Surface-functionalized activated sericite for the simultaneous removal of cadmium and phenol from aqueous solutions: Mechanistic insights

Accepted Manuscript Surface-functionalized activated sericite for the simultaneous removal of cadmium and phenol from aqueous solutions: Mechanistic insights Lalhmunsiama, Diwakar Tiwari, Seung-Mok Lee PII: DOI: Reference:

S1385-8947(15)01154-7 http://dx.doi.org/10.1016/j.cej.2015.08.072 CEJ 14064

To appear in:

Chemical Engineering Journal

Received Date: Revised Date: Accepted Date:

16 June 2015 12 August 2015 13 August 2015

Please cite this article as: Lalhmunsiama, D. Tiwari, S-M. Lee, Surface-functionalized activated sericite for the simultaneous removal of cadmium and phenol from aqueous solutions: Mechanistic insights, Chemical Engineering Journal (2015), doi: http://dx.doi.org/10.1016/j.cej.2015.08.072

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Surface-functionalized activated sericite for the simultaneous removal of cadmium and phenol from aqueous solutions: Mechanistic insights

Lalhmunsiamaa, Diwakar Tiwarib, Seung-Mok Leea,*

a

Department of Environmental Engineering, Catholic Kwandong University, Gangneung, Gangwon-do 210-701, Republic of Korea

b

Department of Chemistry, School of Physical Sciences, Mizoram University, Aizawl–796004, Mizoram, India

*Corresponding author Telephone:

+82-33-649-7535

Fax:

+82-33-642-7635

E.Mail:

[email protected]

ABSTRACT The present communication addresses the development of a novel hybrid material precursor to natural sericite. The hybrid material is then successfully utilized for efficient removal of cadmium and phenol from aqueous solutions. Initially, sericite was annealed and activated with hydrochloric acid. The activation caused a significant increase in specific surface area of sericite, thereby provided a suitable surface structure for grafting of organosilanes.

The

activated

sericite

(AS)

was

functionalized

with

3-

aminopropyltriethoxysilane (APTES), and the resultant AS-APTES along with pristine sericite and AS were characterized using SEM-EDX, BET, XRD and FT-IR analyses. Batch reactor studies showed that increases in sorptive pH, contact time, initial concentration and temperature significantly favored the sorption of Cd(II), and a 100-fold increase in background electrolyte concentration did not significantly affect the uptake of Cd(II) or phenol from aqueous solutions. Cd(II) removal was found to be spontaneous and the uptake process was endothermic in nature. Further, the intra-particle diffusion was found to be the rate-limiting step in the sorption of Cd(II). Various physico-chemical parametric studies enabled to discuss the sorption mechanism of these contaminants at the solid/solution interface. In a single pollutant sorption studies, it was deduced that sorbed Cd(II) formed inner-sphere complexes, whereas phenol was sorbed through the hydrogen bonding with the amino groups or partitioned within the interspace region. Simultaneous sorption studies suggested that these two pollutants were possibly removed simultaneously by the prevailing hydrophilic and hydrophobic groups. Furthermore, desorption and reusability studies as well as the applicability of the material for real wastewater treatment demonstrated that ASAPTES is a promising solid material for the efficient removal of two important water pollutants i.e., Cd(II) and phenol from aqueous waste. Keywords: Sericite, organosilane, mesopore, Cd(II), phenol, intra-particle diffusion

1.

Introduction Wastewater produced from process industry generally contains inorganic pollutants and

other organic toxin [1, 2]. Heavy metals such as cadmium, copper, lead, chromium etc along with phenol and its derivatives are important pollutants generated in petroleum refinery [3]. The presence of these pollutants in the aquatic environment cause chemical, physical or biological changes to the quality of water and consumption of polluted water can cause detrimental effects on living organisms [4]. Cadmium (Cd) and phenol are the most deleterious pollutants due to their detrimental effects to living organisms even at trace level concentrations. Cadmium (Cd) is a toxic heavy metal commonly present in wastewater. It is highly toxic in any of its chemical forms, and ingestion of cadmium causes nausea, salivation, diarrhoea, muscular cramps, chronic pulmonary problems, skeletal deformity, hypertension, proteinuria, kidney stone formation and itai itai disease etc. in human body [5, 6, 7]. Moreover, Cd(II) ions can replace Zn(II) ions in some enzymes, thereby affecting the enzyme activity in biological systems. Cd is also classified as a category 1 carcinogen (human carcinogen) by the International Agency for Research on Cancer (IARC) [8]. Among organic pollutants, phenol and its derivatives are considered the most hazardous due to their adverse effects in living organisms [9]. Exposure to phenol-contaminated water is associated with protein degeneration, tissue erosion, paralysis of the central nervous system and damage to the kidney, liver and pancreas [10]. Therefore, waste water contaminated with Cd(II) and phenol requires proper and adequate treatment prior to its discharge into the ecosystem. Several techniques employed in the removal of Cd(II) from the wastewater include: chemical precipitation, coagulation/flocculation, ion exchange,

solvent extraction,

cementation, complexation, electrochemical operations, biological operations, adsorption, evaporation, filtration, and membrane processes [11]. Similarly, various methods such as distillation, liquid-liquid extraction, adsorption, pervaporation, and membrane extraction,

thermal oxidation, catalytic oxidation and photocatalytic degradation etc. were employed for the removal of phenol from aqueous solutions [12]. Among these techniques, adsorption has advantages over other technique for removing pollutants at low level concentrations due to their efficiency. Additionally, use of naturally abundant materials in sorption process is potentially suitable as it offers cost effective and environmentally benign treatment technologies [13]. Natural clay and its derivatives have recently received great interest due to their potential applications in diverse areas of research. A huge effort has been exerted to improve the efficiency of natural clay materials for waste water treatment through functionalization [14].

Clay is

most commonly modified

using

surfactant

molecules.

However,

inorgano/organo-clay possesses a certain drawback in that surfactants bind to the clay surface with a weak electrostatic interaction, sometimes resulting in desorption of intercalated surfactants and, hence, causes secondary pollution [15]. Recently, organosilane grafting has been demonstrated to be an efficient method for functionalizing clay minerals, as the formation of covalent bonding between the clay minerals and the organosilane results in stable immobilization of organic moieties and prevents their leaching into the surrounding environment [16]. The successful immobilization of silane is influenced by various factors including concentration of surface hydroxyl groups, type of surface hydroxyl groups, hydrolytic stability of the bonds formed and characteristics of silane. Moreover, substrate features such as broken edges in the clay play an important role in silylation. It is reported that silanes are readily intercalated into the interlayer spaces of various types of swelling clay, such as montmorillonite, whereas the successful grafting of silane with non-swelling clays requires more specific properties including a high reaction temperature and an inert atmosphere [17]. Several investigators have reported success when grafting organosilanes onto natural clays including bentonite [18], montmorillonite [19] and kaolinite [20].

Sericite is mica-based natural clay that is abundant in Korea, China and Japan. It is a mineral of fine-grained muscovite, whose chemical formula is KSi3Al3O10(OH)2 with a very high layer charge density close to ≤2 e per unit cell. Its unit cell is composed of 2:1 T:O phyllosilicate layers (where ‘T’ is tetrahedral [SiO4]4- and ‘O’ is octahedral [AlO3(OH)3]6-) [21, 22, 23]. The strong unit charge of sericite causes a strong electrostatic force to attract the aluminosilicate layers strongly along with the available cations within the interspace. This attraction caused muscovite a non-swollen type mineral in water and the cation exchange capacity as well as specific surface area of sericite is relatively low [21]. Therefore, the present study employed a novel activation method in order to increase the specific surface area of sericite which provides a suitable surface structure for the introduction of organosilane.

The

activated

sericite

was

further

functionalized

using

3-

aminopropyltriethoxysilane (APTES), and the functionalized activated sericite material was employed for efficient removal of cadmium and phenol from aqueous solutions.

2.

Materials and methods

2.1. Materials Sericite was collected from the Keumnam deposit, Gagokmyun, Kangwon province, Korea. Nitric acid, sodium hydroxide, hydrochloric acid and sodium chloride were obtained from Duksan Pure Chemicals Co. Ltd., Korea, and cadmium sulfate was obtained from Kanto Chemical Co. Inc., Japan. Phenol was obtained from Samchun Pure Chemical Co. Ltd., Korea, and 3-aminopropyltriethoxysilane (APTES) was procured from Sigma Aldrich Co., USA. The de-ionized water used for all the experiments was purified in the Millipore water purification system (Milli-Q+). 2.1.1. Preparation of activated sericite (AS) Sericite rock was crushed and sieved to obtain fine powder (0.075-0.053 mm) using a mechanical sieve. The chemical composition of the sericite is given elsewhere [24]. The

sericite was further annealed at 800 °C for 4 h and then activated using 3.0 mol/L HCl at 100 °C for 1 h with constant stirring. The slurry was filtered and washed several times with distilled water to remove excess acid. The AS sample was dried at 80 °C and stored in a plastic bottle. 2.1.2. Preparation of functionalized activated sericite The surface functionalization of AS was performed using a method similar to one described previously [19]. Briefly, 15 ml of APTES was added to a mixture of 150 ml ethanol and 50 ml distilled water. Approximately 15 g of AS was dispersed in the above solution and refluxed at 80 °C with continuous stirring for 7 h. The slurry was filtered and dried at 60 °C overnight. Subsequently, the dried material was washed with a mixture of water and ethanol using a Soxhlet extractor for 24 h to remove residual silanes. The slurry was collected, washed with distilled water, and completely dried in a hot air oven at 70 °C. The material was named AS-APTES. 2.2. Characterization The surface morphology and elemental composition of the samples were characterized using a scanning electron microscope (SEM; FE-SEM-Model: SU-70, Hitachi, Japan) equipped with an energy dispersive X-ray spectroscopy (EDX) system. The specific surface area of the samples was analyzed using a Protech Korea BET surface area analyzer (Model ASAP 2020). X-ray diffraction (XRD) data were recorded with an X-ray diffraction instrument (PANalytical, Netherland; Model X’Pert PRO MPD) using Cu Kα radiation at a wavelength of 1.5418 Å. Moreover, the functional groups present in the solid samples were identified using Fourier transform-infrared spectrometry (FT-IR) with the KBR disk method (Bruker, Tensor 27, USA). 2.3. Batch reactor studies Batch experiments were carried out to optimize the effects of pH, initial solute

concentration, contact time, temperature and background electrolyte concentration on the sorption of Cd(II) and phenol by AS-APTES. Moreover, simultaneous sorption of Cd(II) and phenol was studied under the optimized conditions. The solution pH was adjusted with 1.0 M HCl or 1.0 M NaOH for all the experiments. The solution mixture was equilibrated using an automatic shaker (KUKJE, Shaking Incubator, Korea, model 36-SIN-125) for 12 h at 25±1 °C (except effect of contact time and effect of temperature studies). The solution was then filtered with a 0.45-µm syringe filter; the bulk Cd concentration was analyzed using a Fast Sequential Atomic Absorption Spectrometer (Model: AA240FS, Varian), and the phenol concentration was measured using a UV-visible spectrophotometer (Model: Humas HS 3300). The results are presented as percentage of pollutants removed. 2.4. Desorption and reusability study Desorption experiment was carried out using hydrochloric acid (HCl) and the HCl concentration was varied between 0.01 to 0.3 mol/L at 25±1 oC.

Initially, batch reactor

experiment for simultaneous removal of Cd(II) and phenol was performed as described in section 2.3. The initial concentration of Cd(II) and phenol were taken as 10.8 and 7.16 mg/L respectively and pH was maintained to 5.0. After sorption process, the solid sample was recovered by centrifugation and it was suspended in 50.0 mL of hydrochloric acid solutions. The bottle was kept in an automatic incubator shaker for 2 h to complete desorption; again the suspension was centrifuged and the supernatant solution was analyzed by AAS for Cd(II) and UV-Vis spectrophotometer for phenol measurements. The regenerated material was washed several times with distilled water, dried in an oven and used for next adsorptiondesorption studies. 2.5. Applications to real water sample Furthermore, batch experiment was performed to remove Cd(II) and phenol simultaneously from the simulated aqueous waste prepared using tap water. The

concentration of Cd(II) and phenol in tap water was maintained at 10.47 and 7.45 mg/L, respectively and pH of tap water was 7.76. Sorption experiment was performed as describe in Section 2.3.

3.

Results and discussion

3.1. Characterization SEM micrographs of sericite, AS and AS-APTES are shown in Fig. 1. The SEM image shows that pristine sericite possessed a compact and ordered layered structure and the AS showed a heterogeneous surface structure with widely distributed mesopores on its surface. However, the introduction of APTES onto AS did not produce further significant changes in surface morphology. The elemental compositions of the materials were characterized with EDX, and the spectra are shown in Fig. S1 (supplementary data). The major elements present in the solid samples were carbon (C), oxygen (O), silicon (Si), aluminium (Al) and potassium (K). Sodium (Na) and magnesium (Mg) were also present in trace amounts [24]. It was noted from EDX analysis that the weight percentage of C increased from 6.74 to 10.88 % after grafting the organosilane onto AS, and that of Si increased from 18.86 to 19.65 %. The nitrogen (N2) adsorption-desorption isotherms of pristine sericite, AS and ASAPTES are given in Fig. S2 (supplementary data). A type IV isotherm with a characteristic hysteresis loop (H3 type) was shown by all three samples under a wide range of relative pressures, indicating the existence of mesopores in the materials [25, 26]. A very small hysteresis loop was observed for the N2 adsorption-desorption isotherm of pristine sericite; the isotherms for AS and AS-APTES were similar but traced a wide hysteresis loop at P/Po>4. The presence of this loop infers the development of mesopores on the surfaces of the sericite samples [27]. The BET specific surface area, pore volume and pore size obtained for sericite, AS and AS-APTES are summarized in Table 1. The pore size and pore volume of sericite were greatly increased after activation. Moreover, the activation caused a very large increase

in specific surface area from 3.65 to 62.92 m2/g. The incorporation of organosilane did not lead to significant changes in the textural properties of AS; however, slight decreases in pore size, pore volume and specific surface area were observed as the interlayer space of the AS was occupied by the grafted organosilane [28]. The X-ray diffraction data obtained for sericite, AS and AS-APTES are graphically displayed in Fig. 2. The disappearance of several diffraction peaks in AS and AS-APTES compared to sericite suggests lower crystallinity of the material after activation. It was interesting to note that a diffraction peak corresponding to basal spacing of the 001 plane was observed for sericite and AS at 2θ values of 8.81 and 8.63, respectively. This result inferred that the activation caused a significant increase in interlayer space of sericite [28]. However, no additional expansion of the interlayer was observed after incorporation of APTES onto AS [29]. The solid materials were further characterized using FT-IR analysis (Fig. 3). Absorption bands occurred at 3440 and 1630 cm-1 and were associated with –OH stretching and bending frequencies, respectively. The stretching band at 3600 cm-1 was attributed to the presence of a hydroxyl group within the layers of the solid sample [30, 31]. The peak obtained due to hydroxyl groups at 3600 cm-1 was strongly reduced after activation, and a weak broad peak at 3440 cm-1 due to -OH stretching was obtained for the AS sample; the size of this peak decreased significantly or even disappeared in samples of AS-APTES. These results illustrate the successful grafting of organosilanes onto the surface hydroxyl groups of AS [21]. In all samples, weak absorption bands obtained at 2920 and 2850 cm-1 corresponded to the CH3 asymmetric stretching mode and CH2 symmetric stretching mode, respectively [32]. A frequency band at 1060 cm-1 was assigned to asymmetric stretching vibrations of Si-O-Si, while the absorption bands at 830 and 720 cm-1 confirmed the presence of quartz. The bands observed at 560 and 480 cm-1 were ascribed to O-Si-O bending vibrations [21]. 3.2. Batch reactor operations

3.2.1. Effect of pH The effect of pH was studied between pH 2.0 to 10.0 with 9.99 mg/L Cd(II) and 7.88 mg/L phenol, and the results are shown in Fig. 4. Increasing the initial pH from 2.0 to 10.0 caused an increase in Cd(II) removal from 8.90 % to approximately 100 %. Moreover, it was noted that Cd(II) removal was particularly poor at a pH lower than 3. At low pH, the amino functional groups of AS-APTES were protonated due to the presence of excess hydrogen ions. Therefore, the metal ions experienced a repulsive force from the surface’s protonated amino groups, and the percentage of Cd(II) removed was limited. However, a gradual increase in solution pH cause decrease in H+ ion concentrations and there was greater electrostatic attraction of metal cations toward the negatively-charged surface of the ASAPTES [33, 34]. As the pH was higher than 5.0, the removal of Cd(II) from the solution correlated positively with pH and reached approximately 100 % removal at pH 10.0. In the aqueous solution, Cd started to precipitate as solid Cd(OH)2 beyond pH 8.5; therefore, the complete removal of Cd from very alkaline solutions might be a mixed effect of sorption and co-precipitation of the metal ions onto the solid surfaces of materials [6]. On the other hand, phenol removal slightly increased from pH 2.0 to 4.0 and remained almost constant between pH 4.0 and 10.0. Phenol mostly remains in undissociated form and exists as phenol molecule only in the pH region between 2 to 10; therefore, phenol removal was not significantly affected by the initial sorptive pH. A similar effect of sorptive pH on the removal of phenol was observed using activated carbon fibers and granular activated carbon [35, 36]. The sorption capacity for phenol was very low in comparison to Cd(II); and the maximum removal was found to be only approximately 35 %. 3.2.2. Effect of initial concentration The effect of initial sorptive concentration was studied by varying the initial concentration of Cd(II) from 2.0 mg/L to 30.0 mg/L and the initial concentration of phenol

from 2.0 mg/L to 15.0 mg/L at pH 5.0. The equilibrium sorption data obtained for Cd(II) at 298, 308 and 318 K and phenol at 298 and 318 K are displayed in Fig. 5. In general, both Cd(II) and phenol exhibited higher sorption percentages at lower starting concentrations. The percent removal gradually decreased since relatively fewer active sites were available as a result of increasing the initial concentration of Cd(II) or phenol for the same dose of the solid [37]. Indeed, increasing the initial concentrations of both Cd(II) and phenol increased the amount adsorbed of both pollutants. Moreover, the amount of Cd(II) removal significantly increased with increasing temperature; the amount of Cd(II) adsorbed was increased from 0.80 to 5.69 mg/g, 0.80 to 6.93 mg/g and 0.80 to 8.41 mg/g at 298, 308 and 318K, respectively, while increasing the initial concentration of Cd(II) from 1.60 to 30.09 mg/L. On increasing the initial concentration of phenol from 2.0 to 15.0 mg/L, the amount of phenol removal was also slightly increased from 1.17 to 1.62 mg/g. However, an increase in temperature did not show a significant effect on phenol removal. It is interesting that the ASAPTES showed a significantly higher sorption capacity for Cd(II) or phenol compared to the pristine sericite or organo-modified sericite reported previously [6, 10]. The concentration dependence data obtained at equilibrium between solid and solution interfaces were further analyzed by Langmuir (Eq.(1)), Freundlich (Eq. (2)) and Temkin (Eq. (3)) adsorption isotherm models [38]. The adsorption isotherm equations were utilized in their linear forms:


 1
 = + … (1)     log  =  =

1 log
 + log  … (2) 

   + 
 … (3)  

where q e is the amount of solute adsorbed per unit weight of adsorbent (mg/g); Ce the

equilibrium bulk concentration (mg/L); qo is the Langmuir monolayer adsorption capacity (mg/g) and KL is the Langmuir constant (L/g). KF and 1/n are the Freundlich constants referring to adsorption capacity and adsorption intensity respectively. BT and KT are Temkin constants related to the heat of adsorption (J/mol) and the equilibrium binding energy (L/mg), respectively. Graphs were plotted depicting Ce/qe versus Ce for the Langmuir model (Fig. S3, supplementary data), log ae versus log Ce for the Freundlich model (Fig. S4, supplementary data) and qe versus ln Ce for the Temkin adsorption isotherm (Fig. S5, supplementary data). Straight lines with fairly good fits were obtained for the three isotherms; the isotherm constants and the associated R2 values are shown in Table 2. The regression coefficients, i.e., the R2 values, indicate a good fit of the Langmuir and Freundlich adsorptions model for Cd(II). The equilibrium data obtained for phenol showed a considerably better fit to the Langmuir adsorption isotherm rather than Freundlich or Temkin isotherms. The fractional values of 1/n (0 <1/n< 1) obtained with the Freundlich isotherm for both Cd(II) and phenol suggest that the AS-APTES has a heterogeneous surface structure with an exponential distribution of active sites. Moreover, higher values of the Langmuir constant and relatively higher values of the Freundlich adsorption capacity further reflect the improved strength and affinity of AS-APTES for Cd(II) and phenol [33, 37]. Langmuir monolayer sorption capacity of Cd(II) and phenol was obtained for various low cost adsorbents along with the AS-APTES and was returned in Table 3a and 3b. The sorption capacity shown by various sorbents is varied due to difference in physico-chemical properties of adsorbents as well as experimental factors including the concentration range of pollutants (Cd(II) and phenol), pH, temperature, ionic strength etc. [52]. 3.2.2. Effect of contact time The effect of contact time on the removal of Cd(II) was studied at 298, 308 and 318 K

with an initial Cd(II) concentration of ~10 mg/L and a pH of 5.0. On the other hand, the effect of contact time on phenol removal was studied only at 298 K with an initial phenol concentration of 7.88 mg/L and a pH of 5.0. Fig. 6 shows that increasing the temperature accelerated the sorption rate of Cd(II), and the time to reach equilibrium was reduced with increased temperature. When the temperature increased from 298 to 308 K, the time needed for Cd(II) sorption to reach apparent equilibrium was reduced from 360 min to 300 min. Further, a quasi-equilibrium at 318 K was achieved at 180 min, which was half of the time required at 298 K. On the other hand, the removal of phenol was found to be very fast, and the equilibrium sorption was attained within a contact time of 30 min. This result confirmed that the active sites on the adsorbent surfaces were readily available to adsorb phenol from aqueous solution [53]. Pseudo-first-order (Eq. (4)) and pseudo-second-order (Eq. (5)) kinetic models were exploited to clarify the sorption kinetics of Cd(II) and phenol onto AS-APTES. The equations were used in linear form [54]. ( −  ) =   −

 " … (4) 2.303

" 1 " = + … (5) $  $   In Eq. 4 and 5, q e is the maximum sorption capacity (mg/g), qt (mg/g) is the amount adsorbed at time t, and k1 (1/min) and k2 (g/mg/min) are the adsorption rate constants of the pseudofirst order and pseudo-second order equations, respectively. The pseudo-first order plot of ln (q e–q t) versus t is shown in Fig. S6 (supplementary data), and the pseudo-second order kinetic model of t/qt versus t is shown in Fig. S7 (supplementary data). The kinetic data fit better to the pseudo-second order kinetic model for both Cd(II) and phenol. Sorption capacity (qe) and the rate constants (k1 and k2) were calculated and are shown in Table 4. The suitability of a

pseudo-second order kinetic model infers that there are strong chemical forces caused by sharing or exchanging electrons at the solid/solution interface [55]. 3.2.3. Intra-particle diffusion Sorption onto porous materials is controlled in part by the intra-particle diffusion rate of the adsorbate. In order to evaluate the time-dependent intra-particle diffusion rate of adsorbates from the surface sorption sites into the interior sites of the material, a kineticbased model developed by Weber and Morris was employed [56]. The intra-particular diffusion model (Eq. (6)) was expressed as: (

 = &' " ) + * … (6) where kid is the intra-particle diffusion rate constant (g/mg/min1/2), and Z is the intercept. According to this theory, the plot of the amount of Cd(II) adsorbed per unit mass of adsorbent (qt) against the square root of time (t1/2) should be linear and, if particle diffusion is the ratelimiting step, the line will pass through the origin. Fig. S8 (supplementary data) shows the intra-particle diffusion plot for Cd(II) at 298, 308 and 318 K and phenol at 298 K. The linear portion of the plot which covers a wide range of contact times between Cd(II) and ASAPTES fit well to the intra-particle diffusion model. This suggests that intra-particle diffusion plays an important role in the sorption of Cd(II) onto AS-APTES [57]. However, a slight deviation from the origin was observed, which may be due to the difference in mass transfer during the initial and final stages of sorption process [58]. In the case of phenol, the deviation from the origin and the low R2 values indicate that intra-particle diffusion was not the ratelimiting step. Indeed, phenol was likely to be mainly adsorbed on the outer surface of the solid material [59]. The intra-particle diffusion constants (Kid) and the R2 values are given in Table 4. 3.2.5. Effect of background electrolyte concentration The presence of background electrolyte ionic species can affect the extent of sorption in

aqueous solution. However, the influence on sorption depends on the nature and concentration of the electrolytes [60]. Sodium chloride (NaCl), one of the most abundant salts present in natural water, was chosen to provide background electrolytes for the present study. The background electrolyte concentrations varied between 0.001 mol/L to 0.1 mol/L; the pH was maintained at 5.0, and the initial concentrations of Cd(II) and phenol were 10.49 mg/L and 7.88 mg/L, respectively. Fig. 7 clearly shows that increasing the NaCl concentration from 0.001 mol/L to 0.1 mol/L decreased the sorption of Cd(II) by 15.76 %. This result infers that the Cd(II) ions were adsorbed with strong chemical forces and, perhaps, formed inner-sphere complexes with the amino groups present on the solid surface [37]. Moreover, significant changes in phenol removal were not observed when increasing the background electrolyte concentrations from 0.001 mol/L to 0.1 mol/L. This result suggests that phenol was also selectively adsorbed with strong forces on the solid surfaces [61]. 3.2.6. Thermodynamic studies The effect of temperature on the sorption of Cd(II) by AS-APTES was studied at 298, 308 and 318 K, and three basic thermodynamic parameters of Gibbs free energy of adsorption (∆G), change in enthalpy (∆H) and change in entropy (∆S), were evaluated. The Gibbs free energy of adsorption (∆G) was calculated as ∆G= -RT ln Kc

... (7)

where R is the universal gas constant (8.314 J/mol·K), T is the reaction temperature in K and Kc is the equilibrium constant taken from the Langmuir isotherm constant [57]. The ∆G values for Cd(II) sorption were -2.596, -2.830 and -3.899 at 298, 308, and 318 K, respectively. The negative values of ∆G infer that the process is thermodynamically feasible and spontaneous in nature. Moreover, the decreasing ∆G values with increasing temperature indicate more efficient sorption at higher temperature [62]. Furthermore, the change in enthalpy (∆H) and change in entropy (∆S) were determined using Eq. (8).

∆G = ∆H – T∆S

… (8)

The plot of ∆G versus T (Fig. S9, supplementary data) was utilized to obtain the values of ∆H and ∆S [33]. ∆H was found to be 16.95 kJ/mol, and ∆S was found to be 65.11 J/mol·K. The positive value of ∆H indicates that the sorption process is endothermic in nature and, again, suggests that higher temperatures favor sorption of Cd(II). The higher ∆H values imply a strong interaction between Cd(II) and AS-APTES [38]. The positive values of ∆S further indicate the high affinity of AS-APTES toward Cd(II) and increasing randomness at the solid-solution interface during the sorption process [59]. 3.2.7. Simultaneous sorption of Cd(II) and phenol The simultaneous sorption of Cd(II) and phenol was carried out by varying the phenol concentration from 2.0 to 15.0 mg/L and fixing the Cd(II) concentration at 10.0 mg/L. In another experiment, the phenol concentration was maintained at 8 mg/L and the Cd(II) concentration was varied from 2.0 to 30 mg/L. The pH of the solution was kept at 5.0, and the temperature was kept constant at 25±1 °C. Fig. 8(a) shows the percent removal of Cd(II) and the extent of phenol removed as a function of initial phenol concentration; Fig. 8(b) shows the percent removal of phenol and the amount of Cd(II) removed as a function of initial Cd(II) concentration. Increasing the phenol concentration from 2.0 to 15.0 mg/L did not noticeably affect the removal of Cd(II), though the amount of phenol removed slightly increased. Similarly, the percentage of phenol removal remained nearly constant when the Cd(II) concentration increased from 2.0 to 30.0 mg/L. These results suggest that AS-APTES is a potential solid material for simultaneous removal of Cd(II) and phenol from aqueous solutions [63]. 3.3. Desorption and reusability Several eluents are reported to desorb the pre-adsorbed metals cations and other organic pollutants efficiently; however it is necessary to find a suitable eluent which allows the

reusability of the adsorbents. In this study, HCl is found to be a suitable eluent and the percentage of Cd(II) and phenol desorbed using various concentrations of HCl is given in Fig. S10.

It is observed that 0.05 mol/L HCl could desorb 92.25 % of Cd(II) and 96.17 % of

phenol within 2 h of contact. Although ~99.9 % of phenol is desorbed by 0.2 mol/L of HCl, further increased in HCl acid concentration above 0.05 mol/L did not show significant enhancement in desorption of Cd(II). Therefore, 0.05 mol/L HCl solution was chosen for regeneration studies. Furthermore, reusability of AS-APTES for the removal of Cd(II) and phenol was studied and the results were shown in Fig. 9. The percentage removal of Cd(II) was slightly decreased as the number of adsorption-desorption cycles increased; however, Ca 64 % of Cd(II) was removed even after the material passed through six replicates of adsorption-desorption cycles. On the other hand, the phenol removal was remained constant throughout the six replicates of adsorption-desorption cycles. Therefore, this study indicates that the AS-APTES is a suitable and excellent material for multiple use in the removal of Cd(II) and phenol simultaneously from aqueous solutions. 3.4. Application of hybrid materials in real water matrix treatment In order to assess practical implication of AS-APTES in real wastewater treatment technologies, the tap water spiked with Cd(II) and phenol was treated by AS-APTES. The tap water usually contains several alkali and alkaline earth metals which may compete with Cd(II) or phenol towards the active sites of AS-APTES. However, it was observed that there was no significant change in percentage uptake of Cd(II) or phenol from tap water compared to the purified water and the percentage removal of Cd(II) and phenol was found to be 86.03 % and 37.18 %, respectively. This is due to the fact that common cations such as calcium, sodium, potassium etc. present in real water sample do not form complexes with amino group; therefore, AS-APTES possessed a good selectivity and high sorption efficiency for

these two pollutants even in tap water solutions [64]. This study further reaffirms the efficient use of this hybrid material in the tertiary treatment of real wastewater. 3.5. Sorption mechanism Based on the sorption data collected under various physico-chemical parametric studies, it is assumed that, in a single pollutant sorption studies, Cd(II) undergoes direct sorption on the solid surface, followed by intra-particle diffusion inside the mesopores of AS-APTES and finally forming an ‘inner-sphere complexes’ with the available amino groups on the solid sample [65]. It is also suggested that a small number of Cd(II) ions that enter within the interspace of AS-APTES undergo ion exchange process with the K+ or other cations present within the interspace as a minor sorption process [6]. Similarly, phenols are also mainly sorbed to the AS-APTES through the hydrogen bonds interaction with the amino group present on the solid material [66, 67]. Moreover, there is a possibility that phenol sorption occurs through the formation of hydrogen bonding with the -OH group present onto the solid materials [1, 68] or the phenol is partitioned within the interspace with the introduced hydrophobic organo silane [21]. The sorption capacity for phenol is very low comparing to the Cd(II) which is, perhaps, due to hydrogen bonding donator or acceptor nature of water. It is probable that water molecule could form a hydrogen bond either with the –NH2 group of AS-APTES or with phenol; this hinders the probability of phenol for hydrogen bonding with amino group and consequently reduces the sorption of phenol [69]. Simultaneous sorption studies suggests that these two different types of pollutants are sorbed through different binding sites avaiable on the solid surface. Therefore, in binary system, Cd(II) ions are preferably sorbed specifically as bonding with the amino group and finally forming an ‘innersphere complexes’ whereas phenols are mainly bound to the aluminol or silanol group of ASAPTES through a strong hydrogen bond or even partitioned within the interspace with the introduced hydrophobic end. Previous report also has shown that nitrogen-functionalized

magnetic ordered mesoporous carbon and carboxylic functionalized porous materials were successfully utilized for the simultaneous removal of Cu(II) with phenol and Pb(II) with phenol respectively [1, 63].

4.

Conclusions Natural mica-type clay, sericite (S), was annealed at 800 °C for 4 h, followed by acid

activation using 3.0 mol/L of HCl at 100 °C in order to obtain activated sericite (AS). The activation caused an immense increase in specific surface area, providing a suitable surface structure

for

grafting

of

organosilane.

The

AS

was

functionalized

using

3-

aminopropyltriethoxysilane (APTES), and the successful grafting of APTES onto AS was confirmed through EDX, BET, XRD and FT-IR analyses. Further, the functionalized activated sericite (AS-APTES) was assessed for suitability in the remediation of wastewaters contaminated with Cd(II) and phenol from aqueous solutions. Batch reactor experiments were conducted to assess the effects of pH, initial pollutant concentration, contact time, temperature, and background electrolyte concentrations. Moreover, AS-APTES was successfully utilized for the simultaneous removal of Cd(II) and phenol. Equilibrium sorption data for Cd(II) sorption was fitted well to the Langmuir and Freundlich adsorption isotherms, whereas phenol sorption data was fitted well to the Langmuir adsorption isotherm. The kinetic data for both pollutants reasonably fitted well to the pseudo-second order kinetic model. Further, increase in temperature was favored the removal of Cd(II) and the sorption process was found to be spontaneous and endothermic in nature. The desorption and reusability studies as well as applicability of the material for real water sample treatment demonstrated that AS-APTES could be a promising solid material for the efficient and effective removal of two important water pollutants i.e., Cd(II) and phenol from aqueous wastes. Acknowledgements

This work was supported by the Korea Ministry of Environment as a "Converging technology project" (Proposal No. 2013001450001). References [1]

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•

Natural sericite clay was activated and functionalized with organosilane

•

Surface-functionalized activated sericite showed enhance sorption capacity

•

Simultaneous removal of Cd(II) and phenol from aqueous solutions

•

Sorption mechanism was extensively investigated

Figure Legends: Fig. 1: SEM micrograph of (a) Sericite (b) AS (c) AS-APTES Fig. 2. XRD patterns of Sericite, AS and AS-APTES Fig. 3. FT-IR spectra obtained for Sericite, AS and AS-APTES Fig. 4. Effect of pH on the removal of Cd(II) and phenol Fig. 5. Effect of initial concentration and temperature on the removal of Cd(II) and phenol Fig. 6. Effect of contact time and temperature on the removal of Cd(II) and phenol Fig. 7. Effect of background electrolytes concentration on the removal of Cd(II) and phenol Fig. 8. Simultaneous removal of Cd(II) and phenol (a) percentage removal of Cd(II) at various concentration of phenol (b) percentage removal of phenol at various concentrations of Cd(II). Fig. 9. Percentage removal of Cd(II) and phenol in sequential adsorption-desorption cycles.

(a)

(b)

(c)

Fig. 1

Intensity (a.u.)

AS-APTES

AS

Sericite 0

10

20

30

40

2θ

50

60

70

80

Fig. 2

Transmitance (%)

AS-APTES

AS

Sericite

0

500

1000

1500

2000

2500

3000

3500

4000

Wavenumbers (cm-1)

Fig. 3

100 Cd(II) 80 Removal (%)

Phenol 60

40

20

0 0

2

4

6 pH

Fig. 4

8

10

Cd(II) 318 K Cd(II) 308 K Cd(II) 298 K Phenol 298 K Phenol 318 K

100

Removal (%)

80 60 40 20 0 0

5

10

15

20

25

30

35

Initial concentration (mg/L)

Fig. 5

100

Removal (%)

80

60

Cd(II) 318 K

Cd(II) 308 K

Cd(II) 298 K

Phenol 298 K

40

20

0 0

60

120 180 240 300 360 420 480 540 600 660 720 780 Time (min)

Fig. 6

100

80

Removal (%)

Cd(II) 60

Phenol

40

20

0 0.001

0.01

0.1

Background electrolytes concentration (mol/L)

Fig. 7

100

2.0

80

1.8

60

1.6

Cd(II) Phenol

40

1.4

20

1.2

0

1.0 0

2

4 6 8 10 12 Concentration of phenol (mg/L)

14

16

Amount of phenol removed (mg/g)

Cd(II) removal (%)

(a)

(b) 7 6

Phenol removal (%)

40

5 30

4

20

3

Phenol

2

Cd(II)

10

1

0

0 0

5

10

15

20

25

30

35

Concentration of Cd(II) (mg/L)

Fig. 8 100

Cd(II) Phenol

Removal (%)

80

60

40

20

0

1

Fig. 9

2

3 4 Number of cycles

5

6

Amount of Cd(II) removed (mg/g)

50

Table 1: Textural properties of sericite, AS and AS-APTES. Materials

BET specific

Pore size

Pore volume

surface area

(nm)

(cm3/g)

(m2/g) Sericite

3.65

2.45

0.019

AS

62.92

7.89

0.112

AS-APTES

59.63

7.66

0.110

Table 2 Langmuir, Freundlich and Temkin constants along with R2 values obtained for the removal of Cd(II) and phenol . Temp . (K) Syste m

Langmui r qo (mg/g)

Freundlic h KL (L/g)

Temki n R2

Cd(II)

298

5.747

2.852

0.999

Cd(II)

308

6.897

3.021

0.988

Cd(II)

318

8.475

4.370

0.998

Pheno l

298

1.502

3.742

0.996

1/n

0.16 7 0.17 3 0.20 0 0.06 0

KF (mg/g ) 3.648 4.498 5.572

1.265

R2

0.98 7 0.98 3 0.98 4 0.93 3

BT (J/mol ) 0.041 9 0.043 4 0.034 7 0.361

KT (L/mg)

R2

2410.4 5 21903. 3 12517. 8 4.098

0.95 8 0.96 8 0.96 5 0.85 7

Table 3a Langmuir sorption capacity of Cd(II) obtained for various low cost materials. Materials

Sorption

pH

capacity

Temperature

Initial

(K)

concentration

(mg/g)

Reference

range (mg/L)

Sericite

3.464

6.5

298

2 to 25

[6]

AC obtained

1.851

6.15

293

15 to 45

[7]

0.528

4.5

298

1 to 20

[37]

from olive stone Iron-oxide immobilized sand 37

Iron oxide

0.116

6.0

298

0.1 to 0.5

[39]

Kaolinite clay

0.88

-

303

10 to 150

[40]

Bamboo

12.08

-

298

20 to 100

[41]

Porous resin

3.506

7.0

298

20 to 500

[42]

AC obtained

2.493

4.5

298

1 to 20

[43]

2.135

4.5

298

1 to 20

[43]

5.0

298

10 to 55

[44]

1.82

5.0

298

10 to 55

[44]

1.24

6.0

303

2 to 20

[45]

AS-APTES

5.747

5.0

298

2 to 30

Present study

AS-APTES

8.475

5.0

318

2 to 30

Present study

activated red mud

charcoal

from rice hulls AC obtained from areca nut waste Formaldehyde 4.76 modified bean husk Pyridine modified bean husk Bagasse Fly ash

Table 3b Langmuir sorption capacity of phenol obtained for various low cost materials. Materials

Bentonite Neutralized red mud H2SO4 treated Aspergillus

Sorption capacity (mg/g) 1.712 4.127

pH

Temperature (K) 298

Initial concentration range (mg/L) 25 to 500 5 to 200

5 6

0.328

5.1

Reference

[46] [47]

294

<1

[48]

38

niger Bacillus sp. Immobilized onto tea waste Kaolinite HDTMAKaolinite PTMAKaolinite Rice husk Rice husk ash AS-APTES

7.62

-

-

50 to 250

[49]

0.52 3.846

6.5 6.5

298 298

10 to 100 10 to 100

[50] [50]

1.111

6.5

298

10 to 100

[50]

0.002 0.886 1.502

7 7 5

298 298 298

0.15 to 0.5 0.5 to 1.3 2.0 to15.0

[51] [51] Present study

Table 4 Kinetic parameters obtained from pseudo-first order, pseudo-second order kinetic and Intra particle diffusion model for Cd(II) and phenol. Syste m

Tem p. (K)

Pseudo first order kinetic

Pseud o secon d order kineti c k1 qe (1/min) (mg/g ) 9.21×1 2.102 0-3

Intra particle diffusio n

R2

0.944

k2 (g/mg/mi n) 5.4×10-4

qe (mg/ g) 7.352

R2

0.97 4

Kid (mg/g/min1 /2 ) 0.266

Cd(II)

298

Cd(II)

308

6.91×1 0-3

2.040

0.986

9.2×10-4

7.042

0.98 1

0.311

Cd(II)

318

6.91×1 0-3

1.752

0.893

2.84×10-3

6.024

0.98 4

0.396

Phen ol

298

2.30×1 0-3

1.759

0.923

17.19 10-2

× 1.277

0.99 9

0.011

Z

R2

0.62 9 0.48 8 0.18 4 1.06 6

0.99 6 0.99 6 0.95 8

0.93 8

39

Equation (1)
 1
 = +    

Equation (2) 1 log  = log
 + log   Equation (3)    = ln  + ln
  Equation (4) log ( −  ) = log  −

 " 2.303

Equation (5) " 1 " = + $  $  Equation (6) (  = &' " ) + * Equation (7) ∆G= -RT ln Kc Equation (8) ∆G = ∆H – T∆S

40