Removal of Heavy Metal Ions from Water by Magnetic Cellulose

Research Article pubs.acs.org/journal/ascecg

Removal of Heavy Metal Ions from Water by Magnetic CelluloseBased Beads with Embedded Chemically Modified Magnetite Nanoparticles and Activated Carbon Xiaogang Luo,* Xiaojuan Lei, Ning Cai, Xiuping Xie, Yanan Xue, and Faquan Yu* Key Laboratory for Green Chemical Process of Ministry of Education; Hubei Key Laboratory for Novel Reactor and Green Chemistry Technology; School of Chemical Engineering and Pharmacy, Wuhan Institute of Technology, Wuhan 430073, China S Supporting Information *

ABSTRACT: For the removal of heavy metal ions from water, new resources should be exploited to design more efficient, environmentally friendly adsorbents. To tackle this challenge, millimeter-scale magnetic cellulose-based beads with microand nanopore structure were fabricated via an optimal extrusion dropping technology from NaOH/urea aqueous solution. The composite beads incorported with carboxyl decorated magnetite nanoparticles and nitric acid modified activated carbon have convenient operation based on sensitive magnetic response and highly effective removal performance for Cu2+, Pb2+, and Zn2+. Their structure and properties were investigated. Moreover, the adsorption equilibrium, kinetics, and thermodynamics of Cu2+, Pb2+, and Zn2+ by the prepared composite adsorbents were examined. The results revealed that these adsorption processes were spontaneous endothermic reactions and determined by combination of physical and chemical adsorptive mechanisms. KEYWORDS: Magnetic cellulose-based beads, Green technology, Heavy metal ions, Adsorption mechanism, Safe drinking water

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INTRODUCTION Heavy metals like copper, zinc, and cobalt are essentially necessary for the growth of the functions of living organisms and the normal body, and the high concentrations of other heavy metals such as lead, cadmium, manganese, and chromium are highly toxic to human and aquatic life, which may lead to kidney and liver problems and genotoxic carcinogen,1 ultimately deteriorating public human health. Many literature works have reported the serious pollution of groundwater and surface water caused by the heavy metals in the world, which has led to the deterioration of the quality of drinking and irrigation water.2−5 A challenging research trend in biology and environmental science and technology is in the area of preferably controling the water contamination and strengthening the purification of the aquatic environment.6 The demands for efficient, handy, and cost-effective technologies for the decontamination of groundwater and surface water without imperiling public human health are still quite urgent.7 Recently, various approaches have been proposed to develop costeffective and more highly effective adsorbents containing natural polymers because of their biocompatibility and biodegradability.8,9 The polysaccharides, as an abundant, biodegradable, and renewable resource, have a capability to interact with various molecules by physical and chemical interactions. Hence, adsorption on polysaccharide-based adsorbents can be a low-cost procedure in water decontami© 2016 American Chemical Society

nation for extraction and separation of compounds and an efficient approach for environmental protection.8 Among them, cellulose is considered to be the most widespread, renewable polysaccharide on earth. The employment of cellulose as a bioaffinity carrier material may improve its adsorption capability for contaminants on account of its numerous hydroxyl groups, which could also be utilized for coupling reactions or for other functional modification.10 It is known that functional cellulose-based materials have been considered as promising biocompatible adsorbents for contamination removal from water.11 Restricted by their low adsorption capacities, developing new promising cellulose-based composite adsorbents is an interesting and very important area of research. There are a lot of featured adsorbents, or efficient adsorbents, which can be used as potential candidates for organic, inorganic, biological, or mineral contaminations.12 In comparison with a traditional adsorbent, a magnetic adsorbent is considered to be an effective and rapid technique for separating inorganic nanoparticles from aqueous solutions because of the better separation ability and lower energy requirement.13 Currently, iron oxide nanoparticles have been diffusely researched in the separation technologies field because Received: April 19, 2016 Revised: June 3, 2016 Published: June 6, 2016 3960

DOI: 10.1021/acssuschemeng.6b00790 ACS Sustainable Chem. Eng. 2016, 4, 3960−3969

Research Article

ACS Sustainable Chemistry & Engineering

The nitric acid decorated activated carbon (AC-NA) used in this paper was decorated by the concentrated nitric acid.19 The appropriate amount of activated carbon (400 meshes) was added to diluted nitric acid solution with a known volume of 1:1 and heated for 3 h at 90 °C to promote the formation of functional groups.19 The residual samples were rinsed several times with water. MCB were prepared with an optimal extrusion dropping technology.20 An 8.0 g portion cellulose was dissolved into a 200 mL mixed solvent of NaOH/urea/H2O of 7:12:81 (by weight) at −12.3 °C under vigorous mechanical stirring for 2 min, until to get the transparent cellulose solution. A 5 g portion of MN-CA and 1 g of ACNA were successively put into the prepared transparent cellulose solution, stirred for 1.5 h with a mechanical stirrer, and then with ultrasonic homogenizer treatment 3 times for 15 min. The resulting solution was added dropwise into sodium chloride solution (250 mL, 10 wt %) at a 5 mL min−1 constant rate with a syringe pump. The prepared composite beads were cured in water overnight and washed with the double distilled water several times. The resultant beads samples freeze-dried for 48 h in a freeze drier for sample characterization or were stored in water for adsorption experiments. Characterization. The structures of surface and cross section of the freeze-dried wet beads were observed with scanning electron microscope (JSM-5510LV), and the elemental concentration was measured with the energy dispersive X-ray (EDX) mappings. The pore structure of the prepared wet MCB such as wet density (ρw), mean pore volume (Vp), and porosity (Pr) was determined by drainage method with a dilatometer.21 The surface area of BET (Brunauer− Emmett−Teller) and the total pore volume were measured with a computer-controlled nitrogen gas adsorption analyzer (ASAP2010). The FTIR spectra of the samples were measured with a Fouriertransform infrared spectrometer (model 1600, PerkinElmer Co). The XRD measurement of MCB was measured with an X-ray diffraction diffractometer (D8-Advance, Bruker). The patterns with the Co Kα radiation (λ = 0.15406 nm) at 40 kV and 30 mA were recorded in the region of 2θ from 10 to 80°. Digitized photographs of the products were taken by a digital camera. The magnetic properties of the magnetite and MCB were determined with a superconducting quantum interference device (SQUID, MPMS XL-7, QUANTUM DESIGN, USA) at room temperature. Adsorption Experiments. Adsorption experiments were performed by the batch technique at three different temperatures (298, 303, and 313 K).22 A 5 g portion of MCB was added to 30 mL of Cu2+, Pb2+, and Zn2+ solution. The range of the concentration of adsorbate was 500−750 mg L−1. The mixture was shaken to ensure good contact between MCB and adsorbate. After the adsorption experiment process, the beads were separated by a magnet, and then the remaining concentration of adsorbate was analyzed by AAS. The repeating experiments were performed three times under the same identical conditions. The average values were calculated to make sure the reproducibility of the experimental results, and the standard deviations of the experimental results were within ±5.0% of average values. The amounts of adsorption of MCB samples were calculated as the following equation:

of their cost-effective preparation, simple surface modification or coating and the ability to manipulate or control materials on nanoscale dimensions, which exhibit excellent versatility in the field of separation and environmental techniques.14 In addition, super paramagnetic iron oxide nanoparticles have large surface area, surface modifiability, biocompatibility, low toxicity, and chemical inertness which is highly apposite in adsorption technology for environmental remediation.15 Among other inorganic materials, the activated carbon (AC) has been considered as a highly effective and frequently used adsorbent for adsorptive removal of various inorganic and organic contaminations existed in water. The excellent adsorption performance of them is relevant with the existence of various surface functional groups, the well-developed porosity and internal pore structure, and the large surface area.16 In our previous work, magnetic cellulose beads entrapping activated carbon were applied to adsorb organic dyes.17 However, the treatment of the contaminants by undecorated activated carbon as filler mainly relies on physical adsorption without selectivity. Moreover, good adsorption capacity of the composite adsorbents was presented only at the very beginning and would significantly decline after several cycles. In this research, the combination of specifically modified magnetic nanoparticles and AC to cellulose may create new nanocomposite adsorbents, which possess high adsorption capacity for heavy metals removal and easy recovery by magnetic field. It could be deduced that electrostatic attraction between negatively charged magnetic cellulose-based beads (MCB) samples and positively charged Cu2+, Pb2+, and Zn2+ was the initial driving force of heavy metals binding to the adsorption site of the adsorbents. A simple technology was also attempted to prepare this new type of cellulose based beads, in which both of the nitric acid modified activated carbon and carboxyl decorated Fe3O4 nanoparticles was embedded into cellulose matrix. The structure and adsorption behaviors, as well as kinetic, equilibrium, and thermodynamic properties and reusability of MCB, were studied to evaluate the performance of their practical applications in heavy metal removal from water.

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EXPERIMENTAL SECTION

Materials. Cotton linter pulp samples (α-cellulose) were obtained from Hubei Chemical Fiber Group Ltd. (Xianfan, Hubei, China) with the content of α-cellulose beyond 95%. The viscosity-average molecular weight of the used cellulose sample was measured by viscometry in cadoxen to be 12.5 × 104 Dalton. Activated carbon (AC, C3345) powder was purchased from Sigma-Aldrich. Activated carbon was passed through 400-mesh sieve before used. Fe3O4 nanoparticle powder (I109514, 99.5%, 20 nm) was purchased from Aladdin industrial corporation (Shanghai, China). The stock solutions of Cu2+, Pb2+, and Zn2+ were obtained by dissolving the required amount of CuSO4, Pb(NO3)2, ZnCl2 (Sinopharm Chemical Reagent Co., Ltd., CN), and then the stock solutions were diluted in double distilled water to get various initial concentrations measured by an atomic absorption spectroscope (AAS, SOLAAR M6, Thermo, USA). All of other chemical agents were analytical grade. Preparation Procedure of MCB. The carboxyl decorated magnetite (Fe3O4) nanoparticles (MN-CA) were obtained by postmodifying the magnetite surface with citric acid.18 The nanoparticles of Fe3O4 were added to 0.1 M citric acid under ultrasonic condition for 45 min, and the reaction was kept for 4 h at room temperature. A magnet was used to separate the MN-CA, and then the prepared MN-CA samples were rinsed thoroughly with acetone and double distilled water.

qeq =

C0 − Ceq m

V

(1)

qe represents the adsorption capacity of heavy metals onto the composite adsorbents (mg g−1), C0 (mg L−1) and Ceq (mg L−1) represent the initial concentration and the equilibrium concentration of heavy metals, respectively, V (L) represents the volume of the aqueous solution, and m (g) represents the weight of MCB. Regeneration Experiments. After the adsorption of heavy metals, the loaded MCB was regenerated in 2 mol L−1 sodium citrate solutions. The MCB was rinsed with distilled water after elution to remove trace amounts of salt, and then, the regenerated MCB samples were assessed by adsorption tests to study their recyclability. 3961


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Figure 1. SEM images of the surface and cross section of native cellulose beads (a, b) and MCB (c, d). (inset) EDS pattern of MCB after Cu2+ adsorbed.

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RESULTS AND DISCUSSION Structure Characterizations. SEM images of native cellulose beads (a, b) and MCB (c, d) were exhibited in Figure 1. Both of the surface and interior of MCB expressed micro- and nanopore structure. The formation of pore structure was due to the H2O-induced phase separation during the preparation process of the sol−gel method, where the regions of solvent-rich resulted in the formation of pore.20 Compared to native cellulose beads (Figure 1a, b), MCB (Figure 1c, d) had similar structure. Alteration in the pore morphology is believed to result from inorganic powders that affect the process of pore nucleation. The physical properties of native cellulose beads and MCB samples are presented in Table S1. From the values of the table, it could be deduced BET surface area of MCB increased, while the average pore diameter of MCB reduced in comparison with that of the native cellulose beads. The BET-specific surface area and average pore diameter of native cellulose beads and MCB are 88.58 m2/g and 864 nm and 90.05 m2/g and 220 nm, respectively. MCB (Table S1) displayed high water content, porosity, and pore volume. The average pore diameter of MCB samples was nanoscale, which indicated that MCB had good porous structure. Due to high hydrophilicity and rich porous structure of MCB, the dissolved heavy metal ions can lightly permeate into the MCB samples and interact with the organic fillers to form a complex, which will facilitate the adsorption of these heavy metal ions. For example, EDS pattern (Figure 1d inset) revealed the presence of the Cu in the cross-section of MCB after adsorption of Cu2+. The presence of Cu distributed

throughout the sample confirmed that not only the surface of MCB but also the reactive material within the core of the beads structure came into contact with water and contaminants to maximize reactivity. Figure 2 indicates the FTIR spectra of MN (a), MN-CA (b), AC (c), AC-NA (d), native cellulose beads (e), and MCB (f), respectively. In Figure 2a, there was a strong peak at 572 cm−1

Figure 2. FTIR spectra of MN (a), MN-CA (b), AC (c), AC-NA (d), native cellulose beads (e), and MCB (f). 3962


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ACS Sustainable Chemistry & Engineering because of Fe−O vibrations of the Fe3O4.23 Additionally, the peaks at 1624 cm−1 was attributed to absorbance of the symmetric stretching vibration of COO− of the modifier of citric acid, and the peaks at 1364 cm−1 were due to absorbance of the symmetric stretching vibration of the COO−, revealing that the Fe3O4 nanoparticles were coated by COO− groups, which could enhance the adsorption of the adsorbents to heavy metal ions including Cu2+, Pb2+, and Zn2+ through the electrostatic attraction.24 FTIR spectra for AC and AC-NA are depicted in Figure 2c and d. The results show the peak at 1720 cm−1 only exists in the oxidized samples. It is difficult for the peak to be identified positively due to its position on the spectra of the native activated carbon. However, López et al.25 obtained the similar results when modified carbon materials using the air oxidation method, which were assigned to lactone groups, carbonyl groups, and free carboxyl groups close to the hydroxyl groups. Then the most logical analyzation for the peak of AC-NA at 1720 cm−1 could be the presence of carboxyl groups which are produced by nitric acid oxidation method. The oxidation of nitric acid have demonstrated its effectiveness by producing numerous acidic surface groups on carbon.26 The new function groups introduced into these inorganic particles provided more active sites located at the surface and the inside of MCB, resulting in the increase of heavy metals adsorption on the beads. Figure 2e exhibits characteristic absorbance of cellulose II, which indicates that the structure of the cellulose had been shifted to cellulose II in the process of the preparation of MCB. The peaks at 3000−3700 cm−1 shifted to a little higher wavenumbers and became a bit broader because of the stretching vibrations of hydroxyl groups of the cellulose, suggesting the strong interaction among the cellulose matrix; the fillers of MN-CA and AC-NA through hydrogen bonding occurred in MCB. These new peaks appearing at 1422, 1724, and 2900 cm−1 are assigned to the stretching vibrations of C− H of −CH3 and C−O, which confirmed the bounding of the cellulose chain onto the surface of MN-CA and AC-NA particles by the electrostatic force. These results suggested that the stability of activated carbon and Fe3O4 nanoparticles in the cellulose matrix was enhanced, which is beneficial for the applications of the magnetic cellulose nanocomposite. The XRD was analyzed to examine the structure and interaction between fillers and cellulose matrix in Figure 3. The intense diffraction peaks of the MCB indexed to (220), (311), (400), (422), (511), and (440) planes were located at 2θ = 30.15°, 36.30°, 43.35°, 53.89°, 57.18°, and 62.29°, respectively (Figure 3a, b), corresponding to the standard XRD data for the cubic phase nanoparticles of Fe3O4 with a face-centered cubic (fcc) structure (JCPDS No. 89-3854). And, the broad diffraction peaks exhibit the very small size of Fe 3 O 4 nanoparticles, which changed only a little after modification. The mean size of MN and MN-CA determined using Scherer’s equation were 296 and 331 nm, respectively. The TEM image of the Fe3O4 nanoparticles was expressed in Figure S1. The XRD patterns exhibited that both the characteristic reflection peaks of activated carbon of the two samples are around 43°. All the peaks of the three samples are shifted due to the distortion of the lattice.27 The peaks of both the raw AC and AC-NA demonstrated only a little change, which may be attributed to the high stability of turbostratic structure of carbon during the modification. The MCB (Figure 3f) samples displayed the characteristic peaks of cellulose II. The diffraction peaks at 2θ = 12.1, 19.8,

Figure 3. Powder X-ray diffraction patterns of native MN (a), MN-CA (b), AC (c), AC-NA (d), native cellulose beads (e), and MCB (f).

and 22.6° for (110̅ ), (110), and (200) planes are peaks of cellulose II crystal.28 In Figure 3f, the existence of characteristic diffraction peaks of Fe3O4 and AC indicates that the two fillers have been introduced into magnetic cellulose nanocomposite. It can be concluded from the obtained results that the AC-NA and MN-CA were embedded into the matrix of cellulose, and the structure and properties of AC-NA and MN-CA were hardly changed. Kinetics Studies on Heavy Metal Adsorption. The batch adsorption kinetics of Cu2+, Pb2+, and Zn2+ by MCB were assessed with three models: pseudo-first-order kinetic, pseudosecond-order kinetic, and intraparticle diffusion models, to decipher the controlling mechanism of the adsorption process. The pseudo-first-order and pseudo-second-order models could be defined as in eqs 2 and 3 below.29 ln(qe − qt ) = ln qe − k1t

(2)

t 1 t = + 2 qt qe k 2qe

(3)

qe and qt represent the amounts of adsorbed (mg g−1) of equilibrium and time t (min), respectively. k1 (min−1) and k2 (g mg−1 min−1) represent the rate constants of adsorption. These constants are obtained by linear regression of t/qt versus t. The initial adsorption rate is decided by the following equation: k 0 = k 2qe 2

(4)

Figure 4 displays the pseudo-second-order kinetic of Cu2+, Pb2+, and Zn2+ onto MCB at different temperatures. As presented in Table 1, parameters (R2) of pseudo-second-order kinetic model (R2 > 0.99) are bigger than those of both of the other two models for all the temperatures. Furthermore, the values of experimental qe agree well with the calculated qe values by using the pseudo-second-order model. The result reveals that the adsorption of Cu2+, Pb2+, and Zn2+ onto MCB accords with the pseudo-second-order model very well. The rate constant k2 values were 0.014, 0.07, and 0.220 g mg−1 min−1 for the adsorption of Cu2+, Pb2+, and Zn2+ onto MCB at temperature of 303 K and initial concentration of 600 mg L−1 (Table 1). 3963


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Figure 4. Adsorption kinetics of pseudo-second-order model of Cu2+ (a), Pb2+ (b), and Zn2+ (c) onto MCB (mg L−1 initial solution concentration). (inset) Linear form of the pseudo-second-order rate model of the kinetics study.

Table 1. Kinetic Parameters of Cu2+, Pb2+, and Zn2+ Removal onto MCB (303 K and 600 mg L−1) qe,exp (mg g−1)

pseudo-first-order kinetic model 2+

Cu Pb2+ Zn2+ pseudo-second-order kinetic model

qe,exp

2+

Cu Pb2+ Zn2+ intraparticle diffusion model Cu2+ Pb2+ Zn2+

qe,exp

29.267 29.834 16.906 (mg g−1)

29.267 29.834 16.906 (mg g−1)

29.267 29.834 16.906

qe,cal (mg g−1)

qe,cal (mg g−1)

qe,cal

35.348 36.010 20.251 (mg g−1)

25.886 36.558 25.665

k1 (min−1)

7.681 15.289 20.401 k2 (g mg−1 min−1)

0.045 0.054 0.010 k0 (g mg−1 min−1)

0.014 0.007 0.220 Ki (mg g−1 min−1/2)

17.188 8.890 90.252 C (mg g−1)

1.415 0.981 0.069

25.084 24.473 19.575

R2 0.964 0.957 0.279 R2 0.999 0.998 0.999 R2 0.973 0.628 0.692

Figure 5. Adsorption kinetics of intraparticles diffusion of Cu2+ (a), Pb2+ (b), and Zn2+ (c) onto MCB (mg L−1 initial solution concentration).

The transferring of the Cu2+, Pb2+, and Zn2+ from water to the surface of MCB samples comprises several steps (e.g., adsorption on the pore surface, external diffusion, surface diffusion, and pore diffusion).22 During the process of the adsorption at the outside surface of the MCB samples, the heavy metals may move from the surface of beads to the inner pores through intraparticle diffusion, which is generally a relatively slow process. And the intraparticle diffusion model has been evaluated with the following equation:22,30 qt = K it 1/2 + c

constant of intraparticle transport (Ki) is obtained by the slope of the linear curve. Figure 5 shows the intraparticles diffusion kinetics of Cu2+, Pb2+, and Zn2+ onto MCB at three different temperatures. The rate of adsorption is possibly affected by various factors including the concentration of the adsorbate, the degree of mixing, the size of the adsorbate molecule and its affinity to the adsorbents, the size distribution of the adsorbent, the diffusion coefficient of the adsorbate in the bulk phase, and so on. The values of R2 (Table 1) obtained by intraparticle diffusion kinetic model are a little lower compared with pseudo-second-order model. The results reveals that although the process of adsorption of Cu2+, Pb2+, and Zn2+ onto the MCB cannot be simulated by the model of the intraparticle diffusion very well, the pore diffusion is still included in the process of adsorption because of the porous nature of MCB.

(5)

in which Ki is the rate constant of the intraparticle diffusion (mg g−1 min−1/2) and c is the intercept. If a straight line passing through the origin in this model is presented from plotting a graph of qt versus t1/2, it is possible that the process of adsorption is comprised of diffusion of heavy metals. The rate 3964


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Figure 6. Adsorption equilibrium isotherm of Cu2+ (a), Pb2+ (b), and Zn2+ (c) onto MCB at 298 K: Langmuir isotherm and Freundlich isotherm fitting curve (equilibrium time 180 min).

Table 2. Equilibrium Adsorption Isotherm Parameters of Cu2+, Pb2+, and Zn2+ Removal onto MCB Cu2+ isotherm equation Langmuir

Freundlich

DR

−1

q0 (mg g ) K (L mg−1) R2 RL kf n R2 qm (mol g−1) K′ (mol2 kJ−2) E (kJmol−1) R2

Pb2+

298 K

303 K

313 K

298 K

303 K

313 K

298 K

303 K

313 K

47.573 0.127 0.999 0.928 15.770 3.948 0.987 40.034 8.758 0.238 0.930

47.641 0.135 0.997 0.924 16.605 4.143 0.998 39.179 0.355 1.185 0.944

46.926 0.192 0.999 0.896 19.431 4.783 0.975 40.372 4.238 0.343 0.858

37.994 0.163 0.999 0.910 17.496 4.444 0.944 40.232 8.108 0.248 0.984

31.066 0.176 0.998 0.904 18.782 4.582 0.973 40.064 4.197 0.345 0.764

33.681 0.224 0.998 0.881 19.652 5.029 0.899 39.992 4.988 0.316 0.938

20.800 0.003 0.918 0.998 0.778 1.730 0.925 28.838 4247.126 0.010 0.977

22.300 0.011 0.983 0.993 0.235 1.251 0.983 33.897 5564.562 0.009 0.986

21.30 0.042 0.993 0.975 0.112 1.077 0.991 35.879 6344.689 0.008 0.956

model or not. The linearized Langmuir isotherm29,31 is shown as follows:

Activation Energy. An Arrhenius equation is applied to deduce the apparent activation energy for the adsorption process of Cu2+, Pb2+, and Zn2+ onto MCB.

⎛ E ⎞ k = A exp⎜ − a ⎟ ⎝ RT ⎠

Zn2+

ce c 1 = + e qe q0 K q0

(6)

(7)

where K represents the Langmuir adsorption constant (L mg−1), Ce represents the solute concentration at equilibrium (mg L−1), q0 represents the maximum concentration retained by the adsorbent (mg g−1), and qe represents the adsorption capacity in equilibrium (mg g−1). The maximum adsorption amounts were calculated to be 47.64, 37.99, and 22.30 mg g−1 for Cu2+, Pb2+, and Zn2+ adsorption onto MCB, respectively. The dimensionless constant separation factor (RL) could be applied to judge whether the adsorption process is favorable or not for the Langmuir isotherm model, and the calculating equation is shown as follows:32

Ea is activation energy (kJ mol−1), k is the rate constant of adsorption (g mg−1 min−1), R, the universal gas constant, is 8.314 J mol−1 K−1, A (g mg−1 min−1), is the Arrhenius constant, and T (K) is the temperature of the solution. Ea and A were obtained by plotting a graph of k (the rate constant obtained from the pseudo-second-order kinetic) versus T at the three different temperatures. The values of Ea and A obtained from the Arrhenius plots (figures not shown) for the adsorption of Cu2+, Pb2+, and Zn2+ onto MCB are 19.77 kJ mol−1 and 5.36, 14.32 kJ mol−1 and 12.42, and 11.82 kJ mol−1 and 13.26, respectively. The values of Ea are very low for Cu2+, Pb2+, and Zn2+/MCB sorption systems. Hence, the adsorption process may include both chemical adsorption and physical adsorption. Equilibrium Studies on Heavy Metals Adsorption. Langmuir, Freundlich, and Dubinin−Radushkevich (DR) isotherm models were chosen to evaluate the adsorption process. The equilibrium time of the adsorption process of Cu2+, Pb2+, and Zn2+ onto MCB was 180 min. Figure 6 presents the plots of qe versus Ce for the adsorption isotherms of Cu2+, Pb2+, and Zn2+ removal onto MCB at 298 K. The correlation coefficients (R2) and constant parameters are summarized in Table 2. The Langmuir adsorption isotherm could predict the maximum monolayer adsorption capacity of MCB and determine whether the adsorptive process is favorable for this

RL =

1 1 + KC0

(8)

where C0 represents the initial concentration of adsorbate. The values of RL are calculated in the range of 0.89−0.92, 0.88− 0.91, and 0.97−0.99 for Cu2+, Pb2+, and Zn2+ onto MCB at 298−313 K, respectively, showing that this process is a favorable adsorption.33 The Freundlich isotherm model29,31 is grounded on multilayer adsorption process, and the equation is shown as the following equation ln qe = ln k f + 3965

1 ln Ce n

(9) DOI: 10.1021/acssuschemeng.6b00790 ACS Sustainable Chem. Eng. 2016, 4, 3960−3969

Research Article


thermodynamic parameters of ΔH and ΔS can be represented by the following Van’t Hoff correlation:

where kf and n represent the isotherm constant of this model which suggest the intensity and capacity of the adsorption, respectively. In this model the constants kf and n be correlated with the adsorption mechanism and the rate of adsorption of heavy metals. The parameters and the correlation coefficients of the Freundlich model were listed in Table 2. It indicates that the values of kf and n changed with the increase of temperature, and the value of n was bigger than 1, suggesting that Cu2+, Pb2+, and Zn2+ are favorably adsorbed by MCB in the whole temperature range. In addition, the higher value of n suggests the formation of stronger bond between heavy metal ions and MCB adsorbents. The DR model is independent of the experimental temperature. It can be applied to predict a maximum adsorption capacity of the MCB and the energy of adsorption per unit of adsorbate. The equation of the DR model34 is shown as follows: ln qe = ln qm − K ′ξ 2

ln K =

(10)

1 (2K ′)1/2

(11)

The values of E procured from the plot of ln qe versus ξ2 at three different (298, 303, and 313 K) adsorption temperatures are shown in Table 2. The DR isotherm was displayed in Figure S2. The value of E provides information about physical and chemical adsorption.35 It was in the range of 0.238−1.185, 0.248−0.345, and 0.008−0.010 kJ mol−1 for Cu2+, Pb2+, and Zn2+ by MCB, respectively, which is lower than the range of adsorption reaction 8−16 kJ mol−1, and these results also showed that the adsorption mechanism of Cu2+, Pb2+, and Zn2+ onto MCB are the combination of chemical and physical sorption. Thermodynamic Studies. The type of adsorption can be decided by some thermodynamic quantities, such as Gibbs free energy (ΔG), the entropy change (ΔS), and enthalpy change (ΔH), these thermodynamic quantities of heavy metals removal onto MCB are given in Table 3. ΔG is calculated as follows: (12)

ΔG = −RT ln K

K represents the adsorption equilibrium constant (from Langmuir model). The relationship between K and the Table 3. Thermodynamic Parameters of Cu2+, Pb2+, and Zn2+ Removal onto MCB heavy metal ions 2+

Cu

Pb2+

Zn2+

T (K) ΔG (kJ mol−1) 298 303 313 298 303 313 298 303 313

−4.076 −5.184 −5.199 −3.892 −4.302 −4.559 −7.85 −11.2 −14.4

ΔH (kJ mol−1)

ΔS (kJ mol−1 K−1)

15.197

0.065

14.872

0.062

116.951

0.420

(13)

ΔS and ΔH were procured from the intercept and slope of Van’t Hoff plots, respectively. The negative values of ΔG indicated the thermodynamical feasibility and spontaneity of the adsorption process for the MCB despite of the increasing of the testing temperature. In addition, the positive values of ΔH suggested that the adsorption process of Cu2+, Pb2+, and Zn2+ ions was endothermic. Thus, a large amount of heat is required to be consumed for the Cu2+, Pb2+, and Zn2+ ions to transfer from water into the surface of MCB. In addition, the positive value of ΔS suggested the increased disorder in the system during the process of adsorption. Hefne et al.36 noted that a positive ΔS value indicates the redistribution of energy between the heavy metals and MCB. Thus, the distribution of translational and rotational energy among a few molecules will augment the positive ΔS and the randomness will increase at the interface of the solid−solution phase during all of the adsorption processes. Desorption of Heavy Metals and Reuses. Figure 7b shows (a) the magnetic hysteresis loops of the native magnetite nanoparticles (MN), (b) the carboxyl decorated magnetite nanoparticles (Fe3O4) (MN-CA), and (c) MCB as a function of the applied magnetic field at 298 K. The saturation magnetization intensity of MCB was about 73.0 emu g−1, with a relatively small hysteresis loop. Hence the magnetite nanoparticles respond very well to magnetic fields without any permanent magnetization. MCB samples could be easily separated with a magnet, suggesting the sensitive magnetic response (inset: photograph of MCB attracted by a magnet). This characteristic is very vital for the industrial and water treatment applications where the secondary pollution should be avoided and the operation should be convenient during the process of adsorption and desorption in the water treatment. In Figure 7a, the desorbed MCB were reused for Cu2+ removal. Heavy metals adsorbed onto MCB were quantitatively desorbed with sodium citrate. The Cu2+ adsorption and desorption capacities fluctuated with just a slight decrease after the second cycle, indicating that MCB could retain its functionality for more than five cycles without significant loss of original adsorptive capacity. These results indicated that heavy metal loaded MCB could be regenerated and further reutilized for heavy metals adsorption from water, which could significantly reduce the operation cost of this adsorptive technology. Design of the Beads and Mechanism of Adsorption. In view of all the results above, the schematic depiction of preparation of MCB and the mechanism of adsorption of heavy metal ions onto MCB were proposed and illustrated in Figure 8. In this work, it is attempted to design a novel cellulose-based adsorbent for the removal of heavy metals from water through a simple optimal extrusion dropping technology, in which the magnetic nanoparticles and activated carbon will make the adsorbent possess high levels of operating convenience in adsorption and desorption processes, and the extremely high adsorption capacity for heavy metals. In the process of adsorption, the high surface negative charges of MCB facilitated the fast migration of positive charged heavy metals to the periphery of the MCB through the electrostatic attraction. Furthermore, the carboxylic functional groups of surface and

where qm represents the theoretical monolayer saturation capacity (mol g−1), the value of ξ is equal to RT ln(1 + 1/Ce), and K′ is the constant of the adsorption energy (mol2 kJ−2), which is concerned with the average adsorption energy (E, kJ mol−1) calculated by the following formula: E=

ΔS ΔH − R RT

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Figure 7. Cycle adsorption of batch experiments for MCB (a), magnetic hysteresis loops of the native magnetite nanoparticles (MN), MN-CA, and MCB (b). (inset) MCB attracted by a magnet.

Figure 8. Schematic depiction of preparation of MCB and the adsorption mechanism of heavy metal ions by MCB.

and Zn2+ by MCB is predominantly by electrostatic attraction between the adsorbents surface and heavy metals. A simple preparation prcedure, cheap cellulose feedstock, their availability on an industrial scale, fast adsorption speed, great adsorption capacity, and good reusability make the beads an economically viable and environmentally friendly cellulosebased adsorbent for highly efficient removal of the tested heavy metal ions from the aqueous environment.

inside of MCB are beneficial to the highly efficient adsorption of heavy metals. The heavy metal ions coordinated by the carboxyl groups can be further adsorbed onto the MCB samples. Thus, it can be concluded that both physical adsorption and chemical adsorption are greatly essential in heavy metal removal from water, which highly improve the absorptive capacity of MCB for heavy metals.37

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CONCLUSIONS Environmentally friendly, magnetic, micro- and nanostructured, cellulose-based beads were fabricated via an optimal extrusion dropping technology. Functional fillers of carboxyl decorated Fe3O4 nanoparticles and nitric acid modification of activated carbon imparted to the beads a convenient, operating-based, sensitive magnetic response and highly effective adsorption performance for Cu2+, Pb2+, and Zn2+. Adsroption experiments showed that these adsorption processes were spontaneous endothermic reactions, controlled by combining physical and chemical adsorptive mechanisms. The adsorption of Cu2+, Pb2+,

ASSOCIATED CONTENT

S Supporting Information *

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acssuschemeng.6b00790. Physical properties of native cellulose beads and MCB (Table S1), TEM image of Fe3O4 nanoparticles (a) and modified Fe3O4 nanoparticles (Figure S1), DR isotherm of Cu2+ (a), Pb2+ (b), and Zn2+ (c) onto MCB at 298 K (Figure S2) (PDF) 3967


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(16) Mohan, D.; Sarswat, A.; Ok, Y. S.; Pittman, C. U., Jr Organic and inorganic contaminants removal from water with biochar, a renewable, low cost and sustainable adsorbent − A critical review. Bioresour. Technol. 2014, 160 (0), 191−202. (17) Luo, X.; Zhang, L. High effective adsorption of organic dyes on magnetic cellulose beads entrapping activated carbon. J. Hazard. Mater. 2009, 171 (1−3), 340−347. (18) Chen, F. H.; Gao, Q.; Ni, J. Z. The grafting and release behavior of doxorubincin from Fe 3 O 4 @ SiO 2 core?shell structure nanoparticles via an acid cleaving amide bond: the potential for magnetic targeting drug delivery. Nanotechnology 2008, 19 (16), 165103. (19) Chingombe, P.; Saha, B.; Wakeman, R. J. Surface modification and characterisation of a coal-based activated carbon. Carbon 2005, 43 (15), 3132−3143. (20) Luo, X.; Zhang, L. High effective adsorption of organic dyes on magnetic cellulose beads entrapping activated carbon. J. Hazard. Mater. 2009, 171 (1−3), 340−347. (21) Luo, X.; Zhang, L. Creation of regenerated cellulose microspheres with diameter ranging from micron to millimeter for chromatography applications. J. Chromatogr. A 2010, 1217 (38), 5922−5929. (22) Kul, A. R.; Koyuncu, H. Adsorption of Pb(II) ions from aqueous solution by native and activated bentonite: Kinetic, equilibrium and thermodynamic study. J. Hazard. Mater. 2010, 179 (1−3), 332−339. (23) Habibi, N. Preparation of biocompatible magnetite-carboxymethyl cellulose nanocomposite: Characterization of nanocomposite by FTIR, XRD, FESEM and TEM. Spectrochim. Acta, Part A 2014, 131 (0), 55−58. (24) Nigam, S.; Barick, K. C.; Bahadur, D. Development of citratestabilized Fe3O4 nanoparticles: Conjugation and release of doxorubicin for therapeutic applications. J. Magn. Magn. Mater. 2011, 323 (2), 237−243. (25) López, F.; Medina, F.; Prodanov, M.; Güell, C. Oxidation of activated carbon: application to vinegar decolorization. J. Colloid Interface Sci. 2003, 257 (2), 173−178. (26) Vinke, P.; van der Eijk, M.; Verbree, M.; Voskamp, A. F.; van Bekkum, H. Modification of the surfaces of a gasactivated carbon and a chemically activated carbon with nitric acid, hypochlorite, and ammonia. Carbon 1994, 32 (4), 675−686. (27) Karthikeyan, S.; Sekaran, G. In situ generation of a hydroxyl radical by nanoporous activated carbon derived from rice husk for environmental applications: kinetic and thermodynamic constants. Phys. Chem. Chem. Phys. 2014, 16 (9), 3924−3933. (28) Luo, X.; Zhang, L. Immobilization of Penicillin G Acylase in Epoxy-Activated Magnetic Cellulose Microspheres for Improvement of Biocatalytic Stability and Activities. Biomacromolecules 2010, 11 (11), 2896−2903. (29) Pitakteeratham, N.; Hafuka, A.; Satoh, H.; Watanabe, Y. High efficiency removal of phosphate from water by zirconium sulfatesurfactant micelle mesostructure immobilized on polymer matrix. Water Res. 2013, 47 (11), 3583−3590. (30) Singh, S. K.; Townsend, T. G.; Mazyck, D.; Boyer, T. H. Equilibrium and intra-particle diffusion of stabilized landfill leachate onto micro- and meso-porous activated carbon. Water Res. 2012, 46 (2), 491−499. (31) Naeem, A.; Westerhoff, P.; Mustafa, S. Vanadium removal by metal (hydr)oxide adsorbents. Water Res. 2007, 41 (7), 1596−1602. (32) Weber, T. W.; Chakravorti, R. K. Pore and solid diffusion models for fixed-bed adsorbers. AIChE J. 1974, 20 (2), 228−238. (33) Başar, C. A. Applicability of the various adsorption models of three dyes adsorption onto activated carbon prepared waste apricot. J. Hazard. Mater. 2006, 135 (1−3), 232−241. (34) Foo, K. Y.; Hameed, B. H. Insights into the modeling of adsorption isotherm systems. Chem. Eng. J. 2010, 156 (1), 2−10. (35) Rieman, W.; Walton, H. F. Ion Exchange in Analytical Chemistry: International Series of Monographs in Analytical Chemistry; Elsevier, 2013.

AUTHOR INFORMATION

Corresponding Authors

*Tel.: +86-139-86270668. E-mail: [email protected]; [email protected] (X.L.). *Tel.: (+86-27) 8719-4980. Fax: (+86-27) 8719-4465. E-mail: [email protected]; [email protected] (F.Y.). Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS The authors thank the National Natural Science Foundation of China (51303142), the Natural Science Foundation of Hubei Province (2014CFA011, 2014CFB775), and the Open Foundation of Hubei Collaborative Innovation Center of Wuhan Institute of Technology (E201105, P201109) for the financial support.

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