Introduction

With the fast advancement of industrial civilization worldwide, as well as the rapid global economic development, a large number of contaminants of metal(loid)s have been widely detected in the natural soil and water bodies (Fan et al. 2023; Gu et al. 2022; Hao et al. 2023; Li et al. 2021a; Ling et al. 2017). These contaminants of metal(loid) ion, which mainly included U(VI) (Bone et al. 2017; Hu et al. 2020; Kou et al. 2022; Pan et al. 2023), Se(IV)/Se(VI) (Hong et al. 2020; Kou et al. 2022; Lu et al. 2018; Wu et al. 2021), Cr(VI) (Lu et al. 2018; Wu et al. 2021), Mo(VI) (Li et al. 2020), Re(VII) (Li et al. 2020), 99Tc(VII) (Boglaienko et al. 2019), Eu(III) (Dong et al. 2018, 2021), Co(II) (Xing et al. 2016), As(III)/As(V) (Wang et al. 2014), Pb(II) (Zhao et al. 2018), Cd(II) (Awual et al. 2018; Zhao et al. 2018), Ni(II) (Flynn and Catalano 2017), rare earth (Li et al. 2021b), and so on, have been commonly reported to pose severe threats to aquatic ecosystems and human health even at trace levels due to their great toxic effects and have been increasingly becoming an ecological concern (Fan et al. 2023; Ling et al. 2017). So, their decontamination from various soil and water bodies has been an important and constant concern. In this respect, a variety of treatment approaches which included adsorption, coagulation, chemical, and biological treatment have been developed to remove metal(loid)s from water (Wan et al. 2018).

Among these metal(loid)s as-mentioned above, chromium (Cr) is one of the most common contaminants that could be found in the hazardous waste sites, often entering groundwater and soil environment from industrial effluents (Fan et al. 2023; Ling et al. 2017; Liu et al. 2023). Because of its high physiological toxicity, Cr has been regarded as a priority contaminant and environmental hazard, and thus it is very imperative to secure an effective method for the quick and complete removal of Cr from contaminated ecological environment (Ling et al. 2017). For Cr in the natural environment, there are two major oxidation states namely Cr(III) that is slightly soluble and considerably less toxic, as well as Cr(VI) that is more toxic, soluble, and mobile (Chen et al. 2021; Kang et al. 2020; Li et al. 2016, 2022; Saslow et al. 2018; Wu et al. 2020). These are three reasons for the fact that Cr(VI) is more toxic than Cr(III). First, Cr(VI) is labile but Cr(III) is inert. Then, Cr(VI) enters the cell via sulfate uptake pathway because of structural similarity of chromate with sulfate but Cr(III) cannot. Finally, Cr(VI) is mobile but Cr(III) is not (Saha et al. 2013a, b). Besides, Cr(VI) could enter into the natural environment via different sources (Saha and Orvig 2010; Saha et al. 2011). Thereby it is more important to find an effective and convenient method to control Cr(VI) exposure and reduce its toxic effect, and reduction of Cr(VI) into Cr(III) by biological material has been regarded as a very useful and cheap process (Mukherjee et al. 2014, 2015a, b, 2016; Nandi et al. 2017; Saha and Saha 2014). Using a high-performance photocatalyst for the efficient photocatalytic reduction of aqueous Cr (VI) into Cr(III) was also reported to be an important method (Ge et al. 2021; Yao et al. 2022a, b; Zhang et al. 2014, 2018, 2022). Besides, lots of other materials have been fabricated for reduction of Cr(VI) into Cr(III) from water bodies including iron carbide loaded on the N-doped carbon nanotubes, the FeS and titanate nanotubes nanocomposites, the graphene oxide adsorbed Fe(II), the h-BN supported nanoscale iron sulfide composite, etc. (Chen et al. 2021; Kang et al. 2020; Li et al. 2016, 2022; Liu et al. 2023; Saslow et al. 2018; Wu et al. 2020).

Among these materials and Cr(VI) remediation systems as-mentioned above, utilizing iron (Fe0) nanoparticles and their nanocomposites has been generally regarded as one of the most promising methods for the remediation of Cr(VI) and other related metal(loid)s from contaminated soil and water as a result of their high specific surface area and high reactive surface sites (Chen et al. 2017, 2023; Gu et al. 2007; Li et al. 2010; Shi et al. 2011a, b; Soliemanzadeh and Fekri 2017; Wei et al. 2021; Zhang et al. 2013). Besides, iron nanoparticles and their nanocomposites were also considered a reactive material in permeable reactive barriers (PRBs), which could provide enormous flexibility for both in situ and ex situ remediation applications of metal(loid)s (Li et al. 2010). So, considerable research on Cr(VI) remediation has been focused on the interaction kinetics and reaction mechanisms between Fe0-based materials and Cr(VI). The remediation of Cr(VI) in Fe0-based interaction systems was mainly ascribed to a procedure that involved the reductive precipitation of Cr(VI) into Cr(III), which was resulted from an electron transfer interaction between Cr(VI) and Fe(0)/Fe(II) at a solid/water interface (Shi et al. 2011a, b; Soliemanzadeh and Fekri 2017; Zhang et al. 2013). In addition, during the potential applications, decorated iron nanoparticles could be improved in the dispersion and persistence of iron nanoparticles in water, and thus enhancing the speed and efficiency of a Fe0-based remediation system. In this regard, clay minerals, which are environmental-friendly and abundant in nature, are promising modifying reagent, and therefore lots of clay minerals like bentonite and montmorillonite have been widely utilized to decorate iron nanoparticles (Shi et al. 2011a, b; Soliemanzadeh and Fekri 2017; Zhang et al. 2013). As an inert and porous material, celite was mainly composed of silica (SiO2), as well as some other inorganic oxides (Abbasi et al. 2023; Chang et al. 2007; Jabli et al. 2020; Liu et al. 2009; Meunier et al. 2014; Satar and Husain 2009). Since celite has desirable physical properties, i.e., inexpensive, chemical inertness, non-biodegradable, as well as interconnected pore structure, celite is very suitable for support of reactive materials. Thereby, there is a growing interest in the utilization of celite as support material of the catalyst by providing a better distribution to enhance reaction rates (Abbasi et al. 2023; Chang et al. 2007; Jabli et al. 2020; Liu et al. 2009; Meunier et al. 2014; Satar and Husain 2009). Nevertheless, according to our literature survey, no attention has been paid to the usage of celite as a modifying reagent to decorate iron (Fe0) nanoparticles.

Therefore, in the present paper, we aimed to prepare celite decorated iron nanoparticles (C-Fe0) and evaluate the sequestration performance of hexavalent chromium by C-Fe0 from aqueous solution. The main objectives of this paper were: (1) to prepare celite decorated iron nanoparticles and characterize the surface structure and properties using scanning electron microscopy (SEM), transmission electron microscopy (TEM), Fourier transform infrared (FTIR) spectra, X-ray diffraction (XRD), etc., (2) to investigate the adsorption kinetics and isotherms of Cr(VI) on C-Fe0 material, and (3) to reveal the interaction mechanisms between C-Fe0 and Cr(VI) using X-ray photoelectron spectroscopy (XPS).

Materials and methods

Chemicals and equipment

All chemicals including potassium dichromate (Shanghai Zhanyun Chemical Co. LTD), diphenylarbazone (Shanghai Maclin Biochemical Technology Co., LTD), sulfamic acid (Shanghai Maclin Biochemical Technology Co., LTD), ferrous sulfate heptahydrate (Sinopharm Chemical Reagent Co. LTD), phosphoric acid (Shanghai Zhanyun Chemical Co. LTD), sodium hydroxide (Xilong Chemical Co., LTD, hydrochloric acid (Zhejiang Zhongxing Chemical Reagent Co. LTD), sulfuric acid (Hangzhou Shuanglin Chemical Reagent Co. LTD), acetone (Hangzhou Shuanglin Chemical Reagent Co. LTD) were purchased in analytical purity without further treatment. The equipment used in this work included electronic analytical balance (EL204, Mettler Toledo (Shanghai) Instrument Co., LTD), precision pH meter (Five Easy Plus, Mettler Toledo (Shanghai) Instrument Co., LTD), CNC ultrasonic cleaner (KQ5200DA, Kunshan Ultrasonic Instrument Co., LTD), ultraviolet–visible spectrophotometer (SP-756-P, Shanghai Spectrometer Co. LTD), constant temperature heating magnetic agitator (DF-101S, Gongyi Yuhua Instrument Co., LTD), electric blast drying oven (GZX-9070MBE, Shanghai Boxun Industrial Co., LTD. Medical equipment factory), table top high speed centrifuge (TG16-WS, Hunan Xiangyi Laboratory Instrument Development Co., LTD), temperature controlled shaker(IKA KS4000i control, German aika), etc.

Experimental methods

To prepare celite decorated iron nanoparticles (C-Fe0), first, a certain amount of celite was put in a 500-mL flask, then 250 mL of deoxidized deionized water was added, with constant stirring under nitrogen conditions to ensure deoxidization. After that, a certain amount of FeSO4·7H2O was added, and a peristaltic pump was used to add 50 mL of NaBH4 solution at a rate of 4 mL/min (the molar ratio of NaBH4 to Fe2+ is ~ 3:1). The reaction is as: Fe2+ + 6H2O + 2BH−4 → Fe0↓ + 2B(OH)3 + 7H2↑ (Liu et al. 2014; Shi et al. 2011a, b). During the reaction, nitrogen was always introduced for deoxidation, and after the adding completion of NaBH4 solution, the stirring is continued for another 15 min, and thus the generated H2 could be completely discharged. The as-prepared material was centrifuged and washed to remove impurity ions, then freeze-dried, and finally vacuum freeze-dried to obtain powdered C-Fe0 material.

To evaluate the sequestration performance of Cr(VI) by C-Fe0 from aqueous solution,the batch adsorption experiment was used, wherein the solutions of Cr(VI) with different concentrations are diluted by the stock solution, and these solutions were added in a certain volume of reaction bottles, then the pH of the reaction solution was adjusted with 0.1 mol/L of NaOH or 0.1 mol/L of HCl. Sodium salt (Na) was used for the adsorption studies. In order to obtain an optimization conditions for the adsorption of Cr(VI), the effects of adsorbent dosage, initial pH value of solution, initial concentration of Cr(VI), and contact time on the adsorption process were studied respectively. After shaking for a certain time with a rotating speed of 220 r/min, a certain volume of reaction filtrate was diluted in a 25-mL colorimetric tube to determine the concentration of Cr(VI) by spectrophotometry method (Liu et al. 2014; Shi et al. 2011a, b).

Determination methods

The surface morphology was observed with a JEOL, JSM-6360LV scanning electron microscope (SEM) and transmission electron microscope (TEM, JEO, JEM-1011). The elemental composition of the reaction samples was characterized by energy-dispersive X-ray spectroscopy (EDS, Oxford instruments X-Max). The phase analysis was determined by the powder X-ray diffraction (XRD) measurements (PANalytical B.V., Empyrean, NL). The Fourier transform infrared (FTIR) spectra were measured on a Nicolet 6700 FTIR spectrometric analyzer using KBr pellets. Surface electronic states were analyzed by X-ray photoelectron spectroscopy (XPS, ESCALAB 250xi of SEMER Fisher Scientific and Technological Co., Ltd), with a Ka-Al radiation (hv = 1486.6 eV). XPS spectra were analyzed by XPS peak fitting program for WIN95/98 (XPSPEAK 4.0 Version 4.1) using the following asymmetric Gaussian–Lorentzian sum function. Line shapes of GL (30) were used for individual constituents (i.e., O1s, Fe2p, and Cr2p) (Chen et al. 2022, 2023; Li et al. 2021a; Wu et al. 2021).

Results and discussion

Batch adsorption results

Figure 1 displayed the experimental results of the sequestration performance of Cr(VI) by C-Fe0 via an adsorption process from aqueous solution as a function of solution pH, adsorbent dosage, and contact time. The optimization of pH played an important role in Cr(VI) adsorption on C-Fe0 due to the direct determination of the species of Cr(VI) and surface charge of C-Fe0 in water by the initial pH value. Previous studies have showed that Cr(VI) was mainly existed in the form of HCrO4− at low pH values and CrO42− was in a dominant position with pH increasing (Su et al. 2020). The effect of pH on sequestration performance of Cr(VI) by C-Fe0 was presented in Fig. 1A. The efficiency of Cr(VI) sequestration slightly increased with pH increased from 2 to 3 and then obviously decreased with pH increasing from 3 to 8. This was mainly because the corrosion of C-Fe0 was accelerated at lower pH values, and the rate of reaction was also accelerated (Li et al. 2012). The products of Fe2+ promoted the reductive conversion of Cr(VI) into Cr(III); thereby, the sequestration of Cr(VI) on C-Fe0 might involve a combined reduction and co-precipitation processes. Figure 1B exhibited the efficiency of Cr(VI) sequestration as a function of C-Fe0 dosage in the range from 0.1 to 1.0 g/L, which exhibited that the increment in C-Fe0 dosage was beneficial to the improvement in Cr(VI) sequestration efficiency. This result was mainly due to the factor that more C-Fe0 dosage could supply more surface reactive sites for Cr(VI) sequestration (Li et al. 2021a; Soliemanzadeh and Fekri 2017). Besides, more C-Fe0 addition would lead to the reduction of C-Fe0 utilization and the inconvenience of separation. So it is necessary to choose a suitable C-Fe0 dosage in the real application of Cr(VI) sequestration. The effect of contact time in Cr(VI) sequestration was also conducted in the adsorption experiment, and the results are displayed in Fig. 1C. The tend of Cr(VI) sequestration on C-Fe0 surface increased sharply from 5 to 60 min, then slowed down from 60 to 120 min, and finally reached adsorption equilibrium. There was a large number of adsorption sites on C-Fe0, and the high Cr(VI) concentration in solution makes it very easy to be removed on C-Fe0 at the initial adsorption stage. With the adsorption progressing, the adsorption sites on C-Fe0 surface and the Cr(VI) concentration decreased, and the adsorption slowed down until it reached equilibrium (Wang et al. 2022).

Fig. 1
figure 1

The effect of pH (A), sorbent dosage (B), and reaction time (C), on the sequestration of Cr(VI) on C-Fe0 material

Full size image

Figures 2 and 3 displayed the adsorption kinetics of Cr(VI) sequestration on C-Fe0 as a function of pH, Cr(VI) concentration, and adsorbent dosage, and the related fitting of pseudo-first-order kinetic model, pseudo-second-order kinetics model, intraparticle diffusion model. It was generally reported that the adsorption kinetics of metal(loid)s can be fitted accurately by the pseudo-first-order kinetic model, pseudo-second-order kinetics model, intraparticle diffusion model (Kong et al. 2016; Shi et al. 2011a, b).

Fig. 2
figure 2

The sequestration of Cr(VI) on C-Fe0 as a function of pH and Cr(VI) concentration (A), and the related kinetic fitting of (B) pseudo-first-order kinetic model, (C) pseudo-second-order kinetics model, (D) intraparticle diffusion model

Full size image
Fig. 3
figure 3

The sequestration of Cr(VI) on C-Fe0 as a function of pH and sorbent dosage (A), and the related kinetic fitting of (B) pseudo-first-order kinetic model, (C) pseudo-second-order kinetics model, (D) intraparticle diffusion model

Full size image

The pseudo-first-order kinetic model could be depicted as Eq. (1):

$$\mathrm{log}\left({q}_{e}-{q}_{t}\right)=\mathrm{log}{q}_{e}-\frac{{k}_{1}t}{2.303}$$
(1)

The pseudo-second-order kinetics model could be depicted as Eq. (2):

$$\frac{t}{{q}_{t}}=\frac{1}{{k}_{2}{q}_{e}^{2}}+\frac{1}{{q}_{e}}t$$
(2)

The intraparticle diffusion model could be depicted as Eq. (3):

$${q}_{t}={k}_{p}{t}^{1/2}+I$$
(3)

where qe (mg/g) and qt (min) are the adsorption capacities at equilibrium and at time t, respectively, k1 (min−1) and k2 (g/(mg min)) are the pseudo-first-order rate constant and pseudo-second-order rate constant, respectively, kp (mg/(min1/2 g)) is the intraparticle diffusion rate constant, and I is the intercept (Kong et al. 2016; Shi et al. 2011a, b). And the kinetic fitting parameters were all presented in Tables 1, 2, 3 and 4, respectively. Comparing the correlation coefficient value (R2) for different modes, we can see that the pseudo-second-order model fitted the adsorption Cr(VI) on C-Fe0 the best. So, the pseudo-second-order was dominant, and the potential rate-determining step in Cr(VI) adsorption on C-Fe0 was chemical interaction which involved sharing and exchanging of electrons between the binding site and Cr(VI) ions (Gerente et al. 2007; Luo et al. 2015; Su et al. 2020).

Table 1 Fitting parameters of pseudo-first-order kinetic model for Cr(VI) adsorption
Full size table
Table 2 Fitting parameters of pseudo-second-order model for Cr(VI) adsorption
Full size table
Table 3 Fitting parameters of intraparticle diffusion model for Cr(VI) adsorption
Full size table
Table 4 Fitting parameters of XPS analysis of C-Fe0 before and after reaction
Full size table

The isotherm data is very important for depicting the adsorption state at equilibrium, which can provide the basic information about thermodynamic performance (Niu et al. 2013; Zhao et al. 2018). The isotherm adsorption curves of Cr(VI) sequestration on C-Fe0 materials were shown in Fig. 4. We can clearly see that the adsorption increased with the increase of initial Cr(VI) concentration, suggesting that the adsorption of Cr(VI) on C-Fe0 favored high concentration because of the larger driving force that arose from high concentration gradient (Zhao et al. 2018). Herein, the Langmuir, Dubinin-Radushkevich (D-R), and Freundlich isotherm models were used to described the isotherm adsorption data, in order to reveal the isotherm adsorption mechanism (Niu et al. 2013, 2014; Zhao et al. 2018).

Fig. 4
figure 4

The adsorption isotherm of Cr(VI) sequestration on C-Fe0 material (A), Langmuir model (B), Freundlich model (C), and D–R model (D) fitting results

Full size image

The linear equation of Langmuir model could be depicted by Eq. (4):

$$\frac{{C}_{e}}{{q}_{e}}=\frac{{C}_{e}}{{q}_{m}}+\frac{1}{{q}_{m}{K}_{L}}$$
(4)

The linear equation of Freundlich model could be depicted by Eq. (5):

$$ln{q}_{e}=ln{K}_{F}+\frac{{lnC}_{e}}{n}$$
(5)

The linear equation of D-R model could be depicted by Eq. (6):

$$ln{q}_{e}=ln{q}_{m}-\beta {\varepsilon }^{2}$$
(6)

where Ce (mg L−1) and qe (mg g−1) are the equilibrium Cr(VI) concentration and adsorption capacity, respectively, qm (mg g−1) is the maximum adsorption amount, KL (L mg−1) is the Langmuir constant, KF (mg·g−1) is the Freundlich constant, and n is adsorption intensity index related to adsorption intensity. Besides, ε (kJ2 mol−2) is a Polanyi potential that could be obtained by ε = RTln(1 + 1/Ce), β (mol2 J−2) is an activity coefficient that was related to a mean free energy (E, kJ mol−1). The E value could be derived by this relationship, E \(=\frac{1}{\sqrt{2\beta }}\). We can use the E value to determine whether the adsorption of Cr(VI) is physical or chemical. When the value of E is below 8 kJ mol−1, it indicates a physical adsorption. When the E value was in the range of 8–16 kJ mol−1, it suggests a chemical adsorption (Zhao et al. 2018). According to the fitting parameters, we can see that the correlation coefficient of Langmuir (RL2, 0.994) was higher than Freundlich (RF2, 0.985) and the D-R model (RD-R2, 0.962), indicating the adsorption of Cr(VI) on C-Fe0 can be the best depicted by Langmuir model with a monolayer adsorption. In addition, the E value was determined to be in the range of 8–16 kJ mol−1, which indicated the adsorption of C-Fe0 for Cr(VI) was chemical interaction in nature.

Characterization results and mechanism insights

Herein, various characterization methods were used to reveal the structural changes of the materials before and after reaction. Figure 5 presented the SEM of celite, and C-Fe0, as well as TEM and EDS mapping of C-Fe0 before reaction. It could be seen from SEM that there existed a little of pores on celite particles with some discal structure, which make it a good possibility for Fe0 to be decorated on celite surfaces. According to the SEM and TEM of C-Fe0, we could observe that the shaped Fe0 particles were dispersed on celite surfaces. Elemental analysis from EDS mapping showed the presence of Fe, Si, O, and to a smaller extent of Ca, which further indicated the successful combination of Fe0 and celite. Figure 6 presented the FTIR spectra, and XRD patterns of C-Fe0 before and after reaction with Cr(VI). In the FTIR spectra, the band at ~ 3400 cm−1 might be caused by the stretching vibration of Si–OH group, the band at ~ 1020 cm−1 might be attributed to the bending vibration of Si–OH, the band at ~ 540 cm−1 might be attributed to the bending vibration of Fe–O. The intensity change of these bands before and after reaction suggess the chemical interaction between Cr(VI) and C-Fe0. Meanwhile, the XRD patterns indicated that C-Fe0 before and after reaction was poorly crystallized. It revealed that both samples consist of SiO2 with some other oxides. Besides, the reflection at 2θ ~ 44.5° was indicative of iron (Jing et al. 2015; Xu et al. 2014). Figure 7 showed the nitrogen adsorption–desorption isotherms and pore distributions of C-Fe0. The BET surface area of C-Fe0 was 9.24 m2/g, and the corresponding pore size (adsorption average pore width) is 11.8 nm, respectively. Figure 8 showed the SEM, TEM, and EDS mapping of C-Fe0 after reaction with Cr(VI). Compared with the EDS mapping of C-Fe0 before reaction, elemental analysis showed additional Cr in EDS mapping of C-Fe0 after reaction, suggesting the surface reaction of C-Fe0 and Cr(VI). The surface became scabrous after reacting with Cr(VI) and the chain-like aggregates of Fe0 became more clear, which might be resulted from the gradual cover of iron oxide layers like FeOOH and Fe2O3 on C-Fe0 surface (Chen et al. 2011).

Fig. 5
figure 5

SEM images of (A) celite, and (B) C-Fe0 samples, and TEM images of (C) C-Fe0, as well as SEM–EDS mapping of (D) C-Fe0 before reaction

Full size image
Fig. 6
figure 6

FTIR spectra (A), and XRD patterns (B) of C-Fe.0 samples before and after reaction with Cr(VI)

Full size image
Fig. 7
figure 7

Nitrogen adsorption–desorption isotherms and pore distributions of C-Fe.0

Full size image
Fig. 8
figure 8

SEM images (A), and TEM images (B), as well as SEM–EDS mapping (C) of C-Fe.0 after reaction with Cr(VI)

Full size image

It has been widely believed that Cr(VI) sequestration by iron and its composites involved a combined processes of physical adsorption and chemical reduction (Wang et al. 2020). So, in the present work, XPS analysis was conducted for characterization of C-Fe0 before and after reaction of Cr(VI). Figure 9 showed the surveyed XPS spectra of C-Fe0 before and after reaction with Cr(VI), as well as the corresponding high XPS spectra of Cr2p, O1s, and Fe2p, before and after reaction with Cr(VI). Binding energies of O1s at ~ 528 eV, ~ 529 eV, ~ 531 eV, ~ 532 eV, and 533 eV were assigned to O–Cr, O–Fe, O–C, O–H, and O = C, respectively (Li et al. 2022; Wang et al. 2020; Wu et al. 2020), suggesting that surface complexation had an important effect on Cr(VI) sequestration. The Fe2p spectrum has three peaks, namely, Fe0 at ~ 706 eV, Fe(III) at ~ 710 eV, and FeOOH at ~ 724 eV; meanwhile, the Cr2p spectrum has four peaks, i.e., Cr2O3 at ~ 576 eV, Cr(III)-Fe(III) at ~ 577 eV, Cr(OH)3 at ~ 586 eV, and Cr(VI) at ~ 587 eV (Lyu et al. 2017, 2018; Wang et al. 2020; Wu et al. 2020), which proved the reduction reaction between Cr(VI) and C-Fe0. It was reported by Wang et al. (2020) that Cr(VI) sequestration on Fe0-based composites followed a common process. Namely, when Fe0-based composites contacted with Cr(VI), electrons could directly transfer from Fe0 to Cr(VI); thus, reduction of Cr(VI) by Fe0 is favorable (Eq. (7)). Then, reduction between Cr(VI) and Fe(II) could spontaneously happen, which reduces Cr(VI) indirectly into Cr(III) (Eq. (8)). Meanwhile, these reactive Fe(II) species could be constantly generated through electron transfer among different Fe species (Eqs. (9) and (10)). Finally, Cr(VI)–Fe(III) and Cr(III)–Fe(III) precipitation, as well as iron oxides could be formed on solid surface because of pH variations during the reaction (Eqs. (11)–(13)) (Wang et al. 2020).

Fig. 9
figure 9

The surveyed XPS spectra of C-Fe.0 before and after reaction with Cr(VI) (A), as well as the high XPS spectra of Cr1s (B), O1s before (C), and after (D), reaction with Cr(VI), Fe2p before (E), and after (F), reaction with Cr(VI)

Full size image
$${2\mathrm{HCrO}}_{4}^{-}+{14\mathrm{H}}^{+}+{3\mathrm{Fe}}^{0}\to {3\mathrm{Fe}}^{2+}+{8\mathrm{H}}_{2}\mathrm{O}$$
(7)
$${\mathrm{HCrO}}_{4}^{-}+{7\mathrm{H}}^{+}+{3\mathrm{Fe}}^{2+}\to {\mathrm{Cr}}^{3+}+{4\mathrm{H}}_{2}\mathrm{O}+{3\mathrm{Fe}}^{3+}$$
(8)
$${\mathrm{Fe}}^{0}-{2\mathrm{e}}^{-}\to {\mathrm{Fe}}^{2+}$$
(9)
$${2\mathrm{Fe}}^{0}-{\mathrm{Fe}}^{0}\to {3\mathrm{Fe}}^{2+}$$
(10)
$${2\mathrm{Fe}}^{3+}+{6\mathrm{OH}}^{-}\to 2\mathrm{Fe}{\left(\mathrm{OH}\right)}_{3}\left(\mathrm{s}\right)\to {\mathrm{Fe}}_{2}{\mathrm{O}}_{3}\left(\mathrm{s}\right){3\mathrm{H}}_{2}\mathrm{O}$$
(11)
$${\mathrm{Fe}}^{2+}+{\mathrm{Cr}}_{2}{\mathrm{O}}_{4}^{2-}\to {\mathrm{FeCr}}_{2}{\mathrm{O}}_{4}\left(\mathrm{s}\right)$$
(12)
$${2\mathrm{Cr}}^{3+}+{6\mathrm{OH}}^{-}\to 2\mathrm{Cr}{\left(\mathrm{OH}\right)}_{3}\left(\mathrm{s}\right)\to {\mathrm{Cr}}_{2}{\mathrm{O}}_{3}\left(\mathrm{s}\right){3\mathrm{H}}_{2}\mathrm{O}$$
(13)

Combining previous reports (Lv et al. 2017) and the observed results herein, we could conclude that Cr(VI) sequestration on C-Fe0 was mainly composed of both adsorption and reduction. Firstly, abundant Cr(VI) in solution could be quickly adsorbed onto C-Fe0 surface and gradually diffused into interior of C-Fe0, which confirmed with rapid decrease of Cr(VI) at the initial stage. When contacted with C-Fe0, these adsorbed Cr(VI) were reduced into Cr(III). After that, some of these Cr species released back into solution. Besides, C-Fe0 were oxidized to Fe2+ during reaction, which could continue to participate in Cr(VI) reduction. Finally, released Cr(III), Fe2+, and Fe3+, as well as remaining Cr(VI) co-precipitated as Cr(III)-Fe(III) (oxy)hydroxides to further remove Cr(VI) from aqueous solution (Lv et al. 2017). The findings indicated that C-Fe0 was a good material for Cr(VI) sequestration.

Conclusions

In the present paper, novel composites namely celite decorated iron nanoparticles (C-Fe0) were prepared by an in situ reduction method, and the sequestration performance of Cr(VI) by C-Fe0 from aqueous solution was evaluated. The influence of ambient conditions, including solution pH, adsorbent dosage, and initial Cr(VI) concentration on Cr(VI) sequestration performance, was studied. The results indicated that increasing pH exhibited the most significantly negative effect on Cr(VI) sequestration. Kinetics study indicated that pseudo-second-order adsorption model was more suitable to describe the Cr(VI) sequestration, and the Langmuir adsorption model fitted the best with the isotherm data of Cr(VI) adsorption on C-Fe0. Finally, the possible Cr(VI) sequestration path by C-Fe0 was analyzed. In general, the C-Fe0 exhibits many advantages as low cost and environmental benignity, providing an admirable alternative over the common methods in Cr(VI) remediation.