Effects of humic acids on the aggregation and sorption of nano-TiO2

Chemosphere 119 (2015) 171–176

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Effects of humic acids on the aggregation and sorption of nano-TiO2 Yanjie Li a, Chen Yang a,b,⇑, Xuetao Guo a, Zhi Dang a,c, Xiaoqin Li c, Qian Zhang d a

College of Environment and Energy, South China University of Technology, Guangzhou, China Guangdong Provincial Key Laboratory of Atmospheric Environment and Pollution Control, South China University of Technology, Guangzhou, China c The Key Lab of Pollution Control and Ecosystem Restoration in Industry Clusters, Ministry of Education, Guangzhou, China d School of Life and Environmental Science, Guilin University of Electronic Technology, Guilin, China b

h i g h l i g h t s  Aromatic-rich HAs stabilized the nano-TiO2 particles better than aliphatic-rich HA.  Phenolic groups of HAs generated higher charge density on the nano-TiO2 surfaces.  Aromatic-rich HA-TiO2 complexes had higher sorption and nonlinearity.

a r t i c l e

i n f o

Article history: Received 22 January 2014 Received in revised form 29 April 2014 Accepted 2 May 2014 Available online 30 June 2014 Handling Editor: J. de Boer Keywords: Nanoparticles Humic acids Aggregation Hydrophobic organic compounds Sorption

a b s t r a c t In this study, humic acids (HAs) from three sources, peat, sediment and straw, used to coat nano-TiO2 were investigated. The results indicated that HAs isolated from peat were aromatic-rich, whereas those isolated from sediment and straw were aliphatic-rich. The nano-TiO2 sedimentation experiments indicated that the presence of aromatic-rich HAs was more capable of stabilizing nano-TiO2 particles than was the presence of aliphatic-rich HAs. This result is because the deionized phenolic groups in the HAs were preferentially adsorbed on the nano-TiO2 surfaces, which generated a higher charge density on the nano-TiO2 surfaces and caused stronger repulsive forces among particles. Furthermore, the aromatic-rich TiO2-HA complexes exhibited a greater sorption capacity than the aliphatic-rich TiO2-HAs complexes and nonlinear phenanthrene sorption because of their higher affinity and the condensed state of aromatic fractions. Note that natural organic matters, such as humic acids, in aquatic environments can not only increase the stability of nanoparticles but can also influence the mobility of hydrophobic organic compounds (HOCs). Ó 2014 Published by Elsevier Ltd.

1. Introduction Nanoparticles (NPs) are widely used in industry and in daily life because of their unique electronic, optical, thermodynamic and catalytic properties (Cohen, 2001). Because of the large number of NPs that are produced and consumed, NPs are also released into the environment and might potentially enter the food chain or drinking water sources. Furthermore, NPs might accumulate in organisms, which may consequently cause undesirable environmental and health risks (Limbach et al., 2005; Baun et al., 2008; Liu et al., 2009). The environmental transportation and transformation, as well as the organism toxicity, of NPs in aquatic environments might be strongly dependent on their sizes, surface ⇑ Corresponding author at: College of Environment and Energy, South China University of Technology, Guangzhou, China. Tel.: +86 2039380512. E-mail address: [email protected] (C. Yang). http://dx.doi.org/10.1016/j.chemosphere.2014.05.002 0045-6535/Ó 2014 Published by Elsevier Ltd.

properties and interactions with other substances in water (Battin et al., 2009; Hofmann and Von der Kammer, 2009). When NPs are released into aquatic environments, they interact with natural organic matters (NOM), such as humic acids (HAs), which are abundant in the environment. Previous studies have reported that NOM can significantly increase the stability of NPs in aquatic environments (Chen and Elimelech, 2007; Hyung et al., 2007; Zhang et al., 2009; Keller et al., 2010). The NOM adsorbed on the surfaces of NPs can help to maintain smaller particles sizes and more negative surface charges on the NPs (Zhang et al., 2009). Furthermore, the electrostatic double layer (EDL) repulsive energy of particles and the high energy barrier among these particles might destroy the previous balance between the attractive and repulsive potential of particles (Keller et al., 2010). Additionally, note that NPs can be stabilized by HAs through complexation between the acidic functional groups (mainly carboxylic acid) in HAs and the surface of NPs (Hajdú et al., 2009).

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On the other hand, NOM can also influence the transportation of hydrophobic organic contaminants (HOCs) through sorption processes on minerals, soils and sediments (Kang and Xing, 2005; Liang et al., 2006). It was reported that oxides coated with NOM exhibited significant sorption abilities, suggesting that NOM possessed an important role in the sorption of HOCs (Wang et al., 2008; Yang and Xing, 2009). Previous studies suggested that the sorption of HOCs on minerals coated with NOM might be related to the solution chemistry and properties of the minerals and HAs (Perminova et al., 1999; Kang and Xing, 2005). The orientations and structures of adsorbed HAs might affect the surface area of the organic phase and the accessibility of hydrophobic domains that control the sorption of HOCs (Murphy et al., 1990). The adsorbed HAs can also increase the sorption affinity and isotherm nonlinearity of phenanthrene on the surfaces of nano-TiO2 and nano-ZnO by increasing the p-polarity/polarizability of adsorbed HAs (Yang and Xing, 2009). It is known that HAs in aquatic environments contain heterogeneous chemical compositions and functional groups because of the various sources from which they are obtained (Gonzalez-Vila et al., 1992; Filella, 2009) and that they might exert different influences on the stabilities and sorption abilities of NPs. HAs are an important fraction of NOM; they are macromolecules composed of multifunctional aromatic components linked by a variety of aliphatic and acidic (mainly carboxylic and phenolic) functional groups (Hayes et al., 1989; Tombácz, 1999). In the present study, HAs extracted from three different sources were employed as the target NOM. Nano-titanium dioxide (nano-TiO2) was selected as the target NPs. Nano-TiO2 is an important inorganic nano-material that is widely used in many commercial applications, such as cosmetics, photocatalysts and pigments (Macwan et al., 2011). The effects of HAs on the stabilities and sorption abilities of nano-TiO2 were investigated using sedimentation and sorption experiments to elucidate the relationships between the properties of HAs with heterogeneous functional groups and the aggregation and sorption behaviors of nano-TiO2.

2. Materials and methods 2.1. Materials Nano-TiO2 (anatase) was purchased from the Aladdin Reagent Company, Shanghai, China. The purity, diameter and specific surface area of the nano-TiO2 were >99.8%, 40 nm and 97.46 m2/g, respectively. Three HAs were obtained from different sources (peat (HA1), river sediment (HA2), and straw (HA3)). HA1 and HA2 were extracted from peat and Pearl river sediment, respectively, according to the method recommended by The International Association of Humic Acid (Swift et al., 1996). HA3 was extracted from aqueous suspensions of rice straw that had been soaked for several weeks (Thurman and Malcolm, 1981). Phenanthrene was purchased from the Sigma–Aldrich (Shanghai) Trading Co., Ltd. All of the other solutions were prepared using analytical grade chemicals (National Medicine Corporation Ltd., Shanghai, China).

2.2. Preparation of TiO2-HAs The TiO2-HA complexes were synthesized according to a previous study (Yang and Xing, 2009). Briefly, 5 g of nano-TiO2 was added to 1 L of a 1 g/L HA solution in a bottle and shaken at 150 rpm for 2 d; then, the suspension was centrifuged at 4000 rpm for 30 min. The precipitated materials were then rinsed three times with Milli-Q water, freeze-dried, ground and stored for use.

2.3. Characterization of HAs and TiO2-HAs The HAs were characterized using Fourier transform infrared spectroscopy (Vector 33, Bruker, Germany), and solid-state 13C nuclear magnetic resonance spectroscopy (Avance AV 400, Bruker, Switzerland) was employed to investigate the functional groups. 2.4. Effects of HAs on the Aggregation of Nano-TiO2 Stock solutions of HAs (1 g/L) were prepared by dissolving the HAs in a small amount of 0.1 M KOH followed by adjusting the pH to 7 with 0.1 M HCl. To prepare the nano-TiO2 dispersions in the HA solutions, HA solutions were prepared for the aggregation experiments with concentrations of 5 mg/L using stock solutions, and then the ionic strength was adjusted to 0.01 M with KNO3; finally, a specific mass of nano-TiO2 was added to the HA solutions. The three HA solutions were prepared under identical conditions. Sedimentation experiments were conducted using a UV spectrometer (Shimadzu UV-2550) to monitor the dynamics of the aggregation process. Briefly, after ultrasonication for 15 min, a suspension of nano-TiO2 in HA solution or a blank solution (0.01 M KNO3) was placed in a cuvette and measured at 508 nm. There was close relationship between the stability of the nanoTiO2 suspension and its electrokinetic properties. In electrolyte solutions, NPs can obtain a high surface charge density, which generates strong repulsive forces. Therefore, it was important to observe the electrophoretic behavior through measurements of the zeta potential to understand the dispersion behavior of NPs in aqueous solution (Li et al., 2007). To investigate the influence of HAs on the zeta potential of the nano-TiO2 aggregates, the suspensions were measured in different HA solutions and in a blank solution using a Malvern Zetasizer Nano-ZS after ultrasonication for 15 min. 2.5. Sorption experiments The sorption experiments were conducted using a batch equilibrium technique at 25 °C and pH 7.0. Phenanthrene was mixed at a high concentration in methanol before being added to the background solution. The concentration of methanol was maintained at less than 0.1% of the total solution volume to avoid the cosolvent effect. The background solution contained 0.003 M NaN3 to minimize bioactivity and 0.01 M KNO3 to adjust the ionic strength. Ten to fifty milligrams of the TiO2-HA complex with 15 mL of phenanthrene solution was added to a 20 mL screw cap vial equipped with a Teflon gasket and mixed for 7 d on a shaker at 150 rpm. Our preliminary experiment indicated that sorption equilibrium was reached within 7 d. After the sorption experiments, the 20 mL screw cap vial was centrifuged at 4000 rpm for 30 min, and 1 mL of the supernatant was transferred to a pre-weighed 1.5 mL amber glass vial for chemical analyses. Each concentration level, including blanks, was measured in three replicates. The aqueous concentrations of phenanthrene were determined with reverse-phase high-performance liquid chromatography (Agilent 1200) with C18 column (5 lm, 250  4.6 mm; Agilent) using a fluorescence detector with an excitation wavelength of 254 nm and an emission wavelength of 385 nm. The mobile phase (flow rate of 1.0 mL/min) was a mixture of Milli-Q water and acetonitrile in a 10:90 ratio by volume. The solid phase concentration was calculated based on the mass balance of the solute between the two phases. The classical Henry (Eq. (1)) and Freundlich (Eq. (2)) sorption equations were used to fit the equilibrium sorption data:

qe ¼ kd  C e

ð1Þ

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qe ¼ kf  C ne

ð2Þ

where Ce (mg/L) and qe (lg/g) are the equilibrium concentrations of phenanthrene in the liquid phase and in the solid phase, respectively; Kd (L/kg) is the distribution coefficient of solute between solid and water; Kf (lg/g)/(mg/L)n is the Freundlich distribution coefficient; and n is the non-linearity exponent. 3. Results and discussion 3.1. Solid-state

13

C NMR spectra

The results of the solid-state 13C NMR spectra for the three HAs are shown in Table 1. In the 0–220 ppm chemical shift region, the C atoms were attributed to alkyl C (0–50 ppm), O-alky C (50– 110 ppm), aromatic C (110–143 ppm), phenolic C (143–158 ppm), carboxyl C (158–190 ppm), and carbonyl C (190–220 ppm) (Kang et al., 2003). The ratio of the aliphatic C region (0–110 ppm) in HA2 or HA3 was larger than that in HA1, whereas the ratio of aromatic C (110–158 ppm) in HA1 was larger than that in HA2 or HA3. The aromaticity of the three HAs, in descending order, was HA1 > HA2 > HA3. This result indicated that HA2 and HA3 were aliphatic-rich, whereas HA1 was aromatic-rich. According to Gondar’s study, the content of carboxylic groups according to the 13C NMR spectra was equal to that obtained by potentiometric titration (Gondar et al., 2005). Therefore, the content of carboxylic groups in HA2 and HA3 was greater than that in HA1 according to the carboxyl C ratios of the three HAs. Similarly, the ratio of the phenolic C of HA1 was larger than that of HA2 and HA3, which suggested that the aromatic-rich HA might contain more phenolic groups. 3.2. FTIR spectra The difference in organic functional groups among the three HAs was also clearly observed in the FTIR spectra, as shown in Fig. 1a. The common features that could be observed from the comparable relative intensities of the three HAs included (a) the intense and broad band centered between 3450 cm1, which was attributed to the stretching of hydrogen-bonded OH, (b) a shoulder at 1245 cm1, which is typically attributed to the CAO stretching of aryl ethers or phenols, (c) a weak peak in the region of 1080– 1030 cm1, assigned to the CAO stretching of polysaccharide-like components or to the SiAO stretching of silicate impurities and (d) a strong peak at approximately 1652 cm1, generally attributed to aromatic C@C stretching or COOA symmetric stretching. These features indicated that the three HAs contained hydroxyl, phenolic and carboxyl groups. The different features of the three HAs included (a) a shoulder at 2925–2850 cm1, attributed to the CAH stretching of methyl or methylene groups of aliphatic chains, whose relative intensities were very weak (a shoulder) for HA1 but strong for HA2 or HA3 (Santos and Duarte, 1998), (b) a stretching vibration at 1598– 1509 cm1, generally attributed to aromatic C@C stretching, NAH deformation or C@N stretching of amides (amide II band), whose relative intensities were strong in the HA1 spectrum but very weak in the HA2 and HA3 spectra, (c) a weak peak at 1460 cm1, mainly Table 1 Structural groups analysis from solid-state Sample

HA1 HA2 HA3

Fig. 1. FTIR spectra of three HAs (a) and their complexes (b).

attributed to the aliphatic CAH deformation, which was present in the FTIR spectra of HA2 and HA3 but very weak in the HA1 spectrum (Senesi et al., 2003) and (d) a weak peak at 1189 cm1 due to the CAOH stretching of aliphatic OH, which was only present in the spectra of HA2 and HA3. Therefore, HA1 contained more aromatic structures with phenolic groups than HA2 and HA3, which contained more aliphatic chains with hydroxyl groups; this result was consistent with the results of the solid-state 13C NMR spectra. The changes in organic functional groups that were caused by surface complexation–ligand exchange reactions between the HAs and nano-TiO2 surfaces are shown in Fig. 1b. Through a comparison of the three HAs, the observed disappearance of the phenolic OH peak at 1245 cm1 implied the presence of strong interactions of phenolic OH groups with the nano-TiO2 surfaces (Yang and Xing, 2009). Peak shifts of COOH stretching at 1652 cm1 and of aliphatic OH groups at 1460 cm1 were observed in the FTIR spectra of the coated HAs, which suggested that the

13

C NMR spectra for three HAs.

Distribution of C chemical shift (ppm), %

Aromaticity

0–50 ppm

50–110 ppm

110–143 ppm

143–158 ppm

158–190 ppm

190–220 ppm

7.16 33.18 18.28

5.16 16.39 45.16

68.67 22.79 20.53

4.66 3.18 3.78

9.03 19.97 9.14

5.30 4.49 3.11

Aromaticity: aromatic C (110–158 ppm)/(aliphatic C (0–110 ppm) + aromatic C (110–158 ppm)).

85.62 34.37 27.70

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interactions of COOH and aliphatic OH groups with nano-TiO2 were insignificant (Kang and Xing, 2008; Yang and Xing, 2009). Therefore, based on the FTIR spectra, the adsorbed HAs had the greatest contribution for the interaction of phenolic OH with nano-TiO2, which was consistent with previous studies (Yang et al., 2009; Yang and Xing, 2009). 3.3. Sedimentation experiments The results from the sedimentation studies with nano-TiO2 suspensions in HA solutions and in the blank solution are shown in Fig. 2. The nano-TiO2 in the HA solutions did not settle, and the suspension remained stable over a longer period of time than in the blank solution. In the blank solution, the concentration of suspended particles decreased by more than 70% in 180 min. This concentration of suspended particles decreased rapidly thereafter because the NP aggregates readily settled and were removed from the solution, which decreased the concentration of dispersed NPs. However, in the three HA solutions, the concentrations of suspended particles decreased by only approximately 20% in 180 min. Our results were consistent with those of previous studies, which also observed that the NOM significantly reduced the rate of sedimentation of metal oxide NPs (Pettibone et al., 2008; Keller et al., 2010). This phenomenon was due to the NOM adsorbed on the surface of the NPs, which formed an electrostatic double layer (EDL) on the surface of the particles to significantly increase the repulsive energy between the NPs, thereby enhancing the stability of NPs in water (Pettibone et al., 2008; Zhang et al., 2009). Therefore, in aquatic environments containing NOM, NPs might be transported relatively longer distances, which increases their potential risks. The rates of sedimentation of the three HAs were largely similar, although there were subtle differences. The rate of sedimentation was the lowest in the HA1 solution, followed by the HA2 and HA3 solutions. The sedimentation curves of HA2 and HA3 exhibited a difference at 100 min, which might be related to an alteration of the nanoparticles influenced by the adsorbed HA as a function of time. As time increased, the amount of adsorbed HAs increased and the effects of heterogeneous of HA would be dominant, which might result in the impact of HAs becoming stronger. This phenomenon was consistent with that reported in previous studies (Pettibone et al., 2008; Keller et al., 2010). 3.4. Zeta potential of nano-TiO2 To further investigate the impact of HAs on the stability of nano-TiO2, zeta potential measurements were employed to gain

Fig. 2. Sedimentation plots for nano-TiO2 in the HA solutions and blank solution.

insight into the aggregates formed in different HA solutions and in the blank solution (Fig. 3). The zeta potential of nano-TiO2 in the blank solution was 6.90 mV. However, in the HA solutions, the zeta potentials of the particle surface were lower: 27.70 mV for HA1, 21.15 mV for HA2 and 14.25 mV for HA3. In the presence of HAs, the nano-TiO2 became more negatively charged. This observation was consistent with the literature on the study of particle aggregation, which reported that the addition of NOM resulted in more stabilized particle suspensions (Fairhurst and Warwick, 1998; Zhang et al., 2009). In the blank solution, the force of electrostatic repulsion between particles was not sufficient to overcome the attractive force, so the NPs were always aggregated. HAs can be considered to be a polyelectrolyte that is generally negative in aqueous solution due to the dissociation of the acidic groups (mainly carboxylic and phenolic hydroxyl) (Tombacz et al., 2000). Therefore, when the negatively charged HAs were adsorbed onto the surfaces of the NPs through ligand exchange (Yang et al., 2009), the surface charge density of the NPs increased and generated strong repulsive forces, which stabilized the NP suspension. In comparison, the zeta potentials of nano-TiO2 in the three HA solutions were, in increasing order, HA1 < HA2 < HA3, which indicated that the stabilities of nano-TiO2 in the three HA solutions were, in descending order, HA1 > HA2 > HA3. This trend was the same as that in the sedimentation experiments, which revealed decreasing aggregation in HA solutions, thereby suggesting that the three HAs from the three sources possessed quite different surface functional groups. According to the solid-state 13C NMR spectral analysis, the aromaticity and phenolic C percent of HA1 were significantly higher than those of HA2 and HA3. This trend was the same as the trend of the stabilities of nano-TiO2 in the three HA solutions, which indicated that the stabilities of nanoTiO2 in the three HA solutions were most likely related to the presence of negative phenolic acid groups. The higher aromaticity of HAs that might contain a higher percentage of phenolic C can stabilize nano-TiO2 particles better. This result was somewhat contradictory to the literature on magnetite, which reported that magnetite was stabilized by citric acid and HAs mainly because the carboxylic groups of the citric acid and HA were able to form surface complexes on the „FeAOH sites of iron oxides (Hajdú et al., 2009). This discrepancy was most likely due to the different nanoparticle properties. According to the FTIR spectral analysis, the mechanism for the sorption of HAs on the nano-TiO2 surface was likely ligand exchange between the nano-TiO2 surfaces and the phenolic OH of the HA fractions. Therefore, the decreasing zeta potentials of the nano-TiO2 in the three HA solutions were most likely due to the presence of negative phenolic acid groups and, to a lesser extent, carboxylic groups in the HAs because they were not important in the sorption process. The zeta potentials in the

Fig. 3. Zeta potentials of nano-TiO2 in the HA solutions (5 mg/L) and blank solution.

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HA2 solution were significantly different from those in the HA3 solution, which indicated that the negative groups on the nanoTiO2 surfaces in the HA2 solution were significantly higher than those in the HA3 solution. The adsorbed HAs could modify, either entirely or partially, the surface charge properties of the nanoTiO2 (Hajdú et al., 2009). The charges carried above the adsorption saturation depended on the amount and the dissociation degree of the unbound negative groups in the adsorption layer (Tombácz, 1999). The significant difference might be due to the higher dissociation degree of HA2 and the lower steric hindrance in the binding groups of HA2. This result was consistent with the sedimentation results. 3.5. Sorption of phenanthrene by nano-TiO2 and complexes Phenanthrene isotherms of nano-TiO2 and TiO2-HA complexes are presented in Fig. 4. As indicated in this figure, the sorption of phenanthrene by nano-TiO2 was significantly enhanced by coating with HAs, which can be primarily attributed to the introduction of hydrophobic sites in the coated HA phase (Wang et al., 2008). Previous studies reported that HAs showed strong sorption affinity to HOCs (Chin et al., 1997; Kang and Xing, 2005; Ran et al., 2007). As shown in Fig. 4, there was a significant difference in the sorption affinity of phenanthrene on different complexes, which was in the order of TiO2-HA1 > TiO2-HA2 > TiO2-HA3. This result suggested that the different properties of HAs can cause different orientations and structures of adsorbed HAs, subsequently affecting the surface area of the organic phase and the accessibility of hydrophobic domains that control the sorption of HOCs (Murphy et al., 1990). All isotherms of the HA-coated nano-TiO2 were fitted using the Henry and Freundlich isotherm models, and the isotherm parameters are listed in Table 2. As shown in Table 2, the Henry model could fit the sorption data of the three complexes well according to the regression coefficients (R2 ranged from 0.983 to 0.994), which indicated that the sorption of phenanthrene onto the complexes might be related with linear partitioning. Hydrophobic interactions were considered to be responsible for the linear partitioning (Chiou, 2003). Similar results were also observed for the sorption data of pyrene on NOM-NPs (R2 ranged from 0.973 to 0.994) (Wang et al., 2008). The Freundlich model could also fit the sorption data well (R2 ranged from 0.982 to 0.993). The Freundlich model is an empirical equation that assumes an exponential variation in site energies (Ng et al., 2002), and it is employed to explain the sorption behavior of organic chemicals on heterogeneous surfaces, such as HA substances (Benedetti et al., 1996). This indicated that the sorption properties would be influenced by heterogeneous surface of HA and that the sorption mechanism might not be the only partition process.

Fig. 4. Sorption isotherms of phenanthrene on nano-TiO2 and complexes.

Table 2 Sorption isotherms parameters of phenanthrene on nano-TiO2 and complexes. Sample

TiO2 TiO2-HA1 TiO2-HA2 TiO2-HA3

Henry model

Freundlich model 2

Kd (L/kg)

R

6.713 715.113 348.302 216.480

0.920 0.994 0.994 0.983

Kf (lg/g)/(mg/L)n

n

R2

79.799 2018.366 157.036 110.154

0.554 0.785 1.167 1.115

0.728 0.993 0.992 0.982

The Freundlich parameter n, which reflects the sorption nonlinearity, was determined. The values of this parameter were 0.785 for the aromatic-rich sorbents TiO2-HA1, whereas they were 1.167 and 1.115 for the aliphatic-rich sorbents TiO2-HA2 and TiO2-HA3, respectively, which indicated that the isotherm of the aromatic-rich sorbent was more nonlinear than that of the aliphatic-rich sorbent. This result indicated that the aromatic fractions played a key role in sorption nonlinearity, which was in good agreement with previous studies (Xing, 2001; Yang and Xing, 2009). Dual-mode sorption or the dual-reactive domain model is generally used to explain the nonlinearity of NOM. These models describe NOM as a heterogeneous substance that consists of two types of domains: an expanded domain and a condensed domain (Kang and Xing, 2005; Yang et al., 2005). Linear isotherms were observed for the expanded domain due to the hydrophobic partitioning mechanism, whereas nonlinear isotherms were observed for the condensed domain (Gunasekara and Xing, 2003). Moreover, aromatic fractions were considered to be an important contributor for the formation of the condensed state of HAs, whereas aliphatic fractions were considered to be an important contributor for the formation of the expanded state. Therefore, the aromatic fractions were most likely responsible for the nonlinear sorption of phenanthrene. The nonlinear sorption isotherm of phenanthrene by TiO2-HA1 indicated that its sorption onto the coated NOM phase was dominated by condensed domains created at their interfaces, whereas the linear sorption isotherms of phenanthrene by the TiO2-HA2 and TiO2-HA3 complexes were dominated by a partitioning mechanism, and no condensed domains were created at their interfaces (Wang et al., 2008). The estimated Kd values for the Henry model obtained in this study were 6.713 L/kg for nano-TiO2 and ranged from 216.480 to 715.113 L/kg for the complexes, which were consistent with a previous study (Wang et al., 2008). The Kd values of pyrene have been reported to range from 26 to 63 for NPs and from 226 to 1530 for NOM-NP complexes (Wang et al., 2008). The sorption coefficients (Kf) were 79.799 (lg/g)/(mg/L)n for nano-TiO2 and ranged from 110.154 to 2018.366 (lg/g)/(mg/L)n for the complexes. These results were consistent with those of a previous study (Yang and Xing, 2009). By comparison, the TiO2-HA1 complex had a significantly larger Kd value for phenanthrene than did the TiO2-HA2 or TiO2-HA3 complexes, which suggested that the amount of phenanthrene adsorbed on the TiO2-HA1 complex was greater than that adsorbed on the TiO2-HA2 or TiO2-HA3 complex. The correlation coefficient between the adsorption coefficients (log Kf) and aromaticity of the target HAs was 0.979, which indicated a good linear relationship. This phenomenon may be due to the stronger affinity to the high aromaticity of HA1 (Perminova et al., 1999; Ahmad et al., 2001), which indicated that the aromatic fractions played a key role in the sorption of phenanthrene after being coated on TiO2. Furthermore, because the aromatic fractions were preferentially adsorbed by nano-TiO2 and because HA1 showed higher aromaticity, a larger amount of HA1 can be coated on the nano-TiO2 surfaces, which would form larger sizes of the hydrophobic domains in the organic phase and increase the accessibility to phenanthrene molecules (Garbarini and Lion, 1985; Murphy et al., 1990).

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4. Conclusion In the present study, the stability and the sorption of phenanthrene on nano-TiO2 coated with HAs were investigated. The HAs significantly enhanced the stability and the sorption ability of nano-TiO2, which implied that NOM coated on NPs can alter the environmental behaviors of NPs and HOCs. The influences of HAs on the stability and sorption ability of nano-TiO2 were different because of the heterogeneous structures in the HAs extracted from different sources. The aromatic-rich HAs might have a more substantial effect on the behavior of NPs in aquatic environments and should be considered when assessing the risks of nanomaterials. Acknowledgements The study was financially supported by the National High Technology Research and Development Program of China (2012AA101403), by the China National Science Fund Program (Nos. 41072268 and 41173104) and by the Pearl River Young Scientist Project of Guangzhou (2011J2200060). The authors appreciate the suggestions regarding the language in the manuscript from Dr. Thomas Bucheli of the Agroscope Reckenholz-Tänikon ART. References Ahmad, R., Kookana, R.S., Alston, A.M., Skjemstad, J.O., 2001. The nature of soil organic matter affects sorption of pesticides. 1. Relationships with carbon chemistry as determined by 13C CPMAS NMR spectroscopy. Environ. Sci. Technol. 35, 878–884. Battin, T.J., Kammer, F.V., Weilhartner, A., Ottofuelling, S., Hofmann, T., 2009. Nanostructured TiO2: transport behavior and effects on aquatic microbial communities under environmental conditions. Environ. Sci. Technol. 43, 8098– 8104. Baun, A., Hartmann, N., Grieger, K., Kusk, K.O., 2008. Ecotoxicity of engineered nanoparticles to aquatic invertebrates: a brief review and recommendations for future toxicity testing. Ecotoxicology 17, 387–395. Benedetti, M., Van Riemsdijk, W., Koopal, L., 1996. Humic substances considered as a heterogeneous Donnan gel phase. Environ. Sci. Technol. 30, 1805–1813. 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, 126–134. Chin, Y.-P., Aiken, G.R., Danielsen, K.M., 1997. Binding of pyrene to aquatic and commercial humic substances: the role of molecular weight and aromaticity. Environ. Sci. Technol. 31, 1630–1635. Chiou, C.T., 2003. Partition and adsorption of organic contaminants in environmental systems. John Wiley & Sons. Cohen, M.L., 2001. Nanotubes, nanoscience, and nanotechnology. Mater. Sci. Eng. C 15, 1–11. Fairhurst, A.J., Warwick, P., 1998. The influence of humic acid on europium–mineral interactions. Colloids Surf., A 145, 229–234. Filella, M., 2009. Freshwaters: which NOM matters? Environ. Chem. Lett. 7, 21–35. Garbarini, D.R., Lion, L.W., 1985. Evaluation of sorptive partitioning of nonionic pollutants in closed systems by headspace analysis. Environ. Sci. Technol. 19, 1122–1128. Gondar, D., Lopez, R., Fiol, S., Antelo, J., Arce, F., 2005. Characterization and acid-base properties of fulvic and humic acids isolated from two horizons of an ombrotrophic peat bog. Geoderma 126, 367–374. Gonzalez-Vila, F.J., Martin, F., Del Rio, J., Fründ, R., 1992. Structural characteristics and geochemical significance of humic acids isolated from three Spanish lignite deposits. Sci. Total Environ. 117, 335–343. Gunasekara, A.S., Xing, B., 2003. Sorption and desorption of naphthalene by soil organic matter. J. Environ. Qual. 32, 240–246. Hajdú, A., Illés, E., Tombácz, E., Borbáth, I., 2009. Surface charging, polyanionic coating and colloid stability of magnetite nanoparticles. Colloids Surf., A 347, 104–108. Hayes, M.H.B., MacCarthy, P., Malcolm, R.L., Swift, R.S., 1989. Humic Substances II. In Search of Structure. John Wiley & Sons Ltd.

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