Simultaneous Removal of Hg(II) and Phenol Using Functionalized Activated Carbon Derived from Areca Nut Waste

Areca nut waste was utilized to obtain high surface area activated carbon (AC), and it was further functionalized with succinic anhydride under microwave irradiation. The surface morphology and surface functional groups of the materials were discussed with the help of scanning electron microscope(SEM) images and fourier transform infra-red (FT-IR) analysis. The specific surface area of the AC and functionalized-AC was obtained by the Brunauer-Emmett-Teller (BET) method, and found to be 367.303 and 308.032 m2/g, respectively. Batch experiments showed that higher pH favoured the removal of Hg(II), whereas the phenol removal was slightly affected by the changes in the solution pH. The kinetic data followed pseudo-first order kinetic model, and intra-particle diffusion played a significant role in the removal of both pollutants. The maximum sorption capacity of Hg(II) and phenol were evaluated using Langmuir adsorption isotherms, and found to be 11.23 and 5.37 mg/g, respectively. The removal of Hg(II) was significantly suppressed in the presence of chloride ions due to the formation of a HgCl2 species. The phenol was specifically adsorbed, forming the donor–acceptor complexes or π–π electron interactions at the surface of the solid. Further, a fixed-bed column study was conducted for both Hg(II) and phenol. The loading capacity of the column was estimated using the nonlinear Thomas equation, and found to be 2.49 and 2.70 mg/g, respectively. Therefore, the study showed that functionalized AC obtained from areca nut waste could be employed as a sustainable adsorbent for the simultaneous removal of Hg(II) and phenol from polluted water.


Introduction
Water pollution is a serious environmental concern, and is growing rapidly due to the extreme use of harmful substances/chemicals by various industries. The wastewaters produced from manufacturing processes and petroleum refineries usually contain heavy metals and other toxic organic pollutants [1,2]. The partly treated or untreated effluents from these industries greatly contaminate the receiving water bodies. Heavy metals and phenolic compounds, which are the most common water pollutants, are found to be highly toxic towards living organisms and are difficult to eliminate from water bodies or even from the biosystem once entered [3][4][5]. Mercury has received greater environmental concern due to its high toxicity, persistence in the environment, and high bioaccumulation [6,7]. Dissolved Hg(II) in an aquatic environment is readily transformed into methylmercury, which is a highly toxic, persistent, and bioaccumulative form of mercury present in assessed for its capability to remove Hg(II) and phenol simultaneously from aqueous solutions under batch and fixed-bed adsorption systems.

Materials
The areca nut waste was collected from Aizawl city, Mizoram, India. It was repeatedly washed with pure water to remove any impurity, and then dried in an oven (Chang Shin Scientific Co., Busan, South Korea) (at 80 • C. Dry areca nut waste was soaked in concentrated H 2 SO 4 for 2 h at the temperature of 120 • C. The solid biomass was then completely transformed into its carbon form. The obtained carbon was carefully washed with pure water to eliminate excess acids, and dried at 70 • C. Any excess acid remaining within the carbon solid was then titrated with ammonia solution for neutralization. The carbon was again washed with distilled water several times until the pH reached 7.0, and then finally dried at 70 • C. The solid sample was cooled and then ground to obtain fine powder. Furthermore, the carbon powder was activated in a muffle furnace (Chang Shin Scientific Co., Busan, South Korea) at 500 • C for 4 h under an N 2 atmosphere. The activated carbon obtained was then kept inside a close plastic container for further use.

Modification of Activated Carbon
The functionalization of the activated carbon was carried out as follows: 5.0 g of succinic anhydride was dissolved in 100 mL of acetone. Ten grams (10 g) of activated carbon obtained from areca nut waste was added to this succinic anhydride solution. The mixture was kept under microwave irradiation at 60 • C for 60 min with constant stirring (Model: MAS II, Sineo Microwave Chemistry Technology Co., Ltd., Shanghai, China). The functionalized activated carbon was separated and washed with distilled water three times and then dried completely at 80 • C. The functionalized-AC was used for the simultaneous removal of Hg(II) and phenol under batch and fixed-bed column systems.

Characterization of Materials
The surface morphology of the AC before and after functionalization was observed by the SEM (scanning electron microscope) analysis (FE-SEM-Model: SU-70, Hitachi, Tokyo, Japan). The specific surface area of the activated carbon before and after functionalization was measured by the Brunauer-Emmett-Teller (BET) surface area analyzer (Model: ASAP 2020, Protech Korea Co. Ltd, Seoul, South Korea). Moreover, Fourier Transform Infra-Red (FT-IR) spectroscopy (Tensor 27, Bruker Co., Billerica, MA, USA) was used for the identification of various functional groups present in the AC and functionalized-AC solids.

Batch Adsorption Experiment
Batch experiments were performed to study the effect of pH, initial Hg(II)/phenol concentrations, contact time, and background electrolyte concentrations on the removal of Hg(II) and phenol using functionalized-AC. The experiments were performed using 0.10 g of functionalized-AC taken into 50 mL of Hg(II)/phenol solution and kept in an automatic shaker (Kukje Machinery Co., Ltd., Okcheon, South Korea) for 12 h at 25 • C. The solution was filtered using a 0.45 µm syringe filter, and the final concentration of Hg(II) was analyzed by a fast sequential atomic absorption spectrometer (AAS) (Model: AA240FS, Varian, Australia), whereas the phenol concentration was measured using a UV-visible spectrophotometer (Model: Humas HS 3300, Humas Co., Ltd., Daejeon, South Korea). The pH of the solutions was adjusted by adding drops of 1 M HCl or 1 M NaOH.

Fixed-Bed Column Adsorption
A glass column with a 1 cm inner diameter was used for performing a fixed-bed column experiments. Accurately 1.0 g of functionalized-AC was firmly packed in the middle of the column using glass beads. The aqueous solution containing both Hg(II) and phenol was pumped upward by a high-pressure liquid chromatographic pump (Acuflow Series II, Dongwoo Science Co., Ltd., Hanam, South Korea), and the flow rate was maintained at 0.5 mL/min. The effluent samples were collected using a Spectra/Chrom CF-2 fraction collector (Model: Retriever 500, Teledyne ISCO, Lincoln, NE, USA), and then filtered with a 0.45 µm syringe filter. The concentration of Hg(II) and phenol in the effluent was measured using an AAS and a UV-vis spectrophotometer, respectively.

Characterization of Materials
The SEM micrographs of the areca nut waste AC and functionalized-AC are shown in Figure 1. The macropores present in the AC solids are clearly visible, and the pore diameters are found to be in the range of 5-15 µm as obtained from the SEM images. Further, the smaller pores were also visible in the SEM image of the AC. Figure 1b shows an SEM image of the functionalized-AC. It is observed that the incorporation of succinic anhydride significantly changed the surface morphology of the AC, and the pores are less visible after the functionalization of the AC. A previous report also had shown that the surface morphology of maize straw was significantly changed after the incorporation of succinic anhydride [35]. The nitrogen adsorption isotherms of the AC and functionalized-AC are shown in Figure 2. According to International Union of Pure and Applied Chemistry (IUPAC) classification, the two solid samples display a H4 type hysteresis loop [39]. The specific surface area of the AC was found to be 367.30 m 2 /g. The surface area slightly decreased to 308.03 m 2 /g after the incorporation of succinic anhydride. This decrease in specific surface area is ascribed due to the accumulation of succinic anhydride within the pores of the AC. Moreover, the incorporation of succinic anhydride within the pores of AC caused a slight decrease in the pore size and pore volume of the AC, from 3.13 to 2.78 nm and from 0.028 to 0.016 cm 3 /g, respectively.
Metals 2017, 7, 248 4 of 14 collected using a Spectra/Chrom CF-2 fraction collector (Model: Retriever 500, Teledyne ISCO, Lincoln, NE, USA), and then filtered with a 0.45 µm syringe filter. The concentration of Hg(II) and phenol in the effluent was measured using an AAS and a UV-vis spectrophotometer, respectively.

Characterization of Materials
The SEM micrographs of the areca nut waste AC and functionalized-AC are shown in Figure 1. The macropores present in the AC solids are clearly visible, and the pore diameters are found to be in the range of 5-15 µm as obtained from the SEM images. Further, the smaller pores were also visible in the SEM image of the AC. Figure 1b shows an SEM image of the functionalized-AC. It is observed that the incorporation of succinic anhydride significantly changed the surface morphology of the AC, and the pores are less visible after the functionalization of the AC. A previous report also had shown that the surface morphology of maize straw was significantly changed after the incorporation of succinic anhydride [35]. The nitrogen adsorption isotherms of the AC and functionalized-AC are shown in Figure 2. According to International Union of Pure and Applied Chemistry (IUPAC) classification, the two solid samples display a H4 type hysteresis loop [39]. The specific surface area of the AC was found to be 367.30 m 2 /g. The surface area slightly decreased to 308.03 m 2 /g after the incorporation of succinic anhydride. This decrease in specific surface area is ascribed due to the accumulation of succinic anhydride within the pores of the AC. Moreover, the incorporation of succinic anhydride within the pores of AC caused a slight decrease in the pore size and pore volume of the AC, from 3.13 to 2.78 nm and from 0.028 to 0.016 cm 3 /g, respectively. The FT-IR analytical results are graphically shown in Figure 3. The predominant peak that appeared at around 3434 cm −1 is attributed to the stretching vibration of -OH groups. Several weak peaks obtained at 2856, 2925, and 2972 cm −1 are attributed to the presence of aldehyde groups and assigned to the stretching and scissoring of C-H bonds. Moreover, the bands that occurred at 1620 and 1700 cm −1 are assigned to the C=C and C=O groups [40]. Further, the vibration peaks in the region of 1042 cm −1 are due to the C-O-C stretching [36]. The weak peaks that occurred at around 1458 cm −1 showed the C-H vibration in -CH2-deformation. Moreover, other vibration bands that occurred at 874 and 740 cm −1 are due to an external bending of -C-H for various substituted benzene rings [41]. It is worth mentioning that the intensity of the peak occurred at 1700 cm −1 (it appeared as a shoulder), which indicates that the stretching vibration of the C=O group is significantly increased after the functionalization of AC. These changes eventually indicate the successful incorporation of succinic anhydride into the AC solids [42]. Previous studies have shown that a new distinguished peak was The FT-IR analytical results are graphically shown in Figure 3. The predominant peak that appeared at around 3434 cm −1 is attributed to the stretching vibration of -OH groups. Several weak peaks obtained at 2856, 2925, and 2972 cm −1 are attributed to the presence of aldehyde groups and assigned to the stretching and scissoring of C-H bonds. Moreover, the bands that occurred at 1620 and 1700 cm −1 are assigned to the C=C and C=O groups [40]. Further, the vibration peaks in the region of 1042 cm −1 are due to the C-O-C stretching [36]. The weak peaks that occurred at around 1458 cm −1 showed the C-H vibration in -CH 2 -deformation. Moreover, other vibration bands that occurred at 874 and 740 cm −1 are due to an external bending of -C-H for various substituted benzene rings [41].
It is worth mentioning that the intensity of the peak occurred at 1700 cm −1 (it appeared as a shoulder), which indicates that the stretching vibration of the C=O group is significantly increased after the functionalization of AC. These changes eventually indicate the successful incorporation of succinic anhydride into the AC solids [42]. Previous studies have shown that a new distinguished peak was observed at around 1730 cm −1 after the incorporation of succinic anhydride into posidonia and maize straw [35,43]. However, in this study, a new peak overlapped with the pre-existing peaks and caused an increase in the peak intensity. Furthermore, the decrease in the intensities of the absorption band at 1042 cm −1 assigned for the C-O-C bond stretching suggests the succinylation of the AC [35,36]. observed at around 1730 cm −1 after the incorporation of succinic anhydride into posidonia and maize straw [35,43]. However, in this study, a new peak overlapped with the pre-existing peaks and caused an increase in the peak intensity. Furthermore, the decrease in the intensities of the absorption band at 1042 cm −1 assigned for the C-O-C bond stretching suggests the succinylation of the AC [35,36].

Batch Adsorption Experiments
3.2.1. Effect of pH on Hg(II)/Phenol Adsorption pH is a crucial parameter influencing the adsorption of Hg(II) and phenol from aqueous solutions, since it determines the speciation of Hg(II) and/or the ionization of phenol as well as the surface charge of the adsorbents [44,45]. Therefore, the effect of pH was studied between pH 2.0 and 8.0 using a solution containing 5.4 mg/L of Hg(II) and 9.36 mg/L of phenol. The percentage removal of Hg(II) and phenol as a function of initial pH is shown in Figure 4. The pH dependence study showed that Hg(II) removal was significantly increased while increasing the solution pH from 2.0 to 4.0, and remained almost unchanged beyond pH 4.0. At low pH, the percentage sorption of Hg(II) was relatively low, due to the competition between Hg 2+ and H + ions towards the same active sites on the functionalized-AC [38]. At higher pH values, the H + ions concentration is decreased, and observed at around 1730 cm −1 after the incorporation of succinic anhydride into posidonia and maize straw [35,43]. However, in this study, a new peak overlapped with the pre-existing peaks and caused an increase in the peak intensity. Furthermore, the decrease in the intensities of the absorption band at 1042 cm −1 assigned for the C-O-C bond stretching suggests the succinylation of the AC [35,36].

Effect of pH on Hg(II)/Phenol Adsorption
pH is a crucial parameter influencing the adsorption of Hg(II) and phenol from aqueous solutions, since it determines the speciation of Hg(II) and/or the ionization of phenol as well as the surface charge of the adsorbents [44,45]. Therefore, the effect of pH was studied between pH 2.0 and 8.0 using a solution containing 5.4 mg/L of Hg(II) and 9.36 mg/L of phenol. The percentage removal of Hg(II) and phenol as a function of initial pH is shown in Figure 4. The pH dependence study showed that Hg(II) removal was significantly increased while increasing the solution pH from 2.0 to 4.0, and remained almost unchanged beyond pH 4.0. At low pH, the percentage sorption of Hg(II) was relatively low, due to the competition between Hg 2+ and H + ions towards the same active sites on the functionalized-AC [38]. At higher pH values, the H + ions concentration is decreased, and

Effect of pH on Hg(II)/Phenol Adsorption
pH is a crucial parameter influencing the adsorption of Hg(II) and phenol from aqueous solutions, since it determines the speciation of Hg(II) and/or the ionization of phenol as well as the surface charge of the adsorbents [44,45]. Therefore, the effect of pH was studied between pH 2.0 and 8.0 using a solution containing 5.4 mg/L of Hg(II) and 9.36 mg/L of phenol. The percentage removal of Hg(II) and phenol as a function of initial pH is shown in Figure 4. The pH dependence study showed that Hg(II) removal was significantly increased while increasing the solution pH from 2.0 to 4.0, and remained almost unchanged beyond pH 4.0. At low pH, the percentage sorption of Hg(II) was relatively low, due to the competition between Hg 2+ and H + ions towards the same active sites on the functionalized-AC [38]. At higher pH values, the H + ions concentration is decreased, and consequently enhances the uptake of Hg(II) ions from aqueous solutions. A previous report also showed that Hg(II) removal using activated biocarbon was relatively low at lower pH values and increased under higher pH conditions [46]. Conversely, phenol removal was slightly affected by changing the solution pH from 2.0 to 8.0. It was reported that phenol predominantly existed in its undissociated form within the pH region between 2.0 and 8.0; therefore, the adsorption of phenol was insignificantly influenced by the solution pH [3]. A similar trend was previously observed in the removal of phenol as a function of pH using activated carbon fibers and surface functionalized activated sericite [47,48]. consequently enhances the uptake of Hg(II) ions from aqueous solutions. A previous report also showed that Hg(II) removal using activated biocarbon was relatively low at lower pH values and increased under higher pH conditions [46]. Conversely, phenol removal was slightly affected by changing the solution pH from 2.0 to 8.0. It was reported that phenol predominantly existed in its undissociated form within the pH region between 2.0 and 8.0; therefore, the adsorption of phenol was insignificantly influenced by the solution pH [3]. A similar trend was previously observed in the removal of phenol as a function of pH using activated carbon fibers and surface functionalized activated sericite [47,48].

Adsorption Kinetics of Hg(II) and Phenol
In order to obtain the adsorption equilibrium time of Hg(II) and phenol by functionalized-AC, the extent of the removal at various intervals of time was analyzed, while keeping a constant solution pH, and the initial concentration of Hg(II) and phenol at ~10 mg/L. The percentage removal of Hg(II)/phenol at various intervals of time is given in Figure 5. The two pollutants exhibit similar sorption kinetics, which arerelatively slow. The percentage adsorption of Hg(II) and phenol continuously increases, and adsorption equilibrium is achieved within 300 min of contact time for both Hg(II) and phenol.

Adsorption Kinetics of Hg(II) and Phenol
In order to obtain the adsorption equilibrium time of Hg(II) and phenol by functionalized-AC, the extent of the removal at various intervals of time was analyzed, while keeping a constant solution pH, and the initial concentration of Hg(II) and phenol at~10 mg/L. The percentage removal of Hg(II)/phenol at various intervals of time is given in Figure 5. The two pollutants exhibit similar sorption kinetics, which arerelatively slow. The percentage adsorption of Hg(II) and phenol continuously increases, and adsorption equilibrium is achieved within 300 min of contact time for both Hg(II) and phenol. consequently enhances the uptake of Hg(II) ions from aqueous solutions. A previous report also showed that Hg(II) removal using activated biocarbon was relatively low at lower pH values and increased under higher pH conditions [46]. Conversely, phenol removal was slightly affected by changing the solution pH from 2.0 to 8.0. It was reported that phenol predominantly existed in its undissociated form within the pH region between 2.0 and 8.0; therefore, the adsorption of phenol was insignificantly influenced by the solution pH [3]. A similar trend was previously observed in the removal of phenol as a function of pH using activated carbon fibers and surface functionalized activated sericite [47,48].

Adsorption Kinetics of Hg(II) and Phenol
In order to obtain the adsorption equilibrium time of Hg(II) and phenol by functionalized-AC, the extent of the removal at various intervals of time was analyzed, while keeping a constant solution pH, and the initial concentration of Hg(II) and phenol at ~10 mg/L. The percentage removal of Hg(II)/phenol at various intervals of time is given in Figure 5. The two pollutants exhibit similar sorption kinetics, which arerelatively slow. The percentage adsorption of Hg(II) and phenol continuously increases, and adsorption equilibrium is achieved within 300 min of contact time for both Hg(II) and phenol.  The pseudo-first order kinetic (Equation (1)) and pseudo-second order kinetic (Equation (2)) models were employed to perform the kinetic modelling using the time dependent adsorption data of Hg(II) and phenol by the functionalized-AC. The equations were taken in their linear form: where q e is the maximum amount of Hg(II)/phenol adsorbed at equilibrium (mg/g), and q t (mg/g) is the amount of the Hg(II) or phenol adsorbed at time t. k 1 (min −1 ) and k 2 (g/mg/min) are the adsorption rate constants of the pseudo-first order and the pseudo-second order models, respectively [49].
The unknown values such as q e , k 1 , and k 2 were calculated from a plot of ln(q e − q t ) versus t, and the plots of t q t against t, respectively, for the pseudo-first order kinetic and pseudo-second order kinetic models, and returned in Table 1. The R 2 values indicated that the adsorption kinetic data fitted well to the pseudo-second order kinetic model rather than the pseudo-first order kinetic model for both the pollutants.
In porous materials, intra-particle diffusion usually plays a crucial role in controlling the adsorption of various pollutants from aqueous solutions. Therefore, the time-dependent intra-particle diffusion rate of Hg(II)/phenol from the outer surface active sites to the inside pores of the functionalized-AC was evaluated using an intra-particle diffusion model (Equation (3)) developed by Weber and Morris [50]. The equation is taken as: where k id is the intra-particle diffusion rate constant (g/mg/min 1/2 ), and Z is the intercept. The graph was plotted for the amount of Hg(II)/phenol adsorbed at various contact times (q t ) against the square root of time (t 1/2 ) and illustrated in Figure 6. The intra-particle diffusion plots for Hg(II) and phenol showed a good linearity with a high R 2 values, which indicated that intra-particle diffusion contributed a vital role in the sorption of Hg(II) and phenol onto functionalized-AC. However, a minor divergence from the origin was observed in the case of phenol, which was perhaps due to the variation in the solutes movement during the initial and final steps of the adsorption process [51]. The intra-particle diffusion constants along with the R 2 values are summarized in Table 1. The pseudo-first order kinetic (Equation (1)) and pseudo-second order kinetic (Equation (2)) models were employed to perform the kinetic modelling using the time dependent adsorption data of Hg(II) and phenol by the functionalized-AC. The equations were taken in their linear form: where qe is the maximum amount of Hg(II)/phenol adsorbed at equilibrium (mg/g), and qt (mg/g) is the amount of the Hg(II) or phenol adsorbed at time t. k1 (min −1 ) and k2 (g/mg/min) are the adsorption rate constants of the pseudo-first order and the pseudo-second order models, respectively [49]. The unknown values such as qe, k1, and k2 were calculated from a plot of ln(qe − qt) versus t, and the plots of against t, respectively, for the pseudo-first order kinetic and pseudo-second order kinetic models, and returned in Table 1. The R 2 values indicated that the adsorption kinetic data fitted well to the pseudo-second order kinetic model rather than the pseudo-first order kinetic model for both the pollutants.
In porous materials, intra-particle diffusion usually plays a crucial role in controlling the adsorption of various pollutants from aqueous solutions. Therefore, the time-dependent intraparticle diffusion rate of Hg(II)/phenol from the outer surface active sites to the inside pores of the functionalized-AC was evaluated using an intra-particle diffusion model (Equation (3)) developed by Weber and Morris [50]. The equation is taken as: where kid is the intra-particle diffusion rate constant (g/mg/min 1/2 ), and Z is the intercept. The graph was plotted for the amount of Hg(II)/phenol adsorbed at various contact times (qt) against the square root of time (t 1/2 ) and illustrated in Figure 6. The intra-particle diffusion plots for Hg(II) and phenol showed a good linearity with a high R 2 values, which indicated that intra-particle diffusion contributed a vital role in the sorption of Hg(II) and phenol onto functionalized-AC. However, a minor divergence from the origin was observed in the case of phenol, which was perhaps due to the variation in the solutes movement during the initial and final steps of the adsorption process [51]. The intra-particle diffusion constants along with the R 2 values are summarized in Table 1.

Adsorption Isotherm Studies
The effect of initial Hg(II) concentration was studied, increasing the initial concentration from 2.0 to 25.0 mg/L in the presence of 10.0 mg/L phenol at pH 5.0. Similarly, the effect of initial phenol concentration was studied, increasing the initial concentration from 1.0 to 20.0 mg/L in presence of 10.0 mg/L Hg(II) at pH 5.0. The equilibrium stage sorption data obtained were plotted between the percentage removals against the initial bulk sorptive concentrations (mg/L) and graphically shown in Figure 7a,b. An enhanced percentage of adsorption was observed at lower concentration of solutes, especially in the case of phenol, and gradually reduced with increasing the initial sorptive concentration. The uptake amount of Hg(II) and phenol was found to be increased from 0.65 to 5.65 mg/g and 0.58 and 5.44 mg/g, respectively, while increasing the initial sorptive concentrations as mentioned above. It is interesting to observe that the percentage removal of phenol remained almost unchanged while increasing the initial concentration of Hg(II) from 2.0 to 25.0 mg/L (Figure 7a). Similarly, there was no significant effect on Hg(II) removal by increasing the phenol concentration from 1.0 to 20 mg/L (Figure 7a). Therefore, these studies infer that the functionalized-AC is able to remove Hg(II) and phenol simultaneously from water, and suggest the presence of different binding sites available for these two different types of pollutants [48]. Table 1. Rate constants (k1 and k2), sorption capacity (qe), and R 2 values obtained from the pseudo-first order and pseudo-second order kinetic models, along with the intra-particle diffusion constants obtained for the adsorption of Hg(II) and phenol by functionalized-AC.

Adsorption Isotherm Studies
The effect of initial Hg(II) concentration was studied, increasing the initial concentration from 2.0 to 25.0 mg/L in the presence of 10.0 mg/L phenol at pH 5.0. Similarly, the effect of initial phenol concentration was studied, increasing the initial concentration from 1.0 to 20.0 mg/L in presence of 10.0 mg/L Hg(II) at pH 5.0. The equilibrium stage sorption data obtained were plotted between the percentage removals against the initial bulk sorptive concentrations (mg/L) and graphically shown in Figure 7a,b. An enhanced percentage of adsorption was observed at lower concentration of solutes, especially in the case of phenol, and gradually reduced with increasing the initial sorptive concentration. The uptake amount of Hg(II) and phenol was found to be increased from 0.65 to 5.65 mg/g and 0.58 and 5.44 mg/g, respectively, while increasing the initial sorptive concentrations as mentioned above. It is interesting to observe that the percentage removal of phenol remained almost unchanged while increasing the initial concentration of Hg(II) from 2.0 to 25.0 mg/L (Figure 7a). Similarly, there was no significant effect on Hg(II) removal by increasing the phenol concentration from 1.0 to 20 mg/L (Figure 7a). Therefore, these studies infer that the functionalized-AC is able to remove Hg(II) and phenol simultaneously from water, and suggest the presence of different binding sites available for these two different types of pollutants [48]. The equilibrium state adsorption data obtained at various sorptive concentrations were further modeled using Langmuir (Equation (4)) and Freundlich (Equation (5)) adsorption isotherms. The adsorption models were taken to its nonlinear forms [52]: The equilibrium state adsorption data obtained at various sorptive concentrations were further modeled using Langmuir (Equation (4)) and Freundlich (Equation (5)) adsorption isotherms. The adsorption models were taken to its nonlinear forms [52]: where q e is the amount of Hg(II)/phenol adsorbed (mg/g) by functionalized-AC, and C e is the bulk Hg(II)/phenol concentration (mg/L) at equilibrium. K L is the Langmuir rate constant (L/g), and q m is the Langmuir monolayer sorption capacity (mg/g). K F is the Freundlich constant, which corresponds to the adsorption capacity (mg/g), and 1/n is the sorption intensity. The nonlinear plot of the Langmuir and Freundlich adsorption isotherms for the removal of Hg(II) and phenol are shown in Figure 8a,b, respectively. Moreover, the adsorption capacity of functionalized-AC for Hg(II) and phenol along with the isotherm constants values are given in Table 2. The adsorption of Hg(II) onto functionalized-AC at various concentrations followed the Freundlich isotherm rather than the Langmuir model, whereas the experimental data obtained for phenol adsorption on functionalized-AC were reasonably fitted well to both the Langmuir and Freundlich models. Furthermore, the adsorption isotherm studies were conducted for the bare activated carbon under the identical experimental conditions. Langmuir and Freundlich adsorption isotherm studies were conducted for the adsorption of Hg(II) and phenol by the activated carbon obtained from areca nut waste, and the results are shown in Figure 8a,b. The isotherm constants values are summarized in Table 2. It is observed that the Langmuir monolayer adsorption capacity for Hg(II) was doubled after the functionalization with succinic anhydride. Moreover, the functionalization of the AC caused an increase in the adsorption capacity of phenol from 3.846 to 5.555 mg/g. where qe is the amount of Hg(II)/phenol adsorbed (mg/g) by functionalized-AC, and Ce is the bulk Hg(II)/phenol concentration (mg/L) at equilibrium. KL is the Langmuir rate constant (L/g), and qm is the Langmuir monolayer sorption capacity (mg/g). KF is the Freundlich constant, which corresponds to the adsorption capacity (mg/g), and 1/n is the sorption intensity. The nonlinear plot of the Langmuir and Freundlich adsorption isotherms for the removal of Hg(II) and phenol are shown in Figure 8a,b, respectively. Moreover, the adsorption capacity of functionalized-AC for Hg(II) and phenol along with the isotherm constants values are given in Table 2. The adsorption of Hg(II) onto functionalized-AC at various concentrations followed the Freundlich isotherm rather than the Langmuir model, whereas the experimental data obtained for phenol adsorption on functionalized-AC were reasonably fitted well to both the Langmuir and Freundlich models. Furthermore, the adsorption isotherm studies were conducted for the bare activated carbon under the identical experimental conditions. Langmuir and Freundlich adsorption isotherm studies were conducted for the adsorption of Hg(II) and phenol by the activated carbon obtained from areca nut waste, and the results are shown in Figure 8a,b. The isotherm constants values are summarized in Table 2. It is observed that the Langmuir monolayer adsorption capacity for Hg(II) was doubled after the functionalization with succinic anhydride. Moreover, the functionalization of the AC caused an increase in the adsorption capacity of phenol from 3.846 to 5.555 mg/g.

Effect of Background Electrolyte Concentrations
The effect of background electrolyte concentrations is a significant parameter to study the binding nature of pollutants onto the solid materials, and it is usually exploited to make a distinction between non-specific and specific adsorption. Specific sorption is unaffected by varying the background electrolyte concentration, whereas non-specific adsorption is significantly affected by increasing the background electrolyte concentrations [53,54]. In order to study the adsorption nature of Hg(II) and phenol onto functionalized-AC, an effect of background electrolyte concentration study was conducted, using NaCl as a background electrolyte. Similarly, the removal behaviour of Hg(II) was studied in the presence of NaNO3. The concentration of Hg(II) and phenol was maintained at

Effect of Background Electrolyte Concentrations
The effect of background electrolyte concentrations is a significant parameter to study the binding nature of pollutants onto the solid materials, and it is usually exploited to make a distinction between non-specific and specific adsorption. Specific sorption is unaffected by varying the background electrolyte concentration, whereas non-specific adsorption is significantly affected by increasing the background electrolyte concentrations [53,54]. In order to study the adsorption nature of Hg(II) and phenol onto functionalized-AC, an effect of background electrolyte concentration study was conducted, using NaCl as a background electrolyte. Similarly, the removal behaviour of Hg(II) was studied in the presence of NaNO 3 . The concentration of Hg(II) and phenol was maintained at approximately 5 mg/L at pH 5.0, and the concentration of NaCl and NaNO 3 was increased from 0.001 to 0.1 mol/L. The percentage removal of mercury and phenol as a function of background electrolytes is shown in Figure 9. It was observed that an increase in NaCl concentrations from 0.001 to 0.1 mol/L caused a considerable decrease in the uptake of Hg, while an increase in the concentration of NaNO 3 to the same extent did not affect the Hg(II) removal. A large decrease in mercury removal in the presence of NaCl was explained with the help of mercury speciation. In aqueous solution, Hg(II) forms various complexes with the chlorides, i.e., HgCl + , HgCl 2 , HgCl 3 − , and HgCl 4 2− , along with the mixed species Hg(OH)Cl. In the pH region between 3.5 and 5.5, the HgCl 2 is a predominant species (i.e., over 80%) [18,19]. Meanwhile, the adsorption experiment was carried out at pH 5.0 in the presence of chloride, in which the predominant species is HgCl 2 . Therefore, the HgCl 2 species were not efficiently removed by the functionalized-AC, and caused a drastic decrease in the removal of mercury from the aqueous solutions. Alternatively, the uptake of Hg(II) remained almost constant while increasing the NaNO 3 concentrations from 0.001 to 0.1 mol/L. These results suggested that the Hg(II) ions were specifically adsorbed and formed 'inner sphere complexes' on the surface of functionalized-AC [53].
On the other hand, the percentage removal of phenol was remained constant while increasing the NaCl concentration from 0.001 to 0.1 mol/L. This result suggested that phenols were also specifically adsorbed through the formation of donor-acceptor complexes, in which the carbonyl group of the functionalized-AC is an electron donor and the aromatic rings of phenol are electron acceptors [55]. Moreover, phenols were trapped within the pores of the solid materials, and possibly forming a strong bond through the π-π electron interaction of the benzene ring in phenol and π-electrons present with the solid materials [56,57].
Metals 2017, 7, 248 10 of 14 approximately 5 mg/L at pH 5.0, and the concentration of NaCl and NaNO3 was increased from 0.001 to 0.1 mol/L. The percentage removal of mercury and phenol as a function of background electrolytes is shown in Figure 9. It was observed that an increase in NaCl concentrations from 0.001 to 0.1 mol/L caused a considerable decrease in the uptake of Hg, while an increase in the concentration of NaNO3 to the same extent did not affect the Hg(II) removal. A large decrease in mercury removal in the presence of NaCl was explained with the help of mercury speciation. In aqueous solution, Hg(II) forms various complexes with the chlorides, i.e., HgCl + , HgCl2, HgCl3 − , and HgCl4 2− , along with the mixed species Hg(OH)Cl. In the pH region between 3.5 and 5.5, the HgCl2 is a predominant species (i.e., over 80%) [18,19]. Meanwhile, the adsorption experiment was carried out at pH 5.0 in the presence of chloride, in which the predominant species is HgCl2. Therefore, the HgCl2 species were not efficiently removed by the functionalized-AC, and caused a drastic decrease in the removal of mercury from the aqueous solutions. Alternatively, the uptake of Hg(II) remained almost constant while increasing the NaNO3 concentrations from 0.001 to 0.1 mol/L. These results suggested that the Hg(II) ions were specifically adsorbed and formed 'inner sphere complexes' on the surface of functionalized-AC [53]. On the other hand, the percentage removal of phenol was remained constant while increasing the NaCl concentration from 0.001 to 0.1 mol/L. This result suggested that phenols were also specifically adsorbed through the formation of donor-acceptor complexes, in which the carbonyl group of the functionalized-AC is an electron donor and the aromatic rings of phenol are electron acceptors [55]. Moreover, phenols were trapped within the pores of the solid materials, and possibly forming a strong bond through the π-π electron interaction of the benzene ring in phenol and π-electrons present with the solid materials [56,57].

Fixed-Bed Column Adsorption
In addition to batch reactor experiments, a fixed-bed column adsorption study was conducted. In this study, the pH of the influent solution was maintained at pH 5.0, and the initial concentrations of Hg(II) and phenol was taken 4.5 and 7.89 mg/L, respectively. It was observed that the functionalized-AC could efficiently remove the Hg(II) and phenol even in a continuous flow system. The nonlinear Thomas equation (Equation (6)) was utilized to fit breakthrough data obtained for Hg(II) and phenol to optimize the loading capacity of functionalized-AC under dynamic conditions [58]. The equation is taken as:

Fixed-Bed Column Adsorption
In addition to batch reactor experiments, a fixed-bed column adsorption study was conducted. In this study, the pH of the influent solution was maintained at pH 5.0, and the initial concentrations of Hg(II) and phenol was taken 4.5 and 7.89 mg/L, respectively. It was observed that the functionalized-AC could efficiently remove the Hg(II) and phenol even in a continuous flow system. The nonlinear Thomas equation (Equation (6)) was utilized to fit breakthrough data obtained for Hg(II) and phenol to optimize the loading capacity of functionalized-AC under dynamic conditions [58]. The equation is taken as: where C e is the concentration of Hg(II)/phenol in the effluent (mg/L); C o is the initial concentration of Hg(II)/phenol in the feed solution (mg/L); K T is the Thomas rate constant (L/min/mg); q o is the loading capacity of Hg(II)/phenol (mg/g); m is the mass of functionalized-AC packed inside the column (g); V is the volume of water passed through the column (L); and Q is the influent flow rate (L/min). As shown in Figure 10, the column data were reasonably fitted well to the nonlinear Thomas equation, and the loading capacities (q o ) of Hg(II) and phenol were found to be 2.49 and 2.70 mg/g, respectively. Moreover, the Thomas rate constants (K T ) were calculated as 2.13 × 10 −3 and 1.09 × 10 −3 L/min/mg for Hg(II) and phenol, respectively. The loading capacities obtained from the fixed-bed column adsorption were comparatively lower than that obtained from the batch reactor studies. This is perhaps due to insufficient contact time provided for the solutes Hg(II)/phenol towards the functionalized-AC solid inside the column [59]. where Ce is the concentration of Hg(II)/phenol in the effluent (mg/L); Co is the initial concentration of Hg(II)/phenol in the feed solution (mg/L); KT is the Thomas rate constant (L/min/mg); qo is the loading capacity of Hg(II)/phenol (mg/g); m is the mass of functionalized-AC packed inside the column (g); V is the volume of water passed through the column (L); and Q is the influent flow rate (L/min). As shown in Figure 10, the column data were reasonably fitted well to the nonlinear Thomas equation, and the loading capacities (qo) of Hg(II) and phenol were found to be 2.49 and 2.70 mg/g, respectively. Moreover, the Thomas rate constants (KT) were calculated as 2.13 × 10 −3 and 1.09 × 10 −3 L/min/mg for Hg(II) and phenol, respectively. The loading capacities obtained from the fixed-bed column adsorption were comparatively lower than that obtained from the batch reactor studies. This is perhaps due to insufficient contact time provided for the solutes Hg(II)/phenol towards the functionalized-AC solid inside the column [59].

Conclusions
The succinic anhydride was successfully incorporated with the activated carbon obtained from areca nut wastes. The functionalized-AC was employed for the simultaneous removal of Hg(II) and phenol from aqueous solutions. The materials were characterized by the SEM, FT-IR, and BET analyses. The batch adsorption experiment showed that increasing the solution pH favored Hg(II) removal whereas phenol removal was not significantly affected by the variation of the solution pH. The kinetic studies showed that an apparent sorption equilibrium was achieved within 300 min of contact, and it was further observed that intra-particle diffusion contributed a significant role in the removal of Hg(II) and phenol. The equilibrium state sorption data for Hg(II) were well simulated with the Freundlich adsorption isotherm, whereas the sorption data obtained for phenol were reasonably fitted to both the Langmuir and Freundlich models. The maximum sorption capacity was evaluated using Langmuir adsorption isotherms and found to be 11.235 and 5.376 mg/g for Hg(II) and phenol, respectively. The removal of Hg(II) was significantly suppressed in the presence of NaCl due to the formation of HgCl2 species. On the other hand, the phenol removal was insignificantly changed with an increase in NaCl concentrations, which indicated that phenols were adsorbed specifically onto the functionalized-AC. Meanwhile, an increase in the concentration of another background electrolyte, i.e., NaNO3, from 0.001 to 0.1 mol/L did not affect the uptake of Hg(II), and it was assumed that the adsorbed Hg(II) formed 'inner sphere complexes' on the surface of the

Conclusions
The succinic anhydride was successfully incorporated with the activated carbon obtained from areca nut wastes. The functionalized-AC was employed for the simultaneous removal of Hg(II) and phenol from aqueous solutions. The materials were characterized by the SEM, FT-IR, and BET analyses. The batch adsorption experiment showed that increasing the solution pH favored Hg(II) removal whereas phenol removal was not significantly affected by the variation of the solution pH. The kinetic studies showed that an apparent sorption equilibrium was achieved within 300 min of contact, and it was further observed that intra-particle diffusion contributed a significant role in the removal of Hg(II) and phenol. The equilibrium state sorption data for Hg(II) were well simulated with the Freundlich adsorption isotherm, whereas the sorption data obtained for phenol were reasonably fitted to both the Langmuir and Freundlich models. The maximum sorption capacity was evaluated using Langmuir adsorption isotherms and found to be 11.235 and 5.376 mg/g for Hg(II) and phenol, respectively. The removal of Hg(II) was significantly suppressed in the presence of NaCl due to the formation of HgCl 2 species. On the other hand, the phenol removal was insignificantly changed with an increase in NaCl concentrations, which indicated that phenols were adsorbed specifically onto the functionalized-AC. Meanwhile, an increase in the concentration of another background electrolyte, i.e., NaNO 3 , from 0.001 to 0.1 mol/L did not affect the uptake of Hg(II), and it was assumed that the adsorbed Hg(II) formed 'inner sphere complexes' on the surface of the functionalized-AC. Further, the loading capacities of Hg(II) and phenol under dynamic conditions were evaluated using the nonlinear Thomas equation, and were found to be 2.49 and 2.70 mg/g, respectively. Therefore, the present study suggested that the low cost activated carbon obtained from areca nut waste and further functionalized with succinic anhydride under microwave irradiation could be effectively employed in the simultaneous removal of Hg(II) and phenol from the wastewater samples.