Arsenic(iii) removal from aqueous solution using TiO2-loaded biochar prepared by waste Chinese traditional medicine dregs

Oxidation of As(iii) to As(v) is an effective way to improve the performance of most arsenic removal technologies. In this study, a new alternative biosorbent, TiO2-loaded biochar prepared by waste Chinese traditional medicine dregs (TBC) was applied in remediation for As(iii) from aqueous solution. Compared with unmodified biochar, the specific surface areas and total pore volumes of TBC increased while the average aperture decreased due to the loading of nano-TiO2. The X-ray diffraction (XRD) of TBC confirmed that the precipitated titanium oxide was primarily anatase. pH did not have a significant effect on the adsorption capacity at 10 mg L−1 As(iii) in suspension with a pH ranging from 2 to 10. Adsorption kinetics data were best fitted by the pseudo-second-order model (R2 > 0.999). The Sips maximum adsorption capacity was 58.456 mg g−1 at 25 °C, which is comparable with other adsorbents reported in previous literature. The Gibbs free energy (ΔG) of As(iii) adsorption was negative, indicating the spontaneous nature of adsorption. The results of free radical scavenging and N2 purging experiments indicated that O2 acted as an electron accepter and O2˙− dominated the oxidation of As(iii). The oxidation of As(iii) obviously affected the adsorption capacity for As(iii) by TBC. X-ray photoelectron spectroscopy (XPS) studies showed that As(iii) and As(v) existed on the surface of TBC, suggesting that the oxidation of As(iii) occurred. TBC played multiple roles for As(iii), including direct adsorption and photocatalytic oxidation adsorption. Regeneration and stability experiments showed that TBC was an environment-friendly and efficient adsorbent for As(iii) removal.


Introduction
Arsenic (As) is a known toxic element whose compounds can cause acute and chronic arsenic poisoning, such as skin cancer, lung cancer and other diseases. 1 Elemental arsenic has been widely studied for a long time due to its biological toxicity and carcinogenic effect. 2 Arsenic in the environment mainly comes from (1) natural factors such as volcanic eruptions and rock weathering and (2) human activities such as industrial and agricultural activities, especially arsenic-containing wastewater from industrial production. 3,4 In nature, arsenic occurs in the +5, +3, 0, and À3 states, which are mainly in the form of sulde and oxide. 5 Oxidation states +3 and +5 commonly exist in water environments, depending on redox potential formation, but states À3 and 0 are very rare. 6 As(III) is more toxic and migratory than As(V). 5 under oxidizing conditions. 6 Due to the high toxicity of arsenic, As contamination is considered a recognized global environmental problem. 8 To minimize the biological toxicity of arsenic to human health, environmental restrictions and regulations strictly control the production and use of arsenic and its compounds. Arsenic is listed as a pollutant of priority concern by the European Union (EU) and the Environmental Protection Agency in the United States of America (US EPA). 9 The maximum permissible limit of inorganic arsenic in safe drinking water is 10 mg L À1 according to the World Health Organization (WTO). 1,2,5,8,9 US EPA stipulates that the arsenic emission standard of wastewater is 0.2 mg L À1 . 5 Technologies of arsenic removal from aqueous solution have been widely studied. At present, the main methods of arseniccontaminated water remediation include coagulation, ion exchange, membrane separation techniques, adsorption, sulfatereducing bacteria (SBR) and oxidation. 1,3,5,6,9 However, these methods have many disadvantages, such as limitation of removal efficiency, environmental unfriendliness or high cost. 5,10 Adsorption is an efficient water purication technology, with advantages such as simple operation, cost-effectiveness, minimal sludge production and regeneration. 10

Preparation of modied biochar
Biochar (BC) was prepared by waste Chinese traditional medicine dregs from the manufacture of an anticancer injection produced by Yibai Pharmaceutical Co., Ltd. The dregs mainly included a certain proportion of Acanthopanax senticosus, ginseng and Astragalus. The materials were washed twice with tap water, then washed with deionized water and dried in a drying oven (101-2AB, Taisite) at 80 C for 48 h. The dried dregs were crushed to less than 2 mm. Then, the biomass powder (80.00 g) was pyrolyzed in a tube furnace (SG-GL1200, Siom) under owing N 2 at an increasing rate of 10 C min À1 . The tube furnace was held at a target temperature of 450 C for 1 hour. The biochar powder was cooled at room temperature, ground and sied in 100 mesh.
TBC was synthesized at room temperature by a modied solgel method as reported previously. 20 In brief, a certain amount of anhydrous ethanol, butyl titanate and concentrated hydrochloric acid were fully mixed and stirred at a certain rotating speed. Then, 3.00 g BC was added to the mixture as solution A, which was stirred at a certain rotating speed. A certain amount of anhydrous ethanol and deionized water were mixed completely as solution B. Aer solution A was stirred for 30 min, solution B was added dropwise into solution A. The mixture continued to be stirred slowly until the solution formed a gel. Aer aging at room temperature for 24 hours without light, the gel was placed in an oven at 60 C and dried for 24 hours. Aer passing through a 100 mesh sieve, TBC was kept for reserve.

Adsorption experiments
The effect of the initial solution pH on the adsorption capacity at a pH range from 2 to 10 (at intervals of 2 pH units) was carried out at 25 C. pH adjustment was accomplished by using of 1 M, 0.1 M and 0.01 M HCl and NaOH solutions. Aer shaking for 24 hours, the suspension was ltered, and the nal pH values were recorded by a pH metre (PHS-3C, INESA, China). For adsorption kinetics experiments, initial concentrations of As(III) were 10, 20, 40 and 80 mg L À1 at 25 C. At each sampling time (5,10,20,40,60,120,240,480, 720, 1440 and 2880 min), the suspension was immediately ltered through 0.45 mm pore size membrane lters. In isothermal equilibrium experiments, the effect of the initial concentration of As(III) was studied for 24 hours by varying concentrations from 5 to 200 mg L À1 at 15 C, 25 C and 35 C. All experiments were carried out with a biochar-weight-tosolution-volume ratio of 2 mg mL À1 .
To understand the photocatalytic removal mechanism of As(III) onto TBC, two radical scavengers (SOD and TBA) were used to ensure that the major radical was responsible for As(III) removal and oxidation. The present investigation aimed to provide evidence for the mechanism of the oxidation and removal of trivalent arsenic from aqueous solution by TBC. SOD (superoxide dismutase) catalyses the dismutation of superoxide radicals to hydrogen peroxide and reacts with hydroxyl radicals. 21 t-Butanol alcohol (TBA) can selectively quench hydroxyl radicals. The contents of total As and As(III) were analysed by atomic uorescence spectrophotometry (AFS) (Haiguang, 8501).

Apparatus and measurements
The point of zero charge (PZC) of TBC and BC was examined using a zeta potential distribution/nanoparticle size analyser (DelsaNanoC particle analyser, USA) by the pH dri method. The surface areas of BC and TBC were examined by a 3H-2000PS4 (BSD, China) surface area analyser using BET N 2 adsorption-desorption isotherms. Scanning electron microscopy/energy-dispersive X-ray (SEM-EDX) was carried out to examine BC and TBC surface architectures by Hitachi SU8010 (Japan) scanning electron microscopy (SEM). Transmission electron microscopy (TEM) study was undertaken using a TEM JEOL JEM-2100. Fourier transform infrared (FT-IR) spectra was recorded using a spectrometer (FTIR-850 spectrometer, China). X-ray photoelectron spectroscopy (XPS) analysis was performed using a Thermo Fisher K-Alpha XPS system. The composition and crystal structure of the adsorbates were analysed with X-ray diffraction (Bruker D8 Advance, Germany). The test method of As(III) was carried out according to "speciation of trivalent arsenic and pentavalent arsenic in seawater by atomic uorescence spectroscopy". 22 The content of titanium in aqueous solution was detected by inductively coupled plasma mass spectrometer (ICP-MS) (Thermo Fisher, XII).

Biochar elemental analysis, surface areas and morphologies
As shown in Fig. 1(a) and (b), scanning electron micrograph results exhibited that the surfaces of BC and TBC were full of rough and clear pores. Aer modication, it can be seen in Fig. 1(c) and (d) that a large number of TiO 2 particles were loosely arranged on the surface of biochar fragments. Compared with BC, EDS data proved a decrease in surface carbon content and an increase in oxygen and titanium content in TBC due to TiO 2 covering portions of BC's carbonaceous surface ( Fig. 2 and Table 1). Data of BET N 2 adsorption and desorption isotherms for BC and TBC were shown in Fig. 3. The surface areas of BC and TBC were 4.5888 m 2 g À1 and 128.2171 m 2 g À1 , respectively. The total pore volumes of BC and TBC were 0.0235 mL g À1 and 0.0859 mL g À1 , respectively. The average apertures of BC and TBC were 20.4847 nm and 2.6798 nm, respectively. It was proven that nano-TiO 2 particles were successfully loaded onto BC. The specic surface area and total pore volumes of TBC increased while the average aperture decreased due to the loading of nano-TiO 2 onto BC. Fig. 1(n), Fig. 2(c) and Table 1 showed SEM-EDX elemental maps of arsenic adsorbed onto TBC, which proved the presence of As.
The XRD spectrums of BC showed that the ne and sharper peaks appeared at 2q values of nearby 20-30 , indicating the formation of aromatic carbon compounds in the biochar materials and an additional amorphous phase coexisted with the crystalline phase (Fig. 4). 23 Peaks of calcium carbonate were observed in BC. Unmodied biochar exhibited a crystallized peak at 26.64 (quartz), indicating that BC contained some SiO 2 . The XRD of TBC conrmed that the precipitated titanium oxide was primarily anatase (TiO 2 ). The major peak at 2q ¼ 25.303 was for the crystalline plane with Miller indices of (101). Other low-intensity peaks were observed at 37.792 (004), 49.035 (200) and 53.753 (105), matching the standard diffraction patterns of TiO 2 . The broadening of the anatase (TiO 2 ) peaks suggested that anatase (TiO 2 ) onto TBC were poorly crystalline domains due to the composite TBC prepared at low temperature. Aer modication, peaks at 2q values of nearby 20-30 were submerged, the quartz peaks became relatively diminished. The peaks of calcium carbonate disappeared. The reason may be that the peaks of calcium carbonate were masked by the peaks of the anatase (TiO 2 ) or calcium carbonate was dissolved in the synthesis of modied biochar. These results revealed successful impregnation of BC with TiO 2 . The points of zero charge of BC and TBC are $2.30 and $4.40, respectively (Fig. 5).  values ranging from 2 to 10. 5 The inuence of initial pH on the As(III) adsorption capacity of BC and TBC was researched at pH values ranging from 2 to 10. The result was shown in Fig. 6. For BC, the adsorption capacity of arsenic was poor, ranging from 0.068 mg g À1 to 0.408 mg g À1 at an initial concentration of 10 mg L À1 As(III). It can be seen in Fig. 6(a) that BC had a buffering effect on solution pH. For TBC, the initial pH had little effect on the adsorption capacity of As(III), which ranged from 4.778 mg g À1 to 4.901 mg g À1 . It was worth noting that the pH values of the suspensions decreased signicantly aer adsorption due to TiO 2 loaded onto BC. Previous studies reported that pH had little inuence on the effect of TiO 2 photocatalytic oxidation (PCO) and the adsorption capacity of As(III). 24 For TiO 2 particles, its surface is positively charged at pH < pH PZC (eqn (1)), whereas at pH > pH PZC, the TiO 2 surface is negatively charged (eqn (2)). 24 Other studies also proposed the complexation reaction (eqn (3)

Adsorption kinetics
The relationship between the adsorption capacity and equilibrium time was plotted at initial concentrations of 10, 20, 40 and 80 mg L À1 in Fig. 7(a). At the beginning of the reaction, the adsorption capacity (q t ) increased quickly with increasing adsorption time due to a large number of active sites and large concentration gradients. 27 As the concentration gradient and adsorption activity centrality gradually decreased, the degree of increase in adsorption capacity gradually decreased with increasing adsorption capacity until equilibrium. When the initial concentration of As(III) increased from 10 mg L À1 to 80 mg L À1 , the adsorption capacity increased, while the removal percentage decreased ( Fig. 7(b)). The average maximum removal percentages at initial concentrations of 10, 20, 40 and   80 mg L À1 were 98.62%, 94.56%, 89.79% and 82.56%, respectively. The reason would be attributed to the gradual reduction and saturation of available binding sites at modied biochar. 20 To further explore the adsorption behaviour of arsenic on TBC, kinetics data at As(III) levels of 10, 20, 40 and 80 mg L À1 were simulated with the pseudo-rst-order (PFO), pseudosecond-order (PSO), intra-particle diffusion (IPD) and Elovich (ELH) kinetic models. Linearized forms of aforementioned models have been presented in eqn (4)-(7). 1,5,27,28 ln(q e À q t ) ¼ ln q e À k 1 t (4) In eqn (4)- (7), t (min) is the contact reaction time; q t (mg g À1 ) is the adsorption capacity at different contact reaction times (t, min); q e,exp (mg g À1 ) is the absorption capacity at adsorption equilibrium in experiments; q e,cal (mg g À1 ) is the absorption capacity calculated according to the model; k 1 (min À1 ) is the PFO rate constant, and k 2 (g mg À1 min) is the PSO rate constant; a (mg g À1 min À1 ) is the initial adsorption rate, b (g mg À1 ) is the Elovich constant. Parameters (k dif and C) of the IPD model are estimated from the slope and intercept of q t vs. t 0.5 . Parameters in the ELH model (a and b) were obtained from the slope (1/b) and intercept 1/b ln(ab) of the curve q t vs. ln(t). In general, the Elovich kinetic model has been usually used for describing the adsorption kinetic of contaminants on the surface of a non-uniform solid adsorbent. 28 The calculated parameters were shown in Table 2. The adsorption kinetics were closer to the PSO model due to higher correlation coefficients of above 0.99 at all concentrations than to the PFO model, suggesting that the adsorption capacity was controlled by the number of available active sites on TBC. 27 A comparatively high As(III) uptake rate (k 2 ) decreases with the increases of initial As(III) concentration. The q e,cal values of the PSO model at 10, 20, 40 and 80 mg L À1 were 4.748, 9.194, 17.844 and 35.224 mg g À1 respectively, which perfectly coincided with the experimental data (4.869, 9.284, 17.851 and 32.924 mg g À1 ).
The results indicated that the adsorption process of As(III) onto TBC was dominated by chemical adsorption, because the PSO model was considered that the adsorption rate was controlled by the chemical adsorption between the adsorbate and the adsorbent. 18 According to the IPD model, two distinct stages of adsorption were clearly observed in Fig. 7(e). At the beginning of the adsorption process, the adsorption capacity increased rapidly with contact reaction time, then increased slowly, and the adsorption nally reached equilibrium. The rst stage showed a good linear relationship between q t vs. t 0.5 , corresponding to the diffusion of As to the outer surface of TBC through external diffusion. The second stage showed that As entered through intragranular diffusion and was adsorbed inside the pore size, and the adsorption rate at this stage was controlled by diffusion in the particles. The tting curve did not pass through the origin, indicating that the diffusion process in particles was not the only control step of the adsorption process. 28 Compared with the second stage, the rst stage had a larger slope, indicating a faster surface diffusion rate. 27,28 The second stage adsorption was related to the thickness of the boundary layer. 28 There was a larger intercept in the second stage, implying that the intra-particle diffusion played an important role in the rate-limiting step. 27,28 The sequence of the suitability of kinetic models based on the value of R 2 was PSO > ELH > PFO > IPD.

Adsorption isotherms
To further understand the adsorption mechanism and the maximum capacity of As(III) removal by TBC, adsorption isotherm data were obtained at 15 C, 25 C and 35 C. The data were tted by the Langmuir (eqn (8)), Freundlich (eqn (9)) and Sips models (eqn (10)), respectively. 1,5,28-30 In eqn (8)- (10), c e is the equilibrium concentration (mg L À1 ); q e is the adsorption capacity (mg g À1 ), and q m is the maximum adsorption capacity (mg g À1 ). Here, K L , K F and K S are the constants of the Langmuir, Freundlich and Sips models, respectively. In the Freundlich and Sips models, n is the adsorption intensity. The Freundlich model is suitable for low initial concentrations, while the Langmuir model is suitable for high adsorbate concentrations. To avoid the limitation of two parameter models, the three parameter Sips model is also applied. At low As(III) concentrations, the Sips model is simpli-ed to the Freundlich isotherm. At high concentrations, it predicts the monolayer adsorption capacity characteristics of the Langmuir model. 5 The tting data were listed in Table 3. As shown in Fig. 8, the experimental data were well tted to three isotherm models with R 2 (>0.90). The absorption procedure of TBC for As(III) can be better described by the Langmuir isotherm model than the Freundlich isotherm model. The favourability of adsorption can be characterized with a dimensionless constant (R L ) of the Langmuir isotherm dened with eqn (11). 5,29,30 where c 0 is the initial concentration (mg L À1 ). In this study, the value of R L was in the range of 0.046 to 0.341, indicating the favourable adsorption of As(III) onto TBC at 15 C, 25 C and 35 C. 29 Freundlich's intensity factor n F in the Freundlich isotherm was also used to evaluate the favourability of As(III) onto TBC, representing the intensity of the adsorption process. In this study, n F was calculated to be 3.361, 3.643 and 3.520 at 15 C, 25 C and 35 C, respectively, indicating the favourable adsorption of As(III) onto TBC. 29 The best t isotherm was selected by its goodness of t, correlation coefficient (R 2 ) and chi-square (c 2 ) value of tting to experimental data. The results showed that the best correlation coefficient and chi-square values were obtained by the Sips model. The results suggested that the adsorption process was controlled by diffusion at low As(III) concentration, while the adsorption was monolayer adsorption at As(III) concentration and had a saturation value.

Adsorption thermodynamics
From Fig. 8 and Table 3, the experimental data proved that the adsorption reaction temperature inuenced the As(III) adsorption equilibrium and capacity. The adsorption capacity (q e,exp ) signicantly increased from 45.526 mg g À1 to 50.394 mg g À1 with increasing adsorption reaction temperature from 15 C to 25 C. At higher temperatures (35 C), it was slightly higher than that at 25 C. The results indicated that the adsorption capacity increased as the temperature increased. The Sips adsorption capacity (q S ) for As(III) onto TBC increased from 15 to 25 C (57.963 to 58.456 mg g À1 ), while it slightly dropped to 35 C (57.453 mg g À1 ). The Langmuir model capacity (q m ) increased from 43.642 to 53.539 mg g À1 with increasing temperature from 15 to 35 C, also indicating increasing capacity as temperature rose.
The changes in Gibbs free energy (DG), enthalpy (DH) and entropy (DS) were calculated using van't Hoff's equations, as shown in Table 4. The consistently negative values of DG over the temperature range from 288 to 308 K indicated that the adsorption process of TBC for As(III) was spontaneous. A continuous drop in DG with temperature from 288 K to 308 K implied that the adsorption capacity increased with increasing temperature. Meanwhile, the value of DH was positive. This meaned that the adsorption process of As(III) onto TBC was an endothermic reaction, and an increase in temperature was conducive to the reaction. This supported the inuence of temperature on the adsorption properties. The Sips maximum adsorption capacity of 58.456 mg L À1 at 25 C is much higher than that of other adsorbents reported in previous literature, as shown in Table 5. Table 5 listed a summary of some studies on the adsorption performance of adsorbents for As(III).

Reusability and stability
Generally, a high recycling performance and low-cost regeneration solution for adsorbents is required. In this study, 0.1 mol NaOH, 0.1 mol HCl, 0.1 mol EDTA and H 2 O were employed to evaluate the reusability of As(III) removal by TBC. There was no signicant difference in the effects of the four regeneration solutions. As shown in Fig. 9(a), the adsorption capacity decreased slightly aer 5 cycles, from 9.274, 9.306, 9.305 and 9.274 mg g À1 to 6.422, 5.649, 6.755 and 6.145 mg g À1 with 0.1 mol NaOH, 0.1 mol HCl, 0.1 mol EDTA and H 2 O, respectively. The sequence of the recycling performance was 0.1 mol EDTA > 0.1 mol NaOH > H 2 O > 0.1 mol HCl. Overall, TBC is a cost-effective adsorbent with good reproducibility and reusability.
In order to consider the stability of TBC, the release of titanium from TBC into the solution was also detected. The solution with the ratio of modied biochar weight to deionized water volume of 2 mg mL À1 was shaken for 24 hours at 25 C, then the solution was ltered to detect the content of titanium. The shaking experiment was carried out ve cycles. The experimental result was shown in Fig. 9(b). Aer ve adsorption cycles, the release rate of titanium was only 0.011%. It suggested that the titanium release rate of TBC was low and had little impact on the environment, indicating that TBC had good stability and was a friendly adsorbent.

Adsorption mechanism
3.7.1 FT-IR. FT-IR spectra of BC and TBC were shown in Fig. 10. The functional groups in biochar clearly changed aer loading with TiO 2 . Peaks at 3402-3432 cm À1 corresponded to the symmetric and asymmetric O-H vibrations of water molecules. 40 The hydroxyl groups were observed with a maximum peak at 3432 cm À1 in BC. Aer activated by TiO 2 , peaks at 570 cm À1 corresponding to Ti-O or Ti-O-Ti stretching vibrations and peaks at 1620 cm À1 corresponding to Ti-OH vibrations appeared. [40][41][42] The hydroxyl groups experienced a redshi from 3432 cm À1 to 3409 cm À1 in TBC. Following arsenic adsorption, the hydroxyl groups in TBC once again experienced a redshi from 3409 to 3402 cm À1 , and the peak ascribing to the Ti-OH vibration shied from 1620 to 1612 cm À1 . The peak at 1057 cm À1 was attributed to -OH deformation vibrations. 23 The peak at 1057 cm À1 was present on BC and TBC but disappeared aer As was loaded onto TBC, and in the case of TBC a new peak appeared at 771 cm À1 , corresponding to As-O stretching vibration. The adsorption band between 650 cm À1 and 950 cm À1 corresponding to As(V)-O stretching vibrations was reported in previous literature. 43 The peak at 1589 cm À1 in BC could be ascribed to the C]O asymmetric stretching of ketones, aldehydes, lactones or carboxyl groups. 32,44,45 The peak at 2900 cm À1 was the C-H stretching vibration or CH 2 group. 46,47 The peaks between 1427 cm À1 and 1440 cm À1 represented the presence of aromatic rings or rings with C]C bonds. 23 3.7.2 The role of TiO 2 modication in TBC on the adsorption and oxidation of As(III). TiO 2 nanomaterials are widely used in arsenic remediation. However, its adsorption and photocatalytic capacity have been controversial. 19,48 To further explore the mechanism of As(III) removal by TBC, two radical scavengers (SOD and TBA) were used to reveal the major radicals responsible for the oxidation of As(III). At the beginning of the adsorption reaction, the concentration of As in As-TBC immediately decreased from 10 mg L À1 to 0.659 mg L À1 , as shown in Fig. 11(a). The arsenic removal rate reached 93.41% within 5 minutes. Aer that, the arsenic concentration gradually decreased to 0.083 mg L À1 , and the removal rate was 99.17% within 4 hours. The decreasing tendency of As concentration in As-TBA-TBC was similar to that in As-TBC. In comparison with As-TBA-TBC and As-TBC, there was an obvious difference from the decreasing tendency of the As concentration in As-SOD-TBC. At the beginning of the adsorption reaction, the concentration of As in As-SOD-TBC (1.533 mg L À1 ) was higher than that in As-TBA-TBC and As-TBC, and the arsenic removal rate in As-SOD-TBC was 84.66%. Then, the arsenic concentration in As-SOD-TBC gradually decreased to 0.122 mg L À1 . The arsenic removal rate was 98.77% within 4 hours.
As shown in Fig. 11(b), the oxidation of As(III) in solution of three reaction groups was observed. In As-TBA-TBC and As-TBC, As(III) was completely converted to As(V) in 2 hours. However, it took three hours in As-SOD-TBC. During the 4 hour adsorption reaction, the concentrations of As and As(III) in As-SOD-TBC were always higher than those in the other two groups. With increasing reaction time, the inhibition of trivalent arsenic oxidation became weak due to the instability and inactivation of SOD. The results showed that SOD obviously weakened the oxidation and adsorption capacity of As(III) onto TBC. This implied that O 2 c À signicantly affected the oxidation and adsorption of As(III) by TBC. Under UV and visible light illumination, TiO 2 photocatalyst absorbs photons with energy equal to or higher than its band gap energy, leading to the formation of electron (conduction band) and hole (valence band) pairs. 19,48 O 2 acts as an electron accepter in aerated aqueous suspensions, leading to superoxide anion radicals (O 2 c À ions), which oxidizes As(III) to As(IV). 19,[48][49][50] To further explore the effect of electron acceptor O 2 on photocatalytic oxidation of As(III), a N 2 stripping experiment was carried out. The experimental results were shown in Fig. 12. Aer the rst ve minutes of the reaction, the concentration of arsenic and the arsenic removal rate were 1.304 mg L À1 and 86.96%, 0.921 mg L À1 and 90.78%, respectively, with or without continuous N 2 purging. Then, the concentration of arsenic gradually decreased to 0.364 mg L À1 and 0.352 mg L À1 , while  the arsenic removal rate increased to 96.36% and 96.47% with or without N 2 purging, respectively. Overall, the concentration of arsenic with N 2 purging was higher, while the arsenic removal rate was lower than that without N 2 purging. Since the contact surface between the solution and atmospheric environment increased during N 2 stripping, the reaction system with nitrogen purging was not completely oxygen-free. This led to the oxidation of As(III) in solution with N 2 purging. Within one hour of the adsorption reaction, the concentration of As(III) with N 2 purging was slightly lower than that without N 2 purging. In the latter reaction, there was no difference in the concentration of As(III) between solutions with or without N 2 stripping. The reduction of oxygen content in the solution caused by N 2 purging slowed down the oxidation of As(III) to a certain extent and affected the removal percentage of arsenic and the adsorption performance of TBC for arsenic. The experimental results of scavenging free radicals and N 2 purging strongly indicated that O 2 c À dominated As(III) photooxidation on TiO 2 . 19,49,50 The oxidation of As(III) obviously affected the removal percentage of arsenic and the adsorption performance of TBC for arsenic. 3.7.3 XPS. XPS was used to analyse the surface chemical compositions of the adsorbents. Specically, Fig. 13(a) showed that characteristic peaks of C 1s, O 1s and Ti 2p appeared at 284.68 eV, 531.08 eV and 458.98 eV in TBC, respectively. Aer adsorption, a new peak of As 3d appeared at a binding energy of 44.98 eV. In the high-resolution XPS spectra of Ti 2p, peaks at approximately 458.98 eV and 464.68 eV belonged to Ti 2p 3/2 and Ti 2p 1/2 of TiO 2 , respectively, as shown in Fig. 13(b). The binding energy difference (E B ) between Ti 2p 3/2 and Ti 2p 1/2 of TiO 2 was 5.7 eV, meeting the requirement of the valence state of Ti 4+ in Fig. 11 Effect of scavengers on oxidation and adsorption of As(III). (a) Concentration and removal percentage of arsenic versus time in As-SOD-TBC, As-TBA-TBC and As-TBC; (b) concentration of As(III) and As(V) versus time in As-SOD-TBC, As-TBA-TBC and As-TBC. As-SOD-TBC: As(III) ¼ 10 mg L À1 , SOD ¼ 500-1500 units per mL, TBC ¼ 2 mg mL À1 ; As-TBA-TBC: As(III) ¼ 10 mg L À1 , t-butanol alcohol (TBA) ¼ 100 mg L À1 , TBC ¼ 2 mg mL À1 ; As-TBC: As(III) ¼ 10 mg L À1 As, TBC ¼ 2 mg mL À1 . Fig. 12 Changes of the adsorption capacity and oxidation of As(III) by TBC (a) with or (b) without continuous nitrogen purging conditions. TiO 2 . 51 The results further conrmed the existence of Ti as Ti 4+ (TiO 2 ).
The strong carbon signal was mainly due to the presence of BC. As shown in Fig. 13(c), four peaks of carbon species were observed in the C 1s spectrum before adsorption. The peaks at binding energies of 284.14 eV and 284.69 eV corresponded to C-C and C-C/C-H, 51,52 respectively. The peak at a binding energy of 285.45 eV was assigned to C-O or C]O. 26,51 The peak at a binding energy of 288.85 eV was likely a signal for C in the lattice of TiO 2 (Ti-O-C). 26,51,53 Aer adsorption, four peaks of carbon species also appeared, as shown in Fig. 13(d). In Fig. 13 26,51,52 and Ti-O-C, 26 respectively. As shown in Fig. 13(f), a new peak at 531.47 eV ascribed to As-O appeared aer adsorption. 6 The decrease in Ti-OH area from 23.45% to 21.57.8% indicated that the hydroxyl groups on Ti-OH were involved in the adsorption of arsenic. Aer arsenic adsorption, new peaks at 44.98 eV binding energies appeared in TBC. The As 3d spectrum was decomposed into two individual component peaks, as shown in Fig. 13(g). The As 3d peaks for As(III) and As(V) responded to binding energy ranges of 44.24 eV and 45.11 eV, respectively. 5,6,34,39 As 3d spectra conrmed that oxidation of As(III) occurred.
These reactions caused a signicant drop in solution pH. Under visible light illumination, the TiO 2 photocatalyst absorbed photons with energy equal to or higher than its band gap energy, leading to the formation of electron (conduction band) and hole (valence band) pairs (eqn (16)). 48 O 2 acted as an electron accepter in aerated aqueous suspensions and was converted to a superoxide anion radical (O 2 c À ), which transformed As(III) into the intermediate product of As(IV) and then into less toxic As(V). 19

Conclusions
In this study, TiO 2 -modied biochar produced by waste Chinese traditional medicine dregs (TBC) which was a friendly and inexpensive biosorbent showed efficient capacity in As(III) removal. Due to loading with TiO 2 , the microstructure and functional groups of the adsorbent obviously changed and the adsorption capability for As(III) removal reached up to 58.456 mg g À1 . This excellent adsorption performance may be attributed to multiple roles for As(III) removal, including direct adsorption and photocatalytic oxidation adsorption. There was little effect on the adsorption capacity among a broad pH range at an initial concentration of 10 mg L À1 As(III). During and aer As(III) adsorption on TBC, most As(III) was converted to less toxic As(V). The results of free radical scavenging and N 2 purging experiments strongly indicated that O 2 played a role as an electron accepter in aerated aqueous suspensions and O 2 c À dominated As(III) photooxidation. The oxidation of trivalent arsenic obviously affected the adsorption capacity of TBC for As(III).