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Characterisation of bioaccumulation dynamics of three
differently coated silver nanoparticles and aqueous
silver in a simple freshwater food chain
Judit Kalman,
A
,
D
Kai B. Paul,
A
,
B
Farhan R. Khan,
C
Vicki Stone
A
and Teresa F. Fernandes
A
,
D
A
School of Life Sciences, Heriot-Watt University, Edinburgh, EH14 4AS, UK.
B
Blue Frog Scientific, Scott House, South St Andrews Street, Edinburgh, EH2 2AZ, UK.
C
Department of Environmental, Social and Spatial Change (ENSPAC), Roskilde University,
Universitetsvej 1, PO Box 260, DK-4000 Roskilde, Denmark.
D
Corresponding authors. Emails: judit.kalman@uca.es; t.fernandes@hw.ac.uk
Environmental context. Nanoparticles may be passed from primary producers to predators higher up the food
chain, but little is currently known about this transfer. We studied the accumulation dynamics of silver
nanoparticles by algae, and then from algae to zooplankton. Using the biodynamic approach, we reconstructed
the accumulation process to show that diet is the primary route of uptake for silver nanoparticles.
Abstract. This study investigated the bioaccumulation dynamics of silver nanoparticles (Ag NPs) with different
coatings (polyvinyl pyrrolidone, polyethylene glycol and citrate), in comparison with aqueous Ag (added as AgNO
3
), in a
simplified freshwater food chain comprising the green alga Chlorella vulgaris and the crustacean Daphnia magna. Algal
uptake rate constants (k
u
) and membrane transport characteristics (binding site density, transporter affinity and strength of
binding) were determined after exposing algae to a range of either aqueous Ag or Ag NP concentrations. In general, higher
k
u
values were related to higher toxicity in the algae. Transmission electron microscopy images were used to investigate
the internalisation of Ag NPs in algal cells following exposure to low concentrations for 72 h (mimicking inhibition tests)
or high concentrations for 4 h (mimicking preparation for daphnia dietary exposure). Ag NPs were only visualised in algal
cells exposed to high Ag NP concentrations. To establish D. magna biodynamic model constants, organisms were fed
Ag-contaminated algae and depurated for 96 h. Assimilation efficiencies ranged from 10 to 25 % and the elimination of
accumulated Ag followed a two-compartmental model, indicating lower loss rate constants for polyvinyl pyrrolidone-, and
polyethylene glycol-coated Ag NPs. Biodynamic model results revealed that in most cases, food is the dominant pathway
of Ag uptake in D. magna. Despite the predicted low steady-state body burdens in D. magna, dietary uptake of Ag was
possible from aqueous and particulate forms of Ag.
Additional keywords: Chlorella vulgaris,Daphnia magna, dietary uptake, internalization.
Received 17 February 2015, accepted 20 May 2015, published online 6 October 2015
Introduction
Silver nanoparticles (Ag NPs) pose a potentially significant
environmental problem owing to their increased use and
release.
[1]
As such, nanoecotoxicology has received increasing
attention from government, regulatory bodies and academics.
[2,3]
In order to set environmental quality standards or perform risk
assessment for these materials, it is essential to determine their
bioavailability, which is considered a prerequisite for toxicity.
[4]
The biodynamic model (BDM) deconstructs trace metal accu-
mulation into a unidirectional process of uptake and efflux from
food and water respectively.
[5]
In doing so, metal body burdens
can be predicted under a variety of environmental conditions.
Moreover, the BDM suggests that toxicity occurs when the metal
influx rates exceed combined rates of loss and detoxification.
[6]
Thus, factors that affect the bioavailability of the metal species
will also affect the uptake rate constant (k
u
) and likely toxicity.
Indeed, under some circumstances, k
u
has been found to be a
successful predictor of acute toxicity.
[7]
The BDM has been
successfully applied to a wide range of trace metals and aquatic
organisms
[5]
and the suitability of this approach for metal-
containing NPs has now been tested for a range of invertebrate
species.
[8–11]
It is assumed metals must associate with receptor
sites at a biological membrane in order for uptake to occur. The
binding affinity and the binding capacity at this site, when
accounting for the environmental conditions, may correlate to
toxicity as has been shown for trace metals.
[12–14]
This rela-
tionship was recently studied on zooplankton in relation to metal
NPs
[15]
using the biotic ligand model, but there is still little
research on the application of these models for metal-containing
NPs to freshwater algae.
The importance of alga as a primary producer of energy and a
food source highlights its priority in risk assessment, because
CSIRO PUBLISHING
Environ. Chem.
http://dx.doi.org/10.1071/EN15035
Journal compilation ÓCSIRO 2015 www.publish.csiro.au/journals/envA
Research Paper
RESEARCH FRONT
impact at this level may affect the higher trophic levels. Thus, it
is not surprising that there have been a few studies on the toxic
effect of Ag NPs on freshwater microalgae.
[16–19]
Furthermore,
as an important food source, algae may facilitate the uptake of
NPs as organisms incidentally ingest NPs during feeding,
[9,20]
which may result in biomagnification and accumulation, with
potentially deleterious effects.
In principle, if an Ag NP is taken up by the organism, it can be
accumulated (in metabolically available or detoxified forms) or
excreted. As mentioned previously, toxicity of Ag NP would
occur if the rate of Ag NP uptake exceeded the combined rates of
excretion and detoxification. Whether the toxicity is caused by
the novel properties of the NP or its dissolved fraction is
uncertain. It has been documented that dissolved Ag will
associate with waterborne ligands
[12]
and debris, including
algae, an important source of nutrition for the cladoceran
Daphnia magna, and Ag NPs are no different.
[21–24]
For trace
metals, including aqueous Ag, the diet is an important route of
metal uptake and bioaccumulation.
[25]
Croteau and Luoma
[6]
and Zhao and Wang
[8]
both showed diet to be the most important
route of uptake during both dissolved Ag and Ag NP exposures
respectively, with over 70 % of accumulated tissue burdens
attributed to dietary intake. Despite this, little work has been
done on the dietary uptake of Ag NPs in D. magna and the
proportional contribution they will have to total Ag body
burdens in comparison with waterborne Ag NPs.
In the present study, we use a simple food web comprising
Chlorella vulgaris and D. magna to explore the dietary uptake of
Ag NPs and use previously reported waterborne biodynamic
parameters
[7]
to assess the importance of each route of uptake.
Such questions are integral to the current risk assessment of
Ag NPs and the future prediction of their risks, because they
allow us to differentiate the importance of the different available
exposure routes (food and water). Furthermore, we can assess
whether the current Ag NP levels in the environment are likely
to significantly accumulate and cause toxicity. It has also been
suggested if one route of uptake is predominant (food or water),
then this may aid in the simplification of such predictions
because only the major uptake route need to be accounted
for.
[6]
Thus, the main objectives of the present study were to
(i) investigate the toxicity of aqueous Ag (added as AgNO
3
) and
Ag NPs coated with polyvinyl pyrrolidone (PVP-Ag NPs),
polyethylene glycol (PEG-Ag NPs) and citrate (Cit-Ag NPs)
to C. vulgaris; (ii) investigate the uptake of Ag NPs by the algae,
define metal-binding characteristics where appropriate; (iii)
determine dietary uptake and assimilation efficiencies of Ag
in D. magna fed C. vulgaris that were previously exposed to Ag
as either NP or in aqueous form; and (iv) simulate a basic food
web and determine the potential for Ag NPs to move between the
trophic levels.
Experimental methods
Test chemicals and nanoparticle characterisation
Stock suspensions of Ag NPs coated with PVP, PEG and citrate
(nominal particle core size 10 nm) were synthesised and pro-
vided by the University of Birmingham (UK) (within the
nanoBEE consortium). AgNO
3
(analytical reagent grade) was
purchased from Fisher Scientific (Loughborough, UK) and the
stock solution was prepared by dissolving the appropriate
amount of silver nitrate in Milli-Q water (Millipore, Livingston,
UK) with a grade of 18.2 MOcm. All AgNO
3
solutions and
silver NP suspensions were made immediately before use.
Dynamic light scattering (DLS, Zetasizer Nano Series, Malvern
Instruments, Malvern, UK) was used to monitor particle size and
zeta potentials in algal medium (Jaworski’s medium, JM
[26]
)at
4mgL
1
. Hydrodynamic diameters and surface-related charge
of the Ag NPs were determined in triplicate after 0, 24, 48 and
72-h incubation.
For transmission electron microscopy (TEM) images, Ag
NPs were incubated in algal medium at 4 mg L
1
for 72 h.
A drop of the suspensions was deposited on carbon-coated
copper TEM grids, then allowed to evaporate at room tempera-
ture overnight before analysis. NPs were viewed in a Philips
CM120 TEM (Philips Electron Optics, Eindhoven, Nether-
lands) or FEI Tecnai F20 FEGTEM (Hillsboro, OR, USA) fitted
with a Gatan Orius Charge Couple Device camera. TEM energy-
dispersive X-ray spectroscopy (EDX) analysis was performed
using an Oxford Instruments 80 mm
2
X-Max SDD fitted to the
microscope and running INCA software (Oxford, UK).
Dissolution of the Ag NPs was measured using ultracentri-
fugation methods.
[27]
Briefly, Ag NPs at 2 mg L
1
were
suspended in algal medium for 4 h and then centrifuged at
51 200g,208C, for 1 h. After centrifugation, the supernatant was
separated from the pellet. The supernatant represented the
dissolved fraction released from the NPs, and the pellet, the
particulate Ag. Both fractions were acidified (2 % HNO
3
), and
total Ag was measured using flame atomic absorption spectro-
photometry (AAS, Perkin Elmer, AAnalyst 200 atomic absorp-
tion spectrometer, Beaconsfield, UK). Dissolution studies were
conducted at a higher concentration than those used in the
waterborne exposures to ensure accurate measurement of the
dissolved fraction.
[28]
Test organisms
Chlorella vulgaris (Culture Collection of Algae and Protozoa
211/12, Scottish Marine Institute, Oban, UK) culture was grown
in JM in 250-mL Erlenmeyer flasks (Scientific Laboratory
Supplies, Coatbridge, UK) under constant rotary agitation
(225 rpm), illumination (120 mmol m
2
s
1
) and temperature
(23 8C). When the cell density reached ,10
6
cells mL
1
, the
stock culture was maintained under static conditions (illumi-
nation of 50 mmol m
2
s
1
) in a 16 : 8 h light : dark photoperiod
at 20 8C. Cultures were regularly maintained by transferring a
small aliquot into fresh sterile medium and were periodically
checked for bacterial contamination by plating on nutrient agar
(Oxoid Ltd, Basingstoke, UK).
Daphnia magna used within the current study were from an
established laboratory culture originally purchased from Blades
Biological (Edenbridge, UK). D. magna were maintained in
aerated aquaria with Elendt M7 media,
[29]
which was adapted by
the removal of ethylenediaminetetraacetic acid (EDTA) to
reduce the effect of dissolved Ag or Ag NP chelation,
hereafter called aM7 (adapted M7) media. Cultures were
maintained with a 16 : 8 h light : dark photoperiod at 20 8C.
Dissolved oxygen (.8mgL
1
) and pH (8) remained within
the range recommended by standard Organization for Economic
Co-operation and Development (OECD) test protocols.
[29]
Daphnids were maintained on an algal diet of C. vulgaris
[29]
at 5 10
5
cells day
1
and medium was changed every 2–3 days.
D. magna used in the present study were 5–10 days old to ensure
peak growth rate had been reached, but that no neonate release
had occurred. After exposure, daphnids were 9–10 days old,
which ensured that individuals used in the study were in the
same developmental and physiological state.
J. Kalman et al.
B
Algal growth inhibition test
Algal growth inhibition assays were performed according to the
OECD test guideline 201
[30]
using C. vulgaris in the exponential
growth phase. Temperature and light conditions for the toxicity
test were identical to those used for culture growth. Experiments
were carried out in triplicates using five concentrations for
each Ag form. The concentration range chosen allowed us to
construct the logistic curve from which IC
50
values (the con-
centration required to cause a 50 % inhibition in growth
after 72 h compared with the controls) were calculated.
The exposure concentrations ranged between 1 and 7 mgL
1
or
9 and 65 nmol L
1
(aqueous Ag), 3 and 10 mgL
1
or 28 and
93 nmol L
1
(PVP-Ag NPs), 3 and 100 mgL
1
or 28
and 935 nmol L
1
(PEG-Ag NPs) and 1 and 15 mgL
1
or 9 and
140 nmol L
1
(Cit-Ag NPs). The initial concentration of the
inoculum was 10
4
cells mL
1
, which was required to ensure
exponential growth. At 24, 48 and 72 h, 1 mL of sample was
taken and chlorophyll was extracted from the cells in 4 mL of
pure acetone (High Performance Liquid Chromatography grade,
Fisher Scientific) along with 0.1 mL of 1.5 mg mL
1
locust
bean gum (Sigma–Aldrich, Gillingham, UK) to aid in the
precipitation of particles. After 72 h, cell density was deter-
mined in the supernatant by measuring in vitro fluorescence of
acetone-extracted chlorophyll (Trilogy laboratory fluorome-
ter, Chlorophyll-aNon-Acidification module). IC
50
values
were calculated using Sigma Plot 10.0 (Systat Software, Inc.,
San Jose, CA, USA).
Ag accumulation in C. vulgaris
To study the relationship between Ag uptake and toxicity in
C. vulgaris, algae were exposed to ranges of concentrations
(below IC
50
values) of aqueous Ag (0.1–2.0 mgL
1
or 0.9–18.6 nmol L
1
), PVP-Ag (0.5–5.0 mgL
1
or 4.6–
46.4 nmol L
1
), PEG-Ag (0.5–10.0 mgL
1
or 4.6–92.8
nmol L
1
) and Cit-Ag (0.5–4.0 mgL
1
or 4.6–37.1 nmol L
1
)
for 72 h under the test conditions described for the growth
inhibition assays. Speciation modelling (using Visual MINTEQ
ver. 3.0, KTH Royal Institute of Technology, Stockholm,
Sweden) estimated that 93% of the total Ag added as AgNO
3
occurred as Ag
þ
. Post-exposure, algal cells were harvested
by centrifugation (2230gat 20 8C for 30 min) and resuspended
in Milli-Q water three times to remove weakly bound Ag or
Ag NPs. The final pellet was dried to constant weightat 60 8C, and
then digested in 60 mL of concentrated HNO
3
at 90 8C. Each
digest was made up to 3 mL with Milli-Q water and analysed
for Ag by inductively coupled plasma mass spectrophotometry
(ICP-MS, 7500ce, Agilent Technologies, Santa Clara, CA) with
rhodium as an internal standard. Standard Reference Material
(SRM 1566b Oyster tissue, National Institute of Standards and
Technology, Gaithersburg, MD, USA) and blanks were pro-
cessed and analysed alongside experimental samples for quality
assurance. Recovery of reference material was 91.3 4.9 %
(n¼3).
TEM was used to investigate possible internalisation of the
NPs in algal cells. Two different treatments were carried
out. First, algae were exposed to Ag NPs at 2 mg L
1
for 4 h
(mimicking the preparation for daphnia dietary exposure; see
below). Second, algae were exposed to lower Ag concentrations,
e.g. 7 mgL
1
for PVP-Ag and Cit-Ag NPs, and 10 mgL
1
for
PEG-Ag NP for 72 h (mimicking the toxicity tests). After the
exposures, cells were centrifuged (2230gat 20 8C for 30 min)
and washed three times with Milli-Q water. Algal cells were
fixed using 3 % glutaraldehyde in 0.1 M sodium cacodylate
buffer (pH 7.3) for 2 h. Post-fixation was carried out in 1 %
osmium tetroxide in 0.1M sodium cacodylate for 45 min,
followed by dehydration in graded acetones, then embedded
in Araldite resin. Ultrathin sections were stained in uranyl
acetate and lead citrate.
Dietary Ag uptake and elimination in D. magna
To study the dietary uptake of Ag by D. magna, daphnids were
fed algae that had been previously exposed to the different Ag
forms. Algal cells in exponential growth phase were exposed to
either AgNO
3
(0.25 mg L
1
) or Ag NPs (2 mg L
1
) for 4 h.
These exposure concentrations were chosen to ensure a suffi-
ciently high tissue concentration was achieved in the daphnids
without causing any mortality. After exposure, algae were
centrifuged and rinsed as previously described. The algal pellet
was resuspended in aM7 medium and cell density was counted
with a hemocytometer (Sigma–Aldrich). Prior to exposure,
daphnids were depurated in clean aM7 medium for 2–4 h to
empty their guts, and then they were transferred into plastic cups
with 120 mL clean aM7 medium. Each exposure and subsequent
elimination was performed in groups of 60 daphnids. Daphnids
were exposed to Ag-contaminated food at a cell density of
510
5
cells mL
1
for either 40 min (shorter than the gut pas-
sage time for dietary Ag assimilation study) or 24 h (ensuring
sufficiently high initial body concentration of Ag for Ag elim-
ination study). To establish assimilation efficiency (AE; the
measure of the amount of metal retained from a diet) and efflux
rate constants (k
ef
), daphnids were transferred to new cups with
clean aM7 medium and allowed to depurate in the presence of
algae (unexposed to aqueous Ag or Ag NPs). The depuration
lasted 96 h. At 0, 1, 2, 4, 6, 24, 48, 72 and 96 h of depuration,
three replicate groups per treatment were sampled, thoroughly
rinsed with Milli-Q water and collected on filter paper. At each
time point, water and food were renewed to avoid re-uptake of
Ag eliminated into the water by daphnids. Animals were dried to
constant weight at 60 8C and digested in concentrated HNO
3
.
Each digest was made up to 3 mL with Milli-Q water (with a
final concentration of HNO
3
of 2 %) and analysed for Ag on
either AAS or ICP-MS.
k
ef
was calculated from the slope of the linear regression
between the natural logarithm of the percentage Ag retained
in the daphnids and the depuration time during the slow-
elimination phase (24–96 h). AE (percentage) was estimated
as the yintercept of the linear regression between the percentage
of Ag retained in the daphnids and the time of the slowly
exchanging pool.
[31]
The metal loss from the organisms is
described as follows:
½C¼½C0eðketÞð1Þ
where [C] is the bioaccumulated Ag concentration (mgg
1
)
during the physiological loss from the slow elimination phase,
[C]
0
is the bioaccumulated Ag concentration (mgg
1
) at the
start of the depuration, k
e
is the rate of constant of loss (day
1
)
and tis the depuration time (days).
Biomagnification factors (BMF) were determined by divid-
ing the Ag concentration in the daphnids by the concentration in
the diet.
Model equations
Metal influx (Influx
organism
, nmol g
1
day
1
) was interpreted in
terms of membrane transporter characteristics (Eqn 2) where
Bioaccumulation of Ag NPs in a simple food chain
C
B
max
is the binding site density (nmol g
1
), K
d
is the transporter
affinity of each binding site (nmol g
1
) from which the log Kis
derived as an affinity constant (i.e. strength of binding), and
[M]
exposure
is the exposure concentration (nmol L
1
).
[32]
Influxorganism ¼Bmax½Mexposure =ðKdþ½Mexposure Þð2Þ
Biodynamic modelling assumes that net bioaccumulation of
a metal at steady state (C
ss
) in an organism can be described as:
Css ¼ðkuw CwÞ=ðkew þgÞþðAE IR CfÞ=ðkef þgÞ
ð3Þ
where k
uw
, uptake rate constant from water (L g
1
day
1
); C
w
,
concentration in water (mgL
1
); AE, assimilation efficiency
(%); IR, ingestion rate (g food g
1
day
1
); C
f
, metal concentra-
tion in food (mgg
1
); k
ew
, rate constant of loss after uptake from
water (day
1
); k
ef
, rate constant of loss after uptake from food
(day
1
); and g, growth rate constant (day
1
).
Parameters were incorporated into the model to determine
the relative importance of waterborne and dietary uptake routes,
and predict the steady-state concentrations of Ag in daphnids.
Waterborne parameters (k
uw
and k
ew
) were determined in a
previous study by Khan et al.
[7]
Ranges of dissolved Ag
[33]
and
Ag concentrations in phytoplankton
[34,35]
reported in the litera-
ture were considered when attempting to validate the models.
Given that it is not possible to determine concentrations of Ag
NPs in the environment, we predicted the concentrations of Ag
NPs in water and food using the estimation by Gottschalk
et al.,
[36]
in which 25 % of the total Ag released is in nanopar-
ticles. The growth rate constant determined by Guan and
Wang
[37]
and IR applied by Zhao and Wang
[8]
were used in
the model.
Statistical analysis
All data were checked for normality of distribution (Kolmogorov–
Smirnov test) and homogeneity of variances (Levene’s test).
Regression analyses followed by analysis of covariance
(ANCOVA) were used to establish significant differences
between uptake rate constants. Analysis of variance (ANOVA)
was used for all other analyses. ANCOVA and ANOVA were
followed by Tukey’s honestly significant difference (HSD) post
hoc tests for a posteriori comparisons using STATISTICA
(Statsoft, version 4, Tulsa, OK, USA). Percentage data were
arcsine-transformed before statistical analysis to ensure com-
pliance with the tests’ assumptions.
Results
Characterisation of Ag NPs
The physicochemical characteristics of the pristine Ag NPs have
been previously reported.
[38]
Briefly, all Ag NPs had a core size
of ,10–11 nm. The hydrodynamic diameters (HDD) of nano-
particles suspended in Milli-Q water were 22.0 nm for PEG-Ag
NPs and Cit-Ag NPs. and 28.3 nm for PVP-Ag NPs.
[38]
The
mean zeta-potentials in Milli-Q water of the Ag NPs were
31.5, 21.0 and 31.6 mV for PVP-Ag NPs, PEG-Ag NPs
and Cit-Ag NPs respectively.
The hydrodynamic diameters of the Ag NPs over the course
of 3 days in the algal medium as measured by DLS were in a
range of 23–54, 46–97 and 15–35 nm for PVP-coated, PEG-
coated and Cit-coated NPs respectively. Particle size analysis
showed a high polydispersity index (.0.5) for each particle
type, suggesting that DLS may not be suitable for particle size
estimation. Qualitatively, the TEM images showed particle size
ranges of ,15 to 20 nm (PVP-Ag), 10 to 50 nm (PEG-Ag) and
20 to 30 nm (Cit-Ag) after 72 h of incubation (Fig. S1). The
average zeta potentials (s.e.) of the Ag NPs in JM for up
to 3 days were 11.3 0.9 mV (PVP-Ag), 10.8 1.1 mV
(PEG-Ag) and 23.2 1.0 mV (Cit-Ag), indicating moderate
NP stability in the algal medium. Changes in particle sizes or
zeta potentials measured by DLS over a period of 3 days did not
follow any specific or significant trend (ANOVA, P.0.05)
(Fig. S2). The mean percentage of dissolved Ag in the algal
medium (s.e.) after 4 h was lowest for Cit-Ag (1.2 0.0 %),
followed by PVP-Ag (4.3 0.6 %) and highest for PEG-Ag
(12.4 0.2 %).
Toxicity of aqueous Ag and Ag NPs to Chlorella vulgaris
The acute toxicity of aqueous Ag and Ag NPs to C. vulgaris is
shown in Fig. 1. Inhibitory effects of dissolved and nanosized
silver were observed after 72 h, with higher inhibitory response
in the case of dissolved silver (IC
50
of 5.3 0.5 mgL
1
). PVP-
and citrate-coated Ag NPs exhibited very similar toxicity (IC
50
values of 9.3 0.1 and 9.2 1.0 mgL
1
respectively), whereas
PEG-coated Ag NP exhibited the lowest toxicity to the algae
(IC
50
of 49.3 5.2 mgL
1
). Exposure concentrations causing
growth inhibition of more than 100 % indicated lethality of the
algae. Statistical comparison of the 72-h IC
50
s revealed signi-
ficant differences (ANOVA, P,0.05) between aqueous Ag and
PEG-coated Ag NP and both differed significantly from PVP-
and citrate-coated Ag NPs.
Uptake of Ag in C. vulgaris
A linear relationship was observed between the uptake rate of
Ag by C. vulgaris and the lower range of exposure concentration
after 72 h (Fig. 2). The silver uptake rate constant from solution
(k
u
s.e.) described by the slope of the best-fit regression
between the uptake rate and exposure concentration was greatest
for aqueous Ag (1.701 0.406 L g
1
day
1
), suggesting that
aqueous silver was the most bioavailable form to the algae. In
the case of Ag NP exposures, the uptake rate constants were
similar for PVP-Ag NP (0.233 0.024 L g
1
day
1
) and Cit-Ag
NP (0.246 0.031 L g
1
day
1
) and higher for PEG-Ag NP
(0.339 0.020 L g
1
day
1
) (ANCOVA, P.0.05).
Exposure as total Ag concentration (µg L
⫺1
)
0 20406080100120
Growth rate inhibition (%)
0
20
40
60
80
100
120
140
160
Aqueous Ag
PVP-Ag NPs
PEG-Ag NPs
Cit-Ag NPs
0 102030
0
10
20
30
40
Fig. 1. Growth inhibition of Chlorella vulgaris after 72 h of exposure to
aqueous Ag and Ag nanoparticles (NPs) (mean s.d., n¼3).
J. Kalman et al.
D
Ligand binding constants showed that Cit-Ag NP had the
lowest binding capacity (B
max
s.e.) of 17 5 nmol g
1
,
followed by PVP-Ag NP (28 13 nmol g
1
) and aqueous Ag
(44 4 nmol g
1
) and PEG-Ag NP (51 33 nmol g
1
). The K
d
(transporter affinity) for Ag form and Ag NP were 1.4 0.6,
32 18, 64 46 and 95 89 nmol L
1
for aqueous Ag, Cit-Ag
NP, PVP-Ag NP and PEG-Ag NP respectively. The binding site
affinity constant (log Ks.e.) was highest for aqueous Ag
(8.86 0.27), followed by Cit-Ag NP (7.49 0.35), PVP-Ag
NP (7.19 0.55) and PEG-Ag NPs (7.02 1.20).
TEM images of control and silver-exposed algal cells are
shown in Fig. 3. Ag NPs were not observed in the control cells or
in the cells exposed to low Ag NP concentrations (,IC
50
) for
72 h. PVP- and PEG-coated Ag NPs were, however, visualised
in the algae cells exposed at 2 mg L
1
for 4 h. The presence of
these Ag NPs in the cells was further confirmed by EDX
spectrum analysis.
Dietborne Ag uptake and subsequent elimination
by Daphnia magna
For dietborne exposures, the accumulated Ag concentrations
in algae were 1200–3300 mgg
1
dry weight. The measured
initial Ag body burden (102 28, 128 23, 124 13 and
107 5mgg
1
for aqueous Ag, PVP-, PEG- and Cit-Ag NPs
respectively) in the daphnids indicated that Ag was bioavailable
to D. magna from the algal diet for all forms. The percentage
retention of Ag (mean s.e.) by the daphnids feeding on algae
that had been exposed to either the different Ag NPs or aqueous
Ag (added as AgNO
3
) is presented in Fig. 4. Daphnids initially
eliminated ,78–88 % of the total Ag depurated from dissolved
or Ag NPs during the first 6 h of depuration, followed by a
slow excretion thereafter with k
ef
ranging between 0.279
and 0.591 day
1
(Fig. 4). D. magna apparently assimilated
less Ag from the AgNO
3
-exposed algae (10.3 0.9 %) than
from the PVP-, PEG- and Cit-Ag NP-exposed diets (24.5 3.6,
20.2 1.9, 19.6 1.1 % for PVP-, PEG- and Cit-Ag NPs
respectively).
Ag assimilated from algae exposed to PEG- and PVP-Ag NP
was eliminated by daphnids with efflux rate constants (k
ef
,
mean s.e.) of 0.279 0.022 and 0.321 0.074 day
1
, corre-
sponding to biological half-lives of 2.5 and 2.2 days respective-
ly. Retention of Ag after uptake from diet containing AgNO
3
or Cit-Ag NP was significantly lower, with k
ef
values of
0.529 0.109 and 0.591 0.043 day
1
respectively, and corre-
sponded to retention half-lives of 1.3 and 1.1 days.
BMF were calculated as 0.0013 (PVP-Ag NP), 0.0016
(PEG-Ag NP), 0.0004 (Cit-Ag NP) and 0.0005 (aqueous Ag).
Determining the importance of dietary and waterborne
uptake
Model-predicted Ag concentrations in D. magna are presented
in Table 1. Different scenarios were tested by applying ranges of
dissolved Ag (C
w
) and Ag concentrations in algae (C
f
), in other
words, varying one parameter (C
w
or C
f
) while all others are
kept constant to see how the chosen ranges modify the relative
contribution of water and food to the total Ag accumulation as
well as the prediction. Incorporating the mean and ranges of
biodynamic parameters developed for D. magna, the model
shows that food is the dominant pathway of Ag uptake in the
cases of Ag NPs (56–99 %). For aqueous Ag (i.e. added as
AgNO
3
), water becomes the major source in the overall accu-
mulation (60–85 %) when Ag concentration in food decreases
from 1.5 to 0.075 mgg
1
. The choices of growth rate constant
and ingestion rate were based on previous studies for the same
PEG-Ag NP
0
5
10
15
20
25
PVP-Ag NP
0
2
4
6
8
10
12
14
16
Cit-Ag NP
0
2
4
6
8
10
12
14
16
Uptake rate (nmol Ag g
⫺1
day
⫺1
)
Waterborne Ag concentration (nmol Ag L
⫺1
)
Aqueous Ag
0 10 20 30 40 50 0 10 20 30 40 50
0 10 20 30 40 50 0 10 20 30 40 50
0
10
20
30
40
50
60
Fig. 2. Ag uptake rates (nmol g
1
day
1
s.d.) in Chlorella vulgaris after waterborne exposures to aqueous Ag and
Ag nanoparticles (NPs) (n¼3). Linear regression (dashed line) was used to determine the uptake rate constants.
Non-linear regression (Michaelis–Menten) fits were used to derive membrane transport characteristics.
Bioaccumulation of Ag NPs in a simple food chain
E
species under similar conditions. The effects of change in these
constants over a range typically found in the literature on the
predicted Ag concentrations in D. magna are shown in Fig. S3
for completeness and shows the robustness of the biodynamic
model as conclusions remain the same.
Discussion
In the present study, we modelled transfer of different Ag NPs
and aqueous Ag in a simple freshwater food chain. The bio-
dynamic approach was used to determine the accumulation
dynamics of each Ag form to the algal food source and then used
(a) (b)
(c)
(d )
0
C
O
Cu
SAg
Ag
1 µm
1 µm
200 nm
100 nm 10 nm
EDX1
EDX2
EDX
50 nm
Cu
Cu
Cu
Cu
Cu
Cu
CuCu
Full scale 1628 cts cursor: 9.697 (6 cts)
Full scale 5033 cts cursor: 10.144 (6 cts) keV
keV
Full scale 5096 cts cursor: 9.777 (11 cts)
CAg
Ag
0123456789
keV
C
Si SAg
Ag
PVP 2ppm 0005 EDX1
PVP 2ppm 0005 EDX2
PEG 2ppm 0009
1234 5678 9
012345678910
Fig. 3. Transmission electron microscopy (TEM) observations of ultrathin slices of Chlorella vulgaris after exposure to
Ag nanoparticles (NPs): (a) control; (b) exposed to 0.007 mg L
1
Cit-Ag NP for 72 h; (c) exposed to 2 mg L
1
PVP-Ag NP
for 4 h; (d) exposed to 2 mg L
1
PEG-Ag NP for 4 h. Energy-dispersive X-ray (EDX) spectra show Ag peaks for PVP-Ag
NP (c), and PEG-Ag NP (d) exposed algae, which are seen as electron-dense dark spots (arrows).
J. Kalman et al.
F
to investigate the dietary transfer of Ag to D. magna. By para-
meterising the model with the results in the present study and
those that previously determined the waterborne accumulation
dynamics for the same set of NPs in D. magna,
[7]
we were able to
assess the relative importance of both uptake routes.
Ag uptake and toxicity to C. vulgaris
The intracellular uptake of Ag NPs has been observed in the
freshwater alga Ochromonas danica by Miao et al.
[39]
; however,
it is unclear whether growth was inhibited by Ag NPs inside the
cells directly or indirectly (by the release of Ag
þ
internally).
Aqueous Ag
Percentage of Ag retained
1
10
100
PVP-Ag NP
012345
012345
012345
012345
1
10
100
PEG-Ag NP
Time (day)
1
10
100
Cit-Ag NP
Time (h)
1
10
100
Fig. 4. Percentage of Ag retained in Daphnia magna after uptake from ingested Ag-contaminated algae.
Data are mean s.e.
Table 1. Selected model-predicted silver concentrations in Daphnia magna (C
ss
predicted)
k
uw
(6.240, 1.653, 0.264 and 0.871 L g
1
day
1
for aqueous Ag, PVP-, PEG- and Cit-Ag NPs respectively) and K
ew
(0.607, 0.274, 0.464 and 0.505 day
1
for
aqueous Ag, PVP-, PEG- and Cit-Ag NPs respectively) from Khan et al.,
[32]
g¼0.08 day
1
,
[39]
IR ¼0.91 g g
1
day
1
.
[7]
AE (10.3, 24.5, 20.2 and 19.6 % for
aqueous Ag, PVP-, PEG- and Cit-Ag NPs respectively) and K
ef
(0.529, 0.321, 0.279 and 0.591 day
1
for aqueous Ag, PVP-, PEG- and Cit-Ag NPs respectively)
from the present study. Ranges of dissolved Ag (C
w
) and Ag concentrations in food (C
f
) were used to estimate the relative contribution of water (C
ss
water) and
food (C
ss
food) to the overall Ag accumulation at steady state (C
ss
predicted). See Experimental methods for more details
C
w
(mgL
1
)C
f
(mgg
1
)C
ss
water (%) C
ss
food (%) C
ss
predicted (mgg
1
)
Aqueous Ag
0.0020 0.075 60.3 39.7 0.03
0.0020 1.5 7.0 93.9 0.25
0.0072 0.075 84.8 15.2 0.08
0.0072 1.5 21.9 78.1 0.30
PVP-Ag NP
0.0007 0.025 18.1 81.9 0.02
0.0007 0.5 1.1 98.9 0.29
0.0024 0.025 44.9 55.1 0.03
0.0024 0.5 3.9 96.1 0.30
PEG-Ag NP
0.0007 0.025 2.4 97.6 0.01
0.0007 0.5 0.1 99.9 0.27
0.0024 0.025 8.2 91.8 0.01
0.0024 0.5 0.4 99.6 0.27
Cit-Ag NP
0.0007 0.025 12.6 87.4 0.01
0.0007 0.5 0.7 99.3 0.14
0.0024 0.025 34.8 65.2 0.01
0.0024 0.5 2.6 97.4 0.14
Bioaccumulation of Ag NPs in a simple food chain
G
Internalisation of NPs in O. danica was observed by TEM in
cells from treatments with no obvious toxic effects. Despite
carrying out specific membrane permeability assessments, the
authors excluded the possibility that the increase of cell mem-
brane permeability resulted in a passive uptake of NPs. How-
ever, polymer-coated CuO NPs (with an average size of 65 nm)
were able to penetrate the cell of Chlamydomonas reinhardtii in
particulate form after 6 h of exposure as observed by Perreault
et al.
[40]
Internalisation of CuO NPs (,5 nm) was also revealed
in the prokaryotic alga Microcystis aeruginosa by Wang
et al.,
[41]
with no increase in membrane permeability observed
during 24-h exposure, the time during which internalisation
could have taken place. However, given that algal cell walls are
often porous in their structure (usually between 5 and 20 nm),
and their permeability changes during mitosis, NPs less than
20 nm in size may pass freely through the cell wall. Moreover,
NPs may induce the formation of larger new pores permeable
to bigger NPs,
[42]
which could be the case in the present study. It
is important to note that no NPs were revealed by TEM in
Chlorella vulgaris after 72 h at low exposure concentrations
(,IC
50
). When cells were exposed to 200 or 285 times higher
concentrations of Ag NPs, intracellular uptake was observed
after 4 h of exposure. Thus, there is a possibility that such a high
concentration of Ag NPs may lead to an increase in membrane
permeability and subsequent internalisation of NPs. Increases in
cell membrane permeability have been shown across many
phyla.
[43–47]
In the present study, Ag NPs were localised in starch granules
within the chloroplast of C. vulgaris as shown by TEM images,
suggesting that granules may act as a storage site for NPs in this
alga. Zhou et al.
[48]
found an increased number of starch
granules in Chlorella pyrenoidosa after exposure to dissolved
zinc and copper; this is likely a defence mechanism against such
ionic metals and is likely used to sequester these toxicants.
Another possible mechanism that can protect microorganisms
against metal toxicity is the secretion of exopolymeric sub-
stances (EPS). The external surface of C. vulgaris contains
EPS, proteins and carbohydrates, which facilitate the binding of
metal ions.
[49]
Miao et al.,
[50]
for instance, found a higher Ag
þ
tolerance due to the secretion of EPS by the marine diatom
Thalassiosira weissflogii, suggesting that EPS may provide
binding ligands for toxicants released from NPs.
In regards to uptake, dissolved Ag exposures resulted in the
highest k
u
values. In comparison with the Ag NPs, the aqueous
Ag uptake rate constant was ,4–7 times higher and aqueous Ag
toxicity was 2–10 times greater. Higher uptake rate constants
of aqueous Ag in comparison with Ag NPs corroborates the
current literature.
[7,51]
Thus, it can be concluded that dissolved
Ag is the most bioavailable form to the algae. For the Ag NPs,
both PVP-Ag NPs and Cit-Ag NPs showed similar uptake rates,
which corresponded to the similar toxicity and bioavailability
of these Ag NPs. In other aquatic species, it has been shown
that the Ag uptake rate constant of particulate (i.e. Ag NP and
food-bound Ag)
[6,7]
and dissolved Ag can be directly linked to
toxicity. These studies show that uptake rate constants may be
used to reliably predict organism toxicity across many aquatic
species. A further advantage of this approach (i.e. using uptake
rate constants to predict toxicity) is that uptake rate constants are
not biased towards the route of uptake. Therefore, the burden of
the toxicant from ingested food and through a biological ligand
as is the case with dissolved metal, is accounted for. However,
although PVP-Ag NPs and Cit-Ag NPs conformed to such a
relationship, PEG-Ag NPs showed the most rapid uptake rate
constant of all Ag NPs, yet the lowest toxicity. PEG- and PVP-
coated Ag NPs did not reach maximal Ag saturation at any
studied concentration; thus, it is likely inappropriate to deter-
mine any relationships between the derived membrane transport
parameters (B
max
,K
d
and log K) and toxicity. Cit-Ag NP
dissolution was more than four times lower than that of PVP-
Ag NPs and PEG-Ag NPs, suggesting that dissolved Ag from Ag
NPs is not the only Ag form that may interact with or be
bioavailable to the algae (Table 2). Therefore, toxicity does
not seem to relate entirely to the dissolved Ag NP fraction but
more likely the specific interactions between both dissolved
Ag (from the Ag NPs) and the Ag NPs (and inherent physico-
chemical characteristics in the medium) themselves. How-
ever, given that solubility studies were carried out at higher
total Ag concentration compared with those of the exposure,
results cannot be directly extrapolated to the lower exposure
concentrations.
[7,52]
D. magna biodynamic modelling
D. magna assimilated Ag (presented to the daphnids in the form
of Ag-exposed algae) from ingested dietary Ag NPs more effi-
ciency than aqueous Ag (AgNO
3
). These results are similar to
other studies using the same species.
[8,31]
However, this is in
opposition to Croteau et al.,
[9]
who showed higher assimilation
of foodborne AgNO
3
in Lymnaea stagnalis fed Ag-spiked dia-
toms. Several factors have been shown to influence the assim-
ilation efficiency of contaminants from ingested food, including
the food quantity and quality, digestive physiology of the ani-
mals and behaviour of the contaminants in the animal’s gut.
[53]
Although the AE of Ag is highly dependent on the algal cell
density and Ag concentration in the diet,
[31]
our data are com-
parable with those reported in the literature.
[2]
Significant dif-
ferences between efflux rate constants in the present study
indicate possible different metabolic pathways for the different
Table 2. Chlorella vulgaris. Summary of biodynamic parameters (k
u
, uptake rate constant; mean ± s.e.), metal binding characteristics (B
max
, binding
site density; K
d
transporter affinity; log K, strength of binding; mean ± s.e.), growth inhibition (IC
50
, concentration required to cause a 50 %
inhibition; mean ± s.d.) and particle dissolution (mean ± s.e.)
PVP, polyvinyl pyrrolidone; PEG, polyethylene glycol; cit, citrate; NP, nanoparticle
Aqueous Ag PVP-Ag NP PEG-Ag NP Cit-Ag NP
k
u
(L g
1
day
1
) 1.701 0.406 0.233 0.024 0.339 0.020 0.2460.031
B
max
(nmol g
1
)4442813 5133 17 5
K
d
(nmol L
1
) 1.4 0.6 64 46 9589 32 18
log K8.86 0.27 7.19 0.55 7.02 1.20 7.49 0.35
IC
50
(mgL
1
) 5.3 0.5 9.3 0.1 49.3 5.2 9.2 1.0
Dissolution (%) – 4.3 0.6 12.4 0.2 1.2 0.0
J. Kalman et al.
H
types of particle coatings. Complete depuration of Ag NPs from
daphnids was not obtained within 96 h indicating that Ag NPs
can be passed on through trophic transfer. If Ag NPs were to
reside within the gut for extended periods of time, this might
result in feeding depression, which has been shown to be caused
by Ag NPs.
[54]
The impairment of feeding can directly affect
growth and reproduction of the individual, which may have
consequences at the population level.
[55,56]
Biodynamic modelling was applied to determine the routes
of aqueous Ag and Ag NPs uptake in D. magna (i.e. food and
water), and to predict the accumulated concentrations of Ag in
the daphnids. Different scenarios tested show that food is the
major source of accumulated Ag in D. magna in the cases of
Ag NPs. This is consistent with results previously recorded for
D. magna
[8]
and for L. stagnalis.
[9]
Predominance of a dissolved
source of Ag in daphnids reported by Lam and Wang
[31]
was
explained by the quite low dietary AE accompanied by high
k
u
; however, the authors noted a strong dependency of AE on
food concentration. Taking into account the relevant pathways
of metal uptake can improve predictions of metal bioaccumula-
tion in an organism. Nevertheless, defining the relative contri-
bution of metals to the overall steady-state body burden may
allow simplification of model, particularly when one uptake
route prevails.
[6]
Our model-predicted accumulated steady-state
concentrations of Ag ranged from 0.01 to 0.30 mgg
1
(dry
weight). These levels are below levels of Ag NP accumulation
reported to cause acute toxicity for other species. For example,
Zhao and Wang
[57]
using the same species investigated 48-h
50 % lethal concentrations and 48-h bioaccumulation of Ag
NPs. The authors noted that at high exposure concentration
(200 mgL
1
) and despite the high accumulated Ag concentra-
tions in daphnids (475–6618 mgg
1
dry weight), no signs of
toxicity were observed. Thus, current levels of Ag within the
environment may not lead to adverse effect on D. magna,
a highly sensitive organism within freshwater aquatic biota.
However, using parameters such as body burdens to predict
toxicity may be ill-advised for those species that ingest a large
portion of Ag NPs, such as D. magna.
[7]
In the present study,
there was no appreciable Ag biomagnification by the daphnids
(BMF ,1); however, this route of exposure is still important
within the environment.
Conclusions
Aqueous Ag was more toxic than Ag NPs to Chlorella vulgaris.
The internalisation of PVP- and PEG-Ag NPs into the algal
cells seemed to occur only at high exposure concentration,
probably owing to increased membrane permeability. In gen-
eral, the uptake rate constants (k
u
) derived for the algae corre-
lated well with toxicity. According to the dissolution studies, Ag
released by the Ag NPs did not wholly account for toxicity.
When Ag-exposed algae were fed to the next trophic level,
Daphnia magna assimilated Ag from ingested algae for all
forms of Ag. However, Ag NPs were assimilated more effi-
ciently than aqueous Ag. Moreover, Ag NPs were not eliminated
completely from daphnids over the depuration period, which
may lead to further possible transport of Ag NPs along the
food chain. Testing different scenarios (minimum and maxi-
mum Ag concentrations in the water and food), the model
showed that for Ag NPs, food is the dominant pathway of Ag
uptake in D. magna. When the concentration of Ag in food is
low, water becomes the major route of Ag uptake in the case of
aqueous Ag.
Author contributions
J. Kalman and K. B. Paul are joint first authors and conducted
the experiments. All authors contributed to the design of the
experiment, discussed the results and commented on the
manuscript.
Acknowledgements
This study was financially supported by the US Environmental Protection
Agency–UK Natural Environment Research Council nanoBEE project. The
authors thank Birmingham University for the Ag NPs, Steve Mitchell
(University of Edinburgh) for TEM sample preparation and Dr Mike Ward
(University of Leeds) for the TEM images and EDX analysis (sponsored by a
Biotechnology and Biological Sciences Research Council grant).
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