Ecotoxicity of nanoparticles of CuO and ZnO in natural water
ABSTRACT The acute toxicity of CuO and ZnO nanoparticles in artificial freshwater (AFW) and in natural waters to crustaceans Daphnia magna and Thamnocephalus platyurus and protozoan Tetrahymena thermophila was compared. The L(E)C50 values of nanoCuO for both crustaceans in natural water ranged from 90 to 224 mg Cu/l and were about 10-fold lower than L(E)C50 values of bulk CuO. In all test media, the L(E)C50 values for both bulk and nanoZnO (1.1–16 mg Zn/l) were considerably lower than those of nanoCuO. The natural waters remarkably (up to 140-fold) decreased the toxicity of nanoCuO (but not that of nanoZnO) to crustaceans depending mainly on the concentration of dissolved organic carbon (DOC). The toxicity of both nanoCuO and nanoZnO was mostly due to the solubilised ions as determined by specific metal-sensing bacteria.
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ABSTRACT: We examined the influence of dissolved organic carbon (DOC) on the bioavailability of waterborne Cu to rainbow trout (Oncorhynchus mykiss) during chronic sublethal exposure. Juvenile rainbow trout were exposed to Cu (as CuSO(4)) and DOC as humic acid (HA, as sodium salt) for one month in synthetic soft water to give treatments with varying combinations of free ionic and HA complexed Cu. The total Cu concentration was 7 microg/l for all treatments (except controls) and HA was added at levels of 0, 2.5 and 7.5 mg/l which corresponded to DOC levels of 1.2, 2.2 and 4.0 mg/l. Fish grew well in all treatments and no mortalities occurred. Cu was highly bioavailable in the treatment with no added HA; gill and liver Cu accumulation occurred as well as a disruption of Na(+) regulation. In Cu treatments with additions of both 2.5 and 7.5 mg/l HA, there was no significant tissue accumulation of Cu. The addition of HA alleviated and delayed the disruption of iono-regulatory mechanisms. A recovery of plasma Na(+) losses was observed and this was associated with an increase in gill Na(+)/K(+) ATPase activity by the end of the exposure. Following the month of chronic exposure the uptake and turnover rates of Cu at the gills and into various tissue compartments were measured through radioisotopic techniques ((64)Cu). While chronic Cu exposure did not result in acclimation (i.e. increased LC50), the uptake rate and extent of Cu uptake into the gills and liver was increased. This study demonstrates that growth and tissue accumulation of Cu are poor predictors of the chronic effects of Cu, and illustrates that HA moderates chronic Cu bioavailability. The lack of a link between Cu bioaccumulation and Cu impact and the role of organic matter in reducing the bioavailability of Cu are important considerations in the context of ecological risk assessment.Comparative Biochemistry and Physiology Part C Toxicology & Pharmacology 10/2002; 133(1-2):147-60. · 2.71 Impact Factor
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ABSTRACT: Metal body burdens of larval and adult chironomids varied nonlinearly with H+. Metal loss during emergence was an indicator of metal adsorption to the exoskeleton. This adsorbed portion was lower for chironomids collected from lakes of pH 4.4 or 6.1 than for those from lakes of pH 5.1 to 5.5. Pb accumulation was inversely proportional to pH, but Cd, Al, Cu and Ni body burdens were lower at pH 4.4 than at pH greater than 5.1. Short-term larval transplant experiments supported the field observations that surface adsorption can be an important component of the total metal burden of larval chironomids and that adsorption processes may be suppressed at low pH. Changes in metal content in live and dead chironomids exposed to different sediment types were not significantly different. Larvae from a lake of pH 4.4 showed an initial rapid phase of accumulation of Cd, Al, Mn, Ni, Zn and Cu when transplanted to sediments from a lake of pH 5.1. Laboratory and field data supported the hypothesis that surface adsorption contributes to total metal content in chironomids and is responsive to lake pH.Environmental Toxicology and Chemistry 07/1988; 7(8):653 - 670. · 2.62 Impact Factor
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ABSTRACT: Batch experiments were performed to investigate the feasibility of humic acid (HA) removal by synthetic nanoscale zerovalent iron (NZVI) and its interaction with As(III) and As(V), the most poisonous and abundant of groundwater pollutants. High-resolution transmission electron microscopy (HR-TEM) and X-ray diffraction (XRD) were used to characterize the particle size, surface morphology of the pristine NZVI and HA-treated NZVI (NZVI-HA), and the zero valence state of the pristine NZVI. It was determined that HA was completely removed by NZVI (0.3 g/L) within a few minutes, at a wide range of initial pH values (approximately 3.0-12.0). Fourier transform infrared (FTIR) and laser light scattering (zeta potential measurement) studies confirmed that NZVI-HA forms inner-sphere surface complexation at different initial pH conditions. The effects of competing anions showed that there was complete removal of HA in the presence of 10 mM NO(-3) and SO4(2-) whereas HA removal was observed 0%, 18% and 22% in presence of 10 mM H2PO4(2-), HCO(3-) and H4SiO4(0), respectively. However, the presence of 2 mM CA2+ and Mg2+ enhanced HA removal from 17 mg g(-1) to 76 mg g(-1) and 55 mg g(-1), respectively. Long-term time-resolved studies of XRD and field emission scanning electron microscopy (FE-SEM) with energy-dispersive X-ray (EDX) revealed the formation of various types of new iron oxides (magnetite, maghemite, and lepidocrocites) during the continuous reaction of HA in the presence of water and NZVI at 1, 30, 60, and 90 days. In addition, the surface-area-normalized rate constant (ksa) of adsorption of As(III) and As(V) onto NZVI was reduced in the presence of HA (20 mg L(-1)), from 100% to 43% and 68%, respectively. Our results show the potential use of NZVI in removing HA and its possible effects on arsenic removal during the application of NZVI in groundwater remediation.Environmental Science and Technology 04/2007; 41(6):2022-7. · 5.26 Impact Factor
Ecotoxicity of nanoparticles of CuO and ZnO in natural water
I. Blinovaa,*, A. Ivaska, M. Heinlaana,b, M. Mortimera,c, A. Kahrua
aLaboratory of Molecular Genetics, National Institute of Chemical Physics and Biophysics, Akadeemia tee 23, Tallinn 12618, Estonia
bEstonian University of Life Sciences, Kreutzwaldi 5, Tartu 51014, Estonia
cTallinn University of Technology, Akadeemia tee 15, Tallinn 12618, Estonia
Natural waters remarkably reduced the toxicity of nanoCuO but not nanoZnO.
a r t i c l e i n f o
Received 13 March 2009
Received in revised form
1 July 2009
Accepted 11 August 2009
Metal oxide nanoparticles
Recombinant luminescent sensor bacteria
a b s t r a c t
The acute toxicity of CuO and ZnO nanoparticles in artificial freshwater (AFW) and in natural waters to
crustaceans Daphnia magna and Thamnocephalus platyurus and protozoan Tetrahymena thermophila was
compared. The L(E)C50values of nanoCuO for both crustaceans in natural water ranged from 90 to 224 mg
Cu/l and were about 10-fold lower than L(E)C50values of bulk CuO. In all test media, the L(E)C50values for
remarkably (up to 140-fold) decreased the toxicity of nanoCuO (but not that of nanoZnO) to crustaceans
depending mainly on the concentration of dissolved organic carbon (DOC). The toxicity of both nanoCuO
and nanoZnO was mostly due to the solubilised ions as determined by specific metal-sensing bacteria.
? 2009 Elsevier Ltd. All rights reserved.
The last decade is distinguished by the drastic growth of produc-
such as ZnO and TiO2are already widely used in personal care prod-
ucts (e.g., sunscreens), coatings and paints;CuO is used in gas sensors,
photovoltaic cells, in catalyst applications and in heat transfer nano-
fluids. Subsequently, the risk of natural water contamination by
synthetic NPs continuously increases (Klaine et al., 2008).
It should be stressed that natural NPs, including nano-sized
particles of metal oxides, exist in all ecosystems and play important
role in biogeochemical processes (Wigginton et al., 2007). During
the evolution living organisms have adapted to the presence of
natural NPs in the environment. For synthetic NPs, however, it is
recognized that their potential harmful properties on ecosystems
have to be evaluated (Handy et al., 2008; Nowack, 2009). Despite
the rapidly increasing amount of nanotoxicological peer-reviewed
papers (Medina et al., 2007) data on ecotoxicity of synthetic NPs
(Baun et al., 2008; Handyet al., 2008) and especially on metal oxide
NPs, except nanoTiO2, are rare (Kahru et al., 2008). As water is an
essential compartment in ecosystems and natural vehicle for
pollutant migration, the data on fate and behavior of synthetic NPs
in different types of natural waters as well as their potential eco-
toxic effects are essential for evaluation of the environmental risks
of nanotechnologies (Nowack and Bucheli, 2007).
The main goal of evaluation of ecotoxicological properties of
chemicals is to prevent the hazard to the ecosystems via estab-
lishing respective environmental standards, which guarantee the
absence of negative effects of those compounds on living organ-
isms. At the same time, it is widely accepted and also shown in our
earlier works (Aruoja et al., 2004; Blinova, 2004; Kahru et al., 2005)
that due to the environmentally non-relevant conditions used
in regulatory testing, most of the standardized bioassays do not
appropriately characterize the potential impacts of hazardous
substances on the environment, in particular, on water ecosystems
(Allen and Hansen, 1996; Hyung et al., 2007; Lewis, 1995). Most of
the ecotoxicity data on chemicals available for standard freshwater
test organisms such as crustaceans, algae and fish have been
generated using so-called artificial freshwater (AFW), which
composition differs from natural waters. However, as bioavail-
ability and toxic effect of a chemical depend on its speciation and
hence, on water composition (Witters, 1998), the hydrochemical
parameters of water used as test medium are very important.
biological effects of chemicals in the aquatic ecosystems has not been
adequately explored. For example, in spite of the intensive investiga-
tion of the effects of natural water composition on bioavailability of
heavy metals during the last decades leading even to the elaboration
of several models, which are used for the prediction of metal toxicity
* Corresponding author. Tel.: þ372 6 398 361.
E-mail address: email@example.com (I. Blinova).
Contents lists available at ScienceDirect
journal homepage: www.elsevier.com/locate/envpol
0269-7491/$ – see front matter ? 2009 Elsevier Ltd. All rights reserved.
Environmental Pollution 158 (2010) 41–47
in natural water (Jager et al., 2006; Kim et al.,1999; Long et al., 2004;
McGeer et al., 2002; Niyogi and Wood, 2004; Pagenkopf, 1983, etc.),
understandingof the behaviorand biological effects of trace metals in
natural waters is still limited (Borgmann, 2000; Handy et al., 2008;
Town and Filella, 2002; Van Leeuwen et al., 2005). For synthetic NPs,
it is known that the type and amount of natural organic matter in the
water affects their stability and bioavailability (Giasuddin et al., 2007;
their behavior in surface waters (Klaine et al., 2008; Lead and
Wilkinson, 2006). However, the effect of organic ligands as well as of
other hydrochemical parameters (pH, hardness, ionic strength, etc.)
on the bioavailability of NPs to aquatic organisms is still inadequately
results obtained with AFW (Handy et al., 2008; Nanotechnology,
2006; Velzeboer et al., 2008).
In order to be relevant for the use in the risk assessment and
establishment of environmental quality standards, ecotoxicity tests
should give information on a chemical’s bioavailability and toxicity
in the given environment (McGeer et al., 2002; Van Assche et al.,
2002). Additional knowledge required to extrapolate laboratory
test results to field populations may be received by extending the
standard protocolsof ecotoxicological testing (Jageret al., 2006), for
example, by replacing the AFW with a natural one.
CuO and ZnO NPs towards particle-ingesting aquatic species (two
crustaceans and one protozoan) in AFW and in six natural waters
the respective soluble salts (CuSO4and ZnSO4$7H2O) were used as
controls for size-dependent and solubility effects. Bioavailability of
Cu and Zn as well as the solubilisation of metal oxides in natural
waters was studied by recombinant sensor bacteria.
2. Materials and methods
2.1. Natural water
Natural water samples were taken during November–December 2007 from six
Estonian rivers with different hydrochemical characteristics. Sampling places were
chosen according to the data of the national monitoring program. The chemical
analysis of water samples (Table 1) was performed in a certified laboratory using
the following standard analytical methods: EN 1899-1:1998 for biochemical
oxygen demand (BOD7), ISO 9963-1:1994 for alkalinity, ISO 11905-1:1997 for total
nitrogen (Ntot), ISO 6878:1998 for total phosphorus (Ptot), ISO 8245:1999 for dis-
solved organic carbon (DOC), ISO 17294-2:2003 for Zn and Cu, ISO 10304-1:1992
for sulphate, ISO 10304-1:1992 for chloride, ISO 6058:1984 for calcium, SFS 3032
(1976) for ammonium. Before the biotesting, suspended solids and plankton
were separated from the water samples by filtration through a 0.45 mm pore size
standard filter (Millipore).
NanoCuO (advertised particle size 30 nm), nanoZnO (advertised particle size
70 nm) and ZnSO4$7H2O were purchased from Sigma–Aldrich, the bulk form of ZnO
from Fluka, thebulk CuOandCuSO4from Alfa Aesar. The stock solutionsofmetalsalts
and suspensions of metal oxides were prepared in MilliQ water. The suspensions of
metal oxides (40 g/l) were sonicated for 30 min and stored in the dark at þ4?C.
2.3. Electron microscopy imaging
The aqueous suspensions of the studied metal oxides (both nano- and bulk
formulations) have been previously characterized by scanning electron microscopy
(SEM): despite of agglomeration, individual nanoscale particles were present in
nanoZnO and nanoCuO suspensions (Kahru et al., 2008). For the current study, size
distribution of CuO NPs (40 mg CuO/l) was analysed by transmission electron
microscopy in Daphnia magna test medium (AFW) using a JEOL 1230 TEM at 120 kV.
For the characterization of NPs after ingestion by D. magna, organisms were exposed
to nanoCuO (4 mg/l) and live crustaceans were fixed for TEM observations using
a JEOL 1011 TEM at 100 kV. TEM photographs were taken from the gut after thin-
sectioning. Individual particles were sized (based on 200 measurements) from TEM
photographs using the freeware ImageJ (NIH, USA).
2.4. Aquatic bioassays
The crustacean D. magna acute immobilisation assay (Daphtoxkit F?) adhering to
OECD 202 guidelines, crustacean Thamnocephalus platyurus acute mortality test
(Thamnotoxkit F?) and ciliate protozoan Tetrahymena thermophila growth inhibition
test (Protoxkit F?) were used. The Toxkits were purchased from MicroBioTests, Inc.
(Nazareth, Belgium). In bothcrustaceanassaysviableand deadorganismswerecounted
under dissection microscope after 24 h (T. platyurus) or 48 h (D. magna) of exposure. For
protozoan growth inhibition test, the investigated compound and T. thermophila culture
protozoan culture clears the substrate suspension in 24 h, inhibition of the growth of
protozoa is reflected by residual turbidity of the food substrate measured by optical
density (OD) of the tests samples at 440 nm. All the tests were performed in the dark
at constant temperature (20?C for D. magna, 25?C for T. platyurus and 30?C for
T. thermophila) according to the respective guidelines of the Toxkits.
The test organisms were exposed to different concentrations of CuO and ZnO
(both, nano- and bulk forms), CuSO4and ZnSO4$7H2O. The filtered river waters
(Table 1) were used as basic test medium in the tests, i.e. for the dilution of studied
compounds and as a control. Evaluation of the toxicity was performed in two steps:
i) determination of 0–100% tolerance range of the test species to the respective
compound and ii) determination of the 50% effect values L(E)C50. The toxicity was
evaluated from 2 to 3 independent experiments, each in several replicates (four for
D. magna, three for T. platyurus and two for T. thermophila).
The AFW (test medium used in the standard test procedure) for crustaceans has
following composition (mg/l): for D. magna – CaCl2$2H2O – 294, MgSO4$7H2O –
123.25, NaHCO3– 64.75, KCl – 5.75, pH ?7.8 ? 0.2 and for T. platyurus – CaSO4$2H2O
– 60, MgSO4$7H2O – 123, NaHCO3– 96, KCl – 4 mg/l; pH – 7.8 ? 0.2 dissolved in
MilliQ water, i.e. AFW does not contain organic compounds. MilliQ water was used
as standard test medium for T. thermophila.
2.5. Bacterial metal-specific biosensors
In parallel to the aquatic biotests, dissolved bioavailable Zn2þand Cu2þin the
solutions/suspensions of tested compounds were quantified using recombinant
bioluminescent Zn-sensor bacteria Escherichia coli MC1061(pSLzntR/pDNPzntAlux)
and Cu-sensor bacteria E. coli MC1061(pSLcueR/pDNPcopAlux), respectively (Ivask
et al., 2009). Bioluminescence of those sensor bacteria increases proportionally with
the concentration of bioavailable Cu2þ(Cu-sensor) or Zn2þ(Zn-sensor) in the test
medium (Ivask et al., 2002). A constitutively luminescent control strain E. coli
MC1061(pDNlux) (Leedja ¨rv et al., 2006) not induced by heavy metals, but otherwise
similar to sensor strains, was used to take into account the potential quenching of
bacterial bioluminescence by the turbid suspensions of metal oxides. 100 ml of the
suspension of Zn- or Cu-sensor bacteria or the constitutively luminescent control
bacteria in 9 g/l of NaCl supplemented with 1 g/l of cas-aminoacids (acid hydrolysate
of casein, LabM) and 0.9 g/l of glucose was mixed with 100 ml of the solution or
suspension of studied metal compound diluted either in MilliQ, AFW for D. magna or
T. platyurus or in the natural river waters and incubated for 2 h at 30?C as described
previously (Heinlaan et al., 2008). The amount of bioavailable Cu and Zn was
quantified assuming that CuSO4and ZnSO4$7H2O were 100% bioavailable to the
sensor bacteria, when compounds were diluted in MilliQ. Detection limits of this
method were 2 mg Cu2þ/l and 20 mg Zn2þ/l.
2.6. Statistical analysis
statistical significance of the differences between toxic effects of the compounds
in different test media. The differences were considered significant, when p < 0.05.
Characterization of natural waters used as test media.
ParameterUnitRiver 1River 2River 3River 4River 5River 6
aDOC – dissolved organic carbon.
bBOD7– biochemical oxygen demand.
cNtot– total nitrogen.
dPtot– total phosphorus.
eZntot– total zinc.
fCutot– total copper.
I. Blinova et al. / Environmental Pollution 158 (2010) 41–4742
3. Results and discussion
3.1. Chemistry of natural water samples
The main water quality parameters, which may affect bioavail-
ability of studied compounds, are presented in Table 1. As consid-
erable part of Estonian rivers and lakes, similarly to typical boreal
ones, contain relatively high concentration of dissolved organic
matter (DOM), this study was mainly focused on the effect of DOM
on bioavailability and toxicity of CuO and ZnO NPs. Table 1 shows
that the content of DOM (expressed as dissolved organic carbon,
of nutrients (phosphorous and nitrogen) in the water samples was
relatively high. The values of biochemical oxygen demand (BOD7),
remained within the same range during the toxicity testing. The
background values of total copper and zinc in river waters (Table 1)
were negligible compared to respective L(E)C50values of CuSO4and
detect bioavailable Cu and Zn in any of the natural water samples.
3.2. Bioavailability of Zn and Cu to recombinant sensor bacteria
At the low concentrations used for the toxicity testing (up to
10 mg/l) ZnO (bulk and nano) was almost fully dissolved and
bioavailable for bacterial biosensors in all test media (Fig. 1).
Differently from both ZnO formulations, nanoCuO and especially
bulk CuO were of remarkably lower solubility: only about 12% of
copper from nanoCuO (the tested concentration range was 0.006–
20 mg/l) and about 0.3% of copper from bulk CuO suspensions (the
tested concentration range was 0.2–2000 mg/l) proved bioavailable
to sensor bacteria in AFW (Fig. 1). Thus, in AFW the solubility of
nanoCuO exceeded the solubility of bulk CuO about 40-fold. Anal-
ogously, the toxicity of nanoCuO in AFW was about 50-fold (D.
magna) and 45-fold (T. platyurus) higher than the toxicity of bulk
CuO (Table 2) proving again that the toxicity of copper oxides for
crustaceans are due to solubilised Cu-ions as also previously shown
for bacteria (Heinlaan et al., 2008) and algae Pseudokirchneriella
subcapitata (Aruoja et al., 2009). In comparison with AFW, the
natural waters reduced the bioavailability of all studied Cu-
compounds to recombinant sensor bacteria by 89–99% (calculated
from data presented in Fig. 1), whereas the effect was most
remarkable in case of nanoCuO. Analogous remarkable reduction in
bioavailability was not observed for Zn compounds (Fig. 1).
3.3. Toxicity of copper compounds in natural waters to crustaceans
The effect of the natural water composition (especially DOM) on
toxicity of the studied copper compounds towards D. magna and
T. platyurus was analysed by comparison of the toxicity results in
river water (this work) and AFW (Heinlaan et al., 2008). In case of
both crustaceans there was a significant decrease in the toxicity of
both forms of CuO as well as CuSO4in natural water samples as
compared with the AFW (Table 2) and the decrease was most
remarkable for nanoCuO. Indeed, the D. magna EC50 values for
nanoCuO, bulk CuO and CuSO4increased by 50–140, 6.5–13 and
3–13-fold, and T. platyurus EC50 values by 55–130, 7.3–21 and
10–100-fold, respectively, depending on river water used for testing
(Table 2). Thus, the data obtained by biotests (Table 2) as well as
recombinant sensor bacteria (Fig. 1) both showed that natural
waters have higher mitigation effect for nanoCuO than bulk CuO.
It is well known that natural organic matter (mainly humic and
fulvic acids) is strongly complexing copper and reducing the
bioavailable copper ion concentrations (Allen and Hansen, 1996).
The relationship between dissolved organic matter in test media
and bioavailability of copper to aquatic organisms has been shown
previously by many investigators (De Schamphelaere et al., 2004a;
Kim et al.,1999; Kramer et al., 2004; Oikari et al.,1992; Hyne et al.,
2005; Santos et al., 2008). The data of the current study (Fig. 2A) are
coherent with data previously reported by Kramer et al. (2004) and
De Schamphelaere et al. (2004a). Statistical analysis of the data on
CuSO4toxicity for D. magna in river waters (Tables 1 and 2) revealed
significantdifferencesbetweenrivers withrelativelylow (rivers 1,2)
and high DOC levels (rivers 4, 5 and 6). Also, the differences
between D. magna EC50values for nanoCuO obtained for rivers 1
and 2 (lower DOC content) and rivers 5 and 6 (high DOC content)
were statistically significant (p < 0.05). A good correlation
Toxicity of nanoCuO and nanoZnO, their bulk forms and respective soluble salts to aquatic organisms.
Test mediaNanoCuO Bulk CuOCuSO4
Crustacean Daphnia magna, 48 h EC50(mean ? SD, mg metal/l)
2.6 ? 1.3
River 1 92.7 ? 12.4
River 2 149 ? 16.6
River 3160 ? 28.4
River 4200 ? 17.5
River 5224 ? 15.9
132 ? 19.7
844 ? 14.4
799 ? 19.2
1566 ? 34.0
1737 ? 0
0.07 ? 0.01
0.24 ? 0.03
0.38 ? 0.04
0.58 ? 0.18
0.81 ? 0.06
0.84 ? 0.03
0.92 ? 0.04
2.6 ? 1.04
3.3 ? 1.15
9.0 ? 0.28
1.7 ? 0.27
3.5 ? 0.30
2.8 ? 0.39
3.4 ? 1.56
7.1 ? 1.1
9.5 ? 1.8
12.0 ? 2.1
6.9 ? 0.55
16.2 ? 2.3
10.8 ? 1.4
1.4 ? 0.24
1.8 ? 0.32
2.0 ? 0.19
1.6 ? 0.14
2.5 ? 0.18
1.4 ? 0.18
1.7 ? 0.13
Crustacean Thamnocephalus platyurus, 24 h LC50(mean ? SD, mg metal/l)
1.7 ? 0.4
River 1152 ? 22.7
River 2217. ? 19.3
River 3 92.7 ? 8.8
River 4112 ? 0
River 5129 ? 23.7
River 690.3 ? 12.5
75.6 ? 4.5
593 ? 94.0
874 ? 25.2
1054 ? 42.8
1518 ? 49.9
1550 ? 42.6
546 ? 108.7
0.04 ? 0.02
0.41 ? 0.06
0.60 ? 0.06
2.1 ? 0.28
2.1 ? 0.32
1.1 ? 0.05
4.6 ? 0.18
0.14 ? 0.02
1.1 ? 0.23
6.0 ? 0.71
3.6 ? 0.74
1.5 ? 0.32
5.3 ? 0.50
1.4 ? 0.35
0.19 ? 0.03
1.9 ? 0.46
3.0 ? 0.37
2.1 ? 1.9
1.2 ? 0.16
0.22 ? 0.06
0.92 ? 0.11
1.6 ? 0.2
0.61 ? 0.26
0.75 ? 0.06
1.1 ? 0.40
1.7 ? 0.88
Protozoa Tetrahymena thermophila, 24 h EC50(mean ? SD, mg metal/l)
0.40 ? 0.12
0.46 ? 0.02
0.27 ? 0.03
9.4 ? 3.0
16.4 ? 0.41
12.4 ? 0.75
26.5 ? 2.52
27.1 ? 1.0
16.6 ? 0.91
12.0 ? 0.45
14.5 ? 1.09
7.1 ? 0.66
21.1 ? 1.37
18.6 ? 0.59
aL(E)C50values in artificial freshwater from Heinlaan et al. (2008).
bnd – not determined.
I. Blinova et al. / Environmental Pollution 158 (2010) 41–4743
(r2¼ 0.81) between the toxicity of nanoCuO to D. magna and DOC in
natural waters used as testing medium (Fig. 2B) is in agreement
with the assumption that toxicity of CuO NPs was mostly caused by
In all natural waters at EC50values of CuSO4for D. magna the
concentration of copper bioavailable to sensor bacteria was 0.07 ?
0.03 mg Cu/l, corresponding to the EC50value of CuSO4in AFW
(Table 2). However, not all copper available to sensor bacteria from
Bioavailable Cu, mg/l
Nominal Cu, mg/l
0.001 0.010.11 10 100100010000
Nominal Cu, mg/l
0.0010.01 0.11 10 100100010000
Nominal Cu, mg/l
Bioavailable Cu, mg/l
AFW D. magna
AFW T. platyurus
Nominal Zn, mg/l
Nominal Zn, mg/l
Nominal Zn, mg/l
Bioavailable Zn, mg/l
Bioavailable Zn, mg/l
Bioavailable Cu, mg/l
Bioavailable Zn, mg/l
Fig. 1. Nominal (x-axis) versus bioavailable Cu and Zn ions (y-axis) in six natural waters and artificial freshwater (AFW). Dissolved bioavailable metals were quantified by
recombinant Escherichia coli Cu-sensor (left panel) and Zn-sensor (right panel).
1015 20 2530 3540
48 h. EC 50, mg Cu/l
48 h. EC 50, mg Cu/l
Fig. 2. Effect of dissolved organic carbon (DOC) concentration in natural waters on toxicity of copper salts (A) and nanoCuO (B) to Daphnia magna. (A): 1 – this study; 2 – De
Schamphelaere et al. (2004a); 3 – Kramer et al. (2004). The natural waters were spiked with different copper salts (1 – CuSO4; 2 – CuCl2; 3 – CuCl2$2H2O).
I. Blinova et al. / Environmental Pollution 158 (2010) 41–47 44
bulk and nanoCuO proved bioavailable to crustacean: the concentra-
tions of bioavailable copper ions in bacterial assay exceeded signifi-
cantly the measured EC50in tests with D. magna (data not shown).
Several factors may be involved in this apparent discrepancy.
For example, assays were done in optimal respective temperature
conditions, i.e. sensor assay at 30?C and crustacean tests at 20?or
25?C. It is also possible that some of the copper complexes formed
(Apte et al., 2005). In addition, the ingestion and excretion of CuO by
crustaceans during the test may modulate the bioavailable fraction of
CuO. Indeed, the copper oxide accumulation in the gut (one of the
most crucial exposure routes for a particle-feeding organism) of both
crustaceans was clearly visible under the microscope as black gut
content (T. platyurus as an example; Fig. 3). Interestingly, the gut of
daphnids exposed either to sub-toxic concentrations of CuO (i.e. no
mortality observed) or toxic concentrations (>50% mortality) were
filled with visible CuO. This gives additional evidence that dissolved
CuO (copper ions) and not the gut-accumulated particles are main
contributors to the toxicity of nano and bulk CuO.
TEM analysis performed in AFWas well as in the gut of D. magna
clearly showed that differently from bulk CuO (data not shown), no
clear agglomerates of nanoCuO were observed in the gut lumen
(Fig. 4). Average nanoCuO particle size in the gut of D. magna –
31 ?12.8 nm remains similar with that in AFW – 30 ?10 nm (Fig. 4).
The test results of T. platyurus for CuSO4in natural water showed
(Table 2) analogous trends as in the test with D. magna with the
exception of river 5 (discussed below). However, in contrast to
D. magna, dependence between toxicity of nanoCuO towards crus-
tacean T. platyurus and DOC concentrations in test water was not
revealed (the differences between EC50values for all rivers except for
river 2 were not statistically significant). It should be noted that
though both D. magna and T. platyurus belong to planktonic crusta-
ceans, their sensitivity towards the same substances may differ (Bli-
nova, 2004). For example, T. platyurus was much more sensitive than
D. magna towards ammonium ions (I. Blinova, unpublished data).
is more tolerant to copper and zinc than other cladocerans. A greater
ability of D. magna to adapt to toxic stress may be linked to the fact
that differently from other cladocerans D. magna has adapted to
withstand wider fluctuations in environmental conditions, e.g.,
hypoxia, turbidity, ionic strength (Koivisto et al.,1992). Furthermore,
the uptake routes of soluble metals by crustaceans is directly from
water through the body surface (Krantzberg and Stokes, 1988; Rob-
inson et al., 2003). In case of NPs, the toxic effect may also depend on
the mechanical adhesion of NPs on the organism’s surface, which in
turn is related to the body size and the structure of exoskeleton.
Indeed, adhesion of aggregates of NPs to the exoskeleton of crusta-
ceans has been noted by Baun et al. (2008) and also observed (under
the microscope) in our experiments. As a result, the combination of
abovementioned factors leads to different probability of survival of
D. magnaand T. platyurus inthe same conditions.This canexplainthe
lack of relationship between DOC concentration and toxicity of
nanoCuO to T. platyurus and comparatively high toxicity of CuSO4in
river 5 (high ammonium concentration).
3.4. Toxicity of zinc compounds in natural waters to crustaceans
Differently from Cu compounds, the change in toxicity of Zn
compounds in natural waters as compared with AFW was less
remarkable (Table 2). The D. magna EC50for nanoZnO, bulk ZnO and
ZnSO4increased only by 1–1.75,1–2.3 and 0.7–3.6-fold, respectively
and in most cases the changes were not statistically significant.
These results are in accordance with the data obtained with Zn-
sensor bacteria showing only small differences in bioavailable Zn-
ion concentrations in AFW and river waters (Fig. 1). Thus, it can be
concluded that the toxicity of nano and bulk ZnO in natural waters
was due to the solubilised Zn-ions.
The other crustacean (T. platyurus) was in general more than
10 times more sensitive than D. magna towards Zn compounds in
AFW but the difference between L(E)50 values was not so remark-
able in natural waters (Table 2). To explain the reasons for higher
mitigating effect of natural waters for T. platyurus than for D. magna
further studies are needed.
It is obvious that natural waters used in the current study
remarkably affected the bioavailability of copper but not zinc. Also,
there was no clear relationship between DOC and toxicity of Zn
compounds (Tables 1 and 2). The results published byother authors
also indicate that in modulation of the toxicity of zinc ions the DOM
is not so important, as, for example, calcium (Clifford and McGeer,
2009; De Schamphelaere et al., 2004b; Hyne et al., 2005). However,
we have not revealed any relationship between toxicity of Zn
compounds and Ca, probably due to the masking effect by other
Fig. 3. CrustaceanThamnocephalus platyurus under the light microscope. Arrows mark the presence (A: exposed organism) or absence (B: control) of black CuO nanoparticles in the gut.
I. Blinova et al. / Environmental Pollution 158 (2010) 41–47 45
3.5. Toxicity of Zn and Cu compounds to T. thermophila
Toxicity of nano- and bulk ZnO formulations, ZnSO4$7H2O
and CuSO4to T. thermophila in standard test conditions and three
river waters was compared (Table 2). Toxicity of ZnSO4$7H2O
decreased in natural waters (p < 0.05), but for nanoZnO the
statistically significant decrease was observed only in the water of
river 5. In case of bulk ZnO, however, the toxicity in river waters
even increased (p < 0.05).
Differently from crustaceans, toxicity of CuSO4to protozoa in
river waters 2 and 4 was not statistically different from EC50values
in standard test conditions (Table 2). It could be assumed that
binding of copper ions by food particles (added to test medium at
the beginning of the assay) was masking the mitigating effect of
organic matter in river waters.
with OD measurements impairing the reliability of EC50determi-
nation. Therefore, no EC50values for copper oxides are presented in
Table 2. However, visual microscopic observations of the protozoa
after exposure to the sub-toxic concentrations of nanoCuO showed
important for investigation of the fate of nanoCuO in the aquatic
This paper shows that ecotoxicological tests are indispensable
tools for hazard evaluation of synthetic NPs as they integrate both
the harmful and mitigating effects and show the net influence of
the tested compounds in the given experimental conditions. The
use of natural water as test medium in ecotoxicological assays can
increase the predictive power of these tests for the environmental
This study showed the remarkable potential of natural water to
mitigate the toxic effects of CuO NPs but not ZnO NPs. In addition,
as in standard test conditions, the toxic effect of CuO and ZnO NPs
in natural waters was mainly due to dissolved metal ions. Thus, to
understand the mechanisms of ecotoxicological action of metal
oxides NPs and its ecological consequences, solubility and specia-
tion are the crucial aspects contrarily, for example, to manufactured
carbon NPs, where size and aggregation seem to be the key factors.
Lastly, this study confirms that biology and physiology of the
target organisms may be of primary importance in transfer of
synthetic NPs along the food-web and environmental compart-
ments and, thus, also important in risk assessment. For example,
interaction and adhesion of NPs on the surface of living organisms
and accumulation in particle-ingesting organisms should be taken
into account, when behavior and transfer of NPs in the water
ecosystems is investigated.
This work was supported by Estonian targeted funding project
SF0690063s08, Estonian Science Foundation Projects 8066, 6974,
6956, 7686 and Maj and Tor Nessling project 2008416 and Archi-
medes scholarship for M. Heinlaan. We thank Dr. B. Arbeille and
Prof. G. Prensier (University of Tours, France) for electron micros-
copy and Prof. H.-C. Dubourguier for critical comments.
1-19 20-2930-39 40-4950-59 60-69 70-79 80-89 90-99
particle size range, nm
% of total
1-19 20-29 30-39 40-49 50-59 60-69 70-79 80-89 90-99
particle size range, nm
% of total
Fig. 4. Nanoparticles of CuO in artificial freshwater (A: TEM; B: particle size distribution) and in the gut of Daphnia magna (C: TEM; D: particle size distribution).
I. Blinova et al. / Environmental Pollution 158 (2010) 41–4746
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