Next Article in Journal
Physicochemical Properties and Storage Stability of Food Protein-Stabilized Nanoemulsions
Next Article in Special Issue
Pesticide Encapsulation at the Nanoscale Drives Changes to the Hydrophobic Partitioning and Toxicity of an Active Ingredient
Previous Article in Journal
Wave Propagation of Porous Nanoshells
Previous Article in Special Issue
Cytotoxic, Genotoxic, and Polymorphism Effects on Vanilla planifolia Jacks ex Andrews after Long-Term Exposure to Argovit® Silver Nanoparticles
Article

Combined Effects of Test Media and Dietary Algae on the Toxicity of CuO and ZnO Nanoparticles to Freshwater Microcrustaceans Daphnia magna and Heterocypris incongruens: Food for Thought

1
Laboratory of Environmental Toxicology, National Institute of Chemical Physics and Biophysics, Akadeemia tee 23, 12618 Tallinn, Estonia
2
Department of Materials and Environmental Technology, Tallinn University of Technology, Ehitajate tee 5, 19086 Tallinn, Estonia
3
Estonian Academy of Sciences, Kohtu 6, 10130 Tallinn, Estonia
4
Institute for Medical Research and Occupational Health, Ksaverska cesta 2, 10001 Zagreb, Croatia
5
Institute of Agricultural and Environmental Sciences, Estonian University of Life Sciences, Fr.R. Kreutzwaldi 5, 51006 Tartu, Estonia
*
Authors to whom correspondence should be addressed.
Nanomaterials 2019, 9(1), 23; https://doi.org/10.3390/nano9010023
Received: 5 December 2018 / Revised: 19 December 2018 / Accepted: 20 December 2018 / Published: 25 December 2018
(This article belongs to the Special Issue Toxicity and Ecotoxicity of Nanomaterials)

Abstract

The chemical composition of the test medium as well as the presence of algae (microcrustaceans’ food) affects the bioavailability and thus the toxicity of metal nanoparticles (NP) to freshwater microcrustaceans. This study evaluated the effect of the addition of algae (Rapidocelis subcapitata at 7.5 × 106 cells/mL) on the toxicity of CuO (primary size 22–25 nm) and ZnO NP (10–15 nm) to planktic Daphnia magna and benthic Heterocypris incongruens in artificial (mineral) and natural freshwater (lake water). The toxicity of ionic controls, CuSO4 and ZnSO4, was evaluated in parallel. When algae were added and the toxicity was tested in mineral medium, 48 h EC50 of CuO and ZnO NP to D. magna was ~2 mg metal/L and 6-day LC50 of H. incongruens was 1.1 mg metal/L for CuO and 0.36 mg metal/L for ZnO. The addition of algae to D. magna test medium mitigated the toxicity of CuO and ZnO NP 4–11-fold when the test was conducted in natural water but not in the artificial freshwater. The addition of algae mitigated the toxicity of CuSO4 (but not ZnSO4) to D. magna at least 3-fold, whatever the test medium. In the 6-day H. incongruens tests (all exposures included algae), only up to 2-fold differences in metal NP and salt toxicity between mineral and natural test media were observed. To add environmental relevance to NP hazard assessment for the freshwater ecosystem, toxicity tests could be conducted in natural water and organisms could be fed during the exposure.
Keywords: aquatic toxicology; nanomaterials; water flea; ostracod; zooplankton; feeding; natural waters aquatic toxicology; nanomaterials; water flea; ostracod; zooplankton; feeding; natural waters

1. Introduction

Nanoparticles (NP), defined as particles with at least one dimension in the range of 1–100 nm [1], may pose a hazard to biota when released into the environment. Assessing the environmental hazards of manufactured NP has been a real challenge for the scientific community due to the unique physicochemical properties of NP. Uncertainties regarding the NP behaviour during the toxicity testing and the questionable ecological relevance of respective experimental setups (e.g., unnatural test conditions and limited number of test species) of the standardised laboratory tests complicate the extrapolation of the laboratory test results to the real ecosystem [2,3]. As safety regulations for nanomaterials (consisting of ≥50% of NP) [1] are still under development, new knowledge on the effects of testing conditions on NP behaviour and toxicity is needed for the correct interpretation of the laboratory test results [4].
Compared to other traditional aquatic test species, microcrustaceans have been shown to be especially sensitive to metal-based NP [5]. However, the considerable variety of the toxicity values between different test species and also crustaceans species can be explained not only by species’ sensitivity pattern but partly or even predominantly by the test medium [2,6]. The chemical composition of the test medium and the feeding of aquatic test organisms during the exposure are the main factors which may affect metal bioavailability. ZnO and CuO NP reach the environment mainly due to their use in cosmetics, coatings, paints and pigments [7,8]. Modelling results suggest that ZnO NP pollution in some freshwater bodies may have already reached toxic concentrations while CuO NP pollution can impose localised hazards [9]. In freshwater, CuO and ZnO NP induce toxicity mostly via bioavailable toxic metal ions [5,10,11,12,13,14], causing ionic and osmoregulatory disturbances [13,15]. The dissolved organic matter in natural waters plays an important role in mitigating the toxic effects of not just metal ions [16,17,18], but also of metallic NP [19,20,21]. Water hardness is another important NP toxicity mitigator, promoting NP aggregation, decreasing dissolution, and allowing outcompeting of the metal ions at the biotic uptake sites [22,23,24].
Some standardised (sub)chronic microcrustacean toxicity tests (ISO 14371 [25], OECD 211 [26]) require adding high concentrations of algae in the test medium in order to feed the test organisms. Acute tests such as OECD 202 [27] often do not. As eutrophication and algal blooms are becoming more and more common in freshwater lakes [28], the addition of algae in the test medium helps to mimic environmental conditions. On the other hand, the toxicity results obtained in the presence of algae may not be valid for the periods outside of algal blooms, when algal concentrations may be up to 400 times lower compared to those used in OECD 211 tests [29]. ISO 14371 freshwater sediment toxicity testing with ostracods requires using even more elevated algal concentrations (7.5 × 106 cells/mL) that exceed even the highest possible algal concentrations in nature [30,31].
This study compares the toxicity of two metal-based NP (CuO and ZnO) to planktic and benthic crustaceans. Daphnia magna is the most common aquatic invertebrate used for NP toxicity testing [5], while Heterocypris incongruens is a relatively novel alternative species to conventional sediment toxicity test organisms [32]. H. incongruens is especially relevant for the safety evaluation of NP, which may impose elevated risk to sediment biota by settling quickly in the waterbodies [9,33]. However, data on NP toxicity to benthic organisms are lacking [34,35]. The artificial freshwaters recommended for the standardised D. magna and H. incongruens toxicity tests have different ionic contents and do not contain dissolved organic matter. In order to add environmental relevance to the experiments, additional tests in natural waters with different organic matter contents were carried out. In addition, algae were added to some of the D. magna tests that normally do not require feeding the test organisms. The different combinations of the test media and the presence or absence of dietary algae in the toxicity experiments will give additional information on the environmental relevance of CuO and ZnO NP toxicity results from the laboratory tests. To our best knowledge, no studies have explored the combined effects of dietary algae and test media with different nutrient profiles or compared these effects for metal NP and respective soluble salts.

2. Materials and Methods

2.1. Chemicals

Nanoparticles used in this study were CuO NP (NNV-011; Intrinsiq Materials; powder form) and ZnO NP (NNV-003; Nanogate; powder form) obtained from the EU FP7 NanoValid project (“Developing of reference methods for hazard identification, risk assessment and LCA of engineered nanomaterials”, 2011−2015). CuSO4∙5H2O and ZnSO4∙7H2O (both Alfa Aesar) were used as ionic controls. CuO and ZnO NP stock suspensions (à 20 mL) were prepared at the concentration of 5 g metal/L in MQ water (MilliQ, >18 MΩ cm, Merck Millipore, Darmstadt, Germany) as has been previously described [36]. Stock suspensions were probe-sonicated for 4 min at 20 kHz (40 W) at continuous mode using 450 Ultrasonifier (Branson Ultrasonics Corporation, Danbury, CT, USA) after preparation and used for up to 4 weeks. The primary sizes of the CuO NP and ZnO NP (both uncoated) according to manufacturers’ data were 22−25 nm and 10−15 nm, respectively, which was in agreement with the TEM analysis showing 24.5 nm and 13.6 nm particle sizes.

2.2. Test Media

OECD 202 artificial freshwater (AFW) [27] and US EPA moderately hard reconstituted water (MHW) [37] were used as standard exposure media for Daphnia magna and Heterocypris incongruens, respectively. In addition, water from two Estonian lakes, Lake Ülemiste and Lake Raku, was used (Table 1). Lake Ülemiste is a natural eutrophic lake while Lake Raku is an artificial sandpit lake with a similar phosphorus concentration but lower dissolved organic carbon concentration and hardness. Lake water was collected between September and March and filtered through Millipore nitrocellulose filters (pore size 0.45 µm) and stored in the dark at 4 °C. The chemical analysis of lake waters was performed by an accredited laboratory (Tallinna Vesi Laboratories). The speciation of metal salts in test media was calculated using Visual MINTEQ version 3.1 [38].

2.3. Physico-Chemical Characterisation of Nanoparticle Suspensions

Dynamic light scattering (DLS) and electrophoretic light scattering (ELS) methodology was employed to measure the hydrodynamic diameter, zeta potential and polydispersity index of NP using ZetaSizer Nano ZS (Malvern Instruments, Malvern, UK) equipped with the 4.0 mW 633 nm laser (Model ZEN3600; 173° angle). NP samples (10 mg metal/L) were incubated for 0 h, 48 h, and 144 h at the toxicity test conditions and vortexed before the measurement.
The dissolution of NP was measured as initially described elsewhere [39]. Briefly, NP suspensions were prepared at 10 mg metal/L and incubated for 48 h (the duration of the Daphnia magna acute toxicity test) in OECD 202 AFW, Lake Raku water and Lake Ülemiste water with (7.5 × 106 cells/mL) and without added algae. The samples were then ultracentrifuged at 362,769 × g for 30 min (duration of the whole cycle 60 min). For metal recovery control, salt solutions were always used in parallel. Metal concentrations in the supernatants were measured using graphite furnace atomic absorption spectroscopy (GF-AAS) analysis in the accredited laboratories of the Institute of Medical Research and Occupational Health (Zagreb, Croatia) and Estonian University of Life Sciences (Tartu, Estonia).

2.4. Test Formats of Bioassays

Three different experimental setups in three different test media were used to test the potential toxicity of NP and soluble salts (Table 1). The 48 h Daphnia magna acute immobilisation tests were carried out in accordance with the OECD 202 testing guidelines [27]. Briefly, D. magna neonates were pre-fed before the test and exposed to NP or metal salts at 21 °C in the dark. After 48 h of exposure, immobilised daphnids were counted. In addition to the standard D. magna test format, a modified one was applied, which included use of natural test media and/or addition of green algae Raphidocelis subcapitata at the concentration of 7.5 × 106 cells/mL (the concentration required in the Ostracodtoxkit [30]). Tests were repeated 2 to 8 times with 2 to 4 technical replicates each including 5 daphnids. Based on the immobilisation results, 50% effect concentrations (EC50) were calculated.
Six-day subchronic Heterocypris incongruens toxicity testing was performed with a modified version of the OSTRACODTOXKIT F [30] (similar to ISO 14371 guidelines [25]) test procedure. Briefly, ostracod neonates (<24 h) were exposed to the NP or salts in the test media at 25 °C in the dark. Ostracods were pre-fed before the test and food was added to the test media (7.5 × 106 cells of R. subcapitata/mL). After 6 days, mortality was recorded. Also, growth inhibition was calculated based on the length measurements under the dissection microscope (Olympus IMT-2, CellB software, Electro Optics, Cambridge, UK) before and after the incubation. As a modification, standard sand was not added to the test to exclude the metal adsorption on sand, which can significantly mitigate metal toxicity [40]. Each test was repeated 2 to 4 times with 2 technical replicates (10 ostracods in each) of all the tested concentrations to calculate EC50 values.

2.5. Statistical Analysis

MS Excel macro REGTOX [41] based on non-linear regression was used to calculate EC50 values. The “optimal” EC50 values were obtained from the log-normal model. Statistically significant differences between EC50 values were determined based on the 95% confidence intervals provided by the REGTOX program. The statistically significant differences between metal recovery results were also determined by the absence of overlap between the 95% confidence intervals.

3. Results

3.1. Behaviour of CuO and ZnO Nanoparticles in the Test Media

3.1.1. Stability of Nanoparticle Suspensions

DLS and ELS data (Table S1) showed the low stability of both CuO and ZnO NP suspensions in all the test media (Table 2). Indeed, in lake water, the zeta potential (ζ) measured at the nominal concentration of 10 mg metal/L ranged from −15 to −19 mV, indicating that suspensions were relatively unstable. In both artificial freshwaters, suspensions were also unstable (ζ values ranged from −2 to −3.4 mV) [42].
The hydrodynamic diameters (Dh) of CuO and ZnO NP as well as the polydispersity index (pdi) increased in time in all the test media, indicating intensive aggregation of NP during the exposure (Table S1). After 48 h incubation, the Dh values for CuO and ZnO NP were on average 4 times greater in artificial freshwaters compared to lake waters. Less intensive aggregation of the studied NP in organics-containing natural water probably occurred due to the NP-stabilising effect of dissolved organic matter (DOM) [43,44]. ZnO NP were more stable in Lake Ülemiste water (with higher DOM content) but the stability of CuO NP in both lake waters was similar. This may be explained by the counteraction of DOM concentration and water hardness [23,43,45], with both parameters being higher in Lake Ülemiste water.
Along with aggregation, sedimentation of NP has been shown to increase by high ionic strength and decrease by high organics content [45,46]. Phosphates have also been shown to stabilise CuO NP [47] but at higher concentrations than present in the natural waters of this study. A higher aggregation of NP in artificial waters compared to natural waters indicates the facilitated sedimentation of NP [44] in the former (Table S1). Sedimentation of algae also occurred in CuO and ZnO NP suspensions in mineral test medium but not in organics containing test medium (Figure S1). By contrast, algae in CuSO4 solution settled in all test media (slightly less in AFW) and algae in ZnSO4 solution showed comparable moderate sedimentation in all the test media.
Similarly to the homoaggregation of NP, the heteroagglomeration of NP and algae potentially occurred in artificial test media [19,48,49]. High ionic strength and low DOM concentrations can facilitate homoaggregation as well as heteroagglomeration [50,51,52]. Compared to dispersed NP, homoaggregated NP themselves enhance algae sedimentation [53], which at the same time is dependent on both algal [54] as well as on NP concentrations [50]. Both these parameters were high in the experimental setup, but the entrapment of algae by CuO NP at lower concentrations (2 mg/L) has also been previously shown [48].

3.1.2. Dissolution of CuO and ZnO Nanoparticles in the Test Media

The dissolution of NP was evaluated by measuring levels of soluble metal forms released in the NP suspension (10 metal mg/L) after 48 h and 6 days of incubation in the test media (see Section 2.3). The total concentration of Zn and Cu in the supernatants, obtained after the ultracentrifugation of NP suspensions, represents the proportion of dissolved metal species (percentage of nominal concentration) in the test media (Table 3). The soluble forms of metal can be inorganic and organic complexes and free metal ions [56] (Table S2). The concentration of Cu dissolved from CuO NP was ≤2% of the nominal concentration in all the test media (Table 3), indicating a very low dissolution of CuO NP compared to ZnO NP. CuO NP dissolution is usually the highest in water characterised by the lowest pH, DOM and hardness values [47,57], but sometimes the link between these characteristics and dissolution is less straightforward [44]. The presence of algae and incubation duration did not have a significant effect on the CuO NP dissolution or Cu recovery from CuSO4 solutions. The Cu recovery from CuSO4 solutions was only 30−45% in all the test media (Table 3), indicating precipitation/adsorption as was discussed in our earlier work [58]. Visual MINTEQ modelling suggested the precipitation of Cu as tenorite in all the test media (Table S2), which implies that the release of copper ions may be underestimated due to speciation effects and subsequent metal recovery. Accordingly, a common term, “metal recovery”, will be used to refer to both metals recovered from NP dissolution experiments as well as from metal recovery control experiments (with metal salts) in the following discussion. Speciation effects were further demonstrated by the parallel analysis of metals in MQ water with the lowest pH value, which showed a significantly higher dissolution of CuO NP (6.9%) and Cu ion recovery (84%) from CuSO4 than in any of the exposure media (Table 3). The addition of algae (7.5 × 106 cells of R. subcapitata/mL) increased the Cu recovery from CuSO4 by 3 to 12% in all test media after 48 h incubation. However, this increase was statistically significant only in lake waters (Table 3).
The dissolution of ZnO NP was 21−25% depending on the test media, while Zn recovery from Zn salt was 90−102% (Table 3) in all the media, and these values were comparable to Zn recovery in MQ water. ZnO NP dissolution has previously been shown to be high at a variety of ionic strength, pH, and DOM concentration values [57]. The presence of a high concentration of humic acids increases and high medium hardness decreases the ZnO NP dissolution [59]. However, we did not observe similar dissolution behaviour. The addition of algae increased the metal recovery from ZnO NP by 20−36% after 48 h of incubation, but the increase was significant only in AFW (Table 3). According to VisualMINTEQ modelling, the precipitation of Zn as hydrozincite occurred in all the test media (Table S2).

3.2. Toxicity of Cu and Zn Compounds to Daphnia magna and Heterocypris Incongruens

3.2.1. Toxicity of Cu Compounds to D. magna

In D. magna acute toxicity tests without algae, lake waters significantly mitigated the toxicity of both CuO NP (up to 18-fold) and Cu salt (up to 4-fold) compared to AFW. The toxicity of Cu compounds (especially of CuO NP), was significantly lower in Lake Ülemiste than in Lake Raku water (Table 4). These results are in accordance with earlier published data on toxicity of other types of CuO NP and CuSO4 in natural waters with different DOM concentrations [20]. The direct link between Cu recovery in different media and toxicity was not revealed. This shows that not all dissolved copper species are equally bioavailable to microcrustaceans.
The addition of algae mitigated CuO NP toxicity in lake waters (5 to >10-fold) but not in AFW. Similar to the tests without algae, NP toxicity was lower in Lake Ülemiste water compared to Lake Raku water. The toxicity of Cu salt was mitigated more in AFW (8-fold) compared to lake waters (3-fold) by the addition of alga. Earlier studies have also shown that the addition of algae (Chlorella) to mineral test medium decreases Cu toxicity to a lower level than in organics-containing media (without algae) [60].

3.2.2. Toxicity of Zn Compounds to D. magna

The effect of lake waters on the toxicity of ZnO NP and Zn salt to D. magna was quite different from Cu compounds: a 3‒4-fold increase in toxicity was observed compared to AFW in acute tests without algae (Table 4). Humic acids can increase ZnO NP toxicity to daphnids while fulvic acids [19] slightly mitigate Zn toxicity [61]. Dissolved organic matter also has much less affinity to Zn ions compared to Cu ions (Table S2). The water hardness, which mitigates Zn toxicity [22], was the highest in AFW, potentially explaining the higher Zn toxicity in natural water with lower hardness. Natural waters also had slightly higher pH which can increase Zn toxicity [61]. Despite the 4-fold difference in Zn recovery upon ultracentrifugation (in the absence of algae), toxicity of ZnO NP and ZnSO4 to D. magna was comparable. This may be explainable by the enhanced dissolution of the metal NP upon contact with the living cell [18].
The presence of algae slightly but significantly reduced the toxicity of Zn compounds to D. magna in lake waters, but did not change or even increased the toxicity (in case of ZnSO4) in AFW (Table 5). The toxicity of ZnSO4 was significantly higher in Lake Raku water compared to Lake Ülemiste water in the experiments with algae, possibly due to the difference in water hardness [59].

3.2.3. Toxicity of Cu and Zn Compounds to H. incongruens

Similar to tests with D. magna, CuO NP were significantly less toxic than CuSO4 in the 6-day ostracod toxicity tests (Table 4), but there was only one case where toxicity was significantly affected by the test medium. The CuSO4 was up to 2-fold less toxic in Lake Ülemiste water compared to MHW and Lake Raku water, probably due to the highest DOM content being in Lake Ülemiste water (Table 2). Sublethal (mortality <20%) concentrations of CuSO4 enhanced the growth of ostracods (up to 41%) in all test media (Table S3).
As for Cu-compounds, the chemical composition of the test media had very little effect on the toxicity of Zn compounds to H. incongruens (Table 4). There were no statistically significant differences between the toxicity of ZnO NP and Zn salt. The toxicity of ZnO NP was significantly lower in Lake Ülemiste water compared to MHW, but no medium effect was observed for Zn salt. Sublethal concentrations of ZnSO4 increased the body length of ostracods by 17% in MHW (Table S3), probably due to the absence of Zn in this water.
Despite being a less common test organism compared to D. magna, the data on H. incongruens sensitivity to metals obtained in this study were consistent with those available in the literature. The previously published LC50 values for ZnSO4 and CuSO4 were in the range from 0.7 to 12 mg Zn/L and from <0.3 to 0.9 mg Cu/L, respectively, despite the fact that reference sediment with possible toxicity mitigating effect was applied [62,63]. Surprisingly, Zn toxicity was the lowest (LC50 12 mg Zn/L) in distilled water as exposure medium [63].

4. Discussion

4.1. Combined Effect of the Media and Feeding on D. magna Toxicity Test Results

The effect of the addition of algae on the toxicity of copper and zinc compounds using modified formats of D. magna acute immobilisation testing (OECD 202) are summarised in Table 5. The toxicity mitigating effect of the added algae (as seen in both metal NP exposure in lake water and CuSO4 in all the test media) was anticipated, because feeding on organic compounds and the presence of extracellular polymeric substances of some algae have been shown to mitigate metal toxicity [64,65]. In addition, uncontaminated algae may help clear the gut of daphnids of metal NP [66]. The lack of effect and even the increased toxicity of the tested compounds in the presence of algae, as seen for CuO and ZnO NP in artificial freshwater and for Zn salt, was unexpected (Table 5).
The addition of algae could potentially change the toxicant exposure for the test organism. One possible explanation may be the internalisation of metals by algae, leading to foodborne metal exposure [62,66,67,68] especially in lake Raku water with lower ionic strength [22]. The concurrent sedimentation (or heteroagglomeration) (see 3.1.1) of NP and algae that was observed in AFW could have increased their simultaneous uptake due to daphnids turning to bottom-feeding. Almost no clear correlations were observed for toxicity and changes in the dissolution of either the metal NP or metal salt upon the addition of algae in the test medium (Table 5). As an exception, the absence of a mitigating effect of algae on ZnO NP toxicity in AFW can partly be due to the increased Zn recovery in the presence of algae (Table 5). Altogether, the effects of addition of algae cannot be considered analogous to the addition of dissolved organic matter in the test medium in case of CuO and ZnO NP in AFW.
The use of microalgae for feeding the crustaceans during NP exposure increases the environmental relevance of laboratory testing. The effect of algae concentrations on the interactions between NP and algae is yet to be determined, but high algal concentrations may lead to the heteroagglomeration of algae and NP. According to Stevenson et al. [29], using environmentally relevant algae concentrations in a chronic Daphnia pulicaria exposure to nano Ag significantly increased the adverse effects compared to the normal feeding rate recommended by the D. magna chronic toxicity test (OECD 211) guidelines.

4.2. Differences between CuO and ZnO Nanoparticle Toxicity to D. magna and H. incongruens

Compared to D. magna test with algae, the toxicity of Zn and Cu salts and ZnO NP to ostracod was only slightly higher in artificial freshwaters, but significantly higher (more than 30-fold) for CuO NP in lake waters (Table 4). The differences for Zn and Cu salts, ZnO NP and CuO NP in AFW may be explained by the 3-fold-longer test duration (Table 1) and the different bioavailability of metal species formed in the two different artificial test media (OECD AFW and US EPA MHW, see Table 2). For example, the toxicity of CuO and ZnO NP to D. magna has been shown to be higher in MHW compared to OECD AFW despite the higher dissolution in OECD AFW [69].
The negligible effect of the natural water on CuO toxicity to ostracod can be partially explained by the non-permanent nature of changes in metal bioavailability, induced by water parameters such as hardness [70,71] and the presence of humic acids [45]. For instance, the mitigative effects can last long enough to be evident in D. magna exposure (48 h) but not in H. incongruens exposure (6 days). The behaviour of the test organisms can also influence their exposure to toxicants. The planktic species D. magna is mostly exposed to soluble or suspended metal species, whereas benthic H. incongruens is more exposed to settled agglomerates of metal compounds. Agglomeration and sedimentation does not necessarily mean that metal compounds are less bioavailable to the test organisms [72]. Ostracods have been shown to be a more vulnerable organism to metal-polluted river sediment compared to water fleas despite the latter ones being more sensitive to metals in an exposure without sediment [73]. As concurrent sedimentation of NP and algae occurred in artificial freshwaters, daphnids may have turned to bottom feeding to access the settled algae. As a result, both daphnids and ostracods could have had similar exposure to the toxicants, feeding on settled NP aggregates along with settled algae. As algae remained suspended in natural waters, daphnids could feed in the upper layers of the test vessel, avoiding potentially higher metal NP concentrations on the bottom of the test vessel.
The effect of the addition of algae in D. magna and H. incongruens test medium cannot be directly compared in this study as H. incongruens toxicity tests without the addition of algae were not conducted to avoid starvation of the test organisms. Based on the literature, the effect of the addition of algae on toxicity of metals in organics-free test medium seems to be similar for both daphnids and ostracods despite the differences seen in natural waters for CuO as explained in the previous paragraph. Toxicity tests carried out for 48‒96 h with adult ostracods Cypris subglobosa and Stenocypris major in tap water and well water with no addition of algae (and no sediment) resulted in lower LC50 for Cu (0.025 to 0.055 mg Cu/L) and slightly to much higher LC50 for Zn (1.2 to 85 mg Zn/L) compared to the results obtained in this study, indicating the toxicity mitigating effect of algae on Cu and toxicity enhancing effects on Zn in artificial freshwater [35,74].

5. Conclusions

It is well known that the chemical composition of toxicological test media affects the bioavailability and thus the toxicity of metal nanoparticles (NP) to aquatic test organisms. However, it is poorly understood to what extent the addition of algae—food for microcrustaceans—into the test medium modulates the toxic effect. This aspect must be addressed since, in the (sub)chronic microcrustacean toxicity tests, algae are added as food by default. In this study, the combined effects of artificial versus natural test media and addition of algae (Raphidocelis subcapitata at 7.5 × 106 cells/mL) on the toxicity of CuO and ZnO NP to planktic Daphnia magna and benthic Heterocypris incongruens was evaluated in standardised and modified test formats. Natural freshwater and addition of dietary algae in 48 h D. magna exposure were used as modifications.
  • Subchronic (6 day) H. incongruens LC50 for CuO NP was 1.1 mg Cu/L, and 0.22 mg Cu/L for CuSO4, in US EPA mineral water. For both ZnO NP and ZnSO4, the respective 6-day LC50 was 0.36 mg Zn/L. For comparison, upon the addition of dietary algae in mineral medium, 48 h EC50 of CuO and ZnO NP for D. magna was ~2 mg metal/L;
  • Compared to standard mineral media, natural freshwater mitigated CuO NP toxicity (4–18-fold) and increased ZnO NP toxicity (3–4-fold) for D. magna. For Cu and Zn salts, the toxicity change followed the same pattern with 3–4-fold mitigation and an increase in natural water. In H. incongruens tests (all including algae), toxicity was mitigated only up to 2-fold (CuO NP and Cu salt) or remained the same (Zn compounds) in natural water;
  • Upon the addition of algae to D. magna for 48 h in OECD mineral medium, no toxicity mitigating effect was recorded for CuO NP, possibly due to the sedimentation of algae and NP. CuSO4 48 h EC50, however, decreased 8-fold. For ZnO NP and ZnSO4, the added algae resulted in comparable or even increased (ZnSO4) toxicity;
  • Algae in natural freshwater mitigated both CuO NP (5 to >10 fold) and Cu salt (3-fold) toxicity. The toxicity of ZnO was also significantly reduced (4-fold) but Zn salt toxicity remained unchanged.
According to the results of the modified D. magna and H. incongurens test formats, toxicity data from standardised acute/subchronic exposures may be overestimated for Cu-compounds and underestimated for Zn-compounds for eutrophic freshwaters during algal blooms. Also, our experiments once more demonstrated that the extrapolation of toxicity values obtained using planktic test species to other groups of aquatic microcrustaceans (e.g., benthic) may lead to significant mistakes in the environmental hazard evaluation, especially in the case of metal-based NP.

Supplementary Materials

The following are available online at https://www.mdpi.com/2079-4991/9/1/23/s1, Table S1: Characterisation of CuO and ZnO nanoparticle suspensions at 0 h, 48 h, and 6 days in five different test media. Table S2: Percentage of soluble and solid fraction predicted by Visual MINTEQ simulation results. Tabel S3. Change in H. incongruens growth (%) at the end of the 6-day experiment at sublethal concentrations. Figure S1: Examples of typical sedimentation of algae after 6 days of incubation with CuO and ZnO NP and respective soluble salts at 10 mg metal/L.

Author Contributions

Conceptualization, M.M. and I.B.; Data curation, M.M.; Formal analysis, M.M.; Funding acquisition, A.K. and M.H.; Investigation, M.M., I.V.V., B.P., K.O. and M.H.; Methodology, M.M. and I.B.; Project administration, M.M.; Resources, I.B. and M.H.; Software, M.M.; Supervision, I.B. and M.H.; Validation, M.M., I.B. and M.H.; Visualization, M.M.; Writing—original draft, M.M.; Writing—review & editing, M.M., I.B., A.K., I.V.V., B.P., K.O. and M.H.

Funding

This research was funded by Estonian Ministry of Education and Research, grants number ETF9347, IUT23-5 and PUT1512 (Personal Research Funding Grant to Margit Heinlaan, "Evaluation of the potential hazardous effects of microplastic to marine and freshwater zooplankton") and by ERDF project Centre of Technologies and Investigations of Nanomaterials (NAMUR+), project number 2014-2020.4.01.16-0123.

Acknowledgments

This work has been partially supported by ASTRA “TUT Institutional Development Programme for 2016-2022” Graduate School of Functional Materials and Technologies” (2014-2020.4.01.16-0032) and by “Center of Excellence” project TK134 “Emerging orders in quantum and nanomaterials (1.08.2015−31.08.2023)”. We thank AS Tallinna Vesi for providing access to Lake Ülemiste for water collection and chemical analysis results for Lake Ülemiste and Lake Raku water.

Conflicts of Interest

The authors declare no conflict of interest. The funders had no role in the design of the study; in the collection, analyses, or interpretation of data; in the writing of the manuscript, or in the decision to publish the results.

References

  1. European Commission. Communication (COM(2012) 572 Final) from the Commission to the European Parliament, the Council and the European Economic and Social Committee: Second Regulatory Review on Nanomaterials; European Commission: Brussels, Belgium, 2013. [Google Scholar]
  2. Holden, P.A.; Gardea-Torresdey, J.L.; Klaessig, F.; Turco, R.F.; Mortimer, M.; Hund-Rinke, K.; Cohen Hubal, E.A.; Avery, D.; Barceló, D.; Behra, R.; et al. Considerations of Environmentally Relevant Test Conditions for Improved Evaluation of Ecological Hazards of Engineered Nanomaterials. Environ. Sci. Technol. 2016, 50, 6124–6145. [Google Scholar] [CrossRef] [PubMed][Green Version]
  3. Blinova, I.; Vija, H.; Lukjanova, A.; Muna, M.; Syvertsen-wiig, G.; Kahru, A. Assessment of the hazard of nine (doped) lanthanides-based ceramic oxides to four aquatic species. Sci. Total Environ. 2018, 612, 1171–1176. [Google Scholar] [CrossRef] [PubMed]
  4. Garner, K.L.; Suh, S.; Lenihan, H.S.; Keller, A.A. Species sensitivity distributions for engineered nanomaterials. Environ. Sci. Technol. 2015, 49, 5753–5759. [Google Scholar] [CrossRef] [PubMed]
  5. Juganson, K.; Ivask, A.; Blinova, I.; Mortimer, M.; Kahru, A. NanoE-Tox: New and in-depth database concerning ecotoxicity of nanomaterials. Beilstein J. Nanotechnol. 2015, 6, 1788–1804. [Google Scholar] [CrossRef] [PubMed][Green Version]
  6. Scientific Committee on Emerging and Newly Identified Health Risks. Risk Assessment of Products of Nanotechnologies; European Commission: Brussels, Belgium, 2009. [Google Scholar]
  7. Keller, A.A.; McFerran, S.; Lazareva, A.; Suh, S. Global life cycle releases of engineered nanomaterials. J. Nanopart. Res. 2013, 15, 1692. [Google Scholar] [CrossRef]
  8. European Chemicals Agency. Literature Study on the Uses and Risks of Nanomaterials as Pigments in the European Union; European Chemicals Agency: Helsinki, Finland, 2018. [Google Scholar]
  9. Garner, K.L.; Suh, S.; Keller, A.A. Assessing the Risk of Engineered Nanomaterials in the Environment: Development and application of the nanoFate model. Environ. Sci. Technol. 2017, 51, 5541–5551. [Google Scholar] [CrossRef] [PubMed]
  10. Notter, D.A.; Mitrano, D.M.; Nowack, B. Are nanosized or dissolved metals more toxic in the environment? A meta-analysis. Environ. Toxicol. Chem. 2014, 33, 2733–2739. [Google Scholar] [CrossRef] [PubMed]
  11. Adam, N.; Leroux, F.; Knapen, D.; Bals, S.; Blust, R. The uptake and elimination of ZnO and CuO nanoparticles in Daphnia magna under chronic exposure scenarios. Water Res. 2015, 68, 249–261. [Google Scholar] [CrossRef] [PubMed]
  12. Adam, N.; Leroux, F.; Knapen, D.; Bals, S.; Blust, R. The uptake of ZnO and CuO nanoparticles in the water-flea Daphnia magna under acute exposure scenarios. Environ. Pollut. 2014, 194, 130–137. [Google Scholar] [CrossRef]
  13. Heinlaan, M.; Ivask, A.; Blinova, I.; Dubourguier, H.-C.; Kahru, A. Toxicity of nanosized and bulk ZnO, CuO and TiO2 to bacteria Vibrio fischeri and crustaceans Daphnia magna and Thamnocephalus platyurus. Chemosphere 2008, 71, 1308–1316. [Google Scholar] [CrossRef]
  14. Bondarenko, O.; Juganson, K.; Ivask, A.; Kasemets, K.; Mortimer, M.; Kahru, A. Toxicity of Ag, CuO and ZnO nanoparticles to selected environmentally relevant test organisms and mammalian cells in vitro: A critical review. Arch. Toxicol. 2013, 87, 1181–1200. [Google Scholar] [CrossRef] [PubMed]
  15. Thurberg, F.P.; Dawson, M.A.; Collier, R.S. Effects of Copper and Cadmium on Osmoregulation and Oxygen-Consumption in 2 Species of Estuarine Crabs. Mar. Biol. 1973, 23, 171–175. [Google Scholar] [CrossRef]
  16. Erickson, R.J.; Brooke, L.T.; Kahl, M.D.; Vende Venter, F.; Harting, S.L.; Markee, T.P.; Spehar, R.L. Effects of laboratory test conditions on the toxicity of silver to aquatic organisms. Environ. Toxicol. Chem. 1998, 17, 572–578. [Google Scholar] [CrossRef][Green Version]
  17. De Schamphelaere, K.A.; Vasconcelos, F.M.; Tack, F.M.G.; Allen, H.E.; Janssen, C.R. Effect of dissolved organic matter source on acute copper toxicity to Daphnia magna. Environ. Toxicol. Chem. 2004, 23, 1248–1255. [Google Scholar] [CrossRef]
  18. Käkinen, A.; Bondarenko, O.; Ivask, A.; Kahru, A. The effect of composition of different ecotoxicological test media on free and bioavailable copper from CuSO4 and CuO nanoparticles: Comparative evidence from a Cu-selective electrode and a Cu-biosensor. Sensors 2011, 11, 10502–10521. [Google Scholar] [CrossRef] [PubMed]
  19. Akhil, K.; Sudheer Khan, S. Effect of humic acid on the toxicity of bare and capped ZnO nanoparticles on bacteria, algal and crustacean systems. J. Photochem. Photobiol. B Biol. 2017, 167, 136–149. [Google Scholar] [CrossRef] [PubMed]
  20. Blinova, I.; Ivask, A.; Heinlaan, M.; Mortimer, M.; Kahru, A. Ecotoxicity of nanoparticles of CuO and ZnO in natural water. Environ. Pollut. 2010, 158, 41–47. [Google Scholar] [CrossRef]
  21. Cupi, D.; Hartmann, N.B.; Baun, A. The influence of natural organic matter and aging on suspension stability in guideline toxicity testing of silver, zinc oxide, and titanium dioxide nanoparticles with Daphnia magna. Environ. Toxicol. Chem. 2015, 34, 497–506. [Google Scholar] [CrossRef]
  22. Barata, C.; Baird, D.J.; Markich, S.J. Influence of genetic and environmental factors on the tolerance of Daphnia magna Straus to essential and non-essential metals. Aquat. Toxicol. 1998, 42, 115–137. [Google Scholar] [CrossRef]
  23. Chen, K.L.; Elimelech, M. Influence of humic acid on the aggregation kinetics of fullerene (C60) nanoparticles in monovalent and divalent electrolyte solutions. J. Colloid Interface Sci. 2007, 309, 126–134. [Google Scholar] [CrossRef]
  24. Cupi, D.; Hartmann, N.B.; Baun, A. Influence of pH and media composition on suspension stability of silver, zinc oxide, and titanium dioxide nanoparticles and immobilization of Daphnia magna under guideline testing conditions. Ecotoxicol. Environ. Saf. 2016, 127, 144–152. [Google Scholar] [CrossRef] [PubMed]
  25. International Organization for Standardization. Water Quality—Determination of Freshwater Sediment Toxicity to Heterocypris incongruens (Crustacea, Ostracoda); ISO 14371 (E); ISO: Geneva, Switzerland, 2012. [Google Scholar]
  26. Organisation for Economic Co-operation and Development. Test No. 211: Daphnia magna Reproduction Test, OECD Guidelines for the Testing of Chemicals, Section 2; OECD Publishing: Paris, France, 2012. [Google Scholar]
  27. Organisation for Economic Co-operation and Development. Test No. 202: Daphnia sp. Acute Immobilisation Test, OECD Guidelines for the Testing of Chemicals, Section 2; OECD Publishing: Paris, France, 2004. [Google Scholar]
  28. Zhang, Y.; Jeppesen, E.; Liu, X.; Qin, B.; Shi, K.; Zhou, Y.; Thomaz, S.M.; Deng, J. Global loss of aquatic vegetation in lakes. Earth-Sci. Rev. 2017, 173, 259–265. [Google Scholar] [CrossRef]
  29. Stevenson, L.M.; Krattenmaker, K.E.; Johnson, E.; Bowers, A.J.; Adeleye, A.S.; McCauley, E.; Nisbet, R.M. Standardized toxicity testing may underestimate ecotoxicity: Environmentally relevant food rations increase the toxicity of silver nanoparticles to Daphnia. Environ. Toxicol. Chem. 2017, 36, 3008–3018. [Google Scholar] [CrossRef]
  30. MicroBioTests Inc. OSTRACODTOXKIT F “Direct Contact” Toxicity Test for Freshwater Sediments. Standard Operational Procedure. Available online: www.microbiotests.be/SOPs/Ostracodtoxkit F SOP - A5.pdf (accessed on 24 December 2018).
  31. Toxic Cyanobacteria in Water. A Guide to Their Public Health Consequences, Monitoring, and Management; Chorus, I., Bartram, J., Eds.; St Edmundsbury Press: Suffolk, UK, 1999; ISBN 0419239308. [Google Scholar]
  32. Belgis, Z.C.; Persoone, G.; Blaise, C. Cyst-based toxicity tests XVI—Sensitivity comparison of the solid phase Heterocypris incongruens microbiotest with the Hyalella azteca and Chironomus riparius contact assays on freshwater sediments from Peninsula Harbour (Ontario, Canada). Chemosphere 2003, 52, 95–101. [Google Scholar] [CrossRef]
  33. Gondikas, A.P.; Von Der Kammer, F.; Reed, R.B.; Wagner, S.; Ranville, J.F.; Hofmann, T. Release of TiO2 nanoparticles from sunscreens into surface waters: A one-year survey at the old danube recreational lake. Environ. Sci. Technol. 2014, 48, 5415–5422. [Google Scholar] [CrossRef] [PubMed]
  34. Coll, C.; Notter, D.; Gottschalk, F.; Sun, T.; Som, C.; Nowack, B. Probabilistic environmental risk assessment of five nanomaterials (nano-TiO2, nano-Ag, nano-ZnO, CNT, and fullerenes). Nanotoxicology 2016, 10, 436–444. [Google Scholar] [CrossRef]
  35. Khangarot, B.S.; Das, S. Acute toxicity of metals and reference toxicants to a freshwater ostracod, Cypris subglobosa Sowerby, 1840 and correlation to EC50 values of other test models. J. Hazard. Mater. 2009, 172, 641–649. [Google Scholar] [CrossRef] [PubMed]
  36. Bondarenko, O.M.; Heinlaan, M.; Sihtmäe, M.; Ivask, A.; Kurvet, I.; Joonas, E.; Jemec, A.; Mannerström, M.; Heinonen, T.; Rekulapelly, R.; et al. Multilaboratory evaluation of 15 bioassays for (eco)toxicity screening and hazard ranking of engineered nanomaterials: FP7 project NANOVALID. Nanotoxicology 2016, 10, 1229–1242. [Google Scholar] [CrossRef] [PubMed]
  37. United States Environmental Protection Agency. Short-Term Methods for Estimating the Chronic Toxicity of Effluents and Receiving Waters to Freshwater Organisms, 4th ed.; United States Environmental Protection Agency: Washington, DC, USA, 2002.
  38. Gustafsson, J.P. Visual MINTEQ Version 3.1. Available online: https://vminteq.lwr.kth.se/download/ (accessed on 24 December 2018).
  39. Ivask, A.; Kurvet, I.; Kasemets, K.; Blinova, I.; Aruoja, V.; Suppi, S.; Vija, H.; Käkinen, A.; Titma, T.; Heinlaan, M.; et al. Size-dependent toxicity of silver nanoparticles to bacteria, yeast, algae, crustaceans and mammalian cells in vitro. PLoS ONE 2014, 9, e102108. [Google Scholar] [CrossRef] [PubMed]
  40. Blinova, I.; Lukjanova, A.; Muna, M.; Vija, H.; Kahru, A. Evaluation of the potential hazard of lanthanides to freshwater microcrustaceans. Sci. Total Environ. 2018, 642, 1100–1107. [Google Scholar] [CrossRef] [PubMed]
  41. Vindimian, E. MSExcel Macro REGTOX_EV7.0.5.xls. Available online: http://www.normalesup.org/~vindimian/en_index.html (accessed on 24 December 2018).
  42. Bhattacharjee, S. DLS and zeta potential—What they are and what they are not? J. Control. Release 2016, 235, 337–351. [Google Scholar] [CrossRef] [PubMed]
  43. Grillo, R.; Rosa, A.H.; Fraceto, L.F. Engineered nanoparticles and organic matter: A review of the state-of-the-art. Chemosphere 2014, 119C, 608–619. [Google Scholar] [CrossRef] [PubMed]
  44. Liu, Z.; Wang, C.; Hou, J.; Wang, P.; Miao, L.; Lv, B.; Yang, Y. Aggregation, sedimentation, and dissolution of CuO and ZnO nanoparticles in five waters. Environ. Sci. Pollut. Res. 2018, 25, 31240–31249. [Google Scholar] [CrossRef] [PubMed]
  45. Bian, S.-W.; Mudunkotuwa, I.A.; Rupasinghe, T.; Grassian, V.H. Aggregation and Dissolution of 4 nm ZnO Nanoparticles in Aqueous Enviroments: Influence of pH, Ionic Strength, Size, and Adsorption of Umic Acid. Langumir 2011, 27, 6059–6068. [Google Scholar] [CrossRef] [PubMed]
  46. Han, Y.; Kim, D.; Hwang, G.; Lee, B.; Eom, I.; Kim, P.J.; Tong, M.; Kim, H. Aggregation and dissolution of ZnO nanoparticles synthesized by different methods: Influence of ionic strength and humic acid. Colloids Surf. A Physicochem. Eng. Asp. 2014, 451, 7–15. [Google Scholar] [CrossRef]
  47. Conway, J.R.; Adeleye, A.S.; Gardea-Torresdey, J.; Keller, A.A. Aggregation, dissolution, and transformation of copper nanoparticles in natural waters. Environ. Sci. Technol. 2015, 49, 2749–2756. [Google Scholar] [CrossRef]
  48. Aruoja, V.; Pokhrel, S.; Sihtmäe, M.; Mortimer, M.; Mädler, L.; Kahru, A. Toxicity of 12 metal-based nanoparticles to algae, bacteria and protozoa. Environ. Sci. Nano 2015, 2, 630–644. [Google Scholar] [CrossRef]
  49. Ji, J.; Long, Z.; Lin, D. Toxicity of oxide nanoparticles to the green algae Chlorella sp. Chem. Eng. J. 2011, 170, 525–530. [Google Scholar] [CrossRef]
  50. Ma, S.; Zhou, K.; Yang, K.; Lin, D. Heteroagglomeration of oxide nanoparticles with algal cells: Effects of particle type, ionic strength and pH. Environ. Sci. Technol. 2015, 49, 932–939. [Google Scholar] [CrossRef] [PubMed]
  51. Quik, J.T.K.; Velzeboer, I.; Wouterse, M.; Koelmans, A.A.; van de Meent, D. Heteroaggregation and sedimentation rates for nanomaterials in natural waters. Water Res. 2014, 48, 269–279. [Google Scholar] [CrossRef]
  52. Lin, D.; Ji, J.; Long, Z.; Yang, K.; Wu, F. The influence of dissolved and surface-bound humic acid on the toxicity of TiO2 nanoparticles to Chlorella sp. Water Res. 2012, 46, 4477–4487. [Google Scholar] [CrossRef] [PubMed]
  53. Röhder, L.A.; Brandt, T.; Sigg, L.; Behra, R. Influence of agglomeration of cerium oxide nanoparticles and speciation of cerium(III) on short term effects to the green algae Chlamydomonas reinhardtii. Aquat. Toxicol. 2014, 152, 121–130. [Google Scholar] [CrossRef] [PubMed]
  54. Schlesinger, A.; Eisenstadt, D.; Bar-Gil, A.; Carmely, H.; Einbinder, S.; Gressel, J. Inexpensive non-toxic flocculation of microalgae contradicts theories; overcoming a major hurdle to bulk algal production. Biotechnol. Adv. 2012, 30, 1023–1030. [Google Scholar] [CrossRef]
  55. Muna, M.; Heinlaan, M.; Blinova, I.; Vija, H.; Kahru, A. Evaluation of the effect of test medium on total Cu body burden of nano CuO-exposed Daphnia magna: A TXRF spectroscopy study. Environ. Pollut. 2017, 1–9. [Google Scholar] [CrossRef]
  56. Stiff, M.J. The chemical states of copper in polluted fresh water and a scheme of analysis to differentiate them. Water Res. 1971, 5, 585–599. [Google Scholar] [CrossRef]
  57. Odzak, N.; Kistler, D.; Behra, R.; Sigg, L. Dissolution of metal and metal oxide nanoparticles under natural freshwater conditions. Environ. Chem. 2015, 12, 138–148. [Google Scholar] [CrossRef]
  58. Heinlaan, M.; Muna, M.; Knöbel, M.; Kistler, D.; Odzak, N.; Kühnel, D.; Müller, J.; Gupta, G.S.; Kumar, A.; Shanker, R.; et al. Natural water as the test medium for Ag and CuO nanoparticle hazard evaluation: An interlaboratory case study. Environ. Pollut. 2016, 216, 689–699. [Google Scholar] [CrossRef]
  59. Li, M.; Lin, D.; Zhu, L. Effects of water chemistry on the dissolution of ZnO nanoparticles and their toxicity to Escherichia coli. Environ. Pollut. 2013, 173, 97–102. [Google Scholar] [CrossRef]
  60. Borgmann, U.; Charlton, C.C. Copper complexation and toxicity to Daphnia in natural waters. J. Great Lakes Res. 1984, 10, 393–398. [Google Scholar] [CrossRef]
  61. Hyne, R.V.; Pablo, F.; Julli, M.; Markich, S.J. Influence of water chemistry on the acute toxicity of copper and zinc to the cladoceran Ceriodaphnia cf dubia. Environ. Toxicol. Chem. 2005, 24, 1667–1675. [Google Scholar] [CrossRef]
  62. Sevilla, J.B.; Nakajima, F.; Kasuga, I. Comparison of aquatic and dietary exposure of heavy metals Cd, Cu, and Zn to benthic ostracod Heterocypris incongruens. Environ. Toxicol. Chem. 2014, 33, 1624–1630. [Google Scholar] [CrossRef] [PubMed]
  63. Kudłak, B.; Wolska, L.; Namieśnik, J. Determination of EC50 toxicity data of selected heavy metals toward Heterocypris incongruens and their comparison to “direct-contact” and microbiotests. Environ. Monit. Assess. 2011, 174, 509–516. [Google Scholar] [CrossRef] [PubMed]
  64. Biesinger, K.E.; Christensen, G.M. Effects of Various Metals on Survival, Growth, Reproduction, and Metabolism of Daphnia magna. J. Fish. Res. Board Can. 1972, 29, 1691–1700. [Google Scholar] [CrossRef]
  65. Xu, H.; Pan, J.; Zhang, H.; Yang, L. Interactions of metal oxide nanoparticles with extracellular polymeric substances (EPS) of algal aggregates in an eutrophic ecosystem. Ecol. Eng. 2016, 94, 464–470. [Google Scholar] [CrossRef]
  66. Dalai, S.; Iswarya, V.; Bhuvaneshwari, M.; Pakrashi, S.; Chandrasekaran, N.; Mukherjee, A. Different modes of TiO2 uptake by Ceriodaphnia dubia: Relevance to toxicity and bioaccumulation. Aquat. Toxicol. 2014, 152, 139–146. [Google Scholar] [CrossRef] [PubMed]
  67. Gong, N.; Shao, K.; Li, G.; Sun, Y. Acute and chronic toxicity of nickel oxide nanoparticles to Daphnia magna: The influence of algal enrichment. NanoImpact 2016, 3–4, 104–109. [Google Scholar] [CrossRef]
  68. Wu, F.; Bortvedt, A.; Harper, B.J.; Crandon, L.E.; Harper, S.L. Uptake and toxicity of CuO nanoparticles to Daphnia magna varies between indirect dietary and direct waterborne exposures. Aquat. Toxicol. 2017, 190, 78–86. [Google Scholar] [CrossRef]
  69. Seo, J.; Kim, S.; Choi, S.; Kwon, D.; Yoon, T.-H.; Kim, W.-K.; Park, J.-W.; Jung, J. Effects of Physiochemical Properties of Test Media on Nanoparticle Toxicity to Daphnia magna Straus. Bull. Environ. Contam. Toxicol. 2014, 93, 257–262. [Google Scholar] [CrossRef]
  70. De Schamphelaere, K.A.; Janssen, C.R. Effects of dissolved organic carbon concentration and source, pH, and water hardness on chronic toxicity of copper to Daphnia magna. Environ. Toxicol. Chem. 2004, 23, 1115–1122. [Google Scholar] [CrossRef]
  71. Bianchini, A.; Wood, C.M. Does sulfide or water hardness protect against chronic silver toxicity in Daphnia magna? A critical assessment of the acute-to-chronic toxicity ratio for silver. Ecotoxicol. Environ. Saf. 2008, 71, 32–40. [Google Scholar] [CrossRef]
  72. Zhu, X.; Wang, J.; Zhang, X.; Chang, Y.; Chen, Y. The impact of ZnO nanoparticle aggregates on the embryonic development of zebrafish (Danio rerio). Nanotechnology 2009, 20, 195103. [Google Scholar] [CrossRef] [PubMed]
  73. Havel, J.E.; Talbott, B.L. Life history characteristics of the freshwater ostracod Cyprinotus incongruens and their application to toxicity testing. Ecotoxicology 1995, 4, 206–218. [Google Scholar] [CrossRef] [PubMed]
  74. Shuhaimi-Othman, M.; Yakub, N.; Ramle, N.A.; Abas, A. Toxicity of metals to a freshwater ostracod: Stenocypris major. J. Toxicol. 2011, 2011. [Google Scholar] [CrossRef] [PubMed]
Table 1. Experimental setup.
Table 1. Experimental setup.
Experiment TypeTest OrganismTest DurationAlgae R. subcapitata (Cells/mL)Test Medium
AcuteDaphnia magna48 hnoAFW, two lake waters
AcuteDaphnia magna48 h7.5 × 106AFW, two lake waters
SubchronicHeterocypris incongruens6 days7.5 × 106MHW, two lake waters
AFW—OECD 202 artificial freshwater; MHW—US EPA moderately hard reconstituted water.
Table 2. Chemical composition of the test media. Hardness values for artificial freshwaters were calculated based on Ca2+ and Mg2+ concentrations. Water was collected twice from Lake Raku and 4 times from Lake Ülemiste. The mean (SD) of the parameters of lake waters, collected at different times, is given. Conductivity in AFW and MHW was measured; other values were calculated based on the ionic composition.
Table 2. Chemical composition of the test media. Hardness values for artificial freshwaters were calculated based on Ca2+ and Mg2+ concentrations. Water was collected twice from Lake Raku and 4 times from Lake Ülemiste. The mean (SD) of the parameters of lake waters, collected at different times, is given. Conductivity in AFW and MHW was measured; other values were calculated based on the ionic composition.
D. magna AFWH. incongruens MHWLake RakuLake Ülemiste
pH7.87.68.3 (0.035)8.2 (0.45)
Conductivity 25 °C (μS/cm)640 1343 2283 (5.7)399 (60)
Total organic carbon (mg/L)005.1 (0.21)10 (0.45)
Total hardness (mg-ekv/L)51.72.7 (0.10)3.9 (0.56)
Total phosphorous (mgP/L)000.035 (0.00071)0.030 (0.012)
Total nitrogen (mgN/L)000.62 (0.13)1.4 (0.40)
Cl (mg/L)731.93.4 (0.28)11 (2.1)
SO42− (mg/L)489322 (0)29 (4.5)
Ca2+ (mg/L)801444 (2.5)66 (11)
Mg2+ (mg/L)12124.6 (0.10)7.8 (0.50)
Na+ (mg/L)18262.7 (0.021)6.7 (1.0)
Cu2+ (μg/L)001.0 (0.25)0.64 (0.19)
Zn2+ (μg/L)000.66 (0.45)0.69 (0.36)
MHW—US EPA moderately hard reconstituted water; AFW—OECD 202 artificial freshwater; 1 value from [55]; 2 measured using ZetaSizer Nano ZS (Malvern Instruments, UK).
Table 3. Metal recovery (%) from metal nanoparticles and salt upon ultracentrifugation after 24 h, 48 h or 6 days of incubation of test media at 10 mg metal/L without the test organisms. The addition of algae Raphidocelis subcapitata (7.5∙× 106 cells/mL) was used in part of the analysed samples. The mean (standard deviation) based on 1‒2 experiments is presented.
Table 3. Metal recovery (%) from metal nanoparticles and salt upon ultracentrifugation after 24 h, 48 h or 6 days of incubation of test media at 10 mg metal/L without the test organisms. The addition of algae Raphidocelis subcapitata (7.5∙× 106 cells/mL) was used in part of the analysed samples. The mean (standard deviation) based on 1‒2 experiments is presented.
MQAFWMHWLake RakuLake Ülemiste
Incubation Time24 h48 h48 h 16 Day 148 h6 Day 148 h6 Day 1
AlgaeNo No YesNo YesYesNoYesYesNoYesYes
CuO6.9 (1.6)0.67 2 (0.47)1.5 (0.8)0.421.71.80.90 2 (0.42)1.1 (0.53)1.01.2 2 (0.58)1.8 (0.29)2.0
CuSO484 (5.6)37 (6.2)47 (4.9)37402633 (0.76)45 (4.1)3132 (0.81)42 (1.3)37
ZnO27 (1.9)24 (6.3)44 (6.8)25555121 (2.7)57 (37)8323 (1.5)54 (28)76
ZnSO488 (9.1)102 (14)94 (5.3)97978190 (0.012)86 (19)9791 (2.0)90 (1.3)90
MQ—Milli-Q water; AFW—OECD 202 artificial freshwater; MHW—US EPA moderately hard reconstituted water; 1 one experiment was conducted; 2 the values include previously published data [55,58].
Table 4. Acute and subchronic toxicity of different metal formulations to Daphnia magna and Heterocypris incongruens based on nominal concentrations. Data are presented as E(L)C50 (95% confidence interval), mg metal/L based on REGTOX “optimal” model. N = 2–8 (D. magna) and n = 2–4 (H. incongruens).
Table 4. Acute and subchronic toxicity of different metal formulations to Daphnia magna and Heterocypris incongruens based on nominal concentrations. Data are presented as E(L)C50 (95% confidence interval), mg metal/L based on REGTOX “optimal” model. N = 2–8 (D. magna) and n = 2–4 (H. incongruens).
D. magna Acute EC50 (48 h)D. magna Acute EC50 (48 h) with AlgaeH. incongruens Subchronic LC50 (6 days) with Algae
AFWLake RakuLake ÜlemisteAFWLake RakuLake ÜlemisteMHWLake RakuLake Ülemiste
CuO NP1.6 * (1.1−3.4)6.3 (3.9−13)28 (18−53)2.0 (1.7−2.2)68 (57−80)>1501.1 (1.1−1.6)1.9 (1.4−3.3)2.2 (0.77−4.2)
CuSO40.053 * (0.047−0.059)0.15 (0.089−0.18)0.22 (0.20−0.25)0.41 (0.35−0.51)0.50 (0.35−0.70)0.65 (0.56−0.76)0.22 (0.20−0.24)0.25 (0.23−0.25)0.44 (0.42−0.49)
ZnO NP1.9 * (1.7−2.2)0.50 (0.46−0.58)0.71 (0.59−0.97)1.7 (1.6−2.2)1.9 (1.4−2.6)3.1 (2.1−4.5)0.36 (0.30−0.49)0.51 (0.38−0.62)0.65 (0.54−0.70)
ZnSO42.3 * (1.9−2.9)0.59 (0.53−0.79)0.76 (0.66−0.91)1.5 (1.4−1.7)0.84 (0.82−0.89)1.3 (1.3−1.4)0.36 (0.12−0.46)0.43 (0.39−0.48)0.43 (0.40−0.50)
EC50—concentration immobilising 50% of test organisms; LC50—concentration lethal to 50% of test organisms; NP—nanoparticles; AFW—OECD 202 artificial freshwater; MHW—US EPA moderately hard reconstituted water; * The calculation of these values included previously published data [36].
Table 5. The effect of addition of algae on the toxicity of copper (CuO NP and CuSO4) and zinc (ZnO NP and ZnSO4) compounds in 48 h D. magna acute immobilisation assay and on the metal recovery (reflects dissolution for NP). Background colour coding is explained below the table and shows statistically significant effects.
Table 5. The effect of addition of algae on the toxicity of copper (CuO NP and CuSO4) and zinc (ZnO NP and ZnSO4) compounds in 48 h D. magna acute immobilisation assay and on the metal recovery (reflects dissolution for NP). Background colour coding is explained below the table and shows statistically significant effects.
Copper CompoundsZinc Compounds
AFWLake RakuLake ÜlemisteAFWLake RakuLake Ülemiste
Change in toxicity 1 (EC50 with algae/EC50 no algae)NP1.311>50.93.84.4
salt7.73.33.00.651.41.7
Change in metal recovery 2 (no algae/with algae)NP0.450.820.670.550.370.43
salt0.790.730.761.11.01.0
Nanomaterials 09 00023 i001  increase   Nanomaterials 09 00023 i002  no effect   Nanomaterials 09 00023 i003 ≤5 fold decrease   Nanomaterials 09 00023 i004 ≥5 fold decrease
1 calculation based on data in Table 4. 2 calculation based on data in Table 3. AFW—OECD202 artificial freshwater; NP—nanoparticles.
Back to TopTop