MnFe2O4 composite

Groundwater for Sustainable Development 2-3 (2016) 53–72

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Research paper

Kinetic, mechanistic and thermodynamic studies of removal of arsenic using Bacillus arsenicus MTCC 4380 immobilized on surface of granular activated carbon/MnFe2O4 composite$ M.S. Podder n, C.B. Majumder Department of Chemical Engineering, Indian Institute of Technology, Roorkee, Roorkee 247667, India

art ic l e i nf o

a b s t r a c t

Article history: Received 21 April 2015 Received in revised form 25 April 2016 Accepted 30 May 2016 Available online 8 June 2016

The ability of biofilm of Bacillus arsenicus MTCC 4380 supported on Granular activated carbon/MnFe2O4 composite for removing As(III) and As(V) ions from wastewater were investigated, in terms of kinetics, mechanistic and thermodynamics. Optimum biosorption/bioaccumulation conditions were estimated as a function of contact time and temperature. The equilibrium was achieved after about 180 min at 30 °C temperature. Non-linear regression analysis was performed for determining the best–fit kinetic model based on three correlation coefficients and three error functions and for forecasting the parameters involved in kinetic models. The results exhibited that Brouser–Weron–Sototlongo for both As(III) and As (V) was able to provide realistic explanation of biosorption/bioaccumulation kinetic. Applicability of mechanistic models in the current study exhibited that the rate controlling step in the biosorption/ bioaccumulation of both As(III) and As(V) was film diffusion rather than intraparticle diffusion. The estimated thermodynamic parameters ΔG0, ΔH0 and ΔS0 revealed that the biosorption/bioaccumulation of both As(III) and As(V) on the adsorbent attached with biofilm was feasible, spontaneous and exothermic in nature. The activation energy (Ea) assessed from Arrhenius equation indicated the nature of biosorption/bioaccumulation being ion exchange type. Increasing concentration of As(III) and As (V) furthermore improved the initial sorption rate, h. Acquired results indicated that the biofilm supported on adsorbent tried are very encouraging for removal of arsenic. & 2016 Published by Elsevier B.V.

Keywords: Arsenic Biosorption/bioaccumulation Biofilm Kinetic Mechanistic Thermodynamic

1. Introduction Arsenic is a persistent, bio-accumulative, toxic element (Langsch et al., 2012) and is ubiquitous in the environment (soil, air, water, and also in living matters) (Singh et al., 2015). Arsenic is considered as a category 1 and group A human carcinogen by the International Association For Research on Cancer (IARC, 2004) and the US Environmental Protection Agency (US EPA, 1997), respectively. Its occurrence in the environment is not only because of biological metabolism, volcanic deposits, rock and soil weathering, geothermal sources, however also to numerous anthropogenic activities comprising smelting, mining activities, petroleum refining, pesticide manufacturing, agricultural use of pesticides and herbicides, production of wood preservatives, pigmentation and combustion of fossil fuels (Smedley and Kinniburgh, 2002; Mondal et al., 2006; Mohan and ☆

GAC/MnFe2O4 composite (MGAC). Corresponding author. E-mail addresses: [email protected] (M.S. Podder), [email protected] (C.B. Majumder). n

http://dx.doi.org/10.1016/j.gsd.2016.05.005 2352-801X/& 2016 Published by Elsevier B.V.

Pittman, 2007). There has been a severe environmental challenge for the metallurgical industries, particularly for copper smelters, because of pressures from public judgement and many environmental regulations enacted. Copper smelting wastewater bearing arsenic may contain highly elevated concentrations of potentially toxic oxyanion, As(III) and As(V) (Basha et al., 2008). As(III) is more poisonous, more mobile and more challenging to be scavenged from water compared to As(V) (Smedley and Kinniburgh, 2002; Hong et al., 2014). In copper smelting wastewater concentration of arsenic is as high as 1979 mg/L (Basha et al., 2008). The occurrence of heavy metal ions, like lead, copper, zinc, iron, cadmium and bismuth, nickel, chromium restrict the solubility of arsenic due to the formation of sparingly soluble metal arsenates, poses a serious threat towards man and the flora and fauna of our ecosystem contaminating the natural water tables (ground and surface water) in the vicinity. The industrial wastewater discharged from various industries also pollutes the groundwater. So, the treatment of toxic contaminants discharged by the industries should be removed from the wastewater before discharging to the nearby rivers or ponds for the sustainable development.

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Nomenclature aE bE A Bbt C0 Ce Cs Cint d D1 D2 De Df Dp Ea F f2 h kAV kb kd kDW kEXP kint kM kMOE knBWS,α kPF kPFO kPSO k′1,0 k′2R k′′2R k′EXP

Elovich coefficient representing the initial adsorption rate (mg/g min) Elovich coefficient representing desorption constant (g/mg) the frequency factor a mathematical function of F (F¼qt/qe) initial concentration of arsenic in the solution (mg/L) equilibrium concentration of arsenic in the solution (mg/L) the concentration of arsenic on the biosorbent and in the solution (mg/L) the intercept of the intraparticle diffusion plot (mg/g) the thickness of the water film adhered to the biosorbent (cm) film diffusion constant (cm2/s) pore diffusion constant (cm2/s) diffusivity (m2/s) film diffusion coefficient (cm2/s) pore diffusion coefficient (cm2/s) the activation energy of adsorption (kJ/mol) the adsorption progress (F¼qt/qe) the involvement of pseudo second order model the initial sorption rate (mg/g min) the Avrami kinetic rate coefficient (1/min) the Bangham model constant (L/g) the equilibrium constant (L/g), the Dumwald–Wagner rate constant (1/min) the exponential rate coefficient (mg/g min) the intraparticle diffusion rate coefficient (mg/g min0.5) the film diffusion rate constant (1/min) the mixed 1,2 order rate coefficient (1/min) the Brouser–Weron–Sototlongo reaction constant (1/min) the fractional power model rate coefficient (mg/g min) the pseudo first order rate coefficient (1/min) the pseudo second order rate coefficient (g/mg min) the fractal–like mixed 1,2 order rate coefficient α (1/min) the modified second order rate coefficient (1/min) the Ritchie second order rate coefficient (1/min) division of kEXP by qe (1/min)

Long term exposures to arsenic contaminated drinking water can cause permanent and severe human health hazards such as gastrointestinal, skin problems (lesions and keratosis), hypertension, black–foot disease, degenerative diseases (affecting the nervous, digestive and respiratory systems) and also cancer (affecting the skin, liver, respiratory, bladder, lungs and kidney) and ultimately death (Yoshida et al., 2004; Wang et al., 2007). Based on the investigation of fatal effect of arsenic on human body, the World Health Organization (WHO) in 1993 and the European Commission in 2003 revised the maximum contaminant level (MCL) of arsenic in drinking water from 50 μg/L to 10 μg/L (WHO, 1993; European commission Directive, 98/83/EC, 1998). The most usually employed arsenic scavenging techniques from water and wastewater are precipitation/coagulation, oxidation/ precipitation, precipitation/adsorption, ion exchange, membrane techniques, reverse osmosis, electrocoagulation, electrodialysis (Mondal et al., 2006; Mohan and Pittman, 2007; Basha et al., 2008; Singh et al., 2015). These technologies are often incompetent and/

k′′EXP k′PFO k′FPSO k′PSO k′′FPSO m mb n nAV nBWS nR p qe qt r R Re t α t T v V

α

the fractal–like exponential rate coefficient (1/min) the fractal–like pseudo first order kinetic rate coeffiα cient (1/min) is the fractal–like pseudo second order kinetic rate α coefficient (g/mg min) multiplication of qe and kPSO (1/min) multiplication of qe and k′FPSO (1/min) the weight of the biosorbent per liter of solution (g/L) an integer the number of observations in the experimental study constant corresponding to the mechanism of adsorption a fractional reaction order number of surface sites the number of parameters to be estimated the amount of adsorbate adsorbed on the biosorbent surface at equilibrium (mg/g) the amount of adsorbate adsorbed on the biosorbent surface at time t (mg/g) mean radius of biosorbent particle (cm) the universal gas constant (8.314 J/mol K) % removal of adsorbate contact time (min) a fractal time the solution temperature (K) adjustment parameter the volume of the solution (mL)

Greek symbols

α αb θ θ0 θe τ1/2 τnBWS,α ΔG0 ΔH0 ΔS0

the fractal time exponent Bangham model constant the biosorbent surface coverage at pre-adsorbed stage, (θ ¼qt/qe) dimensionless the biosorbent surface coverage at time t, (θ0 ¼q0/qe) dimensionless the equilibrium surface coverage half–reaction time (min) the time required for adsorbing half the maximum amount Gibbs free energy change (kJ/mol) enthalpy change (kJ/mol) entropy change (J/mol K)

or costly, typically producing large amount of sludge containing high concentration of heavy metals, which must be discarded. The scientific search for exclusive techniques has been concentrated on the use of biosorption or bioaccumulation process. Many microorganisms are famous to be competent to uptake metal ions from less concentrated aqueous phase and for accumulating them within their cell structure. Microorganisms can physically scavenge heavy metals from aqueous phase via either bioaccumulation or biosorption. Bioaccumulation can be stated as the uptake of contaminants by living cells. The contaminants can transport from the exterior of the microbial cell via the cellular membrane where the metal is sequestered, accumulate intracellularly through the cell membrane as well as via the cell metabolic cycle (Wong and So, 1993; Malik, 2004). Biosorption is a passive uptake of contaminants by definite categories of dead/inactive biological materials or by materials derived from biological sources, may bind and accumulate contaminants from aqueous phase (Naja and Volesky, 2006).

M.S. Podder, C.B. Majumder / Groundwater for Sustainable Development 2-3 (2016) 53–72

Biosorption is because of various metabolism-independent processes that basically occur within the cell wall, where the uptake mechanisms of contaminant will vary consistently with the type of biomass. Biofilms can be well-defined as microorganisms communities attached to a surface (Le Cloirec et al., 2003). There are four possible motivations for the formation of biofilm: defense (shield from risky situations), colonization (formation of biofilm as a mechanism for remaining in a favorable position), community (use of cooperative profits) and default mode of growth. Bacteria devote the most of their normal presence growing as a biofilm. It is probable that the existence of an appropriate substrate for attachment is all that is essential for prompting formation of biofilm (Jefferson, 2004). Biofilms have been effectively utilized for the water treatment for over a century (Atkinson, 1975). The advantages of this type of bioprocess became a focus of interest for a considerable number of researchers, not only in the field of water and wastewater treatment, but also in many other areas of biotechnology (Adler, 1987). A large number of research projects are currently being conducted on biofilm reactors for plant and animal cell cultures, for the production of bioactive substances, drinking water production, groundwater and wastewater treatment. One key advantage of biofilms is the positive influence of solid surfaces on bacterial activity observed by ZoBell (1943) and confirmed by other researchers (Audic et al., 1989; Doran and Bailey, 1986; Klein and Ziehr, 1990; Mondal et al., 2008). Microbial biofilms can also be applied in-situ for the treatment of groundwater. Immobilized biofilms, viable and nonviable, essentially trapped metal ions as contaminated water is pumped through (Macaskie et al., 1992; Summers, 1992). Bacteria are documented for generating macromolecules typically termed extracellular polymeric substances or exopolysaccharides (EPSs). EPSs are bacterial metabolic products and may originate from their hydrolysis or lysis. They are correspondingly related to the organic matter exists in the wastewater to be treated (Comte et al., 2008). These EPS building molecules include numerous ionizable functional groups for example hydroxyl, carboxyl, sulphydryl, amino and phosphoric groups (Denizli et al., 2004). The accumulation of elements from environment and the adhesion to surfaces are two main functions of EPS on supported biosorption processes. The interactions mechanisms between biofilms and metal ions are properly defined by Le Cloirec et al. (2003) and can be recommenced as follows: bulk diffusion (diffusion of the metal ions existing in solution to the external surface of the biofilm), external mass transfer (occurring of mass transfer via the layer of high concentration around the biofilm), rapid metal ion interactions with surface of solid and specially with the bacterial wall (these interactions can be bioaccumulation, enzyme production, biosorption on the bacterial surface, extracellular precipitation by metabolites produced by bacteria, oxidation and/or reduction and extracellular complexation), sluggish surface diffusion, diffusion into the biofilm prior to the interaction reaction with bacteria and lastly, interactions with bacteria existing inside the biofilm. Most of the researches have focused on either the biosorption of metal ion on the surface of non-living biomass or biosorption on the surface living microbial cells followed by intracellular and extracellular accumulation of metal ion (Yee and Fein, 2001; Velasquez and Dussan, 2009) in batch and continuous column study. A novel technique for an efficient metal ion removal mediated by immobilized bacterial cells is designed. This metal ion removal system is called simultaneous biosorption and bioaccumulation (SBB) system. In this system, both non-living biomass and living microbial cells are used simultaneously. For effective removal of toxic metal ions, the biomass needs to be immobilized to increase the mechanical strength and resistance to the various chemical

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constituents of aqueous waste (Shashirekha et al., 2008). Due to practical difficulties in solid–liquid separation, the free biomass are immobilized on the support. Immobilized biomass also shows better potential in packed/fluidized bed reactors and continuous stirred tank reactors due to minimal clogging under continuous flow conditions as well as effective biosorbent regeneration and metal recovery (Gadd, 1992; Shashirekha et al., 2008). Adsorption/ biosorption and biological treatment either operated separately or simultaneously in one-unit results in a better removal and process performance (Dash et al., 2009). Arsenic removal efficiency of bacteria improves when it is immobilized on a solid support like GAC (Mondal et al., 2006). Mondal et al. (2008) reported the bio-removal of arsenic from contaminated water by using Ralstonia eutropha MTCC 2487 and activated carbon in a batch reactor. They observed simultaneous adsorption bioaccumulation (SABA) when fresh GAC was used as supporting media for bacterial immobilization. Mishra et al. (2013) designed the simultaneous biosorption and bioaccumulation (SBB) batch system for the removal of Zn(II) ion from liquid phase. Cedrus deodara sawdust was used as carrier to immobilize Zinc sequestering bacteria “VMSDCM” accession number HQ108109. In the current research, GAC/MnFe2O4 composite was used as carrier due to its higher porosity, surface area, easy availability and cost effectiveness and Bacillus arsenicus MTCC 4380 was used for formation of biofilm due to its arsenic resistance property. Finally the immobilized bacterial cells were used to remove a wide range of arsenic. In the heterogeneous systems, the choice of the support is definitely very vital for the development of a good catalyst and activated carbon (AC) has been mainly utilized for this purpose since the 1970 s (Reinoso, 1998). Owing to its high surface area, special surface reactivity and porous structure, activated carbon, it can efficiently adsorb gases and compounds dispersed or dissolved in liquids (Oliveira et al., 2002). In the water treatment arena, activated carbon is widely utilized as an adsorbent for the removal of arsenic (Mondal et al., 2007a, 2008). This adsorbent is highly inert and thermally stable and it can be utilized over a wide pH range (Zhang et al., 2007a) and moreover activated carbon can easily be functionalized (Shim et al., 2001). In the present work, GAC/MnFe2O4 composite was synthesized in order to hybridize high adsorption capacity of MnFe2O4 with adsorptivity of GAC. MnFe2O4, a familiar soft material, has outstanding chemical stability and frequently controls the concentration of free metal and organic matter in water or soil through adsorption reactions (Muroi et al., 2001; Wu and Qu, 2005). GAC can also remove arsenic, but to increase the adsorption capacity of GAC, it was modified with MnFe2O4. MnFe2O4 is mostly responsible for developing the charge on the surface of adsorbent when GAC/MnFe2O4 composite comes in interaction with water. The surfaces of both B. arsenicus MTCC 4380 and GAC/MnFe2O4 composite achieves the coordination shells with the prevailing OH group with the materials under hydration (Giri et al., 2013). With the change in pH, these surface-active OH groups may again bind or release H þ where the surface remains positive. As the affinity of AsO33  or H2AsO3  and AsO43  or H2AsO4  anions with metal oxide is higher than that of hydroxide with metal oxide, the AsO33  or H2AsO3  and AsO43  or H2AsO4  can substitute the hydroxide from the surface of the hydrolyzed metal oxides. Formation of MnFe2O4 layer enhanced the net positive charge (NPC) of the GAC/MnFe2O4 and improved the arsenic adsorption capacity. The iron and manganese in GAC/MnFe2O4 composite triggered the oxidation of As(III) to As(V), which can easily be adsorbed by the adsorbent in the experimental pH range (Goldberg and Johnston, 2001). In case of GAC, As(III) did not oxidize to As(V). So the binding of the As(III) or As(V) to the surface of the GAC/MnFe2O4 adsorbent generally resulted in higher binding

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capacities owing to the probable various electron transfers from either the Mn or Fe ions (Parsons et al., 2009). Furthermore the arsenic resistant bacteria are normally gram negative in nature, hence can biosorb positively charged metal ions by electrostatic attraction (Hossain and Anantharaman, 2005). Positively charged metal ions like Mn and Fe, after chemisorption on cell surface decrease its surface negative charges as well as generate some additional positive charges for further biosorption of negatively charged arsenic ions. Moreover the bioaccumulation of other metals i.e. Mn and Fe can take place by some specific protein in the bacterial cells. For example Mn is taken up by Mg transport system (Nies and Silver, 1989). Fe normally follows siderophore-mediated uptake (Cornelis and Matthijs, 2002). So if Mn or Fe is mobilized from these composites to the aqueous media it will be finally biosorbed and accumulated by the immobilized bacterial cells. The purposes of the present study, divided into six parts, were 1) to characterize the prepared fresh adsorbent (GAC/MnFe2O4 composite) and adsorbent attached with biofilm before and after metal loading with SEM-EDX, 2) to inspect the effect of contact time in addition to temperature for removing As(III) and As (V) from synthetically prepared copper smelting wastewater, 3) to calculate the kinetics and mechanism of present biosorption/ bioaccumulation process, 4) to study the thermodynamics of biosorption/bioaccumulation process to identify the mechanism, 5) to examine the effect of initial arsenic concentration on kinetics of biosorption/bioaccumulation, 6) to investigate the impact of temperature on kinetics of biosorption/bioaccumulation.

2. Materials and methods 2.1. Materials All the chemicals and reagents were of analytical reagent grade and utilized devoid of additional refinement. Standards, matrix modifier and wash solutions were prepared by deionized double distilled water. The stock solutions of As(III) (1000 mg/L) and As(V) (1000 mg/L) were prepared by dissolving 1.734 g of sodium arsenite (NaAsO2) and 4.16 g of sodium arsenate (Na2HAsO4, 7H2O), purchased from Himedia Laboratories Pvt. Ltd. Mumbai India, in 1 L of double distilled water, respectively. All other necessary chemicals used in the experiments, were purchased from Himedia Laboratories Pvt. Ltd. Mumbai India. Glassware utilized for experimental purposes was washed in 10% nitric acid and rinsed with deionised water for removing any probable interference by other metals. 2.2. Microorganism and growth medium The microorganism utilized was the arsenic resistant bacterium B. arsenicus (MTCC 4380) (Microbial Type Culture Collection and Gene Bank (MTCC), Chandigarh, India). Culture media was prepared as per the guidelines of microbial type cell culture (MTCC). Composition of growth medium and cultivation conditions are exhibited in Table 1. 2.3. Acclimatization The acclimatization of B. arsenicus MTCC 4380 in arsenic (either As(III) or As(V)) environment was carried out as follows (Mondal et al., 2008). The revived culture was initially grown in MTCC prescribed growth media in a 250 ml round bottom flask tightly closed with cotton plug as follows: B. arsenicus MTCC 4380 was cultivated in 250 mL flask containing 100 mL of the growth media with As(III) and As(V). The

Table 1 Composition of growth medium and cultivation conditions. Component (g/L) or condition Beef Extract Yeast extract Peptone NaCl pH Temperature (°C)

1.0 2.0 5.0 5.0 7.0 30

cultures were acclimatized to As(III) and As(V) individually by exposing the culture in a series of shake flasks. The bacterial inoculum was prepared by transferring a loop full of bacterial culture from the nutrient agar tubes to the flask containing sterilized growth media, incubated at 30 °C for 24 h with moderate agitation in an incubator cum orbital shaker. Then the acclimatization of B. arsenicus MTCC 4380 in arsenic environment was carried out as follows: After 24 h the synthetic medium in the flask had turned milky specifying significant bacterial growth in the flask. Appropriate amount of arsenic (As(III) or As(V)) was added into the flask having 100 mL sterilized growth media to acquire a concentration of 50 mg/L of arsenic. Firstly growth of B. arsenicus MTCC 4380 was inhibited and the growth started after 2 h. After 24 h of incubation at 30 °C, 5 mL of the arsenic resistant bacterial inoculum was periodically added in a series of 250 ml flasks containing 100 mL of arsenic containing sterilized growth media (As(III) or As (V) concentration, 100 mg/L, 200 mg/L, 500 mg/L, 800 mg/L, 1000 mg/L, 1200 mg/L, 1500 mg/L and 1800 mg/L) under sterile conditions in a laminar hood chamber. After 24 h later, another fresh growth media (As(III) or As(V) concentration, 2000 mg/L) was also inoculated with 5 mL of the last culture (As(III) or As (V) concentration, 1800 mg/L ) to ensure that the bacteria was already adapted to both As(III) or As(V). For inoculum, a further sub culturing was performed and all the inoculum transfers were done in exponential phase (OD value ∼1 at 600 nm). 2.4. Methods 2.4.1. Adsorbent preparation Granular activated carbon (GAC) was washed systematically with double distilled water and oven dried at 105 °C for 4 h. 25 g of dried GAC was added into 500 mL conical flask containing 250 mL of 0.5 M HCl solution; next was shaken for 4 h at 120 rpm at 25 °C. Then the mixture was left overnight. The mixture then was filtered to separate GAC which were repeatedly washed with double distilled water to provide neural pH. Then the adsorbents were dried at 110 °C for 3 h for removing moisture, cooled to room temperature and kept in plastic bags for further use. GAC/MnFe2O4 composites were prepared by chemical coprecipitation method with few modifications (Shao et al., 2012). In this procedure, a fixed quantity of acid treated GAC was mixed into 200 mL solution containing dissolved ferric (III) chloride (FeCl3) (0.05 mol) and manganese (II) chloride (MnCl2) (0.025 mol) at room temperature. The amount of acid treated GAC was fixed for acquiring GAC/MnFe2O4 mass ratios of 2:1. The solution temperature was raised to 60 °C under energetic magnetic stirring and then 5 mol/L of NaOH solution was added drop wise to the above mixture till the pH of the solution attained 11. Thereafter next 1 h agitation was carried on. Then the suspension was heated in a water bath at 100 °C for 4 h. After cooling, the prepared composite was constantly washed with double distilled water for eliminating the contaminations (e.g. Na þ , Cl  ) accompanied with the

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processes. Next as-prepared composite was collected from the washed solution by filtering the mixture and then was oven dried at 110 °C. The reaction ionic equation (Bianfang et al., 2007) is as follows: Mn2 þ þ 2Fe3þ þ 8OH  -Mn(OH)2↓þ2Fe(OH)3↓-MnFe2O4 þ4H2O (1) MnFe2O4 þGAC þ4H2O-GAC–MnFe2O4

(2)

2.4.2. Immobilization of microbial cells on the adsorbent To immobilize B. arsenicus MTCC 4380 on prepared GAC/MnFe2O4 composite initially 90 mL culture media was inoculated with 5 mL of bacterial suspension of B. arsenicus MTCC 4380 from both As(III) and As(V) acclimatized 24 h old culture. The flasks were incubated at 30 oC for another 24 h with moderate shaking at 120 rpm. Then immobilization of bacterial cell was performed by adding weighed amount of prepared GAC/MnFe2O4 composite to the above suspension containing 24 h old culture. Then the flasks were again incubated at 30 °C for next 24 h with moderate shaking at 120 rpm. Bacterial cell immobilization was established by observing a small amount of bacterial treated GAC/MnFe2O4 composite through scanning electron microscopy. 2.4.3. Characterization The measurements of SEM were done for observing the surface morphologies of the GAC/MnFe2O4 composite attached with biofilm before and after biosorption/bioaccumulation process (SEM; LEO electron Microscopy, England). The images were taken with an accelerator voltage¼15 kV and an emission current¼100 mA by the Tungsten filament. 2.4.4. Batch experimental studies and analytical methods A medium with 1.0 g/L of beef extract and 2.0 g/L of yeast extract, 5.0 g/L of peptone and 5.0 g/L of NaCl was used for the growth of the microorganism. The media was sterilized at 121 °C for 15 min, cooled to room temperature, inoculated with bacteria and kept at 30 °C for 24 h with moderate agitation (120 rpm) in an incubator cum orbital shaker. All biosorption/bioaccumulation experiments were done by shaking optimum adsorbent dose of 0.7 g/L of the GAC/MnFe2O4 composite with 100 mL of B. arsenicus MTCC 4380 bacterial suspension as a test solution and required amount of arsenic (As(III) or As(V)) was supplemented to give final concentration As(III) or As(V) of requisite concentration, at an optimum initial pH value around 7 and at a preferred temperature in an incubator cum orbital shaker (REMI Laboratory instruments) at 120 rpm. 1.0 N NaOH and 1.0 N HCl solutions were used to adjust the initial pH of the solution using a digital pH meter (HACHs India). To study the biosorption/bioaccumulation kinetics, batch experiments were conducted by contacting an optimum adsorbent dose (0.7 g/L) of the GAC/MnFe2O4 composite with As(III) or As (V) solution of a fixed concentration (50 mg/L) at a constant temperature (30 oC) at a range of contact time (5–500 min). To inspect the impact of temperature on biosorption/bioaccumulation, batch experiments were done by contacting an optimum adsorbent dose (0.7 g/L) of the GAC/MnFe2O4 composite with As (III) or As(V) solution of a fixed concentration (50 mg/L) at an optimum contact time of 180 min at a range of temperatures (20, 25, 30, 40, 45 and 50 °C). To examine the effect of initial concentration on the kinetics of biosorption/bioaccumulation of both As(III) and As(V) by B. arsenicus MTCC 4380 immobilized on GAC/MnFe2O4 composite, all batch experiments were performed by agitating 0.7 g/L of GAC/MnFe2O4 composite with 100 mL of B. arsenicus MTCC 4380

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bacterial suspension as a test solution supplemented with required amount of arsenic (As(III) or As(V)) to give final concentration As(III) or As(V) of 50, 100, 500, 1000, 1500 and 2000 mg/L at a constant temperature (30 °C). To investigate the influence of temperature on the kinetics of biosorption/bioaccumulation of both As(III) and As(V) by B. arsenicus MTCC 4380 immobilized on GAC/MnFe2O4 composite, 0.7 g/L GAC/MnFe2O4 composite was added to each round bottom flasks containing B. arsenicus MTCC 4380 bacterial suspension as a test solution and required amount of arsenic (As(III) or As(V)) was supplemented to give final concentration As(III) or As(V) of 50 mg/ L. Experiments were conducted at three temperatures (30, 40 and 50 °C) by shaking the flasks. The samples were withdrawn from the flasks through filtration by Whatman Filter paper (Cat No 1001 125) (Remi Instruments ltd., Mumbai India) after fixed contact time for thermodynamic studies as well as at predetermined time intervals and then centrifuged at 10,000 rpm for 10 min for kinetic studies, a portion of filtrate was diluted with HNO3 solution (10%, v/v). The filtrate was analyzed for determination of arsenic concentration using ThermoFisher Scientific iCE 3000 Series AA graphite furnace atomic absorption (GFAA) spectrometer (detection limit 20 mg/L).

3. Theoretical background With the goal of assessing biosorption capacity by mass balance, detailed as the amount of adsorbate molecules adsorbed per unit mass of biosorbent at time t (mg/g) was calculated as follows:

qt =(Co−Ct )

V M

(3)

The amount of adsorbate molecules adsorbed in terms of percentage was calculated as follows:

R e ( %)=

( Co−Ct ) ×100 Co

(4)

where C0 and Ct are arsenic concentrations at time 0 and t (mg/L), respectively, V is the volume of the solution (mL) and M is the mass of biosorbent used (g). 3.1. Determining adsorption kinetic parameters by non-linear regression The kinetic parameter sets are computed by non-linear regression because of the inherent bias causing after linearization. This offers a mathematically laborious method to evaluate kinetic parameters utilizing the original kinetic equation (Khan et al., 1996; Ncibi, 2008; Hadi et al., 2012). Commonly Gauss–Newton methods or Levenberg–Marquardt based algorithms (Edgar and Himmelblau, 1989; Hanna and Sandall, 1995) are utilized. The biosorption kinetic data of arsenic on immobilized bacterial cells were analyzed by non-linear curve fitting analysis utilizing professional graphics software package OriginPro (8.5.1 version) for fitting the kinetic models. The optimization method requires the selection of a Goodness– of–Fit Measure (GoFM) with the purpose of valuing the fitting of the kinetics to the experimental data. In the present research, six GoFM (Residual Sum of Squares (SSE), reduced Chi–square test (Reduced χ2), coefficient of determination (R2), Adjusted R–square ( R̅2), R value (R) and Root–MSE value) were employed for evaluating isotherm parameters utilizing the OriginPro software by considering 95% confidence interval.

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and exponential (EXP) kinetic models are two parametric empirical models to analyze kinetic data at near to equilibrium. As well other empirical models for instance Avrami, modified pseudo second order (MPSO) and mixed 1,2 order model (MOE) equations have been recommended for modeling the adsorption kinetics at the interface of solid/solution. These models are mainly three parametric equations. The prior kinetic models of adsorption at the interface of solid/ solution can be generalized by taking into consideration the fractal–like methodology i.e. time dependence of adsorption rate coefficient. A physical significance of the fractal–like idea has been considered for adsorption kinetics on surfaces of solid that is energetically heterogeneous. The fractal–like study exhibits that the kinetics of adsorption at interface of the solid/solution in a real system with various kinds of surface sites and with various affinities for adsorption can be demarcated by a fractal–like methodology. Based on this study, the history of process can influence the process besides the achieved rate coefficient of adsorption is a function of time. Table 2 presents representative kinetic models for a biosorptive reaction (The details of adsorption kinetic modeling are provided with supplementary materials).

3.2. Adsorption kinetic modeling Biosorption kinetics reveals the rate of adsorbates bonding on the surface of the biological materials. Kinetics studies deliver the vital information about the probable mechanism of biosorption that includes the diffusion (bulk, external, and intraparticle) and chemical reactions. In general, it is supposed that adsorbate transport happens in the few following steps. The first step includes the transport of adsorbate in the bulk of the solution, second step comprises the external diffusion (the substrates diffuse from the bulk solution to the external surface of the biosorbent), the third step was because of the transport of the adsorbate across the boundary layer, the fourth step involves the transfer of compounds in the pores to the internal parts of the biosorbent and finally uptake of molecules by the active sites, and the fifth step comprises biosorption and desorption of adsorbate (Plazinski et al., 2009; Michalak et al., 2013). In many experimental biosorption systems, the influence of transport in the solution is rejected by fast mechanical mixing, so, it is not supposed to be involved in governing of the overall sorption rate and can be overlooked, as a rule (Plazinski et al., 2009). With the purpose of inspecting dynamic biosorption behavior of As(III) and As(V) onto immobilized bacterial cells i.e. governing the current biosorption process mechanisms and the probable rate limiting steps for example mass transport and/or chemical biosorption processes, kinetic models have been utilized for fitting the experimental data. Fourteen kinetic models were employed in the current research supposing that concentrations on the adsorbent surface are equal to the measured concentrations. There are many models for kinetics of biosorption at the interface of solid/solution. Fractional power model (FP), pseudo first order (PFO) and pseudo second order (PSO) are common empirical models, however their major drawback is that they may simply define the kinetics of adsorption at some restrictive situations (Haerifar and Azizian, 2013). Elovich is a two parametric semiempirical model and fractional power (FP), Ritchie second order

3.3. Adsorption mechanistic modeling The scavenging of adsorbate species from the liquid by the solid phase is taken place in three successive steps as follows (McKay et al., 1981; Singh and Pant, 2006; Al-Degs et al., 2008). The three stages tangled in the mechanism of biosorption process are as follows: (1) Firstly the adsorbate mass transfer from the aqueous phase on the biosorbent surface i.e. film diffusion or surface diffusion takes place. (2) Secondly internal diffusion of adsorbate via either a pore diffusion model or homogeneous solid phase diffusion model i.e. particle diffusion occurs.

Table 2 Adsorption kinetic models for biosorption/bioaccumulation. Sr. no.

Expression

Equation form

Remarks

1

Fractional power model (FP)

qt =kFP t v

2

Pseudo first order model (PFO)

qt =qe 1 − exp ( −kPFO t )

3

Pseudo second order model (PSO)

Based on adsorption capacity

(

qt =

(

Based on adsorption capacity

)

Chemisorption and based on adsorption capacity and

)

qe2 kPSO t

1 +( qe kPSO t )

4

Elovich model

⎛ 2.3 ⎞ ⎛ t +1 ⎞ ⎛ 2.3 ⎞ ⎛ 1 ⎞ ⎟−⎜ ⎟ × log ⎜ ⎟ qt =⎜ ⎟×log ⎜ ⎝ bE ⎠ ⎝ aE bE ⎠ ⎝ bE ⎠ ⎝ aE bE ⎠

Chemisorption

5

Avrami model

n qt =qe 1 − exp ⎡⎣ −(kAV t ) ⎤⎦ AV

Multiple kinetic orders

6

Modified second order model (MSO)

⎧ ⎡ ⎤⎫ ⎪ ⎪ 1 ⎥⎬ qt =qe ⎨ 1−⎢ ⎪ ⎪ ⎣⎢ β2R +k2′ R t ⎥⎦ ⎭ ⎩

Number of surface sites is two

7

Ritchie second order model

⎧ ⎡ 1 ⎤⎫ ⎥⎬ qt =qe ⎨ 1 − ⎢ ⎣ 1 +k2′ ′R t ⎦ ⎭ ⎩

Adsorbent surface coverage is presumed to be zero

8

Exponential kinetic model (EXP)

9

Mixed 1,2 order model (MOE)

10

Fractal like mixed 1,2 order model (FMOE)

(

Fractal like pseudo first order model (FPFO)

12

Fractal like pseudo second order (FPSO)

⎪

⎪

⎪

qt =qe ln ⎡⎣ 2.72 − 1.72 exp −kEXP ′ t ⎤⎦

(

qt =qe

qt = 11

⎪

)

(

)

⎡ ⎤ qe ⎢⎣ 1 −exp (−k ′ t α ) ⎥⎦ 1,0 1 −feq exp (−k ′ t α ) 1,0

(

(

Exponential form of kinetic equation A combined form of pseudo first and second order equations

1 −exp ( −kMOE t ) 1 −f2 exp ( −kMOE t )

qt =qe 1 − exp −kFPFO ′ tα qt =

)

Time dependency of the rate coefficient

Time dependency of the rate coefficient

))

Time dependency of the rate coefficient

qe2 k′ tα FPSO 1 +k′ qe t α PSO

13

Fractal like exponential (FEXP)

⎡ ⎤ qt =qe ln ⎢⎣ 2.72 − 1.72 exp −kExp ′ ′′ t α ⎦⎥

Time dependency of the rate coefficient

14

Brouser–Weron–Sototlongo model (BWS)

⎡ ⎛ ⎛ ⎞α ⎞−1/ (nBWS −1) ⎤⎥ t qt =qe ⎢ 1 − ⎜ 1+( nBWS −1) ⎜ ⎟ ⎟ τ nBWS , α ⎠ ⎠ ⎢ ⎥ ⎝ ⎝ ⎣ ⎦

Complex nature of adsorption

(

)

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Table 3 Adsorption mechanistic models for biosorption/bioaccumulation. Sr. no.

Expression

Equation form

Remarks

1

Weber and Morris model

qt =kint t 0.5+Cint

Intra-diffusion

2

Rate limiting step

⎛ q ⎞ 6 ⎛ D π2 ⎞ ln ⎜ 1 − t ⎟=ln 2 −⎜ 22 t ⎟ qe ⎠ ⎝ π ⎝ r ⎠

Based on diffusion

3

Dumwald–Wagner model

log 1 − F2 =

4

Richenberg model

F values40.85 Bb t F valueso 0.85 ⎛ π 2F Bb t = ⎜⎜ π − π − 3 ⎝

(

)

Intra-diffusion

−kDW t 2.303

= − 0.4977 − ln (1 − F )

Based on diffusion

⎞ ⎟⎟ ⎠

5

McKay plot

ln ( 1 − F )= − kM t

Film diffusion

6

Bangham”s model

⎡ ⎛ C0 ⎞ ⎤ ⎛ kb m ⎞ ⎟ =log ⎜ ⎟ + α log (t ) log ⎢ log ⎜ ⎝ 2.303V ⎠ b ⎝ C0 −qt m ⎠ ⎥⎦ ⎣

Intra-diffusion

7

Diffusion coefficient

8

Based on diffusion

t(1/2)=

0.030r 2 Dp

t(1/2)=

0.230rδ Cs × Df Ce

Diffusivity model

F=1−

( ) 6

π2

⎛ De tπ 2m 2 ⎞ ⎛ 1 ⎞ b ⎟ ∑1m ⎜ 2 ⎟ exp ⎜⎜ − ⎟ b ⎝ mb ⎠ r2 ⎝ ⎠

Chemisorption

(3) The third stage is the biosorption of adsorbate onto the surface sites. Due to its very fast in nature, it cannot be considered for the rate determining step.

4. Results and discussion

So as to know the rate controlling step, the following different models (Table 3) were applied using the experimental data of kinetic study (The details of mechanistic model equations are provided with the supplementary material).

The acid treated GAC itself has the lower capacity to adsorb As(III) and As(V). Its main function was to provide a template with high specific area for MnFe2O4 loading. Untreated GAC (GAC), utilized in water treatment facilities, occupies mainly negatively charged surface due to hydroxyl group at neutral pH and therefore was not a good adsorbent for negatively charged/neutral arsenic. GAC is rich in hydroxyl groups.The hydroxyl groups can offer chemical reaction sites and adsorb iron and manganese ions to grow MnFe2O4 particles. The reason may be that the template can prevent MnFe2O4 particles from aggregating in adsorption process and enhance the effective adsorption area, resulting in the highly enhancement of adsorption capacity (Mohan and Pittman, 2007). Loading of MnFe2O4 increased the positive charge density on the adsorbent surface by neutralizing negative surface charge and creating positive charge in its place (Mondal et al., 2009). Mondal et al. (2007a, 2010) reported that impregnation of Ca on the surface of GAC also increased the positive charge on the surface of GAC and finally improved the adsorption capacity of GAC. By comparing the arsenic adsorption capacity of MnFe2O4 with and without the template in Table 5, it was observed the adsorption capacities of MnFe2O4 loading on the acid treated GAC are much higher than the bare ones. So, the introduction of acid treated GAC template can efficiently improve the adsorption capacity. The reason may be that the template can prevent MnFe2O4 particles from aggregating in adsorption process and enhance the effective adsorption area, resulting in the highly enhancement of adsorption capacity. A similar finding has been found out for the loading of Fe3O4 on pure wheat straw (Tian et al., 2011).

3.4. Adsorption thermodynamic modeling The entropy and Gibbs free energy parameters should be measured with the intention of deciding if the processes will happen spontaneously. Thermodynamic parameters for example ΔG0, ΔH0 and ΔS0 can be calculated utilizing equilibrium constant while the temperature varies (Table 4) and their evaluation gives an insight into the possible nature of adsorption (The details of thermodynamic equations are provided with supplementary material).

3.5. Adsorption activation energy The adsorption activation energy (Ea) was calculated by considering the equilibrium constants under the different experimental conditions (Table 4) (The detailed equation is provided with supplementary material).

Table 4 Thermodynamic equations and their parameters. Sr. no. Expression

Equation form

1

∆G 0= − RT ln Kd where kd=

2

Gibbs free energy van’t Hoff

k d= 3

Arrhenius

4.2. Effect of contact time Remarks qe Ce

Free energy change

(80)

ln ( kd )= −

∆ H 0 ∆ S0 + RT R

where

Enthalpy change Entropy change

qe Ce

ln kPSO= −

Ea +ln RT

A

4.1. Effect chemical treatment on biosorption

Apparent activation energy

Contact time is also a significant factor for the biosorption/ bioaccumulation process. Fig. 1 represents the influence of contact time on the % removal of As(III) and As(V) using immobilized bacterial cells. The time required to attain equilibrium for both As (III) and As(V) biosorption/bioaccumulation on immobilized bacterial cells was 180 min. With further increase in time, no noteworthy improvement was found in scavenging arsenic. So, further biosorption/bioaccumulation studies were continued for a contact time of 180 min.

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Table 5 Comparison of the % removal of As(III) and As(V) using different adsorbent/biosorbent. Adsorbents

Operating conditions

B. arsenicus MTCC 4380 immobilized on surface of GAC/MnFe2O4 composite Co: 50 mg/L; M: 4 g/L; pH: 7.0; T: 30 oC; t: 90 min GAC/MnFe2O4 composite Co: 50 mg/L; M: 4 g/L; pH: 7.0 (As(III)) and 4.0 (As (V)); T: 30 oC; t: 90 min Acid treated GAC Co: 50 mg/L; M: 4 g/L; pH: 10.0 (As(III)) and 7.0 (As(V)); T: 30 oC; t: 90 min Co: 50 mg/L; M: 4 g/L; pH: 7.0 (As(III)) and 4.0 (As Bare MnFe2O4 (V)); T: 30 oC; t: 90 min

% removal of As (III)

% removal of As (V)

86.522 88.696

91.769 91.231

83.478

85.077

66.087

57.385

the later stage of the process was due to the lowering of concentration of metal ions (Mishra et al., 2010). Furthermore, the characteristic of the equilibrium time curve exhibited that the SBB process approaches the equilibrium in short span of time (Mishra et al., 2013). Thus, the curves found were single, smooth and continuous leading to equilibrium and suggested the possibility of monolayer coverage of the adsorbate on the surface of immobilized bacterial cells (Ranjan et al., 2009). 4.3. Biosorption kinetic studies

Fig. 1. Effect of contact time on As(III) and As(V) removal in SBB studies (Co: 50 mg/ L; T: 30 °C; pH: 7.0; M: 0.7 g/L) (Error bars represent means 7 standard errors from the mean of duplicate experiments).

From the results it is further noticeable that in all the systems, the saturation time does not depend on the adsorbate concentration in the solution. The change in the rate of removal might be due to the fact that initially all sites of immobilized bacterial cells are easily obtainable and also the concentration gradient of adsorbate is very high. At optimum pH, the rapid kinetics of interaction of adsorbate-immobilized bacterial cells might be agreed to improve availability of the active sites of the immobilized bacterial cells. Therefore, the scavenging of adsorbate was rapid in the early stages and gradually decreases with the interval of time until equilibrium in each case. The decrease in removal of metal ion at

Information on the kinetics of adsorbate uptake is vital to identify optimum operating conditions for full scale batch process. As a non-linearizable kinetic model and with the purpose of comparing its fitting ability to the previous considered models, a nonlinear regression analysis was performed to fourteen biosorption/bioaccumulation kinetic models. Table S1 of supplementary materials displays the values of kinetic constants of all the models for the biosorption/bioaccumulation of both As(III) and As (V) by immobilized bacterial cells. The results showed that there was no noteworthy relationship between the kinetic data for both As(III) and As(V) (Fig. 2) with low correlation coefficients and high error values suggesting that these models (fractional power, pseudo first order, Elovich and exponential) are not appropriate in the current case. On the basis of maximum correlation coefficients (R, R2 and R̅2) and the lowest error values (SSE, Reduced χ2 and Root–MSE), the superior and perfect fitting of the experimental results was observed using Brouers–Weron–Sotolongo for both As(III) and As (V) among all the verified kinetic models. The equilibrium uptake, qe (67.532 mg/g and 70.653 mg/g for As(III) and As(V), respectively), acquired from Brouers–Weron– Sotolongo model was closely in line with the experimental value

Fig. 2. Kinetic modeling of As(III) and As(V) biosorption/bioaccumulation on B. arsenicus MTCC 4380 immobilized on surface of GAC/MnFe2O4 composite (t: 5–500 min; Co: 50 mg/L; M: 0.7 g/L; pH: 7.0; T: 30 °C) (Error bars represent means 7 standard errors from the mean of duplicate experiments).

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(67.391 mg/g and 70.495 mg/g for As(III) and As(V), respectively). Going to the theory beyond the best fitting model (i.e. BWS), the biosorption/bioaccumulation phenomenon of both As(III) and As (V) immobilized bacterial cells would be governed by chemisorption interactions type, revealing that the rate controlling step might be chemical biosorption/bioaccumulation including valency forces via exchange or sharing of electrons between As(III) or As (V) anions and immobilized bacterial cells. Also BWS model expresses another remarkable data which is the time important to adsorb half the maximum amount (τnBWS,α). With respect to initial arsenic concentration, 22.615 min and 13.756 min (lowest value) were enough for immobilized bacterial cells to achieve half of the As(III) and As(V) uptake capacities, respectively which is important as well as valuable parameter for computing the reaction speed. As for the BWS model itself, it has a good fitting behavior and more implicitly, it conveys such quality data (i.e. biosorption/bioaccumulation capacity nearest to experimental value, qe(BWS), and the time of half reaction, τnBWS,α) which are very important aimed at industrial treatment design purposes (Ncibi et al., 2014). On the basis of correlation coefficient (R, R2 and R̅2) and error values (SSE, Reduced χ2 and Root–MSE), it can be concluded that other kinetic models, such as Avrami, mixed 1,2 order, modified second order, Ritchie second order, Fractal–like pseudo first order, Fractal–like pseudo second order, Fractal–like exponential, and Fractal–like mixed 1,2 order exhibited good fitting of biosorption/ bioaccumulation kinetic data for both As(III) and As(V) on immobilized bacterial cells. The value of good correlation coefficients (R, R2 and R̅2) and low error values (SSE, reduced χ2 and Root–MSE) for pseudo second order model forecasts that mechanism of biosorption/bioaccumulation is chemisorption type. Avrami exponent (nAV) (0.575 for As(III) and 0.509 for As(V)) is a fractionary number connected with the possible variations of the biosorption/bioaccumulation mechanism that happens all over the biosorption/bioaccumulation process (Vaghetti et al., 2009). In place of following only an integer-kinetic order, the mechanism of biosorption/bioaccumulation could follow multiple kinetic orders that are altered throughout the contact of the adsorbate with immobilized bacterial cells (Vaghetti et al., 2009). nAV is a resultant of the multiple kinetic order of the biosorption/bioaccumulation procedure. It is understood that the mixed 1,2-order rate equation is linear in the Lagergren coordinates (i.e. it behaves like the first order equation) near to the equilibrium and somewhat in the initial portion of the experiment. Precisely, we may conclude that it contains two linear segments connected with a curved one. Actually, the second-order equation shows linear behavior in the initial portion of the experiment also, though the linear section is quite short. On the other hand, MOE may be may be treated as Langmuir equation for energetically homogeneous surfaces or as purely empirical equation for energetically heterogeneous surfaces. The vital assumption of modified second order model was that a number of surface sites, nR, are used by each adsorbate. It was supposed that the pre-adsorbed stage occurred on immobilized bacterial cells biosorption/bioaccumulation. The vital supposition of Ritchie second order model was that one adsorbate was adsorbed on two surface sites. It was supposed that the pre-adsorbed stage has not occurred on immobilized bacterial cells biosorption/bioaccumulation (Cheung et al., 2001). The biosorbent surface coverage has been typically presumed to be zero. In case of Fractal–like models (Fractal–like pseudo first order, Fractal–like pseudo second order, Fractal–like exponential, and Fractal–like mixed 1,2 order), the biosorption/bioaccumulation

61

rate coefficient is considered a function of time by using the fractal–like idea. One of the probable physical significances of fractal–like kinetics was that the biosorption/bioaccumulation of As(III) and As(V) occurred at solid/solution interface. In this approach, it was showed that by passing time, various paths for biosorption/bioaccumulation of As(III) and As(V) on surface appears (Haerifar and Azizian, 2012). 4.4. Final remarks on biosorption/bioaccumulation kinetic studies The value of correlation coefficient (R, R2 and R̅2) and error values (SSE, Reduced χ2 and Root–MSE) for Brouers–Weron–Sotolongo kinetic model was better than that observed using the other kinetic models for both As(III) and As(V). So, the kinetics of As(III) and As(V) biosorption/bioaccumulation using immobilized bacterial cells as an biosorbent can be well clarified by Brouers– Weron–Sotolongo kinetic model. It signifies that the mechanism of biosorption/bioaccumulation of both As(III) and As(V) is complex. The complex phenomenon of biosorption/bioaccumulation may include chemical interactions between the adsorbate and the functional groups on the biosorbent surface, which may involve hydrogen bounding, ligand exchange, Van der Waals, electrostatic interactions and hydrophobic interactions. Among the models pseudo second order model also possessed a good fitness with experimental kinetic data. This tendency comes as a suggestion that the rate limiting step in biosorption/ bioaccumulation of arsenic (As(III)/As(V)) are chemisorption involving valence forces through the sharing or exchange of electrons between immobilized bacterial cells and arsenic (As(III)/As (V)) ions (Miretzky et al., 2008), complexation, coordination and/ or chelation (Lu and Gibb, 2008). It is vital for obtaining the rate at which As(III) or As(V) is biosorbed/bioaccumulated on the surface of immobilized bacterial cells that is important to design fixed-bed biosorption/bioaccumulation column. Using the biosorption/bioaccumulation rate, kinetic constants are estimated to determine the equilibrium capacity of immobilized bacterial cells and mass transfer coefficient at various aqueous phase concentrations. Amount of As(III) or As (V) biosorbed/bioaccumulated on solid surface is calculated using the kinetic equation which is important for valuing the concentration of the aqueous phase in fixed-bed column operation. The main design factors of fixed-bed biosorption/bioaccumulation column, the breakthrough time as well as the shape of breakthrough curve are dependent on biosorption/bioaccumulation rate. For the faster biosorption/bioaccumulation rate, breakpoint time is achieved prior besides the breakthrough curve shape is steeper. The descriptive models from the best to worst for As(III) and As (V) were sorted according to GoFM values and shown Table S2 and Table S3 of supplementary materials, respectively. Among fourteen different isotherm models, Brouers–Weron–Sotolongo isotherm was observed to be appropriate to predict the kinetic data of biosorption/bioaccumulation of both As(III) and As(V) on immobilized bacterial cells according to GoFM values presented in the Table S2 and Table S3 of supplementary materials. According to correlation coefficient R̅2 values the fitness of the models for all kinetic models is almost equivalent to each other. So on the basis of equivalent biosorption/bioaccumulation capacity, the orders followed by the models in decreasing manner are shown in Table S4 of supplementary materials. From the Table S4 of supplementary materials it can be concluded that on the basis of equivalent biosorption/bioaccumulation capacity Fractal–like pseudo second order kinetic model is the best fitting model for both As(III) and As(V) owing to the highest value.

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4.5. Biosorption mechanistic studies 4.5.1. Intraparticle diffusion model Results of intraparticle model for both As(III) and As(V) are presented in Table S5 of supplementary materials. Fig. 3 specified the multi-linear nature of intraparticle model. Since the plot did not pass through the origin, thus intraparticle diffusion was not the only rate controlling step. Thus there were three processes governing the rate of biosorption/bioaccumulation, nonetheless only one was rate governing in any certain time range. The intraparticle diffusion rate constant kint2 for both As(III) and As(V) was estimated from the slope of the second linear portion. The multi-linear curve of the intraparticle model with intercept Cint2 indicated the fact that both intraparticle diffusion of adsorbate through the mesoporus openings filled with liquid and film/external mass transfer across the thickness of boundary layer were the rate controlling steps in the biosorption/ bioaccumulation of both As(III) and As(V) kinetics on the surface of immobilized bacterial cells. The value of intercept Cint obtained through the model delivered the value of thickness of boundary layer of liquid surrounding the immobilized bacterial cells (Kavitha and Namasivayam, 2007). The interpretation of the random data points observed in Fig. 3 stated that intraparticle mechanistic model was suitable to costeffectively and skilfully define the biosorption/bioaccumulation of both As(III) and As(V) on the surface of immobilized bacterial cells

in terms of linear regression coefficient Adjusted R–square ( R̅2), ranging between 0.923 and 0.993 in both case. Larger value of intercepts achieved for second linear portion i.e. Cint2 recommended that the film diffusion had played a superior role as the rate controlling step (Rengaraj et al., 2007). 4.5.2. Determination of rate limiting step The values of the film diffusion coefficient D1 and the pore diffusion coefficient D2 for both As(III) and As(V) are shown in Table S5 of supplementary materials. The high negative exponential of almost nearest values stated that both pore diffusion and film diffusion had governed the mechanism of biosorption/ bioaccumulation for both As(III) and As(V). 4.5.3. Dumwald–Wagner model The plot (Fig. 4 of log (1–F2) versus t has provided almost linear curves ( R̅2 Z0.956) for the removal of both As(III) and As(V) by immobilized bacterial cells, respectively but did not pass through the origin signifying that the diffusion of adsorbate into pores of the immobilized bacterial cells was not the only rate limiting step in both case. Table S5 of supplementary materials shows the calculated results of Dumwald–Wagner model. In the present research the linearity of the plots intersected the origin of coordinates; so film diffusion process was the rate governing step.

Fig. 3. Intraparticle diffusion modeling of As(III) and As(V) biosorption/bioaccumulation on B. arsenicus MTCC 4380 immobilized on surface of GAC/MnFe2O4 composite (t: 5– 500 min; Co: 50 mg/L; M: 0.7 g/L; pH: 7.0; T: 30 °C) (Error bars represent means 7 standard errors from the mean of duplicate experiments).

Fig. 4. Dumwald–Wagner modeling of As(III) and As(V) biosorption/bioaccumulation on B. arsenicus MTCC 4380 immobilized on surface of GAC/MnFe2O4 composite (t: 5– 500 min; Co: 50 mg/L; M: 0.7 g/L; pH: 7.0; T: 30 °C) (Error bars represent means 7 standard errors from the mean of duplicate experiments).

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4.5.4. Richenberg model or Boyd plot The results of the corresponding model for both As(III) and As (V) are showed in Table S5 of supplementary materials and Fig. 5, respectively. From the plots of Bbt versus t, it was observed that the plots had dynamic linear form with a correlation coefficient Adjusted R–square ( R̅2) of 0.975 for As(III), 0.956 for As(V) at studied temperature, but did not pass through origin which had chosen the ongoing biosorption/bioaccumulation processes to be controlled by film diffusion mechanism. 4.5.5. Film diffusion mass transfer rate equation or McKay plot The values achieved for ln(1 F) as a function of time t, were plotted in Fig. 6 for As(III) and As(V), respectively and the outcomes of the respective model for both As(III) and As(V) are shown in Table S5 of supplementary materials. The rate constant of the initial quick process (kM1) was assessed from the slope of the first straight line. As can be understood from Fig. 6, the rate constant of the sluggish process (kM2) was achieved from the slope of the second linear portion. It is obvious that the initial rapid process was governed by film diffusion. Moreover, the values of kM2 are smaller than the values of kM1 for both As(III) and As(V). Consequently the acquired values of kM2 were proof of a pore diffusion mechanism in the second stage of the biosorption/bioaccumulation (Atun and Hisarli, 2003). The presence of two straight lines for

63

both As(III) and As(V) specified that two processes i.e. film diffusion and pore diffusion had engaged in these processes. The plots were approximately linear for both As(III) and As(V), but it did not return perfect linearity. So, pore diffusion was not the rate governing step. In the present system film diffusion was the rate controlling step.

4.5.6. Bangham”s model The outcomes of the respective model for both As(III) and As (V) are revealed in Table S5 of supplementary materials and Fig. 7. The goodness of fit of curve for Bangham”s model was related in terms of correlation coefficient Adjusted R–square ( R̅2). Values of Bangham parameters kb and α are 29.943 L/g and 0.5 for biosorption/bioaccumulation of As(III) and 40.915 L/g and 0.493 for biosorption/bioaccumulation of As(V) with a correlation coefficients of 0.996 and 0.984, respectively. The double logarithmic plot as stated by above equation did not yield perfect linear curves and some data were scattered ( R̅2 r0.996) for the scavenging of As(III) and As(V) by immobilized bacterial cells clarifying that the diffusion of adsorbate within pores of the by immobilized bacterial cells was not only the rate controlling step (Weber and Morris, 1963), film diffusion had influence on the rate limiting step.

Fig. 5. Richenberg modeling of As(III) and As(V) biosorption/bioaccumulation on B. arsenicus MTCC 4380 immobilized on surface of GAC/MnFe2O4 composite (t: 5– 500 min; Co: 50 mg/L; M: 0.7 g/L; pH: 7.0; T: 30 °C) (Error bars represent means 7 standard errors from the mean of duplicate experiments).

Fig. 6. McKay plot of As(III) and As(V) biosorption/bioaccumulation on B. arsenicus MTCC 4380 immobilized on surface of GAC/MnFe2O4 composite (t: 5–500 min; Co: 50 mg/ L; M: 0.7 g/L; pH: 7.0; T: 30 °C) (Error bars represent means 7 standard errors from the mean of duplicate experiments).

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4.5.7. Determination of Diffusion coefficient The Dp and Df values for As(III) were 1.283  10  7 cm2/s and 2.783  10  7 cm2/s, respectively. Likewise these values for As (V) were 1.808  10  7 cm2/s and 1.773  10  6 cm2/s, respectively (Table S5 of supplementary materials). Thus in the present case pore diffusion was not the only rate controlling step for both As(III) and As(V). For As(III) the value of Df (2.783  10  7 cm2/s) falls within the range of 10  6 and 10  8 cm2/s and for As(V) the Df value (1.773  10  6 cm2/s) is also in the range. So, it was noticed that the film diffusion was the rate controlling step for both As(III) and As (V). Higher value of Df of As(V) than that of As(III) was due to the attendance of As(V) as totally positively charged species and As(III) as typically neutral species at the experimental pH. Owing to negative charge of As(V) it was freely transported from the bulk solution to mostly the positively charged immobilized bacterial cells and got biosorbed/bioaccumulated on the active sites of immobilized bacterial cells instead of the interior pores. 4.5.8. Determination of diffusivity From the slope π2De/r2 of the plot of ln[1/(1  F2)] versus t (Fig. 8; Table S5 of supplementary materials), the value of diffusion coefficient, De as evaluated, was observed to be 6.863  10  14 m2/s and 6.818  10  14 m2/s for As(III) and As(V) biosorption/bioaccumulation on immobilized bacterial cells, respectively. For the present systems, the value of De falls within the range of 10  9 to 10  17 m2/s, so the system was chemisorption system. The values of the diffusion coefficient, De fell well within the magnitudes reported in literature (Naiya et al., 2009; Singha and Das, 2011). 4.6. Final remarks on biosorption/bioaccumulation mechanistic studies Biosorption/bioaccumulation kinetics mechanism can determined from the mechanistic study. The above considered models for biosorption/bioaccumulation of both As(III) and As (V) indicated that two processes i.e. film diffusion (diffusion of adsorbate through the solution to the external surface of biosorbent or boundary layer diffusion of adsorbate) and pore diffusion (diffusion of the adsorbate from the surface film into the pores) were tangled in the current biosorption/bioaccumulation processes for both As(III) and As(V) and pore diffusion was not the single rate governing step. Generally, film diffusion controlled the rate monitoring step in both cases. The uptake of arsenic (either As (III) or As(V)) species from the liquid to the solid phase was carried out in three consecutive steps (Al-Degs et al., 2008). Firstly, film

diffusion occurred. Secondly, pore diffusion took place. The third step is the biosorption/bioaccumulation which is being very rapid in nature and cannot be taken into account for the rate determining step (Singh and Pant, 2006). 4.7. Effect of temperature The temperature has two major effects on the biosorption/ bioaccumulation process. Temperature dependence of the biosorption/bioaccumulation system dictates the biosorption/bioaccumulation as endothermic or exothermic. Increasing the temperature is recognized to increase the rate of diffusion of the adsorbate, owing to the decrease in the viscosity of the solution. Besides varying the temperature will change the equilibrium biosorption/bioaccumulation capacity of the immobilized bacterial cells for a specific biosorbent (Nouri et al., 2007). The effect of temperature on the scavenging efficiency of As(III) and As(V) was inspected in the range of 20–50 °C throughout the equilibrium time. The results of effect of temperature have been revealed in Fig. 9. The favorable temperature for the growth of B. arsenicus MTCC 4380 is 30 °C (Shivaji et al., 2005). The result indicated that the maximum removal was reached at a temperature of 30 oC in biosorption/bioaccumulation system. Further increase in temperature resulted in lower removal efficiency for arsenic removal in biosorption/bioaccumulation system. This can be clarified by the exothermicity of the biosorption/ bioaccumulation process. Temperature influences the interaction between the biomass and the metal ions, generally by impacting the stability of the metal-sorbent complex and the ionization of the cell wall moieties (Sag and Kutsal, 1995). The temperature of the biosorption/bioaccumulation solution could be vital for energy dependent mechanisms in metal binding process (Green-Ruiz et al., 2008). Energy independent mechanisms are less expected to be influenced by temperature because the processes responsible for biosorption/bioaccumulation are mainly physico-chemical in nature (Kacar et al., 2002; Ozdemir et al., 2009). The results acquired in the current study indicated that the metal ion sorption in SBB system was exothermic. Kacar et al. (2002) and Ozdemir et al. (2009) reported in their work about the exothermic behavior of metal ion sorption on the bacterial surface. The authors correlated their research outcomes as an energy independent mediated sorption of metal ions. This can also be clarified by the spontaneity of the biosorption/ bioaccumulation process. Initial rise in removal efficiency up to

Fig. 7. Bangham modeling of As(III) and As(V) biosorption/bioaccumulation on B. arsenicus MTCC 4380 immobilized on surface of GAC/MnFe2O4 composite (t: 5–500 min; Co: 50 mg/L; M: 0.7 g/L; pH: 7.0; T: 30 °C) (Error bars represent means 7 standard errors from the mean of duplicate experiments).

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Fig. 8. Diffusivity of (a) As(III) and (b) As(V) biosorption/bioaccumulation on B. arsenicus MTCC 4380 immobilized on surface of GAC/MnFe2O4 composite (t: 5–500 min; Co: 50 mg/L; M: 0.7 g/L; pH: 7; T: 30 °C) (Error bars represent means 7 standard errors from the mean of duplicate experiments).

might be due to many issues: the relative increase in the avoidance tendency of the arsenic ions from the solid phase of immobilized bacterial cells to the bulk liquid phase; deactivating the surface of immobilized bacterial cells or destructing some active sites on the surface of immobilized bacterial cells because of bond ruptures (Meena et al., 2005; Romero-Gonz´alez et al., 2005) or due to the weakening of adsorptive forces between the adsorbate species and the active sites of the immobilized bacterial cells and also between the adjacent molecules of biosorbed/bioaccumulated phase for high temperatures or movement of immobilized bacterial cells with higher speed, so lower interaction time with the active sites of immobilized bacterial cells was obtainable for them (Yadav and Tyagi, 1987; Roy et al., 2014). 4.8. Biosorption thermodynamic studies

Fig. 9. Effect of temperature on As(III) and As(V) removal in SBB studies (Co: 50 mg/ L; pH: 7.0; M: 0.7 g/L; t: 180 min) (Error bars represent means7 standard errors from the mean of duplicate experiments).

30 °C is mainly because of the rise in collision frequency between biosorbent and adsorbate. Further rise in temperature (430 °C) resulted in lower removal efficiency for arsenic removal by immobilized bacterial cells. This decrease in scavenging efficiency

The evaluated values of the thermodynamic parameters from the plot of ln kd versus 1/T (Fig. 10) are represented in Table 6. The equilibrium constant kd was calculated; while the temperature was changed between 20 and 50 °C for both As(III) and As(V). The maximum removal of adsorbate was attained at 30 °C in both case. The scavenging efficiency firstly increased with increasing the temperature from 20 to 30 °C. Then, it decreased with increasing the temperature from 30 to 50 °C. The ΔH0 values acquired for the biosorption/bioaccumulation of both As(III) and As(V) were negative owing to the exothermic nature of the biosorption/

Fig. 10. Thermodynamic modeling of (a) As(III) and (b) As(V) biosorption/bioaccumulation on B. arsenicus MTCC 4380 immobilized on surface of GAC/MnFe2O4 composite (T: 20–50 °C; Co: 50 mg/L; pH: 7.0; M: 0.7 g/L; t: 180 min) (Error bars represent means 7 standard errors from the mean of duplicate experiments).

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Table 6 Thermodynamic constants for As(III) and As(V) biosorption/bioaccumulation on B. arsenicus MTCC 4380 immobilized on the surface of GAC/MnFe2O4 composite. Inorganic form

T (K)

 ΔG0 (KJ/mol)

-ΔH0 (KJ/mol)

 ΔS0 (J/mol K)

As(III)

293 298 303 308 313 318 323

23.855 24.595 25.391 25.076 24.622 24.325 24.121

24.934

0.001

As(V)

293 298 303 308 313 318 323

26.296 27.545 29.192 27.642 27.448 26.906 26.934

28.805

0.004

bioaccumulation process. The value attained for As(V) biosorption/ bioaccumulation was more than the value attained for As(III). A negative value of ΔG0 stated the spontaneous nature of the biosorption/bioaccumulation process; however the negative value of ΔS0 presented a reduction in the randomness at the solid/solution interface during the biosorption/bioaccumulation process (Ngah and Hanafiah, 2008). Higher negative value of ΔG0 at a temperature of 30 °C, as was attained in the study, concluded more driving force for biosorption/bioaccumulation at 30 °C (Crini and Badot, 2008). The values of ΔG0 achieved in the present study were between  23.855 and  25.391 KJ/mol and 26.296 and  29.192 KJ/mol for As(III) and As(V), respectively. So, it specified that the mechanism of present biosorption/bioaccumulation process had occurred through ion exchange and/or surface complexation mechanism. 4.9. Effect of initial arsenic concentration on biosorption/bioaccumulation kinetics To examine the effect of initial concentration on kinetic studies, a series of contact time experiments for both As(III) and As(V) was conducted at various initial concentration (50–2000 mg/L) at temperature of 30 °C. Fig. 11 exhibited the contact time essential for both As(III) and As(V) with initial concentration of 50– 1000 mg/L for attaining equilibrium was 180 min. But for both As (III) and As(V) with higher initial concentration ( 41000– 2000 mg/L), higher equilibrium time of 260 min. was needed. As can be decided from Fig. 11 the amount of the biosorbed/

bioaccumulated both As(III) and As(V) on the surface of immobilized bacterial cells changed with time and at some point in time attained a constant value beyond which no more As(III) or As (V) was removed from solution. At the moment, the amount of the both As(III) and As(V) desorbing from the biosorbent was in a state of dynamic equilibrium with the amount of the both As(III) and As (V) being biosorbed/bioaccumulated on the surface of immobilized bacterial cells. The time essential for attainment of this state of equilibrium is called the equilibrium time. Thus, the rate of biosorption/bioaccumulation declined with time till it slowly inclined to a plateau because of the constant reduction in the concentration driving force. 4.10. Determination of initial sorption rate From Fig. 12 and Table 7, it is noticed that the pseudo second order rate constants (kPSO) were observed to reduce and the initial sorption rates (h) were understood to increase from 6.083 to 1319.303 mg/g min and 9.565 to 1470.811 mg/g min with increased initial concentration of As(III) and As(V) from 50 to 2000 mg/L, respectively. The initial sorption rate was superior for higher initial As(III) and As(V) concentration, as the resistance to uptake both As(III) and As(V) reduced, as the mass transfer driving force improved. The evaluated qe(PSO) values decided just suitable with the experimental data and high correlation coefficient (R, R2 and R̅2) and low error values (SSE, Reduced χ2 and Root–MSE) shown that the model can be used for the whole biosorption/bioaccumulation process and permitted the chemisorption of both As(III) and As (V) on immobilized bacterial cells. 4.11. Effect of temperature on biosorption/bioaccumulation kinetic Temperature is an important factor leading the biosorption/ bioaccumulation process. The effect of temperature on the of biosorption/bioaccumulation of both As(III) and As(V) using immobilized bacterial cells was done from 30 to 50 °C at C0 is 50 mg/ L and GAC/MnFe2O4 composite loading is 0.7 g/L. A decreasing biosorption/bioaccumulation rate of both As(III) and As(V) with increasing temperature from 30 to 50 °C recognized the process to be exothermic (Fig. 13). The time needed to reach biosorption/ bioaccumulation equilibrium is 180 min at temperatures ranging between 30 °C and 50 °C. The % removal of both As(III) and As (V) was however comparatively unaffected to temperature, reducing from 94.348 at 30 °C to 84.783 at 50 °C for As(III) and from 98.692 at 30 °C to 94.077 at 50 °C for As(V).

Fig. 11. Effect of initial concentration on contact time for As(III) and As(V) biosorption/bioaccumulation on B. arsenicus MTCC 4380 immobilized on surface of GAC/MnFe2O4 composite (Co: 50–2000 mg/L; t: 5–500 min; pH: 7.0; M: 0.7 g/L; T: 30 °C) (Error bars represent means 7 standard errors from the mean of duplicate experiments).

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Fig. 12. Determination of the initial sorption rate for biosorption/bioaccumulation of As(III) and As(V) on B. arsenicus MTCC 4380 immobilized on surface of GAC/MnFe2O4 composite (Co: 50–2000 mg/L; t: 5–500 min; pH: 7.0; M: 0.7 g/L; T: 30 °C) (Error bars represent means 7 standard errors from the mean of duplicate experiments). Table 7 Initial sorption rate for As(III) and As(V) biosorption/bioaccumulation on B. arsenicus MTCC 4380 immobilized on the surface of GAC/MnFe2O4 composite. Inorganic form

Concentration (g/L)

kPSO (g/mg min)

h (mg/g min)

As(III)

50 100 500 1000 1500 2000

1.250E  03 8.886E  04 9.326E  04 5.757E  04 4.753E  04 3.558E  04

6.083 16.453 361.194 770.585 1174.918 1319.303

50 100 500 1000 1500 2000

1.850E  03 9.617E  04 9.556E  04 6.420E  04 4.227E  04 3.474E  04

9.565 19.206 399.216 959.031 1210.600 1470.811

As(V)

4.12. Biosorption activation energy Biosorption/bioaccumulation rate constants kPSO of both As(III) and As(V) were calculated from experimental data at a different temperatures supposing non-linear form of pseudo second order kinetics. Parameters of Arrhenius equation were fitted using these

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Fig. 14. Determination of the activation energy for As(III) and As(V) biosorption/ bioaccumulation on B. arsenicus MTCC 4380 immobilized on surface of GAC/MnFe2O4 composite (T: 30–50 °C; t: 5–500 min; Co: 50 mg/L; pH: 7.0; M: 0.7 g/ L) (Error bars represent means 7 standard errors from the mean of duplicate experiments).

rate constants for estimating temperature independent rate parameters and biosorption/bioaccumulation type. A plot of ln kPSO versus 1/T showed a straight line with slope  Ea/R (Fig. 14; Table 8). The magnitude of activation energy for As(III) and As (V) biosorption/bioaccumulation on the immobilized bacterial cells were 11.238 KJ/mol and 10.204 KJ/mol suggesting that both the biosorption/bioaccumulation of As(III) and As(V) on the surface of the immobilized bacterial cells was activated chemisorption. 4.13. Characterization 4.13.1. FT-IR Analysis The biosorption/bioaccumulation capacity of heavy metal on various biosorbents depends on the existence of various active functional groups on the surface of biosorbent. The Fourier Transform Infrared spectra (FT-IR) study of fresh GAC/MnFe2O4 composite and As(III) acclimatized immobilized bacterial cells at unloaded and metal loaded stage (Fig. 15(a)) and fresh GAC/MnFe2O4 composite and As(V) acclimatized immobilized bacterial cells at unloaded and metal loaded stage (Fig. 15(b)) at optimized batch experimental condition were examined for detecting the functional groups responsible mainly for the process of biosorption/bioaccumulation. Table S6 and Table S7 of supplementary materials indicate the wavenumber for various functional

Fig. 13. Effect of temperature on contact time for biosorption/bioaccumulation of As(III) and As(V) on B. arsenicus MTCC 4380 immobilized on surface of GAC/MnFe2O4 composite (T: 30–50 °C; t: 5–500 min; Co: 50 mg/L; pH: 7.0; M: 0.7 g/L) (Error bars represent means 7 standard errors from the mean of duplicate experiments).

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Table 8 Activation energy for As(III) and As(V) biosorption/bioaccumulation on B. arsenicus MTCC 4380 immobilized on the surface of GAC/MnFe2O4 composite. Inorganic form

T (K)

kPSO (g/mg min)

Ea (KJ/mol)

As(III)

303 313 323 303 313 323

1.250E  03 1.350E 03 1.650E  03 1.850E  03 2.000E  03 2.380E  03

11.238

As(V)

10.204

groups existing in the fresh GAC/MnFe2O4 composite and As(III) acclimatized immobilized bacterial cells at As(III) unloaded and loaded stage and fresh GAC/MnFe2O4 composite and As (V) acclimatized immobilized bacterial cells at As(V) unloaded and loaded stage, respectively. Surface –OH and –NH groups were main active functional groups responsible for biosorption/bioaccumulation as the wavenumber shifted from 3442.481 cm  1 to 3457.985 cm  1 (As(III)) and from 3437.519 cm  1 to 3457.985 cm  1 (As(V)) which may be probably owing to the complexation of –OH groups with As(III) or As(V) ions (Allievi et al., 2011; Singha and Das, 2011). Some researchers have also stated that after the adsorption arsenic on the Fe–Ce and Fe–Mn adsorbents, the peak of hydroxyl groups reduced or vanished (Zhang et al., 2005, 2007b). Aliphatic C– H stretching may be liable for biosorption/bioaccumulation of As(III) and As(V) on As(III) and As(V) acclimatized B. arsenicus MTCC 4380

Fig. 15. FT-IR spectra of (a) fresh GAC/MnFe2O4 composite (MGAC), As(III) acclimatized immobilized bacterial cells before and after As(III) biosorption/bioaccumulation (b) fresh GAC/MnFe2O4 composite (MGAC), As(V) acclimatized immobilized bacterial cells before and after As(V) biosorption/bioaccumulation.

immobilized on surface of GAC/MnFe2O4 composite as wavenumber shifted from 2940 cm  1 to 2930 cm  1 and from 2919.69 cm  1 to 2945.736 cm  1, respectively possibly owing to the complexation of C–H stretching vibration of alkyl chains. Table S6 and Table S7 of supplementary materials also display the accountability of aliphatic acid C¼O stretching for As(III) and As(V) biosorption/bioaccumulation by shifting the wavenumber from 1740 cm  1 to 1760 cm  1 and from 1744.496 cm  1 to 1764.961 cm  1, respectively (Singha, 2011). The next biosorption/bioaccumulation peaks at 1650 cm  1 shifted to 1660 cm  1 for As(III) and 1640 cm  1 shifted to 1680 cm  1 for As(V), perhaps owing to the complexation of amide group (N–H stretching and C¼O stretching vibration) with As(III) and As(V) ions (Seki et al., 2005). Table S6 and Table S7 of supplementary materials also show the intense bands at 1550 cm  1 which shifted to 1530 cm  1 (As(III)) and at 1573.953 cm  1, shifted to 1542.326 cm  1 (As(V)) that indicated the responsibility of aromatic –NO2 group for the biosorption/bioaccumulation process, respectively. Another shift was observed from 1449.302 cm  1 to 1440 cm  1 (As(III)) and from 1449.302 cm  1 to1490.853 cm  1 (As (V)), corresponding to the complexation of nitrogen with As(III) and As(V) ions of the N–H group (François et al., 2012). Wavenumber shifted from 1380 cm  1 to 1390 cm  1 (As(III)) and from 1340 cm  1 to 1377.364 cm  1 (As(V)) assigned the reactivity of carboxylate anion C¼O stretching for the biosorption/bioaccumulation process (Baig et al., 2010). Wavenumber 1230 cm  1 shifted to 1237.829 cm  1 and 1232.248 cm  1 shifted to 1237.209 cm  1 assigned for –SO3 stretching for the biosorption/bioaccumulation of As(III) and As(V), respectively. The peaks at 1071.628 cm  1 (As(III)) and 1082.171 cm  1 (As(V)) may be attributed to the C–N stretching vibrations of amino groups which shifted to higher frequency and appeared at 1100 cm  1 and 1130 cm  1, respectively due to the interaction of nitrogen from the amino group with As(III) and As (V) ions (Giri et al., 2011). Wavenumber 1060 cm  1 shifted to 1040 cm  1 (As(III)) and shifting from 1060 cm  1 to 1050 cm  1 (As (V)) indicated the Si–O stretching, active for the biosorption/bioaccumulation process. The other weak biosorption/bioaccumulation peak/bioaccumulation shifted from 890.543 cm  1 to 875.039 cm  1 (As(III)) and from 885.581 cm  1 to 870.078 cm  1 (As(V)), corresponding to the O–C–O scissoring vibration of polysaccharide (Pokhrel and Viraraghavan, 2007). The band at 592 cm  1 (Fig. 15(a)) and 590.388 cm  1 (Fig. 15(b)) could be attributed to the existence of Fe–O bond (McCafferty, 2010; Ren et al., 2012), but then it shifted to 596 cm  1 after biosorption/bioaccumulation of As(III) (Fig. 15(a)) and to 615.814 cm  1 after biosorption/bioaccumulation of As(V) (Fig. 15(b)), respectively. A typical peak at 563 cm  1 (Fig. 15(a)) and 564.341 cm  1 (Fig. 15(b)) could be assigned to Mn–O bond (Kohler et al., 1997; Parida et al., 2004) and it had a different variability to 569 cm  1 (Fig. 15(a)) and 574.884 cm  1 (Fig. 15(b)) for biosorption/ bioaccumulation of As(III) and As(V), respectively. The change in wavenumber of Me–O bonds after biosorption/bioaccumulation of As(III) and As(V) indicated that both Fe–O and Mn–O bonds were responsible for both MnFe2O4–As(III) and MnFe2O4–As(V) (Li et al., 2010; Ren et al., 2012). Presence of As(III) and As(V) on the surface of immobilized bacterial cells can be assured from the bands appeared at 797.519 cm  1 (Fig. 15(a)) and 828.527 cm  1 (Fig. 15(b)), respectively (Mondal et al., 2007b; Aryal et al., 2010). It has to be cited here, that a clear band was very hard to be got in the case of As (III), compared with the distinctive band of As(III) found at 797.519 cm  1 and As(V) found at 828.527 cm  1. This may be because of different mechanisms involved in As(III) and As (V) biosorption/bioaccumulation. It should be distinguished that the As–O band after biosorption/bioaccumulation of arsenic was not clearly observed because of the broad overlapping peaks in this region (Li et al., 2010).

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Fig. 16. Scanning electron micrographs (SEM) (500  ) and EDX of (a) fresh GAC/MnFe2O4 composite, (b) As(III) acclimatized Bacillus arsenicus MTCC 4380 immobilized on GAC/MnFe2O4 composite at unloaded stage, (c) As(III) acclimatized Bacillus arsenicus MTCC 4380 immobilized on GAC/MnFe2O4 composite at loaded stage, (d) As (V) acclimatized Bacillus arsenicus MTCC 4380 immobilized on GAC/MnFe2O4 composite at unloaded stage and (e) As(V) acclimatized Bacillus arsenicus MTCC 4380 immobilized on GAC/MnFe2O4 composite at loaded stage.

4.13.2. SEM-EDX analysis The SEM images of the prepared fresh GAC/MnFe2O4 composite (Fig. 16(a)) and As(III) acclimatized immobilized bacterial cells at stage of unloaded and loaded with As(III) (Fig. 16(b) and (c)) and As(V) acclimatized immobilized bacterial cells at stage of unloaded and loaded with As(V) (Fig. 16(d) and (e)) were presented. It can be observed from Fig. 16(a), manganese ferrite (MnFe2O4) particles with several diameters were randomly distributed onto the acid treated GAC surface. It is evident from Fig. 16(b) and (d) that most of the active sites of GAC/MnFe2O4 composite were shielded because of the formation of biofilm on it (Mondal et al., 2008). The bacterial mass partially occupies the void spaces of the biosorbent surface, so formation of bio-layer on the surface of biosorbent decreases its surface porosity (Mondal et al., 2008). It can be observed from Fig. 16(b) and (d), the immobilized bacterial cells have bulky structure with no porosity. A change in surface morphology from being smooth to rough and occupation of pores specified the As(III) and As(V) biosorption/bioaccumulation on the surface and pores of GAC/MnFe2O4 composite giving it a rough texture (Fig. 16 (c) and (e)) (Agarwal et al., 2013). The corresponding EDX spectra of the unloaded and loaded immobilized bacterial cells was collected and given in Fig. 16(a)– (e). The presence of iron, manganese and oxygen on the unloaded composite surface and iron, manganese and oxygen, arsenic on the loaded immobilized bacterial cells were exposed evidently. This outcome again established the occurrence of MnFe2O4 particles on the acid treated GAC surface as well as biosorption/bioaccumulation of arsenic on the surface of immobilized bacteria cells.

5. Conclusions

 The present research exhibited that the B. arsenicus. MTCC 4380

   



immobilized on surface of GAC/MnFe2O4 composite was applied successfully for the biosorption/bioaccumulation of both As(III) and As(V) from synthetically prepared copper smelting wastewater. The optimum contact time and temperature for biosorption/ bioaccumulation of both As(III) and As(V) were 180 min and 30 °C, respectively. Contact time needed for attainment of equilibrium enhanced with rising concentration, though remained practically unaffected by increasing temperature. The rate of biosorption/bioaccumulation of both As(III) and As(V) by immobilized bacterial cells reduced with rising concentration. By applying 14 different kinetic models and employing method of the nonlinear regression for curve fitting analysis (maximizing the correlation coefficient (R, R2, R̅2) and minimizing the error values (SSE, Reduced χ2 and Root–MSE) to estimate optimum parameter sets, the Brouser–Weron–Sototlongo was found suitable to predict the kinetic of biosorption/bioaccumulation of both As(III) and As(V) on B immobilized bacterial cells according to GoFM values. The current experimental data also followed pseudo second order model very well, hence it can be concluded that chemisorption was the possible route of biosorption/bioaccumulation on immobilized bacterial cells.

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 The results attained from different mechanistic models speci

  

 



fied the controller of film diffusion over pore diffusion. The effective diffusivity estimated by using Vermeulen”s approximation was indicated that the interaction between arsenic ions (either As(III) or As(V)) and immobilized bacterial cells were chemical in nature. The related thermodynamic parameters showed that the biosorption/bioaccumulation of both As(III) and As(V) had been a feasible and exothermic process and spontaneous in nature. From Arrhenius equation, it was verified that the mechanism of biosorption/bioaccumulation of As(III) and As(V) by immobilized bacterial cells might be an ion exchange type. EDX analysis documented the presence of iron and manganese in the GAC/MnFe2O4 composite and it also confirmed the fact that both As(III) and As(V) was biosorbed/bioaccumulated on the immobilized bacterial cells. Spectroscopic studies (Fe–SEM and FT-IR) confirmed that ion exchange process was responsible for the uptake of arsenic (As (III) or As(V)) onto immobilized bacterial cells. B. arsenicus MTCC 4380 immobilized on surface of GAC/MnFe2O4 composite can be employed as a capable biosorbent for removal of both As(III) and As(V) from contaminated water sources. Therefore, the simultaneous biosorption and bioaccumulation is an eco-friendly technique for the arsenic removal and play a significant role for sustainable development. Groundwater can be treated in-situ or in biofilm reactor by applying this technique. The present technique can also be used by the rural poor is most efficient with low cost, sustainability, adaptation and adoption using locally available material, reagents and it requires very low maintenance.

Acknowledgment Our thanks to Indian Institute of Technology, Roorkee for providing necessary facilities and to Ministry of Human Resource Development, Government of India for financial support. The thoughtful comments by Prof. Dr. Prosun Bhattacharya, the Editorin-Chief, Prof. Dr. Jochen Bundschuh, the Editor-in-Chief and two anonymous reviewers are highly appreciated.

Appendix A. Supplementary material Supplementary data associated with this article can be found in the online version at http://dx.doi.org/10.1016/j.gsd.2016.05.005.

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