Adsorption of anionic dye onto magnetic Fe 3 O 4 /CeO 2 nanocomposite: Equilibrium,kinetics,and thermodynamics

Abstract

In this study, a magnetically separable Fe₃O₄/CeO₂ (Fe/Ce) nanocomposite is synthesized by sol-precipitation method and characterized by field-emission scanning electron microscopy, transmission electron microscopy, X-ray diffraction, energy dispersive spectrometer , vibrating sample magnetometer, atomic absorption spectrometer, and zeta potential measurements. The Fe/Ce is used as sorbent to adsorb anionic dye of Acid Black 210 (AB210) from aqueous solutions, and the maximum adsorption capacity is about 90.50 mg/g, which is six times higher than that of the commercial CeO₂. Dependence of absorption performance on essential factors, such as initial dye concentration, temperature and initial pH, are experimentally examined. The result shows that the adsorption kinetic of Fe/Ce follows pseudo-second-order model and the adsorption isotherm is well described by the Langmuir adsorption model. Furthermore, the thermodynamic analysis indicates that the adsorption of Fe/Ce for AB210 is spontaneous and endothermic.

Keywords

Magnetic anionic dye kinetics thermodynamics

Introduction

In recent years, organic dye contamination has attracted much attention due to their large amount of emissions and increasing damage to natural environment (Hao et al., 2014; Song et al., 2015). Even very low concentration of dyes in water could reduce the light penetration through the water surface, precluding photosynthesis of the aqueous flora (Aguayo-Villarreal et al., 2013). Additionally, some dyes may be highly toxic, potentially carcinogenic, allergenic, mutagenic, and teratogenic on human and aquatic life (Peng et al., 2016; Zhang et al., 2014). These molecules or their metabolites (i.e. aromatic amines) are chemically stable and poorly biodegraded in natural conditions (Ratna, 2012; Zahrim et al., 2011). Therefore, it is highly desirable to develop an economical solution to rapidly and efficiently dispose these dyes from aqueous solutions.

To date, various methods are available for the removal of these dyes from wastewater including adsorption, ultrafiltration, chemical oxidation, ion exchange, photocatalysis, electrodialysis, catalytic degradation and coagulation/flocculation (Boumaza et al., 2014; Fiorentin et al., 2010; Fu and Wang, 2011; Kasgöz and Durmus, 2008; Mane and Babu, 2011). Among these methods, adsorption is verified to be one of the most promising alternative techniques to remove the nonbiodegradable dyes from wastewater due to its simple operation, superior adsorption effectiveness, relatively low cost, and almost harmless byproducts (Maiti et al., 2017; Xiao et al., 2016). So, it is highly approving to design and fabricate superior adsorbents using an economic and facile approach. Until now, various adsorbents such as activated carbon (Tan et al., 2008), fly ash (Dizge et al., 2008), bentonite clay (Tahir and Rauf, 2006), mesoporous silica (Huang et al., 2011), zeolite (Meshko et al., 2001), carbon nanotubes (Machado et al., 2016), graphene oxide (Konicki et al., 2017), and magnetic nanomaterials (Li et al., 2017) have been adopted to remove dyes from aqueous solutions.

One of the promising and efficient adsorbents are magnetic nanocomposites because they can be easily separated from aqueous solutions by applying an external magnetic field (Xu et al., 2010). Therefore, many magnetic nanocomposites based on Fe₃O₄ have been tested for the possibility to reduce dye concentration in aqueous solutions, owing to Fe₃O₄ having strong ferromagnetic resonance property and easy recyclability (Kim and Song, 2014). Pan et al. (2016) have synthesized Fe₃O₄@HCP core–shell porous magnetic microspheres for Methyl Orange adsorption and fast separation from aqueous solutions. Shan et al. (2014) have studied the adsorption of dyes such as Reactive Red, Congo Red, and Acid Red 1 from aqueous solutions by magnetic Fe₃O₄/MgAl-layered double hydroxide composite. Yao et al. (2012) have prepared magnetic Fe₃O₄@graphene nanocomposite for the removal of methylene blue and Congo Red from water. Chong et al. (2016) have studied the removal of Methyl Orange, Methylene Blue, and Crystal Violet from aqueous solutions by magnetic Fe(0)/Fe₃O₄/graphene composite.

However, some magnetic nanocomposites are difficult to be reused because of their unstable performance and are easily affected by environmental conditions, which limit their actual application (Liu et al., 2008; Xu et al., 2010). Considering the factors mentioned above, CeO₂ can be one of the efficient and easily reusable adsorbent due to its outstanding thermal stability (Ouyang et al., 2013; Zhang et al., 2016). Because Fe₃O₄ is easily aggregated leads to the decrease of surface area and affected by the environmental conditions, coated-CeO₂ could improve the dispersion and stability of Fe₃O_4. (Chong et al., 2016; Yao et al., 2012). Additionally, CeO₂ can be used not only as a adsorbent but also as a support in the Fe/Ce composite system to protect Fe₃O₄ against acid and alkaline environments due to the presence of the CeO₂ in the outside (Xu and Wang, 2012). Therefore, the CeO₂-coated Fe₃O₄ is a promising alternative adsorbent. It could be easily separated from the aqueous solutions, which is beneficial for multiple reuse, simplified post-processing steps, reduced costs, etc. (Xing et al., 2011; Yu et al., 2012). To date, only a few studies have been conducted on the adsorption of dyes by Fe₃O₄/CeO₂ composites (Li et al., 2017). However, the scale productions of these adsorbents are also limited because of the tedious process of preparation, time-consumption, and costly production.

In the present work, we used the Fe/Ce as an adsorbent to remove anionic dye Acid Black 210 (AB210) from aqueous solutions. The effects of main parameters like initial dye concentration, temperature, and initial pH solution for the adsorption capacity of Fe/Ce have been investigated. Additionally, the kinetic data are analyzed by applying the pseudo-first-order, pseudo-second-order, and intra-particle diffusion kinetic models. The experimental equilibrium adsorption data are fitted to the Langmuir, Freundlich, and Temkin isotherm adsorption models. Thermodynamic parameters, such as ΔG°, ΔH°, and ΔS°, are also calculated.

Experiment

Materials

Materials were purchased from Shanghai Titan Scientific Co., Ltd (China): HNO₃ (AR, 65–68%), C₂H₅OH (AR), NaOH (AR), Fe₃O₄ (AR), Ce(NO₃)₃•6H₂O (AR), HCl (AR, 36–38%), NH₃•H₂O (AR, 25–28%), commercial CeO₂ (AR). AB210 was supplied by Alibaba.com. Deionized water (DI) supplied by our university was used in all the experiments. All chemicals used were of analytical grade and without further purification.

Preparation of Fe/Ce

Fe/Ce was prepared via sol-precipitation method (Gao et al., 2018). The specific experimental process is as follows. Firstly, 0.0020 mol Ce(NO₃)₃•6H₂O and 20.0 mL of deionized water (DT) were added to a three-necked flask and stirred evenly at 65 °C. Then, 20.0 mL of ammonia solution (0.40 mL NH₃•H₂O) was added dropwise to the reaction solution to form Ce(OH)₃ sol with the injection pump (20.0 mL/h). Subsequently, pretreated Fe₃O₄ (0.0007 mol) was added to the reaction and 20.0 mL ammonia solution (0.40 mL NH₃•H₂O) was also added dropwise into the above reaction solution. Thereafter, the reaction was stirred for 5 h at 65 °C in order to prompt Ce(OH)₃ sol formed to completely precipitate on the surface of Fe₃O₄. Then heating was stopped and cooled to room temperature naturally. Finally, the resulting sediment was collected with a magnet and washed alternately with water and ethanol several times, and the Fe/Ce was obtained after drying at 70 °C overnight. The preparation conditions of CeO₂ are the same as that of Fe/Ce except for the addition of Fe₃O₄.

Adsorption experiments

The as-obtained Fe/Ce was employed as an adsorbent for the removal of AB210 from the mimic waste water. Adsorption experiments were carried out in a flask, where the dye solution (50.0 mL) with initial dye concentration was placed. Initial concentrations of dye were varied from 20 mg/L to 80 mg/L for AB210. The flask with dye solution was sealed and placed in a temperature-controlled shaking water bath and agitated at a constant speed of 180 r/min. To observe the effect of temperature, the experiments were carried out at five different temperatures, i.e. 20 °C, 30 °C, 40 °C, 50 °C, and 60 °C. Before mixing with the adsorbent, the pH levels of various dye solutions were adjusted by adding hydrochloric acid (0.1 M HCl) or sodium hydroxide (0.1 M NaOH). When the desired temperature was reached, about 50.0 mg of Fe/Ce was added into the flask. At the end of the equilibrium period, aqueous sample was taken out from the flask with a syringe and the liquid was separated by the external magnetic field. The dye concentration was calculated by UV–visible spectrophotometer (λ = 461 nm).

The amount of dye adsorbed at equilibrium q_e (mg/g) was calculated applying the following equation (Radia et al., 2018)

q_{e} = \frac{(C_{o} - C_{e}) V}{m}

(1)

where C_o (mg/L) is the initial dye concentration, C_e (mg/L) is the dye concentration at equilibrium, V (L) is the volume of the solution, and m (g) is the mass of the adsorbent.

The procedures of kinetic experiments were identical with those of the equilibrium tests. At predetermined moment, aqueous sample was taken from the solution, and a UV–visible spectrophotometer was used to determine the liquid, which was separated from the adsorbent and dye. The adsorption capacity of Fe/Ce at time (t) q_t (mg/g) was calculated by the following equation

q_{t} = \frac{(C_{o} - C_{t}) V}{m}

(2)

where C_t (mg/L) is the dye concentration at any time t.

Characterization

The morphology and microstructure of the sample were captured by a field-emission scanning electron microscopy (FE-SEM, JSM 7401F, JEOL, Japan) at 3.0 kV and transmission electron microscopy (TEM, Tecnai F30). The crystal structure and phase of the as-obtained sample were identified by an X-ray diffractometer (XRD, Mini Flex600, Rigaku, Japan). The elementary compositions of the sample were captured by energy dispersive spectrometer (EDS) (EDS:IE 300 X, Oxford, UK). The zeta potentials of Fe/Ce were measured at different pH values on zeta potential and particle size analyzer (90plus PALS, US) at room temperature. The magnetic hysteresis loops were measured by a vibrating-sample magnetometer (VSM, 7407, Quantum Design Company, USA). The released iron concentration of the sample was measured by double-beam flame atomic absorption spectrometer (AAS) (AA240, Agilent, USA). The adsorption experiments were carried out at the electric agitation condition (SHA-C, Sai De Li Si Experimental Analytical Instrument Company, Tianjin, China). The absorption spectra of the initial AB210 solution as well as the filtrate after adsorption were monitored by a UV–visible spectrophotometer (Beijing Purkinje General Instrument production of TU-1900) with a characteristic absorption at the wavelength of 461 nm to quantify the concentration of AB210 dye.

Results and discussion

Characterization of Fe/Ce

Figure 1 shows the high-magnification FE-SEM images (a, b) of Fe₃O₄ and Fe/Ce and TEM images (c, d) of Fe/Ce and CeO₂. The morphology of Fe₃O₄ is irregular and some approximate spherical particles can be found in Figure 1(a). Owing to the strong interaction among magnetic particles, such as the magnetic dipolar interaction and van der Waals force, an obvious agglomeration and sticking together of irregular Fe₃O₄ nanoparticles also can be observed (Wang et al., 2015). Compared with Fe₃O₄, the morphology and size of Fe/Ce have no obvious change after coating of CeO₂ (Figure 1(b)), which can be ascribed to the poor uniform size of Fe₃O₄ nanoparticles and uneven CeO₂ coated on the surface of Fe₃O₄ (Figure 1(c)). Besides, the Fe/Ce particles still remain aggregate, which reveals that the sample still maintains the characteristic of magnetism. In Figure 1(c), obvious delamination of Fe/Ce particles can be found, where the black area in the middle is Fe₃O₄, and the light-colored area on the edge is the coated CeO₂. Moreover, the particle size of the coated CeO₂ is substantially the same as that of the CeO₂ particles (prepared under the same conditions) (Figure 1(d)). It further confirms that CeO₂ is successfully loaded onto Fe₃O₄ surface. Figure 1(e) and (f) shows the EDS element mapping images of Fe₃O₄ and Fe/Ce. In Figure 1(e), pure Fe₃O₄ contains iron and oxygen and it can be concluded from the table inside the figure that the atomic ratio of iron and oxygen is in accordance with that of Fe₃O₄. In Figure 1(f), Fe/Ce contains cerium apart from iron and oxygen, and the positions of iron and oxygen are basically the same as that of Fe₃O₄ (Figure 1(e)). According to the table inside the figure, the molar ratio of CeO₂ to Fe₃O₄ can be roughly calculated to be about 3:1. This indicates that Fe/Ce material is successfully prepared.

Figure 1.

FE-SEM images of (a) Fe₃O₄, (b) Fe/Ce. TEM images of (c) Fe/Ce, (d) CeO₂. EDS elemental patterns of (e) Fe₃O₄, and (f) Fe/Ce.

Figure 2 shows the XRD pattern of the pure CeO₂, single Fe₃O₄, and Fe/Ce. For single Fe₃O₄, the stronger diffraction peaks and the corresponding crystal plane of the single Fe₃O₄ at 30.10°(220), 35.42°(311), 56.94°(511), 62.52°(440) are assigned to Fe₃O₄ (JCPDS No. 19–0629). The stronger diffraction peaks and the corresponding crystal plane of pure CeO₂ at 28.55°(111), 33.08°(200), 47.48°(220), 56.33°(311) match well with those of CeO₂ (JCPDS No. 34–0394). As a contrast, the XRD pattern of Fe/Ce contains obvious diffraction peaks of pure CeO₂ and single Fe₃O₄. This indicates that these substances of Fe₃O₄ and CeO₂ are included in the as-obtained material. Moreover, it is worth noting that the decrease in the intensity of diffraction peaks corresponds to Fe₃O₄ and CeO₂, respectively, which reveals reduced crystallinity of as-prepared Fe/Ce due to the interaction between CeO₂ and Fe₃O₄ and agglomeration of particles. In addition, no other peaks could be detected, which indicates that there is no impurity in Fe/Ce. The results obtained are in good agreement with that of EDS element mapping images and TEM images, which further confirms that the as-prepared sample is the CeO₂-coated Fe₃O₄.

Figure 2.

The XRD of CeO₂, Fe₃O₄, and Fe/Ce.

Magnetic properties of bare Fe₃O₄ and Fe/Ce are investigated via VSM at room temperature. Figure 3 shows that Fe/Ce has near-zero remanence and coercivity at room temperature, which reveals a super-paramagnetic characteristic (Wu et al., 2016). The saturation magnetization values of bare Fe₃O₄ and Fe/Ce are 73.58 emu/g and 28.06 emu/g, respectively. Compared with bare Fe₃O₄, the decreased 45.52 emu/g of saturation magnetization value of Fe/Ce is attributed to the existence of loaded CeO₂, which weakens the magnetic property of bare Fe₃O₄ (Xu and Wang, 2012). Simultaneously, it illustrates the successful synthesis of Fe/Ce. So, the result of the VSM is consistent with the TEM, EDS, and XRD data. Moreover, the as-prepared Fe/Ce still shows excellent magnetic sensitivity, as shown in illustration 2, and the dye-loaded adsorbent can be rapidly separated from the mixture (illustration 1) within 10 s with the help of the magnet. This excellent magnetic response behavior of Fe/Ce contributes to a rapid, easy, and high throughput way to enrich and separate adsorbents (Xu et al., 2014).

Figure 3.

Magnetic hysteresis loops of Fe₃O₄ and Fe/Ce. Photographs (inset) (1-AB210 and Fe/Ce mixing, 2-the magnet collects Fe/Ce after adsorption of AB210).

In order to evaluate the potential application in dye wastewater treatment, the Fe/Ce is employed as the adsorbent for the removal of AB210 from mimic waste water. As can be seen in Figure 4, the adsorption capacity (q_t) of CeO₂ and Fe/Ce rapidly increases to 38.64 mg/g and 34.34 mg/g within the first 2 min and continues to increase in the next 10 min to 30 min, respectively, then further increases at a relatively minor rate till 120 min. Nevertheless, q_t of Fe/Ce is 41.18 mg/g in 120 min, which is 2.89 mg/g lower than that of CeO₂. Since the CeO₂ particles is coated on the surface of the Fe₃O₄, its specific surface area is reduced, which leads to the decrease in the adsorption performance. Simultaneously, we can clearly see that q_t of commercial Fe₃O₄ is only 12.18 mg/g in 120 min, which is 29.00 mg/g lower than that of Fe/Ce. This demonstrates that Fe/Ce still maintains good adsorption property in addition to the magnetic property. Additionally, the Fe/Ce (illustration 3) and CeO₂ (illustration 4) intuitively show that the dye solution is almost colorless after adsorption 120 min, respectively. However, the Fe₃O₄ (illustration 2) is not significantly different from the color of the initial concentration of the dye (illustration 1) within 120 min. This also indicates that the Fe/Ce still has a better adsorption performance.

Figure 4.

Instantaneous adsorption capacity of different materials within 120 min (the amount of all the materials are the same: 50.0 mg 50.0 mg/L). Photographs (inset) (1-AB210, 2, 3, 4 is the corresponding colors of the commercial Fe₃O₄, Fe/Ce, and CeO₂ in 120 min, respectively).

Effect of initial dye concentration on adsorption kinetics

To determine the effect of the initial concentration of AB210 on the adsorption capacity of Fe/Ce, the adsorption experiments are performed in several initial concentrations ranging from 20 mg/L to 80 mg/L at 30 °C and pH = 7.0. As seen in Figure 5, the adsorption capacity of Fe/Ce increases with increasing AB210 initial concentration. And the adsorption capacity growth trends of different initial concentrations are basically the same. This indicates that the initial AB210 concentration plays a vital role in the adsorption capacity of Fe/Ce because the initial dye concentration provides a momentous driving force to overcome the mass transfer resistance of the dye between solid and aqueous phases. Hence, the adsorption capacity increases with increasing initial dye concentration.

Figure 5.

Effect of initial dye concentration on adsorption capacity of the AB210 onto Fe/Ce. Experimental conditions: T = 30 °C, pH = 7.0.

In this study, the experimental data for the adsorption of AB210 onto Fe/Ce are tested with three different adsorption kinetic models: pseudo-first-order model, pseudo-second-order kinetic model, and intra-particle diffusion model. The pseudo-first-order kinetic model presumes that the ratio of occupation of adsorption sites is proportional to the number of unoccupied sites. The pseudo-second-order kinetic model assumes that the adsorption mechanism depends on the adsorbent and adsorbate, and the rate-limiting step may be the chemical adsorption involving valence forces through the exchanging or sharing of electrons between adsorbent and adsorbate. The intra-particle diffusion model is used to research whether the intra-particle diffusion is the rate controlling step in the process of adsorption of AB210 dye onto Fe/Ce.

The pseudo-first-order kinetic model is shown as the following (Mellah et al., 2006)

\ln (q_{e} - q_{t}) = \ln (q_{e}) - k_{1} t

(3)

where k₁ (min⁻¹) is the pseudo-first-order rate constant adsorption and t (min) is the time. The values of q_e and k₁ are calculated from the linear plots of ln(q_e − q_t) versus t (Figure 6(a)).

Figure 6.

Adsorption kinetic for adsorption of AB210 onto Fe/Ce. Pseudo-first order (a) and Pseudo-second order (b). Experimental conditions: T = 30 °C, pH = 7.0.

The pseudo-second-order kinetic model can be represented in the following linear form equations (Ho and Mckay, 1999)

\frac{t}{q_{t}} = \frac{1}{k_{2} q_{e}^{2}} + \frac{t}{q_{e}}

(4)

where k₂ [g/(mgmin)] is the pseudo-second-order rate constant adsorption. The values of k₂ and q_e are calculated from the intercept and slope of the linear plots obtained by linearly fitting t/q_t versus t (Figure 6(b)). The kinetic parameters and correlation coefficients (R²) are listed in Table 1. Taking the correlation coefficient into consideration, the experimental data for adsorption of AB210 onto Fe/Ce is best fit with the pseudo-second-order model (R² ≥ 0.9993). Furthermore, the experimental values q_e,exp are basically close to the values q_e,cal obtained in the pseudo-second-order kinetic model except when the dye concentration is 80 mg/L.

Table 1.

Comparison of the pseudo-first-order, pseudo-second-order, and the intra-particle diffusion models for different initial concentrations of AB210 at 30 °C and pH = 7.0.

Dye	C_o	20	40	50	60	80
	(mg/L)	20	40	50	60	80
	q_e,exp	19.43	34.26	46.29	49.87	64.44
	(mg/g)	19.43	34.26	46.29	49.87	64.44
Pseudo-first-order model	k ₁	0.04405	0.02066	0.00684	0.00821	0.00463
	(min⁻¹)	0.04405	0.02066	0.00684	0.00821	0.00463
	q_e,cal	1.44	4.30	10.87	9.55	19.50
	(mg/g)	1.44	4.30	10.87	9.55	19.50
	R ²	0.8873	0.9628	0.9286	0.9322	0.8634
Pseudo-second-ordermodel	k ₂	0.11884	0.02422	0.01342	0.01341	0.01011
	[g/(mgmin)]	0.11884	0.02422	0.01342	0.01341	0.01011
	q_e,cal	19.46	34.04	41.44	46.32	52.97
	(mg/g)	19.46	34.04	41.44	46.32	52.97
	R ²	0.9999	0.9997	0.9995	0.9993	0.9994
Intra-particle diffusionmodel	k_p	0.30402	0.68961	0.96949	1.06479	1.57949
	[mg/(gmin^0.5)]	0.30402	0.68961	0.96949	1.06479	1.57949
	C	17.59	28.58	32.85	37.54	40.45
	(mg/g)	17.59	28.58	32.85	37.54	40.45
	R ²	0.8524	0.9613	0.9885	0.9807	0.9818

The intra-particle diffusion model is expressed by the following equation (Wu et al., 2009)

q_{t} = k_{p} t^{0.5} + C

(5)

where k_p [mg/(g·min^0.5)] is the intra-particle diffusion rate constant and C (mg/g) is the constant connected with the thickness of the boundary layer. The values of k_p and C are calculated from the linear plot of q_t by t^0.5 (Figure 7) and are given in Table 1. As shown in Figure 7, the plots clearly show two different portions, which indicate the different stages of adsorption. The larger slope portion is ascribed to the diffusion of adsorbate through the solution to the external surface of adsorbent. The smaller slope portion displays the gradual adsorption stage, which corresponds to the diffusion of adsorbate molecules inside the pores of adsorbent. If the plot of q_t versus t^0.5 is linear and passes through the origin (C = 0), intra-particle diffusion is the sole rate-limiting step. Whereas when the plots for AB210 do not pass through the origin (C ≠ 0), this deviation of plots from the origin demonstrates that although intra-particle diffusion is involved in the adsorption process, it is not the rate-controlling step. Values of intercept (C) provide a concept of the thickness of the boundary layer; the larger the C, the greater the boundary layer effect. The C values (17.59–40.45 mg/L for AB210) increases with the increase in the initial dye concentrations (20–80 mg/L). At the same time, this also explains the reason for the larger deviation between q_e,exp and q_e,cal when the initial concentration is 80 mg/L. Therefore, these results demonstrate that the boundary layer diffusion effect is promoted with the increase in the initial dye concentration (Konicki et al., 2018).

Figure 7.

Intra-particle diffusion model of adsorption the AB210 onto Fe/Ce. Experimental conditions: T = 30 °C, pH = 7.0.

Adsorption isotherm

The equilibrium adsorption data are analyzed using Langmuir, Freundlich, and Temkin isotherm models. The Langmuir isotherm model assumes that the adsorption occurs on the homogeneous surface of the adsorbent. The Freundlich isotherm model is commonly used to describe the surface heterogeneity. The Temkin isotherm presumes that the heat of adsorption of all molecules in the layer reduces linearly with coverage because of the adsorbate–adsorbent interactions and the adsorption is characterized by means of the uniform distribution of bonding energies, up to some maximum binding energy.

The Langmuir adsorption isotherm can be described by the following (Qiu et al., 2014)

\frac{C e}{q e} = \frac{1}{q_{m} b} + \frac{C_{e}}{q_{m}}

(6)

where b (L/mg) is a constant related to the energy of adsorption and q_m (mg/g) is the monolayer adsorption capacity. The values of b and q_m are calculated from the intercept and slope of the linear plot C_e/q_e versus C_e.

The linear form of Freundlich isotherm is shown by the following (Zhou et al., 2015)

\ln (q_{e}) = \ln (K_{F}) + \frac{1}{n} \ln (C_{e})

(7)

where n and K_F (mg/g) are Freundlich constants, which represent adsorption strength and adsorption capacity, respectively. The values of n and K_F can be determined by the slope and intercept of the linear plot ln(q_e) versus ln(C_e). When n > 1, it indicates that the adsorption process is favorable.

The linear form of Temkin isotherm equation can be described as the following (Mashhadi et al., 2016)

q_{e} = B \ln (K_{T}) + B \ln (C_{e})

(8)

where K_T (L/mg) and B are equilibrium binding constants and related to the heat of adsorption, respectively. The values of K_T and B are defined by the intercept and slope of the linear plot q_e versus ln(C_e), respectively.

Figure 8(a) to (c) shows the linearized forms of the Langmuir, Freundlich, and Temkin isotherm models based on the experimental data, respectively. The values of these adsorption isotherm parameters are shown in Table 2. According to the table, Langmuir model is best fit as compared with the other two models, which suggests a homogeneous adsorption of Fe/Ce for AB210. This result also indicates the homogeneous nature of Fe/Ce surface, i.e. each AB210 dye molecule adsorption has equal adsorption activation energy. Based on the Langmuir equation, the value of q_m is calculated as 90.50 mg/g, which is considerably higher than that of the commercial CeO₂ and other CeO₂ with different morphologies previously reported (Table 3). In particular, the q_m of Fe/Ce is six times higher than that of the commercial CeO₂. At the same time, this result also clearly shows the excellent adsorption performance of Fe/Ce for AB210 except for the magnetic property.

Figure 8.

(a) Langmuir, (b) Freundlich, and (c) Temkin isotherms for AB210 adsorption onto Fe/Ce_. Experimental conditions: T = 30 °C, pH = 7.0.

Table 2.

Langmuir, Freundlich, and Temkin parameters for the adsorption of AB210 onto Fe/Ce.

Isotherm model	Isotherm parameters	Fe₃O₄/CeO₂
Langmuir isotherm model	q_m (mg/g)	90.50
	b (L/mg)	0.23
	R_L	0.05
	R ²	0.9939
Freundlich isotherm model	k_f (mg/g)	24.83
	n	3.12
	R ²	0.9578
Temkin isotherm model	K_T (L/mg)	5.08
	B	14.87
	R ²	0.9802

Table 3.

Maximum adsorption capacities of Ceria with different morphologies adsorbents to AB210 and other dyes.

Adsorbents	Dyes	Q_m (mg/g)	Reference
Fe/Ce	Acid Black 210	90.50	This work
CeO₂	Acid Black 210	94.23	This work
Commercial CeO₂	Acid Black 210	14.81	This work
Ceria nanoparticles	Acid Black 210	35.29	Hu et al. (2017)
Ceria nanofibers	Methyl Orange	10	Zhang et al. (2016)
Spindle CeO₂	Methyl Orange	3.2	Zheng et al. (2018)
Ceria nanoparticles	Congo Red	18	Li et al. (2012)
Octahedral CeO₂	Congo Red	58.4	Zheng et al. (2018)
Ceria hollow sphere	Acid Orange 7	22	He et al. (2014)
Ceria multiple layers	Acid Orange 7	25	He et al. (2014)
Ceria hollow spheres	Active Red X3B	82.16	Hu et al. (2017)

The essential characteristics of Langmuir isotherm model can be described according to the dimensionless equilibrium parameter (R_L), which is defined by the following (Kaur et al., 2008)

R_{L} = \frac{1}{1 + b C_{o}}

(9)

where C_o (mg/L) is the highest initial concentration of the adsorbate and b (L/mg) is the Langmuir constant. The value of R_L indicates the adsorption in nature to be unfavorable (R_L > 1), linear (R_L = 1), favorable (0 < R_L < 1), or irreversible (R_L = 0) (Ghasemi et al., 2015). The calculated value (R_L) is also listed in Table 2. The value (R_L) is 0.05, indicating a favorable adsorption of AB210 onto Fe/Ce under the investigated conditions.

Effect of temperature

The effect of temperature on adsorption of AB210 onto Fe/Ce is shown in Figure 9. The experiments are implemented at five different temperatures (20 °C, 30 °C, 40 °C, 50 °C, and 60 °C) under a constant initial dye concentration of 50 mg/L and pH = 7.0. With the increasing of the temperature from 20 °C to 40 °C, adsorption capacity increases relatively significant from 41.18 mg/g to 43.90 mg/g. However, the rate of increase in the adsorption capacity decreases slowly from 43.90 mg/g to 44.78 mg/g with the temperature increasing from 40 °C to 60 °C. This phenomenon may be that the adsorption process is divided into two stages, which is consistent with the intra-particle diffusion model. The experimental results show that the adsorption capacity increases with temperature in the range of 20–60 °C, which indicates that the adsorption of AB210 onto Fe/Ce is endothermic in nature.

Figure 9.

Effect of temperature on adsorption of the AB210 onto Fe/Ce. Experimental conditions: C_o = 50 mg/L, pH = 7.0.

The values of thermodynamic parameters of change in Gibbs free energy (ΔG°), enthalpy (ΔH°), and entropy (ΔS°) are determined using the following equations (Gad and El-Sayed, 2009)

\ln (K_{a}) = \frac{Δ S^{o}}{R} - \frac{Δ H^{o}}{R T}

(10)

Δ G^{o} = - R T \ln (K_{a})

(11)

K_{a} = \frac{q_{e}}{C_{e}}

(12)

where T (K) and K_a are the solution temperature and the adsorption equilibrium constant, respectively. The values of ΔS° and ΔH° can be calculated from the intercept and slope of the van’t Hoff plot shown in Figure 10. The values of ΔG° are calculated by equations (11) and (12). The obtained values are shown in Table 4.

Figure 10.

Van’t Hoff plot for the adsorption of the AB210 onto Fe/Ce. Experimental conditions: C_o = 50 mg/L, pH = 7.0.

Table 4.

Thermodynamic parameters for the adsorption of AB210 onto Fe/Ce.

T (°C)	ΔG° (kJ/mol)	ΔS° [J/(molK)]	ΔH° (kJ/mol)	R ²
20	−5.20	28.90	2.88	0.9926
30	−5.31
40	−5.40
50	−5.47
60	−5.53

The positive value of ΔH° (2.88 kJ/mol) indicates that the adsorption process of Fe/Ce for AB210 is endothermic in nature. The positive value of ΔS° (28.90 kJ/mol) suggests that the randomness increases at the solid–liquid interface during the adsorption of Fe/Ce for AB210. The negative values of ΔG° decrease from −5.20 kJ/mol to −5.53 kJ/mol with the increase in temperature, which indicates the feasibility and spontaneity of adsorption process. Moreover, this phenomenon also suggests that the adsorption process is more favorable at higher temperature. Similar result is also reported, e.g. Zheng et al. (2018) have synthesized shape-dependent CeO₂ nanostructures to adsorb the anionic dye Congo Red.

Generally, the adsorption process is considered to be physisorption in nature when the ΔG° value is in the range of −20 kJ/mol to 0 kJ/mol, while the ΔG° value range from −400 kJ/mol to −80 kJ/mol implies a chemisorption process (Zeng et al., 2014). Furthermore, if the value of ΔH° is lower than 40 kJ/mol, then it is certified that it conforms to the physisorption process (Kara et al., 2003). Considering the obtained values of ΔH° and ΔG°, the adsorption process of Fe/Ce for AB210 is considered as physical adsorption.

Effect of initial pH

The pH of the aqueous solutions is a key parameter in the adsorption process. Figure 11(a) shows the effect of initial pH on the released iron concentration of Fe₃O₄ and Fe/Ce in treated solution, respectively. As the pH value increases from 3.0 to 9.0, the released iron concentration of Fe₃O₄ and Fe/Ce decreases from 2.04 × 10⁻¹ mg/L to 6.46 × 10⁻⁴ mg/L, 9.05 × 10⁻³ mg/L to 2.58 × 10⁻⁴ mg/L⁻¹, respectively. However, when pH = 9.0–11.0, the effect is almost nonexistent. When pH = 3.0, the released iron concentration of Fe/Ce is about 22.5 times lower than that of Fe₃O₄. Because the loaded CeO₂ layer of Fe/Ce protects Fe₃O₄ inside from being affected by acidic condition. At the same time, it also indicates that CeO₂ is successfully loaded on the surface of Fe₃O₄. Similarly, the released iron concentration of Fe/Ce (C_Fe=6.46 × 10⁻⁴ mg/L) is about 15 times lower than that of Fe₃O₄ (C_Fe=9.70 × 10⁻³ mg/L) at pH = 7.0 condition. The results obtained are consistent with those of TEM. The effect of initial pH on adsorption of anionic dye AB210 has been studied in the range 3.0–11.0 at 30 °C with the initial dye concentration always fixed at 50 mg/L. As shown in Figure 11(b), as the pH value increases from 3.0 to 11.0, the adsorption capacity at equilibrium decreases from 44.02 mg/g to 1.52 mg/g after 120 min. It is more important that the adsorption capacity declines dramatically from 41.18 mg/g to 3.05 mg/g with the increase in the pH value from 7.0 to 9.0. It can be explained by the electrostatic attractions (Wu et al., 2009). Meanwhile, the zeta potential of Fe/Ce also shows a decreasing trend with the effect of initial pH too. The value (pH_pzc) of Fe/Ce is 7.7, at which the net surface charge on adsorbent is zero. The adsorbent surface has a net positive charge at pH < pH_pzc, while has a net negative charge at pH > pH_pzc. The AB210 is anionic dye, which dissociates into the cationic form (–OH₂⁺, –NH₃⁺) and the negatively charged dyes. At acidic pH, the presence of the hydroxyl (–OH) on the Fe/Ce surface is protonated to the cationic form (–OH₂⁺) and generates electrostatic attraction force with dyes anions. Therefore, the adsorption of AB210 onto Fe/Ce surface is more favorable at lower pH value. However, at basic pH, the hydroxyl (–OH) is deprotonated to anionic form (–O⁻), causing a decrease in the number of positively charged sites, which in turn does not favor the adsorption of negatively charged dyes molecules. Thus, adsorption capacity of AB210 onto Fe/Ce tended to decrease with the increasing pH value, which can be attributed to the electrostatic repulsion between the negatively charged sites on the adsorbent surface and the negatively charged dye molecules. Therefore, the adsorption of AB210 is favorable when the pH value is smaller than pH_pzc (Ahmad and Kumar, 2011; Konicki et al., 2017).

Figure 11.

(a) The effect of initial pH on iron release from Fe/Ce, commercial Fe₃O₄, respectively. Iron standard curve (inset). (b) The effect of initial pH of dye solution on adsorption capacity of the AB210 (the black line) and the effect of initial pH on zeta potentials of Fe/Ce (the red line). Experimental conditions: C_o=50.0 mg/L, V = 50.0 mL, T = 30 °C.

Conclusions

The removal of AB210 from aqueous solutions is studied by the Fe/Ce sorbent. Fe/Ce can be well dispersed in water and can be simply separated from water under external magnetic field. The adsorption process is investigated under key parameters such as initial dye concentration, temperature, and initial pH. The adsorption kinetic of Fe/Ce for AB210 follows the pseudo-second-order model well. The absorption isotherm of AB210 on Fe/Ce best fits the Langmuir adsorption model. The maximum adsorption capacity of Fe/Ce is 90.50 mg/g, which is six times higher than that of the commercial CeO₂. The thermodynamic parameters indicate that the adsorption of AB210 onto Fe/Ce is spontaneous and endothermic process in the experimental conditions. The thermodynamic analysis data indicates that the adsorption of AB210 dye onto Fe/Ce is confirmed as a physisorption process. In addition, the effect of initial pH also shows that the adsorption of AB210 on Fe/Ce is favorable at acid condition. Therefore, the research on adsorption equilibrium, thermodynamics, and kinetics of Fe/Ce could enrich the adsorption mechanism research of adsorbents and Fe/Ce may be a convenient and efficient adsorbent for the treatment of dye wastewater.

Footnotes

Declaration of Conflicting Interests

The author(s) declared no potential conflicts of interest with respect to the research, authorship, and/or publication of this article.

Funding

The author(s) received no financial support for the research, authorship, and/or publication of this article.

References

Aguayo-Villarreal

Ramírez-Montoya

Hernández-Montoya

et al . (2013) Sorption mechanism of anionic dyes on pecan nut shells (Carya illinoinensis) using batch and continuous systems. Industrial Crops & Products 48(3): 89–97.

Ahmad

Kumar

(2011) Adsorption of amaranth dye onto alumina reinforced polystyrene. Clean - Soil, Air, Water 39(1): 74–82.

Boumaza

Kaouah

Hamane

et al . (2014) Visible light assisted decolorization of azo dyes: Direct Red 16 and Direct Blue 71 in aqueous solution on the p-CuFeO₂/n -ZnO system. Journal of Molecular Catalysis A Chemical 393(18): 156–165.

Chong

Zhang

Tian

et al . (2016) Rapid degradation of dyes in water by magnetic Fe(0)/Fe₃O₄/graphene composites. Journal of Environmental Sciences 44(6): 148–157.

Dizge

Aydiner

Demirbas

et al . (2008) Adsorption of reactive dyes from aqueous solutions by fly ash: Kinetic and equilibrium studies. Journal of Hazardous Materials 150(3): 737–746.

Fiorentin

Trigueros

DEG

et al . (2010) Biosorption of reactive blue 5G dye onto drying orange bagasse in batch system: Kinetic and equilibrium modeling. Chemical Engineering Journal 163(1): 68–77.

Wang

(2011) Removal of heavy metal ions from wastewaters: A review. Journal of Environmental Management 92(3): 407–418.

Gad

El-Sayed

(2009) Activated carbon from agricultural by-products for the removal of Rhodamine-B from aqueous solution. Journal of Hazardous Materials 168(2–3): 1070–1081.

Gao

Zhang

Zhou

et al . (2018) Magnetic composite Fe₃O₄/CeO₂ for adsorption of azo dye. Journal of Rare Earths 36: 986–993.

10.

Ghasemi

Khosroshahy

Abbasabadi

et al . (2015) Microwave-assisted functionalization of Rosa Canina-L fruits activated carbon with tetraethylenepentamine and its adsorption behavior toward Ni(II) in aqueous solution: Kinetic, equilibrium and thermodynamic studies. Powder Technology 274: 362–371.

11.

Hao

Yang

Rao

et al . (2014) Facile additive-free synthesis of iron oxide nanoparticles for efficient adsorptive removal of Congo red and Cr(VI). Applied Surface Science 292(4): 174–180.

12.

Feng

et al . (2014) Solvothermal synthesis and characterization of ceria with solid and hollow spherical and multilayered morphologies. Applied Surface Science 322(322): 147–154.

13.

Mckay

(1999) Pseudo-second order model for sorption processes. Process Biochemistry 34(5): 451–465.

14.

Deng

Chen

(2017) Ceria hollow spheres as an adsorbent for efficient removal of acid dye. ACS Sustainable Chemistry & Engineering 5(4): 3570–3582.

15.

Huang

Chang

et al . (2011) Adsorption of cationic dyes onto mesoporous silica. Microporous & Mesoporous Materials 141(1): 102–109.

16.

Kara

Yuzer

Sabah

et al . (2003) Adsorption of cobalt from aqueous solutions onto sepiolite. Water Research 37(1): 224–232.

17.

Kasgöz

Durmus

(2008) Dye removal by a novel hydrogel-clay nanocomposite with enhanced swelling properties. Polymers for Advanced Technologies 19(7): 838–845.

18.

Kaur

Walia TPS and Kansal

(2008) Removal of Rhodamine-B by adsorption on walnut shell charcoal. Journal of Surface Science & Technology 24(3): 179–193.

19.

Kim

Song

(2014) Precise adjustment of structural anisotropy and crystallinity on metal–Fe₃O₄ hybrid nanoparticles and its influence on magnetic and catalytic properties. Journal of Materials Chemistry C 2(25): 4997–5004.

20.

Konicki

Aleksandrzak

Mijowska

(2017) Equilibrium, kinetic and thermodynamic studies on adsorption of cationic dyes from aqueous solutions using graphene oxide. Chemical Engineering Research & Design 123: 35–49.

21.

Konicki

Hełminiak

Arabczyk

et al . (2018) Adsorption of cationic dyes onto Fe@graphite core-shell magnetic nanocomposite: Equilibrium, kinetics and thermodynamics. Chemical Engineering Research & Design 129: 259–270.

22.

Wang

Zhang

et al . (2012) Surfactant-assisted synthesis of CeO₂ nanoparticles and their application in wastewater treatment. RSC Advances 2(32): 12413–12423.

23.

Zhao

Song

et al . (2017) Magnetic ordered mesoporous Fe₃O₄/CeO₂ composites with synergy of adsorption and Fenton catalysis. Applied Surface Science 425: 526–534.

24.

Liu

Han

Cui

et al . (2008) Synthesis and characterization of Fe₃O₄/polystyrene microspheres. Polymer Materials Science & Engineering 24(3): 137–140.

25.

Machado

Carmalin

Lima

et al . (2016) Adsorption of Alizarin Red S dye by carbon nanotubes: An experimental and theoretical investigation. Journal of Physical Chemistry C 120(32): 18296–18306.

26.

Maiti

Mukhopadhyay

Devi

(2017) Evaluation of mechanism on selective, rapid, and superior adsorption of Congo Red by reusable mesoporous α-Fe₂O₃ nanorods. ACS Sustainable Chemistry & Engineering 5(12): 11255–11267.

27.

Mane

Babu

PVV

(2011) Studies on the adsorption of Brilliant Green dye from aqueous solution onto low-cost NaOH treated saw dust. Desalination 273(2–3): 321–329.

28.

Mashhadi

Sohrabi

Javadian

et al . (2016) Rapid removal of Hg (II) from aqueous solution by rice straw activated carbon prepared by microwave-assisted H₂SO₄ activation: Kinetic, isotherm and thermodynamic studies. Journal of Molecular Liquids 215: 144–153.

29.

Mellah

Chegrouche

Barkat

(2006) The removal of uranium(VI) from aqueous solutions onto activated carbon: Kinetic and thermodynamic investigations. Journal of Colloid & Interface Science 296(2): 434–441.

30.

Meshko

Markovska

Mincheva

et al . (2001) Adsorption of basic dyes on granular activated carbon and natural zeolite. Water Research 35(14): 3357–3366.

31.

Ouyang

Xie

et al . (2013) Hierarchical CeO₂ nanospheres as highly-efficient adsorbents for dye removal. New Journal of Chemistry 37(3): 585–588.

32.

Pan

Liu

et al . (2016) Facile method for the synthesis of Fe₃O₄@HCP core–shell porous magnetic microspheres for fast separation of organic dyes from aqueous solution. RSC Advances 6(53): 47530–47535.

33.

Peng

Zeng

et al . (2016) Superabsorbent cellulose-clay nanocomposite hydrogels for high efficient removal of dye in water. ACS Sustainable Chemistry & Engineering 4(12): 7217–7224.

34.

Qiu

Jiang

Zhang

et al . (2014) Polyethylenimine facilitated ethyl cellulose for hexavalent chromium removal with a wide pH range. ACS Applied Materials & Interfaces 6(22): 19816–19824.

35.

Radia

Oumessaad

Adh’ Ya

et al . (2018) Adsorption of hexavalent chromium by activated carbon obtained from a waste lignocellulosic material (Ziziphus jujuba cores): Kinetic, equilibrium, and thermodynamic study. Adsorption Science & Technology 36(3–4): 1066–1099.

36.

Ratna

PBS.

Pollution due to synthetic dyes toxicity & carcinogenicity studies and remediation.

International Journal of Environmental Sciences 2012; 3: 940–955.

37.

Shan

Yan

Yang

et al . (2014) Magnetic Fe₃O₄/MgAl-LDH composite for effective removal of three red dyes from aqueous solution. Chemical Engineering Journal 252: 38–46.

38.

Song

Yang

Jia

et al . (2015) Self-assembled synthesis of urchin-like AlOOH microspheres with large surface area for removal of pollutants. RSC Advances 5(42): 33155–33162.

39.

Tahir

Rauf

(2006) Removal of a cationic dye from aqueous solutions by adsorption onto bentonite clay. Chemosphere 63(11): 1842–1848.

40.

Tan

Ahmad

Hameed

(2008) Adsorption of basic dye on high-surface-area activated carbon prepared from coconut husk: Equilibrium, kinetic and thermodynamic studies. Journal of Hazardous Materials 154(1): 337–346.

41.

Wang

Ding

Zhao

et al . (2015) Microstructure and magnetic properties of MFe₂O₄ (M =Co, Ni, and Mn) ferrite nanocrystals prepared using colloid mill and hydrothermalmethod. Journal of Applied Physics 117(17): 25–324.

42.

Tseng

Juang

(2009) Initial behavior of intraparticle diffusion model used in the description of adsorption kinetics. Chemical Engineering Journal 153(1): 1–8.

43.

Chen

Wang

et al . (2016) Magnetic material grafted poly(phosphotungstate-based acidic ionic liquid) as efficient and recyclable catalyst for esterification of oleic acid. Industrial Engineering Chemistry Researh 55(7): 1833.

44.

Xiao

Zhang

Xia

et al . (2016) Rapid and high-capacity adsorption of sulfonated anionic dyes onto basic bismuth(III) nitrate via bidentate bridging and electrostatic attracting interactions. RSC Advances 6(46): 39861–39869.

45.

Xing

Zhou

et al . (2011) Characterization and reactivity of Fe₃O₄/FeMnOx core/shell nanoparticles for methylene blue discoloration with H₂O₂. Applied Catalysis B Environmental 107(3–4): 386–392.

46.

Liu

Che

et al . (2014) Polarization enhancement of microwave absorption by increasing aspect ratio of ellipsoidal nanorattles with Fe₃O₄ cores and hierarchical CuSiO₃ shells. Nanoscale 6(11): 5782–5790.

47.

Wang

(2012) Magnetic nanoscaled Fe₃O₄/CeO₂ composite as an efficient Fenton-like heterogeneous catalyst for degradation of 4-chlorophenol. Environmental Science & Technology 46(18): 10145–10153.

48.

Deng

Gao

et al . (2010) Synthesis of magnetic microspheres with immobilized metal ions for enrichment and direct determination of phosphopeptides by matrix-assisted laser desorption ionization mass spectrometry. Advanced Materials 18(24): 3289–3293.

49.

Karmakar

Wang

et al . (2010) Multifunctional Fe₃O₄ cored magnetic-quantum dot fluorescent nanocomposites for RF nanohyperthermia of cancer cells. Journal of Physical Chemistry C 114(11): 5020–5026.

50.

Yao

Miao

Liu

et al . (2012) Synthesis, characterization, and adsorption properties of magnetic Fe₃O₄@graphene nanocomposite. Chemical Engineering Journal 184(1): 326–332.

51.

Tong

et al . (2012) One-step synthesis of magnetic composites of cellulose@iron oxide nanoparticles for arsenic removal. Journal Materials Chemistry A 1(3): 959–965.

52.

Zahrim

Tizaoui

Hilal

(2011) Coagulation with polymers for nanofiltration pre-treatment of highly concentrated dyes: A review. Desalination 266(1): 1–16.

53.

Zeng

Duan

Tang

et al . (2014) Magnetically separable Ni_0.6Fe_2.4O₄ nanoparticles as an effective adsorbent for dye removal: Synthesis and study on the kinetic and thermodynamic behaviors for dye adsorption. Chemical Engineering Journal 258(2): 218–228.

54.

Zhang

Shi

Yang

et al . (2016) Fabrication of electronspun porous CeO₂ nanofibers with large surface area for pollutants removal. Ceramics International 42(12): 14028–14035.

55.

Zhang

Zhu

Wang

et al . (2014) Ionothermal confined self-organization for hierarchical porous magnesium borate superstructures as highly efficient adsorbents for dye removal. Journal of Materials Chemistry 2(45): 19167–19179.

56.

Zheng

Wang

Long

et al . Shape-dependent adsorption of CeO₂ nanostructures for superior organic dye removal. Journal of Colloid and Interface Science 525: 225–233.

57.

Zhou

Yang

et al . (2015) Room-temperature crystallization of hybrid-perovskite thin films via solvent–solvent extraction for high-performance solar cells. Journal of Materials Chemistry A 3(15): 8178–8184.