Nanotoxicology, 2010; Early Online, 1–13
Effect of silver nanoparticle surface coating on bioaccumulation and
reproductive toxicity in earthworms (Eisenia fetida)
WILLIAM A. SHOULTS-WILSON1, BRIAN C. REINSCH2, OLGA V. TSYUSKO1,
PAUL M. BERTSCH1, GREGORY V. LOWRY2, & JASON M. UNRINE1
1
Department of Plant and Soil Sciences, University of Kentucky, Lexington, KY, and 2Department of Civil and
Environmental Engineering, Carnegie Mellon University, Pittsburg, PA, USA
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(Received 11 August 2010; accepted 29 October 2010)
Abstract
The purpose of this study was to investigate the effect of surface coating on the toxicity of silver nanoparticles (Ag NPs) soil.
Earthworms (Eisenia fetida) were exposed to AgNO3 and Ag NPs with similar size ranges coated with either polyvinylpyrrolidone (hydrophilic) or oleic acid (amphiphilic) during a standard sub-chronic reproduction toxicity test. No significant effects
on growth or mortality were observed within any of the test treatments. Significant decreases in reproduction were seen in
earthworms exposed to AgNO3, (94.21 mg kg-1) as well as earthworms exposed to Ag NPs with either coating (727.6 mg kg-1
for oleic acid and 773.3 mg kg-1 for polyvinylpyrrolidone). The concentrations of Ag NPs at which effects were observed are
much higher than predicted concentrations of Ag NPs in sewage sludge amended soils; however, the concentrations at which
adverse effects of AgNO3 were observed are similar to the highest concentrations of Ag presently observed in sewage sludge in
the United States. Earthworms accumulated Ag in a concentration-dependent manner from all Ag sources, with more Ag
accumulating in tissues from AgNO3 compared to earthorms exposed to equivalent concentrations of Ag NPs. No differences
were observed in Ag accumulation or toxicity between earthworms exposed to Ag NPs with polyvinylpyrrolidone or oleic acid
coatings.
Keywords: Eisenia fetida, silver nanoparticles, reproductive toxicity, bioaccumulation, X-ray absorption spectroscopy
Introduction
Silver nanoparticles (Ag NPs) have become an integral component in a wide array of consumer products
(Klaine et al. 2008; Luoma 2008). Products currently
being produced that utilize Ag NPs include medical
devices, food-storage containers, soaps and sanitizers,
wound dressings, and fabrics; Ag NPs are frequently
utilized in polymeric, colloidal, spun and powder
forms (Luoma 2008). With greater Ag NP production
and incorporation into consumer products, it is
expected that Ag NPs will provide an increasing
contribution of anthropogenic Ag into the environment (Blaser et al. 2008). Much of the Ag from
fabrics is expected to enter the wastewater stream
and eventually treatment plants in dissolved or
nano-particulate form (Benn and Westerhoff 2008)
where an estimated 90% will be removed from the
waste stream and partitioned to sewage sludge
(Blaser et al. 2008; Mueller and Nowack 2008;
Gottschalk et al. 2009). The sludge-associated Ag
NPs may be disposed of through incineration, landfilling or application to terrestrial environments as
biosolids, depending on the prevailing practices of
the country or region (Mueller and Nowack 2008). In
the United States and much of Europe, approximately
60% of sewage sludge is applied to agricultural lands
as biosolids (United States Environmental Protection
Agency [USEPA] 2009). Thus, this represents a
potential exposure pathway to terrestrial organisms
of high significance.
Despite increased production of Ag NPs, relatively
little is known about their environmental fate and
potential effects, particularly in terrestrial environments (Klaine et al. 2008; Unrine et al. 2008). The
toxicity of Ag NPs has thus far been mainly studied in
the aquatic environment, with toxicity demonstrated
for 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.
Correspondence: Dr Jason Unrine, Department of Plant and Soil Sciences, University of Kentucky, N212-N Agricultural Science Center North, Lexington, KY
40546, USA. Tel: +1 859 257 1657. Fax: +1 859 257 3655. E-mail: [email protected]
ISSN 1743-5390 print/ISSN 1743-5404 online 2010 Informa UK, Ltd.
DOI: 10.3109/17435390.2010.537382
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2
W. A. Shoults-Wilson et al.
2009), algae (Griffitt et al. 2008; Navarro et al. 2008),
daphnia (Griffitt et al. 2008), Japanese medaka
(Chae et al. 2009), fathead minnow embryos
(Laban et al. 2010), and various zebrafish life stages
(Lee et al. 2007; Asharani et al. 2008; Griffitt et al.
2008; Yeo and Kang 2008; Bar-Ilan et al. 2009). In
the terrestrial environment, Ag NPs have not been
investigated to such an extent and Ag toxicity itself
is not well studied (Ratte 1999; Nahmani et al.
2007b). The soil nematode Caenorhabditis elegans
(Rhabditidae, Maupas 1900) has been used to assess
Ag NP toxicity but these experiments were performed
in K-medium (NaCl and KCl solution) rather than in
soils (Roh et al. 2009). To address the need for an
understanding of the terrestrial toxicity of Ag NPs, the
current study investigated the toxicity of Ag NPs in
soil to a model soil organism, the epigeic earthworm
Eisenia fetida (Lumbricidae, Savigny 1826).
Another aspect of Ag NP toxicity that has not been
extensively examined is the role that the NP surface
chemistry (coating) plays in toxicity. Several studies
have shown differences in toxicity (Griffitt et al. 2008;
Navarro et al. 2008; Bar-Ilan et al. 2009; Kvitek et al.
2009; Panacek et al. 2009) and transcriptomic
responses (Roh et al. 2009) between Ag NP and
AgNO3, with ionic forms typically exhibiting greater
toxicity. However, Ag NPs have different surface
chemistries depending on coating or functionalization, which may alter their toxicity and/or mode of
action. For instance, Ahamed et al. (2008) found that
Ag NP particles coated with polysaccharides
caused greater damage to mammalian cell lines
than similarly-sized Ag NP that were uncoated. Previous studies also suggest that Ag NPs stabilized with
surfactants cause greater toxicity to microorganisms
than unmodified Ag NPs (Kvitek et al. 2009;
Panacek et al. 2009). In contrast, zerovalent iron
nanoparticles have been shown to be less toxic to
bacteria when coated with polymers or natural organic
matter (Li et al. 2010).
It is unclear if these differences in surface chemistry
are important determinants of toxicity in environmentally realistic exposure scenarios, such as in soil. For
example, it is widely acknowledged that naturally
occurring colloids in soil environments are coated
with substances such as iron oxohydroxides and natural organic matter (Bertsch and Seaman 1999).
Therefore, engineered surface coatings may be extensively modified in the natural environment, and the
properties of pristine coatings may become less
important in terms of bioavailability and toxicity.
Furthermore, if toxicity is primarily related to the
release of free Ag ions, then surface coatings may
only be important as they relate to oxidation and
dissolution.
The objectives of this study were: (i) To determine
concentrations of Ag (both ionic and nanoparticulate)
that caused adverse effects on E. fetida growth, mortality and reproduction, and (ii) to determine whether
Ag NP toxicity could be influenced by surface chemistry. In order to achieve this, we performed standardized reproductive toxicity tests (Organisation for
Economic Cooperation and Development [OECD]
2004) on E. fetida exposed to various concentrations
of ionic Ag and Ag NPs in artificial soil media. The Ag
NPs were coated with either polyvinyl pyrolidone
(PVP) or oleic acid (OA). The PVP-coated Ag NPs
are hydrophilic, primarily forming stable suspensions
in polar solvents (Yan et al. 2009), while OAcoated Ag NPs are amphiphilic, forming stable suspensions in polar solvents, non-polar solvents or
polar/non-polar interface layers depending on pH
(Li et al. 2002; Seo et al. 2008). The log Kow values
of PVP and OA are 3.4 and 7.73, respectively
(BASF 2008; International Programme on Chemical
Safety [IPCS] 2010; in press). Therefore, comparing
the two should provide an understanding of the
importance of surface chemistry in Ag NP toxicity
to E. fetida under an environmentally relevant exposure scenario.
Materials and methods
Material characterization
Silver powders, each with a stated nominal size
distributions of 30–50 nm, with polyvinylpyrolidone
(PVP) and oleic acid (OA) coatings were purchased
from NanoAmor (Nanostructured & Amorphous
Materials, Inc; Houston, TX, USA). The powders
were both high purity (>95% Ag), which we verified
by dissolving in concentrated, ultra-high purity
HNO3 and analyzing for metallic impurities using
inductively coupled plasma mass spectrometry (ICPMS; see description of methods below). Primary
particle diameter distributions were determined
using transmission electron microscopy (TEM) as
described in detail in (Shoults-Wilson et al. 2010;
in press). A Philips BioTWIN 12 Transmission Electron Microscope (FEI Company; Hillsboro, OR,
USA) was used to examine the particles, at
23,000–6,800 magnification and an accelerating
potential of 100 kV. Particle diameter 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,
Japan). The hydrodynamic radii of the particles in
stock suspensions (in deionized water) used to dose
the soils were determined by dynamic light scattering
Effects of silver nanoparticles on earthworms
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(DLS) using a Malvern Zeta-Sizer Nano-ZS (Malvern Instruments; Worcestershire, UK). Intensity
weighted distributions were transformed to volume
weighted distributions using the complex refractive
index components 0.135 (real) and 3.99 (imaginary).
The same instrument was used to determine apparent zeta potential in stock suspensions using phase
analysis light scattering (PALS). The mass weight
percentage of surface coating material was determined using thermogravimetric analysis (TGA;
SDTQ600, TA Instruments, New Castle, DE,
USA). A correction for moisture content was
made by determining mass loss upon heating to
120 C. The mass loss due to the coatings was determined after heating to 700 C.
Characterization of Ag speciation in soil
X-ray absorption spectroscopy (XAS; specifically
extended X-ray absorption fine structure, or EXAFS)
was conducted at Stanford Synchrotron Radiation
Laboratory (SSRL; Menlo Park, CA USA) wiggler
magnet beamline 11-2. Compounds and methods
used to perform XAS analysis of Ag NPs in soil
have been previously described in detail (Reinsch
et al. 2010; Shoults-Wilson et al. 2010). Briefly, Ag
K-edge (25.514 keV) XAS spectra were collected over
the energy range 25.284–26.635 keV at room temperature. Data were analyzed using SixPACK software version 0.67 (Webb 2005). 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 then averaging multiple scans. Background subtraction was performed using SixPACK with E0
defined as 25.530 keV, and R-background = 1.0 Å.
The resulting spectra were converted to frequency (k)
space and weighted by k3 (Webb 2005). The Linear
least-squares combination fitting (LCF) procedure
employed was adapted from Kim et al. (2000)
and Reinsch et al. (2010) whereby only components
that contributed significantly (a > 10% decrease in
fitting correlation based on reduced Chi squared
value) to the fit were included.
EXAFS spectra were collected for a library of
model compounds including, Ag0, Ag2S, AgO,
Ag2O, AgPO4, AgCO3, AgSO4, and AgCH3CO2
and were used in LCF analysis. To help constrain
the possible number of Ag species, we also analyzed
both stock Ag NP powders for total S content using a
high temperature combustion sulfur analyzer. The
method was verified when we observed excellent
agreement between theoretical and measured values
for high purity Ag2S and AgSO4 samples.
3
Sub-chronic and reproductive toxicity test
The design of the earthworm reproduction, bioaccumulation and subchronic lethality assays followed
published guidelines (OECD 2004). The artificial
soil medium consisted of 69.58% dry mass quartz
sand, 0.43% dry mass crushed limestone, 20% dry
mass kaolin clay, and 10% dry mass sphagnum peat
moss (sieved to < 2 mm). We selected a range of Ag
NP exposure concentrations based on range finding
studies where citrate coated Ag NPs did not cause
significant mortality in E. fetida at concentrations as
high as 40 mg kg-1 (Unrine et al. unpublished data).
To treat dried soils, NPs were first suspended in
18.2 MW DI water through sonication. We sonicated
1 L of 4 g L-1 of the particles in an ultrasonic bath in
1 L glass media bottles FS20 Ultrasonic Cleaner,
Fisher Scientific, Fairlawn, NJ, USA). The resultant
suspensions were applied to soils at a rate of 50% of
the moisture holding capacity (approximately 26% v/
w). The nominal concentrations of the soils were 10,
100 and 1000 mg Ag kg-1 dry soil. A separate treatment added dissolved AgNO3 to nominal concentrations of 10 and 100 mg Ag kg-1 dry soil. 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. Earthworm
masses ranged from 0.243–0.752 g and averaged
0.450 ± 0.101 g 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 to mimic a realistic diurnal cycle and facilitate
reproduction. After 24 h of exposure, 5 g of a 50/50
mixture of dehydrated alfalfa and commercially available chicken feed was added to the surface and
hydrated with 5 mL of DI water. This feed mixture
had previously been shown to allow rapid, sustained
growth in the earthworms. At the end of each week of
exposure, earthworms were observed for mortality
and another 4 g of feed mixture was spread on the
surface and hydrated. Container weights were monitored and water was added on a biweekly basis to
maintain weight. After 28 days of exposure, earthworms were observed for mortality, gently cleaned on
paper towels, massed and placed on wet filter paper
inside Petri dishes. The earthworms remained in the
dishes for 24 h to void their guts and were then
sacrificed. An additional 5 g of food and 5 mL of
water were added to the containers and they were
incubated for an additional 28 days, with weekly soil
moisture replenishment. The soil was manually
searched and cocoons were collected and counted.
All cocoons were immediately placed in water and
4
W. A. Shoults-Wilson et al.
whether or not they floated [a method of determining
hatching success (OECD 2004)] was recorded.
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ICP-MS analyses
To determine Ag concentration, whole earthworms
and ~ 0.25 g soil samples were digested in concentrated trace metal grade nitric acid using a
MARS Express microwave digestion system (CEM,
Matthews, NC, USA) according to U.S. EPA method
3052 (USEPA 1996). Total Ag concentrations in the
digestates were determined by ICP-MS following
U.S. EPA method 6020A (USEPA 1998), using an
Agilent 7500 CX (Agilent Technologies; Santa Clara,
CA, USA). Digestion sets included standard reference tissues (DOLT-4, National Research Council of
Canada; Ottawa, ON, Canada) or soils directly spiked
with a known mass of the OA and PVP Ag NPs.
Duplicate digestions and reagent blanks were also
included with each digestion set. Analytical runs
included duplicate dilutions, spike recovery samples,
and inter-calibration/cross-calibration verification
samples. The instrument was recalibrated after the
first 10 samples and every 20 samples thereafter.
Previous experiments demonstrated that the Ag
NPs dissolved in concentrated HNO3 at room temperature (‡ 95% recovery). Recovery of Ag after
directly adding Ag NP powder to soils in the digestion
vessels was 94.6 ± 2.1%. Mean recovery of tissue Ag
was 117.3% from DOLT-4. Spike recovery averaged
98.5% and the mean relative percentage difference
(RPD) between replicate subsamples of soil was
1.4%. Concentrations in earthworms exposed to
clean soil as controls did not exceed the method
quantification limit (MQL: 0.217 ng mL-1). Three
earthworms exposed to low (10 mg kg-1 nominal)
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 analyses
All data were log transformed before analysis to
improve normality and homogeneity of variance. Normality and homogeneity of variance of the data were
tested using Shapiro-Wilk’s test and Levene’s test,
respectively. Significant differences of a treatment on
growth, mortality, bioaccumulation and reproduction
between treatments and controls were tested using a
one-way analysis of variance (ANOVA) if normal and
homogenous. Post-hoc multiple comparisons for
ANOVA were performed using Duncan’s Multiple
Range test. Non-normal data were tested using the
Wilcoxon Rank-Sum test. Accumulation of Ag as a
function of Ag concentration in exposure soil was
modeled by transforming the data in order to produce
a linear relationship. Linear relationships were then
analyzed for significance (a = 0.05) using ANOVA.
Bioaccumulation factors (BAFs) were calculated by
dividing concentrations of Ag found in earthworm
tissues by concentrations in exposure soils and compared using ANOVA. All calculations were performed using Statistical Analysis Software (SAS
v9.1, SAS Institute, Cary NC, USA).
Results
Material characterization
Primary particle diameter was determined by measuring images of the particles taken using TEM. The
PVP coated particles had a slightly larger mean diameter (56.35 ± 1.16 nm) than the particles coated with
OA (50.60 ± 1.02 nm). The two particles had very
similar particle size distributions (Figure 1A, 1B) with
the majority of the particles within 10 nm of the
nominal (30–50 nm) size range. Dynamic light scattering (DLS) was used to determine the size distribution of particles in the stock suspensions that were
used to dose the exposure soils. This technique provided similar size distributions to TEM size analysis
for the PVP-coated particles (Figure 1A, 1C). However, the particles coated in OA had a greater percentage of aggregates in suspension which were as
large as 200 nm in diameter (Figure 1B, 1D). The
electrophoretic mobility of particles was found to be
1.9 ± 0.0 mM cm V-1 s-1 for the particles coated in
PVP and 2.4 ± 0.1 mM cm V-1 s-1 for particles coated
in OA. The apparent zeta potentials estimated using
the Hückel approximation (Fka = 1.5) were 35.9 ±
0.8 mV for the PVP-coated particles and 45.4 ±
1.1 mV for the OA-coated particles. No metallic
impurities were detected in the stock suspensions.
The coatings accounted for 6.4 and 1.5 wt % of
the mass of the PVP and OA powders, respectively.
Characterization of Ag speciation
Over the course of 28 d, Ag NPs did not appreciably
oxidize while in the artificial soil (Table I). Silver
nanoparticles exposed only to air were found to be
partially oxidized (Table I, Figure 2A, 2C), with
approximately 14% as Ag2O and 86% as Ag0 for
PVP NPs. For OA coated NPs, only the Ag0 component was statistically significant. Total S analysis
ruled out significant sulfidation of the particles since
5
Effects of silver nanoparticles on earthworms
10
60
20
40
10
20
0
10
20
30
40
50
60
70
80
0
90 100
20
0
0
10
70
100
D
60
40
30
40
% Volume
Cumulative %
% by Volume
50
20
10
0
0
20 40
60
40
50
60
70
80
0
90 100
100
70
60
80
Cumulative %
% by Volume
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60
20
30
Geometric diameter (nm)
Geometric diameter (nm)
C
20
80
50
60
40
30
40
% Volume
Cumulative %
20
10
0
0
80 100 120 140 160 180 200
0
20
40
60
Cumulative %
0
80
Cumulative %
40
Frequency
Cumulative %
30
Frequency
60
20
100
40
80
Frequency
Cumulative %
30
Frequency
B
100
40
Cumulative %
A
20
0
80 100 120 140 160 180 200
Hydrodynamic diameter (nm)
Hydrodynamic diameter (nm)
Figure 1. Histograms of size classes by frequency percent of total particle volume, as well as cumulative percentage of size classes as part of all
nanoparticles. The individual graphs demonstrate particles size of PVP-coated (A) and OA-coated (B) particles counted using transmission
electron microscope; PVP-coated (C) and OA-coated (D) particle sizes calculated using dynamic light scattering.
the PVP Ag NP sample contained only 0.009% S and
the OA Ag NP sample only contained 0.011% S.
Approximately 20% of the initial OA coated particle
spectrum (Figure 2C) was left as unidentified oxidized Ag, because after determining the Ag0 foil was
the initial and dominant component no other model
compound, when applied with Ag0, lowered the
reduced chi squared value of the overall fit below
10% of the reduced chi squared-value of the fit that
only included Ag0. After aging for 28 d in soil, similar
speciation was observed with the PVP Ag NP soil
containing 89% Ag0, and the OA Ag NP soil
containing approximately the same percentage of
Ag0 as the particles exposed to air alone.
Growth and mortality
There were no significant differences in growth
between any treatments and the control (Table II).
There were no significant differences in growth
between E. fetida exposed to PVP- or OA-coated
Ag NPs. Significant increase in mortality was only
found in one treatment (9.331 ± 0.873 mg kg-1 PVP
Table I. A summary of EXAFS analysis of the chemical form of two Ag NPs with different surface coatings. Ag NPs were analyzed directly
from their packaging (exposed only to air) or after aging for 28 d in artificial soil. No oxidation occurred in the soil. Chi2 and reduced Chi
squared (RC2) values are provided as a means to quantitatively compare relative goodness of fit between samples.
Coating
Exposure
%Ag0
%Ag2O
%Ag2S
Unidentified Ag species
Chi square
RC2
PVP
Air
86 ± 3%
14 ± 3%
ND
ND
6.656
0.0257
(30–50 nm)
Soil
89 ± 3%
11 ± 3%
ND
ND
430.5
2.290
Oleic acid
Air
79.5 ± 0.4%
ND
ND
20.5% ± 0.4%
2.035
0.0078
(30–50 nm)
Soil
78.2 ± 0.5%
ND
ND
21.8% ± 0.5%
59.96
0.0362
ND, not determined.
6
W. A. Shoults-Wilson et al.
30
60
Data + Fit
B
25
50
20
40
k-3 c(k)
k-3 c(k)
A
Ag0 = 86 (3)%
15
10
30
Ag0 = 89 (3)%
20
5
0
Data + Fit
Ag2 O = 14 (3)%
4
6
10
8
12
Ag2 O = 11(3)%
10
14
0
16
4
6
10
12
k (Å )
k (Å )
40
Data + Fit
C
40
D
Data + Fit
k-3 c(k)
30
k-3 c(k)
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8
-1
-1
0
Ag = 79.5 (1)%
20
10
30
Ag0 =78.2 (8)%
20
10
4
6
8
10
12
4
-1
6
8
10
k (Å-1)
k (Å )
Figure 2. Extended X-ray absorption fine structure spectroscopy (EXAFS) spectra (solid black line) linear combination fitting results (grey
line), with the uncertainty for the last digit in parenthesis, for 30–50 nm PVP-coated silver nanoparticles (Ag NPs) exposed to air only (A), and
aged for 28 d in artificial soil (B); results for 30–50 nm oleic acid-coated Ag NPs exposed to air only (C) and aged for 28 d in artificial soil (D).
Fitted models are a combination of spectra from Ag0 (dashed line) and Ag2O (dotted line) standards. The fitted model in Panel C is dashed to
better show the underlying data. The quantitative fitting results are also displayed, including goodness of fit analyses, in Table III.
NPs at 28 d; Table III). This concentration was the
lowest concentration for that Ag NP and two concentrations which were one and two orders-ofmagnitude higher demonstrated no significant
mortality. Also, the majority of the mortality observed
occurred in only one replicate container. Therefore,
this increase in mortality is most likely random and
not the result of Ag toxicity.
Table II. Measurements of earthworm growth at different exposure concentrations expressed as mean values plus or minus standard error.
Ag form
Surface coating
None
NA
Ionic
NA
NP
NP
PVP
OA
[Ag]soil (mg kg-1)
Initial mass (g)
Final mass (g)
Growth (g)
0
0.477 ± 0.014
0.674 ± 0.053
0.197 ± 0.045
8.38 ± 1.26
0.454 ± 0.007
0.766 ± 0.024
0.313 ± 0.021
94.12 ± 3.21
0.438 ± 0.028
0.578 ± 0.028
0.140 ± 0.015
9.33 ± 0.87
0.417 ± 0.011
0.688 ± 0.032
0.271 ± 0.021
79.45 ± 1.02
0.494 ± 0.026
0.718 ± 0.052
0.224 ± 0.030
773.3 ± 11.2
0.416 ± 0.011
0.680 ± 0.037
0.264 ± 0.035
8.03 ± 0.30
0.457 ± 0.031
0.720 ± 0.010
0.264 ± 0.021
81.62 ± 2.90
0.446 ± 0.010
0.683 ± 0.020
0.238 ± 0.031
727.6 ± 34.0
0.430 ± 0.019
0.693 ± 0.035
0.262 ± 0.023
NP, Nanoparticle; PVP, Polyvinylpyrolidone; OA, Oleic acid.
7
Effects of silver nanoparticles on earthworms
Table III. Mortality of E. fetid a over the course of the experiment. Values that are significantly different (p < 0.05) than that of the control are
marked by *.
Ag form
Surface coating
None
NA
Ionic
NA
NP
[Ag]soil (mg kg-1)
0
PVP
NP
OA
% Survival
Day 7
% Survival
Day 14
% Survival
Day 21
% Survival
Day 28
93.0%
100%
100%
100%
8.38 ± 1.26
100%
100%
100%
96.7%
94.12 ± 3.21
100%
96.7%
96.7%
96.7%
9.33 ± 0.87
95.0%
90.0%
90.0%
85.0%*
79.45 ± 1.02
100%
96.3%
96.3%
89.3%
773.3 ± 11.2
100%
100%
100%
100%*
8.03 ± 0.30
100%
100%
100%
90.0%
81.62 ± 2.90
100%
100%
100%
100%
727.6 ± 34.0
100%
100%
100%
100%*
or BAF between the two surface functionalized
materials.
Bioaccumulation
Tissue concentrations of Ag measured at the end of
the experiment in earthworms exposed to AgNO3 and
Ag NPs were Ag concentration-dependent (Figure 3).
E. fetida exposed to Ag at similar concentrations
accumulated significantly higher concentrations of
Ag when exposed to AgNO3 compared to either of
the Ag NPs. Bioaccumulation factors were also higher
for AgNO3 than for Ag NPs (Figure 4). The BAFs
recorded varied depending on the exposure concentration, with lower BAFs typically resulting from
greater exposures. At all of the concentrations, accumulation was very similar between the two Ag NPs,
and at comparable exposure concentrations there
were no significant differences in tissue concentration
Reproduction
The number of cocoons produced per earthworm
decreased in a concentration-dependent manner
from exposure to AgNO3 and both Ag NPs (Figure 5).
Exposure to 94.12 ± 5.56 mg kg-1 AgNO3 caused a
significant decrease in cocoon production that was
not observed in treatments of similar concentrations
of Ag NPs. Only at concentrations as high as 773.3 ±
11.2 mg kg-1 (PVP Ag NPs) and 727.6 ± 34.0 (OA Ag
NPs) was a significant decrease in cocoon production
observed for organisms exposed to Ag NPs. At
1.8
10
A
B
e
c
1.4
1.2
d
1.0
0.8
AgNO3
Ag NP (PVP)
Ag NP (OA)
AgNO3
Ag NP (PVP)
Ag NP (OA)
a
0.6
b
0.4
0.2
0.0
0.0
mg Ag kg-1 fresh tissue
1.6
[LN([Ag]Tissue +1)]0.5
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NP, Nanoparticle; PVP, Polyvinylpyrolidone; OA, Oleic acid.
8
6
4
AgNO3
Ag NP (PVP)
Ag NP (OA)
2
0
2.0
4.0
LN([Ag]Soil+1)
6.0
8.0
0
200
400
600
800
1000
mg Ag kg-1 soil
Figure 3. Ag accumulated by E. fetida after 28 days as a function of the exposure concentration of Ag in the soil. Tissue and soil concentrations
of Ag were transformed to provide a significant (p < 0.05) linear relationship in Panel A. The untransformed data are shown in panel B. All error
bars represent standard error. . = Ag+ (r2 = 0.9875; m = 0.2900); ! = Ag NP (PVP) (r2 = 0.9937; m = 0.2226); r = Ag NP (OA) (r2 = 0.9842;
m = 0.2196). a–e: Indicates groups of values that are significantly (p < 0.05) different from groups marked with a different letter.
8
W. A. Shoults-Wilson et al.
BAF ([Ag]Tissue /[Ag]Soil)
0.12
AgNO3
Ag NP (PVP)
Ag NP (OA)
a
0.10
0.08
0.06
c
0.04
b
0.02
b
d bd
e
e
0.00
10
100
1000
Figure 4. Eisenia fetida bioaccumulation factors (BAFs) for Ag
compared to exposure concentrations. Because BAF values can
vary with exposure, values are grouped by nominal exposure concentration. a–e: Indicates groups of values that are significantly
(p < 0.05) different from groups marked with a different letter.
no concentration was cocoon production significantly
different between earthworms exposed to the two
different Ag NPs. Rate of cocoon hatching was not
significantly affected by any Ag treatment when compared to the controls.
Discussion
Toxicity of silver to Eisenia fetida
The present study found no concentration-dependent
changes in growth or mortality in E. fetida caused
by AgNO3 exposure up to 94.12 ± 5.56 mg kg-1or
7
6
Cocoons / Worm
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Nominal [Ag]Soil (mg kg-1)
*
*
5
4
3
AgNO3
Ag NP (PVP)
Ag NP (OA)
Control Mean
Control SE
2
1
0
1
10
*
*
*
100
1000
[Ag] kg mg-1
Figure 5. The number of cocoons produced per individual E. fetida
at the end of the 28-day exposure period at each exposure concentration. The mean cocoons/earthworm produced by control
animals is provided as a benchmark. All error bars represent
standard error. . = Ag+; ! = Ag NP (PVP); r = Ag NP (OA).
*Indicates a mean values that are significantly (p < 0.05) different
than those of the controls.
Ag NP up to 791.7 mg kg-1, although in the case of
AgNO3, there was a non-significant decrease in
growth. In any case, the concentrations of AgNO3
used in the present study were not sufficient to cause
significant mortality in E. fetida. We did find significant, concentration dependent Ag accumulation and
reproductive toxicity was observed in earthworms
exposed to AgNO3 and Ag NPs at the highest concentrations tested for each treatment. The concentrations at which we observed reproductive toxicity was
orders of magnitude higher than currently predicted
concentrations of Ag NPs in sewage sludge (up to
6.24 mg kg-1; Gottschalk et al. 2009). However, the
highest total Ag concentrations found in sewage sludge
in the United States (82195 mg kg-1; USEPA 2009)
are similar to the Ag concentrations from AgNO3 at
which we observed reproductive impairment. Actual
concentrations in agricultural soil would be somewhat
lower depending on the actual application rate of
sewage sludge biosolids and plowing practices. In
forested soils, there may be little or no mixing of
biosolids into the soil profile. Ag may also accumulate
in the soil over long periods of repeated application.
Silver ions have long been known to be toxic to a wide
variety of aquatic organisms, with fewer studies
investigating toxicity to terrestrial organisms (Ratte
1999). Earlier studies with the earthworm species
Lumbricis terrestris (Lumbricidae, Linnaeus 1751)
demonstrated decreased growth but no significant
mortality or even bioaccumulation of silver (Ratte
1999). The studies reviewed by Ratte (1999) appear
to imply that earthworms do not accumulate or suffer
toxic effects from Ag.
One previous study also found a significant relationship between total concentrations of Ag in contaminated soil and E. fetida bioaccumulation, growth,
mortality and reproduction (Nahmani et al. 2007a).
The highest Ag concentration that the organisms were
exposed to in that study was 64.4 mg kg-1, which is
less than the highest concentration used in the present
study. However, the soils investigated by Nahmani
et al. (2007a) varied widely in composition and contained multiple potentially toxic metals, potentially
supplementing Ag toxicity. The study itself found a
strong correlation between soil characteristics (such
as soil pH and % sand), other metals (such as Pb and
Cd) and toxic effects (Nahmani et al. 2007a). Therefore, it is difficult to determine whether this previous
work provides unambiguous evidence for the toxicity
of Ag to E. fetida. A related study found a significant
correlation between Ag concentration in soil porewater and Ag uptake constants (Nahmani et al. 2009).
However, the kinetics of Ag uptake by E. fetida could
not be well described by multiple regression models.
Overall, the results of Nahmani et al. (2007a, 2009)
Effects of silver nanoparticles on earthworms
agree with our finding that earthworms exposed to Ag
accumulate it and suffer toxic effects.
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Toxicity of silver ions versus silver nanoparticles
In the present study, AgNO3 was found to be significantly more toxic to E. fetida than two Ag NPs.
E. fetida exposed to AgNO3 accumulated significantly more Ag and in some cases had significantly
decreased reproduction than those exposed to
comparable concentrations of Ag NPs. This finding
is similar to the conclusions of previous aquatic
toxicity studies. Kvitek et al. (2009) found that
AgNO3 was more toxic to Paramecium caudatum
than several different Ag NPs stabilized by surfactants. A similar study found that Ag NPs stabilized
by surfactants had antifungal activity similar to that
of AgNO3 but that ionic AgNO3 was more toxic than
non-stabilized Ag NPs (Panacek et al. 2009). AgNO3
also caused a greater decrease of photosynthesis in
the algae Chlamydomonas reinhardtii than Ag NPs
(Navarro et al. 2008). With regards to metazoans,
dissolved AgNO3 has been shown to be more toxic
than some Ag NPs to adult zebrafish (Danio rerio)
and adult Daphnia pulex, while the reverse was true
for juvenile D. rerio and neonate Ceriodaphnia dubia
(Griffitt et al. 2008). This latter study also concluded
that dissolved Ag could not account for all Ag NP
toxicity. Another study that used D. rerio embryos as
a model found a lower LC50 for AgNO3 than Ag
NPs of various sizes (Bar-Ilan et al. 2009). Similarly,
a study on the embryos of fathead minnows
(Pimephalas promelas) found AgNO3 to be more toxic
than Ag NPs (Laban et al. 2010) although once
again, dissolved Ag could not account for all Ag
NP toxicity. Overall, the body of evidence from
aquatic toxicity studies supports our finding that
Ag ions are more toxic than Ag NPs.
Some studies have drawn the opposite conclusions.
Exposure to certain Ag NPs has been shown to
produce significantly more inhibition of nitrifying
bacteria (Choi et al. 2008) and a slight decrease in
the reproduction of nematodes (Caenorhabditis elegans) compared to AgNO3 (Roh et al. 2009).
C. elegans also showed up-regulation of different
genes when exposed to ionic Ag than when exposed
to Ag NPs, indicating again that Ag NP toxicity may
not result entirely from exposure to dissolved Ag.
However, this C. elegans study was conducted in
K-medium containing high Cl- concentrations (83
mM) that would have caused precipitation of Ag as
AgCl, limiting Ag ion bioavailability. Such high Clconcentrations are not representative of the majority
of soil solutions. A study by Chae et al. (2009) found
9
increased mortality in Japanese medaka (Oryzias
latipes) exposed to AgNO3 than Ag NPs at similar
concentrations, although they did not quantify dissolved Ag in the Ag NP exposures which was mixed by
1 h of sonication and two weeks of stirring. Long
periods of oxygen-saturated suspension in water can
result in significant dissolution of Ag NPs (Liu and
Hurt 2010). Japanese medaka also exhibited different
patterns of biomarker responses between AgNO3 and
Ag NP exposures, supporting the theory that Ag NPs
have a mode of toxicity separate from dissolved Ag
(Chae et al. 2009). Taking all available information
into consideration, while dissolved Ag is typically
more toxic than Ag NPs, it is unclear whether it
accounts for all Ag NP toxicity.
The varying responses to Ag ions and Ag NPs may
be the result of comparing endpoints in different test
species, in different exposure media. It may also be
due to the different sizes and surface chemistries of
the Ag NPs under investigation. In studies that do not
analyze dissolved Ag in Ag NP exposures, toxicity
between the two sources of Ag may also be conflated.
Few previous studies have been conducted in soil
media or on organisms other than bacteria, making
comparisons with the present study difficult. In
environmentally realistic exposure scenarios, where
aggregation and dissolution are expected to play an
important role and particle surfaces are extensively
modified, particle specific toxicity may be less important. Our EXAFS studies have demonstrated that
only 1120% of the Ag was oxidized after 28 d in
the experimental soil for PVP-Ag NPs and OAAg NPs, respectively. This is almost identical to the
state of the particles prior to introduction to the soil,
making it apparent that little additional oxidation took
place during the exposure.
Scheckel et al. (2010) found that both coated and
uncoated Ag NPs exhibited similar rates of oxidation
and sorption in a kaolin reaction system over
18 months. They also observed a lack of transformation in the presence of NaNO3, but in the presence of Cl-, significant transformation occurred
over 18 months. The approximately 8-fold greater
toxicity of AgNO3 appears to correlate with the
1120% of the Ag NPs that are oxidized to Ag2O.
This suggests that oxidation and the possible release
of free Ag ions plays an important role in toxicity,
although a role of intact oxidized particles cannot be
ruled out. For example, the EXAFS data suggested
that the PVP-Ag NP containing soil contained 11%
Ag2O, however, it is possible that this fraction of
Ag2O may dissolve in the earthworm gut or mucous,
resulting in either dietary or dermal exposure. Furthermore, intact particles may have been taken up
with subsequent dissolution within the tissues
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10
W. A. Shoults-Wilson et al.
(Unrine et al. 2008). Concentrations of dissolved Ag
ions in pore waters were too low to be measured using
either ion specific electrodes or ultrafiltration (3 kDa)
followed by ICP-MS (MDL = 0.05 mg L-1).
E. fetida accumulated higher concentrations of Ag
when exposed to AgNO3 than when exposed to Ag
NPs. However, bioaccumulation does not appear to
be a function solely of Ag ions. If only the oxidized
portion of Ag NPs aged in test soils, one would expect
at least 9- to 12.5-fold less accumulation of Ag NPs
relative to AgNO3. On the contrary, the present study
observed only 2.5- to 5.5-fold less accumulation from
Ag NP-treated soils relative to AgNO3 treated soils.
This indicates that dissolved ions cannot entirely
account for the Ag accumulated by earthworms
exposed to Ag NPs, even when accounting for
decreasing BAFs with increasing exposure concentration. While the principle route of metal uptake by
earthworms is dermal absorption of dissolved ions,
studies have found that a significant amount of metal
is taken up through the digestive tract (Saxe et al.
2001; Vijver et al. 2003). Other recent studies have
indicated that earthworms may take up intact metal
nanoparticles (Unrine et al. 2008, 2010). Therefore,
while toxicity of Ag NPs may be primarily related to
Ag oxidation and dissolution, earthworms may accumulate particles which increase their tissue concentrations of Ag without appreciably contributing to
toxicity. However, Petersen et al. (2008) have estimated that any BAF below 0.0315 ± 0.0001 calculated
for E. fetida could result from incomplete voiding of
gut contents over 24 h. Because none of the BAFs for
Ag NPs calculated in the present study exceeded
0.0315, it may be difficult to distinguish Ag NP
bioaccumulation from Ag NPs retained in the gut
following voiding of gut contents. Future studies of
Ag NP toxicity and accumulation should account for
oxidation and dissolution of the Ag NPs in order to
explain potential toxicity from Ag ions and accumulation of Ag ions and Ag NPs.
Influence of surface coating on silver nanoparticle toxicity
The present study found no difference in toxicity
endpoints between organisms exposed to similarlysized Ag NPs coated with polyvinylpyrrolidone (PVP)
and oleic acid (OA). The primary difference between
the two surface coatings investigated in the present
study was in their hydrophobicity, with PVP providing
a hydrophilic coating as opposed to more hydrophobic behavior of OA-coated NPs (Li et al. 2002;
Seo et al. 2008; Yan et al. 2009). Because hydrophobicity is often correlated to bioavailability, Ag NP
bioaccumulation and toxicity would be predicted to
differ based on coating properties. Previous studies
have primarily made comparisons between noncoated Ag NPs and those functionalized by materials
that produce stable suspensions (e.g., Ahamed et al.
2008; Kvitek et al. 2009; Panacek et al. 2009). For
instance, Ag NPs modified with a surfactant
(TWEEN 80) and PVP caused more toxicity to Paramecium caudatum than unmodified particles, while
particles modified with polyethylene glycol caused
similar toxicity compared to the unmodified particles
(Kvitek et al. 2009). Likewise, Panacek et al. (2009)
found increased antifungal activity from Ag NPs
stabilized by surfactants and PVP than from Ag
NPs that were not stabilized. The authors speculated that by disrupting cell membranes, surfactants
provided Ag NPs with an easier entry into cells, thus
increasing their toxicity over unmodified particles.
Similarly, a study using mammalian cell lines found
evidence for greater apoptosis and DNA damage in
cells exposed to polysaccharide (acacia gum) coated
Ag NPs and those that were uncoated (Ahamed et al.
2008).
Surface coating has been shown to influence toxicity and uptake of Cd/Se and Cd/Te quantum dots by
a murine macrophage cell line (Clift et al. 2008).
Quantum dots coated in a hydrophobic mixture of
organic compounds caused significant toxicity that
was not caused by carboxyl and amino polyethylene
glycol coated particles. A study conducted on Fe0
NPs has likewise demonstrated that uncoated particles were more toxic to Escherichia coli than coated
particles due to fewer direct interactions between
bacterial cells and coated particles (Li et al. 2010).
They also found some differences in toxicity between
NPs with different coatings, with coatings that produced thicker layers around the NP being less toxic.
These studies indicate that the surface modification of
nanoparticles plays a large role in interaction with
cells and, therefore, toxicity. However, all involved
simplified, primarily aqueous systems and assessed
toxicity using unicellular organisms or mammalian
cell cultures. It is not clear that surface coatings would
have such a strong effect on toxicity in complex
environmental systems, especially when nanoparticles
would be expected to be coated in organic matter or
clay minerals through natural processes (Bertsch and
Seaman 1999; Li et al. 2010).
One oversight of many previous studies of surface
coating effects on Ag NP toxicity is a failure to
characterize Ag dissolution. This is an important
consideration, because surface coating may affect
the rates of oxidation and dissolution of Ag. Given
the known toxicity of Ag ions it seems prudent to
attempt to disprove the hypothesis that Ag NP toxicity
is primarily caused by release of free Ag ions before
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Effects of silver nanoparticles on earthworms
claiming that Ag NPs cause particle-specific toxicity.
Because Ag ions are readily sorbed to reactive mineral
surfaces and high molecular weight organic matter,
we were unable to directly measure Ag dissolution in
the soils; however, solid-state EXAFS analysis of the
bulk particles and soils suggested that the two particles were not chemically equivalent. The PVPcoated particles were oxidized to a lesser extent
than OA-coated particles. This indicates that surface
functionalization of Ag NPs can impact the chemical
state of the Ag NPs. While in theory this should affect
the rate of Ag ion release from the particles and
therefore toxicity, our study found no significant
difference between the two. If Ag ions are the cause
of most toxicity, parameters that affect Ag dissolution
such as soil pH, soil organic matter and gut pH or
passage time would be expected to be consistent
between the treatments and therefore lead to similar
toxicity. In any event, it is possible that particle
surface coatings may influence fate and transport of
Ag NPs, including partitioning of Ag particles to
sewage sludge, which could ultimately influence
exposure scenarios. Therefore, studies investigating
the toxicity of Ag NPs should continue to examine the
role of surface functionalization in environmentally
relevant exposures.
Conclusions
We have shown that the earthworm E. fetida can
bioaccumulate Ag from both both AgNO3 and Ag
NPs. We have also demonstrated that AgNO3 is
nearly an order-of-magnitude more toxic to E. fetida
than the two surface functionalized Ag NPs that we
examined. EXAFS analyses found only 11–20% oxidized Ag(I) and no further oxidation over time, and
could not rule out a role for Ag ions in toxicity of Ag
NPs to earthworms. Finally, as a corollary to the
previous conclusion, the relative hydrophobicity of
the Ag NP surface coatings used in this study did
not appear to affect toxicity or bioaccumulation of Ag
in E. fetida exposed in soil. This further lends support
to the idea that the observed Ag NP toxicity was
primarily related to the release of Ag ions and not
to nanoparticle-specific toxicity. More research into
Ag NP toxicity in terrestrial systems is warranted, in
particular the effects of long-term accumulation and
aging processes on Ag NP toxicity in soil are needed.
Acknowledgements
The authors wish to acknowledge the assistance of M.
Lacey, E. Harding, J. Backus, J. Song, O. Zhurbich
11
and two anonymous reviewers. Portions of this
research were carried out at the Stanford Synchrotron
Radiation Laboratory, 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 and the members of his Environmental Geochemistry Lab of Chapman University for
sample preparation and analysis support. Also, thanks
to the staff and beamline scientists at SSRL.
Declaration of interest: Major funding for this
research was provided by the United States Environmental Protection Agency (U.S. EPA) through Science to Achieve Results Grant # RD 833335 and
through support of the National Science Foundation
(NSF) and U.S. EPA for the Center for Environmental Implications of Nanotechnology (CEINT). It was
also funded in part by the National Science Foundation (BES-0608646 and EF-0830093), and the U.S.
EPA R833326. The authors report no conflict of
interest. The authors alone are responsible for the
content and writing of the paper.
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