Adsorptive removal of Cr(III)–EDTA chelate from an aqueous solution by magnetic mesoporous silica microspheres Open Access

Chromium (Cr) pollution of the aquatic environment is widely perceived as an environmental problem due to its nonbiodegradability and toxicity properties (Guan et al. 2017; Hai et al. 2020). The worldwide use of Cr in many industries, such as electroplating, tannery, metallurgy and organic chemical synthesis, has received great concerns because of its potential threats to environmental safety (Egodawatte et al. 2015; Guimaraes et al. 2020). In the environment, chromium was mainly in the form of trivalent chromium (Cr(III)) and hexavalent chromium (Cr(VI)). The toxicity and essentiality of Cr depend on its oxidation state (Novotnik et al. 2016). Many studies have shown that the excess accumulation of Cr(III) in humans can decrease immune system activity (Wang et al. 2019). Furthermore, Cr(III) could be converted into more toxic Cr(VI) during the changes in Eh and pH states (Luther et al. 2013; Su et al. 2016), which improves the toxicity of Cr(III). Thus, the elimination of Cr(III) ions from water is critical (Eyvazi et al. 2019).

For the real-water environment, the dissolved chromium mainly exists in the form of free hydrated ions or complexes (with inorganic or organic ligands). For example, a variety of organic pollutants was found in tannery effluent, which would react with Cr(III) and form diverse Cr(III)-bearing complexes (Chen & Pan 2020). Ethylenediamine tetraacetate (EDTA) is a commonly used chelating agent in industrial production. A stable complex (Cr(III)–EDTA) can be formed by chelation between Cr(III) and EDTA (Novotnik et al. 2016). Unlike free hydrated Cr(III), Cr(III)–EDTA is usually stubborn to a traditional treatment method due to its high chemical stability. Therefore, a variety of treatment technologies including chemical precipitation (Lofrano et al. 2013; Tang et al. 2020), membrane separation (Zhang et al. 2021), adsorption (Wang et al. 2020a), biological treatment (Xu et al. 2022), Fenton-like process (Ye et al. 2018), electrochemical reduction (Durante et al. 2011) have been employed for the elimination of Cr(III)–chelate. Among these techniques, adsorption has been deemed one of the most promising strategies for removing low concentration of Cr(III) due to its high cost-effectiveness, simple implementation and high efficiency (Liu 2021; Zhang et al. 2022).

Recently, many studies have been conducted to remove Cr(III) from water using different adsorbents. For example, montmorillonite-based porous adsorbents were prepared by the gel casting method for the removal of chromium citrate in low concentration (10 mg L^–1) in tanning wastewater (Hao et al. 2021). DETA-functionalized magnetic carbon adsorbent was found to efficiently remove Cr(III)–EDTA in aqueous solution, and the introduced amino active sites contribute to the improved Cr(III)–EDTA adsorption performance (Wang et al. 2021). Note that mesoporous adsorbents were found to be active for heavy metal adsorption even in high-salinity wastewater, for example, EDTA- or DTPA-modified magnetic mesoporous microspheres have a good adsorption performance on Cr(III) in high-salinity organic wastewater (Wang et al. 2020a, 2020b, 2020c). While adsorption of heavy metals on mesoporous silica adsorbents is very limited, surface modification with active groups such as amino groups, carboxyl groups and so on, may improve the heavy metal adsorption on mesoporous silica adsorbents. The amino group-modified adsorbents exhibit good adsorption for Cr(III)–EDTA (Wang et al. 2020a, 2021). Therefore, amino group-modified mesoporous adsorbent may improve the adsorption efficiency for Cr(III)–EDTA chelate in high-salinity water. Magnetic mesoporous silicon microspheres with large surface areas ordered mesoporous structure and easy modification have attracted much attention for the minimization of organic and inorganic pollutants in water. (3-Aminopropyl) trimethoxy silane (APTES) has a high content of nitrogen atoms, which can provide a large number of available adsorption sites. Therefore, in this study, APTES-modified magnetic mesoporous microspheres (FNMs/APTES) were synthesized to remove Cr(III)–EDTA from wastewater. The kinetics, adsorption equilibrium and effect of water chemical conditions are conducted, and related adsorption mechanism was also advised.

The chemicals used in this study, FeCl₃·6H₂O, NH₃·H₂O, NH₄NO₃, CH₃COONa, ethylene glycol, toluene, tetraethoxysilane (TEOS), hexadecyl trimethyl ammonium bromide (CTAB) were of analytical reagent grade and were purchased from Tianjin Kameiou Chemical Reagent Co., Ltd. (Tianjin City, China). (3-Aminopropyl) trimethoxy silane (APTES) was obtained from Alighting reagent (Shanghai) Co., Ltd. (Shanghai, China).

The morphology of the material was characterized by JEM2100f transmission electron microscope (TEM). D/max2200PC X-ray diffraction (XRD) of Rigaku Company in Japan was used to analyze the crystal structure before and after modification of the material, Cu-K radiation, graphite monochromator, tube voltage of 40 kV, tube current of 30 mA, scanning range of 2θ = 1–70°. A Nexus 870 Fourier Transform infrared spectrometer (FT-IR) from Nicolet was used to characterize the adsorbents. The thermogravimetry (TGA) Q500 of TA Company was used to analyze the thermal stability of prepared materials. The ASAP 2460 automatic specific surface area and porosity analyzer was used to characterize the adsorbents’ specific surface area and pore size distribution. The elemental distribution and surface chemical composition were analyzed with VG ESCALB MK-II X-ray photoelectron spectrometer (XPS).

The preparation of Fe₃O₄@nSiO₂@mSiO₂ (FNMs) is based on our previous studies (Wang et al. 2020c). The Fe₃O₄@nSiO₂@mSiO₂ was modified sequentially with APTES to introduce amine groups. Briefly, 1 g of FNMs was dispersed into 50 mL of anhydrous toluene, and 1 mL of APTES with N₂ atmosphere was added and stirred in an oil bath at 110 °C for 12 h. In the end, the products were magnetically separated, cleaned with toluene and ethanol for several times, and vacuum dried at 60 °C (FNMs/APTES).

Typically, 20 mg of FNMs/APTES was dispersed in 50 mL of HCl solution with different concentrations. The concentration of Fe was detected by ICP-OES and the Fe dissolution rate was calculated at different sampling times.

In this study, Cr(III)–EDTA was used to simulate the complex form of Cr(III) in actual wastewater. EDTA–2Na and Cr(SO₄)₃·6H₂O were mixed at 2:1 (molar ratio) and stirred at 25 °C for 24 h. All experiments were conducted at pH 4.0. 20 mg of FNMs/APTES was added to 50 mL solution of Cr(III)–EDTA with different initial concentrations (2.5–20 mg·L⁻¹) at 15, 25 and 35 °C with 140 r min⁻¹ for 12 h, respectively. For kinetic study, 200 mg of FNMs/APTES was added to 500 mL of 15 mg·L⁻¹ of Cr(III)–EDTA solution, sampling at different intervals. Furthermore, the residual Cr(III) concentration of the filtrate was analyzed by ICP-AES after the suspension was filtered by 0.45-μm fiber membrane, and the equilibrium adsorption amount at different temperatures was calculated by Formula (1).

To study the adsorption rate of Cr(III)–EDTA by FNMs/APTES, the adsorption kinetic models (pseudo-first-order kinetic model and pseudo-second-order kinetic model) were used to simulate the experimental data.

Wide-angle XRD patterns (Figure 1(c)) show that the material is made up of inverse apical ferrite (Fe₃O₄) strucutre, confirmed by 2θ peaks at 30.1, 35.5, 43.1, 53.5, 57.0 and 62.7° that correspond to the crystal planes at (220), (331), (400), (422), (551) and (440) (Elmobarak & Almomani 2021). No other impurity peaks appear, indicating that the material is of high purity. The formation of amorphous SiO₂ is inferred from diffraction peak at 23°, indicating that the prepared material is SiO₂-coated Fe₃O₄. The XRD characteristic peaks of FNMs/APTES remain unchanged, indicating that the modified material does not destroy the crystal structure. As shown in Figure 1(d), the peak at 2.88° represents the characteristic peak of the hexagonal mesoporous silica, indicating that the prepared FNMs are hexagonal mesoporous structures and shows the regular arrangement. Compared with FNMs, the intensity of the diffraction peak of FNMs/APTES is significantly reduced. This is because the modifier APTES enters into the mesoporous pore, and the pore effect reduces the intensity of the diffraction peak.

The N₂ adsorption desorption isotherms and the pore size distribution of the materials are shown in Figure 1(e) and 1(f). It was illustrated that a rapid decrease of surface area was obtained after APTES modification on the FNMs (from 519.02 to 393.43 m²·g⁻¹). FNMs are mesoporous materials with an average pore size of 2.26 nm and a pore volume of 0.30 cm³·g⁻¹ that reduced to 2.20 nm and 0.25 cm³·g⁻¹, respectively, after APTES modification, which may be because of the entry of APTES into the mesoporous channel of FNMs.

Thermogravimetric (TGA) curve (Figure 2(b)) indicated the mass loss of FNMs and FNMs/APTES was 8 and 14.05% between 20 and 600 °C. The FNMs exhibited a weight loss below 200 °C, which was due to the loss of adsorbed water. The weight loss observed from 200 to 600 °C is due to the loss of structural water. For FNMs/APTES, the loss in mass below 200 °C is due to departure of adsorbed water molecules, while the weight loss from 200 to 600 °C resulted from the decomposition of the organic compounds anchored on FNMs. The TGA loss of FNMs (8%) was significantly lower than that of FNMs/APTES (14.05%), indicating the successful modification on the surface of FNMs with APTES.

The stability test results of the material under acidic conditions are shown in Table 1. From the results, the dissolution rate of Fe increases with the increase of HCl concentration and immersion time. When the concentration of HCl was below 0.5 mol·L⁻¹, the dissolution rate was less than 0.5% when the immersion time was 60 h. The dissolution rate of FNMs increased significantly at HCl concentration of 1.0 mol·L⁻¹, and the dissolution rate is 6.19% after 60 h of immersion. Thus, the dissolution rate of Fe is low and FNMs demonstrated a good and stability under weak acid conditions.

Table 3 shows that the thermodynamic parameters for Cr(III)–EDTA adsorption on FNMs/APTES were calculated from the Gibbs–Helmholtz equation. The ΔG° is a negative value and decreases with increasing temperature, indicating that the adsorption of Cr(III)–EDTA by FNMs/APTES is a spontaneous process. In combination with increasing adsorption (Figure 4), ΔH° and ΔS° are positive, demonstrating that the adsorption process is an endothermic reaction and increases the disorder of the solid–liquid system.

The effects of inorganic cations (Na⁺, K⁺, Ca²⁺) may interfere in the adsorption of Cr(III)–EDTA onto the FNMs/APTES (Figure 6(b)). The results suggested that Na⁺ and K⁺ had some promoting effects on adsorption, whereas Ca²⁺ had inhibiting effects on adsorption. This is due to the addition of cations competing with the Cr(III)–EDTA in the solution for the adsorption site of the adsorbent. The addition of cations will compete with Cr(III)–EDTA for the adsorption site of the adsorbent. In the solution, due to the incomplete complexation between EDTA and Cr(III), the complexation ability of K⁺, Ca²⁺ and Na⁺ with EDTA is different. The order of the complex stability constants is: K_Cr−EDTA = 23.4>K_Ca–EDTA = 10.69>K_Na–EDTA = 1.6>K_K–EDTA = 0.8. Therefore, Ca²⁺ is more likely to complex with EDTA in the solution, and result in the slightly reduced Cr(III)–EDTA adsorption on the adsorbent.

The reusability experiments of FNMs/APTES were carried out by HCl solution and repeated for four adsorption–desorption cycles (Figure 7). The initial adsorption capacity was 12.01 mg·g⁻¹, and after four adsorption–desorption experiments, the adsorption capacity of FNMs/APTES was 9.97 mg·g⁻¹ (83.03% of the original adsorption amount). The decrease in Cr(Ш) adsorption is probably because of the incomplete desorption. The reusability of FNMs/APTES reflects that FNMs/APTES can be used as a promising candidate for Cr(III)–EDTA adsorption in water.

The XPS spectra of N 1 s before and after Cr(III)–EDTA adsorption is observed in Figure 8(b). The characteristic peaks of the pre-adsorption amino group (–NH₂) and protonated amino group (–NH₃⁺) appeared at 399.19 and 400.32 eV. After EDTA adsorption, the percentage of protonated amino groups on the adsorbent increased, which may be ascribed to the electrostatic interaction between EDTA anions in the solution and the protonated amino groups of the adsorbents. For Cr(III)–EDTA on FNMs/APTES, the ratio of protonated amino groups on FNMs/APTES decreased, which may be due to the electrostatic interaction between [Cr–EDTA]⁻ and [Cr–OHEDTA]²⁻ anions and protonated amino groups of the adsorbent. Moreover, the diffraction peaks of amino and protonated amino groups shifted toward the direction of high binding energy, which may be due to the formation of Cr(III)–amide complex compounds between Cr(III) and –NH₂ on the surface of FNMs/APTES (Wu et al. 2020).

In this work, FNMs/APTES was synthesized and used for the removal of Cr(III)–EDTA in water. The prepared FNMs/APTES exhibited good chemical stability for acid resistance. FNMs/APTES has a high adsorption capacity for Cr(III)–EDTA in water (13.13 mg·g⁻¹) and has a good recycling performance. The XPS analysis confirmed that FNMs/APTEs showed significant differences before and after Cr(III)–EDTA adsorption. In addition, the presence of cations Na⁺, K⁺ and Ca²⁺ had a slight effect on adsorption, reflecting the potential application for heavy metal-organic chelate in high-salinity water. Cr(III)–EDTA in water was removed as a whole by FNMs/APTES through electrostatic attraction between Cr(III)–EDTA anions and the protonated amino groups of FNMs/APTES.

We are thankful for the financial support provided by the National Natural Science Foundation of China (22076111), China.

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

The authors declare there is no conflict.