Sorption, dissolution and pH determine the long-term equilibration and toxicity of coated and uncoated ZnO nanoparticles in soil

Environmental Pollution 178 (2013) 59e64

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Sorption, dissolution and pH determine the long-term equilibration and toxicity of coated and uncoated ZnO nanoparticles in soil Pauline L. Waalewijn-Kool a, *, Maria Diez Ortiz b, Nico M. van Straalen a, Cornelis A.M. van Gestel a a b

Department of Ecological Science, Faculty of Earth and Life Sciences, VU University, De Boelelaan 1085, 1081 HV Amsterdam, The Netherlands Centre for Ecology and Hydrology, Maclean Building, Benson Lane, Crowmarsh Gifford, Wallingford, Oxfordshire OX10 8BB, UK

a r t i c l e i n f o

a b s t r a c t

Article history: Received 26 October 2012 Received in revised form 26 February 2013 Accepted 1 March 2013

To assess the effect of long-term dissolution on bioavailability and toxicity, triethoxyoctylsilane coated and uncoated zinc oxide nanoparticles (ZnO-NP), non-nano ZnO and ZnCl2 were equilibrated in natural soil for up to twelve months. Zn concentrations in pore water increased with time for all ZnO forms but peaked at intermediate concentrations of ZnO-NP and non-nano ZnO, while for coated ZnO-NP such a clear peak only was seen after 12 months. Dose-related increases in soil pH may explain decreased soluble Zn levels due to fixation of Zn released from ZnO at higher soil concentrations. At T ¼ 0 uncoated ZnO-NP and non-nano ZnO were equally toxic to the springtail Folsomia candida, but not as toxic as coated ZnO-NP, and ZnCl2 being most toxic. After three months equilibration toxicity to F. candida was already reduced for all Zn forms, except for coated ZnO-NP which showed reduced toxicity only after 12 months equilibration. Ó 2013 Elsevier Ltd. All rights reserved.

Keywords: Zinc oxide nanoparticles Soil Springtail Long-term exposure Dissolution

1. Introduction Nanotechnology has developed an increasing number of nanobased products that are currently applied in textiles, electronics, pharmaceuticals and cosmetics. This has raised scientific and public concerns about the potential impact of nanomaterials on the environment. Zinc oxide nanoparticles (ZnO-NP) are among the most commonly used nanoparticles having different uses such as environmental remediation and sunscreen application (Wang, 2004; Osmond and McCall, 2010). Due to their increased use and disposal ZnO-NP are likely to enter the environment, with soil being a potential sink, posing a hazard to soil organisms (Tourinho et al., 2012). Adverse effects on the reproduction of earthworms were shown for ZnO-NP (Canas et al., 2011; Hooper et al., 2011). Standardized tests for regular chemicals are useful for determining the toxicity of nanoparticles (Kahru and Dubourguier, 2010). Short-term toxicity tests, however, lack the ability to study environmental fate processes, such as dissolution and sorption. Zn ion release from ZnO-NP was shown to be relatively fast in water (Poynton et al., 2011) and kaolin suspensions (Scheckel et al., 2010). Dissolution of metal-oxide nanoparticles depends on surface area, which is larger for smaller particles (Borm et al., 2006). Different * Corresponding author. E-mail address: [email protected] (P.L. Waalewijn-Kool). 0269-7491/$ e see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.envpol.2013.03.003

aquatic studies show that nanoparticles dissolve faster than larger sized materials of the same mass (Wong et al., 2010; Reed et al., 2012). The dissolution of nanoparticles in soil may be different compared to liquid media. Nanoparticles tend to aggregate and may form coatings over mineral surfaces (Theng and Yuan, 2008). Soil organic matter has a high binding capacity for metal oxides and influences the dissolution of nanoparticles. Soil pH may play an important role in the dissolution of the amphoteric ZnO (Bian et al., 2011). The aqueous solubility of ZnO ranges from several thousand mg per litre at pH 6 to around 1 mg/l at pH 8 (Apte et al., 2009). Also, a coating of nanoparticles is likely to influence dissolution in soils, by preventing the release of metal ions. The majority of the nanoparticles are produced with a coating, but data is lacking on the difference in toxicity between coated and uncoated ZnO-NP. Long-term exposures in the environment have shown to decrease zinc toxicity in soil over time (Lock and Janssen, 2002; Smit et al., 1997). Currently, it is not known whether long-term exposure of ZnO-NP also reduces their bioavailability and potential toxicity in soil. Collembola are an integral part of soil ecosystems and are vulnerable to the effects of soil contamination. Folsomia candida has been used as a model organism for more than 40 years (Fountain and Hopkin, 2005). The 28-day EC50 for the toxicity of ZnCl2 in Lufa 2.2 soil has been reported to be between 348 and 476 mg Zn/kg d.w. (Smit et al., 1997; Nota et al., 2010).

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This study aims to determine the toxicity of coated and uncoated ZnO-NP, non-nano ZnO and ZnCl2 to F. candida in equilibrated soil. Considering an expected slow Zn release, natural soil was equilibrated for one year in the laboratory. Pore water was collected from soils freshly spiked and after three, six and twelve month equilibration. At each sampling time the toxicity of the four Zn forms to F. candida was compared to explain ZnO-NP toxicity by total Zn in the soil or by soluble Zn concentrations in the pore water. We hypothesized that 1) ZnO-NP toxicity is attributable to soluble Zn concentrations rather than to particle size, 2) the coating of ZnONP reduces its dissolution and thereby its toxicity to springtails and 3) that equilibration leads to increased toxicity of ZnO due to an increased dissolution. 2. Materials and methods 2.1. Test compounds Coated ZnO-NP (Z-COTEÒHP1) and uncoated ZnO-NP (Z-COTEÒ) powder, both with a reported diameter of <200 nm, were obtained from BASF SE (Ludwigshafen, Germany). The mass fraction (w/w) of the coated ZnO-NP was 96e99% zinc oxide and 1e4% triethoxyoctylsilane (coating). Triethoxyoctylsilane (CAS 2943-75-1, colourless liquid) was purchased from SigmaeAldrich Chemie BV (97.5%). Non-nano ZnO (Merck, pro analysi, >99%) and ZnCl2 (Merck, zinc chloride pure) were used for comparison of particle size and with free zinc, respectively. ZnO-NP were characterized using Transmission Electron Microscopy (TEM) and Dynamic Light Scattering (DLS); see Supporting Information Fig. SI-1e4. 2.2. Soil Loamy sand soil (LUFA-Speyer 2.2, Sp 2121, Germany, 2009) with pHCaCl2 5.5, total organic carbon content 2.09%, cation exchange capacity 10.0 meq/100 g and water-holding capacity (WHC) 46.5% was used. The soil was oven-dried at 60  C overnight prior to the experiments to eliminate undesired soil fauna. 2.3. Spiking the soil The first experiment consisted of seven concentrations of uncoated ZnO-NP (nominal range 100e6400 mg Zn/kg d.w.), five non-nano ZnO concentrations (400e6400 mg Zn/kg d.w.), six ZnCl2 concentrations (100e3200 mg Zn/kg d.w.) and two controls without added zinc. The test compounds were introduced into the soil as aqueous solutions prepared in soil extracts (Van der Ploeg et al., 2011). Soilewater suspensions were prepared by mixing air-dried soil with Milli-Q water using a soile water ratio of 2:5 (w/v), shaken at 180 rpm at ambient temperature for one hour, and filtered under vacuum (Whatman filter paper, type 595). Uncoated ZnO-NP, nonnano ZnO and ZnCl2 were added to the filtrates (see Kool et al. (2011) for TEM photos of ZnO particles in soil solutions), shaken for two days at 180 rpm and carefully mixed with dry soil. Soil moisture content was adjusted to 23.3% (w/w) with Milli-Q water. In the second experiment seven concentrations of coated ZnO-NP (nominal range 100e6400 mg Zn/kg d.w.), triethoxyoctylsilane and two controls without added zinc were tested. The hydrophobic triethoxyoctylsilane coating makes the coated ZnO-NP insoluble in water, therefore they were mixed with the soil as dry powders. Soil moisture content was adjusted to 23.3% (w/w) with Milli-Q water. For each Zn treatment, a total amount of 600 g soil was prepared. To investigate the toxicity of the coating, five concentrations of triethoxyoctylsilane were tested in a 28-day toxicity test with F. candida. For this purpose, 20 g dry soil was spiked with triethoxyoctylsilane dissolved in acetone. After evaporation of the acetone, 180 g dry soil was added and soils were mixed with a spoon to reach nominal concentrations (75e1200 mg/kg). Milli-Q water was added to achieve a soil moisture content of 23.3% (w/w). An acetone control and a water control were included in the toxicity test. 2.4. Conditions for equilibration of the soil Soils were equilibrated in glass jars with the individual Zn forms for one year, in a climate room at 20  1  C. Twice a month, soil moisture content was checked by weighing the jars, and moisture loss was replenished with Milli-Q water. Soil moisture content never decreased more than 2.5% upon replenishing the water loss. Once every 6e8 weeks and upon sampling, soils were mixed with a spoon. No growth of fungi was visible over the course of the experiment. The lids of the jars were not tightly closed so the jars were aerated continuously. After three, six and twelve months the soil was sampled to collect pore water and to perform a toxicity test with F. candida.

2.5. Toxicity test Toxicity experiments were carried out following ISO-guideline 11267 (ISO, 1999) using survival and reproduction of F. candida as effect parameters. F. candida (“Berlin strain”; VU University Amsterdam) was cultured in plastic containers with a moist bottom of plaster of Paris containing 10% charcoal, at 20  1  C and 12/12 h light/dark. Each experiment was initiated with age-synchronized 10e 12 day old juveniles. At the start of each test (8 in total, four time points in the two experiments), ten synchronized animals were transferred into 100 ml glass test containers containing 30 g soil each. Five replicates for each test concentration and control were prepared. The test jars were filled randomly and before introduction the animals were checked under the microscope for a healthy appearance. At the beginning of the experiments the animals were fed a few grains of dried baker’s yeast. The jars were incubated at 20  1  C and 12/12 h light/dark. Once a week, the moisture content of the test soils was checked by weighing the jars, and moisture loss was replenished with Milli-Q water when necessary. The jars were also aerated by this procedure. After four weeks, the jars were emptied into a 200 ml beaker glass and 100 ml Milli-Q water was added. The mixture was stirred carefully to let all the animals float to the surface. The number of adults and juveniles were counted manually after taking a picture of the water surface using a digital camera (Olympus, C-5060).

2.6. Zinc analyses After spiking and after one year, soil samples were oven-dried at 60  C for total zinc analysis. Approximately 100 mg dried soil (two replicates per test concentration and time point) was digested in a mixture of Milli-Q water, concentrated HCl and concentrated HNO3 (1:1:4 by vol.), in tightly closed teflon bombs in an oven (CEM MDS 81-D) at 140  C for 7 h. The solution was analysed by flame Atomic Absorption Spectrometry (AAS) (Perkin Elmer 1100B). Certified reference material (ISE sample 989, River Clay from Wageningen, The Netherlands) was used to ensure the accuracy of the analytical procedure. Measured zinc concentrations in the reference material were within 10% of the certified concentrations. At each sampling time, pore water from each test soil was collected by centrifuging 50 g soil (Centrifuge Falcon 6/300 series, CFC Free), after saturation with Milli-Q water and one week equilibration time (Tipping et al., 2003). Soils were centrifuged for 50 min at 2000 g over two round filters (S&S 597Ø 47 mm, pore size 11 mm) and a 0.45 mm membrane filter (S&S Ø 47 mm), placed inside the tubes (Hobbelen et al., 2004). Pore water was analysed by flame AAS (Perkin Elmer 1100B). In case of uncoated ZnO-NP, zinc concentrations in the pore water were also determined by flame AAS after ultrafiltration to obtain a particle-free extract. Soil solutions were centrifuged in a 100 kDa ultrafiltration device (Amicon Ultra-15 Filters, Millipore) for 20 min at 2000 g. This did not show differences in Zn porewater concentrations before and after ultrafiltration for the uncoated ZnO-NP. Because of that and because we expected the coated ZnO-NP to be less prone to aggregation, we tried the use of 3 kDa ultrafiltration devices for coated ZnO-NP. The pH in the pore water was only measured at the beginning of the first experiment. After six and twelve month equilibration, pHwater of the soil was measured. In freshly spiked soil with coated ZnO-NP soil pHwater was also measured. Soils were shaken with deionised water (5:1 liquid:soil)for 2 h at 200 rpm. The pHCaCl2 of soils spiked with triethoxyoctylsilane and of soils equilibrated with coated ZnO-NP for twelve months were measured by shaking them with 0.01M CaCl2 (5:1 liquid:soil) for 2 h at 200 rpm. pH was recorded using a Consort P907 meter.

2.7. Data analysis LC/EC50, the concentrations in soil and pore water causing 50% reduction in springtail survival and reproduction, respectively, were estimated applying the logistic model of Haanstra et al. (1985). A generalized likelihood ratio test (Sokal and Rohlf, 1995) was performed to compare EC50 values obtained at different equilibration times. All calculations were performed in SPSS 17. Zn speciation was modelled for the lowest and highest Zn concentrations measured in the pore water using Visual MINTEQ (http://www.lwr.kth.se/English/ OurSoftware/vminteq; Gustafsson, 2007). Speciation was also calculated for the concentrations closest to the EC50 values.

3. Results This study is a follow-up of a previous experiment with ZnO-NP, non-nano ZnO and ZnCl2. The results of the 28-days toxicity tests with F. candida in freshly spiked Lufa 2.2 soil are published in Kool et al. (2011). These data represent time point zero (T ¼ 0) in this study.

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and 36.6 mg Zn/l, respectively. These porewater Zn concentrations were lower than the ones measured for uncoated ZnO-NP at all time points and for coated ZnO-NP a substantial peak at intermediate concentrations was not observed until twelve months equilibration.

3.1. Total soil analysis and pH measurements Soil zinc concentrations measured at the start and end of the experiment are presented in Table SI-1. Recovery was 91% for all soil samples at both sampling times. All results are expressed on the basis of measured concentrations. The pH slightly increased with increasing soil concentrations for coated and uncoated ZnO-NP, and non-nano ZnO, and dose-related decreased for ZnCl2 at all time points (Table SI-2). For all Zn forms, except for the lower ZnCl2 concentrations, pH showed a steady decrease with time. The pHCaCl2 for soil spiked with triethoxyoctylsilane was approx. 4.83 for all test concentrations.

3.3. Ecotoxicity to F. candida EC50 values estimated for the four Zn forms are presented based on measured Zn concentrations in soil (Table 1) and on porewater concentrations (Table 2). Corresponding doseeresponse curves for coated and uncoated ZnO-NP, non-nano ZnO and ZnCl2 are shown in Fig. SI-5. In the first experiment (including uncoated ZnO-NP, non-nano ZnO and ZnCl2), control survival of the Collembola in the four toxicity tests with freshly spiked, three, six and twelve month equilibrated soil was 72, 88, 87 and 62%, respectively. The number of juveniles in the control with three month equilibrated soil was only 58 (18.3, n ¼ 10), but since survival was normal and consistent doseeresponse curves were seen the test was considered valid. In the controls with freshly spiked and with six and twelve month equilibrated soil the number of juveniles exceeded the minimum of 100 per test container set by ISO (1999), and was 140 (56, n ¼ 5), 494 (167, n ¼ 10) and 228 (122, n ¼ 10), respectively. Uncoated ZnO-NP and non-nano ZnO were toxic in freshly spiked soil with no significant difference between the two ZnO powders. No effect on collembolan survival was found in soil equilibrated for three, six or twelve months with uncoated ZnO-NP and non-nano ZnO. After three months equilibration the EC50 of uncoated ZnO-NP increased from 1964 to 2847 mg Zn/kg based on total soil concentrations and from 10.1 to 39.9 mg Zn/l based on porewater concentrations. After six months equilibration no EC50 could be estimated for ZnO-NP due to a flat doseeresponse curve and very high juvenile numbers in the control. However, almost 50% reduction in reproduction was observed for the lowest test concentration (i.e. 116 mg Zn/kg), which was considered unrealistic. After twelve months equilibration, the reduction in reproduction was highest (46%) at 1027 mg Zn/kg, which contained the highest porewater concentration (i.e. 67.1 mg Zn/l). For non-nano ZnO 37% reduction in reproduction was observed at 3628 mg Zn/kg and 65% at 8359 mg Zn/kg after three months equilibration. The data did not allow estimating an EC50, but 50% effect on reproduction is expected between these two exposure concentrations. After six months, no EC50 could be calculated but 62% effect was observed at the highest soil concentration. It was possible to calculate an EC50 of 31.3 mg Zn/l based on porewater concentrations. After 12 months, only 40% reduction in reproduction was found at the highest exposure concentration and therefore

3.2. Soil pore water analysis 3.2.1. ZnCl2 Zn concentrations in pore water collected from soils equilibrated with ZnCl2 increased with exposure concentration at all time points (Table SI-3). The maximum concentration was 612 mg Zn/l in freshly spiked soil at a nominal soil concentration of 1600 mg Zn/kg dry soil. In equilibrated soil the maximum amounted to 2520, 2700 and 2196 mg Zn/l at a nominal soil concentration of 3200 mg Zn/kg dry soil after 3, 6 and 12 months, respectively. 3.2.2. Uncoated ZnO-NP and non-nano ZnO Zinc concentrations in pore water increased from 1.85 to 12.6 mg Zn/l in soil freshly spiked with uncoated ZnO-NP with increasing soil concentrations (Fig. 1, Table SI-3). After three, six and twelve months, porewater concentrations for uncoated ZnO-NP and non-nano ZnO increased with time, but peaked at intermediate concentrations. The highest Zn concentration measured after one year was 67.1 mg Zn/l in soil at the measured total concentration of 1027 mg Zn/kg for uncoated ZnO-NP. Ultrafiltration did not affect the zinc concentrations, suggesting that no intact nanoparticles were present in the soil solutions. For non-nano ZnO a maximum of 66.5 mg Zn/l was measured in the pore water after one year from soil containing a measured concentration of 941 mg Zn/kg. 3.2.3. Coated ZnO-NP For the coated ZnO-NP, zinc concentrations in the pore water increased with increasing soil concentrations during the first three months of equilibration (Fig. 1; Table SI-3). At the highest test concentration, 6.10 mg Zn/l was measured in the pore water in freshly spiked soil, which increased to 22.5, 20.9 and 23.0 mg Zn/l after three, six and twelve months, respectively. The maximum porewater concentration after six and twelve month equilibration was measured in soils spiked with 1600 mg Zn/kg, and was 26.6 T=0

porewater concentration (mg Zn/l)

T=3

coated Zn O-N P

u n coated Zn O-N P

n on -n ano Zn O

T=6

70

T=12 60 50 40 30 20 10 0 0

2000

4000

6000

0

1000

2000

3000

4000

5000

6000 0

2000

4000

6000

8000

measured concentration soil (mg Zn/kg d.w.) Fig. 1. Zinc concentrations measured in pore water (mg Zn/l) as a function of total zinc concentrations in Lufa 2.2 soil (mg Zn/kg d.w.) freshly spiked with coated and uncoated ZnONP, non-nano ZnO and ZnCl2 (T ¼ 0) and after three (T ¼ 3), six (T ¼ 6) and twelve months (T ¼ 12) equilibration.

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Table 1 EC50 values for the effect on the reproduction of Folsomia candida after 28-d exposure to Lufa 2.2 soil freshly spiked (T ¼ 0) with coated and uncoated ZnO-NP, non-nano ZnO and ZnCl2 and after three (T ¼ 3), six (T ¼ 6) and twelve months (T ¼ 12) equilibration. EC50 values are presented as total concentrations in the soil (mg Zn/kg d.w.). Corresponding 95% confidence intervals are presented in between brackets. Time (months) T T T T

¼ ¼ ¼ ¼

0 3 6 12

Coated ZnO-NP

Uncoated ZnO-NP

Non-nano ZnO

ZnCl2

873a 749a 576a 1817b

1964 (1635e2293) 2847 (e)

1591 (e) 3628 < EC50 < 8359 e >8359*

299a (181e415) 912b (e) e 707b (419e996)

(659e1087) (463e1035) (263e888) (1344e2291)

e

>5855*

*No 50% reduction in survival or reproduction was observed at the highest test concentration. e Data did not allow estimating an EC50 value and/or 95% confidence intervals (see text). a,b indicate significant differences between LC/EC50 values at different time points according to a generalized likelihood-ratio test (c2(1) > 3.84; p < 0.05).

no EC50 was calculated. Overall, the toxicity tests with ZnO showed a decrease in springtail toxicity with time, which continued for one year. The toxicity of ZnCl2decreased with time shown by an increase in LC50 values from 955 (95% CI 819e1114) to 1632 (1336e1927) and 1943 (1300e2586) mg Zn/kg in freshly spiked soil and after 3 and 6 months, respectively. After one year equilibration of ZnCl2spiked soils springtail survival was no longer affected (LC50 > highest test concentration). The effect on reproduction also decreased with time shown by a three-fold increase in EC50 after three months equilibration. EC50 values were lowest for freshly spiked soil and increased after equilibration, also when expressed as porewater Zn concentrations. For the test with six month equilibrated soil no EC50 could be estimated due to the very high juvenile numbers in the control. Reproduction was already reduced by 65% at the lowest test concentration (123 mg Zn/kg), which was considered to be unrealistic and therefore no EC50 was calculated. In the second experiment (including coated ZnO-NP), survival of the Collembola in the controls of the four toxicity tests performed with freshly spiked soil, three, six and twelve month equilibrated soil was 88, 85, 57 and 85%, respectively. The numbers of juveniles in the control with freshly spiked, three, six and twelve month equilibrated soil were 483 (96, n ¼ 10), 331 (142, n ¼ 10), 258 (114, n ¼ 10), 315 (96, n ¼ 10), respectively. No effect on collembolan survival was found in soil freshly spiked or equilibrated with coated ZnO-NP. The effect on reproduction in freshly spiked, three and six month equilibrated soil was similar, as shown by the EC50 values of 873, 749 and 576 mg Zn/kg, respectively. After one year equilibration the EC50 increased to 1817 mg Zn/kg. Based on soluble Zn concentrations the EC50 values increased from 4.08 to 15.8e18.0 mg Zn/l after three to twelve months equilibration. For the effect on collembolan survival and reproduction of triethoxyoctylsilane a LC50 of 653 mg/kg (95% CI ¼ 514e793 mg/kg), an EC50 of 505 mg/kg (376e633 mg/kg) and an EC10 of 271 (106e437 mg/ kg) were estimated (Fig. SI-6). 4. Discussion Several studies demonstrate that ZnO rapidly dissolves in soils. Priester et al. (2012) found high Zn accumulation in the leaves of

soybeans after 50 days exposure to ZnO-NP in a soil at pH 6.78. As they measured similar Zn values for ZnCl2-exposed plants, this suggests most of the ZnO was dissolved. In soils with pH < 7 ZnO dissociates into free Zn2þ ions (Bodar et al., 2005). After dissociation, adsorption of Zn to various soil components such as organic matter, clay minerals and Fe and Mn (hydr)oxides occurs. Voegelin et al. (2005) showed a rapid dissolution of ZnO in a near neutral non-calcareous soil (pH 6.5) and rapid formation of Zn precipitates. They also found that dissolved Zn species in the soil do not diffuse from the solid phase where these Zn precipitates are formed rapidly. In soils at field capacity, slow diffusion at the surfaces of metal oxides and humics inhibits further dissolution of ZnO (Crout et al., 2006). This has also been observed in kaolin suspensions at pH 6, where ZnO-NP rapidly converted to Zn2þ that was bound in sorption complexes within one day and these sorption complexes were maintained through a 12-month ageing period (Scheckel et al., 2010). Donner et al. (2010) spiked different field soils with ZnSO4 and aged them for up to two years and found a decrease of available Zn with time. So, it can be assumed that ZnO-NP rapidly dissolved in soil and that the dissolution of ZnO was counteracted by fixation of Zn over time. In this study, the porewater concentrations showed an interesting trend in Zn dissolution from both uncoated ZnO-NP and nonnano ZnO, with increasing Zn dissolution with time at low concentrations but not at high doses. Formation of agglomerates at high exposure concentrations could have hindered dissolution. However, it is more likely that the slightly increased pH levels with increasing exposure concentrations have reduced ZnO solubility. At higher soil concentrations there may be a higher degree of precipitation of released Zn ions, with sulphate, phosphate and other ions (McLaughlin, 2002) and metal fixation increases with pH (Crout et al., 2006). During the one-year equilibrium period the pH slightly decreased with time, explaining the increase of porewater concentrations with time. Several studies have investigated the influence of primary particle size on NP solubility in water. David et al. (2012) observed similar dissolution rates of ZnO-NP with a primary particle diameter above 20 nm and non-nano ZnO, but a higher solubility for 6 nm ZnO-NP. Bian et al. (2011) showed that 15 nm ZnO-NP dissolve more readily in water than 240 nm ZnO. In our soils we found no

Table 2 EC50 values for the effect on the reproduction of Folsomia candida after 28-d exposure to Lufa 2.2 soil freshly spiked (T ¼ 0) with coated and uncoated ZnO-NP, non-nano ZnO and ZnCl2 and after three (T ¼ 3), six (T ¼ 6) and twelve (T ¼ 12) months equilibration. EC50 values are presented as soluble Zn levels in the pore water (mg Zn/l). Corresponding 95% confidence intervals are presented in between brackets. Time (months) T T T T

¼ ¼ ¼ ¼

0 3 6 12

Coated ZnO-NP

Uncoated ZnO-NP

Non-nano ZnO

ZnCl2

4.08a 16.5b 15.8b 18.0b

10.1 (7.83e12.4) 39.9 (e) e >67.1*

7.94 (e) e 31.3 (11.6e51.0) >66.5*

16.8a (e) 180b (e) e 178b (142e213)

(3.86e4.30) (14.4e18.7) (12.4e19.2) (1.85e34.1)

*No 50% reduction in survival or reproduction was observed at the highest test concentration. e Data did not allow estimating an EC50 value and/or 95% confidence intervals (see text). a,b indicate significant differences between EC50 values at different time points according to a generalized likelihood-ratio test (c2(1) > 3.84; p < 0.05).

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clear differences in Zn dissolution between soils spiked with uncoated ZnO-NP and non-nano ZnO, despite the fact that the ZnO-NP were more than 50 times smaller than non-nano ZnO in the initial powders. So, the total surface area or primary particle size does not seem to play a major role in the dissolution of this type of ZnO-NP and this type of soil. We did observe lower porewater Zn levels for coated ZnO-NP. In general, a surface coating is used to prevent aggregation of nanoparticles by steric or electrostatic mechanisms (Christian et al., 2008). Soluble Zn concentrations similar to the ones obtained from uncoated ZnO-NP after three months were reached only after twelve months equilibration. This suggests that a surface coating could delay the dissolution of ZnO-NP in soil. However, it is more likely that the higher pH levels for soils with coated ZnO-NP compared to the uncoated ZnO-NP have caused the lower Zn levels. The fact that the peak in dissolved Zn appeared much later in case of coated ZnO-NP also suggests that the coating must have prohibited Zn release, and that it took at least several months to degrade the coating material. This seems in line with the half-life for triethoxyoctylsilane in soil of approx. 30 days estimated by a level III fugacity model in EPI-SuiteÔ. Both types of ZnO-NP (Z-COTEÒ and Z-COTEÒHP1) have been tested for acute effects on Daphnia magna. Wiench et al. (2009) report 48-h EC50 values of 7.5 and 1.1 mg Zn/l for the uncoated and coated ZnO-NP, respectively. Also, in our toxicity tests EC50 values were higher for uncoated than for coated ZnO-NP, both when expressed on the basis of total soil concentrations and on porewater concentrations (1964 mg Zn/kg or 10.1 mg Zn/l versus 873 mg Zn/kg or 4.08 mg Zn/l; at T ¼ 0). The porewater-based EC50 values are in the same range as the values reported by Wiench et al. (2009) for daphnids. The hydrophobic character of the coated NP may have enhanced its toxicity. The coating triethoxyoctylsilane, which represents approx. 3% of the ZnO-NP, may have contributed to the overall toxicity. When applying a mixture toxicity approach, taking into account that on a mass basis the coating represented 3.85% of the Zn mass, it may be assumed that the EC50 for coated ZnO-NP of 873 mg Zn/kg dry soil corresponds with a triethoxyoctylsilane concentration of 33.6 mg/kg dry soil. Compared with the EC50 values for uncoated ZnO-NP and triethoxyoctylsilane these concentrations correspond with 0.44 and 0.067 Toxic Units (TU), respectively. This means that the toxic strength of the mixture equals 0.5 TU, so the coated ZnO-NP are much more toxic than expected from the EC50 values of uncoated ZnO-NP and triethoxyoctylsilane. The decrease of toxicity of the coated ZnO-NP at the end of the 12-month equilibration period may be explained from the loss of the coating. In our ZnCl2-spiked soils toxicity decreased with time, while the dissolved Zn increased from 17.1% (of total Zn concentration at the highest spiking concentration) for freshly spiked soil to 34.9, 37.4 and 30.4% after three, six and twelve months equilibration, respectively. For the T ¼ 0 samples and the samples from the test with coated ZnO-NP, it seemed that toxicity could fairly well be explained from the Zn concentration in pore water, with EC50 values ranging between 8 and 20 mg Zn/l (Table 2). At later sampling times, toxicity of the uncoated ZnO-NP, non-nano ZnO and ZnCl2 could however, no longer be explained from porewater Zn concentrations. Since this might be due to complexation of Zn in the pore water, we modelled Zn speciation using Visual MINTEQ for measured soluble Zn concentrations of 1.18 and 2700 mg Zn/l (lowest and highest values measured in pore water). Speciation was also calculated for concentrations of 49.2 and 266 mg Zn/l which are closest to the EC50 values estimated at time zero and after six months, using a measured dissolved organic carbon content (NicaDonnan DOC) of 315 mg/l for Lufa 2.2 soil (t ¼ 20  C). For these speciation calculations, we also needed porewater pH values.

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Unfortunately, we applied different methods for determining pH of the soil and pore water. Based on pHwater of pHCaCl2 values, porewater pH values were estimated using regressions derived by De Vries et al. (2005). These porewater pH values (Table SI-2), which showed a consistent decline with time, were used for speciation calculations. For ZnCl2 concentrations up to 49.2 mg Zn/l more than 95% of the dissolved Zn was estimated to be bound to DOC and not available in solution as Zn2þ. At 266 mg Zn/l 65.3% was available as Zn2þ, and at 2700 mg Zn/l 80.9% (Table SI-4). The increase in free zinc concentrations in soil solution coincides with a decrease in pH. According to the principles of the Biotic Ligand Model, competition of Hþ ions and free zinc on the uptake sites of the animal therefore may explain the decreased toxicity of ZnCl2 over time (Thakali et al., 2006). For the ZnO-NP and non-nano ZnO in our study, all Zn measured in the pore water was bound to DOC. When plotting the estimated EC50 values based on porewater concentrations against pH, a significant negative correlation was found for all four Zn forms (Fig. SI-7). For the coated ZnO-NP, EC50 with pH showed a less steep decline, which may be attributed to the degradation of the coating which apparently delayed release of Zn. In this study we investigated the effect of time on ZnO-NP dissolution in soil and found that pH is a main factor. We conclude that the release of Zn in soils spiked with coated and uncoated ZnO-NP continued for one year and that the fixation of Zn contributed to a reduced bioavailability and springtail toxicity with time. Acknowledgement The work reported here was conducted in the context of NanoFATE, Collaborative Project CP-FP 247739 (2010e2014) under the 7th Framework Programme of the European Commission (FP7NMP-ENV-2009, Theme 4), coordinated by C. Svendsen and D. Spurgeon of NERC e Centre for Ecology and Hydrology, UKWallingford; www.nanofate.eu. BASF SE (Ludwigshafen, Germany) kindly provided Z-COTEÒ and Z-COTEÒHP1 (nanoscale zinc oxide) as test material through the PROSPECT project. The authors thank Dr. Kerstin Jurkschat (Department of Materials Oxford University UK) for performing the transmission electron microscopy. Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.envpol.2013.03.003. References Apte, S.C., Rogers, N.L., Batley, G.E., 2009. Ecotoxicology of manufactured nanoparticles. In: Lead, Smith (Eds.). Blackwell Publishing Ltd, pp. 267e301. Bian, S.W., Mudunkotuwa, I.A., Rupasinghe, T., Grassian, V.H., 2011. Aggregation and dissolution of 4 nm ZnO nanoparticles in aqueous environments: influence of pH, ionic strength, size, and adsorption of humic acid. Langmuir 27, 6059e6068. Bodar, C.W.M., Pronk, M.E.J., Sijm, T.H.M., 2005. The European Union risk assessment on zinc and zinc compounds: the process and the facts. Integrated Environment Assessment Management 1, 301e319. Borm, P., Klaessig, F.C., Landry, T.D., Moudgil, B., Pauluhn, J., Thomas, K., Trottier, R., Wood, S., 2006. Research strategies for safety evaluation of nanomaterials, Part V: role of dissolution in biological fate and effects of nanoscale particles. Toxicological Sciences 90, 23e32. Canas, J.E., Qi, B., Li, S., Maul, J.D., Cox, S.B., Das, S., Green, M.J., 2011. Acute and reproductive toxicity of nano-sized metal oxides (ZnO and TiO2) to earthworms (Eisenia fetida). Journal of Environmental Monitoring 13, 3351e3357. Christian, P., von der Kammer, F., Baalousha, M., Hofmann, Th, 2008. Nanoparticles: structure, properties, preparation and behaviour in environmental media. Ecotoxicology 17, 326e343. Crout, N.M.J., Tye, A.M., Zhang, H., McGrath, S.P., Young, S.D., 2006. Kinetics of metal fixation in soils: measurements and modelling by isotopic dilution. Environmental Toxicology and Chemistry 25, 659e663.

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David, C.A., Galceran, J., Rey-Castro, C., Puy, J., Companys, E., Salvador, J., Monne, J., Wallace, R., Vakourov, A., 2012. Dissolution kinetics and solubility of ZnO nanoparticles followed by AGNES. Journal of Physical Chemistry 116, 11758e 11767. De Vries, W., Schülze, Lofts, S., Tipping, E., Meili, M., Römkens, P.F.A.M., Groenenberg, J.E., 2005. Calculation of Critical Loads for Cadmium, Lead and Mercury; Background Document to a Mapping Manual on Critical Loads of Cadmium, Lead and Mercury. Alterra-report 1104. Alterra, Wageningen, The Netherlands. Donner, E., Broos, K., Heemsbergen, D., Warne, M.S.J., McLaughlin, M.J., Hodson, M.E., Nortcliff, S., 2010. Biological and chemical assessments of zinc ageing in field soils. Environmental Pollution 158, 339e345. Fountain, M.T., Hopkin, S.P., 2005. Folsomia candida (Collembola): a “standard” soil arthropod. Annual Review Entomology 50, 201e222. Gustafsson, J.P., 2007. A Windows Version of MINTEQA2 Version 4.0, MINTEQA2 was Released by the US Environmental Protection Agency in 1999. Visual MINTEQ Version 2.53. Haanstra, L., Doelman, P., Oude Voshaar, J.H., 1985. The use of sigmoidal dose response curves in soil ecotoxicological research. Plant and Soil 84, 293e297. Hobbelen, P.H.F., Koolhaas, J.E., van Gestel, C.A.M., 2004. Risk assessment of heavy metal pollution for detritivores in floodplain soils in the Biesbosch, the Netherlands, taking bioavailability into account. Environmental Pollution 129, 409e419. Hooper, H.L., Jurkschat, K., Morgan, A.J., Bailey, J., Lawlor, A.J., Spurgeon, D.J., Svendsen, C., 2011. Comparative chronic toxicity of nanoparticulate and ionic zinc to the earthworm Eisenia veneta in a soil matrix. Environment International 37, 1111e1117. ISO, 1999. Soil QualitydInhibition of Reproduction of Collembola (Folsomia candida) by Soil Pollutants. ISO 11267. International Standardization Organization, Geneva. Kahru, A., Dubourguier, H.-C., 2010. From ecotoxicology to nanoecotoxicology. Toxicology 269, 105e119. Kool, P.L., Diez Ortiz, M., van Gestel, C.A.M., 2011. Chronic toxicity of ZnO nanoparticles, non-nano ZnO and ZnCl2 to Folsomia candida (Collembola) in relation to bioavailability in soil. Environmental Pollution 159, 2713e2719. Lock, K., Janssen, C.R., 2002. The effect of ageing on the toxicity of zinc for the potworm Enchytraeus albidus. Environmental Pollution 116, 289e292. McLaughlin, M.J., 2002. Bioavailability of metals to terrestrial plants. In: Allen, H. (Ed.), Bioavailability of Metals in Terrestrial Ecosystems: Importance of Partitioning for Bioavailability to Invertebrates, Microbes and Plants. Society of Environmental Toxicology and Chemistry(SETAC), USA. Nota, B., Verweij, R.A., Molenaar, D., Ylstra, B., van Straalen, N.M., Roelofs, D., 2010. Gene expression analysis reveals a gene set discriminatory to different metals in soil. Toxicological Sciences 115, 34e40. Osmond, M.J., McCall, M.J., 2010. Zinc oxide nanoparticles in modern sunscreens: an analysis of potential exposure and hazard. Nanotoxicology 4, 15e41. Poynton, H.C., Lazorchak, J.M., Impellitteri, C.A., Smith, M.E., Rogers, K., Patra, M., Hammer, K.A., Allen, H.J., Vulpe, C.D., 2011. Differential gene expression in

Daphnia magna suggests distinct modes of action and bioavailability for ZnO nanoparticles and Zn ions. Environmental Science and Technology 45, 762e768. Priester, J.H., Ge, Y., Mielke, R.E., Horst, A.M., Moritz, S.C., Espinosa, K., Gelb, J., Walkerg, S.L., Nisbet, R.M., Ani, Y.-J., Schimel, J.P., Palmer, R.G., HernandezViezcas, J.A., Zhao, L., Gardea-Torresdey, J.L., Holden, P.A., 2012. Soybean susceptibility to manufactured nanomaterials with evidence for food quality and soil fertility interruption. Proceedings of the National Academy of Sciences, Early Edition, 1e6. Reed, R.B., Ladner, D.A., Higgins, P., Westerhoff, P., Ranville, J.F., 2012. Solubility of nano-zinc oxide in environmentally and biologically important matrices. Environmental Toxicology and Chemistry 31, 93e99. Scheckel, K.G., Luxton, T.P., El Badawy, A.M., Impellitteri, C.A., Tolaymat, T.M., 2010. Synchrotron speciation of silver and zinc oxide nanoparticles aged in a kaolin suspension. Environmental Science and Technology 44, 1307e1312. Smit, C.E., van Beelen, P., van Gestel, C.A.M., 1997. Development of zinc bioavailability and toxicity for the springtail Folsomia candida in an experimentally contaminated field plot. Environmental Pollution 98, 73e80. Sokal, R.R., Rohlf, F.J., 1995. Biometry, third ed. W.H. Freeman, San Francisco, CA, USA. Thakali, S., Allen, H.E., Di Toro, D.M., Ponizovsky, A.A., Rooney, C.P., Zhao, F.J., McGrath, S.P., 2006. A Terrestrial Biotic Ligand Model. 1. Development and application to Cu and Ni toxicities to barley root elongation in soils. Environmental Science and Technology 40, 7085e7093. Theng, B.K.G., Yuan, G., 2008. Nanoparticles in the soil environment. Elements 4, 395e399. Tipping, E.J., Rieuwerts, Pan, G., Ashmore, M.R., Lofts, S., Hill, M.T.R., Farago, M.E., Thornton, I., 2003. The solidesolution partitioning of heavy metals (Cu, Zn, Cd, Pb) in upland soils of England and Wales. Environmental Pollution 125, 213e225. Tourinho, P.S., van Gestel, C.A.M., Lofts, S., Svendsen, C., Soares, A.M.V.M., Loureiro, S., 2012. Metal-based nanoparticles in soil: fate, behavior and effects on soil invertebrates. Environmental Toxicology and Chemistry 31, 1679e1692. Van der Ploeg, M.J.C., Baveco, J.M., van der Hout, A., Bakker, R., Rietjens, I.M.C.M., van den Brink, N.W., 2011. Effects of C60 nanoparticle exposure on earthworms (Lumbricus rubellus) and implications for population dynamics. Environmental Pollution 159, 198e203. Voegelin, A., Pfister, S., Scheinost, A.C., Marcus, M.A., Krestzschmar, R., 2005. Changes in zinc speciation in field soil after contamination with zinc oxide. Environmental Science and Technology 39, 6616e6623. Wang, Z.L., 2004. Zinc oxide nanostructures: growth, properties and applications. Journal of Physics: Condensed Matter 16, 829e858. Wiench, K., Wohlleben, W., Hisgen, V., Radke, K., Salinas, E., Zok, S., Landsiedel, R., 2009. Acute and chronic effects of nano- and non-nano-scale TiO2 and ZnO particles on mobility and reproduction of the freshwater invertebrate Daphnia magna. Chemosphere 76, 1356e1365. Wong, S.W.Y., Leung, P.T.Y., Djurisi c, A.B., Leung, K.M.Y., 2010. Toxicities of nano zinc oxide to five marine organisms: influences of aggregate size and ion solubility. Analytical and Bioanalytical Chemistry 396, 609e618.