Nanospecific Phytotoxicity of CuO Nanoparticles in Soils Disappeared

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Nanospecific phytotoxicity of CuO nanoparticles in soils disappeared when bioavailability factors were considered Hao Qiu, and Erik Smolders Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b01892 • Publication Date (Web): 21 Sep 2017 Downloaded from http://pubs.acs.org on September 22, 2017

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Environmental Science & Technology

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Nanospecific phytotoxicity of CuO nanoparticles in soils disappeared when bioavailability

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factors were considered

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Hao Qiu*,†,‡ and Erik Smolders†

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†

Division of Soil and Water Management, KU Leuven, 3001, Heverlee, Belgium

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‡

School of Environmental Science and Engineering, Shanghai Jiao Tong University, 200240,

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Shanghai, China

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*

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E-mail address: [email protected] (Hao Qiu)

Corresponding author. Tel.: +32 1637 6550; Fax: +32 1632 1997

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Word Limits (< 7000)

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Main text: 5012 Words

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Figures: 4 (1200 word equivalents)

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Tables: 2 (600 word equivalents)

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ACS Paragon Plus Environment


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ABSTRACT

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Bioavailability-modifying factors such as soil type and aging have only rarely been

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considered in assessing toxicity of metal-containing nanoparticles in soil. Here, we examined the

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toxicity to barley (Hordeum vulgare) of CuO nanoparticles (CuO-NPs) relative to CuO bulk

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particles (CuO-BPs) and Cu acetate (Cu(OAc)2) in six different soils with or without aging. The

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set up allows identifying whether or not NPs-derived colloidal Cu in soil porewater contributes

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to toxicity. Ultrafiltration (50 kDa) was performed together with geochemical modeling to

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determine {Cu2+} (free Cu2+ activity in soil porewater). Based on total soil Cu concentration,

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toxicity measured with seedling root elongation ranked Cu(OAc)2 > CuO-NPs > CuO-BPs in

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freshly spiked soils. The differences in toxicity among the three toxicants became smaller in soils

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aged for 90 days. When expressing toxicity as {Cu2+}, there was no indication that

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nanoparticulate or colloidal Cu enhanced toxicity. A calibrated bioavailability-based model

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based on {Cu2+} and pH successfully explained (R2 = 0.78, n = 215) toxicity of all Cu forms in

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different soils with and without aging. Our results suggest that toxicity predictions and risk

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assessment of CuO-NPs can be carried out properly using the bioavailability-based approaches

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that are used already for their non-nano counterparts in soil.

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INTRODUCTION

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Copper-containing nanoparticles (NPs) have been widely used in agriculture as fungicides

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or pesticides1, and in industry as catalysts, lubricant additives, conducting polymers and

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antifouling agents2, 3. Due to their widespread applications, there is increasing concern about the

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release of Cu-based NPs into the environment and subsequent side effects on ecosystem and

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human health. An inherent hypothesis when assessing risks posed by metal-based NPs is that the

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nanospecific properties (e.g., high surface area to volume ratio and high reactivity) will cause

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specific interactions with the test media and test organisms compared to the case of

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corresponding bulk particles or salt forms4, 5. As a result, new biological effects may be expected.

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However, a consensus on the existence of nanospecific effects has yet to be reached. For instance,

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the inhibitory effects of different copper forms on the growth of duckweeds (Landoltia punctata)

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followed the order: CuCl2 > CuO NPs > bulk CuO6, whereas Zhao et al. found that CuO NPs

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exhibited a much higher growth inhibition on water hyacinth (Eichhornia crassipes) than that of

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bulk CuO and CuSO47. The toxicities of CuO NPs have been attributed to either the NPs

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themselves8, 9, released metal ions10, or both11, 12. It becomes clear that an accurate assessment of

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NPs dissolution in the exposure media is essential to draw robust conclusions in nanotoxicity

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analysis.

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The fate and toxicity of nanoparticles in aquatic system have been studied extensively13,

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while the current knowledge on the interactions of NPs with soil and the subsequent impact on

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toxicity is clearly lagging behind. In simple solution system, it is relatively straightforward to

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distinguish the effect of NPs from the effect of released ions and even to calculate the

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contributions of NPs and ions released to the overall toxicity14-16. In soils, quantification of the

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amount of ions released remains problematic17,

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. This is because organic matter, reactive



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minerals (e.g., iron and aluminum (hydr)oxides), and clay minerals in soil are likely to play a

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vital role in determining the fate and bioavailability of NPs via sorption interactions or by

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inducing particle aggregation19. The conclusions on nanospecific toxicity were often drawn by

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comparing the toxicity of NPs with the reference toxicant forms on a basis of total soil metal

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concentration or, at best, concentrations in membrane filtered porewater that includes the

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colloidal fraction and, hence the NPs20-22. In such cases, bioavailability factors are difficult to

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analyze and apparent nanospecific toxicity (i.e., NPs are more toxic than other forms on a total

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soil concentration basis) may be just due to confounding factors and do not reflect the real effect

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at the toxicological level. According to the biotic ligand model (BLM) theory for ‘conventional’

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forms of metals, the free metal ion in the soil porewater is the main toxic species responsible for

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inducing toxicity and its toxicity can be mitigated by competing ions such as Ca2+ and protons23,

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0.45 µm filtration step) do not discriminate the mineral colloids from truly dissolved free metal

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ion or small organic metal complexes3. If the NPs in the soil porewater contribute to toxicity in

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addition to the released metal ions, then soil treatments with NPs would cause more toxic effects

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to organisms than treatments with traditional forms of metals when the free ion activities in soil

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porewater of these treatments are the same25. In contrast, if the NPs or, more general, the

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minerals colloids do not contribute to toxicity, then toxicity is merely related to the metal ions

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and existing bioavailability-based models such as BLM for metal salts could be calibrated to

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normalize toxicity data of different forms of metals26. These possibilities remain to be further

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evidenced by experimental data.

. In soils dosed with NPs, the total metal concentration in the soil porewater (normally after a

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Aging is another important aspect to consider in the environmental implications of NPs. In

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toxicity studies of NPs in soil, a 24 h equilibration time (rather than long-term soil incubation) is

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often used to provide a worst-case scenario in terms of less dissolution and greater NPs

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bioavailability27,

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representative of soils subjected to long-term aging, thus may not reveal the full toxicity

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potential of NPs29. Besides, the equilibration processes or rates for different metal forms are

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likely to be different17. Therefore, for comparative toxicity studies (nano versus non-nano forms),

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it is necessary to determine whether different equilibration time (aging time) may provide

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different answers and input data for risk assessment of NPs. The effects of aging on NP

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dissolution may also differ across different soils. To the best of our knowledge, the effects of soil

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type and aging on the toxicity of metal-based NPs have hardly been investigated.

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. However, toxicity of NPs measured in freshly spiked soils may not be

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The aims of this study were 1) to investigate the relative toxicity of Cu derived from CuO

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NPs to barley (Hordeum vulgare L.) in different soils upon different aging time after spiking

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with bulk-sized CuO and Cu acetate as two reference toxicant forms, 2) to identify potentially

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larger toxicity of the NPs than that of Cu ions, either expressed on total soil or soil porewater

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basis, and 3) to test if the BLM concept derived for metal salt spiked soils is applicable to CuO

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NPs across different soils and aging time. All CuO NPs fate and toxicity studies were referenced

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to bulk CuO and Cu acetate because the risk assessment of NPs needs information about

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potential additive toxicity of the metal-containing NPs in soil beyond the obvious effects of the

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Cu2+ in soil solution or of the effects of the total salt form added to soil. In our study, both a

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short-term (24 h) and a long-term (90 d) equilibrium time were used for identifying the aging

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effects. It is postulated that CuO-NPs may show, if any, nanospecific toxicity after short-term

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soil incubation and that this nanospecific effects will disappear upon long-term incubation as

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CuO-NPs dissolve. The nanospecific effects on toxicity can go in both directions, e.g., lower

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toxicity of NP compared to the salt when using total soil Cu concentration as dose or larger

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toxicity when using soil porewater Cu (including colloidal Cu) as dose. In the latter case, the NPs

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serve as reservoir for time-dependent release of ions.

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MATERIALS AND METHODS

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Soils and treatments for toxicity tests

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A total of six soils from five locations in Europe were used for toxicity tests. The top 20 cm

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soil samples were collected from Houthalen (Belgium), Kovlinge (KOV, Sweden), Rhydtalog

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(RHY, Wales), Ter Munck (TMK, Belgium), and Zegveld (ZEG, The Netherlands), respectively

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(Supporting Information, Table S1). These soil samples were air-dried, sieved through a 4-mm

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mesh, homogenized, and stored in barrels at ambient temperature until use. One portion of

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Houthalen soil was amended with CaO to have two pH levels: the original low pH soil (HTL, pH

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= 4.8) and the adjusted high pH soil (HTH, pH = 5.8). Each soil was spiked with 5 to 7

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concentration levels of CuO nanoparticles (CuO-NPs) (Sigma-Aldrich,

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CuO-BPs treatment except for HTL and RHY soil. Dissolution of metal oxides depend on their

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surface area, which is larger for smaller particles37. The higher surface of CuO-NPs compared to

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CuO-BPs increases the rate of dissolution25. Upon aging, total dissolved Cu concentrations

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increased in the soils (except for TMK soil) spiked with CuO-NPs and CuO-BPs but decreased

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in the soils spiked with Cu(OAc)2. This suggests a slow dissolution of Cu oxides during soil

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incubation, counteracted by fixation of dissolved Cu2+ in soil. The increasing porewater metal

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concentrations in soil over time was also reported for ZnO-NPs38 and AgNPs29. Aging reactions

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in soil have shown to decrease the mobility and solubility for many metal salts39, 40.

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The truly dissolved Cu after the ultrafiltration step ranked, at equivalent Cu addition to soil,

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Cu(OAc)2 > CuO-BPs > CuO-NPs. An exception was TMK soil, where adding CuO-NPs

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resulted in higher truly dissolved Cu concentration in the porewater than the case of adding the

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same amount of CuO-BPs. For Cu(OAc)2 treatments, truly dissolved Cu decreased with aging in

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all soils as did the total dissolved Cu, indicating the fixation reactions (see above). For CuO-NPs

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and CuO-BPs treatments, aging had inconsistent effects on truly dissolved Cu concentrations.

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These variable trends contrast more consistent increasing trends in total dissolved Cu and may be

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related to the balance between slow dissolution counteracted by flocculation of colloidal NP and

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by the aging of the Cu2+ ions.

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The colloidal Cu in the 0.45 µm filtered pore water, i.e., Cu CuO-NPs > CuO-BPs when

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using total soil Cu concentration [Cu]tot as the dose descriptor. This corresponds to the order of

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solubility of these Cu forms in soil. After soil incubation for 90 days, EC50[Cu]tot values for the

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three Cu forms were similar in HTL and RHY soils, suggesting that aging reverted these

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different Cu forms to be equally toxic (Figure 2). For the other 4 soils, the differences in values

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of EC50[Cu]tot for the three Cu forms became smaller than those of soils without aging (Table 2).

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It is expected that the toxicity of metal NPs, BPs and salts converge to a common value after

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sufficient aging time that is determined by the equilibrium partitioning of metal species between

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the soil solution and soil solid phases29, 42. In RHY and TMK soil, there were no significant

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effect of aging on the toxicity of CuO-NPs and CuO-BPs. In the rest of the soils investigated,

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values of EC50[Cu]tot for CuO-NPs and CuO-BPs decreased upon aging, indicating an increased

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toxicity. This may result from an increase in dissolution with increasing time. For Cu(OAc)2,

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there were slight increases in EC50[Cu]tot values in all soils after aging (Table 2). This is

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consistent with the findings that toxicity of metal salts decreased with prolonged soil incubation

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time39, 40. The EC50[Cu]tot in soils freshly spiked with Cu(OAc)2 varied by a factor of 5 and

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increased with soil CEC as shown earlier35. The EC50[Cu]tot of soils freshly spiked with NPs

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differed by a factor of 38. In reality, changes in NP toxicity expressed in terms of the total metal

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added to the soil may actually reflect a balance between surface modifications, dissolution

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processes and the speciation and solid phase binding of dissolved ions. When porewater Cu

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concentration [Cu]pw was used as the dose, both truly dissolved Cu and particulate Cu that can

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pass 0.45µm membrane were considered. The EC50[Cu]pw for the three Cu forms increased in

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the following order: Cu(OAc)2 < CuO-BPs < CuO-NPs in all freshly spiked soils (Table 2). This

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cannot be explained simply by the porewater Cu concentrations (order: Cu(OAc)2 > CuO-NPs >

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CuO-BPs). The Cu species in the porewater thus needed to be explicitly considered. Upon aging,

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there were almost no changes in values of EC50[Cu]pw for the three Cu forms in RHY, TMK,

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and ZFG soil, while an evident decrease in EC50[Cu]pw values for CuO-NPs and CuO-BPs were

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observed in HTL, HTH, and KOV soil. When expressing toxicity as free ion activity in soil

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solution which were estimated from the composition of the truly dissolved porewater fraction,

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the EC50{Cu2+} values were almost unaffected by source of Cu and ageing. There was only one

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case in which the Cu2+ toxicity was enhanced when NPs were the source relative to other

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treatments (HTL soil freshly spiked with CuO-NPs). In contrast, the five other freshly spiked

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soils and in all six aged soils, the toxicity (as seen from EC50{Cu2+} values) of the NPs was not

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different from the corresponding salt amended soil (4 cases) or was even lower (5 cases). Hence,

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the evidence for toxicity of colloidal Cu in a CuO-NP amended soil is exceptional and, at most,

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transient (Figure 3 and Table 2).

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Normalization of toxicity of different Cu forms across soils

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As a next step, all soil treatments, including 215 data points of different forms of Cu and

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different soil incubation time, were compiled together to better incorporate bioavailability among

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soils, metal forms and aging. A logistic dose-response equation was applied to fit all data.

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Generally, model fits improved with improved considerations of bioavailability. When we used

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total Cu concentrations as dose, the obtained R2 = 0.43 and RMSE = 26.4 (Figure 4A), and these

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values were 0.51 and 24.7 with the use of porewater Cu concentrations as dose. When we used

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free ion activities as dose, R2 value further went up to 0.60 and RMSE decreased to 22.0. Even

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though, there was still 40% of variance in toxicity that cannot be explained, suggesting other

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bioavailability-modifying factor (porewater chemistry) may also need to be considered. A free

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ion effective dose model, rather than a traditional BLM was used to predicting Cu toxicity43:

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n

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γ effect = log{Cu 2 + }effect − α ⋅ pH − ∑η n ⋅ p{C z + }

(3)

1

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where γeffect is the effective dose that incorporates not only the {Cu2+}, but also the factors (e.g.,

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H+ and other competing cations CZ+) modifying the effects of bioavailability. The coefficients α

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and ηn, describing the effects of cations on Cu toxicity, are assumed to be independent of the

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effect level23. Toxicity of Cu2+ was then related to γeffect via the log-logistic dose-response model

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(Eq. 1). Model coefficients were calibrated by multiple nonlinear regression analysis. The

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activity of H+ (but not other major cations K+, Ca2+, Na+, and Mg2+) in the porewater was

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identified as the sole variable that affects Cu2+ toxicity and was included into the final model.

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The model parameter α for pH was negative (-0.92), indicating increased toxicity with increasing

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pH (i.e., decreasing H+), which is compatible with the competitive binding concept of the BLM44

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and with the findings by Lofts et al.43. With the use of effective dose γeffect, the model provided

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superior fit (R2 = 0.78, RMSE = 17.9) (Figure 4D) in comparison to the fits obtained with other

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dose descriptors (Figure 4A, B, and C). This supports the assumption that free ions released

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from NPs represents the dominant bioavailable form of Cu in soil, confirming the feasibility of

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applying the existing bioavailability-based approaches to assess soil NPs toxicity.

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In soil toxicity tests that tested soluble metal salts as the conventional worst-case source,

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the pore water has been well recognized as a dominant pathway for exposure and free metal ions

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is supposed to be responsible for plant uptake and subsequent toxicity24, 45. In freshwater, the role

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of Cu2+ release from CuO-NPs as the sole cause for CuO-NPs toxicity is under debate. For

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example, the aquatic toxicity of CuO-NPs was 48 times higher than that of CuO-BPs to Vibrio

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fischeri when CuO concentrations were expressed as suspension Cu, while their toxicity was

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almost identical when considering bioavailable Cu (with the use of Cu ion sensor bacteria)46.

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This and the present study acknowledged the predominant role of Cu2+ in determining toxicity of

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different Cu forms. However, other aquatic studies suggested that suspended particles maybe of

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greater importance as these NPs can be internalized and trigger the generation of reactive oxygen

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species (ROS) or dissolve intracellularly12, 47. After exposure to CuO-NPs and CuO-BPs in a

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sand-solution matrix, a fraction of the Cu in the wheat shoots was found to be present as CuO,

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suggesting intact uptake and transport of these materials11. Wang et al.8 also reported xylem- and

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phloem-based transport of CuO NPs in maize in hydroponic culture as observed by TEM-EDS.

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In our experimental condition, i.e., in real soil environment, the resulting CuO-NPs in the

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porewater may be too large to cross the cell wall. At equivalent Cu concentrations added into soil,

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root Cu contents in CuO-NPs treatments were only slightly higher than that in CuO-BPs

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treatments, and were close to (or lower than) that in Cu(OAc)2 treatments (Table S2). In soils, it

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has been demonstrated that increasing Ca concentrations from 0.1 to 10 mM increased the zeta-

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potential of topsoil colloids from -25 to -5 mV, meaning that colloids tend to coagulate or

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flocculate at high Ca levels48. In the HTL soil (low Ca) in which the exceptionally larger toxicity

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of NP was found (free ion activity based, Figure 3), it is expected that colloids are smallest

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potentially explaining the additional toxicity to plants other than the released ions. The majority

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of CuO-NPs (original size < 50 nm) was reported to form large aggregates (> 450 nm) in ISO

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test media49. It has also been reported that bacteria are largely protected against NP entry (no

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transport mechanisms for supramolecular and colloidal particles)46. For example, only < 5 nm

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quantum dots entered the bacterial cells, probably by light-aided oxidative damage of the cell

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membrane50. In our study, the presence of complex soil solution constituents may have

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accelerated the processes of aggregation/agglomeration of NPs. As a result, only the released

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ions were available for uptake and subsequently inducing the toxicity.

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Previously, bioavailability has been successfully used to account for the observed variations

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in interaction patterns and toxicity of a binary mixture (Cu-Zn) across different soils32. In the

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present study, the observed differences in toxicity between forms of Cu were mainly explained

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by differences in solubility. By disentangling the effects of dissolution of NPs and incorporating

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the bioavailability into toxicity assessment, the source and aging specific toxicity reduced. Upon

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even only three months of aging, it appeared that soil properties were by far more important

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factors determining the toxicity (total soil based) than the source of Cu. Our results embrace the

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argument that the nanospecific properties may be less important in inducing toxicity than the

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actual exposure dose and environmental conditions51. This also suggests that risk assessment of

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metal-containing NPs can be conducted based on existing concepts of the metal risk assessment

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in soil.

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Taken together, we found that ultrafiltration was an effective way of separating truly

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dissolved Cu from single nanoparticles and their aggregates in the soil porewater. The toxicity of

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CuO-NPs differed from CuO-BPs and Cu(OAc)2 when soil total Cu concentration was used as

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the dose. Upon aging, toxicity increased for CuO-NPs and CuO-BPs but decreased for Cu(OAc)2.

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The phytotoxicity of CuO nanoparticles was lower than that of Cu2+ salt and was equal (or

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sometimes somewhat larger) than bulk CuO. These differences were reduced when

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bioavailability factors were considered. This suggests the principal mechanism of NPs toxicity in

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soil was dissolution of metal oxides into a metal ion form, not a direct effect of particles acting

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on plant. We cannot exclude that NPs may play an unique role in inducing toxicity via intact

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uptake (including xylem- and phloem-based transport, etc.)8, 11, but speculate that these processes

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and related nanospecific effects are likely to be more evident in aquatic system. In our study, soil

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factors showed to be more potent than Cu sources for toxicity. A larger toxicity of soil solution

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Cu2+ ion activity in the presences of CuO NPs as a source compared to its absence was found in

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only 1 of 12 cases. Hence, the evidence for toxicity of colloidal Cu in a CuO-NP amended soil is

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exceptional and, at most, transient. The developed bioavailability-based model, i.e., the free ion

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effective dose model, can be used to predict and reconcile toxicity of different forms of Cu in the

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investigated soils. Our findings suggest that research efforts towards incorporating

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bioavailability into toxicity predictions are more valuable than putting the emphasis on

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reassessing toxicity of metal-containing NPs separately from toxicity of other forms present in

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soil.

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ASSOCIATED CONTENT

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Supporting Information

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The Supporting Information is available free of charge on the ACS Publications website. Tables

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showing the physicochemical properties of the selected soils, total concentrations of Cu in

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different soil treatments, root Cu uptake, and Cu recovery in different solutions after

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ultrafiltration.

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AUTHOR INFORMATION

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Corresponding Author

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*

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Notes

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The authors declare no competing financial interest.

E-mail: [email protected] (Hao Qiu)

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ACKNOWLEDGEMENTS

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This study was supported by the National Natural Science Foundation of China (project No.

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41701571). H.Q. received a mobility grant from the Belgian Federal Science Policy Office

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(project No. 3E150127).

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Figures and Tables

100

50

HTL SOIL 0

1000

1500

200

100

0

2000

HTH SOIL 0

Total Cu (mg/kg)

Truly dissolved Cu (µM)

(B) 100

CuO-NPs_fresh CuO-BPs_fresh Cu(OAc)2_fresh

80

(D) 20

CuO-NPs_aged CuO-BPs_aged Cu(OAc)2_aged

60 40 20

HTL SOIL

0 0

500

1000

1500



200

100

RHY SOIL 0

2000

4000

6000



8000


40 20

RHY SOIL 4000

6000

8000

Total Cu (mg/kg)

4000

6000

4000

10000


6000

8000

KOV SOIL

0 0

TMK SOIL 6000

3000

4000

5000

CuO-NPs_fresh

CuO-NPs_aged

Cu(OAc)2_fresh

Cu(OAc)2_aged

30 20 10

ZEG SOIL

0 0

50

5

4000

2000

40

(L) 60

10

2000

1000

2000

4000

6000

8000

10000

Total Cu (mg/kg)

15

0


20

10000


0

8000

40

50

TMK SOIL 2000

6000

60

(K) 60

5

0

4000

Total Cu (mg/kg)


0

2000


80

8000

10

(J) 20

60

2000

0

Total Cu (mg/kg)

80

0

2000

15

10000


0

KOV SOIL

0

(F) 100


HTH SOIL

Total Cu (mg/kg)

(H) 100

20

Total Cu (mg/kg)

0

(I) 20

300

565

40

8000

5

0


60

Total Cu (mg/kg)


0

6000

10

2000

Total dissolved Cu (µM)


(G) 400

4000

15

Total Cu (mg/kg)

564

2000


80

Total Cu (mg/kg)


563

566

500

300


0

(E) 100



150



200

(C) 400


8000

10000

Total Cu (mg/kg)





(A) 250


562


CuO-NPs_fresh

CuO-NPs_aged

Cu(OAc)2_fresh

Cu(OAc)2_aged

40 30 20 10

ZEG SOIL

0 0

2000

4000

6000

8000

10000

Total Cu (mg/kg)

567

Figure 1. Total dissolved (

Nanospecific Phytotoxicity of CuO Nanoparticles in Soils Disappeared

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