Chemical Engineering Journal 225 (2013) 128–137

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Chemical Engineering Journal

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Manganese oxide immobilized activated carbons in the remediation ofaqueous wastes contaminated with copper(II) and lead(II)

1385-8947/$ - see front matter � 2013 Elsevier B.V. All rights reserved.http://dx.doi.org/10.1016/j.cej.2013.03.083

⇑ Corresponding author. Tel.: +91 9862323015; fax: +91 389 2330834.E-mail address: diw_tiwari@yahoo.com (D. Tiwari).

Lalhmunsiama a,b, Seung Mok Lee a, Diwakar Tiwari b,⇑a Department of Environmental Engineering, Kwandong University, Gangneung, Gangwondo, Republic of Koreab Department of Chemistry, School of Physical Sciences, Mizoram University, Aizawl 796 004, India

h i g h l i g h t s

� Novel manganese dioxide nano-particles immobilized on activated carbons.� Morphology of materials was obtained by SEM images.� Materials showed enhanced capacity in Cd(II) and Pb(II) removal.� Column reactor operations conducted for remediation.

a r t i c l e i n f o

Article history:Received 24 January 2013Received in revised form 17 March 2013Accepted 19 March 2013Available online 26 March 2013

Keywords:Manganeseoxide immobilized activatedcarbonsActivated carbonsCu(II)Pb(II)SorptionBreakthrough curveKinetics

a b s t r a c t

The aim of the present investigation was to utilize the large specific surface area of activated carbons inthe impregnation of the manganese oxide as further to enhance the suitability of materials in the reme-diation of aquatic environment contaminated with two important heavy metal toxic ions viz., Cu(II) andPb(II). The activated carbons were obtained by exploiting two different agricultural by-product or wastematerials viz., rice hulls (AC-R) or areca nut waste (AC-N) enabled, perhaps, a cost effective treatment. Thein situ impregnation of manganese oxide within the pores and surfaces of ACs, therefore, was resulted toobtain the manganese immobilized ACs viz., MIAC-R and MIAC-N. These solid samples were character-ized by the SEM/EDX analytical methods. Further, these materials were assessed for the removal of Cu(II)and Pb(II) from aqueous solutions under the batch and column reactor operations. The results obtainedfrom batch experiment showed that an increase in sorptive pH was caused apparently to enhance thepercent uptake of Cu(II) and Pb(II). The kinetic data were followed with the pseudo-second order kineticmodels. Equilibrium modelling studies suggested that the data was fitted well to the Freundlich andLangmuir adsorption isotherms and the 1000 times increase in background electrolyte concentrationscould not affect significantly the uptake of these two ions. Further, the breakthrough data was obtainedby column experiment which was utilized to fit into non-linear Thomas equation. The results showedthat ACs and MIACs obtained from these agriculture by-product/wastes were found to be potential solidmaterials for the removal of Cu(II) and Pb(II) from aquatic environment.

� 2013 Elsevier B.V. All rights reserved.

1. Introduction

The remediation of aquatic environment contaminated withheavy metal toxic ions is one of the key issues around the globebecause of the acute toxicity possessed by these heavy metalstowards the aquatic-life, plants and animals including humanbeing. Moreover, the heavy metals pose severe health and environ-mental problems due to their non-biodegradability, persistency innature and tendency to accumulate within biosystem [1–3].Copper is an essential trace elements but potentially toxic at higherconcentrations. According to the WHO, the maximum acceptable

level of Cu(II) in drinking water was 1.5 mg/L. Copper mainly en-ters into the aquatic environment through the discharge of indus-trial effluent including the plating, mining, smelting, brassmanufacturing, electroplating, petroleum refining, Cu-based agro-chemical industries etc. [4,5]. Excessive intake of copper by humanbeing was reported to cause for severe mucosal irritation, hepaticand renal damage, capillary damage, gastrointestinal irritationand central nervous system irritation [6]. On the other hand Pb(II)is known to be non-essential toxic ions showed several bio-toxiceffects. It was reported that even at extremely low concentrations,it is highly toxic to organisms including humans. The maximumpermissible limit of Pb(II) in drinking water was as low as0.010 mg/L demonstrated by the WHO [7]. The main sources oflead were reported to be combustion of fossil fuels, smelting of

Lalhmunsiama et al. / Chemical Engineering Journal 225 (2013) 128–137 129

sulfide ores, process industries such as battery manufacturing,metal plating, industries engaged with lead–acid batteries, paint,oil, metal, phosphate fertilizers, electronic industry, and wood pro-duction. Apart from this, lead was also used in storage batteries,insecticides, plastic water pipes, food, beverages, ointments andmedicinal concoctions for flavoring and sweetening [8,9]. Leadpoisoning in human were caused severe damage to the kidney,nervous system, reproductive system, liver and brain. Severe expo-sure of lead was reported to be sterility, abortion, stillbirth andneonatal deaths etc. [10].

Therefore, because of the occurrence and persistency of thesetwo heavy metal toxic ions viz., copper(II) and lead(II) within theaquatic environment, efforts were intended to the low level re-moval of these ions from aquatic environment. Literature showedthat various methods were suggested for the removal of thesetwo ions included with precipitation, electroplating, evaporation,ion exchange, membrane separation, reverse osmosis, coagulation,adsorption and some biological processes. Among these methods,adsorption is one of most preferable, effective and economical forits capability to remove even trace amount of metal pollutantsfrom aqueous solutions [11,12]. Moreover, the use of waste mate-rials to be employed as sorbent materials may likely to enhance theoverall treatment process more viable and economical.

Rice husk [13] and areca nut waste [14] were abundantly avail-able as a waste in many countries around the globe. Rice huskswere usually used as a low-value energy resource, burned in thefield, or discarded, which were found to be unfavorable to theatmospheric environment [15]. Areca nut waste also can beregarded as low value biomass and so far there is no report forits possible application [16]. Therefore, the application of thesedead biomasses in obtaining the activated carbons and furthersuitable and selective modification may perhaps enhance its rolein the environmental bio-remediation also would alleviate theproblems of disposal and management of these waste by-productsin general.

Recent studies reveal that many nano-sized metal oxidesexhibit significant sorption capacity and selectivity towards sev-eral heavy metal toxic ions. The nano-particles of ferric oxides,manganese oxides, aluminium oxides, titanium oxides, magnesiumoxides, cerium oxides etc. were classified as the promising adsor-bents for heavy metal removal from aqueous systems [17–19]. Itwas further demonstrated that the manganese oxide was exploitedfor the sorption of cationic or even anionic pollutants fromaqueous solutions [20–23]. Previous studies indicated that thenano-sized manganese dioxide which was impregnated to the sur-face of sand were found to be efficient and effective in the removalof Cu(II), Pb(II), Mn(II), Cd(II), Cr(VI) etc. [24–26].

Therefore, keeping in view, the present work deals with the uti-lization of agriculture waste materials viz., rice hulls and areca nutwaste to obtain the activated carbons. Further, the large specificsurface area of activated carbons was used to impregnate the smallor even nano-sized manganese oxide as further to enhance thesuitability of materials in the remediation of aquatic environmentcontaminated with two important heavy metal toxic ions viz.,Cu(II) and Pb(II) under the batch and column reactor operations.

2. Experimental section

2.1. Materials

Rice hulls and the areca nut wastes were collected from Aizawl,Mizoram, India. Sulfuric acid and ammonia used were of AR gradeand the de-ionized water was further purified by the Milliporewater purification system (Milli-Q+). The chemicals viz., nitric acid,sodium hydroxide and copper(II) sulfate were obtained from the

Duksan pure Chemicals Co. Ltd., Korea and manganese nitrate asMn(NO3)2�6H2O; 97% extra pure and lead(II) nitrate both were ob-tained from the Junsei Chemical Co. Ltd., Japan.

2.2. Preparation of activated carbons (ACs)

The raw materials, rice husk and areca nut waste were washedwith de-ionized distilled water to remove any adhering impurities.Dried at room temperature and the solid materials were digestedin conc. H2SO4 and the content was kept at 120 �C for 2 h. Thecarbonized carbon obtained was washed with distilled water tillthe filtrate solution reached pH �4.0. Again dried at 70 �C andthe samples were titrated with small volume of NH3 solution toneutralize any excess acids. The sample was washed with distilledwater and completely dried at 70 �C. The sample was cooled atroom temperature and grounded to obtain fine powders. Further,this carbonized carbon was activated by using muffle furnace at800 �C in the N2 environment for 6 h and the activated carbon ob-tained from rice husk (AC-R) and areca nut waste (AC-N) were usedfor batch experiment as well for the preparation of manganeseoxide immobilized activated carbons (MIAC-R and MIAC-N).

2.3. Preparation of manganese oxide immobilized activated carbon

The manganese oxide immobilized activated carbon (MIAC) wasprepared by taking 60 g of AC in a round bottom flask with 100 mLof 0.025 M manganese nitrate solution (pH �9) and was kept in arotary evaporator at 60 �C at the rotating speed of 30 rpm. Further,almost 90% of the water was removed by applying slowly the vac-uum. The slurry was taken out in a beaker and kept in a dryingoven at 90 �C to dry the solids completely. Further, the samplewas kept for 2 h at 110 �C for the stabilization of immobilized man-ganese-oxide. Samples were taken out from the oven and cooled atroom temperature and washed with distilled water and dried againat 70 �C. These samples were then used for further investigations.

2.4. Characterization of the materials

Characterization of the samples were done by SEM/EDX data inwhich the surface morphology and composition of these solids ACand MIAC were obtained by using SEM (Scanning Electron Micro-scope) machine (FE-SEM-Model: SU-70, Hitachi, Japan) equippedwith energy dispersive X-ray spectroscopy EDX system. Further,the amount of manganese content in MIACs as well as in ACs sam-ples was determined by standard US EPA (3050B) method.

2.5. Speciation studies

The speciation studies was performed for Cu(II) and Pb(II) usingthe MINEQL+ (Version 4.5), a geochemical computer simulationprogram. The input parameters were taken as initial concentrationof Cu(II) as 10 mg/L and Pb(II) as 20 mg/L at a constant tempera-ture 25 ± 1 �C.

2.6. pHpzc determination of the solids

The pHpzc (point of zero charge) was determined using theknown method as was detailed before [27,28]. In brief, 500 mL ofdouble distilled water was added to an Erlenmeyer flask, cappedwith cotton and was slowly and continuously heated until boilingfor 20 min to expel out the dissolved CO2. The flask was cappedimmediately to prevent re-absorption of atmospheric CO2 bywater. Then, 50 mL of 0.01 mol/L NaCl solutions was prepared fromCO2 free water and the pH of each solution in each flask wasadjusted to pH values of 2, 4, 6, 8, 10 and 12 by adding 0.1mol/L HCl or 0.1 mol/L NaOH, solutions. Then, 0.15 g of the solid

130 Lalhmunsiama et al. / Chemical Engineering Journal 225 (2013) 128–137

sample was added and the final pH was measured after 24 h underagitation at temperature of 25 ± 1 �C. The pHpzc was taken as thepoint at which the curve crossed the line pHfinal equals to pHinitial.

2.7. Batch reactor operations

Batch experiments were carried out to obtain the adsorptiondata with the variation of sorptive pH, contact time, initial metalconcentration, and background electrolyte concentrations. Theadsorption of Cu(II) and Pb(II) were investigated by taking 0.25 gof AC or MIAC in 0.10 L of sorptive solution. The solution mixturewas equilibrated by using automatic shaker for 24 h at 25 ± 1 �C.The solution was then filtered by using 0.45 lm syringe filter,the final pH was measured and then the final metal concentrationwas obtained by using atomic absorption spectroscopy (FastSequential Atomic Absorption Spectrometer: Model AA240FS, Var-ian). Adjustment of the pH was done by the drop wise addition ofconc. HNO3 or conc. NaOH. The time dependence data was col-lected at different time intervals. Similarly, the concentrationdependence data was obtained by varying the initial sorptive con-centration from 1.0 mg/L to 14.0 mg/L (for Cu(II)) and from 2 to16 mg/L (for Pb(II)).

2.8. Column studies

The column experiments were performed using a glass column(1 cm inner diameter) packed with 1 g of AC-R/AC-N/MIAC-R/MIAC-N (kept middle in the column); below and above to this,1 g each of bare sand (30–60 BSS in size) was placed and then itwas packed with glass beads. The sorptive solution was pumpedupward from the bottom of the column using Acuflow Series II,High-Pressure liquid chromatograph; at a constant flow rate of1.0 mL/min. Effluent samples were then collected using Spectra/Chrom CF-2 fraction collector. The collected samples were againfiltered with 0.45 lm syringe filter and the total bulk sorptive con-centration was measured using AAS.

Fig. 1. SEM images of (a) AC-R, (b) AC

3. Results and discussion

3.1. Characterization of solids

The SEM images of AC-R, AC-N, MIAC-R and MIAC-N wereobtained and presented in Fig. 1. It was reported that the nano-sizedmanganese oxides exhibit an adsorption superior to its bulkcounterpart because of its polymorphic structures and higherspecific surface area [29]. The SEM images clearly indicated thatthe surface structure of MIAC was greatly changed from the ACsamples. Surface morphology of AC showed very porous surfacestructure and the pores are very unevenly distributed on the surface.The pore size varies significantly. On the other hand, the SEM imagesof MIAC showed that the manganese dioxide particles were signifi-cantly aggregated or even clustered onto the surface of AC or evenwithin the pores of the AC. The particles were distributed unevenlyand predominantly small in size. Earlier, reports indicated that thenano-sized manganese dioxide particles were very orderly aggre-gated onto the sand surface providing distinct specific surface area[24,25]. Further, the EDX data was obtained and shown graphicallyin Fig. 2. Results indicated that composition of the sample MIAC-Rand MIAC-N contained reasonable weight percent of manganesei.e., 15.71% and 9.33% respectively. Whereas, AC samples showedno significant manganese content as also shown in Table 1. Theseresults indicated that the manganese oxide particles were signifi-cantly immobilized on to the surface of AC-R and AC-N. Further,the effective amount of manganese content of the samples were ana-lyzed by US EPA 3050B method and it was obtained as 1577 mg/kgand 1305 mg/kg for MIAC-R and MIAC-N respectively, however, verylow amount of manganese, i.e., 4.5 mg/kg and 5.5 mg/kg was presentin the AC-R and AC-N samples, respectively.

3.2. Speciation of Cu(II) and Pb(II)

The speciation data obtained by the MINEQL geochemicalprogram for Cu(II) and Pb(II) was presented graphically in Fig. 3aand b, respectively. The equilibrium constants used were given

-N, (c) MIAC-R, and (d) MIAC-N.

Fig. 2. EDX analytical results for (a) AC-R, (b) AC-N, (c) MIAC-R, and (d) MIAC-N.

Table 1Weight percent of the different elements shown by different samples (SEM-EDXanalysis).

Materials C O Si S Mn

AC-R 69.63 20.56 9.35 0.48 –AC-N 95.67 2.78 – 1.55 –MIAC-R 49.00 26.87 7.82 0.61 15.71MIAC-N 72.86 16.56 – 1.25 9.33

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elsewhere [24]. The results obtained with speciation studies couldenable to understand the sorption studies or to deduce the mech-anism involved at solid/solution interface. Therefore, a completespeciation studies were carried out in a wide range of pH i.e.,2.0–10.0. The results obtained showed that copper existed asCu2+ up to pH �5.2, beyond that it started in precipitating asinsoluble tenorite. On the other hand, lead existed as Pb2+ up topH 6.2 and beyond this pH it started to precipitate as Pb(OH)2. Boththese ions i.e., Cu2+ and Pb2+ were completely precipitated beyondpH 7.0. It was also observed that insignificant amount of Cu(OH)+

and Pb(OH)+ occurred in between with the maximum of 1.4% and3.9% at the pH around 5.5 (for Cu(II)) and 6.0 (for Pb(II)).

3.3. Batch reactor operations

3.3.1. Effect of pHThe change in solution pH affects the surface charge of the

adsorbents therefore; effect of pH is recognized as an importantparameter helping in to deduce the mechanism involved at solid/solution interface [22,30]. The effect of pH was evaluated betweeninitial pH 2.0–10.0 and the initial Cu(II) or Pb(II) concentration wastaken as 10 mg/L. The results obtained were presented graphicallyin Fig. 4a and b respectively for Cu(II) and Pb(II) sorption. Figuresclearly indicated that, in general, increasing the pH from 2.0 to10.0 of the sorptive solution significantly caused for an apparentincrease in the adsorption of Cu(II) and Pb(II). More quantitatively,it was observed that increasing the pH from 2.23 to 10.34 for AC-R,2.25 to 10.2 AC-N, 2.59 to 10.58 for MIAC-R and 2.54 to 10.63 forMIAC-N caused to increase in Cu(II) removal respectively from15.44% to 100%, 11.81–100%, 29.97–100% and 26.34–100%.

(a)

(b)

Fig. 3. Speciation of Cu(II): 10 ppm and Pb(II): 20 ppm as a function of pH.

(a)

(b)

Fig. 4. Removal of Cu(II) and Pb(II) using AC-R, AC-N, MIAC-R and MIAC-N as afunction of pH.

132 Lalhmunsiama et al. / Chemical Engineering Journal 225 (2013) 128–137

Similarly, increasing the pH from 2.18 to 10.15 for AC-R,2.24–10.21 for AC-N, 2.78–10.33 for MIAC-R and 2.88–10.43 forMIAC-N caused to increase in the Pb(II) removal respectively from18.19% to 100%, 14.47–100%, 31.86–100% and 28.14–100%. Thismay be explicable by using the pHpzc of the adsorbents and thespeciation of these cationic species present in solution. The pHpzcwas obtained as 6.6 for the two activated carbon, AC-R and AC-Nand 6.2 for both MIAC-R and MIAC-N. This inferred that below thispH the surface of these solids carried a net positive charge and be-yond which it contained with net negative charge. Therefore, thelower extent of sorption occurred at lower pH values was becauseof the electrostatic repulsion took place between the metal cations(Cu2+ and Pb2+) and the positively charged solid surface. In addi-tion, at high acidic condition i.e., at very low pH around pH 2–3,perhaps a competition occurred between the excess of hydrogenions and metal ions towards the same sites of solid surface whichsuppressed further the metal ions adsorption. Increasing the pHcaused for a decrease in H+ ion concentrations due to successivede-protonation. This resulted with gradual increase in negativecharge density onto the solid surface which enables to favor moreadsorption of the metal ions [31]. This result was in consistent withthe studies conducted with activated carbon prepared from Ceibapentandra hulls [32], Van apple pulp [33] and manganese oxidecoated carbon nanotubes [34] in the removal of several metalcations. Similarly, the removal of Mn2+ by manganese oxide coatedzeolites [35] and Cu2+ and Pb2+ by manganese modified naturalsand [24] showed somewhat similar results. Moreover, the specia-tion studies showed that Cu(II) exists as Cu2+ upto pH 5.2 and Pb(II)exists as Pb2+ upto pH 6.2, therefore the higher percentage removal

beyond pH 5.2 for copper and pH 6.2 for lead might be a mixedeffect of adsorption and co-precipitation of metal ions on the sur-face of solids resulting a complete removal of these ions i.e., 100%uptake at and above pH �9.0 [24–26]. In this pH range (i.e., �5.2–10.0) the process of adsorption was difficult to explain since resultscould be associated with greater uncertainty [36]; because of thisthe entire further investigation was carried out slightly at lowerpH i.e., pH 4.0 for Cu(II) and pH 5.0 for Pb(II).

3.3.2. Effect of contact timeThe removal of Cu(II) and Pb(II) as a function of time was stud-

ied by varying the contact time from 0 min to 24 h at a fixed metalconcentration of 10.0 mg/L and constant pH 4.0 for Cu(II) and pH5.0 for Pb(II). The results obtained were shown graphically inFig. 5a and b respectively for Cu(II) and Pb(II). These figures clearlyindicated that a sharp initial uptake of these two ions occurredwhich further slowed down with lapse of time and attained a con-stant value at 480 min for Cu(II) and 300 min for Pb(II). It was alsoobserved that the maximum uptake of these two ions took placeduring the initial period of contact which was in a line to the earlierreports as well [37,38]. Further, the data obtained for the effect ofcontact time were employed for pseudo first order rate kinetic andpseudo second order rate kinetic models in order to establish thebest fitted observed sorption data for these two studied metalcations.

The pseudo-first order kinetic model [39] was employed to itslinear form:

(a)

(b)

Fig. 5. Effect of contact time on the removal of Cu(II) and Pb(II) using differentsamples of AC-R, AC-N, MIAC-R and MIAC-N.

(a)

(b)

Fig. 6. Plots of ln(qe � qt) versus t for the pseudo first order kinetic model.

Lalhmunsiama et al. / Chemical Engineering Journal 225 (2013) 128–137 133

lnðqe � qtÞ ¼ ln qe � k1t ð1Þ

where qt (mg/g) is the amount of the metal ion adsorbed at time t.The value of k1 can be obtained from the slope of the plot of ln(qe � qt) versus t as shown in Fig. 6a and b respectively for Cu(II)and Pb(II).

Similarly, the pseudo-second order kinetic equation [39] wasalso employed to its linear form:

tqt¼ 1

k2q2eþ t

qeð2Þ

where q2e is the maximum adsorption capacity (mg/g) and k2 (g/mg/

min) is the adsorption rate constant of pseudo-second order. Fromthe plots of t

qtagainst t, (Fig. 7a and b respectively for Cu(II) and

Pb(II)), qe and k2 were evaluated by the slope and intercept of thestraight lines obtained for these systems.

All kinetic data obtained from the plots of pseudo-first orderkinetic model and pseudo-second order kinetic model were sum-marized in Tables 2 and 3 respectively for Cu(II) and Pb(II). Thevalidity of the exploited model was observed by the values of cor-relation coefficient, R2. It can be observed that the two metal ionsfollowed pseudo-first order as well as pseudo-second order kineticmodel. Further, relatively it was observed that the sorption of thesetwo cations i.e., Cu(II) and Pb(II) onto the surface of these solidswere followed the pseudo-second-order kinetic model. Further,the applicability of pseudo-second-order kinetic model to thetwo metal ions suggested that a ‘chemisorption’ as a predominantprocess involved with valency forces through sharing or exchangeof electrons between the sorbent and the sorbate ions [19,30,40].

These results were in accordance with the other studies viz., effectof sorptive pH and the effect of electrolyte concentrationconducted separately and suggested that the sorbing species weresorbed specifically onto the solid surfaces with strong chemicalbonds.

3.3.3. Effect of initial metal concentrationIn the batch reactor operations, the initial concentration of

metal ions in the solution plays a key role as a driving force to over-come the mass transfer resistance between the aqueous and solidphases. An increase in the amount adsorbed with increasing themetal concentration may indicate potential application of adsor-bents for the treatment of wastewaters contaminated with metalions at high sorptive concentrations [41]. The initial metal concen-tration for Cu(II) was taken between 1 and 14 mg/L at a constantpH 4.0 and 2–16 mg/L for Pb(II) keeping the solution pH 5.0 andat constant temperature 25 ± 1 �C. The data obtained werereturned as showing the percentage removal in Fig. 8a and brespectively for Cu(II) and Pb(II). The figures showed that increas-ing the initial metal concentration caused to decrease the percentremoval for both Cu(II) and Pb(II). This was due to the fact that atlow sorptive concentrations, relatively larger numbers of activesites were available onto the solid surface for lesser number ofsorbing ions however, increasing the sorbtive concentration forthe same dose of solid relatively less number of active sites werepresent resulted in higher percent removal of these metal ions[42]. Moreover, it was calculated and obtained that increasingthe initial metal concentration from 2.0 to 16.0 mg/L for Pb(II)

(a)

(b)

Fig. 7. Plots of tqt

against t for the pseudo second order kinetic model.

(a)

(b)

Fig. 8. Effect of initial sorptive concentration for the removal of Cu(II) and Pb(II)using different samples of AC-R, AC-N, MIAC-R and MIAC-N.

134 Lalhmunsiama et al. / Chemical Engineering Journal 225 (2013) 128–137

caused to increase the amount of Pb(II) uptake respectively from0.55 to 2.94 mg/g (for AC-R), from 0.52 to 2.76 mg/g (for AC-N),from 0.67 to 3.31 mg/g (for MIAC-R) and from 0.64 to 3.18 mg/g(for MIAC-N). Similarly, increasing the Cu(II) concentration from

Table 2Values of qe, k1 and R2 obtained for the sorption of Cu(II) and Pb(II) onto the solidsusing pseudo-first order kinetic model.

Materials Cu(II) Pb(II)

k1 (min�1) qe (mg/g) R2 k1 (1/min) qe (mg/g) R2

AC-R 1.0 � 10�3 1.146 0.953 5.1 � 10�3 1.150 0.973AC-N 1.0 � 10�3 1.265 0.993 4.8 � 10�3 1.341 0.976MIAC-R 1.5 � 10�3 1.073 0.976 6.8 � 10�3 1.350 0.965MIAC-N 1.0 � 10�3 1.169 0.924 5.4 � 10�3 1.060 0.979

Table 3Values of qe, k2 and R2 obtained for the sorption of Cu(II) and Pb(II) onto the solids using

Materials Cu(II)

k2 (g/mg/min) qe (mg/g) R2

AC-R 44.0 � 10�3 0.481 0.9AC-N 60.7 � 10�3 0.427 0.9MIAC-R 42.1 � 10�3 0.807 0.9MIAC-N 82.0 � 10�3 0.532 0.9

1.0 to 14.0 mg/L caused for an increase in Cu(II) removal respec-tively from 0.23 to 0.68 mg/g (for AC-R), from 0.21 to 0.53 mg/g(for AC-N), from 0.27 to 0.91 mg/g (for MIAC-R) and from 0.26 to0.88 mg/g (MIAC-N). The higher uptake of these metal cations bythese solids conferred the potential applicability of these materialsin the attenuation of lead or copper contaminated water even athigher concentrations. Comparatively, Pb(II) was removed withhigher extent to the Cu(II) using these solids indicating higheraffinity of these materials towards the Pb(II).

The equilibrium state adsorption data was further utilized tothe standard Langmuir and Freundlich adsorption isotherms. TheFreundlich equation [43] was taken as:

log ae ¼1n

log Ce þ log Kf ð3Þ

where ae and Ce are the amount adsorbed (mg/g) and bulk sorptiveconcentration (mg/L) at equilibrium, respectively, and Kf and 1/n are

pseudo-second order kinetic model.

Pb(II)

k2 (g/mg/min) qe (mg/g) R2

82 55.3 � 10�3 1.502 0.99634 89.4 � 10�3 1.241 0.99668 16.0 � 10�3 2.518 0.98147 51.2 � 10�3 1.791 0.994

Table 4Freundlich constants and R2 values obtained for the adsorption of Cu(II) and Pb(II)using different samples of AC and MIAC.

Materials Cu(II) Pb(II)

1/n Kf (mg/g) R2 1/n Kf (mg/g) R2

AC-R 0.248 0.370 0.949 0.704 0.727 0.983AC-N 0.220 0.315 0.809 0.699 0.636 0.991MIAC-R 0.270 0.519 0.954 0.554 1.202 0.974MIAC-N 0.265 0.475 0.958 0.574 0.957 0.968

Table 5Langmuir monolayer adsorption capacity (qo), Langmuir constants (b) and R2 valuesestimated for the adsorption of Cu(II) and Pb(II) onto the different samples of AC andMIAC.

Materials Cu(II) Pb(II)

qo (mg/g) b (L/g) R2 qo (mg/g) b (L/g) R2

AC-R 0.684 1.223 0.981 5.348 0.155 0.928AC-N 0.521 1.899 0.979 5.000 0.140 0.969MIAC-R 0.938 1.870 0.998 2.843 0.381 0.878MIAC-N 0.887 1.386 0.984 4.318 0.283 0.947

(a)

(b)

Fig. 9. Effect of background electrolyte concentration for the removal of Cu(II) andPb(II) using different samples of AC-R, AC-N, MIAC-R and MIAC-N.

Lalhmunsiama et al. / Chemical Engineering Journal 225 (2013) 128–137 135

the Freundlich constants referring to adsorption capacity andadsorption intensity or surface heterogeneity, respectively. Andthe linearized Langmuir adsorption model [44] in its usual formwas also adopted for the estimation of maximum monolayer cover-age of metal ions onto the solids (qo) within the studied sorptiveconcentration range:

Ce

q¼ 1

qobþ Ce

qoð4Þ

where q is the amount of solute adsorbed per unit weight of adsor-bent (mg/g) at equilibrium; Ce the equilibrium bulk concentration(mg/L); qo the Langmuir monolayer adsorption capacity, i.e., theamount of solute required to occupy all the available sites in unitmass of solid sample (mg/g) and ‘b’ is the Langmuir constant (L/g).

It was observed that the equilibrium adsorption data obtainedfor these two ions is reasonably fitted well to the Langmuir andFreundlich adsorption models. Therefore, the Freundlich and Lang-muir constants were obtained and returned in Tables 4 and 5,respectively both for Cu(II) and Pb(II) sorption by these solids.The fractional values obtained for Freundlich constant 1/n (0 < 1/n < 1) indicated the heterogeneous surface structure of solids withan exponential distribution of active sites [42,44]. Similarly, theFreundlich sorption capacity obtained is relatively higher showeda strong affinity of the solids towards these two cations. Moreover,the applicability of Freundlich equation pointed it that strongchemical forces were involved at the solid solution interface andthe sorbed ions were likely to be interacted laterally. On the otherhand, the Langmuir constants were estimated for these systemsand the Langmuir monolayer capacity was obtained higher forthe Pb(II) compared to the Cu(II). Moreover, in general the MIACspossessed with higher sorption capacity than their correspondingACs samples, inferred that manganese impregnated samples havegreater affinity towards these ions. Similarly, the lower values ofLangmuir constant (b) for Pb(II) towards these solids comparedto Cu(II) reflected greater affinity of the solid towards Pb(II) [45].

3.3.4. Effect of background electrolyte concentrationThe change in background electrolyte concentrations is an

important parameter revealing the nature of binding of the sorp-tive ions onto the solid surface. The background electrolytesdependence data by soil minerals is usually used to distinguishbetween nonspecific and specific adsorption. Outer sphere

complexes involved with electrostatic or van der Walls interac-tions and are strongly affected by the ionic background electrolytesin the aqueous phase, whereas inner sphere complexes involvedwith stronger covalent or ionic bonds were only weakly affectedby the background electrolyte concentrations [46,47]. Hence,keeping in view the background electrolyte concentrations wasincreased from 0.001 mol/L to 1.0 mol/L NaNO3 in the sorption ofCu(II) and Pb(II) by these solids keeping the initial metalconcentration �10 mg/L and pH 4.0 as constants. The results ob-tained were returned with percent removal versus backgroundelectrolyte concentration in Fig. 9a and b respectively for Cu(II)and Pb(II). It was observed that 1000 times increase in the back-ground electrolytes concentration decreased the percent removalof Cu(II) by 6.13%, 5.42%, 4.91% and 6.34% and for Pb(II) by 3.87%,4.65%, 5.66% and 8.19% for AC-R, AC-N, MIAC-R and MIAC-N,respectively. These results strongly pointed that Cu(II) and Pb(II)ions were specifically adsorbed on the surface of the solid materi-als and predominantly bound by a strong chemical forces resultingin the formation of inner sphere complexes [24,48].

3.4. Column reactor operations

The breakthrough data were obtained by keeping initial sorptiveconcentration at 10 mg/L having a constant pH 4.0. The experimentswere carried keeping the column conditions as stated before. Theresults obtained were presented graphically in Fig. 10a and brespectively for Cu(II) and Pb(II). It was observed that a completebreakthrough was obtained for the throughput volume of Cu(II) with0.13 L, 0.14 L, 0.16 L and 0.15 L respectively of AC-R, AC-N, MIAC-Rand MIAC-N, whereas for Pb(II) at 0.40 L, 0.44 L, 0.50 L, and 0.57 Lrespectively for AC-N, MIAC-N, AC-R and MIAC-R. These resultsinferred that these solid materials, ACs as well as MIACs possessed

Table 6Thomas constants for the removal of Cu(II) and Pb(II) from aqueous solutions by different samples of AC and MIAC.

Materials Cu(II) Pb(II)

qo (mg/g) KT (L/min/mg) Least square sum qo (mg/g) KT (L/min/mg) Least square sum

AC-R 2.316 1.80 � 10�3 2.80 � 10�1 5.960 1.92 � 10�3 1.80 � 10�1

AC-N 2.506 1.76 � 10�3 1.50 � 10�1 4.932 2.26 � 10�3 3.30 � 10�2

MIAC-R 2.449 2.23 � 10�3 2.10 � 10�1 6.242 3.39 � 10�3 5.80 � 10�2

MIAC-N 2.535 1.72 � 10�3 3.20 � 10�1 5.346 1.71 � 10�3 1.30 � 10�1

136 Lalhmunsiama et al. / Chemical Engineering Journal 225 (2013) 128–137

significant high removal capacity for these metal ions even underthe dynamic conditions. The data obtained were further utilized tooptimize the removal capacity of these solids under the dynamicconditions using the Thomas equation [49]:

Ce

Co¼ 1

1þ eðKT ðqom�CoVÞÞ=Qð5Þ

where Ce was Cu(II) or Pb(II) concentration in the effluent (mg/L); C0

was Cu(II) or Pb(II) concentration in the feed (mg/L); KT the Thomasrate constant (L/min/mg); qo was the maximum amount of Cu(II) orPb(II) could be loaded (mg/g) under the specified columnconditions; m the mass of adsorbent loaded (g); V the throughputvolume (L); and Q the flow rate of pumped sorptive solution(L/min). A non-linear regression was conducted using the columndata by the least square fitting for two unknown parameters i.e.,

(a)

(b)

Fig. 10. Breakthrough curves for the removal of Cu(II) and Pb(II) using AC-R, AC-N,MIAC-R and MIAC-N solids.

KT and qo. The values obtained were returned in Table 6. The datashowed that the removal capacity of Pb(II) was higher for all thefour samples than the Cu(II) which were again in accordance tothe results obtained from batch reactor experiments.

4. Conclusions

The activated carbon samples, precursors to the agriculturalby-product or waste biomaterials viz., rice hulls and areca nutwaste, were obtained. Further, relatively high specific surface areaof ACs were utilized in situ impregnation of small sized or evennano-sized particles of manganese oxide; which were immobilizedonto the surface and pores of the ACs. The SEM-EDX analysisshowed that AC-R and AC-N were not contained with manganesewhereas significant and distinct peaks of manganese wereobserved with MIAC-R and MIAC-N solids. Moreover, the SEMimages of these solids (i.e., MIACs) showed that manganese oxidenano-particles were distributed on to the surface as well as withinthe pores of ACs. The batch reactor studies enabled that the con-centration dependence data followed both Langmuir and Freund-lich adsorption models and the kinetic data was fitted well to thepseudo-second-order model. Increasing the background electrolyteconcentrations from 0.001 to 1 mol/L was not affected significantlythe uptake of these metals; conferred the uptake of these cationsonto the solids was proceeded with specific adsorption and boundwith strong chemical forces, perhaps, predominantly forming an‘inner-sphere complexes’ onto the surface of these solids. Themaximum removal capacity of these solids obtained by usingThomas equation referred the suitable application of AC and MIACfor the removal of Cu(II) and Pb(II) even under the dynamic condi-tions. The removal capacity of Pb(II) is comparatively better thanCu(II) for all these materials employed. The study is an attemptfor suitable/efficient, cost-effective, cleaner and environmental be-nign treatment technology in the remediation of Cu(II) and Pb(II)contaminated waters.

Acknowledgement

This work was supported by the National Research Foundationof Korea (NRF) grant funded by the Korea government (MEST) (No.2012R1A2A4A01001539).

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