been found to be especially sensitive to AgNP; more so than algae and bacteria.54 For grazers in the natural environment, it remains to be determined whether the reduction in algal quantity that we see here might offset the simultaneous benefits of improved food quality. In general, there are few studies that have analyzed the effects of AgNPs on single or multiple species of algae, and, to our knowledge, the present study is the first performed with algal communities under natural environmental condition. We found significant responses in both algal growth and elemental composition to AgNPs exposure in the presence of enriched nutrients using algal dilution bioassays over 72 h. Although we did not quantify the specific aspects of the kinetics of cellular damage and repair, our results indicate that the algae would be able to compensate for AgNPs induced damage (by enhancing Puptake), yet at some metabolic cost that was evident from reduced growth and reduced N-uptake compared to the nonexposed groups. The negative effect of AgNPs on natural algal growth might be offset by the improved nutritional quality (C:P and N:P ratios). These interactive effects could be more pronounced under field exposures for a longer period of time. Hence, future in situ studies are needed to determine the extent and duration of AgNPs exposure to establish how this translates into altered growth rates and elemental composition and their potential ecological consequences in the food web.

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ACKNOWLEDGMENTS

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REFERENCES

We thank A. Scott, F. Mustafa, and E. Hoque for assistance in the lab and field. This research was supported by grants from Environment Canada and the Natural Sciences and Engineering Research Council (NSERC) of Canada. P. Das acknowledges support from the Ontario Graduate Scholarship and a Fish and Wildlife Development Fund Student Research Award from the Saskatchewan Ministry of Environment.

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S Supporting Information *

Tables S1−S8, Figures S1−S7 are available as Supporting Information. This material is available free of charge via the Internet at http://pubs.acs.org.

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Corresponding Author

*Phone: 705.748.1011 ×7699; fax: 705.748.1139; e-mail: [email protected]. Notes

The authors declare no competing financial interest. 4579

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