Size-dependent adsorption of antibiotics onto nanoparticles in a field-scale wastewater treatment plant

Environmental Pollution 248 (2019) 1079e1087

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Size-dependent adsorption of antibiotics onto nanoparticles in a field-scale wastewater treatment plant* Kaifeng Yu, Chi Sun, Bo Zhang*, Muhammad Hassan, Yiliang He School of Environmental Science & Engineering, Shanghai Jiaotong University, 800 Dongchuan Road, Shanghai, 200240, China

a r t i c l e i n f o

a b s t r a c t

Article history: Received 9 October 2018 Received in revised form 22 February 2019 Accepted 25 February 2019 Available online 28 February 2019

This work present aims to evaluate the effect of a conventional wastewater treatment process on the number of nanoparticles, and the role of nanoparticles as a carrier of antibiotics. A set of methods based on asymmetrical flow field flow fractionation coupled with multi-angle light scattering to separate and quantify nanoparticles in real wastewater was established. The characterization of nanoparticles was conducted by transmission electron microscopy, energy dispersive spectrometer, UVevisible spectrophotometer and three-dimensional excitation-emission matrix fluorescence spectroscopy. The adsorption of different sizes of nanoparticles separated from the real wastewater for four targeted antibiotics (sulfadiazine, ofloxacin, tylosin and tetracycline) was studied. The results show that the number of nanoparticles were increased in the wastewater treatment process and the size range between 60 and 80 nm was predominant in wastewater samples. The nanoparticles were mainly composed of O, Si, Al and Ca elements and organic components were in the size range of 0e10 nm. Targeted antibiotics were dominantly adsorbed onto nanoparticles with 60e80 nm size range at each stage. The concentrations of tetracycline adsorbed on nanoparticles were surprisingly increased in the end of the treatment process, while ofloxacin and tylosin had the completely opposite phenomenon to tetracycline. The pH and ionic strength definitely affected the aggregation of nanoparticles and interaction with the antibiotics. It is of great significance to give insights into nanoparticle-antibiotic assemblages for the effective treatment and avoiding the water risks due to nanoparticles’ ubiquitous and their risks of carrying antibiotics. © 2019 Elsevier Ltd. All rights reserved.

Keywords: Antibiotics Nanoparticles Adsorption DOM Wastewater

1. Introduction Antibiotics were frequently detected in wastewater treatment plants (WWTPs) (Perez et al., 2005), sludge (Marx et al., 2015), surface water (Li et al., 2018b), underground water (Szekeres et al., 2018), drinking water (Li et al., 2018a) and sediments (Zhou et al., 2011) all over the world, resulting from excessive usage to prevent diseases in livestock, agriculture and industry. WWTPs are considered as crucial routes for antibiotics into other environmental matrices due to relatively low removal efficacy of prevalent traditional wastewater treatment processes for many antibiotics (Le-Minh et al., 2010) and high antibiotic concentrations in the effluents. Moreover, WWTPs are the hotspot reservoirs of antibiotic resistant bacteria (ARB) and antibiotic resistant genes (ARGs) induced by antibiotics or antivirus in the wastewater, which

*

This paper has been recommended for acceptance by Bernd Nowack. * Corresponding author. E-mail address: [email protected] (B. Zhang).

https://doi.org/10.1016/j.envpol.2019.02.090 0269-7491/© 2019 Elsevier Ltd. All rights reserved.

probably pose potential risk to the environment and human health (Le Page et al., 2017). Nanoparticles (NPs), defined as ubiquitous type of matters with the particle size ranging from 1 to 100 nm (nm) (EU, 2011; Sharma et al., 2015), exhibit distinctive physicochemical properties related to their large specific surface area, high reactivity, crystalline structure (Ribeiro et al., 2017) and quantum size effect (Sharma et al., 2015). Therefore, NPs possess the ability to combine with organic pollutants, heavy metals and alter the fate, transportation and toxicity of carrying pollutants in the aquatic environment (Baun et al., 2008). There are many reports about the complex interactions of engineered nanoparticles (ENPs) with the tracing organic contaminants in the aquatic environment. Engineered carbon nanotubes (CNTs) could adsorb lincomycine (LCN), sulfamethoxazole (SMX) (Chen et al., 2017a; Yu et al., 2015), and iopromide (Kim et al., 2014). However, the constituents of NPs in WWTPs which are reservoirs of natural nanoparticles (NNPs) spontaneously produced by microorganisms (Borghese et al., 2017) and possess the properties of co-existence of NNPs and ENPs, and

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also the interactions with antibiotics are more complicated and largely unknown. The separation of NPs with different sizes is the basis of exploring the interaction of NPs in the environment with other pollutants. The filtration and centrifugation can usually be used to separate the NPs of various sizes, but fail in providing the information on size-distribution of aggregates (Isaacson & Bouchard, 2010). Following the separation, several detection methods, such as dynamic light scattering (DLS), electron microscopy image analysis and photon correlation spectroscopy (Calzolai et al., 2012; Wang, 2005) were used to obtain the size distribution. Asymmetrical flow field flow fractionation coupled with multi-angle light scattering (AF4-MALS) is a powerful device to combine the particle size measurement with the effective separation of nanoparticles owing to the interaction of the analytes in the field in a non-uniform flow (Gigault et al., 2013; Schmidt et al., 2011). AF4MALS can separate particles in liquid ranging from a few nanometers to several tens of microns. In contrast to transmission electron microscopy (TEM), AF4-MALS does not cause significant particle aggregation when measuring the particle size distribution. AF4-MALS can also obtain the intensity of small particles accurately, while DLS usually shields the small particles and overestimates the average size of particles. Flow-based field flow fractionation techniques do not need a stationary phase, to which NPs may irreversibly bind, and are selective in the range of 2e800 nm, thus they are ideal for separating primary NPs from aggregates. The use of light scattering and ultravioletevisible (UVevis) detectors provide information on particle sizes and aggregation states. It has been widely applied to measure the size distribution of colloids, nanoparticles and macromolecules in the environmental samples (Dubascoux et al., 2008a; Sadik et al., 2014). Analyzing NPs in real water systems in terms of their sizes, compositions and interaction with antibiotics is not a simple task due to their low concentration and fractionation with different sizes (Li et al., 2012). Our work established a set of method to first give profiles of the size distributions and elemental compositions of aquatic NPs in the effluent from a WWTP by using AF4 coupled with on-line UVevis and MALS, and the interactions between typical selected antibiotics and separated NPs of different size ranges in real wastewater. This work aims to give an insight into size-dependent adsorption of antibiotics onto NPs in real wastewater and provide a reference for the corresponding treatment measures.

2. Materials and methods 2.1. Sample collection The water samples were collected from initial influent (INF), primary sedimentation tank effluent (PSTE), secondary sedimentation tank effluent (SSTE) and final effluent (FE) of the WWTP located in Shanghai, China. An anaerobic-anoxic-oxic (A2/O) process was applied in the plant to treat about 50,000 tons wastewater per day, composed of 70% domestic sewage and 30% industrial wastewater from its 15.5 km2 service areas. At each sampling site, 4 samples (2 L per time) were obtained with a time-interval of 2 h and mixed together when sampling was finished. The parameters of pH and oxidation-reduction potential (ORP) were immediately detected by pH/ORP rapid measurement probe (SX721, Sassin, Shanghai, China), and dissolved oxygen (DO) by portable DO analyzer (HQ30d, HACH, USA). All samples were transported to the laboratory as soon as possible and stored in refrigerator at 4
C before analysis.

2.2. Separation and quantification of NPs 2.2.1. Pretreatment of water samples The raw water samples were filtered through a mixed cellulose membrane (KMCE04570100, Millipore, USA) with a pore size of 0.45 mm. The filtrates were concentrated by using an ultrafiltration stir cell (8400, Millipore, USA) under the pressure of 0.15 MPa and agitation rate at 100 rpm. The regenerated cellulose membrane with a molecular weight of 10 kDa (Wyatt Technology Corporation, USA) was used. Concentration factor (N) was calculated by the following equation:

N¼

VS VR

(1)

Where VS is the volume of raw wastewater sample; VR is the remained volume in the stir cell. 2.2.2. Separation of NPs by AF4-MALS An analytical platform coupling AF4 (Wyatt Technology Corporation, Santa Barbara, CA, USA) with MALS (DAWN EOS, Wyatt Technology Corporation, Santa Barbara, USA) was established and used to determine the size distribution of NPs in the water samples. An Agilent 1260 series manual sampler with an injection loop (1 mL) and a high performance liquid chromatography (HPLC) pump (G1311A; Agilent Technologies, Santa Clara, CA, USA) were used to inject working suspensions and to deliver the carrier flow (deionized water), respectively. The trapezoidal AF4 channel was 27.5 cm long from tip to tip with tapered inlets and outlets. The 10 mM NaCl solution (Super pure, Merck), filtered through 0.1 mm mixed cellulose membranes before usage, was used as the mobile phase. The separation parameters, including the horizontal flow rate, cross flow rate, elution time, and size of separation membrane of AF4, were optimized. After optimization, injection volume was set at 100 mL, horizontal flow rate at 1 mL min1, cross flow rate at 2 mL min1, elution time at 40 min, and separation membrane was selected as 10 kDa regenerated cellulose membrane. Under the optimal conditions, the original geometric diameter as a function of time can be obtained. The samples were firstly separated by AF4 and then entered into the online test module consisted of the HPLC equipped with a UV detector and MALS. The detection data were collected and processed by ASTRA 6.1 software (Wyatt Technology Corporation, USA). The dispersed solution of the separated NPs was collected by a fraction collector. The elution time of NPs with different particle sizes can be obtained by the ASTRA 6.1 software, and the fraction collector can be controlled to collect fractions with different elution times so as to obtain nanoparticles with different sizes. The fractions are collected in the form of droplets and the syringe needle of the collector will always be at the surface of the liquid to avoid loss of volatilization due to liquid coming into contact with the air during dripping. The fractions collected in this study ranged from 20 to 40 nm, 40e60 nm, 60e80 nm and 80e100 nm. It should be noted that the particles below 20 nm could not be quantified due to the detection limits of MALS. 2.2.3. Quantification of NPs If the index of refraction of the sample is known, it is possible to use ASTRA's proprietary number density calculations. The geometric radius and number density of a sample are determined for each eluting slice using just a DAWN. It is possible to measure the total number of particles in a peak. The number density of nanoparticles with single size can be calculated automatically by ASTRA 6.1 software and exported to Excel. Therefore, the total number density of targeted size ranges was summation of that single size in

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the sections according to the follow equation:

Pb Qab ¼

a Qs VN

(2)

Where, Qa-b is the number density of NPs in raw wastewater before concentrated (particles mL1), Qs is the number density of the single size NPs obtained by ASTRA 6.1 and V is the volume of concentrated wastewater sample. The N is calculated by equation (1). 2.3. Characterization of NPs 2.3.1. EDS and TEM The elemental composition of NPs in the wastewater was analyzed by the energy dispersive spectrometer (EDS) system of TEM (JEM-2010HT, JEOL Electronics Co., Ltd. Japan). The morphology of final effluent was examined on a TEM. For sample preparation, one drop of concentrated suspension was placed on a carbon-coated copper grid, followed by ambient drying for approximately one day before TEM analysis. 2.3.2. UVevis analysis and 3D-EEM UVevis analysis was carried out by online UV analysis module of AF4 separation system with detection wavelength set at 254 nm. The abscissa tR is converted to Rh by using the ISIS software to obtain the spectrum with the abscissa of Rh and the ordinate as the UVevis response signal. Three-dimensional emission-excitation matrixes (3D-EEM) analysis was operated to investigate the characteristics and distribution of the fluorescence components of the NPs in each treatment stage by using three-dimensional fluorescence analyzer (F-7000, HITACHI, Japan). Since the NP concentrations were low after AF4 separation, the Rayleigh scattering and the Raman scattering intensities were much higher than the fluorescence intensity of the particles. The fluorescence spectrum of the target substance would be greatly disturbed. Therefore, Rayleigh scattering and Raman scattering were needed to be deducted, and the interpolation method was used to complement the deducted data (Chen et al., 2017b). 2.3.3. Zeta potential and particle size measurements The zeta potentials of wastewater samples, as well as their particle sizes, were obtained by using a DLS Delsa7M Nano C Particle analyzer (Bechman, USA). 2.4. Antibiotic analysis Four targeted antibiotics (sulfadiazine (SDZ), ofloxacin (OFC), tylosin (TYL) and tetracycline (TC)) were selected out to study their adsorption behavior on the surface of NPs with different size ranges. Due to the unique separation principle of AF4, NPs in the concentrated wastewater would be separated from the aqueous phase, dispersed in the carrier liquid, and collected as fractions (Dubascoux et al., 2008b; Isaacson & Bouchard, 2010; Kammer et al., 2011). The AF4 separation mechanism is illustrated in Fig. S1. The detected antibiotics in the separated fractions would be the antibiotics adsorbed on the particles. Antibiotics were measured by a TSQ Quantum triple quadrupole LC - MS/MS instrument (ESI source, Thermo Fisher Scientific, USA) equipped with a column (2.1  50 mm, 1.8 mm) (Agilent, USA). The HPLC conditions were as follows: phase A: 0.1% formic acid solution, phase B: acetonitrile, flow: was 0.3 mL min1; column temperature: 35
C and injection volume: 10 mL. The elution procedure was: 0e2 min for 95% A, 2e8 min for 95 - 60% A, 8e10 min for 60% A and 10e12 min for 95% A. The mass spectrometry conditions were as

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follows: the ionization source was electrospray ion source, positive ion mode with a spray voltage of 3500 V, a sheath gas pressure of 30 kPa, an auxiliary gas pressure of 10 kPa and an ion source temperature of 350
C (Chen et al., 2019; Yue et al., 2018). The precursor ions, daughter ions and collision energy of the four antibiotics can be seen in Table S1 in the supplementary material. 2.5. Influence factors of pH and ionic strength for TC adsorption on NPs Due to its special physicochemical properties (three pKa values: 3.3, 7.7, 9.7) (Anton-Herrero et al., 2018; Figueroa et al., 2004; Yang et al., 2011) and relatively high concentration in effluents of WWTPs, TC was typically selected to explore the influence of pH and ionic strength (IS) on the size-dependent adsorption of antibiotics onto NPs in the laboratory. With a confirmed IS of 0.01 M, pH values of separated FE samples were adjusted to 4, 5, 6, 7, and 8 using hydrochloric acid solution (0.1 M) and liquor natrii hydroxide (0.1 M). Then, standard TC solution was spiked to make the TC concentration of separated samples 10 mg L1 (Table S2). Similar to pH influence experiments, the IS of separated FE samples were adjusted to 0.001, 0.01 and 0.1 M by using sodium chloride solids, respectively. The concentration of TC in each sample was adjusted to 10 mg L1. Liquids were settled for two days and TC concentrations of the samples were measured by LC - MS/MS. 3. Results and discussion 3.1. Separation and quantification of NPs in sewage The separation results of NPs at different wastewater treatment stages are shown in Fig. S2. In the elution time ranging from 14 to 24 min, samples from the INF and PSTE showed larger elution peaks, mainly because the number distribution of particles with smaller particle size level was more irregular; while the SSTE and FE exhibited relatively smooth elution peaks. After the biodegradation process, the number distribution of small sized particles was stabilized, resulting in a smoother peak shape. The number densities of NPs with 5 size ranges are shown in Fig. 1 (a, b) and Table S3. The concentration of NPs reached the highest level (1.76 ± 0.06  107particles mL1) in SSTE, while it was least 1.07 ± 0.06  107 particles mL1) in the INF. It is obviously observed that the total number of NPs were increased in the WWTP, especially after the biological reactor, indicating that the NPs conceivably be produced and escaped from the traditional biological treatment systems. The metabolic activity of bacteria plays a crucial role (Fig. S3). Abundant microorganisms in the wastewater generate NPs through two processes called biologically induced mineralization (BIM) and biologically controlled mineralization (BCM) (Sharma et al., 2015). Solid materials attached to microorganisms were degraded to NPs in the BIM process. NPs generated by the BCM process were generally mineral particles with particular functions for the organisms (Sharma et al., 2015). It is evident that biological wastewater treatment system (A2/O) with plentiful activated sludge suspensions might produce more NPs. Besides, in secondary sedimentation tank (SST), the precipitation behavior of particles can be described by the Stokes settling formula, i.e. the vertical sedimentation velocity of particles is proportional to the square of their diameters. The smaller size of particles are more difficult to precipitate. Particle size is one of the most important properties that affects the mobility of NPs in different environmental matrices (Tosco & Sethi, 2018; Tufenkji & Elimelech, 2004). When the wastewater flowed through the primary sedimentation tank (PST), nanoparticle number densities were also increased

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Fig. 1. Number density (a and b) of size-dependent NPs at different stages in the WWTP, concentrations of four targeted antibiotics in the INF and FE (c) and antibiotics adsorbed by size-dependent NPs in the four sampling sites (d).

which could be ascribed to the hydraulic conditions that affect the formation and size of particles. Water flow in the WWTP was generally turbulent due to the relative higher flow velocity and aeration (in aerobic tank), and thus the disturbance between adjacent flow layers might accelerate the dispersing process of large particles into smaller ones. The particle sizes within 40e60 nm and 60e80 nm ranges were predominant in all the wastewater samples, accounting for 58.92% and 32.09% in the INF, 59.26% and 30.97% in the PSTE, 24.37% and 72.14% in the SSTE, 39.56% and 45.75% in the FE, respectively. In contrast, the particle size less than 40 nm was as low as 7.82%e 8.25% in the front of WWTP, reaching the lowest at 2.65% in the SSTE and adversely increasing to 13.92% in the FE. The particle size range between 80 and 100 nm had the least percentage for less than 1.52%. It seems that NPs among 40e80 nm size were more stable than that of other size ranges. According to the DLVO theory, when the two charged colloidal particles are close to each other, the electrostatic repulsion occurs. In addition, Van der Waals’ forces exist between the colloidal particles that are close to each other. Therefore, whether the colloidal particles can be aggregated depends on the resultant force between the two (Phenrat et al., 2007). NPs also experience dynamic movements of dispersion, aggregation and re-dispersion or aggregation, dispersion, and reaggregation process. Dynamic equilibrium could easily be achieved in middle sized NPs, resulting in more NPs with 40e80 nm size range.

3.2. Characterization of NPs in sewage 3.2.1. TEM and element analysis The small NPs with 10 nm and 20 nm size in the FE were sphereshaped (Fig. 2 (a) and (b)). However, NPs with larger sizes of 50 nm and 100 nm exhibited irregular and inhomogeneous forms (Fig. 2 (c) and (d)), probably due to aggregation of smaller particles (Hu et al., 2018) and adsorption of various materials. Compared to small size NPs, there were more particles in the TEM images of 50 nm and 100 nm size NPs, following larger specific surface areas. Element analysis was conducted in order to identify the components of NPs in wastewater. Four elements of O, Si, Ca and Al were dominant in all the wastewater samples (Table S4). The O and Si elements were over 40% and 30% in the INF and FE, respectively. It could be speculated that NPs were mainly consisted of SiO2, not only because silicates present high background concentrations in the environment, but also silica NPs have been widely synthetized and used in industrial and commercial products with end up in WWTPs (Grass et al., 2014; Otero-Gonzalez et al., 2015). Silica NPs were stable in wastewater and could be transported through the effluent (Jarvie et al., 2009). The SiO2 colloids were negatively charged (Liu et al., 2013) with huge specific surface area and hydroxylated surface which enhance the adsorption of cationic and polar materials onto it. Al was main element in the INF, PSTE and FE, suggesting that aluminum oxide (Al2O3) NPs could also be the important component in wastewater from its extensively

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Fig. 2. TEM images of nanoparticles with different size ((a) 10 nm; (b) 20 nm; (c) 50 nm; (d) 100 nm) in the final effluent of WWTP.

production and application. Some research findings indicated that Al2O3 NPs were mainly removed by adsorption onto activated sludge (Chen et al., 2012), in correspondent with the atom percentage largely decreased in the SSTE. Besides, Al and Si can be composed of aluminosilicate in water, which also has a strong adsorption capacity. The percentage of Cl was much higher in FE than that in SSTE, resulting from the disinfection process. 3.2.2. UVevis analysis As shown in Fig. 3, the maximum UV254 absorbance was in the Rh range of 0e10 nm, which indicates that dissolved organic matters (DOM) or the organic matters with aromatic ring were the main constituents of NPs with the size range of 0e10 nm (Worms et al., 2010). 3.2.3. EEM analysis The EEM characterizations of NPs at different wastewater treatment stages are shown in Fig. 4. The EEM peak values of NPs were at 280e360 nm/200e230 nm ranges and 280e330 nm/ 250e280 nm ranges, in correspondent with aromatic proteins and microbial by-products, respectively. The aromatic proteins in some domestic waste residues would enter into the WWTP, and some of them would exist in the form of NPs. After the biological treatment, the EEM peak colors were darker (Fig. 4 (c) and (d)) than that forward (Fig. 4 (a) and (b)), which indicated that microbial metabolites could be produced in the process of biodegradation of organic matters. Thus, DOM became more aromatic in A2O processes (Maizel & Remucal, 2017; Wang & Chen, 2018). Despite lack of

Fig. 3. Size distributions of NPs in wastewater samples measured by UVevis.

quantitative analysis, qualitative estimate of treatment efficacy could be obtained by comparing decolorizing effect in the fluorescence spectra (Sgroi et al., 2017). The EEM peaks and colors were slightly changed, suggesting that DOM coated on NPs was almost unremoved.

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Fig. 4. Three-dimensional EEM fluorescence spectra of NPs along the wastewater treatment process: (a) INF; (b) PSTE; (c) SSTE; (d) FE. Up right is the color band indicator.

3.3. Antibiotics adsorption on NPs 3.3.1. Occurrence and distribution of antibiotics in the WWTP Fig. 1 (c) and Table S4 show the concentrations of selected antibiotics in the INF and FE. The concentrations of SDZ and OFC were the highest, reaching at 1460 ± 40 ng L1 and 1370 ± 140 ng L1 in the INF, whereas 810 ± 70 ng L1 and 650 ± 130 ng L1 in the FE, respectively. The residue concentration of TYL was the lowest in INF and FE with concentration of 212 ± 61 ng L1 and 45 ± 16 ng L1, reaching at 78.8% removal efficiency. While the removal efficiencies of other selected antibiotics were relative low, reaching at 44.5% (SDZ), 53.8% (TC), and 52.6% (OFC), respectively. OFC were significantly removed by adsorption affected by electrostatic repulsion (Li & Zhang, 2010). Considering the HRT, the function of biodegradation was not a major path to remove OFC (Li & Zhang, 2010). TC were mainly adsorbed by the sludge (Gao et al., 2012) with no biodegradation (Li & Zhang, 2010). To some extent, abiotic degradation of TC (photo-degradation and hydrolysis) also matters (Sarmah et al., 2006). An octanol/water partition coefficients (Kow) governs the partitioning mechanism of the adsorption of hydrophobic organic compounds on particles (soils) (Chang et al., 2009). The high removal efficacy of TYL with relative high log Kow (Table S7) was probably ascribed to sludge adsorption. Removal of SDZ by activated sludge adsorption was negligible because SDZ is refractory and water-soluble (Li & Zhang, 2010). Furthermore, biodegradation of SDZ was also negligible in the WWTP. The biodegradation of SDZ, starting time as long as 6e12 d, was completed within 2e4 d at 20
C in some simulating activated

sludge systems (Ingerslev & Halling-Sorensen, 2000; Perez et al., 2005), which was far beyond the HRT (less than 1 d) in the real WWTPs. Therefore, the removal efficiency of SDZ was the lowest among the four targeted antibiotics. NPs, abundantly existed in the wastewater and sludge, however, probably are overlooked in the role of coaction and transportation of antibiotics in the WWTP. 3.3.2. Antibiotics adsorbed on size-dependent NPs The concentrations of four targeted antibiotics adsorbed onto particles within 60e80 nm size range were significantly higher than that of other particle size ranges (Fig. 1 (d) and Table S3). NPs of 60e80 nm size range were the predominant sorbents for TC, TYL and OFC accounting for 31.2e47.7%, 34.7e72% and 46.3e72% in the four sampling sites, respectively. While other size ranges homogeneously contributed less than 30%. The adsorption capacity is usually positively related to the total adsorption sites provided by the NPs (Peiris et al., 2017). The more NPs, the more adsorption sites. The particle numbers of NPs with 40e60 nm and 60e80 nm size ranges totally accounted for more than 90% (Table S3). Besides, larger size NPs had higher specific surface areas. Consequently, the adsorption of four antibiotics on NPs within 40e80 nm size range accounted for more than 60%. In the INF, concentrations of TC, TYL, and OFC adsorbed onto NPs within 20e100 nm size range accounted for 1.75%, 19.02% and 4.89% of the total TC, TYL and OFC concentrations in aqueous phase, respectively (Table S3 and S4). While TC, TYL, and OFC accounted for 28.35%, 6.91% and 0.48% in the FE, respectively. The concentrations of TC adsorbed on NPs were surprisingly increased in the

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end of treatment process, especially in the SSTE. Interestingly TYL and OFC had the completely opposite phenomenon to TC. The distinct log Kow (Table S7) suggests that the TCs are more hydrophilic, while TYL and OFC are more hydrophobic. But recent studies suggested that TC could be absorbed by montmorillonite and bentonite (Parolo et al., 2008). In the biological reactor, huge amount of activated sludge not only could adsorb NPs coated with antibiotics, but also generate and desorb NPs with components changed (Table S4). Thus, the interaction between activated sludge and NPs, and coaction between NPs and antibiotics with different chemical properties, could contribute different fates of antibiotics in the wastewater. Furthermore, the total NP numbers in SSTE were the largest, which is probably also related to the highest adsorption of TC. In contrast, SDZ was the least one (<1 ng L1) adsorbed by NPs because of its high hydrophilic property (Li & Zhang, 2010) and negative distribution on its surface (Jia et al., 2016). Still, NPs within 60e80 nm were the dominant particles for SDZ adsorption.

3.4. Factor influencing antibiotic adsorption on size-dependent NPs 3.4.1. Influence of DOM DOM interaction with NPs is likely to happen and will significantly influence aggregation and stability of NPs (Chen & Elimelech, 2007; Domingos et al., 2009; Keller et al., 2010). The coaction

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mechanisms of antibiotics with DOM could be interpreted by hydrophobic effect, complexation, hydrogen bonds, pep stacking and charge attraction (Chen et al., 2015; Jia et al., 2016; Qin et al., 2018). However, little attention was attached to DOM-NPs-antibiotics assemblages. NPs in different size range probably act as the important role of cores or carriers of various materials under different conditions. With its colloidal properties, DOM conceivably occupies the surface sites of NPs, inhibiting sorption of trace organic pollutants. Adsorption of antibiotics to NPs might be concentrationdependent, as well as DOM (Hyung et al., 2007; Qin et al., 2018). Different wastewater and treatment processes may contribute different DOM and NPs with various physicochemical properties influencing DOM-NPs-antibiotics assemblages, which needs further in-depth research.

3.4.2. Influence of pH When the pH of concentrated final effluent fractions was adjusted to 5, the total concentration of TC adsorbed on NPs increased to 67.10 ng L1 (Fig. 5 (b)). As pH increased from 5 to 8 with a settled IS of 102 M, the sorption capacity of NPs decreased to 46.38 ng L1. TC possesses the chemical properties with three pKa values, which elicits different speciation at different pH values, such as the cation, zwitterion, monoanion and dianion forms (Table S5) (Li & Zhang, 2010). The surface charge of NPs is generally

Fig. 5. Number density of NPs in the FE and their adsorption for spiked TC influenced by different pH ((a) and (b)) and IS ((c) and (d)).

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negative (Jia et al., 2016). At pH value of 5, TC was dominant by cation form, resulting in adsorption onto negatively charged surface sites of NPs. TC molecule becomes neutral or negatively charged with increasing the pH in the solution (Zhao et al., 2011). Therefore, electrostatic repulsion between TC and NPs resulted in less adsorption capacity. TC was predominantly adsorbed on NPs with 60e80 nm size range although pH varied. As pH was adjusted from 4 to 5, the adsorption capacity of TC on different size ranges increased. When pH was changed from 5 to 8, TC concentration adsorbed onto NPs within 60e80 nm size ranges significantly decreased. DLS size trend was coincident with TC concentrations, indicating that NPs aggregation and dispersion were significantly affected by pH. 3.4.3. Influence of ionic strength As ionic concentration increased from 103 M to 102 M, sorption of TC onto size-dependent NPs decreased. This was because competition sorption between TC and ion on NPs was becoming intense (Parolo et al., 2008). However, TC concentration was slightly decreased when IS changed from 102 M to 101 M. Spiked TC concentrations (10 mg L1) were much higher than that of ionic solutions, so the surface sites of NPs had stronger affinity of TC than Naþ (Parolo et al., 2008). From the viewpoint of particle size distribution, TC was predominantly adsorbed onto NPs within size range of 60e80 nm at different IS. The particle numbers of 60e80 nm and 80e100 nm size ranges slightly augmented along with increased IS, while that of 20e40 nm and 40e60 nm size ranges were largely decreased. A clear aggregation of NPs was observed at higher IS, which is in agreement with other studies (Yang et al., 2013). Powerful IS will decrease the electrostatic repulsive forces between NPs at any given pH, resulting in decreased dispersion of NPs (Domingos et al., 2009; Yang et al., 2013). Effects of divalent cations, such as Ca2þ, Mg2þ etc., which are differing from monovalent cations, need further investigation. 3.5. Environmental implications The conventional wastewater treatment processes should improve the ability to deal with fine particles (especially nanoparticles), which is the significant environmental implications obtained from the research findings. More attention should be paid to tremendous NPs existed in the WWTP because of their potential risk, increased number and ability to escape the traditional processes. Traditional wastewater treatment technologies could effectively remove suspended particles by adsorption on activated sludge, sedimentation and membrane filtration. However, NPs were not easily removed by traditional methods related to their ubiquity, stability and internal sources from activated sludge (microorganisms), which is the cornerstone of traditional biological processes. NPs with certain size ranges tend to adsorb some chemicals (antibiotics, heavy metals etc.), enhancing their mobility in environmental matrices (e.g. river water, reservoir, sediment, soil, plants). Those antibiotics adsorbed on NPs, out of capture by WWTPs, will eventually be discharged into natural water, causing potential hazards to the environment and human health. Toxic risk will become more serious in the presence of NPs through the release of toxic chemicals in wastewater (Nel et al., 2006). It is urgent to take measures to control antibiotics and that bound onto NPs from WWTPs. 4. Conclusions The low removal efficacy and high antibiotic concentrations in

effluent of traditional WWTPs remain serious challenges. However, our results elicit that NPs, due to their ubiquitous and physicochemical properties, act vital roles in the fate and transport of antibiotics in the wastewater, increasing the escape ability of antibiotics to aquatic environment. The interaction between NPs and antibiotics could also be obviously affected by environmental factors. The nanoparticle-antibiotic assemblages need particular concern to identify the interaction mechanisms, assess their environment risk and develop advanced controllable measures. Acknowledgement This work was financially supported by National Natural Science Foundation of China (No. 21677097), National Science and Technology Major Projects of Water Pollution Control and Management of China (2017ZX07207002-05), and Key Laboratory of Jiangxi Province for Persistent Pollutants Control and Resources Recylce (Nanchang Hangkong Univesrity) (No. ST201522002). Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.envpol.2019.02.090. References Anton-Herrero, R., Garcia-Delgado, C., Alonso-Izquierdo, M., Garcia-Rodriguez, G., Cuevas, J., Eymar, E., 2018. Comparative adsorption of tetracyclines on biochars and stevensite: looking for for the most effective adsorbent. Appl. Clay Sci. 160, 162e172. https://doi.org/10.1016/j.clay.2017.12.023. Baun, A., Sorensen, S.N., Rasmussen, R.F., Hartmann, N.B., Koch, C.B., 2008. Toxicity and bioaccumulation of xenobiotic organic compounds in the presence of aqueous suspensions of aggregates of nano-C(60). Aquat. Toxicol. 86 (3), 379e387. https://doi.org/10.1016/j.aquatox.2007.11.019. Borghese, R., Brucale, M., Fortunato, G., Lanzi, M., Mezzi, A., Valle, F., Cavallini, M., Zannoni, D., 2017. Extracellular production of tellurium nanoparticles by the photosynthetic bacterium Rhodobacter capsulatus. J. Hazard Mater. 324 (Pt A), 31e38. https://doi.org/10.1016/j.jhazmat.2016.11.002. Calzolai, L., Gilliland, D., Rossi, F., 2012. Measuring nanoparticles size distribution in food and consumer products: a review. Food Addit. Contam. Part A Chem Anal Control Expo Risk Assess 29 (8), 1183e1193. https://doi.org/10.1080/19440049. 2012.689777. Chang, P.-H., Li, Z., Jiang, W.-T., Jean, J.-S., 2009. Adsorption and intercalation of tetracycline by swelling clay minerals. Appl. Clay Sci. 46 (1), 27e36. https://doi. org/10.1016/j.clay.2009.07.002. Chen, B., Sun, W.L., Wang, C.H., Guo, X.Y., 2017a. Size-dependent impact of inorganic nanoparticles on sulfamethoxazole adsorption by carbon nanotubes. Chem. Eng. J. 316, 160e170. https://doi.org/10.1016/j.cej.2017.01.087. Chen, K.L., Elimelech, M., 2007. Influence of humic acid on the aggregation kinetics of fullerene (C60) nanoparticles in monovalent and divalent electrolyte solutions. J. Colloid Interface Sci. 309 (1), 126e134. https://doi.org/10.1016/j.jcis. 2007.01.074. Chen, Y., Su, J.Q., Zhang, J., Li, P., Chen, H., Zhang, B., Gin, K.Y., He, Y., 2019. Highthroughput profiling of antibiotic resistance gene dynamic in a drinking water river-reservoir system. Water Res. 149, 179e189. https://doi.org/10.1016/j. watres.2018.11.007. Chen, Y., Su, Y., Zheng, X., Chen, H., Yang, H., 2012. Alumina nanoparticles-induced effects on wastewater nitrogen and phosphorus removal after short-term and long-term exposure. Water Res. 46 (14), 4379e4386. https://doi.org/10.1016/j. watres.2012.05.042. Chen, Y., Yu, K., Zhou, Y., Ren, L., Kirumba, G., Zhang, B., He, Y., 2017b. Characterizing spatiotemporal variations of chromophoric dissolved organic matter in headwater catchment of a key drinking water source in China. Environ. Sci. Pollut. Res. Int. 24 (36), 27799e27812. https://doi.org/10.1007/s11356-017-0307-5. Chen, Z., Zhang, Y., Gao, Y., Boyd, S.A., Zhu, D., Li, H., 2015. Influence of dissolved organic matter on tetracycline bioavailability to an antibiotic-resistant bacterium. Environ. Sci. Technol. 49 (18), 10903e10910. https://doi.org/10.1021/acs. est.5b02158. Domingos, R.F., Tufenkji, N., Wilkinson, K.I., 2009. Aggregation of titanium dioxide nanoparticles: role of a fulvic acid. Environ. Sci. Technol. 43 (5), 1282e1286. Dubascoux, S., Heroult, J., Hecho, I.L., Potin-Gautier, M., Lespes, G., 2008a. Evaluation of a combined fractionation and speciation approach for study of size-based distribution of organotin species on environmental colloids. Anal. Bioanal. Chem. 390, 1805e1813. cho, I., Gautier, M.P., Lespes, G., 2008b. Dubascoux, S., Von Der Kammer, F., Le He Optimisation of asymmetrical flow field flow fractionation for environmental nanoparticles separation. J. Chromatogr. 1206 (2), 160e165. https://doi.org/10.

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