Abstract

An efficient biosorbent containing magnetic nanoparticles, walnut shell powder, foam, and alginate (AMWSF) was prepared and used in Pb(II) removal. The adsorption process was performed in an airlift magnetic separation loop system. Optimum adsorption conditions were tested at pH 3–7, biomass dose of 0.03–0.4 g, temperature of 15–35 °C, initial Pb(II) ion concentration of 50–400 mg·L−1, and contact time of 10–480 min. The equilibrium adsorption capacity reached up to 69.45 mg·g−1. The physicochemical properties of AMWSF were analyzed by scanning electron microscopy coupled with energy-dispersive X-ray spectroscopy and Fourier transform infrared spectroscopy. The experimental data were in agreement with the pseudo-second-order kinetic and Langmuir isotherm models. The influences of Cu(II), Cd(II), and Zn(II) on Pb(II) adsorption showed antagonistic effect strength in the order of Cu(II) > Cd(II) > Zn(II). AMWSF was reused seven times and separated rapidly by magnetic field. The results demonstrated the potential of AMWSF in practical applications involving Pb(II).

INTRODUCTION

In recent years, with the fast-growing industrial activities, large numbers of heavy metals have been used in different manufacturing activities. Toxic metal ions have been discharged into aquatic ecosystems due to improper disposal. Heavy metals are non-biodegradable and accumulate in the food chain, so they present serious threats to ecosystems and human health. Therefore, it is necessary to remove heavy metal ions or transform them to less toxic forms in water bodies, in order to promote sustainable development. Pb(II) is an indispensable micronutrient to humans, but excessive Pb(II) would disrupt their metabolism and cause several diseases such as digestive disorder, autoimmune disorder, and liver, kidney, stomach and lung cancer (Gola et al. 2016). Therefore, heavy metal contamination in industrial wastewater is a global concern at present.

To date, many technologies have been used to decontaminate heavy metal ions from wastewaters, such as chemical precipitation, electrolysis, ion exchange, oxidation reduction, biological treatment, reverse osmosis and membrane filtration (Akram et al. 2017; Xu et al. 2017a; Shi et al. 2018). However, these methods have many disadvantages such as high cost, process complexity and secondary pollution (Yilmaz et al. 2011). These drawbacks have made the research focus on an effective technique for heavy metal ions removal. Biosorption has attracted considerable interest in the recent years because of important advantages such as high efficiency, low cost, ease of availability and environmental friendliness (Esmaeili & Khoshnevisan 2016; Safinejad et al. 2017). It has been found to be an appropriate choice and suitable for protection of the environment and human health. Various biosorbents have been applied to validate the removal performance of Pb(II). Still, these biosorbents have some defects such as difficult recycling and poor regeneration performance, preventing them from achieving a good status. To solve this issue, developing a biosorbent with excellent regeneration performance is necessary for Pb(II) removal.

Abundant and cheap walnut shell (WS) is often preferentially used in Pb(II) removal. This natural biomaterial contains lignin, cellulose, and hemicellulose, and its structure has various functional groups (Zhu et al. 2016). The efficiency of Pb(II) removal increases when WS is used in powdered form. In this work, a good biosorbent (referred to as AMWSF) was prepared by using alginate-coated magnetic nanoparticles, WS powder, and foam. Magnetic nanoparticles were used for rapid separation. In previous studies, biosorption was mainly used for the removal of a single heavy metal, but the water body is a complex system where several heavy metals often coexist in wastewater (Li et al. 2017). In Pb(II) removal, other heavy metals influence the process involved.

During adsorption, mechanical stirring may cause serious damage to an adsorbent owing to shear force and may thus affect solid–liquid separation and adsorbent recycling. For this reason, an airlift magnetic separation loop system was designed and an efficient and floating biosorbent was used for Pb(II) removal, as shown in Figure 1.

Figure 1

Schematic route for preparation of AMWSF and airlift magnetic separation loop system.

Figure 1

Schematic route for preparation of AMWSF and airlift magnetic separation loop system.

The aims of the study were as follows: (1) optimize adsorption conditions by investigating the influences of pH, temperature, contact time, biomass dose, and varying Pb(II) concentrations; (2) consider the adsorption mechanism according to the kinetics, isotherm models, and various analysis techniques; (3) investigate the effects of Cu(II), Cd(II), and Zn(II) on Pb(II) adsorption in a binary system; and (4) research the reusability performance of AMWSF through desorption experiments.

EXPERIMENTAL

Materials

The WS was collected from plant residues. Sodium alginate and calcium chloride were supplied by West Long Chemical Co., Ltd (Guangdong, China). Lead nitrate and other chemical agents were all procured from HuaLan Chemical Technology Co., Ltd (Shandong, China). Different concentrations of Pb(II) solutions were prepared from the stock solution of 0.1 mol·L−1 by dissolving 33.121 g of lead nitrate in deionized water until the resulting solution had a volume of 1,000 mL. All the required chemicals used were of analytical grade.

Preparation of magnetic nanoparticles, walnut shell and foam encapsulated in calcium alginate

Dirt from walnut shells was removed through repeated washing with deionized water. Then, the shells were dried in an oven at 80 °C for 12 h before they were ground in a ball grinder. The resulting powder particles with diameters of less than 0.6 mm were stored in a zip-top bag. Magnetic nanoparticles (n-Fe3O4) were prepared by the co-precipitation method (Mahmoud et al. 2016). In this procedure, 8.08 g of Fe(NO3)3·9H2O and 2.78 g of FeSO4·7H2O were dissolved in 25 mL of 0.5 M HCl. The mixture was then placed into a three-necked flask, which contained 250 mL of 1.5 M NaOH, with magnetic stirring (400 rpm) for 30 min at 80 °C. The obtained n-Fe3O4 particles were separated in the presence of a magnetic field, washed three times with deionized water, and finally dried at 60 °C until complete dryness. The particles were finally placed in a zip-top bag for later use. For AMWSF preparation, 2.5 g of sodium alginate was added to 100 mL of deionized water at 80 °C. The mixture was continuously stirred until the components were completely mixed. Five grams of powder was poured into the beaker and stirred for 30 min. Then, 1.0 g of n-Fe3O4 particles was mixed with the above mixture. The final mixture was stirred for another 30 min. The product was dropped into a 3% CaCl2 solution until AMWSF was obtained. AMWSF, which was separated from the CaCl2 solution after 24 h hardening, was rinsed three times with deionized water. The typical size of the AMWSF was about 5.0 mm in diameter. AMWSF was evaluated as Pb(II) adsorbent. As represented in Figure 1, an efficient biosorbent (AMWSF) was obtained, which can be used in Pb(II) removal.

Characterization of AMWSF

Before and after biosorption, the physical and chemical properties of AMWSF were characterized. Surface morphological structure and elemental analysis were conducted with a scanning electron microscope (SEM, Tokyo, Japan) coupled with an energy-dispersive X-ray spectrometer (EDX, Abingdon, Oxfordshire, UK). Fourier transform infrared (FTIR) spectra were obtained, to identify the functional groups on the tested AMWSF, in a PerkinElmer spectrometer using a KBr beam splitter with 4,000–400 cm−1 wavenumber and 4 cm−1 resolution.

Batch adsorption experiments

The experiments of Pb(II) adsorption on the prepared AMWSF were conducted to study the effects of adsorption parameters, such as pH (3–7), temperature (15–35 °C), biomass dose (0.03–0.4 g), contact time (10–480 min), and Pb(II) concentration (50–400 mg·L−1). At room temperature, the experiments were performed in 100 mL solutions in a 250 mL reactor. Uniform bubbles flowed into the aqueous solution from the bottom of the reactor. During the initial analyses, 100 mL of solutions with 100 mg·L−1 of Pb(II) and 0.1 g AMWSF, pH of 5.0, and contact time of 480 min was treated. Suspension pH was adjusted by using 0.1 M HCl and NaOH. The concentration of Pb(II) was determined by flame atomic absorption spectrometry (FAAS).

The adsorption amount could be calculated using equation: 
formula
(1)
where C0 and Ce (mg·L−1) are the initial and equilibrium concentration, respectively; qe (mg·g−1) represents the adsorption capacity; V (L) is the volume of solution and M (g) is the biomass dose (dry weight).

Kinetic and isotherm models

The adsorption curves were interpreted by fitting the pseudo-first-order and pseudo-second-order. These equations were described as follows: 
formula
(2)
 
formula
(3)
where k1 (min−1) and k2 (g·mg·min−1) are the rate constant of pseudo-first-order and pseudo-second-order, respectively; qe and qt (mg·g−1) are the amount of Pb(II) on AMWSF at equilibrium and at time t (min), respectively.
The adsorption mechanism was studied by applying the Langmuir and Freundlich isotherm models. The linearized equations were as follows: 
formula
(4)
 
formula
(5)
where Ce (mg·L−1) and qe (mg·g−1) represent equilibrium Pb(II) concentration and equilibrium adsorption capacity, respectively; qm (mg·g−1) is the maximum adsorption capacity; KL (L·mg−1) and KF (mg·g−1) are the Langmuir and Freundlich adsorption equilibrium constants, respectively.

Effects of co-existing heavy metals ions

Heavy metals (Cu(II), Cd(II), and Zn(II)) were selected in this experiment because they are often detected in industrial wastewater. The competitive effects of Cu(II), Cd(II), and Zn(II) on Pb(II) adsorption were investigated through binary adsorption experiments. Nitrate salts (Cd(NO3)2, Cu(NO3)2 and Zn(NO3)2) were used for the preparation of these metal solutions. The experiments were carried out at a constant Pb(II) concentration of 100 mg·L−1. Meanwhile, the concentration of co-existing heavy metal ions was 50, 100, 150, and 200 mg·L−1. The solution pH was controlled by 0.1 M NaOH and HCl. All the experiments were set up with three parallel groups, and the average was calculated.

Desorption study

HNO3 solution (1.0 M) was used as the desorbent for the Pb(II) desorption experiments. AMWSF with adsorbed Pb(II) was rinsed with deionized water for the removal of residual Pb(II) and then added into the HNO3 solution. The process was allowed to proceed in a 250 mL beaker under the following conditions: 0.1 g·L−1 biomass dose, pH 6, 25 °C, and 30 min. Finally, the supernatant was collected for the estimation of Pb(II) by FAAS. The obtained AMWSF was washed three times with deionized water for the subsequent adsorption, which was performed for checking the reusability of AMWSF. The adsorption process was performed for seven cycles.

RESULTS AND DISCUSSION

Characterization of AMWSF

The surface morphology of AMWSF was investigated with a SEM and is illustrated in Figure 2(a) and 2(b). After adsorption (Figure 2(b)), the AMWSF had a rough surface, and massive precipitation was observed in the section. The precipitates might be the deposited Pb(II). Figure 2(c) and 2(d) show the EDX analysis results for the AMWSF. The observed characteristic peak of Pb(II) after adsorption implied that Pb(II) was successfully adsorbed on the AMWSF. Pb(II) content increased by 14.21% after adsorption (Table 1). Finally, the EDX measurements confirmed the formation of Pb(II) precipitate on the AMWSF.

Table 1

Quantitative EDX analysis results before and after adsorption

Element Mass (%)
 
Before After 
45.52 35.31 
26.11 30.64 
Fe 9.56 9.56 
Pb 14.21 
Others 18.81 10.28 
Element Mass (%)
 
Before After 
45.52 35.31 
26.11 30.64 
Fe 9.56 9.56 
Pb 14.21 
Others 18.81 10.28 
Figure 2

SEM graphs of AMWSF: (a) before and (b) after adsorption; EDX spectra of AMWSF: (c) control and (d) Pb(II) treated.

Figure 2

SEM graphs of AMWSF: (a) before and (b) after adsorption; EDX spectra of AMWSF: (c) control and (d) Pb(II) treated.

For investigation of the removal mechanism, functional groups interacting with Pb(II) were determined by FTIR. The FTIR spectra obtained from AMWSF before and after adsorption were compared (solid and dashed lines, respectively; Figure 3). No peak appeared or disappeared, but the position of the peaks shifted. The bands at 3,422 cm−1 shifted to 3,414 cm−1, implying that –CO and –NH2 participate in Pb(II) adsorption. Another two changes observed from 1,624 cm−1 to 1,598 cm−1 and from 1,417 cm−1 to 1,420 cm−1 represented the amide I band stretching and –COO vibration, respectively, and indicated the binding of Pb(II) with the functional groups in the AMWSF surface. In addition, Fe-O appeared at 583 cm−1, confirming the binding of Fe3O4 nanoparticles into the AMWSF. The above observations demonstrated that carboxylic, carbonyl, hydroxyl, and amino groups participate in adsorption.

Figure 3

FTIR spectra of AMWSF in the absence and presence of Pb(II).

Figure 3

FTIR spectra of AMWSF in the absence and presence of Pb(II).

Effects of environmental factors on Pb(II) adsorption

The performance of AMWSF for Pb(II) removal was mainly affected by pH, biomass dose, temperature, and initial Pb(II) concentration. Hence, maximum adsorption capacity was obtained through several experiments, which were described earlier in the paper.

To prevent precipitation, the experiments were performed at pH ≤7. As shown in Figure 4(a), the adsorption capacity of Pb(II) increased with increasing of pH, and the maximum adsorption ability was 71.87 mg·g−1. At an extremely acidic pH, the number of H+ ions competing with Pb2+ ions increased in the adsorption active sites. At high pH values, the dissociation of protons occurs from the binding sites, and the amount of Pb2+ available for adsorption increased (Tiwari et al. 2017). Moreover, some Pb2+ ions translated to Pb(OH)2 at pH > 6 and impeded the combination between the other Pb2+ ions and surfaces of the functional groups of the biosorbents. Therefore, pH 6 was selected for subsequent adsorption experiments.

Figure 4

Effect of pH (a), biomass dose (b), temperature (c), and Pb(II) concentration (d) on the biosorption of Pb(II).

Figure 4

Effect of pH (a), biomass dose (b), temperature (c), and Pb(II) concentration (d) on the biosorption of Pb(II).

The amount of AMWSF dose was varied from 0.03 g·L−1 to 0.4 g·L−1. As shown in Figure 4(b), the removal efficiency of Pb(II) increased at increased AMWSF dose, whereas the adsorption capacity was reduced. The low adsorption at a high dose may be due to the constant initial Pb(II) concentration, which was insufficient to completely cover the available active sites on the AMWSF (Masoumi et al. 2016). To summarize, an AMWSF dose of 0.1 g·L−1 was the point at which high removal efficiency was observed and was thus selected for the next experiments.

The effect of temperature on Pb(II) adsorption was studied. As shown in Figure 4(c), the temperature of the adsorption was less significant because the biosorption conformed to energy-independent mechanisms (Ozdemir et al. 2009). This result is consistent with those of some studies on Pb(II) biosorption (Masoudzadeh et al. 2011; Khadivinia et al. 2014). The adsorption experiments were performed at room temperature.

The effect of initial Pb(II) concentration, which ranged from 50 mg·L−1 to 400 mg·L−1, was studied. As shown in Figure 4(d), the adsorption capacity of Pb(II) increased at concentrations of up to 300 mg·L−1 and stabilized thereafter. The high influent metal concentration provided a large driving force for overcoming the entire mass transfer resistance between the aqueous solution and biosorbent (Yu et al. 2017). Meanwhile, adsorption stability can be attributed to the lack of binding sites for complexation at high Pb(II) concentrations.

Adsorption kinetics

To predict the rate for Pb(II) removal and interpret the kinetic mechanism of adsorption process, the effect of contact time on the removal efficiency of Pb(II) was studied. The adsorption capacity increased sharply within 60 min and then slowly until it reached equilibrium at 180 min (Figure S1, available with the online version of this paper). Therefore, 180 min was selected as the optimum contact time in this study. The initial fast adsorption was attributed to sufficient unutilized activated sites and low mass transfer resistance (Wang et al. 2017). In the subsequent step, the adsorption process was slowed possibly because of the decrease in available binding sites, and thus the Pb(II) adsorption capacity remained constant.

The pseudo-first- and pseudo-second-order kinetic models were employed to fit with the experimental data. The obtained correlation coefficient (R2) was used for assessing the applicability of each model. The model with the higher R2 value was considered a more suitable model (Islam et al. 2017). The calculated results of kinetic parameter (Figure S2 and Table 2; Figure S2 is available with the online version of this paper) suggested that the Pb(II) data can be simulated better by the pseudo-second-order kinetic model (0.9996) than by the pseudo-first-order kinetic model (0.9641). Moreover, the calculated constants generated by the pseudo-second-order kinetic model were close to the experimental data, implying that Pb(II) adsorption was strongly dependent on available activated sites, and the pseudo-second-order kinetic mechanism was dependent on the concentrations and nature of the adsorbate and adsorbents (Tiwari et al. 2017).

Table 2

Model results of AMWSF adsorption kinetics data

 Pseudo-first-order
 
Pseudo-second-order
 
qe,exp (mg·g−1k1 (min−1qe,1 (mg·g−1R2 k2 (g·mg−1·min−1qe,2 (mg·g−1R2 
69.45 0.0096 32.47 0.9641 0.0011 71.74 0.9996 
 Pseudo-first-order
 
Pseudo-second-order
 
qe,exp (mg·g−1k1 (min−1qe,1 (mg·g−1R2 k2 (g·mg−1·min−1qe,2 (mg·g−1R2 
69.45 0.0096 32.47 0.9641 0.0011 71.74 0.9996 

Adsorption isotherm

To elucidate the Pb(II) adsorption mechanism, adsorption behavior was studied by two experimental adsorption isotherms. The Langmuir and Freundlich models were utilized to fit the experimental data and reflect adsorption type. The Langmuir model was applied for explaining homogeneous adsorption systems involving monolayer adsorption where no interaction among sorbed molecules occur. The Freundlich equation was proposed as an empirical equation for multilayer heterogeneous adsorption (Benettayeb et al. 2017; Zhang et al. 2017).

The corresponding results are summarized in Figure S3 and Table 3 (Figure S3 is available with the online version of this paper). The R2 value obtained from the Langmuir isotherm (0.9874) was higher than that of the Freundlich isotherm (0.7752), indicating that the adsorption behavior followed the Langmuir model equation. Moreover, the theoretical maximum adsorption values from the Langmuir equation were all 170.36 mg·g−1 for Pb(II), which was close to the experimental value (160.45 mg·g−1). The effect of the isotherm shape can be expressed in terms of the equilibrium parameter RL, also known as the separation factor, which can be used to predict whether an adsorption system is unfavorable or favorable. RL can be calculated as 1/(1 + KL · C0). If RL = 0, adsorption is irreversible; if 0 < RL < 1, it is favorable; if RL = 1, it is linear; and adsorption is unfavorable if RL > 1. In this study, RL value was 0.9423 for the initial Pb(II) concentration of 100 mg·L−1, suggesting that the adsorption process was favorable. In summary, the results showed that monolayer adsorption of the active sites on the AMWSF surface played a major role in Pb(II) removal.

Table 3

Langmuir and Freundlich isotherm parameters for adsorption of Pb(II) on AMWSF

 Langmuir
 
Freundlich
 
qm,exp (mg·g−1qm (mg·g−1KL (L·mg−1R2 KF (mg·g−1n R2 
160.45 170.36 0.9423 0.9874 42.476 3.9261 0.7752 
 Langmuir
 
Freundlich
 
qm,exp (mg·g−1qm (mg·g−1KL (L·mg−1R2 KF (mg·g−1n R2 
160.45 170.36 0.9423 0.9874 42.476 3.9261 0.7752 

Effects of co-existing heavy metals ions

In general, industrial effluents are multi-metal ion systems. Other metal ions interfere with the adsorption of target metal ions (Jones et al. 2016; Wang et al. 2016). Compared with common metal ions, heavy metal ions considerably affect adsorption. Several studies reported that the effects are mostly synergetic, antagonistic, and with no interaction. Thus, determining the effects of Cu(II), Cd(II), and Zn(II) on Pb(II) adsorption by AMWSF is necessary. In the present study, the effects of competing ions were studied by using a binary system (Cu/Pb, Cd/Pb and Zn/Pb). As plotted in Figure 5, the adsorption capacity of Pb(II) showed a downward trend in binary systems with increased Cu(II), Cd(II) and Zn(II) concentrations. The increase was attributed to the heavy metal ions that competed with Pb(II) for available sites on the AMWSF. The results indicated that these ions had an inhibiting effect on Pb(II) adsorption, and this effect was intensified when the concentrations of the competing ions increased. The difference between the Pb(II) adsorption capacities of the binary systems was attributed to the different physicochemical characteristics of heavy metal ions (Wang et al. 2017). Some studies suggested that electronegativity plays a vital role in metal adsorption processes. Heavy metals have different ionic radii, indicating that their electronegativity varies. The electronegativity values were 2.33, 1.90, 1.69 and 1.65 for Pb(II), Cu(II), Cd(II) and Zn(II), respectively. The higher electronegativity of heavy metals meant they were more easily attracted by adsorbents (Wang et al. 2018). Thus Pb(II) had a larger covalent index and exhibited a higher affinity than Cu(II), Cd(II) and Zn(II), which was also the reason why the adsorption capacity of Pb(II) showed no significant decline in this study.

Figure 5

Effect of Cu, Cd, and Zn(II) concentration on the Pb(II) adsorption.

Figure 5

Effect of Cu, Cd, and Zn(II) concentration on the Pb(II) adsorption.

Desorption study

For practical application, biosorbent reuse is extremely important because this procedure reduces processing cost (Xu et al. 2017b). To investigate the reusability of AMWSF, the desorption experiments were repeated for seven cycles. In this study, HNO3 solution (1.0 M) was used for the desorption of Pb(II). Then, AMWSF was rinsed with deionized water for reuse in the next cycle. Finally, AMWSF was rapidly separated from the liquid phase by using an external magnetic field. From Figure 6, the adsorption capacity of AMWSF had a decreasing trend with increasing reuse, and the decline was ascribed to the activated binding sites occupied by Pb(II) having not released completely. However, no significant decline was observed, and the airlift loop system reduced the mass loss of AMWSF. Furthermore, the adsorption capacity of AMWSF when used for one and seven times were 65.93 and 41.12 mg·g−1, respectively. Meanwhile, the loss rate of AMWSF decreased by 8.41% after seven consecutive cycles. Finally, AMWSF is suitable for cyclic utilization for Pb(II) removal and this good performance implied its high potential application.

Figure 6

Adsorption capacity and loss rate of biomass after seven repeated cycles for Pb(II) biosorption on AMWSF.

Figure 6

Adsorption capacity and loss rate of biomass after seven repeated cycles for Pb(II) biosorption on AMWSF.

CONCLUSIONS

In this study, the optimization of pH, biomass dose, temperature, initial concentration, and contact time was achieved at 6, 0.1 g, 25 °C, and 180 min, respectively. The equilibrium adsorption capacity of up to 69.45 mg·g−1 was achieved, and the adsorption process can be described well by the pseudo-second-order kinetic and Langmuir isotherm models. The adsorption mechanism demonstrated that the deposition of Pb(II) was due to the functional carboxylic, hydroxyl, carbonyl and amino groups. Co-existing heavy metals (Cu(II), Cd(II) and Zn(II)) have an antagonistic effect during Pb(II) removal, and AMWSF can be efficiently reused up to seven times because of the airlift magnetic separation loop system that does not cause shear force damage to AMWSF. Overall, AMWSF was found to be an efficient adsorbent for Pb(II) removal in an airlift magnetic separation loop system.

ACKNOWLEDGEMENTS

This work was supported by the Natural Science Foundation of Inner Mongolia Autonomous Region (2017MS0217) and Science Research Program of Higher Education Institutions of Inner Mongolia Autonomous Region (NJZY16084).

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Supplementary data