Uptake and impact of silver nanoparticles on the growth of an estuarine dinoflagellate, Prorocentrum minimum

NanoImpact 15 (2019) 100181

Contents lists available at ScienceDirect

NanoImpact journal homepage: www.elsevier.com/locate/nanoimpact

Uptake and impact of silver nanoparticles on the growth of an estuarine dinoflagellate, Prorocentrum minimum Shelby V. Butza,b, James L. Pinckneyb, Simon C. Aptec, Jamie R. Leada,

T

⁎

a

Center for Environmental Nanoscience and Risk, Department of Environmental Health Sciences, Arnold School of and Public Health, University of South Carolina, 921 Assembly Street, Suite 401, Columbia, SC 29208, United States of America b School of the Earth, Ocean and Environment, Marine Science, University of South Carolina, 701 Sumter Street, EWS 617, Columbia, SC 29208, United States of America c Center for Environmental Contaminants Research, CSIRO Land and Water, Private Mail Bag 7, Bangor, NSW 2234, Australia

A R T I C LE I N FO

A B S T R A C T

Editor: Bernd Nowack

Silver nanoparticles (AgNPs) are widely used and have been shown to have relatively high inherent toxicity to a range of living organisms including fish, annelids, daphnia and algae. AgNPs capped with either citrate (citAgNPs) or polyvinylpyrrolidone (PVP; PVP-AgNPs) were exposed to the estuarine dinoflagellate Prorocentrum minimum at a range of concentrations (1–1000 μg L−1). Transformations of the AgNPs (e.g. dissolution), along with their uptake and biological effects were measured. DLS and TEM measurements indicated a particle diameter of 15 nm for both particle types. A multi-method approach showed that both particle types underwent rapid and concentration-dependent dissolution and agglomeration. The effects were most apparent for the citAgNPs. With a larger total Ag addition, a larger percentage and mass of Ag was operationally defined as internalized in or strongly bound to cells, regardless of capping agent. Exposure to AgNP concentrations of 50–100 μg L−1 significantly decreased cellular growth after 72 h resulting in EC50 values of 60.5 μg L−1 (citAgNP), 67.1 μg L−1 (PVP-AgNP), and 30.3 μg L−1 (AgNO3(aq) control). The calculated no observed effect concentrations (NOEC) were 24.0, 19.0, 0.2 μg L−1 for cit-AgNP, PVP-AgNP and AgNO3 respectively. Using a predicted exposure concentration (PEC) of 0.116 ng L−1 for nano-Ag in the environment and predicted no effect concentrations (PNEC) of 0.02, 0.02, 0.002 μg L−1 (respectively), PEC/PNEC risk assessment ratios of < 1 were determined for all treatments. This indicated no immediate concern for AgNPs toxicity to P. minimum.

Keywords: Silver Nanoparticles Uptake Bioaccumulation Toxicity Phytoplankton Algae Estuarine

1. Introduction Nanoparticles (NPs) have become a central component in a wide range of commercial and consumer products having electronic, biomedical, and pharmaceutical drug delivery applications, as well as, cosmetics and personal care products (Piccinno et al., 2012; Nowack and Bucheli, 2007). AgNPs are used frequently in consumer products (Nanotechnologies, T. P. o. E., n.d.) because of their strong antimicrobial properties (Sondi and Salopek-Sondi, 2004). Since AgNPs are known to be toxic and to be present in the environment (and also potentially to have novel properties), they have a potential environmental risk, which might be significantly different to that of the free silver ion. Microalgae have been an organism of interest and potential mechanisms of AgNP toxicity are likely to be largely associated with NP dissolution and the interaction of dissolved Ag with the cells (Ivask et al., 2014; Zhang et al., 2016; Turner et al., 2012). Although there is an

increasing amount of literature on algal-NP interactions, the wide range of species, NP types and media conditions make complete rationalization challenging. An insufficient number of studies capture a full range of data from synthesis/procurement to toxicity mechanisms to fully explain disparate data. Many studies have been performed at high and unrealistic concentrations, although increasing numbers of studies are more realistic (Jia et al., 2019) and concentration dependence of NP behavior is well established (Baalousha et al., 2016; Hadioui et al., 2013). For AgNPs, which are the most studied, NP uptake has both been ruled out with only cell wall sorption found and subsequent uptake of the ion (Yue et al., 2017) and internalization has also been shown to occur (Sekine et al., 2017). Toxicity is often explained as due to ion alone (Ivask et al., 2016), although there is evidence for a smaller nanospecific role for NPs (Kleiven et al., 2019). Therefore, dissolution was not the only process contributing to algal toxicity and therefore the role of NP-cell interactions cannot be discounted (Zhang et al., 2016; Leclerc

⁎ Corresponding author at: Department of Environmental Health Sciences, University of South Carolina, Arnold School of Public Health Research Center, 921 Assembly St., Columbia, SC, 29208, United States of America. E-mail address: [email protected] (J.R. Lead).

https://doi.org/10.1016/j.impact.2019.100181 Received 19 March 2019; Received in revised form 11 July 2019; Accepted 29 July 2019 Available online 30 July 2019 2452-0748/ © 2019 Published by Elsevier B.V.

NanoImpact 15 (2019) 100181

S.V. Butz, et al.

and Wilkinson, 2014; Sorensen and Baun, 2015). Research coupling algae and AgNPs has primarily focused on toxicity of AgNPs to freshwater algal species. Our literature search on web of science, for the last 5 years, shows that there are a relatively small number of papers examining Ag NP and algae interactions. Of those papers, only a small fraction (10–20%) investigate estuarine/seawater algae (Ivask et al., 2014; Sorensen and Baun, 2015; Ribeiro et al., 2014; Navarro et al., 2015; Venkatesan et al., 2016; Shankar et al., 2016; Srikar et al., 2016; Lodeiro et al., 2017; Huang et al., 2016; Sendra et al., 2017) making this an important focus area. In such cases, the detection of Ag is challenging because of the high salt background and its effect on inductively coupled plasma mass spectrometry (ICP-MS) analysis. Dinoflagellates are one the largest groups of marine eukaryotes (Gomez, 2005), mixotrophic, meaning they can obtain energy through photosynthesis or phagocytosis and are more mobile than diatoms because they have flagella which assist with locomotion. Dinoflagellates are also of special interest due to their ability to cause harmful algae blooms such as red tides (Heil et al., 2005). The dinoflagellate Prorocentrum minimum is a mixotrophic, bloom forming estuarine species that is widely distributed geographically and is potentially harmful to humans via shellfish poisoning (Heil et al., 2005). Examining AgNP toxicity on the primary production level is important to ensure the base of estuarine food webs are not contaminated or compromised by toxicants that may drastically alter phytoplankton community cell density, function, or structure. The objective of this study was to synthesize and accurately characterize a range of AgNPs and identify the effects of NP concentration and surface coating on transformations, uptake, and toxicity to P. minimum.

Table 1 NP characterization results. Characteristics of cit-AgNPs and PVP-AgNPs in stock suspension and in exposure media after 72 h. TEM values and peak sizes are reported as mean ± 1 standard deviation. Pdi is the polydispersity index measurement. Note: Results of cit-AgNP are represented by the 72 h time point because they were immediately undetectable upon addition to saltwater media. ND indicates a measurement that was not determined due to changes in exposure. Particle

Media

cit-AgNP cit-AgNP PVP-AgNP PVP-AgNP PVP-AgNP PVP-AgNP

Water L1 -Si Water L1 -Si L1 -Si L1 -Si

Batch conc. (ppm) 8.6 8.6 8.6 8.6 8.6 8.6

Time in media (hrs) 0 72 0 24 48 72

TEM size (nm)

14.0 ± 20.7 ± 14.8 ± ND ND 17.4 ±

1.7 6.5 5.7

7.2

pdi

0.2 ND 0.3 0.6 0.5 0.4

Hydrodynamic diameter (nm)

11.2 ND 17.6 49.6 20.6 16.4

(DLS) and absorbance by UV–vis were measured every 24 h, TEM analysis was conducted after 72 h in media without algae. At 0, 24, 48 and 72 h AgNPs were ultrafiltered (EMD Millipore Amicon™ Ultra-4 Centrifugal Filter Unit, 3 kDa membrane) via centrifugation (400 RPM, 20 °C, 30 min) then preconcentrated using an anion exchange resin to remove the seawater matrix (Butz et al. 2019 in prep) (102.9 ± 1.4% recovery) and measured via inductively coupled plasma mass spectrometry (ICP-MS) (Perkin Elmer NexION 350D) for dissolved Ag. An internal standard was added to all samples and standards measured via ICP-OES and -MS. A nitric acid (1%) blank and mid range standard (10 μg L−1) was measured every 10 samples as a quality assurance and quality check (QA/QC) to account for blank and drift correction on the ICP-OES and -MS.

2. Methods 2.1. NP synthesis

2.3. Biological media and test organism AgNPs were synthesized using a modified chemical reduction of silver salts in sodium citrate (Henglein and Giersig, 1999; Jana et al., 2001; Doty et al., 2005; Cumberland, 2013; Römer et al., 2011). To remove excess reactant species and prevent re-equilibration, cit-AgNPs were washed with a citrate solution using a diafiltration approach to prevent drying and aggregation (3 kDa cellulose membrane ultrafiltration disc, EMD Millipore). Half was retained unchanged and half was recapped with a 500 mg L−1 PVP (10k) solution (Tejamaya et al., 2012). PVP was added and immediately and the solution stirred vigorously for 10 min, stirring was then decreased for an additional 60 min. PVP was chosen as a capping agent to provide different surface properties than citrate (Römer et al., 2011; Tejamaya et al., 2012). Citrate creates a charge stabilized NP which is more likely to be transformed by dissolution, aggregation or other mechanisms. PVP creates a sterically stable NP which is more persistent in complex media. Both solutions were stored at 4 °C in dark. All measurements and calculations for AgNP synthesis, exposures, data interpretation were based on the mass of Ag, not of AgNO3.

A standard culturing media, denoted L1-Si was prepared according to the National Center for Marine Algae and Microbiota (NCMA) instructions. The complete recipe can be found in the supporting information (Tables S1-S3). Natural seawater (25 psu) was collected from North Inlet Estuary, Georgetown, SC (33.3281° N, 79.1668° W) and filtered (0.25 μm) before addition of nutrients according to the NCMA instructions. This media was used for all NP exposures. The estuarine dinoflagellate Prorocentrum minimum (CCMP# 695) was obtained from the NCMA, Bigelow, Maine, USA and grown in L1-Si enriched seawater media (Guillard and Ryther, 1962; Guillard, 1975) at 23 ± 2 °C, on a 12 h light and dark cycle, with irradiance range of 90–100 μmol quanta m−2 s−1 to mimic natural conditions. The dinoflagellate culture was allowed to grow for four days prior to transfer to individual exposure bottles to ensure cells were in exponential growth (supported by cell count and fluorescence measurements) (Fig. S1). 2.4. Exposures Exposures were performed for 72 h in triplicate at concentrations between 0.1 and 1000 μg L−1 (Juristo and Moreno, 2001; Klaus, 2015) at 23 ± 2 °C on a 12-hour light/dark cycle with an irradiance range of 90–100 μmol quanta m−2 s−1 to mimic incubation and natural conditions. This length of time was chosen for exposures based on standard protocol for proper bio-assessment of phytoplankton growth (Munawar and Munawar, 1987; Schafer et al., 1994). For each exposure, 200 mL of media and 25 mL of cultured dinoflagellates (cell density > 100,000 cells mL−1) were added to individual glass bottles along with either an aliquot of stock cit-AgNP, PVP-AgNP, AgNO3 (dissolved Ag control), citrate control solution, or PVP control solution. Citrate and PVP control solutions were made by adding the same amount of citrate or PVP indicated in the AgNP synthesis. Neither solution influenced fluorescence or cell growth (Fig. S2). To ensure consistent light exposure, an

2.2. NP characterization Stock solutions were analyzed for total Ag concentration via inductively coupled plasma optical emission spectrometry (ICP-OES) (Varian 710-ES). Transmission electron microscopy (TEM) preparation and analysis procedure were performed as described in the literature (Prasad et al., 2015). Size, polydispersity index (pdi), and zeta potential measurements of stock NPs were measured using a Malvern Nanosizer Nano-Zs). The mean of at least three measurements were reported (Tejamaya et al., 2012). Absorption spectra of stock NPs after 24, 48, and 72 h in exposure media were measured over a wavelength of 200–800 nm using a UV–visible spectrophotometer (Shimadzu UV2600). In exposures without algae, pdi and size by dynamic light scattering 2

NanoImpact 15 (2019) 100181

S.V. Butz, et al.

A.

B.

120 90

15

60

10

30

5

0

0 0 hr

C.

24 hr

48 hr

72 hr

120

D. Ionic Ag (µg)

90

Ionic Ag (%)

20

60 30 0

0 hr

24 hr

48 hr

72 hr

0 hr

24 hr

48 hr

72 hr

20 15 10 5 0

0 hr

24 hr

48 hr

72 hr

E. 120

20

90

15

60

10

30

5

0

0 0 hr 1 µg/L

24 hr 10 µg/L

48 hr

72 hr

0 hr

100 µg/L

1 µg/L

24 hr 10 µg/L

48 hr

72 hr

100 µg/L

Fig. 1. AgNP dissolution in exposure media. Dissolution results from ultracentrifugation of AgNPs in L1-Si media over 72 h. (A) percent ionic cit-AgNP treatment (B) mass of ionic Ag cit-AgNP treatment (C) percent ionic Ag PVP-AgNP treatment (D) mass of ionic Ag PVP-AgNP treatment (E) percent ionic Ag AgNO3 treatment (F) mass of ionic Ag AgNO3 treatment (not 100% dissolution accounted for due to loss) (note the difference in y-axis maximum value in Fig. F compared to B and D).

2.5. Growth inhibition

ultraviolet spherical quantum probe (Biospherical Instruments Inc., model QSI2101) was used along with the Logger 2100 program to measure the exposure irradiances. Exposure bottles were static during exposure but gently shaken and inverted before samples were taken to resuspend phytoplankton that may have settled to the bottom of the bottle and ensure good sampling. After 72 h, an aliquot was taken from each bottle and centrifuged (2000 rpm, 4 °C, 10 min) to precipitate suspended cells. The supernatant (10 mL) was removed and acidified (1% nitric acid). The pellet of cells was washed with 10 mL of fresh media via centrifugation under same conditions. This process was repeated twice. The washes were removed, combined, and acidified (1% nitric acid). The remaining algal pellet was digested with concentrated nitric acid. All acidified samples were diluted to 1% acid and measured by ICP-MS. The same QA/QC methods as stated in Section 2.2 were implemented for these samples. We assumed the Ag remaining with the precipitate was either, operationally defined as internalized or strongly bound to cells to the cell wall and Ag removed during the washing process was weakly bound to the cell wall. Ag could be internalized through endocytosis or other mechanisms associated with the cell wall. However, analysis of interactions at this level were not conducted, therefore throughout this study we refer to this fraction as internalized or strongly bound.

Four separate inoculum cultures were measured every 24 h for 4 days for chlorophyll a (Chl. a) in raw fluorescence units (RFUs) (Turner Trilogy Fluorometer with Chl a-in vivo module) and cell density enumerated using inverted light microscopy (Olympus IMT-2 microscope, 10 μL Burker counting chamber). A linear regression analysis was performed to relate the parameters of fluorescence (RFU) and cell density (cells mL−1) (adj. R2 = 0.78, p < 0.05) (Fig. S3). Once this relationship was determined only fluorescence was measured and used to calculate cell density during exposures. Analysis of variance (ANOVA) and REGWF, post hoc statistical analysis was performed to test for significant differences between concentration effect on cell density. Sigmoidal concentration-response equations were derived for each treatment (Table S4). From these equations EC50, no effect concentration (NOEC), and predicted no effect concentration (PNEC) (Bureau, 2003) values were determined for each treatment using the non-linear curve fitting procedure in IBM SPSS v. 21 (Motulsky and Christopoulos, 2004). NOEC values were divided by 1000 to give PNEC values to account for uncertainties of inhibition occurring at lower exposure concentrations not identified in this study (ECB, E. C. B., 2003; Mueller and Nowack, 2008). The PEC value of 0.166 ng L−1 was taken from Gottschalk et al. (Gottschalk et al., 2009; 3

NanoImpact 15 (2019) 100181

S.V. Butz, et al.

A.

primarily via dissolution. However, TEM characterization also showed the presence of NPs with a primary NP size of 20.7 ± 6.5 nm, which is significantly larger than the original NPs (p < 0.05). It is possible that the remaining AgNPs increased in size due to smaller particles dissolving faster leaving larger particles in suspension. Most images contained agglomerates with a mean size of 138.6 ± 15.2 nm (Fig. S7), rather than individual NPs, and agglomerates were not observed in the original NP sample, indicating aggregation occurred in media. Therefore, it is likely that a fraction of NPs grew in size via dissolution and ripening, and additionally appeared as larger agglomerates rather than dispersed NPs. This is in agreement with previous studies with substantial transformations which were dependent on NP properties and concentration and media properties (Tejamaya et al., 2012; Misra et al., 2012; Griffitt et al., 2008). Based on the behavior of Ag in seawater and speciation modeling (data not shown), it is likely that dissolved Ag is present as chloro-complexes. After 72 h in exposure media PVP-AgNPs remained largely unchanged with mean TEM size 17.4 ± 7.2 nm (p < 0.05) (Fig. S4B). UV–vis analysis indicated that PVP-AgNPs remained in solution and were detectable in exposure media after 72 h with roughly 20% decrease in concentration (Fig. S5). DLS results confirm the stability of PVP-AgNPs in media (Table 1, Fig. S8). Similar losses of AgNPs in media have also been observed elsewhere (Tejamaya et al., 2012). For both NP types, the masses of dissolved ions increased with increasing added AgNPs, although the percentage of ionic Ag decreases with increasing concentration (Fig. 1). This result is consistent with previous data (Hadioui et al., 2013; Azimzada et al., 2017; Merrifield et al., 2017). A possible explanation is that higher concentrations results in more and rapid agglomeration of the NPs, which reduces the specific surface area of NPs, reducing dissolution. Ionic Ag data combined with TEM, DLS and UV–vis data again suggest agglomeration, dissolution and the complexation of dissolved Ag to silver chlorocomplexes in suspension or nucleation of new NPs, but at much reduced levels compared to the cit-AgNPs, which is in agreement with results from previous research at similar concentrations (Merrifield et al., 2017). The increased stability of PVP-AgNPs compared to citAgNPs was expected given that PVP is sterically stable and less prone to aggregation in high salt media (Huynh and Chen, 2011; Hitchman et al., 2013). Results further suggest that, at lower concentrations, behavior is governed by dissolution while, at higher concentrations aggregation is dominant. Loss of ionic Ag in AgNO3 exposures can be attributed to sorptive losses, with an average of roughly 20% lost (Fig. 1E–F).

Ag (µg / 100,000 cells)

4

3

2

1

0

Percentage of total Ag added (%)

B. 40 30 20 10 0

Internal or strongly bound

Loosely bound

Fig. 2. Ag measured with P. minimum cells after 72 h of exposure. Presented in (A) % of total Ag added. Internal and loosely bound represented- portion no represented includes material in suspension, dissolved, sedimented, or absorbed to the glassware. (B) mass of Ag added in μg per 100,000 cells. Internalization analysis was performed on 1, 40 and 100 ppb for each exposure. Each graph shows the concentrations of 40 and 100 ppb. 1 ppb exposures were below the detection limit of the analytical technique (ICP-OES).

3.2. Ag accumulation

Giese et al., 2018) and used to determine the PEC/PNEC risk assessment ratios (Gottschalk et al., 2015). If the resultant value is < 1 then there is no immediate concern for this substance, however, if the value is > 1, the substance is “of concern”, and requires further research or action (ECB, E. C. B., 2003; Mueller and Nowack, 2008). In addition, if the PEC/PNEC ratio is > 0.1 and ≤1 future comparative analysis is suggested (Bureau, 2003).

In all treatments, Ag was internalized or strongly bound to the dinoflagellate cell wall. For all exposures, the mass of Ag accumulated increased with increasing exposure concentration; as exposure concentration increased percent accumulated decreased (Fig. 2). AgNO3 exposures resulted in the greatest Ag accumulation after 72 h (100 μg L−1 exposure) (1.95 ± 0.86, 0.58 ± 0.02 μg) (internalized or strongly bound, weakly bound), followed by PVP-AgNP (1.40 ± 1.06, 0.26 ± 0.22 μg) then cit-AgNP (1.14 ± 0.54, 0.31 ± 0.02 μg) (Fig. 2). However, statistical analysis indicated no significant difference between treatments (t-tests, p > 0.05). Mass balance calculations indicated that the strongly bound Ag (after 72-hours, 100 μg L−1 exposure) accounts for 5.1% in cit-AgNP, 6.2% in PVP-AgNP, and 8.7% in AgNO3 exposures (Table S5 and Table S6). In addition, after 72 h at the 100 μg L−1 concentration, 35.2% (7.91 ± 1.03 μg) and 18.8% (4.22 ± 0.31 μg) of Ag was measured as ionic Ag remaining in media in the cit- and PVP-AgNP exposures, indicating that the majority of Ag was not taken up but remained in the aqueous phase. Based on mass balance calculations 38.4% (8.64 ± 0.53 μg) and 53.9% (12.12 ± 0.53 μg) of Ag remained as nanoparticulate Ag remaining in the media (Table S5 and Table S6). Citrate is weakly bound to the AgNP

3. Results and discussion 3.1. Nanoparticle characterization TEM of stock NPs confirmed a spherical particle shape for both citand PVP-AgNPs with an average particle size of 14.0 ± 1.7 nm and 14.8 ± 5.7 nm (mean ± 1 standard deviation n = 100) (Table 1, Fig. S4). DLS and UV–vis characterization values for stock NPs can be found in Table 1. Cit-AgNPs underwent substantial changes in the exposure media. UV–vis analysis showed cit-AgNPs were lost from the media rapidly (Fig. S5) which, combined with DLS results, confirm rapid NP transformation in media (Table 1 and Fig. S6). Most likely this loss was 4

NanoImpact 15 (2019) 100181

S.V. Butz, et al.

A.

1.4

*

1.2

0.1 ppb CIT

1.0

1 ppb CIT 10 ppb CIT

0.8

*

*

0.6 0.4

40 ppb CIT 70 ppb CIT

*

*

100 ppb CIT

*

1000 ppb CIT

0.2 0.0

Normalized Cell Density

0 hrs

B.

24 hrs

48 hrs

72 hrs

1.4 1.2

0.1 ppb PVP

1.0

1 ppb PVP 10 ppb PVP

0.8 0.6

*

0.4

40 ppb PVP

*

*

*

70 ppb PVP 100 ppb PVP 1000 ppb PVP

0.2 0.0 0 hrs

C.

24 hrs

48 hrs

72 hrs

1.4 1.2 1.0

0.1 ppb Ag 1 ppb Ag 10 ppb Ag 40 ppb Ag 100 ppb Ag

0.8

*

0.6 0.4

*

*

0.2

*

*

0.0 0 hrs

24 hrs

48 hrs

*

72 hrs

Fig. 3. P. minimum cell density normalized to the control. Graphed P. minimum abundance over 72-hour exposure. Cell density measurements were normalized to the control (0 ppb exposure). The 0 h time point was taken immediately after addition of respective concentration AgNP or AgNO3. (A) Cit-AgNP (B) PVP-AgNP (C) AgNO3 control. Asterisks (*) indicate a significantly different result. Table 2 Toxicity results. Toxicology results of cit-AgNPs, PVP-AgNPs, and the positive control (AgNO3) exposures after 72 h in exposure media. PEC value of 0.611 ng L−1 was used in the equation to determine the PEC/PNEC ratio. Indicator EC50 (μg L−1) NOEC (μg L−1) PNEC (μg L−1) PEC/PNEC ratio

cit-AgNP

95% confidence internal

PVP-AgNP

95% confidence internal

AgNO3

95% confidence internal

60.5 24.0 0.02 0.007

66.3–56.0 30.8–21.0 0.03–0.01 0.04–0.00

67.1 19.0 0.02 0.009

68.9–65.0 22.1–17.6 0.02–0.01 0.04–0.00

30.3 2.0 0.002 0.08

37.6–27.4 4.1–1.2 0.007–0.00 0.51–0.00

and 10 μg L−1 of all treatments and to 40 μg L−1 of cit- and PVP-AgNPs. A constant decrease in cell density was measured in 40 and 100 μg L−1 of AgNO3, 70, 100 and 1000 μg L−1 of cit-AgNPs and PVP-AgNPs. An increase in growth was shown in both NP treatments at different concentrations and times. This indicates that algae are able to recover after initial exposure to low concentrations of AgNPs. These results show that all treatments (cit-AgNPs, PVP-AgNPs, AgNO3) impact dinoflagellate density and show a dose and time relationship with response. In all treatments cell densities decreased over time in high exposure concentrations while at low concentrations there is an initial decrease and increase over time, however this trend of an increase over time is not supported by statistical significance. The similarities and differences between AgNO3 and AgNP results suggest that AgNPs dissolve during exposure and dinoflagellates are exposed to Ag ions and newly formed NPs, possibly containing chloride. AgNPs release Ag, which results in the increase in cell density in cit-AgNPs as

by charge stabilization which makes it more susceptible to dissolve and release Ag+ ions, therefore behaving similarly to an ionic exposure. PVP-AgNPs experience less dissolution than cit-AgNPs, but had 15% dissolution at the end of exposures, which is consistent with other work conducted on PVP-AgNPs (Merrifield and lead, 2016). In 100 μg L−1 exposures there was significantly less Ag weakly bound (in wash) to the surface of the dinoflagellate than strongly bound (p < 0.05, t-test) after 72 h. Results from 1 μg L−1 exposures were below the detection limit (Table S5). 3.3. Growth inhibition As exposure concentration increased, cell density decreased over time (Fig. 3). The significant differences resulting from statistical analysis (ANOVA and REGWF, post hoc test) are shown in Fig. 3. No significant change in cell density was seen over 72 h of exposure to 0.1, 1, 5

NanoImpact 15 (2019) 100181

S.V. Butz, et al.

Fig. 4. P. minimum inhibition dose response curves (72 h). Graphed results of dinoflagellate inhibition dose response curves. To obtain each curve a regression was performed with calculated percent inhibition at 72 h as the dependent variable and treatment concentration as the independent variable fit to the (sigmodal) dose response curve. Curves for all three treatments are shown corresponding R2 values are: cit-AgNP 0.928, PVP-AgNPs 0.921, and AgNO3 0.954.

Inhibition (%)

100 75 50 25 0 50

AgNO3

100

150

Concentration (µg

L-1)

cit-AgNP

200

pvp-AgNP the increased recovery in cit-AgNP as compared to PVP-AgNP exposures could be due to a nano-specific effect. Cit-AgNPs are less stable than PVP-AgNPs in complex media (Tejamaya et al., 2012) due to capping agents and steric versus charge stable bonds, respectively. Therefore, more Ag is dissolved from cit-AgNPs and ionic Ag is likely present as silver-chloride complexes therefore being similar to an ionic Ag exposure than AgNP exposure and thus more toxic than PVP-AgNPs that remain relatively more stable in complex medias. NOEC values were determined to be 20 μg L−1 for PVP-AgNPs, 24 μg L−1 for cit-AgNPs, and 2 μg L−1 for AgNO3 exposures (Table 2). NOEC values for both particle types were not statistically different (p > 0.05), while the NOEC for AgNO3 was an order of magnitude lower, and significantly different. PNEC vales were calculated by risk assessment procedures using a value of 1000 (Gottschalk et al., 2009) due to the limited knowledge of nano-Ag in the environment and calculated as 0.02 μg L−1, 0.02 μg L−1, and 0.002 μg L−1 (PVP-AgNP, citAgNP, AgNO3) (Table 2). A PEC value of 0.166 ng L−1 (1.66 × 10−4 μg L−1), (Sun et al., 2016) was divided by PNEC values to calculate PEC/ PNEC risk assessment ratios for each treatment. For all treatments PEC/ PNEC ratios were < 1 (0.007, 0.009, 0.08) suggesting that for P. minimum risk is limited and AgNPs do not pose an immediate threat to this estuarine species (Table 2). Despite the conclusions that no immediate risk to P. minimum is found, there are considerable uncertainties. For instance, PEC values are poorly known, not validated by analytical techniques, and the uses and discharge of AgNPs are rapidly increasing (Nanotechnologies, T. P. o. E., n.d.; Wiesner et al., 2006). In addition, subtle changes were observed at low concentrations, although populations were able to recover over time if the concentrations were not too high. To properly assess the risk AgNPs present to the aquatic environment analytical analysis of NP characteristics, uptake, accumulation, and accurate values of nano-Ag in the environment must be identified.

compared to AgNO3 treatments. Although not supported by statistical significance, the increase in cell densities over time may suggest that algae have recovery mechanisms which allow them to detoxify added Ag under certain conditions. EC50 values were determined as 60.5 (66.3–56.0), 67.1 (68.9–65.0), and 30.3 (37.6–27.4) μg L−1 (cit-AgNP, PVP-AgNP, AgNO3) (EC50, ± 95% confidence interval) (Table 2, Fig. 4). Although as previously discussed AgNO3 data suggested that cit-AgNPs behave in some ways similarly to the AgNO3 exposures but there was no significant difference between the toxic effects observed by NP type. If we assume that the ion control (AgNO3) is not changing, then data suggests that AgNPs form ions even at low concentrations. However, the Ag ion in the AgNO3 control is subject to chemical changes (formation of new NPs and formation of complexes) making interpretation of data from the control problematic. Silver speciation research and modeling (VisualMinteq) indicates that the majority of Ag (%) is present in seawater as silver-chloride species, mainly as AgCl2. Thus, suggesting that silverchloride complexes may be the cause of toxicity. It has been seen in other studies that as Ag mass accumulated increased, toxicity increased (Taylor et al., 2016). Our results concur that in both ionic and nanoparticulate Ag exposures cell densities decreased with increasing Ag mass accumulated. Furthermore, there was no significant difference in Ag accumulation in or strongly bound to P. minimum between AgNPs and AgNO3 cells but the toxicity between AgNPs and ionic Ag was significantly different with ionic Ag being more toxic than AgNPs. This difference in observed toxicity between Ag ion and NP can be explained as follows. The form of the Ag that is taken up is different and more of the AgNO3 is actually internalized, while more of the NPs are sorbed externally, possibly giving rise to unknown, and potentially longer term, effects as they slowly dissolve and the ions are internalized. A cellular study would be beneficial to identify Ag as internalized within the cell. If Ag is not internalized, and is instead strongly bound or associated with the outside of the cell wall then this could be an explanation of why accumulation and toxicity results do not compliment previous studies. However, previous research has shown that NPs are taken up and accumulated within the cell (Sekine et al., 2017; Taylor et al., 2016; Merrifield et al., 2018). Thus, combined with our results we suggest a nano-specific effect in accumulation and toxicity which there is differential uptake of Ag by cells in low exposure concentrations, as seen in Merrifield et al. 2018 (Merrifield et al., 2018). The ability for cell densities to recover after initial exposure to AgNPs can also be explained by differential uptake between individual cells. We hypothesize that cells take up different amounts of Ag, thus cells may be able to survive with less Ag mass accumulated or may be able to depurate low amounts of Ag mass. It is further understood that

3.4. Conclusion We have shown that Ag can be taken up or strongly bound to P. minimum cells after exposure to AgNPs. We have also shown a decrease in cell density in a time- and dose-dependent manner, with differences between the particles and ions observed. From these data and at current levels, however, no immediate risk of AgNPs to P. minimum is indicated. Further testing and mechanistic understanding of AgNP interactions with marine algae should be continued. Although risk assessment of a single algal species is important for the identification of species-specific responses, especially HAB species, to fully understand the impact and toxicity of AgNPs on the estuarine aquatic environment, multi-species 6

NanoImpact 15 (2019) 100181

S.V. Butz, et al.

algal assays should be conducted.

Juristo, N., Moreno, A.M., 2001. Basics of Software Engineering Experimentation. Springer, Boston, MA. Klaus, B., 2015. Statistical relevance—relevant statistics, part I. EMBO J. 34 (22), 2727–2730. Kleiven, M., Macken, A., Oughton, D.H., 2019. Growth inhibition in Raphidocelis subcapita: evidence of nanospecific toxicity of silver nanoparticles. Chemotherapy 221, 785–792. Leclerc, S., Wilkinson, K.J., 2014. Bioaccumulation of nanosilver by Chlamydomonas reinhardtii-nanoparticle or the free ion? Environ. Sci. Technol. 48, 358–364. Lodeiro, P.; Browning, T. J.; Achterber, G., A.; El-Shahawi, M. S., Mechanisms of silver nanoparticle toxicity to the coastal marine diatom Chaetoceros curvisetus. Nature Scientific Reports 2017, 7 (10777). Merrifield, R.C., lead, J.R., 2016. Preparation and characterization of three-layer, isotopically labelled core-shell nanoparticles; a tool for understanding mechanisms of bioavailability. NanoImpact 2, 54–60. Merrifield, R.C., Stephan, C., Lead, J., 2017. Determining the concentration dependent transformations of ag nanoparticles in complex media: using SP-ICP-MS and Au@Ag Core−Shell nanoparticles as tracers. Environ. Sci. Technol. 51, 3206–3213. Merrifield, R.C., Stephan, C., Lead, J.L., 2018. Quantification of Au nanoparticle biouptake and dstribution to freshwater algae using single cell - ICP-MS. Environ. Sci. Technol. 52 (4), 2271–2277. Misra, S.K., Dybowska, A., Berhanu, D., Luoma, S.N., Valsami-Jones, E., 2012. The complexity of nanoparticle dissolution and its importance in nanotoxicological studies. Sci. Total Environ. 438, 225–232. Motulsky, H., Christopoulos, A., 2004. Fitting Models to Biological Data Using Linear and Nonlinear Regression: A Practical Guide to Curve Fitting. Oxford University Press Inc., New York. Mueller, N.C., Nowack, B., 2008. Exposure modeling of engineered nanoparticles in the environment. Environ. Sci. Technol. 42 (12), 4447–4453. Munawar, M., Munawar, I.F., 1987. Phytoplankton bioassays for evaluating toxicity of in situ sediment contaminants. Hydrobiologia 149 (1), 87–105. Nanotechnologies, T. P. o. E Consumer Products Inventory. http://www.nanotechproject. org/cpi. Navarro, E., Wagner, B., Odzak, N., Sigg, L., Behra, B., 2015. Effects of differently coated silver nanoparticles on the photosynthesis of Chlamydomonas reinhardtii. Environ. Sci. Technol. 49, 8041–8047. Nowack, B.; Bucheli, T. D., Occurrence, behavior and effects of nanoparticles in the environment. Environ. Pollut. 2007, 157, 5–22. Piccinno, F., Gottschalk, F., Seeger, S., Nowack, B., 2012. Industrial production quantities and uses of ten engineered nanomaterials in Europe and the world. J. Nanopart. Res. 14 (9), 1–11. Prasad, A., Lead, J.L., Baalousha, M., 2015. An electron microscopy based method for the detection and quantification of nanomaterial number concentration in environmentally relevant media. Sci. Total Environ. 537, 479–486. Ribeiro, F., Gallego-Urrea, J.A., Jurkschat, K., Crossley, A., Hassellov, M., Taylor, C., Soares, A.M.V.M., Lourerio, S., 2014. Silver nanoparticles and silver nitrate induce high toxicity to Pseudokirchneriella subcapitata, Daphnia magna and Danio rerio. Sci. Total Environ. 466-467, 232–241. Römer, I., White, T.A., Baalousha, M., Chipman, K., Viant, M.R., Lead, J.R., 2011. Aggregation and dispersion of silver nanoparticles in exposure media for aquatic toxicity tests. J. Chromatogr. A 1218, 4226–4233. Schafer, H., Hettler, H., Fritsche, U., Pitzen, G., Roderer, G., Wenzel, A., 1994. Biotests using unicellular algae and ciliates for predicting long-term effects of toxicants. Ecotoxicol. Environ. Saf. 27 (1), 64–81. Sekine, R., Moore, K.L., Matzke, M., Vallotton, P., Jiang, H., Hughes, G.M., Kirby, J.K., Grovenor, C.R.M., Svendsen, C., Lombi, E., 2017. Complementary imagine of silver nanparticle interactions with green algae: dark-field microscopy, electron microscopy, and nanoscale secondary ion mass spectrometry. ACS Nano 11, 10894–10902. Sendra, M., Yeste, M.P., Gatica, J.M., Garrido-Moreno, I., Blasco, J., 2017. Direct and indirect effects of silver nanoparticles on freshwater and marine microalgae (Chlamydomonas reinhardtii and Phaeodactylum tricornutum). Chemosphere 179, 279–289. Shankar, P.D., Shobana, S., Karuppusamy, I., Pugazhendhi, A., Ramkumar, V.S., Arvindnarayanf, S., Kumar, G., 2016. A review on the biosynthesis of metallic nanoparticles (gold and silver) using bio-components of microalgae: formation mechanism and applications. Enzym. Microb. Technol. 95, 28–44. Sondi, I., Salopek-Sondi, B., 2004. Silver nanoparticles as antimicrobial agent: a case study on E. coli as a model for gram-negative bacteria. J. Colloid Interface Sci. 275, 177–182. Sorensen, S.N., Baun, A., 2015. Controlling silver nanoparticle exposure in algal toxicity testing- a matter of timing. Nanotoxicology 9 (2), 201–209. Srikar, S.K., Giri, D.D., Pal, D.B., Mishra, P.K., Upadhyay, S.H., 2016. Green synthesis of silver nanoparticles: a review. Green and Sustainable Chemistry 6, 34–56. Sun, T.Y., Bornhoft, N.A., Hungerbuhler, K., Nowack, B., 2016. Dynamic probabilistic modeling of environmental emissions of engineered nanomaterials. Environ. Sci. Technol. 50, 4710–4711. Taylor, N.S., Merrifield, R.C., Williams, T.D., Chipman, J.K., Lead, J.L., Viant, M.R., 2016. Molecular toxicity of cerium oxide nanoparticles to the freshwater algae Chlamydomonas reinhardtii is associated with the supra-environmental exposure concentrations. Nanotoxicology 10 (1), 32–41. Tejamaya, M., Romer, I., Merrifield, R.C., Lead, J.R., 2012. Stability of citrate, PVP, and PEG coated silver nanoparticles in ecotoxicology media. Environ. Sci. Technol. 46, 7011–7017. Turner, A., Brice, D., Brown, M.T., 2012. Interactions of silver nanoparticles with the marine macroalga, Ulva lactuca. Ecotoxicology 21, 148–154. Venkatesan, J., Kim, S., Shim, M.S., 2016. Antimicrobial, antioxidant, and anticancer

Acknowledgements Dinoflagellate cultures were prepared and maintained in Dr. Tammi L. Richardson's laboratory at the University of South Carolina, we would like to immensely thank her and her laboratory technicians for their assistance. Funding was provided in part by The Slocum-Lunz Foundation, The Guy Harvey Wildlife Foundation, The Center for Environmental Nanoscience and Risk, and The National Science Foundation (1508931). We appreciate this support. Appendix A. Supplementary data Supplementary data to this article can be found online at https:// doi.org/10.1016/j.impact.2019.100181. References Azimzada, A., Tufenkji, N., Wilkinson, K.J., 2017. Transformations of silver nanoparticles in wastewater effluents: links to Ag bioavailability. Environ. Sci.: Nano 4 (6), 1339–1349. Baalousha, M.A., Sikder, M., Prasad, A., Chandler, G.T., Lead, J.L., 2016. The concentrations-dependent behavior of nanoparticles. Environ. Chem. 1–3. Bureau, E.C., 2003. Technical Guidance Document on Risk Assessment. Institute for Health and Consumer Protection, E. C. Dublin. Cumberland, S., 2013. A.; J.R., L., Synthesis of NOM-capped silver nanoparticles: size, morphology, stability, and NOM binding characteristics. Sustain. Chem. Eng. 1, 817–825. Doty, R.C., Tashikhudo, R.T., Brust, M., Fernig, D.G., 2005. Extremely stable water-soluble ag nanoparticles. Chem. Mater. 17, 4630–4635. ECB, E. C. B, 2003. Technical Guidance Document on Risk Assessment. Institute for Health and Consumer Protection Part II. Giese, B., Klaessig, F., PArk, B., Kaegi, R., Steinfeldt, M., Wigger, H., von Gleich, A., Gottschalk, F., 2018. Risk, release and concentrations of engineered nanomaterial in the environment. Nat. Sci. Rep. 8 (1), 1565. Gomez, F., 2005. A list of free-living dinoflagellate species in the world's oceans. Acta Bot. Croat. 64 (1), 129–212. Gottschalk, F., Sonderer, T., Scholz, R.W., Nowack, B., 2009. Modeled environmental concentrations of engineered nanomaterials (TiO2, ZnO, Ag, CNT, fullerenes) for different regions. Environ. Sci. Technol. 43 (24), 9216–9222. Gottschalk, F., Lassen, C., Kjoelholt, J., Christensen, F., Nowack, B., 2015. Modeling flows and concentrations of nine engineered nanomaterials in the Danish environment. Int. J. Environ. Res. Public Health 12, 5581–5602. Griffitt, R. J.; Luo, J.; Gao, J.; Bonzongo, J.; Barber, D. S., Effects of particle composition and species on toxicity of metallic nanomaterials in aquatic organisms. 2008, 27 (9), 1972–1978. Guillard, R.R.L., 1975. Culture of phytoplankton for feeding marine invertebrates. In: Smith, W.L., Chanley, M.H. (Eds.), Culture of Marine Invertebrate Animals: Proceedings — 1st Conference on Culture of Marine Invertebrate Animals Greenport. Springer US, Boston, MA, pp. 29–60. Guillard, R.R.L., Ryther, J.H., 1962. Studies of marine planktonic diatoms. I. Cyclotella nana Hustedt and Detonula confervacea Cleve. Can. J. Microbiol. 8, 229–239. Hadioui, M., Leclerc, S., Wilkinson, K.J., 2013. Multimethod quantification of Ag+ release from nanosilver. Talanta 105, 15–19. Heil, C.A., Gilbert, P.M., Fan, C., 2005. Prorocentrum minimum (Pavillard) Schiller A review of a harmful algal bloom species of growing worldwide importance. Harmful Algae 4, 449–470. Henglein, A., Giersig, M., 1999. Formation of colloid silver nanoparticles: capping action of citrate. J. Phys. Chem. B 103, 9533–9539. Hitchman, A., Smith, G.H.S., Ju-Nam, Y., Sterling, M., Lead, J.R., 2013. The effect of environmentally relevant conditions on PVP stabilized gold nanoparticles. Chemosphere 90, 410–416. Huang, J., Cheng, J., Yi, J., 2016. Impact of silver nanoparticles on marine diatom Skeletonema costatum. Appl. Toxicol. 36, 1343–1354. Huynh, K.A., Chen, K.L., 2011. Aggregation kinetics of citrate and polyvinyalpyrrolidone coated silver nanoparticles in monovalent and divalent electrolyte solutions. Environ. Sci. Technol. 45 (13), 5564–5571. Ivask, A., Kurvet, I., Kasemets, K., Blinova, I., Aruoja, V., Suppi, S., Vija, H., Käkinen, A., Titma, T., Heinlaan, M., Visnapuu, M., Koller, D., Kisand, V., Kahru, A., 2014. Sizedependent toxicity of silver nanoparticles to bacteria, yeast, algae, crustaceans and mammalian cells in vitro. PLOS ONE 9 (7), e102108-1–14. Ivask, A., Visnapuu, M., Vallotton, P., Marzouk, E.R., Lombi, E., Voelcker, N., 2016. Quantitative multimodal analyses of silver nanoparticle-cell interactions: implications for cytotoxicity. NanoImpact 1, 29–38. Jana, N.R., Gearheart, L., Murphy, C.J., 2001. Wet chemical synthesis of silver nanoparticles and nanowires of controllable aspect ratio. Chem. Commun. 617–618. Jia, K., Sun, C., Wang, Y., Li, X., Mu, W., Fan, Y., 2019. Effect of TiO2 nanoparticles and multiwall carbon nanotubes on the freshwater diatom Nitzschia frustulum: evaluation of growth, cellular components and morphology. Chem. Ecol. 35, 69–85.

7

NanoImpact 15 (2019) 100181

S.V. Butz, et al.

of silver nanoparticles with algae and fish cells: a side by side comparison. Nanobiotechnology 15–16. Zhang, L., Li, J., Yang, K., Liu, J., Lin, D., 2016. Physicochemical transformation and algal toxicity of engineered nanoparticles in surface water samples. Environ. Pollut. 211, 132–140.

activities of biosynthesized silver nanoparticles using marine algae Ecklonia cava. Nanomaterials 6 (12), 235. Wiesner, M.R., Lowry, G.V., Alavarez, P., Dionysiou, D., Biswas, P., 2006. Assessing the risks of manufactured nanomaterials. Environ. Sci. Technol. 40 (14), 4336–4345. Yue, Y., Li, X., Sigg, L., Suter, M.J.-F., Pillai, S., Behra, R., Schirmer, K., 2017. Interaction

8