Abstract

Functional magnetic Fe3O4@PPy microspheres were prepared and characterized by XRD, FTIR, SEM, TEM, and magnetometer, and the adsorption of Hg(II) onto Fe3O4@PPy was investigated. The results showed that the adsorption of Hg(II) onto Fe3O4@PPy dramatically increases within 5 min and reaches adsorption equilibrium at 200 min. The adsorption of Hg(II) increases with pH increased, and a removal efficiency (RE) of 90.5% was obtained at pH 7.2. The isotherm studies revealed that the adsorption of Hg(II) onto the Fe3O4@PPy fits well with the Langmuir isotherm model, and the calculated qm value of 232.56 mg/g. The adsorption process of Hg(II) onto the Fe3O4@PPy is well-fitted by the pseudo-second-order model with a high correlation coefficient (R2) of 0.999. The thermodynamic coefficients (ΔH°, ΔS°, and ΔG°) were calculated from the temperature-dependent adsorption isotherms and illustrated that the adsorption of Hg(II) on the Fe3O4@PPy was spontaneous and endothermic. Different desorption agents were used to recover Hg(II) adsorbed onto Fe3O4@PPy, and a satisfactory recovery percentage of 93.0% was obtained by using 0.1 M HCl and 0.05 M NaCl.

HIGHLIGHTS

  • Amino-functional magnetic Fe3O4@PPy microspheres were prepared and characterized.

  • The Hg(II) can be adsorbed by Fe3O4@PPy effectively.

  • The used Fe3O4@PPy could be regenerated by two-component desorbent of HCl and NaCl.

Graphical Abstract

Graphical Abstract
Graphical Abstract

INTRODUCTION

Mercury (Hg) is considered one of the most important environmental contaminants due to its bioaccumulation and strong biological toxicity. The presence of Hg can threaten human health even at a trace level (Zhang et al. 2017; Zhao et al. 2018). For all three species of mercury (elemental (Hg0), metallic (Hg(II)), and organic (MeHg)), Hg(II) and MeHg have higher toxicity to living organisms. Therefore, it is necessary to develop the efficient, fast, and economically feasible technology for Hg(II) removal (Yuan et al. 2014). Many physical, biological, and chemical techniques have been employed to remove the Hg(II) from wastewaters, such as membrane separation, ion exchange, chemical precipitation, coagulation, and adsorption (Zhang et al. 2012, 2020; Sharma et al. 2015; Wang et al. 2020). However, there are many problems with the current technology because of low efficiency, high cost, and complicated operation. Among these techniques, the adsorption method has been widely studied due to its low cost, simple operation, and high efficiency (Liu et al. 2018; Guo et al. 2019). Many adsorbents, such as activated carbon (Alomar et al. 2017; Huang et al. 2021), mesoporous silica (Antochshuk et al. 2003), chitosan (Sampaio et al. 2015), and magnetic nanocomposite (Kim & Park 2017; Xiao et al. 2019), have been used to remove the Hg(II) from the aqueous solution. Among these adsorbents, the amino-functionalized magnetic nanocomposite has gradually attracted the interest of researchers due to their superparamagnetic properties, biocompatibility, and easy to surface modification. The amino groups in nanocomposite can interact with metal ions through ion exchange or chelation and can also improve its adsorption features, such as selectivity and adsorption capacity (Zhou et al. 2019). The investigation of Wang et al. (2012) showed that the amino-functionalized magnetic composite microspheres have a good adsorption effect on the metal chromium ions in the solution. The core–shell Fe3O4@polypyrrole composite (Fe3O4@PPy) is a typical amino-functionalized magnetic nanocomposite which has been widely investigated in recent years (Peng et al. 2015; Tang et al. 2017). However, the applications of Fe3O4@PPy to remove Hg(II) from solution are rare.

In this work, Fe3O4@PPy composite microspheres are prepared and characterized by different techniques. After that, the Hg(II) adsorption behavior under different influence factors (pH, initial Hg(II) concentration, and contact time), the adsorption isotherms, kinetics, thermodynamics, and the reusability of Fe3O4@PPy were investigated. Finally, the optimal conditions and removal mechanism were determined.

MATERIALS AND METHODS

Chemicals and materials

FeCl3·6H2O (>99%), Hg(NO3)2 (>98.5%), and sodium acetate anhydrous (>99%) were purchased from Tianjin Chemical Reagent Third Factory, China. Pyrrole and glycol (>99.5%) were obtained from Shanghai Macklin Biochemical Co., Ltd, China. Absolute ethanol and polyethylene glycol (PEG)-6000 were from Tianjin Fuyu Fine Chemical Co., Ltd, China. HCl solution (36–38%), HNO3 (65%), and NaOH (>96%) were received from Tianjin Northern Tianyi Chemical Reagent Factory, China. All the chemicals were used without further purification. Distilled deionized water (18.4 MΩ cm) was used for material synthesis and aqueous experiments.

Methods of synthesis

Preparation of Fe3O4 microspheres

The Fe3O4 microspheres were prepared through a polyol reduction method (Peng et al. 2015; Figure 1). Firstly, 6 mM FeCl3·6H2O was dissolved in 40 mL glycol and stirred until evenly mixed. Subsequently, 42.9 mM sodium acetate anhydrous and 1.0 g of polyethylene glycol (PEG)-6000 were added and the mixed solution was stirred vigorously for 0.5 h. After that, the mixture was transferred to the 50 mL Teflon autoclave, and heated at 200 °C for 12 h. Then, the sample was collected with a permanent magnet, and washed with distilled water and ethanol. The above separation process is repeated three times. Finally, the resulting Fe3O4 microspheres were dried in a vacuum oven at 40 °C for 6 h.

Figure 1

The synthesis procedure of Fe3O4@PPy composite microspheres.

Figure 1

The synthesis procedure of Fe3O4@PPy composite microspheres.

Preparation of magnetic Fe3O4@PPy microspheres

Fe3O4 microspheres (0.10 g) were dispersed in HCl solution (40 mL, 0.1 M) under sonication, and then, the mixture was sealed in a three-necked flask and left to stand for 12 h. Subsequently, the sample was collected with the help of a magnet, washed with deionized water repeatedly to remove the residual HCl. After that, 10 mL absolute ethanol and 0.15 mL redistilled pyrrole were added into the flask under sonication and N2 atmosphere, and then, the mixture was sealed and left to stand at 5 °C for 12 h. Then, HCl (0.165 mL, 12 M) and FeCl3·6H2O solution (30 mL, 1.1 M) were slowly added into the above mixture in turn under 5 ± 1 °C and N2 atmosphere, stirring and sonication for 1 h. Finally, the precipitate was separated, washed with distilled water and ethanol, and then dried in a vacuum oven at 40 °C for 12 h (Bhaumik et al. 2011).

Instruments

The Fe3O4@PPy composite was characterized by X-ray diffraction (XRD) using Cu Kα radiation (λ = 1.5406 Å) at 45 kV/40 mA, for 2θ values between 10° and 90° (D8 Advance, Bruker, Germany). The XRD patterns obtained were analyzed by Jade 9.1 software program. The infrared spectroscopic information was recorded by an IRPrestige-21 Fourier transform infrared spectroscopic spectrometer (FTIR, Shimadzu, Japan) for the functional group analysis. It was collected using pressed KBr discs with a resolution of 4 cm−1 over the range of 4,000–500 cm−1 on a Fourier transform infrared. The surface morphology of the composite was studied by scanning electron microscopy (SEM; Quanta 200E, FEI, USA) and transmission electron microscopy (TEM; JEM-2100F, JEOL, Japan). Magnetic measurements were performed with a superconducting quantum interference device (SQUID) magnetometer (MPMS XL-7, Quantum Design, USA). To determine Hg(II) ion removal by the adsorbent, the Hg(II) concentration in the remaining solution was measured with an inductively coupled plasma optical emission spectrometer (ICP-OES Agilent 5100, USA).

Adsorption investigation

The different concentrations (5–50 mg/L) of Hg(II)-containing solution were prepared by dissolving the Hg(NO3)2 into deionized water, and the pH was adjusted with 0.1 M HNO3 or NaOH. For each experiment, 5 mg Fe3O4@PPy was dispersed into 20 mL Hg(II) solution, and then kept for a preset time (0.05–5 h) and temperature (25–45 °C). After reaching the adsorption equilibrium, the Fe3O4@PPy were separated with a permanent magnet, and the Hg(II) concentration in the remaining solution was analyzed using ICP-OES. The removal efficiency (RE) and adsorption capacity (qe) of adsorbent are calculated using the following equations (Cao et al. 2020), respectively:
formula
(1)
formula
(2)
where qe is the amounts of Hg(II) adsorbed (mg/g) at equilibrium, c0 is the initial Hg(II) concentration (mg/L), ce is the equilibrium Hg(II) concentration (mg/L), m is the mass of Fe3O4@PPy used (g), and V is the volume of solution (L).

Desorption investigation

1 mg Fe3O4@PPy with adsorbed Hg(II) ions was dispersed in 20 mL solution containing desorption agent and shaken in a conical flask for 3 h. After that, the adsorbent was separated, and the concentration of Hg(II) ions in solution was analyzed by ICP-OES for calculating Hg(II) the recovery rate.

RESULTS AND DISCUSSION

Characterization of Fe3O4@PPy microspheres

Figure 2 shows the XRD pattern of the Fe3O4@PPy microspheres. The diffraction peaks at 31.17°, 35.50°, 43.12°, 53.58°, 57.03°, and 62.63° correspond to (220), (311), (400), (511), and (440) crystal planes of the cubic anti-spinel structure, respectively. It is shown that magnetic Fe3O4 is not oxidized and Fe3O4@PPy microspheres are well crystallized in the preparation process of acidic proton and oxidative polymerization.

Figure 2

XRD pattern of Fe3O4@PPy.

Figure 2

XRD pattern of Fe3O4@PPy.

The FTIR spectrum of Fe3O4@PPy is presented in Figure 3. As shown in Figure 3, the strong absorption band at 579 cm–1 corresponds to the Fe–O stretching vibration of Fe3O4; the band at 1,556 cm–1 is attributed to the C = C vibrations of pyrrole rings; and the bands at 1,245, 1,033, and 773 cm–1 are attributed to C–N stretching vibration, C–H out-of-plane bending vibrations, and N–H in-plane bending vibrations, respectively. From these results, it can be inferred that the monomers have been polymerized successfully to be Fe3O4@PPy microspheres.

Figure 3

FTIR spectrum of Fe3O4@PPy.

Figure 3

FTIR spectrum of Fe3O4@PPy.

The SEM and TEM images of the Fe3O4@PPy microspheres are shown in Figure 4. From the SEM image in Figure 4, the thickness of the clear uniform shell is about 80 nm. It is clearly seen that the Fe3O4 core is encapsulated with polypyrrole coating, indicating the successful polymerization of pyrrole in the present reaction system and the formation of Fe3O4@PPy composite microspheres. It can be shown from TEM images that the Fe3O4@PPy microspheres with a uniform diameter size of about 290 nm have been prepared. The diameter is almost consistent with the Fe3O4@PPy prepared by previous study (Zhang et al. 2017). TEM images also show the magnification image of the Fe3O4@PPy microspheres retaining good dispersibility.

Figure 4

SEM (a) and TEM (b) images of Fe3O4@PPy.

Figure 4

SEM (a) and TEM (b) images of Fe3O4@PPy.

Figure 5 shows the saturation magnetization curves of Fe3O4@PPy microspheres. The saturation magnetization values of Fe3O4@PPy are 49.6 emu g–1, and the coercive force of Fe3O4@PPy microspheres is 28.85 Oe, indicating that the coated microspheres have good magnetic properties (Morel et al. 2008).

Figure 5

Saturation magnetization curve of Fe3O4@PPy.

Figure 5

Saturation magnetization curve of Fe3O4@PPy.

Effect of initial Hg(II) ions concentration on Hg(II) adsorption

The effect of initial ions concentration on the adsorption of Hg(II) by Fe3O4@PPy is shown in Figure 6. It can be seen that the RE and qe decreased and increased with initial Hg(II) concentration increasing from 5.0 to 60 mg/L, respectively. In particular, the qe increased sharply with the initial concentration increasing from 5 to 10 mg/L; however, the adsorption capacity increased slowly at initial concentration above 20 mg/L. This phenomenon may indicate that the adsorption sites of Fe3O4@PPy were not completely occupied at the low initial concentration of Hg(II); therefore, the adsorption capacity has not reached equilibrium. However, the number of free adsorption sites decreased with the Hg(II) ions concentration increases, resulting in part of Hg(II) ions cannot be combined with the adsorbent, and thus, the increase of qe becomes less obvious. Li et al. (2019) found that a novel hierarchical carbon/Fe–Mn composite readily fabricated from biomass was utilized as an adsorbent for Hg(II) removal. The composite exhibited high removal efficiency of 96.8%, and considerable adsorption capacity of 9.8 mg/g. Zabihi et al. (2010) fabricated porous carbons from walnut shells, which exhibited a high monolayer adsorption capacity of 151.5 mg/g for Hg(II) removal. In the results of this study, the maximum RE and qe were 93.86% and 232.55 mg/g, respectively, indicating that the Fe3O4@PPy can effectively adsorb the Hg(II).

Figure 6

Effect of initial ions concentration on the adsorption of Hg(II) by Fe3O4@PPy (adsorption temperature, 25 °C; adsorption time, 5 h; solution pH, 6.0; Fe3O4@PPy dosage, 0.25 g/L).

Figure 6

Effect of initial ions concentration on the adsorption of Hg(II) by Fe3O4@PPy (adsorption temperature, 25 °C; adsorption time, 5 h; solution pH, 6.0; Fe3O4@PPy dosage, 0.25 g/L).

Effect of the adsorbent dosage on Hg(II) adsorption

The effect of the Fe3O4@PPy dosage on Hg(II) removal is shown in Figure 7. As can be seen that the RE increases gradually with the increase of the adsorbent dosage. When the dosage is 0.25 g/L, the RE is 85.12%, and the RE reaches almost 90% at the dosage of 0.4 g/L. This phenomenon can be attributed to more available adsorption sites and the increase of adsorption contact area (Peng et al. 2015; Ma et al. 2020). On the other hand, the qe decreased with the increase of dosage. At low dosage, the number of Hg(II) ion in solution is much more than the binding sites of adsorbents. Therefore, most adsorption sites will be occupied by Hg(II). As the dosage increases, the number of adsorption sites increases rapidly. It indicates that some adsorption sites remained in an unsaturated state on Hg(II) adsorption, resulting in a decrease in the utilization rate of the Fe3O4@PPy (Mollahosseini et al. 2019).

Figure 7

Effect of the adsorbent dosage on Hg(II) adsorption (adsorption temperature, 25 °C; adsorption time, 5 h; solution pH, 6.0; initial Hg(II) concentration, 10 mg/L).

Figure 7

Effect of the adsorbent dosage on Hg(II) adsorption (adsorption temperature, 25 °C; adsorption time, 5 h; solution pH, 6.0; initial Hg(II) concentration, 10 mg/L).

In a similar study to explore the effect of the adsorbent dosage on Hg(II) adsorption, Ghasemi et al. (2019) using polydopamine decorated SWCNTs found that the qe of 55 mg/g and the RE of 82% for Hg(II) removal can be achieved at the adsorbent dosage of 0.3 g/L. Consequently, it can be concluded that Fe3O4@PPy nanocomposite is a more effective adsorption media for the removal of Hg(II) from aqueous solution.

Effect of solution pH on Hg(II) adsorption

The RE at different pHs is presented in Figure 8. As can be seen that the RE of Hg(II) sharply increased from ∼8 to ∼86% as pH was raised from 2.1 to 6.0. Furthermore, when the pH was increased from 6.0 to 7.2, the RE increased gradually from ∼86 to ∼91%. It is attributed that the amino group of the pyrrole will protonate in strong acidic solution to form positive sites such as , resulting in electrostatic repulsion between the Hg(II) ions and the adsorbent. The dominant form of Hg(II) is in the form of stable compounds without electrostatic interaction (Zhang et al. 2012). Therefore, the adsorption capacity of Fe3O4@PPy is very low when the pH is less than 4.

Figure 8

Effect of pH on the Hg(II) adsorption (adsorption temperature, 25 °C; adsorption time, 5 h; initial Hg(II) concentration, 10 mg/L; Fe3O4@PPy dosage, 0.25 g/L).

Figure 8

Effect of pH on the Hg(II) adsorption (adsorption temperature, 25 °C; adsorption time, 5 h; initial Hg(II) concentration, 10 mg/L; Fe3O4@PPy dosage, 0.25 g/L).

At high pH, deprotonation occurred on the surface of Fe3O4@PPy, which reduced the electrostatic repulsion between the Hg(II) ions and the adsorbent. The amino groups of the pyrrole can combine with OH in the solution to form negatively charged NH2OH, significantly enhancing the ability to adsorb Hg(II) (Deb et al. 2017). In the alkaline media, and complexes could be formed, which cannot be adsorbed by Fe3O4@PPy, because these complexes are water-insoluble (Mollahosseini et al. 2019).

Adsorption kinetics

The Hg(II) adsorption on Fe3O4@PPy was investigated as a function of contact time, and the adsorption data are analyzed by the pseudo-first-order kinetics, pseudo-second-order kinetics, and Elovich models. The linear forms of the above three models are expressed as follows (Zhang et al. 2019):
formula
(3)
formula
(4)
formula
(5)
where qt (mg/g) is the amount of the Hg(II) adsorbed on Fe3O4@PPy at time t (min); k1 (min−1) and k2 (g mg−1min−1) are the adsorption rate constant of pseudo-first-order and pseudo-second-order kinetics, respectively; a (mg g−1min−1) is the initial adsorption rate; and b (g mg−1) is related to the extent of surface coverage and activation energy for chemisorption. The adsorption data and fitting curves are shown in Figure 9. The parameters of the three models are listed in Table 1.
Table 1

Parameters pseudo-first-order kinetics, pseudo-second-order kinetics, and Elovich models

Pseudo-first-order
Pseudo-second-order
Elovich
k1qeR2k2qeR2abR2
0.015 36.97 0.811 2.24 × 10−3 174.2 0.999 4.64 × 107 0.160 0.980 
Pseudo-first-order
Pseudo-second-order
Elovich
k1qeR2k2qeR2abR2
0.015 36.97 0.811 2.24 × 10−3 174.2 0.999 4.64 × 107 0.160 0.980 
Figure 9

Adsorption kinetics data (a) and fitted curves of pseudo-first-order kinetics (b), pseudo-second-order kinetics (c), and Elovich model (d).

Figure 9

Adsorption kinetics data (a) and fitted curves of pseudo-first-order kinetics (b), pseudo-second-order kinetics (c), and Elovich model (d).

As shown in Figure 9(a), the amount of Hg(II) adsorption on Fe3O4@PPy sharply increases within 5 min, and then reached equilibrium within approximately 200 min. The R2 (>0.999) of the pseudo-second-order model is higher than that of two other models, and the qe value (174.2 mg/g) is calculated from the pseudo-second-order model approximately equal to the experimental value (173.76 mg/g), indicating that the adsorption process may be chemisorption involving force through sharing or exchange of electrons between the Hg(II) and Fe3O4@PPy (Ghasemi et al. 2019). Elovich model can describe the chemisorption process on the highly heterogeneous adsorbent. The correlation coefficients of the Elovich model in this study is 0.980, indicating that there is a coordination bond between the Fe3O4@PPy and Hg(II) (Anbia et al. 2015).

Adsorption isotherms

The Langmuir, Freundlich, and D-R models were used for analyzing the adsorption data and to understand the adsorption mechanism.

The Langmuir model assumes monolayer adsorption on the homogeneous surface, and adsorption occurs at specific adsorption sites. The linear form of the Langmuir model can be expressed by the following equation (Guo et al. 2018):
formula
(6)
where KL is Langmuir adsorption constant (L/mg), and qm represents the maximum Hg(II) adsorption capacity of Fe3O4@PPy (mg/g). ce and qe are defined above.
The linear form of the Freundlich model can be expressed by the following equation:
formula
(7)
where KF (mg1−nLng−1) and n are the Freundlich constants, the other parameters are defined above.
The linear form of the D-R model can be expressed as follows (Xiao et al. 2020):
formula
(8)
where qDe (mol g−1) is the adsorbed molar mass of Hg(II) per gram of Fe3O4@PPy; qDm (mol g−1) is the maximum adsorption capacity; β is the activity coefficient related to mean adsorption energy (mol2 J−2); and ɛ is the Polanyi potential (kJ2mol−2), as can be calculated from the following equation:
formula
(9)
where R is the gas constant (8.314 J mol–1K–1) and T is the absolute temperature (K). The mean adsorption energy E (kJ mol–1) can be calculated from β, and expressed as follows:
formula
(10)

The fitted curves and parameters of the three models are shown in Figure 10 and Table 2. The calculated R2 values of three models indicated that the adsorption followed the Langmuir model very well. The calculated qm value of 239.76 mg/g is very close to the experimental value of 232.55 mg/g. The results suggested that the binding energy of the Fe3O4@PPy surface was uniform, namely the Fe3O4@PPy surface exhibited the identical adsorption activity. Generally, the E values range from 8 to 16 kJ mol−1 and less than 8 kJ mol−1 correspond to the chemical adsorption and physical adsorption mechanism, respectively (Ghasemi et al. 2019). The E value calculated in this study is 9.13 kJ mol−1, indicating that the chemical adsorption is the dominating mechanism in the Hg(II) adsorption process.

Table 2

Parameters of the Langmuir, Freundlich, and D-R models

Langmuir model
Freundlich model
D-R model
qmKLR21/nKFR2βER2
239.76 0.614 0.997 0.258 3.31 × 105 0.755 6.0 × 10−3 9.13 0.973 
Langmuir model
Freundlich model
D-R model
qmKLR21/nKFR2βER2
239.76 0.614 0.997 0.258 3.31 × 105 0.755 6.0 × 10−3 9.13 0.973 
Figure 10

The fitted curves by the Langmuir, Freundlich (a) and D-R (b) models.

Figure 10

The fitted curves by the Langmuir, Freundlich (a) and D-R (b) models.

Adsorption thermodynamics

The thermodynamic parameter can be used to determine the spontaneity of adsorption and the adsorption mechanism (Liu et al. 2020). The adsorption thermodynamics of Hg(II) adsorption on Fe3O4@PPy was investigated at 298–318 K, and the thermodynamic parameters such as the enthalpy (ΔH°), the entropy (ΔS°), and the Gibbs energy (ΔG°) are calculated by Van der Hoff equation (Equations (11)–(13)) (Zou et al. 2020). The experimental conditions were adsorption equilibrium time 5 h, initial Hg(II) concentration 10 mg/L, adsorbent dosage 0.25 g/L, and pH 6.0.
formula
(11)
formula
(12)
formula
(13)
where Kd is the distribution coefficient, other parameters are defined above. The ΔG° and ΔH° can be calculated by the plot of ln(Kd) versus 1/T according to Equation (11), and the ΔS° can be calculated by Equation (12). The values of three parameters at different temperatures are listed in Table 3.
Table 3

The adsorption thermodynamic parameters of Hg(II) upon Fe3O4@PPy

T (K)ΔG° (kJ mol−1)ΔH° (kJ mol−1)ΔS° (J mol−1K−1)
298 −2.17 40.05 141.84 
303 −2.99 
308 −3.67 
313 −4.37 
318 −5.02 
T (K)ΔG° (kJ mol−1)ΔH° (kJ mol−1)ΔS° (J mol−1K−1)
298 −2.17 40.05 141.84 
303 −2.99 
308 −3.67 
313 −4.37 
318 −5.02 

From Table 3, it can be seen that the values of ΔG° are more negative with temperature increasing from 298 to 318 K, indicating that the Hg(II) adsorption on Fe3O4@PPy is spontaneous. The positive value of ΔH° indicated that the adsorption process is endothermic. The positive value of ΔS° suggested that the adsorbed Hg(II) on the Fe3O4@PPy surface are organized in a more random fashion compared with those in the aqueous phase. Generally, the value of ΔH° for the physical and chemisorption is in the range of 2.1–20.9 kJ mol−1 and 80–200 kJ mol−1, respectively (Li et al. 2019). Therefore, the adsorption of Hg(II) by Fe3O4@PPy has both physical adsorption and chemisorption.

Mechanism speculation

For the Hg(II) adsorption onto the Fe3O4@PPy, the pH is an important factor affecting adsorption by changing the species of mercury ions in the aqueous solution and Zeta potentials of Fe3O4@PPy. Figure 11 shows Zeta potential values and isoelectric point (IP) of Fe3O4@PPy. As can be seen from Figure 11, the IP value of Fe3O4@PPy is about 3.75. Fe3O4@PPy owned more negative charges and negative charges would be carried on their surface in our experimental pH 4–12, showing more potential to adsorb Hg(II) by electrostatic attraction.

Figure 11

Zeta potentials of Fe3O4@PPy under different pHs.

Figure 11

Zeta potentials of Fe3O4@PPy under different pHs.

At pH <4, the main specie of Hg(II) in the solution is Hg2+, which would compete with H+ for adsorption sites (amino groups), as shown in Equations (14) and (15) (Yuan et al. 2014; Zhang et al. 2020). In addition, the protonation of the amino group will make a higher electrostatic repulsion between Hg(II) and Fe3O4@PPy (Deb et al. 2017). However, the part of Hg(II) can form coordinate bonds with Fe3O4@PPy by sharing a lone electron pair of the nitrogen atom of the amino group. Therefore, complexation is the dominating mechanism at pH <4.
formula
(14)
formula
(15)
In the range of pH 4–7, Hg(OH)+ and Hg(OH)2 are the dominant species. The amino groups of the pyrrole can combine with OH in the solution to form negatively charged NH2OH (Kim & Park 2017). Positively charged Hg(OH)+ can be attracted to Fe3O4@PPy with negatively charged groups. On the other hand, the complexation reaction gradually increases due to the weakening of protonation (Alomar et al. 2017; Ghasemi et al. 2019). As the pH increases, the Hg(OH)2 will be formed in the solution. Because Hg(OH)2 is not stable, it quickly decomposes into less soluble HgO precipitation. Therefore, the adsorption mechanism has changed to the coexistence of complexation, electrostatic attraction, and precipitation removal. The reaction mechanisms are formulated as follows (Alomar et al. 2017; Ghasemi et al. 2019; Zhang et al. 2020):
formula
(16)
formula
(17)
formula
(18)
formula
(19)
However, in the range of pH higher than 7, the insoluble complexes of and would be formed due to excess OH, resulting in the transformation of removal mechanism from adsorption to precipitation (Ghasemi et al. 2019). The reactions are summarized in the following equations (Ghasemi et al. 2019):
formula
(20)
formula
(21)

Desorption experiment

In order to recover mercury, avoid secondary pollution and study the reusability of Fe3O4@PPy, the desorption experiment was studied. The previous studies have shown that the addition of HCl will protonate Fe3O4@PPy (Kim & Park 2017) and makes the adsorption slower (Ghasemi et al. 2019; Mohammadnia et al. 2019). In addition, NaCl in solution will also inhibits Hg(II) adsorption because the adsorbed Hg(II) can react with Cl to form complexes soluble in water and the NaCl molecule can occupy the part of adsorption sites of the adsorbent (Wang et al. 2010). Therefore, the desorption investigation was carried out by using two-component desorbent composed of different concentrations of HCl and NaCl (0.05 M). These experimental results are shown in Figure 12.

Figure 12

Recovery efficiency of Hg(II) with different desorption agent concentrations: (a) 0.01 M HCl + 0.05 M NaCl; (b) 0.04 M HCl + 0.05 M NaCl; (c) 0.08 M HCl + 0.05 M NaCl; and (d) 0.1 M HCl + 0.05 M NaCl.

Figure 12

Recovery efficiency of Hg(II) with different desorption agent concentrations: (a) 0.01 M HCl + 0.05 M NaCl; (b) 0.04 M HCl + 0.05 M NaCl; (c) 0.08 M HCl + 0.05 M NaCl; and (d) 0.1 M HCl + 0.05 M NaCl.

As shown in Figure 12, the recovery rate of Hg(II) increased with the increasing concentration of HCl. The recovery rate of Hg(II) was only 31.89% with 0.01 M HCl and 0.05 M NaCl. Furthermore, the recovery rate of 93.03% can be achieved with the HCl and NaCl concentration of 0.1 and 0.05 M, respectively. The experimental results showed that the regeneration by HCl and NaCl is available, and the Fe3O4@PPy appeared the excellent stability and outstanding recycle possibility. The mercury desorbed in this study can be treated centrally to avoid secondary pollution of the water environment, such as making dry batteries, etc.

Comparison with other adsorbents for Hg(II) adsorption

The comparison of Hg(II) adsorption performance between Fe3O4@PPy and other adsorbents are listed in Table 4. It can be seen that Fe3O4@PPy has an excellent adsorption ability for Hg(II), indicating that Fe3O4@PPy has great potential in wastewater treatment applications, especially for Hg(II) removal.

Table 4

The comparison of Hg(II) adsorption performance between Fe3O4@PPy and other adsorbents

AdsorbentExperimental parameters
qmax (mg/g)References
Adsorbent dosage (mg/L)Initial concentration (mg/L)pHEquilibrium time (min)
GO/Fe3O4–Si–Pr–SH 200 20 60 129.7 Mohammadnia et al. (2019)  
SWCNTs/Fe3O4@PDA 200 20 60 249.07 Ghasemi et al. (2019)  
SWCNT-SH 250 30 60 131 Bandaru et al. (2013)  
Carboxylate functionalized bentonite 200 25 5.5 300 113 Anirudhan et al. (2012)  
Porous carbon 80 180 9.8 Li et al. (2019)  
Fe3O4@PPy 250 10 200 232.55 This work 
AdsorbentExperimental parameters
qmax (mg/g)References
Adsorbent dosage (mg/L)Initial concentration (mg/L)pHEquilibrium time (min)
GO/Fe3O4–Si–Pr–SH 200 20 60 129.7 Mohammadnia et al. (2019)  
SWCNTs/Fe3O4@PDA 200 20 60 249.07 Ghasemi et al. (2019)  
SWCNT-SH 250 30 60 131 Bandaru et al. (2013)  
Carboxylate functionalized bentonite 200 25 5.5 300 113 Anirudhan et al. (2012)  
Porous carbon 80 180 9.8 Li et al. (2019)  
Fe3O4@PPy 250 10 200 232.55 This work 

CONCLUSIONS

The magnetic Fe3O4@PPy composite microsphere exhibited great potential on Hg(II) ion removal from solution. The adsorption data was well fitted in the Langmuir and pseudo-second-order kinetic models, and the calculated maximum adsorption capacity was 232.56 mg/g. Thermodynamics studies show that the adsorption of Hg(II) on Fe3O4@PPy is an endothermic and spontaneous process, and the adsorption of Hg(II) by Fe3O4@PPy has both physical adsorption and chemisorption. The Hg(II) ion loaded Fe3O4@PPy could be regenerated by two-component desorbent composed of 0.1 M HCl and 0.05 M NaCl and the maximum recovery of 93.0% was obtained which suggested that the Fe3O4@PPy had good potential to capture and recover Hg(II) ion from aqueous solution.

ACKNOWLEDGEMENTS

The study was supported by the National Natural Science Foundation of China (52074176, 51774200, and 51904174); the Natural Science Foundation of Shandong Province (ZR2020ME106); 2019 Science and Technology Plan of Qingdao West Coast New District (2019-48); ‘Qun xing’ programs of SDUST (QX2018M43); Graduate Tutor Guidance Ability Improvement Program of Shandong Province (SDYY18080)/SDUST; Shandong Province Key Research and Development Project (2019GGX103035); and Young Science and Technology Innovation Program of Shandong Province (2020KJD001).

DATA AVAILABILITY STATEMENT

All relevant data are included in the paper or its Supplementary Information.

REFERENCES

REFERENCES
Alomar
M. K.
,
Alsaadi
M. A.
,
Jassam
T. M.
,
Akib
S.
&
Hashim
M. A.
2017
Novel deep eutectic solvent-functionalized carbon nanotubes adsorbent for mercury removal from water
.
Journal of Colloid and Interface Science
497
,
413
421
.
https://doi.org/10.1016/j.jcis.2017.03.014
.
Anbia
M.
,
Kargosha
K.
&
Khoshbooei
S.
2015
Heavy metal ions removal from aqueous media by modified magnetic mesoporous silica MCM-48
.
Chemical Engineering Research and Design
93
,
779
788
.
https://doi.org/10.1016/j.cherd.2014.07.018
.
Anirudhan
T. S.
,
Jalajamony
S.
&
Sreekumari
S. S.
2012
Adsorption of heavy metal ions from aqueous solutions by amine and carboxylate functionalised bentonites
.
Applied Clay Science
65–66
,
67
71
.
https://doi.org/10.1016/j.clay.2012.06.005
.
Antochshuk
V.
,
Olkhovyk
O.
,
Jaroniec
M.
,
Park
I.
&
Ryoo
R.
2003
Benzoylthiourea-modified mesoporous silica for mercury(II) removal
.
Langmuir
19
(
7
),
3031
3034
.
https://doi.org/10.1021/la026739z
.
Bandaru
N. M.
,
Reta
N.
,
Dalal
H.
,
Ellis
A. V.
,
Shapter
J.
&
Voelcker
N. H.
2013
Enhanced adsorption of mercury ions on thiol derivatized single wall carbon nanotubes
.
Journal of Hazardous Materials
261
,
534
541
.
https://doi.org/10.1016/j.jhazmat.2013.07.076
.
Bhaumik
M.
,
Maity
A.
,
Srinivasu
V. V.
&
Onyangoa
M. S.
2011
Enhanced removal of Cr(VI) from aqueous solution using polypyrrole/Fe3O4 magnetic nanocomposite
.
Journal of Hazardous Materials
190
(
1–3
),
381
390
.
https://doi.org/10.1016/j.jhazmat.2011.03.062
.
Cao
Y. Y.
,
Zhou
G. Z.
,
Zhou
R. S.
,
Wang
C. Z.
,
Chi
B. R.
,
Wang
Y.
,
Hua
C. Y.
,
Qiu
J.
,
Jin
Y. Q.
&
Wu
S.
2020
Green synthesis of reusable multifunctional γ-Fe2O3/bentonite modified by doped TiO2 hollow spherical nanocomposite for removal of BPA
.
Science of The Total Environment
708
,
134669
.
https://doi.org/10.1016/j.scitotenv.2019.134669
.
Deb
A. S.
,
Dwivedi
V.
,
Dasgupta
K.
,
Ali
S. M.
&
Shenoy
K. T.
2017
Novel amidoamine functionalized multi-walled carbon nanotubes for removal of mercury(II) ions from wastewater: combined experimental and density functional theoretical approach
.
Chemical Engineering Journal
313
,
899
911
.
https://doi.org/10.1016/j.cej.2016.10.126
.
Ghasemi
S. S.
,
Hadavifar
M.
,
Maleki
B.
&
Mohammadnia
E.
2019
Adsorption of mercury ions from synthetic aqueous solution using polydopamine decorated SWCNTs
.
Journal of Water Process Engineering
32
,
100965
.
https://doi.org/10.1016/j.jwpe.2019.100965
.
Guo
S. Z.
,
Zhang
J.
,
Li
X. L.
,
Zhang
F.
&
Zhu
X. X.
2018
Fe3O4-CS-L: a magnetic core-shell nano adsorbent for highly efficient methyl orange adsorption
.
Water Science & Technology
77
(
3
),
628
637
.
https://doi.org/10.2166/wst.2017.602
.
Guo
J. H.
,
Yan
C. Z.
,
Luo
Z. X.
,
Fang
H. D.
,
Hu
S. G.
&
Cao
Y. L.
2019
Synthesis of a novel ternary HA/Fe-Mn oxides-loaded biochar composite and its application in cadmium(II) and arsenic(V) adsorption
.
Journal of Environmental Sciences
85
,
168
176
.
https://doi.org/10.1016/j.jes.2019.06.004
.
Huang
Y. M.
,
Li
G.
,
Li
M. Z.
,
Yin
J. J.
,
Meng
N.
,
Zhang
D.
,
Cao
X. Q.
,
Zhu
F. P.
,
Chen
M.
,
Li
L.
&
Lyu
X. J.
2021
Kelp-derived N-doped biochar activated peroxymonosulfate for ofloxacin degradation
.
Science of The Total Environment
754
,
141999
.
https://doi.org/10.1016/j.scitotenv.2020.141999
.
Kim
K. J.
&
Park
J. W.
2017
Stability and reusability of amine-functionalized magnetic-cored dendrimer for heavy metal adsorption
.
Journal of Materials Science
52
(
2
),
843
857
.
https://doi.org/10.1007/s10853-016-0380-z
.
Li
Y.
,
Xia
M. D.
,
An
F. F.
,
Ma
N. F.
,
Jiang
X. L.
,
Zhu
S. M.
,
Wang
D. W.
&
Ma
J.
2019
Superior removal of Hg (II) ions from wastewater using hierarchically porous, functionalized carbon
.
Journal of Hazardous Materials
371
,
33
41
.
https://doi.org/10.1016/j.jhazmat.2019.02.099
.
Liu
T. T.
,
Jing
L.
,
Liu
Q. Y.
&
Zhang
X. M.
2018
Facile one-pot synthesis of a porphyrin-based hydrophilic porous organic polymer and application as recyclable absorbent for selective separation of methylene blue
.
Chemosphere
212
,
1038
1046
.
https://doi.org/10.1016/j.chemosphere.2018.08.122
.
Liu
S. R.
,
Chen
M.
,
Cao
X. Q.
,
Li
G.
,
Zhang
D.
,
Li
M. Z.
,
Meng
N.
,
Yin
J. J.
&
Yan
B. Q.
2020
Chromium (VI) removal from water using cetylpyridinium chloride (CPC)-modified montmorillonite
.
Separation and Purification Technology
241
,
116732
.
https://doi.org/10.1016/j.seppur.2020.116732
.
Ma
L. L.
,
Chen
N.
,
Feng
C. P.
,
Li
M.
,
Gao
Y.
&
Hu
Y. T.
2020
Coupling enhancement of Chromium (VI) bioreduction in groundwater by phosphorus minerals
.
Chemosphere
240
,
124896
.
https://doi.org/10.1016/j.chemosphere.2019.124896
.
Mohammadnia
E.
,
Hadavifar
M.
&
Veisi
H.
2019
Kinetics and thermodynamics of mercury adsorption onto thiolated graphene oxide nanoparticles
.
Polyhedron
173
,
114139
.
https://doi.org/10.1016/j.poly.2019.114139
.
Mollahosseini
A.
,
Khadir
A.
&
Saeidian
J.
2019
Core–shell polypyrrole/Fe3O4 nanocomposite as sorbent for magnetic dispersive solid-phase extraction of Al3+ ions from solutions: investigation of the operational parameters
.
Journal of Water Process Engineering
29
,
100795
.
https://doi.org/10.1016/j.jwpe.2019.100795
.
Morel
A. L.
,
Nikitenko
S. I.
,
Gionnet
K.
,
Wattiaux
A.
,
Lai-Kee-Him
J.
,
Labrugere
C.
,
Chevalier
B.
,
Deleris
G.
,
Petibois
C.
,
Brisson
A.
&
Simonoff
M.
2008
Sonochemical approach to the synthesis of Fe3O4@SiO2 core − shell nanoparticles with tunable properties
.
ACS Nano
2
(
5
),
847
856
.
https://doi.org/10.1021/nn800091q
.
Peng
X. Q.
,
Zhang
W.
,
Gai
L. G.
,
Jiang
H. H.
,
Wang
Y.
&
Zhao
L. C.
2015
Dedoped Fe3O4/PPy nanocomposite with high anti-interfering ability for effective separation of Ag(I) from mixed metal-ion solution
.
Chemical Engineering Journal
280
,
197
205
.
https://doi.org/10.1016/j.cej.2015.05.118
.
Sampaio
C. G.
,
Frota
L. S.
,
Magalhães
H. S.
,
Dutra
L. M. U.
,
Queiroz
D. C.
,
Araújo
R. S.
,
Becker
H.
,
Souza
J. R. R.
,
Ricardo
N. M. P. S.
&
Trevisan
M. T. S.
2015
Chitosan/mangiferin particles for Cr(VI) reduction and removal
.
International Journal of Biological Macromolecules
78
,
273
279
.
https://doi.org/10.1016/j.ijbiomac.2015.03.038
.
Sharma
A.
,
Sharma
A.
&
Arya
R. K.
2015
Removal of mercury(II) from aqueous solution: a review of recent work
.
Separation Science and Technology
50
(
9
),
1310
1320
.
https://doi.org/10.1080/01496395.2014.968261
.
Tang
S.
,
Lan
Q.
,
Liang
J. Y.
,
Chen
S. R.
,
Liu
C.
,
Zhao
J. X.
,
Cheng
Q.
,
Cao
Y. C.
&
Liu
J. Y.
2017
Facile synthesis of Fe3O4/PPy core-shell magnetic nanoparticles and their enhanced dispersity and acid stability
.
Materials & Design
121
,
47
50
.
https://doi.org/10.1016/j.matdes.2017.02.049
.
Wang
Y.
,
Tan
H.
,
Yang
H.
,
Shu
H.
,
Nie
J.
,
Wang
T.
&
He
J.
2010
Effects of NaCl on Hg adsorption in soil and its kinetics
.
Journal of Anhui Agricultural Sciences
38
(
25
),
13750
(in Chinese)
.
Wang
Y. Q.
,
Zou
B. F.
,
Gao
T.
,
Wu
X. P.
,
Lou
S. Y.
&
Zhou
S. M.
2012
Synthesis of orange-like Fe3O4/PPy composite microspheres and their excellent Cr(VI) ion removal properties
.
Journal of Materials Chemistry
22
(
18
),
9034
.
https://doi.org/10.1039/C2JM30440F
.
Wang
J.
,
Wang
Q.
,
Gao
X. L.
,
Tian
X. X.
,
Wei
Y. Y.
,
Cao
Z.
,
Guo
C. G.
,
Zhang
H. F.
,
Ma
Z.
&
Zhang
Y. S.
2020
Surface modification of mesoporous silica nanoparticle with 4-triethoxysilylaniline to enhance seawater desalination properties of thin-film nanocomposite reverse osmosis membranes
.
Frontiers of Environmental Science & Engineering
14
(
1
),
95
104
.
https://doi.org/10.1007/s11783-019-1185-5
.
Xiao
X. F.
,
Deng
Y. Y.
,
Xue
J. L.
&
Gao
Y.
2019
Adsorption of chromium by functionalized metal organic frameworks from aqueous solution
.
Environmental Technology
1
13
.
https://doi.org/10.1080/09593330.2019.1683618
.
Xiao
F.
,
Yan
B. Q.
,
Zou
X. Y.
,
Cao
X. Q.
,
Dong
L.
,
Lyu
X. J.
,
Li
L.
,
Qiu
J.
,
Chen
P.
,
Hu
S. G.
&
Zhang
Q. J.
2020
Study on ionic liquid modified montmorillonite and molecular dynamics simulation
.
Colloids and Surfaces A: Physicochemical and Engineering Aspects
587
,
124311
.
https://doi.org/10.1016/j.colsurfa.2019.124311
.
Yuan
Q.
,
Chi
Y.
,
Yu
N.
,
Zhao
Y.
,
Yan
W. F.
,
Li
X. T.
&
Dong
B.
2014
Amino-functionalized magnetic mesoporous microspheres with good adsorption properties
.
Materials Research Bulletin
49
,
279
284
.
https://doi.org/10.1016/j.materresbull.2013.08.063
.
Zabihi
M.
,
Haghighi Asl
A.
&
Ahmadpour
A.
2010
Studies on adsorption of mercury from aqueous solution on activated carbons prepared from walnut shell
.
Journal of Hazardous Materials
174
(
1–3
),
251
256
.
https://doi.org/10.1016/j.jhazmat.2009.09.044
.
Zhang
C.
,
Sui
J. H.
,
Li
J.
,
Tang
Y. L.
&
Cai
W.
2012
Efficient removal of heavy metal ions by thiol-functionalized superparamagnetic carbon nanotubes
.
Chemical Engineering Journal
210
,
45
52
.
https://doi.org/10.1016/j.cej.2012.08.062
.
Zhang
W.
,
Zhu
W. Y.
,
Xu
W. T.
,
Wang
Y.
,
Li
N.
,
Zhang
T. T.
&
Wang
H.
2017
Fe3O4@Polypyrrole microspheres with high magnetization and superparamagnetism for efficient and fast removal of Pb (II) ions
.
Russian Journal of Physical Chemistry A
91
(
13
),
2657
2665
.
https://doi.org/10.1134/S0036024417130258
.
Zhang
H. H.
,
Cao
X. Y.
,
Wang
H.
,
Ma
Z.
,
Li
J.
,
Zhou
L. M.
&
Yang
G. P.
2019
Effect of black carbon on sorption and desorption of phosphorus onto sediments
.
Marine Pollution Bulletin
146
,
435
441
.
https://doi.org/10.1016/j.marpolbul.2019.06.059
.
Zhang
Z. Z.
,
Xia
K.
,
Pan
Z. W.
,
Yang
C. X.
,
Wang
X.
,
Zhang
G. W.
,
Guo
Y. F.
&
Bai
R. B.
2020
Removal of mercury by magnetic nanomaterial with bifunctional groups and core-shell structure: synthesis, characterization and optimization of adsorption parameters
.
Applied Surface Science
500
,
143970
.
https://doi.org/10.1016/j.apsusc.2019.143970
.
Zhao
G. G.
,
Ye
S. Y.
,
Yuan
H. G.
,
Ding
X. G.
,
Wang
J.
&
Laws
E. A.
2018
Surface sediment properties and heavy metal contamination assessment in river sediments of the Pearl River Delta, China
.
Marine Pollution Bulletin
136
,
300
308
.
https://doi.org/10.1016/j.marpolbul.2018.09.035
.
Zhou
G. Z.
,
Wang
Y.
,
Zhou
R. S.
,
Wang
C. Z.
,
Jin
Y. Q.
,
Qiu
J.
,
Hua
C. Y.
&
Cao
Y. Y.
2019
Synthesis of amino-functionalized bentonite/CoFe2O4@MnO2 magnetic recoverable nanoparticles for aqueous Cd2+ removal
.
Science of The Total Environment
682
,
505
513
.
https://doi.org/10.1016/j.scitotenv.2019.05.218
.
Zou
X. Y.
,
Xiao
F.
,
Liu
S. R.
,
Cao
X. Q.
,
Li
L.
,
Chen
M.
,
Dong
L.
,
Lyu
X. J.
&
Gai
Y. J.
2020
Preparation and application of CPC/Keggin-al30 modified montmorillonite composite for Cr (VI) removal
.
Journal of Water Process Engineering
37
,
101348
.
https://doi.org/10.1016/j.jwpe.2020.101348
.
This is an Open Access article distributed under the terms of the Creative Commons Attribution Licence (CC BY-NC-ND 4.0), which permits copying and redistribution for non-commercial purposes with no derivatives, provided the original work is properly cited (http://creativecommons.org/licenses/by-nc-nd/4.0/).