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Soil Sci. Soc. Am. J. 75:2011 Posted online 25 Oct. 2010 doi:10.2136/sssaj2010.0127nps Received 15 Mar. 2010. *Corresponding author (jason.unrine@uky.edu). © Soil Science Society of America, 5585 Guilford Rd., Madison WI 53711 USA All rights reserved. No part of this periodical may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, recording, or any information storage and retrieval system, without permission in writing from the publisher. Permission for printing and for reprinting the material contained herein has been obtained by the publisher.

Role of Particle Size and Soil Type in Toxicity of Silver Nanoparticles to Earthworms

Nanoparticles in the Environment

A relatively recent development in the manufacturing of consumer products has been the incorporation of nano-sized (<100-nm) particles. Because of

their antimicrobial properties, Ag NPs have been used in medical devices, food-storage containers, soaps and sanitizers, wound dressings, and fabrics (Luoma, 2008). With an increase in Ag NP manufacture and use, it is expected that Ag NPs will become a new and growing source of anthropogenic Ag in the environment (Blaser et al., 2008). Preliminary studies on the life cycle of Ag NPs have shown that Ag NPs incorporated into fabrics (usually in powdered or spun Ag forms) will primarily enter human waste streams and wastewater treatment plants (Benn and Westerhoff, 2008; Luoma, 2008). A majority (as much as 90%) of the particles are projected to be bound to sewage sludge, which may then be applied to terrestrial environments as biosolids (Blaser et al., 2008; Mueller and Nowack, 2008). Therefore, the environmental impacts of Ag NPs are of great concern in terrestrial systems that receive biosolids inputs.

Relatively little is known about Ag NP fate and toxicity relative to ionic forms of Ag, for which the toxicity in aquatic systems has been extensively studied (Hogstrand and Wood, 1998; Klaine et al., 2008). Thus far, Ag NPs have been shown to exhibit toxic actions toward a range of organisms, including bacteria (Morones et al., 2005; Panacek et al., 2006; Choi et al., 2008; Choi and Hu, 2008), yeast (Panacek et al., 2009), paramecia (Kvitek et al., 2009), algae (Griffitt

W. Aaron Shoults-WilsonDep. of Plant and Soil SciencesUniv. of KentuckyLexington, KY 40546

Brian C. ReinschDep. of Civil and Environmental EngineeringCarnegie Mellon Univ.Pittsburg, PA 15213

Olga V. Tsyusko Paul M. Bertsch

Dep. of Plant and Soil SciencesUniv. of KentuckyLexington, KY 40546

Greg V. LowryDep. of Civil and Environmental EngineeringCarnegie Mellon Univ.Pittsburg, PA 15213

Jason M. Unrine*Dep. of Plant and Soil SciencesUniv. of KentuckyLexington, KY 40546

Silver nanoparticles (NPs) are an emerging contaminant of concern due to their increased use. The earthworm Eisenia fetida was exposed to a range of concentrations of AgNO3 and two polyvinylpyrolidone coated Ag NPs with different particle size distributions. They were exposed in two different soils: a naturally occurring sandy loam and a standardized artificial soil. The AgNO3 significantly reduced E. fetida growth and reproduction at 7.41 ± 0.01 mg kg−1 Ag in the sandy loam but only reproduction was affected at concentrations of 94.1 ± 3.2 mg kg−1 in the artificial soil. In the artificial soil, significant (a = 0.05) reproductive toxicity was only observed in organisms exposed to the Ag NPs at concentrations approximately eight times higher than those at which the effects from ionic Ag were observed. Eisenia fetida exposed to either AgNO3 or Ag NPs in the sandy loam accumulated significantly (a = 0.05) higher concentrations of Ag than those exposed in the artificial soil and had higher bioaccumulation factors. Earthworms exposed to AgNO3 also accumulated significantly higher concentrations of Ag than those exposed to Ag NPs. No differences in toxicity were observed between the two size distributions. Extended x-ray absorption fine structure spectroscopy analysis of the soils indicated that the Ag was approximately 10 to 17% Ag(I), suggesting that Ag ions may be responsible for effects on growth and development caused by exposure to Ag NPs. Our results also suggest that soil type is a more important determinant of Ag accumulation from Ag NPs than particle size.

Abbreviations: ART, artificial soil; BAF, bioaccumulation factor; CEC, cation exchange capacity; EXAFS, extended x-ray absorption fine structure spectroscopy; ICP-MS, inductively coupled plasma–mass spectrometry; NP, nanoparticle; PVP, polyvinylpyrolidone; TEM, transmission electron microscopy; YSL, Yeager sandy loam soil.

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et al., 2008; Navarro et al., 2008), nematodes (Roh et al., 2009), daphnia (Griffitt et al., 2008), Japanese medaka (Chae et al., 2009), fathead minnow embryos (Laban et al., 2010), and all life stages of zebrafish (Asharani et al., 2008; Griffitt et al., 2008; Yeo and Kang, 2008; Bar-Ilan et al., 2009). In many of these studies, however, it was not clear if toxicity was caused by the particles themselves or Ag ions released as the result of oxidation and dissolution. Despite the recent interest in Ag NP toxicity, little has been done to investigate its toxicity to soil-dwelling organisms even though recent studies have indicated that land application of biosolids may be a major exposure pathway for Ag NPs. Furthermore, the toxicity of ionic Ag has not been widely studied in soils (Ratte, 1999; Nahmani et al., 2007b). To characterize the toxicity of Ag NPs in the soil environment, we performed bioaccumulation and subchronic toxicity studies using the model organism Eisenia fetida (Lumbricidae, Savigny, 1826). To address the fate of Ag NPs and the relative importance of Ag NP vs. Ag ion related toxicity, we compared the toxicity of Ag NPs to AgNO3 and performed extended x-ray absorption fine structure spectroscopy (EXAFS) to assess the speciation and oxidation of Ag NPs in soil.

It has been widely recognized that the chemical properties of both aquatic and soil systems will cause differences in metal speciation and toxicity (Paquin et al., 2002; Daoust et al., 2006; Rombke et al., 2006; Spurgeon et al., 2006; Bielmyer et al., 2007; Nahmani et al., 2007a). Especially important determinants of metal toxicity in soil systems are the pH, organic matter content, soil texture, and cation exchange capacity (CEC) (Daoust et al., 2006; Rombke et al., 2006; Spurgeon et al., 2006; Nahmani et al., 2007a). While the effects of soil properties on metal toxicity have been investigated for other metals such as Cd, Cu, Pb and Zn, they have not been well studied for Ag, although Ag may be a major factor in the toxicity of metal-contaminated soils (Nahmani et al., 2007a). Because biosolids contaminated by Ag NP may be applied to a variety of soils, understanding the factors that influence Ag NP toxicity will be important in determining the risk that Ag NPs pose to soil systems.

It is also critical to address the added dimensions of nanomaterial toxicity relative to metal ions (Unrine et al., 2008, 2010). For example, metal NPs may oxidize and dissolve, releasing metal ions. Numerous studies have observed significant differences between the toxicity of Ag ions and Ag NPs (Griffitt et al., 2008; Navarro et al., 2008; Bar-Ilan et al., 2009; Kvitek et al., 2009; Panacek et al., 2009; Roh et al., 2009); however, few have investigated the chemical speciation of Ag in exposure media. The physicochemical properties of nanomaterials, such as size, shape, crystallinity, surface area, magnetism, and surface functionalization, may also vary, potentially affecting toxicity (Handy et al., 2008; Nel et al., 2009; Phenrat et al., 2009). Previous studies examining various oxidative, cytotoxic, and developmental endpoints have provided evidence that smaller Ag NPs cause more toxicity than larger Ag NPs (Carlson et al., 2008; Choi and Hu, 2008; Bar-Ilan et al., 2009; Yen et al., 2009). These studies were primarily conducted in tissue culture;

however, there is little evidence to date that Ag NP size affects toxicity toward fully developed multicellular organisms in environmentally realistic exposure scenarios where aggregation may play an important role.

The objectives of the present study were first to compare Ag NP toxicity for particles of differing sizes with the toxicity of Ag ions. This objective was met by performing standardized (Organization for Economic Cooperation and Development [OECD], 2004) subchronic and reproductive toxicity tests for a model soil organism, the earthworm Eisenia fetida. Second, the effects of soil parameters were investigated by testing toxicity and uptake in two different soils, a standard artificially constituted soil (OECD, 2004), which had a relatively high CEC and organic matter and clay contents, and a locally collected, naturally occurring soil (Yeager sandy loam, a sandy, mixed, mesic Typic Udifluvent), which had a lower CEC and organic matter and clay contents. Finally, the influence of particle size on the toxicity and bioaccumulation of Ag was tested by exposing organisms to polyvinylpyrolidone (PVP) coated Ag NPs with different particle size distributions.

MATeRiALS And MeThOdSMaterial Characterization

Two Ag NP powders, with a PVP coating and stated nominal primary particle size distributions of 10 and 30 to 50 nm were purchased from NanoAmor (Nanostructured & Amorphous Materials, Houston, TX). The powders were both high purity (>99% Ag), which we verified by dissolving samples in concentrated, ultra-high-purity HNO3 and analyzing for metallic impurities using inductively coupled plasma–mass spectrometry (ICP-MS; see below). Primary particle size distributions were determined using transmission electron microscopy (TEM). For TEM analysis, the Ag powders were suspended in pure water by sonication, centrifuged briefly to remove large aggregates from suspension, and then both Au and Cu Pelco TEM grids (Ted Pella Inc., Redding, CA) were dipped into the suspensions. All grids were then dried in a laminar flow clean hood. Particle size determinations via TEM were statistically analyzed based on measurements of at least 100 particles per material from at least three separate micrographs using image analysis software (NIS Elements, Nikon, Tokyo). A Philips BioTWIN 12 transmission electron microscope (FEI Co., Hillsboro, OR) was used to examine the particles at 23,000 to 6800× magnification and an accelerating potential of 100 kV. The particle hydrodynamic radius in a deionized (DI) water suspension was determined by dynamic light scattering in stock solutions used to dose the soils using a Malvern Zeta-Sizer Nano-ZS (Malvern Instruments, Malvern, UK). The percentage by volume was calculated using the refractive indices 0.135 (real) and 3.99 (imaginary). The same instrument was used to determine the zeta potential using phase analysis light scattering.

Characterization of Silver Speciation in SoilX-ray absorption spectroscopy (XAS) was conducted at the

Stanford Synchrotron Radiation Laboratory (Menlo Park, CA) wiggler magnet beamline 11–2. Silver model compounds used in this study included: Ag2S (Alfa Aesar, Ward Hill, MA; ³99%), Ag2CO3 (Acros

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Organics, Geel, Belgium; ³99%), AgCl (Fisher Scientific, Hampton, NH; ³99.0%), AgNO3(s) (Fisher Scientific; ³99.7%), Ag2O (Sigma Aldrich, St. Louis, MO; ³99%), Ag2SO4 (Fisher Scientific; ³98.0%). Boron nitride (BN) powder was also used (99.5%, Alfa Aesar). Model compounds were diluted by 1:5 mass of dry powder sample to BN powder. Samples were analyzed undiluted. Both model compounds and samples were placed into 1.5-mm-thick Teflon holders and sealed with Kapton tape. X-ray absorption spectroscopy was performed using a double Si(220), F = 90°, monochromator crystal with the beam detuned 40% from maximum intensity to reduce harmonic signals. The experimental setup also included Ar-filled ionization chambers and a Ag calibration foil. Silver K-edge (25.514 keV) XAS spectra were collected across the energy range 25.284 to 26.635 keV at room temperature. Transmission mode was used for the model compounds and a Lytle (fluorescence) detector was used for the soil samples.

The data were analyzed using SixPACK software version 0.60. X-ray absorption scans of related samples were calibrated for changes in assigned monochromator energy to 25.514 eV by using the first derivative of the Ag calibration standard and averaging them together. These spectra were deglitched when needed. Background subtraction was performed using SixPACK the arbitrary start of extended x-ray absorption fine structure (EXAPS) region (E0) defined as 25.525 keV and background interatomic distance (R) = 1.0 Å. The resulting spectra were converted to frequency (k) space and weighted by k3 (Webb, 2005). The linear least-squares combination fitting procedure used was adapted from Kim et al. (2000), whereby only components that contributed significantly (a >10% decrease in fitting correlation based on reduced c2 value) to the fit were included. The spectra were fit to various k ranges, depending on the quality of the EXAFS data.

Subchronic and Reproductive Toxicity TestsThe design of the earthworm reproduction, bioaccumulation,

and subchronic lethality assays followed published guidelines (OECD, 2004). The artificial soil (ART) medium consisted of 69.57% dry mass quartz sand, 0.43% dry mass powdered CaCO3, 20% dry mass kaolin clay, and 10% dry mass sphagnum peat moss (sieved to <2 mm). The natural soil was a Yeager sandy loam (YSL) collected from a type site in central Kentucky (Estill County), air dried, and sieved to <2 mm. The soils were characterized for particle size distribution (micro–macropipette method; Burt et al., 1993; Sheldrick and Wang, 1993), pH (electrode in 5:1 solvent/soil; OECD 2004), organic matter (by combustion; Nelson and Sommers, 1982) and CEC (1 mol L−1, pH 7, NH4OAc extraction and saturation, bases analyzed via ICP-MS, NH4 analyzed via ion-selective electrode; Soil and Plant Analysis Council, 2000; Table 1). To treat dried soils, NPs were first suspended in 18 MW DI water by sonicating for 15 min at room temperature in a water-bath sonicator (FS20 Ultrasonic Cleaner, Fisher Scientific). The resultant suspensions were applied to soils at a rate of approximately 50% of the water holding capacity (approximately 26% v/w for the ART and 19.2% v/w for the YSL; OECD, 2004). The particle concentrations in the suspensions were adjusted to achieve the desired nominal concentrations while maintaining the soils at the same water holding capacity. The nominal concentrations of treatments in the artificial soils were 10, 100, and 1000 mg Ag kg−1 dry soil. Dissolved AgNO3 was also used to create

treatments with nominal concentrations of 10 and 100 mg Ag kg−1 dry soil. In a separate experiment, YSL buffered with 0.1% w/w CaCO3 (YSL+) was treated to achieve a nominal 10 mg Ag kg−1 dry soil for both Ag NPs and AgNO3. The CaCO3 was added to increase the soil pH and improve organism survival. Three replicate exposure containers of soil (750 g) were prepared for each treatment and allowed to equilibrate in an incubator overnight.

Ten adult earthworms with fully developed clitella were added to each exposure chamber. Worm masses ranged from 0.243 to 0.752 g and averaged 0.450 ± 0.101 g (mean ± standard error) after being gently cleaned on moistened paper towels. The animals were maintained in an environmental chamber at 20°C with 12 h of light per day. After 24 h of exposure, 5 g of a 50:50 mixture of dehydrated alfalfa (Medicago sativa L.) and commercially available chicken feed was added and hydrated with 5 mL of DI water. This feed mixture had previously been shown to allow rapid, sustained growth in the worms. At the end of each week of exposure, the earthworms were observed for mortality and given another 4 g of feed mixture. Container weights were monitored and water was added on a biweekly basis to maintain moisture content. After 28 d of exposure, the worms were observed for mortality, gently cleaned on paper towels, weighed, and placed on moistened filter paper inside petri dishes. The worms remained in the dishes for 24 h to void their guts before being sacrificed by freezing. An additional 5 g of food and 5 mL of water were added to the containers and they were incubated for an additional 28 d, with weekly water replenishment. At the end of 28 d, the exposure containers were submersed in 60°C water, forcing any juveniles to emerge to the surface, where they were counted. The soil was manually searched and cocoons were collected and counted. All cocoons were immediately placed in water and hatching success was determined using the flotation method (hatched cocoons float).

Silver Uptake TestsShort (14-d) exposures were used to investigate the kinetics of Ag

bioaccumulation using different sources of Ag in soils with different properties with time. The soils used were YSL, YSL amended with 0.1% w/w CaCO3 to improve earthworm survival, and ART prepared as described above. For each soil, 150 earthworms were weighed and used to define three representative size classes. While these varied from test

Table 1. Properties of the two soils used in this study: artificial soil (ART) and natural Yeager sandy loam (YSL).

Property ART YSL

pH In H2O 7.00 5.17

In KCl 6.23 4.01

Cation exchange capacity, cmol kg−1

Total 14.45 9.18

K 0.02 0.11

Ca 10.07 0.91

Mg 0.64 0.27

Na 0.05 0.05

Composition, %

Sand 79.12 76.34

Silt 6.71 16.53

Clay 14.17 7.13 Organic matter 7.65 1.77

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to test based on the weight of worms in the culture at the time, the size classes were approximately 0.200 to 0.325, 0.325 to 0.400, and >0.400 g. For each treatment, three replicate containers of spiked soil were made and allowed to equilibrate overnight. Each replicate represented a different size class. At each time point (1, 3, 7, 10, and 14 d), a single worm from each replicate was sampled to provide size-stratified samples and eliminate initial worm mass as a confounding factor in uptake. Worms were gently cleaned on paper towels, weighed, and placed on wet filter paper inside petri dishes for 24 h to void their guts and before being sacrificed.

inductively Coupled Plasma–Mass Spectroscopic Analyses

Whole earthworms and ?0.25-g dry soil samples were digested in concentrated trace metal grade HNO3 using a MARS Express microwave digestion system (CEM, Matthews, NC) according to USEPA (1996). Total Ag concentrations in the digestates were determined by ICP-MS following USEPA (1998), using an Agilent 7500cx (Agilent Technologies, Santa Clara, CA). Digestion sets included standard reference tissues (DOLT-4, National Research Council of Canada, Ottawa, ON) or soils (2710 Montana soil, National Institute of Standards and Technology [NIST], Gaithersburg, MD). Duplicate digestions and reagent blanks were also included with each digestion set. Analytical runs included duplicate dilutions, spike recovery samples, and intercalibration–cross-calibration verification samples. Previous experiments had demonstrated that the Ag NPs dissolved in concentrated HNO3 at room temperature (³95% recovery). The recovery of Ag after directly adding Ag NP powder to soils in the digestion vessels was 94.6 ± 2.1%. Therefore, the digestion procedure was adequate to completely oxidize and dissolve the NPs in the soil and differences between nominal and measured concentrations were due to difficulties spiking the soils with suspensions that were prepared from powders. The mean recovery of acid-leachable Ag was 76.95 ± 3.21% from NIST 2710 and 117.3% from DOLT-4. Recovery of diluted samples spiked with Ag standard before analysis was 100.6%. The mean relative difference between replicate subsamples of soil was 1.9%. Concentrations in earthworms exposed to clean soil as controls did not exceed the method quantification limit (MQL: 0.217 ng mL−1). Three worms exposed to low concentrations of Ag also did not exceed the MQL and were not considered for data analysis. All concentrations are expressed in terms of Ag.

Statistical AnalysesAll data were logarithmically transformed before statistical

analysis to increase normality and heterogeneity of variance and better satisfy the assumptions for parametric statistical analysis. The bioaccumulation factors (BAFs) for Ag were calculated by dividing the tissue concentrations of E. fetida by the concentrations in the exposure soil. Normality and homogeneity of variance for all data were tested using the Shapiro–Wilk test and Levene’s test, respectively. For normal and homogenously distributed data, significant differences of a treatment on growth, mortality, bioaccumulation, and reproduction between treatments and controls were tested using a one-way ANOVA. Non-normal data were tested using the Wilcoxon Rank Sum test. The same approach was used to compare endpoints between treatments in

different soils. Normal data with homogenous variance that had more than two treatments were analyzed using a one-way ANOVA followed by Duncan’s post hoc test. A criteria of a = 0.05 was assumed for all reporting of significance unless otherwise stated. All calculations were performed using SAS version 9.1 (SAS Institute, Cary, NC).

ReSULTSMaterial Characterization

Transmission electron microscopy indicated that the smaller particles had a mean primary particle diameter of 40.88 ± 0.81 nm (mean ± standard error of the mean), while the larger particles had a mean diameter of 56.35 ± 1.16 nm (Fig. 1A and 1B). Dynamic light scattering (Fig. 1C and 1D) indicated that the smaller particles were much closer to their nominal size (10 nm) on average than indicated by TEM, possibly due to better dispersion of particles in suspension than when dried onto TEM grids. The larger particles (nominally 30–50 nm) were found to have a greater proportion of larger particles that could not be dispersed by sonication. The z-average diameters were 114.3 nm for the smaller particles and 154.8 nm for the larger particles; however, due to the polydispersity of the suspensions, these values do not reflect the volume-weighted particle size distributions. The electrophoretic mobility of the particles was found to be −2.6 ± 0.1 mmol cm V−1 s−1 for the smaller particles and −1.9 ± 0.0 mmol cm V−1 s−1 for the larger particles. The zeta potential estimated using the Hückel approximation was −49.5 ± 1.7 mV for the smaller particles and −35.9 ± 0.8 mV for the larger particles. These findings are not surprising. While particles with the same coating would be predicted to have the same surface charge, in actuality particle size has been shown influence surface charge (Hotze et al., 2010).

Characterization of Silver SpeciationSilver nanoparticles exposed only to air but not to soil

were found to be partially oxidized (Table 2; Fig. 2A), with approximately 10 to 11% in the form of Ag2O and 89 to 90% as Ag0. This corresponds to the percentages of Ag2O and Ag0 found for both sizes of Ag NPs aged in the ART for 28 d (Fig. 2B and 2D); however, smaller (10-nm) Ag NPs aged in the YSL soil had a higher percentage of Ag2O (15%) and consequent lower percentage of Ag0 (84%) (Fig. 2C). Larger (30–50 nm) Ag NPs aged in the YSL (Fig. 2E) indicated possibly lowered amounts of Ag0 (74%) than particles exposed to the air or aged in the ART, but the addition of Ag2O did not improve the fit. Unfortunately, the poor fit of the model does not allow us to draw conclusions on whether Ag NP size influences oxidation.

earthworm Mortality, Growth, and ReproductionEisenia fetida exposed to Ag and Ag NPs in the ART showed

no significant concentration-dependent decreases in growth or survival (Table 3). Exposure to 7.41 ± 1.19 mg kg−1 ionic Ag in the YSL, however, resulted in a significant decrease in growth.

A significant decrease in the number of cocoons produced per worm was observed in 94.12 ± 3.21 mg kg−1 AgNO3, 801.0

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± 74.9 mg kg−1 10-nm Ag NPs, and 773.3 ± 11.2 mg kg−1 30- to 50-nm Ag NPs when spiked in the ART (Table 4). There were no significant differences in the hatching rate of cocoons at any concentration in the ART treatments and no juveniles survived to be counted, including in the controls, indicating that

conditions in the ART were not suitable for neonate survival. On the other hand, 7.41 ± 1.19 mg Ag kg−1 dry soil of AgNO3 in the YSL caused significant decreases in the hatching success of cocoons and the number of juveniles counted but not the number of cocoons that were produced.

Fig. 1. Size distributions of polyvinylpyrolidone coated Ag nanoparticles (nPs) determined by two methods: primary particle sizes as determined by transmission electron microscopy (TeM) for (A) 10-nm and (B) 30- to 50 nm Ag nPs; particle sizes in dispersion as measured by dynamic light scattering for (C) 10-nm and (d) 30- to 50-nm Ag nPs; and TeM images of (e) 10-nm and (F) 30- to 50-nm Ag nPs.

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Uptake of Silver

Individuals exposed to AgNO3 consistently accumulated higher concentrations of Ag than worms exposed to Ag NPs. For all concentrations and both soils, at the end of 28-d exposures (Fig. 3A), tissue concentrations of Ag were significantly (a = 0.05) higher in individuals exposed to AgNO3 than those exposed to statistically similar concentrations of Ag NPs. Additionally, individuals exposed to AgNO3 and Ag NPs in the natural YSL soil accumulated significantly higher concentrations of Ag than those exposed to similar concentrations in the ART (Fig. 3A).

Overall, there were few significant differences in accumulation between time points in a given treatment after Day 1 in the 14-d uptake experiments (Fig. 4). In general, higher exposure concentrations resulted in a greater accumulation of Ag by the organisms. For comparison between treatments, accumulation values were pooled across Days 3 to 14. These experiments produced similar results to the bioaccumulation data of the 28-d reproductive toxicity tests (Fig. 5A). Organisms exposed to AgNO3 and Ag NPs in the YSL accumulated significantly higher concentrations of Ag than those exposed in the ART, while there did not appear to be a consistent difference in accumulation between exposures in pure YSL and YSL amended with CaCO3 (Fig. 5A). As with the 28-d exposures, organisms exposed to AgNO3 accumulated significantly more Ag than those exposed to Ag NPs at similar concentrations (Table 5). This was not observed consistently at all time points, however, with later time points being more likely to show a significant difference.

In most instances, E. fetida exposed to Ag NPs of different sizes did not accumulate significantly different concentrations of Ag (Fig. 3A and 5A). In a few instances, however, a significant difference in Ag accumulation was detected between the two Ag NPs. After 28 d in the ART, E. fetida accumulated significantly higher concentrations of Ag when exposed to 79.45 ± 1.02 mg kg−1 of the larger particles than when exposed to 84.15 ± 2.54 mg kg−1 of the smaller particles (Fig. 3A). This same relationship was also found in E. fetida exposed for 14 d to 18.85 ± 0.82 mg kg−1 Ag of the smaller particles and 18.71 ± 1.40 of the larger particles in the YSL amended with CaCO3 (Fig. 5A).

Bioaccumulation FactorsBioaccumulation factors were calculated to better

compare accumulation patterns with time between the different forms of Ag. After 28-d exposures, earthworms exposed to AgNO3 had significantly higher BAFs than those exposed to Ag NPs (Fig. 3B). Additionally, BAFs were significantly higher for earthworms exposed in the YSL rather than the ART (Fig. 3B). At a nominal Ag concentration of 10 mg kg−1 in the ART, BAFs from exposure to smaller (10-nm) Ag NPs were greater than those for the larger (30–50-nm) Ag NPs. The reverse was true at nominal 100 mg kg−1 in the same soil (Fig. 3B).

As with tissue concentrations, BAF values from Days 3 to 14 were pooled for analysis for the uptake exposures. A pattern similar to that observed for the

28-d exposures, with AgNO3 producing greater BAFs and exposures in the ART typically producing lower BAFs (Fig. 5B). In three different exposures (all in the YSL soil), BAFs for the smaller Ag NPs were significantly greater than those for the larger Ag NPs. The reverse was true for one exposure concentration in the ART (Fig. 5B).

diSCUSSiOnOxidation of Silver nanoparticles in Soil

The EXAFS data collected for this study indicate that little oxidation of Ag0 occurred as a result of aging in the test soils, particularly in the ART, as the percentage of Ag0 was similar to the particles exposed to air alone. Oxidation of Ag in the YSL was slightly greater than in the ART, where 5 to 15% more of the Ag0 was oxidized during 28 d than in the ART. This may indicate that soil type plays an important role in determining the rate of Ag NP oxidation. This agrees with the findings of Scheckel et al. (2010), who reacted Ag NPs with kaolin in electrolyte-saturated suspensions. They found that while Ag NPs did not oxidize at all during 18 mo when the electrolyte solution was saturated with NaNO3, exposure to NaCl caused gradual oxidation during 12 mo (Scheckel et al., 2010). While their experimental setup was quite different from that of the present study, it does indicate that Ag NP oxidation may only occur under specific conditions. Future studies using different Ag NPs or soil types should account for Ag oxidation as a factor that may influence the toxicity or accumulation of Ag from Ag NPs.

Toxicity of Silver and Silver nanoparticles to Eisenia fetida

This demonstrated that AgNO3 is more toxic to E. fetida than equivalent amounts of Ag in PVP-coated NP form. It is also consistent with a majority of aquatic toxicity studies that have compared the toxicity of ionic and NP Ag to a variety of species (Griffitt et al., 2008; Navarro et al., 2008; Bar-Ilan et al., 2009; Kvitek et al., 2009; Panacek et al., 2009). To our knowledge, however, no other study has compared ionic Ag and Ag NP toxicity in soils.

Table 2. Least-squares combination fitting speciation analysis results of Ag K-edge x-ray absorption fine structure spectroscopy of two sizes of poly-vinylpyrolidone coated Ag nanoparticles (NPs). The Ag NPs were analyzed directly from their packaging or after aging for 28 d in artificial soil (ART) and natural Yeager sandy loam (YSL). No oxidation occurred in the ART but slight oxidation occurred in the YSL; c2 and reduced c2 values are provided as a means to quantitatively compare the relative goodness-of-fit among samples.

Ag nP Soil type Ag0 Ag2O c2 Reduced c2

———— % ————Small Ag NPs (10 nm)

none 89.8 ± 0.4 11.2 ± 0.4 13 0.047

ART 90 ± 1 11 ± 4 52.70 0.280

YSL 84 ± 2 15 ± 2 125.3 0.602

Large Ag NPs (30–50 nm)

none 86 ± 3 14 ± 3 6.656 0.0257

ART 89 ± 3 11 ± 3 430.5 2.290YSL 74 ± 7 – 1922 11.37

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It is worth noting that significant Ag NP toxicity observed in our study occurred at concentrations that were 10 times greater than the concentrations at which toxicity was observed in ionic Ag treatments. This difference roughly corresponds with the amount of Ag that EXAFS analysis indicated was oxidized from the Ag NPs during a 28-d period. Although EXAFS did not directly measure dissolved Ag, this is the maximum percentage of Ag that we would expect to be available for dissolution. This suggests that Ag ions play an important role in the Ag uptake and toxicity observed for Ag NPs in this study. Other studies

of earthworm uptake of trace metals suggest that dissolved ions absorbed through dermal exposure constitute the principle route of uptake and, therefore, toxicity (Spurgeon et al., 2006).

The effects of Soil Composition on Silver and Silver nanoparticle Toxicity and Bioaccumulation

In this study, two different soils were used to compare the effects of soil composition on Ag and Ag NP toxicity. Silver nitrate added to a naturally occurring sandy loam soil produced toxic effects at lower concentrations than in a standardized

Fig. 2. extended x-ray absorption fine structure spectroscopy (eXAFS) spectra (solid black line) linear combination fitting results (gray line), with the uncertainty for the last digit in parentheses, for 10-nm polyvinyl pyrolidone (PVP) coated Ag nanoparticles (nPs) (A) in their pure form, and aged for 28 d in (B) artificial soil and (C) natural sandy loam; results for 30- to 50-nm PVP-coated Ag nPs aged for 28 d in (d) artificial soil and (e) natural sandy loam. Fitted models are a combination of spectra from Ag0 (dashed line) and Ag2O (dotted line) standards.

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artificial soil, thus indicating greater toxicity due to the properties of the sandy loam. This result agrees with the previous findings of Nahmani et al. (2007a). In their study of contaminated soils collected in the field, they found that the strongest regressions for E. fetida mortality included the Ag content of the soil and the sand content. Their strongest regression for cocoon production included Ag, Cd, and Pb in the soil, as well as soil pH (Nahmani et al., 2007a). Together, their findings indicate that increased Ag, larger proportions of sand, and increased soil acidity lead to increased toxicity. Likewise, our study showed greater toxicity in a sandier, more acidic soil. Similar relationships have been noted for other metals such as Cu, with its toxicity increasing in artificial soils with lower pH, lower organic C contents, and lower clay contents (Daoust et al., 2006). Organic C and pH have also been shown to be the primary soil characteristics

affecting chronic Zn toxicity in Eisenia andrei (Lumbricidae, Bouché, 1972), a species closely related to E. fetida (Rombke et al., 2006). The toxicity of Zn and bioaccumulation of Cd have been shown to strongly correlate with soil pH (Spurgeon et al., 2006).

Differences in subchronic toxicity may also be related to the microbial communities of the two soils, which would have influenced earthworm nutrition and feeding (Curry and Schmidt, 2006). According to fatty acid methyl ester (FAME) analysis of the two dry soils, the YSL had concentrations of most markers of microbial mass that were significantly greater than the concentrations found in the ART. We have observed no significant effects on microbial communities after 48 h of exposure to 10 mg kg−1 AgNO3 or Ag NPs (Shoults-Wilson et al., unpublished data, 2010). Both Ag and Ag NPs exhibit well-documented antimicrobial effects (Morones et al., 2005; Kvitek et al.,

2009; Panacek et al., 2009) that may have affected microbial populations differently in each soil; however, FAME biomarkers have not been measured at the exposure concentrations used in this study for the same durations. Therefore, we can’t rule out that the effects of Ag on microbial activity influenced earthworm fecundity differently in the two soil types.

The bioaccumulation of Ag in this study was also influenced by the type of soil used for the exposure. Because no adverse effects were observed at the low Ag NP concentrations tested for the YSL in this study, we were not able to compare toxicity among soil types; however, we expect greater toxicity in the YSL given the increased uptake of Ag relative to the ART. For both ionic Ag and Ag NPs, E. fetida in the sandy loam soil consistently accumulated significantly higher concentrations of Ag than

Table 3. Growth and survival endpoints for 28-d subchronic and reproductive toxicity tests. Control values are given for each soil type.

Ag formParticle

sizeSoil

type†Ag conc. in soil Growth Survival to day 28

nm mg kg−1 g %None

NA‡ART 0.0 0.197 ± 0.045§ 93.0

YSL 0.0 0.530 ± 0.065 100

AgNO3NA

ART8.378 ± 1.257 0.313 ± 0.021 96.7

94.12 ± 3.21 0.140 ± 0.015 96.7

YSL 7.413 ± 0.015 0.409 ± 0.038* 100

Nanoparticles

10ART

5.669 ± 0.681 0.176 ± 0.040 85.0

84.15 ± 2.54 0.294 ± 0.050 96.7

801.0 ± 74.9 0.270 ± 0.007 100*

YSL 8.033 ± 1.062 0.442 ± 0.065 100

30–50ART

9.331 ± 0.873 0.271 ± 0.021 85.0*

79.45 ± 1.02 0.224 ± 0.030 89.3

773.3 ± 11.2 0.264 ± 0.035 100*YSL 8.556 ± 1.783 0.485 ± 0.028 93.3

* Significant (n = 3, a = 0.05) difference from the control.† ART, artificial soil; YSL, Yeager sandy loam soil.‡ NA, not applicable. § Values are given as mean ± standard error.

Table 4. Reproductive endpoints for 28-d reproductive toxicity tests. Control values are given for each soil type.

Ag form Particle size Soil type† Ag conc. in soil Cocoons and worms Cocoons hatched Juveniles and worms

nm mg kg−1 %None NA‡ ART 0.0 3.84 ± 0.34§ 59.7 ± 2.2 0

YSL 0.0 6.50 ± 0.69 88.5 ± 5.8 35.5 ± 2.8

AgNO3 NA ART 8.378 ± 1.257 5.20 ± 0.36 75.8 ± 4.6 0

94.12 ± 3.21 0.41 ± 0.25* 31.5 ± 15.8 0

YSL 7.413 ± 0.015 5.47 ± 0.09 51.7 ± 2.9* 16.9 ± 3.6*

Nanoparticles 10 ART 5.669 ± 0.681 4.40 ± 0.70 58.5 ± 18.0 0

84.15 ± 2.54 5.00 ± 0.96 53.8 ± 9.3 0

801.0 ± 74.9 2.20 ± 0.42* 65.5 ± 5.2 0

YSL 8.033 ± 1.062 5.60 ± 0.84 87.8 ± 2.1 34.3 ± 2.1

30–50 ART 9.331 ± 0.873 4.30 ± 0.20 61.4 ± 5.3 0

79.45 ± 1.02 3.92 ± 0.19 61.1 ± 4.2 0

773.3 ± 11.2 1.60 ± 0.15* 53.7 ± 2.3 0

YSL 8.556 ± 1.783 6.23 ± 0.79 78.4 ± 3.5 32.4 ± 2.6* Significant (n = 3, a = 0.05) difference from the control.† ART, artificial soil; YSL, Yeager sandy loam soil.‡ NA, not applicable.§ Values are given as mean ± standard error.

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those exposed to similar concentrations in the artificial soil. Because the sandy loam had much lower CEC, organic matter, clay content, and pH, it seems reasonable to conclude that Ag ions were more available for uptake in the YSL than the ART. Again, this is supported by the current understanding of dermal exposure to available ions as the primary route for trace metal uptake in earthworms (Spurgeon et al., 2006). Some recent studies have indicated, however, that while this may be true for some metals such as Cd, Cu, and Pb, a significant (estimated 18–31%) amount of Zn is accumulated via the gut (Saxe et al., 2001; Vijver et al., 2003). This may constitute a route of Ag NP uptake and accumulation of whole particles rather than just the dissolved fraction.

The BAFs for Ag described by Nahmani et al. (2009) were similar to those from this study, ranging from 0.01 to 3.63 when comparing E. fetida tissue concentrations to bulk soil concentrations. That we found BAF values for the same element only in the low end of that range is not surprising because the previous study investigated a wide variety of soils and did not feed the worms as we did (Nahmani et al., 2009). The BAFs that have been calculated for the substrate concentrations of other metals are typically larger than those we found for Ag. These include 0.499 (Cu), 0.191 (Fe), 0.266 (Zn), 0.502 (Pb) (Suthar and Singh, 2009), and 0.6 to 3.1 (Hg) (Burton et al., 2006). Suthar and Singh (2009) found that different proportions of cow manure and sewage sludge produced different BAFs for Cu, Fe, Pb, and Zn but concluded that this was more related to increasing exposure to those metals via the sludge than to changes in bedding composition. It appears that BAFs are also affected by soil properties and therefore it is difficult to compare them between separate studies. Petersen et al. (2008) estimated that for E. fetida, a BAF that is not greater than 0.0315 ± 0.001 cannot be assumed to represent actual accumulation owing to retention of soil in the earthworm gut. In our study, retained soil could account for the unpredictable nature of tissue concentrations seen in the 14-d uptake experiments, especially considering that most Ag NP exposures did not have BAF values >0.0315. It therefore suggests that in some treatments, primarily those performed in the ART and at high Ag concentrations, the observed accumulation may be the result of incomplete gut clearance of soil.

In this study, almost 90% of the tissue concentration from animals exposed to AgNO3 for 28 d in the YSL was attained by animals similarly exposed to Ag NPs. Calculated BAFs for Ag NPs were also higher than the amount estimated to result primarily from incomplete gut clearance of soil (Petersen et al., 2008). According to the EXAFS data, the particles were, at most, 15% oxidized, indicating that oxidized and potentially dissolved Ag cannot account for all of the accumulation. Unrine et al. (2008, 2010) have previously demonstrated that Au and Cu nanoparticles can be taken up from the soil into earthworm tissues. The spectroscopic techniques used in those studies to demonstrate the presence of particles in tissues were not applicable to Ag NP in earthworm tissues in the present

study because of the low concentrations of Ag that were present and high detection limits for Ag NPs relative to Au and Cu. More work needs to be undertaken to determine whether Ag NPs themselves are taken up by earthworms or whether the accumulation is primarily from dissolved Ag. Regardless, as described above, it appears that toxicity is primarily related to the oxidized fraction present as Ag(I) in the soil, which would represent the maximum amount of Ag available for dissolution.

effects of nanoparticle Size on Toxicity and Bioaccumulation

We observed no difference in toxicity between two Ag NPs of different sizes. Despite the fact that nanoparticle size has a potentially large impact on physicochemical and biological properties, little has been done to investigate the relative toxicity of Ag NPs with different sizes. Most work on the potential size-dependence of NP toxicity has focused on cytotoxicity, primarily in tissue cultures under simplified conditions. Rat alveolar

Fig. 3. Bioaccumulation of Ag after 28 d by Eisenia fetida exposed to three sources of Ag: AgnO3, small (10-nm) Ag nanoparticles (S Ag nP), and large (30–50-nm) Ag nanoparticles (L Ag nP). Bioaccumulation was measured as (A) tissue concentrations of Ag and (B) bioaccumulation factors (BAFs). All bars represent mean values ± standard error. Bars labeled with the same letter are not significantly different (n = 3–6, a = 0.05). Bars are grouped by soil type (ART, artificial soil; YSL, Yeager sandy loam) and nominal exposure concentration (e.g., 10 = 10 mg Ag kg−1 dry soil).

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macrophages exposed to three different sizes (10, 30, and 55 nm) of hydrocarbon-coated Ag NPs demonstrated different levels of cytotoxicity (Carlson et al., 2008; Auffan et al., 2009). Smaller Ag NPs significantly decreased mitochondrial function, cell integrity, mitochondrial membrane potential, and glutathione content of cells, while increasing reactive O species (ROS) at lower concentrations than larger particles. A study utilizing uncoated Ag NPs found that smaller (2–4-nm) Ag NPs caused more cytotoxicity in macrophage cells than the medium (5–7-nm) and larger (20–40-nm) particles studied (Yen et al., 2009). Antimicrobial endpoints have also shown that Ag NP size can affect cellular toxicity. Choi and Hu (2008) found that inhibition of nitrifying bacteria was correlated with the fraction of Ag NPs <5 nm but not mean particle size. Inhibition was correlated to an increase in intracellular ROS but size-dependent ROS generation was not explicitly determined. Embryonic zebrafish (Danio rerio; Cyprinidae, Hamilton, 1822) exposed to a range of Ag NP sizes (3-, 10-, 50-, and 100-nm mean diameter) exhibited significantly higher mortality and sublethal defects when exposed to smaller particles than larger particles (Bar-Ilan et al., 2009). However, the researchers were unable to rule out dissolved Ag ions as the primary driver of toxicity. For example, smaller particles may have a higher inherent rate of dissolution.

Previous research has also implicated particle size as a factor in the toxicity of other inorganic nanoparticles. One study found significant differences in cytotoxicity between different sizes of silica nanoparticles but the relationship was dependent on the cell lines tested (Petushkov et al., 2009). Another study using murine “macrophage-like” cells reported size-dependent uptake of polyethylene-coated CdTe–CdSe quantum dots, with smaller (20-nm) particles taken up more readily than larger (200-nm) ones (Clift et al., 2008). Titanium oxide nanoparticles have also shown increased antimicrobial properties at smaller (<25-nm) sizes (Simon-Deckers et al., 2009). Neither size of particle, however, caused detectable cytotoxicity. In contrast, Ni NPs of various sizes (30, 60, and 100 nm) showed no difference in toxicity toward D. rerio embryos (Ispas et al., 2009).

One obvious trend in previous studies investigating size as a factor in NP toxicity is that most studies have been performed using cell and bacterial cultures. Except for a single study using embryonic zebrafish (Bar-Ilan et al., 2009), Ag NPs have not been shown to cause toxicity in multicellular organisms in a size-dependent manner. Our study indicates that under environmentally realistic soil conditions where aggregation and dissolution may play a role, primary particle size (in the size range of the study materials) is not an important factor, except as it may relate to particle oxidation and dissolution. The findings of the present study also suggest that dissolved Ag is primarily responsible for Ag NP toxicity to earthworms for the ecologically relevant endpoints that we examined. Although particle size could affect the rate of Ag dissolution by increasing surface area, it did not appear to cause a difference in the present study. Another explanation

Fig. 4. Bioaccumulation of Ag at different time points during 14 d by Eisenia fetida exposed to three sources of Ag: AgnO3, small (10-nm) Ag nanoparticles (S Ag nP), and large (30–50-nm) Ag nanoparticles (L Ag nP), demonstrating that little consistent uptake of Ag occurred after 24 h, regardless of treatment. each data series represents one Ag source at one nominal exposure concentration. The series are classified by soil, with exposure soils being (A) Yeager sandy loam (YSL), (B) YSL amended with CaCO3 (YSL+), and (C) artificial soil (ART).

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may be that size-dependent toxicity only exists between particles of certain size classes. For instance, the smallest sizes used by studies that demonstrated size-dependent toxicity have had a mean size £10 nm (Carlson et al., 2008; Choi and Hu, 2008; Bar-Ilan et al., 2009; Yen et al., 2009). Below 10 nm, the surface area of particles increases exponentially with respect to volume, meaning that equivalent doses will have much larger surface areas, which can play an important factor in ROS generation and other toxic mechanisms (Carlson et al., 2008). The smaller sized material used in this study did not have a majority of particles £10 nm and therefore differences in surface areas and any associated toxicity would have been negligible.

In this study, tissue concentrations and BAFs of Ag in worms exposed to larger Ag NPs were in some cases significantly different than those calculated for worms exposed to smaller NPs. This was especially true for BAFs, which took into account the fact that test soil concentrations may be different between two exposures with the same nominal concentration. Significantly greater BAFs were frequently found for worms exposed to the

larger Ag NPs, indicating that the earthworms accumulated greater amounts of Ag when exposed to the larger particles. The BAFs calculated for the Ag NPs varied among treatments, however, and in some cases exposure to the smaller particles resulted in significantly greater BAFs. Because this trend was not consistent, no conclusions can be drawn about the role particle size plays in the bioaccumulation of Ag by earthworms exposed to Ag NPs. More research is required to determine whether this is the case and what, if any, primary particle size may influence the bioaccumulation of Ag from NPs.

COnCLUSiOnSThis study provides evidence that Ag NP toxicity in

earthworms is primarily related to the release of Ag ions into the soil solution. Second, soil properties affect the uptake of Ag from both ionic and NP sources. Finally, while no evidence of size-dependent NP toxicity was found, it cannot be ruled out that Ag NP size may be a factor in the bioaccumulation of Ag or in the dissolution of Ag nanoparticles. More studies into the factors affecting the toxicity and bioaccumulation of Ag and Ag NPs need to be performed in soil systems so that we can better understand the potential impacts of Ag NP release into the environment. Of particular importance will be the long-term accumulation, aging, and dissolution of Ag NPs in soil as well as surface functionalization.

ACknOWLedGMenTSWe wish to acknowledge the assistance of M. Lacey, E. Harding, M. Engle, and O. Zhurbich. Major funding for this research was provided by the USEPA through Science to Achieve Results Grant no. RD 833335. This material is also based in part on work supported by the National Science Foundation (NSF) and the USEPA under NSF Cooperative Agreement EF-0830093, Center for the Environmental

Fig. 5. earthworm (A) tissue concentrations and (B) bioaccumulation factors (BAFs) of Ag pooled across 3 to 14 d of exposure from three sources of Ag: AgnO3, small (10-nm) Ag nanoparticles (S Ag nP), and large (30–50-nm) Ag nanoparticles (L Ag nP). All bars represent mean values ± standard error. Bars marked with the same letter are not significantly different (n = 3, a = 0.05). Bars are grouped by soil type (ART, artificial soil; YSL, Yeager sandy loam soil; YSL+, Yeager sandy loam amended with CaCO3) and nominal exposure concentration (e.g., 25 = 25 mg Ag kg−1 dry soil).

Table 5. The nominal and actual concentrations of Ag in each exposure soil during the 14-d uptake experiments.

Ag form Particle size Soil type† nominal Ag Actual Ag

nm ————— mg kg−1 ————AgNO3 NA‡ ART 25 23.73 ± 1.06 a§

YSL+ 25 23.14 ± 2.14 a

YSL 100 87.75 ± 3.13 bc

Nanoparticles 10 ART 25 17.92 ± 0.48 a

100 80.09 ± 4.01 bc

YSL+ 25 18.85 ± 0.82 a

100 83.21 ± 13.85 bc

YSL 100 118.5 ± 52.8 b

500 479.2 ± 47.3 d

30–50 ART 25 19.98 ± 1.15 a

100 78.12 ± 2.82 bc

YSL+ 25 18.71 ± 1.40 a

100 87.55 ± 1.38 bc

YSL 100 61.59 ± 9.90 c500 288.1 ± 38.5 e

† ART, artificial soil; YSL, Yeager sandy loam soil; YSL+, Yeager sandy loam amended with CaCO3.‡ NA, not applicable.§ Values followed by the same letter are not significantly different (n = 3–4, a = 0.05).

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Implications of NanoTechnology (CEINT), and NSF-BES0608646 and USEPA R833326. Any opinions, findings, conclusions or recommendations expressed in this material are those of the author(s) and do not necessarily reflect the views of the NSF or the USEPA. This work has not been subjected to USEPA review and no official endorsement should be inferred. Portions of this research were carried out at the Stanford Synchrotron Radiation Laboratory (SSRL), a national user facility operated by Stanford University on behalf of the U.S. Department of Energy, Office of Basic Energy Sciences. We thank Dr. C. Kim, J. Dale, B. Deeb, and A. Capranico of Chapman University for sample preparation and analysis support. Thanks to the staff and beamline scientists at SSRL.

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