Aggregation behavior of engineered nanoparticles and their impact on activated sludge in wastewater treatment

Chemosphere 119 (2015) 568–576

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Aggregation behavior of engineered nanoparticles and their impact on activated sludge in wastewater treatment Zhou Xiao-hong ⇑, Huang Bao-cheng, Zhou Tao, Liu Yan-chen, Shi Han-chang State Key Joint Laboratory of Environmental Simulation and Pollution Control, School of Environment, Tsinghua University, Beijing 100084, China

h i g h l i g h t s

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 NPs remained stable aggregates in 1000

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wastewater after 4.5 h or more suspension.  Presence of monovalent and divalent electrolytes cannot increase aggregation of NPs.  O2 fluxes of activated sludge were damaged after treating with NPs up to 4.5 h.

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Article history: Received 5 March 2014 Received in revised form 4 July 2014 Accepted 11 July 2014 Available online 13 August 2014 Handling Editor: I. Cousins Keywords: Aggregation Nanoparticles Wastewater Activated sludge Noninvasive measurement

a b s t r a c t The ever-increasing daily use of engineered nanoparticles will lead to heightened levels of these materials in the environment. These nanomaterials will eventually go into the wastewater treatment plant (WWTP), therefore, resulting into a pressing need for information on their aggregation behavior and kinetics in the wastewater aqueous matrix. In this work, we dispersed two different metal oxide nanoparticles (ZnO and TiO2) into the influent of two different WWTPs. Through the time-resolved dynamic light scattering analysis and transmission electron microscopy, the metal oxide nanoparticles (NPs) were quite stably existed in the wastewater matrix with aggregates of diameter 300–400 nm after 4.5 h or more suspension. We confirmed that the dissolved organic matters (DOMs) attributed to the stability of nanoparticles. No propensity of NPs to aggregate were observed in the presence of both monovalent and divalent electrolytes even at high concentrations up to 0.15 M in NaCl or 0.025 M in CaCl2, indicating that the destabilization of nanoparticles in the complicated wastewater matrix was not achieved by the compression of electrical double layer, therefore, their aggregation kinetics cannot be simply predicted by the classic Derjaguin–Landau–Verwey–Overbeek theory of colloidal stability. However, obvious aggregation of nanoparticles in the Al2(SO4)3 solution system was observed with the likely mechanism of bridging of the metal oxide nanoparticles and aggregates due to the formation of hydrous alumina (Al(OH)3
H2O) in the Al2(SO4)3 solution. In the wastewater matrix, we used the noninvasive measurement technology to detect the O2 flux of activated sludge before and after treatment with 1, 10 and 100 mg L 1 NPs. The results confirmed that both ZnO and TiO2 NPs showed an adverse impact on the O2 uptake of activated sludge when the exposure time extended to 4.5 h. Ó 2014 Elsevier Ltd. All rights reserved.

⇑ Corresponding author. Tel.: +86 10 62796953; fax: +86 10 62771472. E-mail address: [email protected] (X.-h. Zhou). http://dx.doi.org/10.1016/j.chemosphere.2014.07.037 0045-6535/Ó 2014 Elsevier Ltd. All rights reserved.

X.-h. Zhou et al. / Chemosphere 119 (2015) 568–576

1. Introduction The increasing use of nanomaterials in consumer products has led to increased concerns about their potential environmental and health impacts; however, many of the answers remain largely unknown. Generally speaking in terms of destination, the nano-scale pollutants would eventually go into the wastewater treatment systems. Therefore better understanding of the transport, fate, and behavior of the nanomaterials in wastewater samples is essential. The aggregation of engineered nanomaterials is a critical factor affecting their fate and toxic effects in the aqueous environment, which have received wide attention. Until now, numerous researches have focused on the aggregation behaviors of nanomaterials in natural or simulated natural water bodies (Chen and Elimelech, 2006, 2007; Zhang et al., 2008; Liu et al., 2009, 2010; Zhang et al., 2009; Keller et al., 2010; Petosa et al., 2010; Weinberg et al., 2011). The composition and characteristics of natural water can change the nanoparticle surface properties, and thus their aggregation and dispersion behavior (Weinberg et al., 2011). The proven factors contain the natural organic matter (NOM), metal ions, pH, ionic strength (IS) and so on. The presence of NOM will change the aggregation degree of nanomaterials. Tendency of aggregation of most metal oxide nanoparticles was weakened by adsorption to NOM, thus resulting in the increasing stability in the natural water bodies (Zhang et al., 2009; Keller et al., 2010). Studies also showed with the addition of 1 mg L 1 NOM, the negative surface charge of nanoparticles increased significantly and therefore their propensity to aggregate is reduced (Zhang et al., 2009). On the other hand, the negative charge that NOM imparted to nanoparticles could be neutralized by divalent cations (Ca2+), especially in the concentration range from 0.04 to 0.06 M (Zhang et al., 2009). In addition, the variation in the pH value of natural water will obviously change the zeta potential (f) of metal oxide nanoparticles, therefore affecting their stability (Liu et al., 2010). Most studies proved that the aggregation kinetics of nanoparticles corresponded with the classic Derjaguin–Landau– Verwey–Overbeek (DLVO) model (Chen and Elimelech, 2006, 2007; Keller et al., 2010; Liu et al., 2009, 2010), however, the bond bridging action among the humic acid, Ca2+ and C60, which was explained by the steric effects was also be revealed (Chen and Elimelech, 2006). Moreover, most common experimental and theoretical approaches used for evaluation of nanomaterial deposition and aggregation are applicable for spherical particles; however, it has been noted above certain limitations for nonspherical or very small particles (Petosa et al., 2010). Although no definitive unified understanding of nanoparticles aggregation in pure water and natural water bodies have been reached, it is no doubt that the related researches have been extensively conducted. However, almost no systematical exploration of aggregation in the wastewater samples are conducted as far as we know. Current researches mostly focus on the impact of nanomaterials on the wastewater treatment system and their removal in WWTPs (Limbach et al., 2008; Zhang et al., 2008; Brar et al., 2010; Gómez-Rivera et al., 2012; Chen et al., 2012). However, there is still dearth of research aiming at the stability of NPs in wastewater environments even though the following reason. Wastewater composition and characteristics show big differences compared with those of both pure water and natural water, such as the type and extent of metal ions, organic composition, chemical agents commonly used in phosphorus removal, which prohibits the direct application of the acquired results in pure water and natural water bodies into the wastewater samples. Noninvasive measurement technology (NMT) is a powerful tool to monitor ions and molecule flux across the biotic membrane. It

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has been used for decades in agriculture (Porterfield and Smith, 2000) and biomedical application (McLamore et al., 2010a), and has recently been applied in environmental monitoring (McLamore et al., 2009). The biggest advantage of NMT is the high signal-to-noise ratio which would significantly improve the measurement accuracy. But the applications of NMT mainly focus on biofilm (McLamore et al., 2010b), few on suspended growth bacteria exposed with nanomaterials are reported. As a consequence, the aim of the study is to examine the characteristics, dispersion and stability of engineered nanomaterials in wastewater as well as their impact on the activities of activated sludge through the NMT analysis. Two nanoparticles, TiO2 and ZnO were selected for the experiments because they are among the most common metal oxide nanoparticles in use (Keller et al., 2013). Results would shed light on the aggregation behaviors of NPs in wastewater matrix and provide basic data to assess their fate in the wastewater environment and potential risks to the wastewater treatment system and received aquatic environment. 2. Material and methods 2.1. Materials and reagents Two engineered nanoparticles, TiO2 nanopowder with 21 nm particle diameter size characterized by transmission electron microscope (TEM) (No. 718467), P99.5% trace metals basis and 35–65 m2 g 1 (BET), and ZnO nanopowder with less than 100 nm particle diameter size (No. 544906), 79.1–81.5% in %Zn and 15–25 m2 g 1 (BET), were obtained from Sigma–Aldrich. Chemicals used in the experiments were analytical grade if not specified. 2.2. Wastewater samples characterization Wastewater samples were taken from the influent of XiaohongMen (XM) and XiaojiaHe (XH) municipal WWTPs in Beijing, China. The particles in the samples were removed through a filter with 0.22 lm nylon membrane and stored at 4 °C for use. The filtered wastewater kept no more than one week during the whole experiments. Dissolved total organic carbon (dTOC) was measured by a Shimadzu TOC-V instrument (Shimadzu scientific instruments) using the combustion-infrared method and represented the dissolved organic matters (DOMs) in wastewater samples. The pH was measured using a pH meter. Dissolved oxygen concentration was measured by using an oxygen electrode (Hach, LDO™). Conductivity was measured with a conductivity meter (Mettler Toledo). The concentrations of various cations and anions were measured by the ion chromatograph (DIOX ICS1000), which were also used to calculate the ionic strength (IS). Inductive coupled plasma-atomic emission spectrometry (ICP-AES, Thermo IRIS) was used to measure the concentrations of metal elements in the wastewater samples and samples containing NPs. The acid decomposition pretreatment stepped as follows: 10 mL sample was added into the 25 mL glass beaker, which was evaporated to dryness (wet salts and particles) on the oven. And then 0.06 g (NH4)2SO4 and 3 mL H2SO4 were added into the beaker and heated on the oven until the mixture solution appeared colorless. The residue was cooled and transferred into a 50 mL cuvette, diluted to volume with DI water and filtered by 0.22 lm membrane before the ICP-AES measurement. The surface tension of wastewater was measured using the Wilhelmy plate method (Drelich et al., 2002). The Excitation-Emission Matrix (EEM) fluorescence spectroscopy was used to characterize the DOM components contained in the wastewater matrix with a luminescence spectrometry (F-7000 FL Spectrophotometer, Hitachi, Japan). The EEM was generated for each sample by scanning over an excitation wavelength (Ex)

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between 220 and 600 nm at an increment of 10 nm, and an emission wavelength (Em) between 230 and 650 nm at an increment of 10 nm. The excitation and emission slit widths were set at 10 nm, and a scanning speed of 1200 nm min 1 was applied for all of the measurements. The spectrum of pure water was used as the blank. In order to explore the factors causing the stability of nanoparticles in wastewater samples, we also prepared the nonDOMs-contained synthetic wastewater samples by using the salts of Na2SO4, NaCl to simulate the IS and pH conditions of real wastewater samples as shown in Table S1.

30 ppm, 50 ppm and 100 ppm). The mixture containing electrolyte or glucose was used for DLS measurement as soon as possible. The hydrodynamic diameter (intensity-based) is calculated using the Stokes–Einstein equation. All DLS measurements were conducted at 25 °C. The well known Helmholtz–Smoluchowski relationship was applied to convert measured electrophoretic mobilities to the f potential (Jacobasch and Schurz, 1988; Liu et al., 2009). All electrophoretic mobility measurements and aggregation experiments were conducted at a pH of 7.8–8.2 (adjusted with NaOH or HCl) to simulate the real wastewater conditions (Table 1). 2.5. Transmission electron microscopy

2.3. Nanoparticle suspension in wastewater matrix Nanomaterial stock suspensions of nanoparticles of 1 g L 1 were prepared by dispersing nanoparticles into the nanopure water (Thermo Barnstead) through sonicated for 1 h at 250 W. The stock suspensions were stored at 4 °C and used for no more than one week in the whole experiments. The nanoparticle suspensions in wastewater matrix were prepared as the following steps. 10 mL glass cuvette were cleaned before use by being soaked overnight in 2% HCl solution, rinsed with excess deionized water, and oven-dried under dust-free conditions. The cuvette was only used once as suggested by Chen and Elimelech (2006). 8 mL 0.22-lm-filtered wastewater was placed into the cuvette. After nanomaterial stock solution was sonicated for 30 min at 250 W, 0.1 mL dispersed stock solution was quickly transferred into the cuvette by using the micropipette. The left volume was filled with the 0.22-lm-filtered wastewater or simulated wastewater to final volume of 10 mL. The final concentration of nanomaterial dispersed in wastewater or synthetic wastewater was 10 mg L 1. The choice of the 10 mg L 1 NPs concentration presented in details was because it had been used in many previous studies of NP aggregation, such as Keller et al. (2010), Zhang et al. (2008) and Zhang et al. (2009), which makes it easier to compare our results with other reported ones.

Transmission electron microscopy (TEM) data for some samples were provided to shed light on the DLS measurements as shown by Liu et al. (2009). Metal oxide nanoparticle morphologies in wastewater conditions and their changes due to the addition of electrolytes were also characterized and observed through TEM imaging. TEM images were taken using a carbon support copper film substrate with a mesh of 200 (Beijing Zhongxing Bary Technology Co., China). One drop of the nanoparticle suspension in wastewater matrix was placed on the substrate in a dust-free condition. The prepared substrate was supported on a fresh filter paper for overnight to make the remaining solution evaporated. For the TEM experiments in the presence of electrolyte of Al2(SO4)3, the procedures were the same, except that a final concentration of 40 mM electrolyte stock solution was introduced into the nanoparticle suspension to stay for 4 h. A layer of white deposit was observed on the bottom of cuvette, which was extracted out for TEM observation. 2.6. Noninvasive measurement by oxygen microelectrode The activated sludge mixture was taken from the aeration tank of XH WWTP. 2 mL mixture was centrifuged at 5000 rpm for Table 1 Characteristics of wastewater samples used in the experiments. Units

2.4. Dynamic light scattering and aggregation experiment The nanoparticle behavior of aggregation was investigated by the particle size distribution in terms of hydrodynamic diameter by dynamic light scattering (DLS) using a particle size analyser (Delsa Nano C Particle Analyzer, Beckman Coulter). It employs an all-solid-state laser with a fixed wavelength of 658 nm as a light source. The intensity of scattered light was measured by a detector at 165° and an auto-correlation function was accumulated for 70 times. For each experiment, 3 mL nanoparticle suspension was introduced into an acrylic disposable cuvette (Brookhaven instruments corporation, USA) for DLS measurement as soon as the suspensions were sonicated for 30 min at 250 W. For the aggregation experiment in the presence of electrolyte in the filtered wastewater matrix containing DOMs, the procedure was same, except that a predetermined amount of electrolyte stock solution was introduced into the sonicated nanoparticle suspension to a series of final concentrations of NaCl (0 M, 0.01 M, 0.05 M, 0.1 M, 0.15 M), CaCl2 (0 M, 0.005 M, 0.01 M, 0.015 M and 0.025 M) and Al2(SO4)3 (0 M, 0.00005 M, 0.0001 M, 0.00025 M, 0.0005 M, 0.00075 M, 0.001 M and 0.0015 M), respectively. For the aggregation experiment in the presence of glucose in the non-DOMs-contained synthetic wastewater, a predetermined amount of glucose stock solution was introduced into the sonicated nanoparticle suspension to a series of final concentrations of glucose (0 ppm, 10 ppm,

pH Conductivity dTOC FClSO24 NO-2 NO3PO34 HCO-3 + K Na+ Mg2+ Ca2+ IS Al As Ba Cd Cr Cu Ni Ti Fe Mn Pb Se Sn Zn a

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N.D. means ‘‘Not detected’’.

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7.82 1503 64.9 14.956 126.334 106.284 N.D. 6.336 2.75 181 30.02 82.55 34.39 77.72 1.49  10 N.D. N.D. 0.0438 N.D. N.D. N.D. N.D. 0.0358 0.0852 0.0417 0.22 0.0711 0.2374 0.0219

8.15 1439 65.2 12.798 102.852 93.28 N.D. 5.388 2.154 204 21.93 69.65 31.11 81.24 1.40  10 0.1149 N.D. 0.0738 0.0003 0.221 N.D. N.D. 0.0122 0.0656 0.0494 N.D. 0.0042 0.128 0.0561

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3. Results and discussion 3.1. Compositions of wastewater samples The compositions of the filtered wastewater samples from two municipal WWTPs are presented in Table 1. A large number of previous studies have proven that the combined effect of IS and DOMs and other characteristics of the solution medium results in either aggregation or stabilization (Zhang et al., 2009; Keller et al., 2010; Liu et al., 2010). Through the determination of pH, conductivity, dissolved TOC (dTOC), anion and cation concentrations, the water quality of influent samples taken from two WWTPs showed slightly difference. Estimated ISs of the wastewater taken from XM and XH were 1.49  10 2 and 1.40  10 2 eq L 1, respectively and dTOC concentrations were 64.9 and 65.2 mg L 1, respectively. Overall the metal elements were at low concentrations, and in particular As, Cu and Ni concentrations were nondetectable in both wastewater samples. Studies show that the IS value in this range is between those of the freshwater and seawater (Keller et al., 2010), and very close to the wastewater property (IS = 1.37  10 2 M L 1) reported by Ganesh et al. (2010). However, the TOC values are significantly variable for different studies (Ganesh et al., 2010; Keller et al., 2010). The surface tension of wastewater samples was measured to be 54.32 ± 0.79 mN m 1 for XM WWTP and 49.04 ± 0.87 mN m 1 for XH WWPT (n = 3). In order to confirm that the interference of the particulate material in the wastewater samples has been excluded, the average diameter of the wastewater samples was determined to be close to zero by DLS. 3.2. Aggregation and stability of nanoparticles in the wastewater samples The cytotoxicities of NPs on a living cell were found to strongly depend on the material’s aggregation properties, making kinetic measurements rather challenging (Limbach et al., 2008). The stability of the two metal oxide nanoparticles in the wastewater samples were tested respectively. Fig. 1 shows a representative aggregation profile of TiO2 and ZnO nanoparticles dispersed in the XM and XH wastewater samples to 10 mg L 1. Based on the time-resolved DLS data in nearly 40 min, TiO2 were observed to remain stable suspensions with aggregates of diameter

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10 min and then the supernatant was discarded. The deposit was transferred to the nylon net, which was cut from a lady stocking and then very tightly stretched on a plexiglass tube with inner diameter of 12 mm and length of 10 mm. The pore size of nylon net was about in the range of 0.5–1 mm. After covering another layer of nylon net, the activated sludge was firmly fixed on the plexiglass tube. And then the tube was placed in a petri dish for NMT measurement. Oxygen flux was measured using a NMT system. The culture medium was the filtrated sludge supernatant supplied with glucose to a final concentration of 2 g L 1. Both ZnO and TiO2 NPs remained stable in the supernatant with the particle size between 200 nm and 400 nm, as similar as their aggregation behavior in the wastewater matrix (Fig. S1). And the medium was aerated to saturation before adding to the petri dish. Firstly, the medium without the metal oxide NPs was added to the dish and the data was collected as the value before exposure for reference. Then, the stock metal oxide NPs was added to the medium and the variation of oxygen flux was recorded immediately. At last, the oxygen flux was monitored again after exposure to NPs lasting 4.5 h. The NPs suspension of 1, 10, and 100 mg L 1 was investigated in this study and the medium without addition of NPs was used as control. All noninvasive measurements were conducted at Xuyue (Beijing) Sci. & Tech. Co., Ltd.

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405 ± 41 nm and 362 ± 21 nm in the XM and XH wastewater samples, respectively. In the meantime, stable suspensions were also observed for ZnO nanoparticles with smaller aggregates of diameter 402 ± 24 nm and 319 ± 24 nm in the XM and XH wastewater samples, respectively. However, the nanoparticle aggregates were larger in the XM wastewater due to the slightly higher IS and lower TOC conditions (Table 1). Whatever, results indicate that in the case of wastewater matrix, the size of the aggregates remained stable at slightly above 300 nm for the two NPs, which accorded with the previous study by Tso et al. (2010) even though they found the stable existed TiO2 and ZnO NP sizes were between 200 nm and 800 nm. Also study reported by Gómez-Rivera et al. (2012) found that CeO2 NP dispersions tend to aggregate in municipal wastewaters. They also claimed that organic and/or inorganic constituents in the real wastewater also contributed to promote the aggregation of the nano-CeO2. The average particle size of CeO2 NPs in the primarily-treated municipal wastewater reached to be in the range of 2500–3000 nm. The divergent results also revealed that the aggregation behaviors of NPs in wastewater matrix are not fully understood. The DLS measurements also revealed that the aggregates particle sizes of ZnO and TiO2 NPs in wastewater matrix increased with their mass concentrations ranging from 1 to 100 mg L 1 (Results not presented) which was also in accordance with the previous study by Tso et al. (2010). Whatever, it can be expected that under these conditions in wastewater matrix, the activated sludge microorganisms would possible be at higher risk due to exposure to the stable dispersions of NPs rather than the quickly aggregated and deposited NPs in some other cases, although the exposure would be in the form of relatively large aggregates (about 300–400 nm), much larger than the vendor reported sizes. At the meantime, the stable aggregates would decrease the removal efficiency of nanomaterials in WWTPs and pose the potential risk to the receiving natural water systems. We also investigate the aggregation profiles of the freshly prepared NPs-wastewater suspension and the suspension allowed to stand for over 4.5 h (Fig. 2). Results showed that no significant changes in the hydrodynamic diameter and propensity to aggregate was observed. It means that the metal oxide NPs were quite stably existed in the wastewater matrix with aggregates of diameter 300–400 nm after 4.5 h or more suspension. In the study of Tso et al. (2010), TiO2 and ZnO NPs were found remaining stable in wastewater with size between 200 and 800 nm even after 20 h of suspension.

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3.3. Factors causing stability of nanoparticles in the wastewater samples Fig. 3 shows a representative aggregation profile of 10 mg L 1 TiO2 and 10 mg L 1 ZnO nanoparticles dispersed in the nonDOMs-contained simulated XM and XH wastewater samples. It can be well seen that the obvious aggregations were observed under the IS and pH value conditions of the simulated wastewater samples. The aggregation rates of TiO2 were found to be 2–3 times quicker than those of ZnO. A comparison between the aggregate sizes in the wastewater and non-DOMs-contained simulated wastewater (Figs. 1–3) indicates that there is very rapid aggregation from <300 nm to over 800 nm in the simulated wastewater within the 40 min of the DLS experiments. Moreover, we also compared the impact of glucose, which is commonly used as the carbon source in the synthetic wastewater treatment (Zhou et al., 2008), on the aggregation process of NPs (Fig. S2). As shown in Fig. S2, glucose at the concentration up to 100 ppm still cannot stabilize the NPs in the synthetic wastewater. This further implies that using the synthetic wastewater to evaluate the ecotoxicity of NPs is unscientific since

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it is unlikely that the organisms will actually be exposed to the NPs in their dispersed sizes in the real wastewater matrix. In synthetic wastewater, NPs tend to form large particles and most of the aggregates will settle out of the suspensions in a few hours, which may reduce their toxicities (Adams et al., 2006; Hsiao and Huang, 2011). Therefore, using synthetic wastewater to assess the NPs toxicity will lead to the facts that the impacts of nanomaterials are overestimated or misjudged in the past studies. Aiming at the real wastewater and the non-DOMs contained wastewater with the IS and pH value close to wastewater, the electrophoretic mobilities (EPM) of the metal oxide nanoparticles were measured at a concentration of 10 mg L 1 (Table S2). In the real wastewater samples, the EPM of the nanoparticles varies from 1.337  10 8 m2 s 1 V 1 to 1.597  10 8 m2 s 1 V 1 accompanying with the f potentials from 17.1 mV to 20.5 mV. The EPM is almost independent of not only the types of metal oxide nanoparticles but also the wastewater samples. However, the EPM in the non-DOMs contained wastewater decreased to a range of 0.708  10 8 m2 s 1 V 1 to 0.821  10 8 m2 s 1 V 1 accompanying with the f potentials from 9.1 mV to 10.5 mV. As mentioned by Keller et al. (2010), the EPM of metal oxide nanoparticles in the natural water is controlled to a large extent by natural organic matters (NOMs), which coat with nanoparticles and change the surface charge on the particles varying from 20 to 30 mV at the pH value ranging from 7 to 9. Through a comparison of TEM images of TiO2 in the nanopure water and wastewater samples (Fig. 4), we confirmed that some transparent gelatinous substances adsorbed onto the particle surfaces and provided a barrier to their further aggregation. Thus, we concluded that the decreased aggregation in wastewater samples was likely related to an increase in the nanoparticles’ surface negative charge, but also to a DOM-based coating of the particles that prevent further aggregation. As a consequence, the nanoparticles stably existed with aggregates of diameter 300–400 nm in wastewater matrix. DOMs is a heterogeneous mixture of aromatic, amino and aliphatic organic compound containing oxygen, nitrogen and sulfur functional groups (Chen et al., 2003), which is composed of complicated components in the municipal wastewater. Fluorescence spectroscopy has commonly been used to discriminate between DOM fractions, including providing information about the structure and functional groups of the components (Chen et al., 2003; Sheng and Yu, 2006; Guo et al., 2010; Osburn et al., 2012; Yu et al., 2013). Therefore, in order to address the issue of DOM components existed in the wastewater matrix, which would consequently cause the stability of nanoparticles in the media, the EEM fluorescence spectroscopy analysis was conducted and presented in Fig. S3. The main fluorescence peaks observed in this study are indicated as peak A and peak B in Fig. S3. The peak A (Ex  275 nm, Em  350 nm) was originated from the protein-like substances, such as tryptophan, tyrosine (Chen et al., 2003). And the peak B (Ex  220 nm, Em  310 nm) were related to xenobiotic-like organic components, such as camphor (Guo et al., 2010).

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The aggregation kinetics of the metal oxide nanoparticles were studied by sequentially adding different concentrations of monovalent, divalent and trivalent electrolytes in forms of NaCl, CaCl2 and Al2(SO4)3 into the wastewater samples containing 10 mg L 1 NPs. Fig. 5 presents a representative aggregation profile of 10 mg L 1 ZnO nanoparticle in the XH wastewater samples with five different CaCl2 concentrations (0 M, 0.005 M, 0.01 M, 0.015 M, 0.025 M). Profiles at addition of monovalent ion are similar to Fig. 5, therefore, not presented in the study.

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A

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ionic strength led to faster aggregation. At higher Al2(SO4)3 concentration (6  10 4 M–1.5  10 3 M), further increase in ionic strength had no effect in enhancing aggregation. TEM imaging of the metal oxide nanoparticles aggregation morphologies in the wastewater matrix revealed that bridging of the metal oxide nanoparticles and aggregates due to the formation of hydrous alumina (Al(OH)3
H2O) in the Al2(SO4)3 system is the likely mechanism for the enhanced aggregation due to the addition of trivalent electrolyte (Fig. 6). These results from this study suggest that various electrolytes play significant and differing roles in the aggregation of metal oxide nanoparticles in wastewater aqueous matrix.

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It can be seen that no obvious aggregation propensity of nanoparticles was observed by adding the monovalent and divalent electrolytes into the wastewater samples even up to the final concentration of 0.025 M of CaCl2. Results are different from the previous studies, where the critical coagulation concentrations (CCC) for NaCl or CaCl2 electrolytes can be found in the natural water bodies or pure water in the lab (Chen and Elimelech, 2006, 2007; Liu et al., 2009, 2010). The enhanced stabilizing propensity might be attributed to the steric barrier imparted by the transparent colloidal substances coverage on the nanoparticles as revealed by the TEM images (Fig. 4), which accorded with the previous results reported by other researches (Chen and Elimelech, 2007; Li and Huang, 2010). Results also suggest that the destabilization of nanoparticles in wastewater is failed by the compression of electrical double layer, therefore, the aggregation kinetics cannot be simply predicted by the traditional DLVO theory. Therefore, it is unscientific that using the present knowledges in the natural water bodies or pure water in the lab to understand and derive the aggregation and transport behavior of nanoparticles in wastewater samples. More explorations should be conducted in the future work. A completely different behavior was observed when the NPs were dispersed with the Al2(SO4)3 solution. Measurement of scattered light intensities over time indicated significant aggregation of the NPs in the solution of Al2(SO4)3 system (Fig. S4). At low Al2(SO4)3 concentration (5  10 5 M–5  10 4 M), an increase in

3.5. Response of activated sludge to NPs acute toxicity In the wastewater matrix, the activated sludge O2 flux changes during exposure to ZnO and TiO2 NPs were investigated. The typical time-resolved oxygen flux profiles of activated sludge after exposure to 1 mg L 1, 10 mg L 1 and 100 mg L 1 TiO2 NPs, respectively, are presented in Fig. 7. The negative value in the oxygen flux indicates the utilization of oxygen due to the physiological activities of activated sludge. Based on the O2 flux profiles, the relative oxygen fluxes of activated sludge after exposure to ZnO and TiO2 NPs compared with the unexposed conditions are summarized in Fig. S5. In generally speaking, the toxicity of NPs is associated with their small size and high specific surface area (Xiong et al., 2011). Although large of studies agreed that the physicochemical characteristics of NPs (e.g., size, shape, surface area, solubility, chemical composition, dispersion factor) play critical roles in determining their toxicities (Adams et al., 2006; Petosa et al., 2010; Hsiao and Huang, 2011), controversies still existed regarding the extents to which the physicochemical properties of NPs influence their toxicity. Therefore, the inconsistency makes for more challenging quantitative comparisons. Several studies have presented that the toxicity to bacterial was increased from TiO2 NPs to ZnO NPs (Adams et al., 2006; Heinlaan et al., 2008; Hsiao and Huang, 2011; Xiong et al., 2011). However, we found that the cytotoxicities of TiO2 and ZnO NPs in terms of the inhibition of oxygen utilization of activated sludge depend on the exposure time. The inhibitory effect of ZnO to activated sludge was prominent as soon as the ZnO NPs suspension was added into activated sludge system and remained relatively stable in the next 4.5 h; however, things were different for TiO2 NPs. As soon as treated with 1 mg L 1, 10 mg L 1 and 100 mg L 1 ZnO NPs, the relative

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Untreated Treated immediately Treated for 4.5 h

C

D

-2

-1

O2 flux [pmol cm s ]

80 40 0 -40 -80 -120 0

1000

2000

3000

t [s] Fig. 7. Typical time-resolved oxygen flux profiles of activated sludge taken from XH WWTP after exposure to TiO2 NPs (A 1 mg L during the NMT analysis (D).

O2 flux of activated sludge responded quickly and dropped to 56%, 56% and 67%, respectively; however the values remained 49%, 50% and 60%, respectively after 4.5 h exposure (Fig. S5). The similar inhibitory degree toward activated sludge was observed by Liu et al. (2011) on the condition of 100 mg L 1 ZnO NPs. Hou et al. (2014) also drew the similar conclusion that 50 mg L 1 of ZnO NPs inhibited the microbial activities in the outer layer (200 lm) of the biofilms. However, a positive relationship between the degree of the inhibitory effect and the inhibitor concentration was not found in our studies. However, TiO2 NPs exhibited negligible toxicities to activated sludge as the O2 flux appeared almost no change at the beginning of treatment with TiO2 NPs. On the contrary, the O2 flux of the

1

; B 10 mg L

1

; C 100 mg L

1

) and the image

relative O2 flux of activated sludge dropped to 30%, 26% and 42%, respectively, after 4.5 h exposure with 1 mg L 1, 10 mg L 1 and 100 mg L 1 TiO2 NPs. It was reported that the toxicity of metal oxide NPs to sludge would be mainly due to the release of metal ion (Hou et al., 2014; Mu and Chen, 2011). In this study, the released Zn2+ and Ti4+ by NPs after exposure for 4.5 h were detected. Released Zn2+ concentration by 1, 10 and 100 mg L 1 ZnO was 0.019 ± 0.002 mg L 1, 0.034 ± 0.006 mg L 1 and 0.122 ± 0.048 mg L 1, respectively. For the activated sludge treated with TiO2 NPs, there was no Ti4+ detected in the supernatant. Mu and Chen (2011) found that the released Zn2+ concentration of 1.2 mg L 1 did not give any significant impact on methane generation. Metal ions released by NPs

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were low; therefore the observed inhibition of the NP aggregates was not likely a result solely of particle dissolution. Instead, the inhibition was presumably due to the generation of reactive oxygen species (ROS) in the presence of NPs (Adams et al., 2006; Zhu et al., 2009). Furthermore, some undetermined mechanisms additional to the ROS production may also be responsible for their toxicity and needed to be revealed in future. We also found that all of the TiO2 NP samples exhibited higher inhibition effect than those caused by the ZnO NP samples after 4.5 h exposure. Moreover, the inhibition of NPs on the aerobic respiration of activated sludge decreased when the NPs concentration is up to 100 mg L 1, which would be attributed to larger aggregates formed in the concentrated NPs concentration (Hsiao and Huang, 2011). In conclusion, compared to the control experiment, the O2 flux decreased greatly after exposure to both TiO2 and ZnO NPs for 4.5 h, which indicated that the NPs inhibited the respiration of activated sludge in the wastewater matrix. A quick inhibitory effect was observed when the activated sludge exposed to ZnO NPs, however, a delay on the inhibition of oxygen uptake rate existed when exposure with TiO2 NPs. Whatever, the adverse impact of NPs on the activated sludge cannot be negligible in the wastewater treatment process. 4. Conclusions In the wastewater matrix, ZnO and TiO2 NPs were found to be quite stable existence with aggregates of diameter 300–400 nm for the investigated time of 4.5 h, which was attributed to the DOMs existed in the wastewater. It implies that the inhibition effect of NP aggregates at the nanoscale to the activated sludge should be taken into consideration in the wastewater treatment. No aggregations of NPs in the wastewater matrix were observed in the increasing presence of monovalent and divalent electrolytes up to 0.15 M in NaCl and 0.025 M in CaCl2, however, obvious aggregation of nanoparticles in the wastewater matrix was observed due to the addition of Al2(SO4)3 with the likely mechanism of bridging of the metal oxide nanoparticles and aggregates due to the formation of hydrous alumina (Al(OH)3
H2O) in the Al2(SO4)3 solution. Based on the results of noninvasive measurement, the O2 fluxes of activated sludge after treated with 1, 10 and 100 mg L 1 NPs were inhibited when the exposure time extended to 4.5 h. The degree of inhibitory effect of TiO2 and ZnO NPs on the oxygen utilization of activated sludge depended on the exposure time. Acknowledgements This study was supported by the China National Natural Science Foundation (51108243), Tsinghua University Initiative Scientific Research Program (20121087922), Program for Changjiang Scholars and Innovative Research Team in University. We also thanked the valuable comments from three anonymous reviewers. Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.chemosphere. 2014.07.037. References Adams, L.K., Lyon, D.Y., Alvarez, P.J.J., 2006. Comparative eco-toxicity of nanoscale TiO2, SiO2, and ZnO water suspensions. Water Res. 40 (19), 3527–3532. Brar, S.K., Verma, M., Tyagi, R.D., Surampalli, R.Y., 2010. Engineered nanoparticles in wastewater and wastewater sludge-evidence and impacts. Waste Manage. 30, 504–520.

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