Irradiated Gao Miao Zi Bentonite for Uranium Retention: Performance and Mechanism

Abstract

Bentonite has been considered as backfill material in the long-term deep geological disposal sites for radioactive waste. The performance of raw and irradiated bentonite based on the retention of radioactive nuclides, such as U(VI), is a critical factor for its application. Herein, the intrinsic features and adsorption behavior of Gao Miao Zi (GMZ) bentonite based on uranyl ions was investigated. In aqueous solutions, bentonite can achieve an adsorption rate of up to 100% for U(VI). The primary mechanism of U(VI) adsorption by GMZ bentonite is ion exchange, supplemented by surface complexation. Strong irradiation can introduce slight structural changes and framework fractures in bentonite, reducing its adsorption capacity for U(VI). This study provides an in-depth analysis of the adverse effects of high doses of radiation (100 kGy) on the microstructure and adsorption properties of bentonite, offering important insights for the safe storage of radioactive waste.

1. Introduction

As human society evolves rapidly, the demand for efficient and sustainable energy, such as nuclear energy, is increasing [1]. The exploration and advancement of nuclear energy requires a great demand for uranium. As a typical contaminant of nuclear wastewater, uranium exists in nature in three main isotopes: ²³⁸U, ²³⁵U, and ²³⁴U, with half-lives of 4.47 × 10⁹ a, 7.04 × 10⁸ a, and 2.46 × 10⁵ a, respectively [2,3]. Uranium is mainly present in the environment in the oxidation states of +4 and +6. U(IV) is found in a low-toxicity, insoluble form under reducing conditions, whereas U(VI) occurs as uranyl (UO₂²⁺), which is highly toxic and exhibits considerable mobility [4,5,6,7,8]. Uranium has radioactive and chemical toxicity. Radiological toxicity arises from α, β, and γ rays emitted during uranium decay, which can cause internal exposure to living organisms, leading to organ damage, genetic mutations, and even cancer. Chemical toxicity is mainly due to the accumulation of uranium in the human body through the food chain, causing harm to various organs [1,9,10,11]. Uranium tailings are the main source of uranium contamination [12]. In the uranium mining area in central Portugal, five open-pit pits operated between 1987 and 1991, producing ca. 74.97 tons of U₃O₈. The concentration of uranium in water peaked at the beginning of spring, primarily due to increased flow that leached uranium from secondary minerals into the water [13]. In the northern uranium mining region of Saskatchewan (Canada), substantial quantities of overburden material are removed and placed in waste rock piles during the open-pit mining process. This can lead to acid mine drainage, resulting in increased concentrations of radionuclides (i.e., uranium), which not only poses a threat to the groundwater system but may also has serious impacts on the ecosystem [14]. Near the uranium mines in the Siavonga region of Zambia’s Southern Province, uranium concentrations in samples from drinking water sources, such as streams, dams, and boreholes, exceed the safety limits established by the World Health Organization. The cancer risk related to uranium in the majority of samples exceeds the acceptable threshold (10⁻⁶), indicating a heightened risk linked to the consumption of this water [15]. The survey showed that the uranium concentration in uranium-containing wastewater at home and abroad was low (

Table 1. Concentration of uranium in uranium-containing wastewater from different sources.

Country	Source	Concentration of Uranium (mg/L)
China	Uranium mine wastewater in western [16]	10.00
	A mine wastewater in Shaoguan [17]	2.68
	Leachate from a uranium tailing	19.80
Australia	A uranium-contaminated groundwater [18]	0.29
United States	Uranium wastewater from the Crow Butte ISR sites [19]	12.20
Brazil	Mine drainage from the Osam Usumi mine [20]	1.05–4.46

). Hence, this study will focus on the treatment of low-concentration wastewater.

The globally accepted approach for the ultimate disposal of high-level radioactive waste (HLNWR) is deep geological disposal, which includes a multi-barrier system composed of natural and engineered barriers to ensure the long-term containment of radioactive waste, thereby protecting public health and the ecological environment [21]. In several countries, bentonite is considered a promising backfill material for deep geological repositories [22]. For example, China’s HLNWR used Gaomiaozi (GMZ) bentonite as the preferred mineral deposit for buffer/backfill material [23]. The strong radiation and high heat generated by high-level waste might penetrate metal containers, affecting the surrounding groundwater and buffer materials. This can lead to the radiation-induced decomposition of groundwater, accelerating the erosion and oxidation of container materials, which in turn may impact the stability and buffering performance of bentonite [24]. Preliminary studies have been conducted both domestically and internationally on the adsorption behavior of bentonite and the clay alterations in structure and properties following irradiation. Zong et al. found that sodium-based bentonite from Zhejiang (China) has high physicochemical stability and adsorption performance, making it an economical and efficient material for U(VI) fixation and recovery from water solutions [25]. Wang et al. found that bentonite effectively enriches and solidifies uranium (VI) ions, with mechanisms including outer-sphere complexation and ion exchange at pH < 7.0 and inner-sphere complexation at pH > 7.0 [26]. Dong et al. found that following γ irradiation aging, the bentonite’s Fe(II) content rose while its Fe(III) content fell, but the mineral varieties and functional groups remained unchanged, and further study is needed for specifics [27]. Cheng et al. observed that high-dose irradiation had a slight effect on the structure of bentonite, causing fractures in the bentonite framework. The U(VI) adsorption capacity decreased relative to non-irradiated bentonite [28]. Cheng et al. investigated the U(VI) adsorption capacity of colloids made from bentonite irradiated with various γ-ray doses [29]. The results indicated that as the irradiation dose increased, the adsorption by bentonite colloids reduced. Furthermore, different bentonites from various regions will have significant variations in geological formation, mineral composition, and chemical properties. There is no clear conclusion regarding the adsorption behavior and reaction mechanisms of GMZ bentonite towards radioactive nuclides. Furthermore, studies on the effects of irradiation on clay minerals, such as montmorillonite, kaolinite, and illite, are still in preliminary stages. In particular, the impact of γ-irradiation on the microstructure and physicochemical characteristics of GMZ bentonite have not yet reached a consensus or provided comprehensive results, indicating a need for further in-depth research [22].

This study aims to explore uranium adsorption by GMZ bentonite from Inner Mongolia based on various factors and its adsorption mechanisms and to examine the impact of irradiation on bentonite’s microstructure and adsorption effectiveness.

2. Experimental

2.1. Materials

In this study, all solutions were prepared with deionized water (resistivity = 18.2 MΩ·cm). All chemical reagents utilized in this experiment were of analytical quality. The uranyl nitrate hexahydrate was obtained from Meryer Biochemical Technology Co., Ltd. (Shanghai, China). Arsenazo III, sodium hydroxide, and sodium chloride were obtained from Shanghai Macklin Biochemical Co., Ltd. (Shanghai, China). Ethylenediaminetetraacetic acid disodium salt was sourced from Tianjin Zhiyuan Chemical Reagents Co., Ltd. (Tianjin, China), nitric acid from Guangzhou Chemical Reagent Factory (Guangzhou, China), and hydrochloric acid from Tianjin Fuyu Fine Chemical Co., Ltd. (Tianjin, China).

2.2. Irradiation and Preparation of Samples

The study utilized bentonite sourced from GMZ in Xinge County, Inner Mongolia, China. Dry the received bentonite at 105 °C for 8 h, then grind it in an agate mortar, and sift it through a 200-mesh screen. At room temperature, the bentonite was irradiated separately using a high-intensity ⁶⁰Co source and a linear electron beam accelerator to achieve a cumulative dose of 100 kGy. The statistics of the test samples are listed in Table S1.

2.3. Characterization

The morphology and chemical composition of GMZ bentonite were recorded using the Zeiss SIGMA300 (Germany) field emission scanning electron microscope with an energy dispersive spectrometer (EDS) system and the Talos F200i (Thermo Scientific, Waltham, MA, USA) transmission electron microscope. The elemental content was assessed using an Inductively Coupled Plasma Optical Emission Spectrometer (Agilent 720, Agilent, Santa Clara, CA, USA). The phase composition was measured using an X-ray diffractometer with Cu Kα radiation (Rigaku Smartlab SE, Rigaku, Japan) in the 2θ range of 5–80°. Fourier-transform infrared (FT-IR) spectroscopy was used to obtain FTIR spectra ranging from 4000 to 400 cm⁻¹ (Thermo fisher Nicolet IS50, Thermo Scientific, Waltham, MA, USA). XPS data, collected from an X-ray photoelectron spectrometer (Thermo Fisher Scientific 250Xi, Thermo Scientific, Waltham, MA, USA), were analyzed using XPS Peak software to provide insights into surface element compositions and valence states. The Brunauer Emmett Teller (BET) specific surface area of the sample was determined using N₂ adsorption–desorption measurements conducted with a specialized surface area and pore size analyzer (TriStar II 3020). Zeta-potential measurements were conducted using the Nanobrook 90Plus Zeta model (Brookhaven, Suffolk, NY, USA).

2.4. Batch Adsorption Experiments

Adding the U(VI)-containing solution, NaCl aqueous solution, humic acid aqueous solution, and GMZ bentonite to the conical flask according to different experimental conditions, adjust the reaction pH by adding a 0.1 M HNO₃ or NaOH solution, with a negligible volume [30]. The suspension was subsequently mixed in a shaker for 1 h and centrifuged at 6000 rpm for 10 min, and the solution was then filtered through a 0.45 μm filter. The uranium concentration in the solution was measured using the Arsenazo III method at a maximum wavelength of 652 nm [31].

The adsorption rate and capacity were determined by employing the following formula [32,33,34]:

$$\mathrm{Adsorption}\mathrm{\%}={\displaystyle \frac{C_0-C_e}{C_0}}\times 100\mathrm{\%}$$

(1)

$$Q_e={\displaystyle \frac{\left(C_0-C_e\right)\times V}{m}}$$

(2)

where Q_e represents the uptake amount (mg/g), C₀ is the initial U(VI) concentration (mg/L), C_e is the U(VI) concentration after adsorption (mg/L), V is the volume of the U(VI) solution (mL), and m is the mass of the adsorbent (g). To obtain credible and reproducible results, all adsorption experiments were repeated at least twice.

3. Results and Discussion

3.1. Structural Characterization of GMZ Bentonite

SEM images of Bento ^N, Bento ^EB, and Bento ^γ are shown in Figure 1a–c. Before irradiation, the GMZ bentonite has a layered structure composed of many stacked flakes with smooth surfaces and some curling at the edges (Figure 1a). Additionally, numerous small particles, initially inferred to be impurities, such as quartz, adhere to the flakes. After irradiation, the orderliness of fresh Bento has been disrupted, and the layered structure has transformed into a flake-like accumulation with a distinct tendency to disperse, while the curling of the edges has become more pronounced (Figure 1b,c) [35,36]. Transmission electron microscope (TEM) analysis further verifies the layered structure of GMZ bentonite, showing that the surface is loose and porous (Figure 1d–e). A lattice spacing of 0.449 nm is associated with the (100) plane of GMZ bentonite particles. The elemental mapping results (Figure 2a), elemental composition (Figure 2b), and elemental content (Figure 2c) show the distribution of elements in bentonite. The primary components in bentonite are Al, Si, O, Na, Mg, K, Ca, and Fe, with the oxygen content significantly higher than that of other elements, indicating that sodium bentonite contains a large amount of oxygen-containing groups [37,38]. In terms of metallic elements, in addition to the Al, Fe, and Mg originally present in the bentonite structure, exchangeable elements, such as Na, K, and Ca, were also detected between the bentonite layers.

The powder XRD pattern (Figure 3a) shows the crystal structure of GMZ bentonite [38]. Montmorillonite, the main component of GMZ bentonite (labeled M), exhibits diffraction peaks at 2θ = 7.10° and 19.76°, corresponding to (001) and (100) crystal planes, respectively, and other peaks corresponding to quartz (labeled Q). The layer spacing d (001) value of the montmorillonite mineral is affected by the type and quantity of interlayer cations, primarily Na⁺ and Ca²⁺. Among them, Ca²⁺ can adsorb double-layer water molecules, while Na⁺ can only adsorb single-layer water molecules [39]. The d (001) value of the GMZ bentonite used in this experiment is 12.559 Å, indicating that the main interlayer cation is Na⁺, which might be a sodium-based bentonite. The main change observed after irradiation was the layer spacing d (001), which increased to 1.4233 Å after γ irradiation and 1.4770 Å after electron beam irradiation. This indicates that irradiation may cause bentonite to undergo dehydroxylation, breaking hydrogen bonds and leading to interlayer expansion [40].

In Figure 3b, the FT-IR spectrum of GMZ bentonite is presented, highlighting the identification of functional groups and their changes before and after irradiation. In the spectrum of the original sample, important peaks for identifying functional groups in bentonite are located at 3628 cm⁻¹ (Al-OH), 3432 cm⁻¹, and 1639 cm⁻¹ (O-H of water molecules), 1035 cm⁻¹ (Si-O), and 796 cm⁻¹ (O-Si-O) [36]. In Bento ^EB, the stretching vibration band of Al-OH shifts from 3628 cm⁻¹ to 3626 cm⁻¹; the stretching vibrations of the H-OH groups shift from 3441 cm⁻¹ and 1645 cm⁻¹ to 3432 cm⁻¹ and 1639 cm⁻¹, respectively; and the Si-O-Si stretching mode at the tetrahedral sites moves from 1035 cm⁻¹ to 1042 cm⁻¹. For Bento ^γ, the stretching vibrations of the H-OH groups in the adsorbed water molecules move from 3441 cm⁻¹ and 1645 cm⁻¹ to 3475 cm⁻¹ and 1654 cm⁻¹, respectively, while the Si-O-Si stretching mode shifts from 1035 cm⁻¹ to 1034 cm⁻¹ [41,42]. The samples after irradiation show a distinct absorption band around 1400 cm⁻¹, corresponding to the bending vibrations of water molecules. Meanwhile, the intensities of the bands located at 3628 cm⁻¹, attributed to Al-OH, and at 1035 cm⁻¹, attributed to Si-O, have changed. These alterations suggest that the skeletal structure of the bentonite (silicon-oxygen tetrahedra and aluminum-oxygen octahedra) may have undergone slight changes.

The nitrogen adsorption–desorption isotherms of GMZ bentonite, measured both before and after irradiation, along with their pore size distribution curves, are presented in Figure 3c,d, with pore structure parameters detailed in

Table 2. Pore structure parameters of GMZ bentonite particles.

Material	SBET (m2/g)	Aperture (nm)	Hole Capacity (cm3/g)
Bento N	19.9368	8.9857	0.0562
Bento EB	22.4455	9.0902	0.0656
Bento γ	23.0011	9.0032	0.0675

. It can be observed that both curves display typical Type IV isotherms and H3-type hysteresis loops. The specific surface area of natural bentonite is 19.9368 m²·g⁻¹, and the average pore diameter of the particles is 8.9857 nm, classifying it as a mesoporous structure [43]. The larger specific surface area and mesoporous structure of GMZ bentonite indicate that it offers multiple adsorption sites for the adsorption of radioactive nuclides in wastewater, signifying its extensive research value in wastewater treatment. Clearly, after irradiation, the S_BET value, pore volume, and pore diameter have all increased. This may be due to the electrostatic repulsion generated during irradiation, which enhanced particle dispersion on the bentonite surface [36].

3.2. Batch Adsorption Studies

3.2.1. Effect of Dosage, Contact Time, and Initial Concentration

Figure 4a shows the effect of the contact time and dosage on the rate of uranium absorption by natural GMZ bentonite. The absorption rate was elevated significantly in the first 5 min and reached equilibrium in about 60 min, so the contact time of 60 min was selected for subsequent experiments. In the initial stage of adsorption, there are numerous adsorption sites on the bentonite, leading to a rapid adsorption process. As these adsorption sites are occupied by uranium ions in the solution, the adsorption reaches saturation [38]. With the increase in the adsorbent dosage from 2 g/L to 10 g/L, the equilibrium adsorption rate of U(VI) increased from 93.78% to 99.74%. This phenomenon is due to the increase in the solid–liquid ratio providing more exchangeable sites [25,44]. However, once the dosage reaches a certain level, further increasing it no longer enhances the adsorption rate. Therefore, considering both safe disposal and reducing pollutant treatment costs, the optimal dosage of GMZ bentonite for U(VI) absorption is 6 g/L.

Figure 4b illustrates how the initial concentration of U(VI) impacts both the rate of adsorption and the adsorption capacity of GMZ bentonite. As the concentration of uranium rises, the rate at which bentonite absorbs uranium progressively diminishes, whereas the equilibrium adsorption capacity consistently increases. This is because when the amount of bentonite added is constant, the adsorption sites present on the bentonite are restricted. When uranium is present in the solution at trace levels, bentonite can rapidly adsorb to it. However, when there is an excess of uranium in the solution, the adsorption sites quickly approach saturation, and the excess uranium ions can only exist as free hydrated ions, resulting in a reduction in the adsorption rate [45]. The inverse correlation between the quantity of uranium adsorbed and the rate of its adsorption can be attributed to the fact that as the uranium concentration increases, the amount of uranium supplied by the solution continues to grow, accumulating on the surface and within the pores of the bentonite, which aids in the efficient adsorption of uranium by bentonite. Considering that the actual concentration of uranium in wastewater generally does not exceed 20 mg/L, it is reasonable to choose 20 mg/L uranium for further studies.

3.2.2. Effect of pH

The adsorption of U(VI) on bentonite was examined in the pH range of 2.0 to 9.0. As shown in Figure 5a, between pH 2.0 and 5.0, the uranium adsorption rate by bentonite gradually increases. However, when the pH rises further from 5.0 to 9.0, the adsorption rate begins to decrease sharply. Therefore, the adsorption of U(VI) is highly influenced by the pH, with weakly acidic conditions being more beneficial for uranium uptake by bentonite. The effect of pH on U(VI) adsorption by bentonite can be attributed to the zero charge point of bentonite and the forms of U(VI) in solution. The variable charge on the surface of bentonite is generated by the proton transfer of surface functional groups, which changes with the pH value of the solution. The amphoteric silicon-oxygen bonds (Si-O) and aluminum-oxygen bonds (Al-O) on the mineral surface can undergo hydrolysis to generate surface functional groups (denoted as ≡SOH, typically including ≡SiOH, ≡AlOH, ≡MgOH, etc.), which serve as active sites for adsorption reactions. These hydrated hydroxyl groups can undergo protonation and deprotonation reactions that exist in different forms at different pH values. At a low pH, the bentonite surface can acquire a positive charge due to the protonation of the surface. In contrast, at a high pH, the surface of the bentonite may acquire a negative charge because of the deprotonation process. As shown in Figure 5d, the zero charge point of GMZ bentonite is around pH = 6.16, indicating that when the pH is lower than 6.16, the bentonite surface has a positive charge, and when the pH is higher than 6.16, it carries a negative charge [46,47]. The complex species of U(VI) in water varies with changes in the pH, thereby affecting the ability of bentonite to remove target pollutants. As shown in Figure S1, when the pH is less than 4, U(VI) predominantly exists as UO₂²⁺; within the pH range of 4.0 to 7.0, due to the increase in hydroxyl groups in the solution, UO₂²⁺ rapidly decreases and begins to form hydrates, primarily existing in the forms of (UO₂)₂(OH)₂²⁺, (UO₂)₃OH⁵⁺, and (UO₂)₃OH⁷⁺ in the solution; when the pH is >7.0, U(VI) primarily exists as UO₂(OH)³⁻ and (UO₂)₃OH⁷⁻ in the solution [48,49].

At a pH of 2.0–5.0, UO₂²⁺ competes with H⁺ for ion exchange sites, and electrostatic repulsion occurs between uranium ions and the positive charge (SOH₂⁺) on the bentonite surface, leading to a low U(VI) adsorption rate. Therefore, the slow increase in the adsorption of U(VI) is due to the gradual decrease in the competing H⁺ and the weakening of the electrostatic repulsion between UO₂²⁺ and the bentonite surface, where the adsorption outer layer complexation and ion exchange are dominant [50]. When the pH is between 5.0 and 7.0, the concentration of UO₂²⁺ in the solution gradually decreases, and the main adsorption reaction occurring with the bentonite surface is the outer-sphere complexation of UO₂²⁺. Thus, the reduction in the adsorption rate may be due to the outer-sphere complexation sites not being fully utilized, with adsorption primarily occurring through ion exchange reactions [42]. At pH > 7.0, the reduction in U(VI) adsorption may be due to the increased repulsive force between the negatively charged (SO⁻) on the bentonite surface and the negatively charged UO₂(OH)³⁻ and (UO₂)₃OH⁷⁻ [51,52].

3.2.3. Effect of Ionic Strength

The ionic strength significantly influences the U(VI) adsorption on bentonite [53]. The influence of the ionic strength on uranium adsorption was studied at four electrolyte concentrations (0, 0.001, 0.01, and 0.1 M NaCl aqueous solution). Figure 5b illustrates that the NaCl concentration greatly reduces the U(VI) adsorption on bentonite. When the ionic strength increases from 0 to 0.1 M NaCl, the overall adsorption rate decreases. Because the ionic strength gradually increases, some ion exchange sites become occupied, enhancing the competitive adsorption among ions. The effect of the ionic strength is more pronounced at a pH of 2.0–5.0, while it is weaker at a pH of 5.0–9.0. This may be because the adsorption mechanism at a pH of 2.0–5.0 is primarily ion exchange [54].

3.2.4. Effect of Natural Organic Matter

Natural organic matter has a strong binding capacity with radioactive nuclides, affecting the mobility, bioavailability, and toxicity of radionuclides in nature [54]. Figure 5c illustrates how varying concentrations of humic acid influence the adsorption rate of GMZ bentonite in a 0.1 M NaCl background electrolyte solution. The findings suggest that HA mildly improves U(VI) adsorption. When the pH value range is 2.0–5.0, 30 mg/L HA enhances uranium adsorption more than 10 mg/L HA. At pH 5.0–9.0, the positive effect of 10 mg/L HA on uranium adsorption exceeded that of 30 mg/L HA. Humic acid (HA) enhances U(VI) adsorption for two main reasons: first, HA has many functional groups, which can bond with bentonite, providing more complexing sites for U(VI) to form a bentonite–HA–U(VI) complex [42,55]; second, HA has a zero point of charge at around pH 2.0. When the pH is above 2.0, HA’s negative charge on bentonite reduces its positive charge, enhancing the electrostatic attraction with U(VI) [26]. As the pH increases, the enhancement of adsorption mediated by higher HA levels diminishes. Because with the increase in pH the surface charge of bentonite gradually shifts from positive to negative, due to electrostatic repulsion, the number of negatively charged HA adsorbed onto the bentonite decreases, leading to fewer complexing sites being formed. Excess HA will combine with uranium in the solution to form soluble HA–U(VI) complexes, which reduces the adsorption of uranium onto the bentonite. However, the amount of uranium remaining in the solution is still less than that bound in the bentonite–HA–U(VI) complexes, so the overall effect on adsorption is still facilitating [25,56,57].

3.2.5. Kinetics and Isotherm Analysis

To further explore the adsorption behavior of U(VI) by GMZ bentonite, both pseudo-first-order and pseudo-second-order kinetics are simulated. Figure 6a shows the nonlinear fitting curves of the two models, and

Table 3. Kinetic parameters of U(VI) adsorption on GMZ bentonite.

Kinetic Parameters	Parameter	Numeric
Pseudo-first-order	K1 (1/min)	1.5966
	qe (mg/g)	4.69
	R2	0.9713
Pseudo-second-order	K2 (g/(mg·min))	0.6414
	qe (mg/g)	4.83
	R2	0.9927
Particle diffusion model	K1 (mg/(kg·min0.5))	3.8068
	K2 (mg/(kg·min0.5))	0.4111
	R32	0.8797

lists the fitting parameters. The adsorption gradually reaches equilibrium over time, which is due to the gradual occupation of the adsorption site and the decreasing concentration of U(VI) in the solution. The correlation coefficient of the pseudo-second-order equation (R² = 0.9927) is notably greater than that of the pseudo-first-order equation (R² = 0.9713), indicating that the adsorption process is rapid and primarily chemical in nature. It can be inferred that the capture of U(VI) may involve a chemical adsorption reaction associated with ion exchange [58,59]. The intraparticle diffusion model and the corresponding constants are shown in Figure 6b and Table 3. The findings indicate that the entire adsorption process can be categorized into three stages: first, uranium in the solution is quickly adsorbed onto the surface of montmorillonite through diffusion; in the second stage, uranium moves from the surface into the interlayers of the montmorillonite; and the third stage represents the adsorption equilibrium stage, indicating that adsorption has reached saturation. In the adsorption equilibrium stage, the correlation coefficient (R² = 0.8797) is less than 0.9, suggesting that the main mechanism of action is chemical adsorption rather than diffusion [60]. This study utilized the widely used Langmuir, Freundlich, and Temkin adsorption isotherm models to perform nonlinear fitting of the experimental data on the U(VI) adsorption by GMZ bentonite, aiming to explore the process of target pollutant ions reaching equilibrium at the solid–liquid interface. The fitting results and related fitting parameters are shown in Figure 6c and

Table 4. Isotherm parameters of U(VI) adsorption on GMZ bentonite.

Isotherm Parameters	Parameter	Numeric
Langmuir	qm (mg/g)	11.7984
	KL (L/mg)	0.0186
	R2	0.9947
Freundlich	KF (mg·g−1) (Lm·g−1)1/n	0.4745
	1/n	0.6352
	R2	0.9948
Temkin	KT (L/mg)	0.1955
	BT (KJ/mol)	1.0110
	R2	0.9804

. Both the Langmuir and Freundlich models can achieve a good fit for adsorption, but the R² value of the Freundlich model (0.9948) is slightly higher than that of the Langmuir model (0.9947). Meanwhile, the Temkin model shows a poorer fit with an R² value of 0.9804. Therefore, the Freundlich model is better suited to describe the adsorption of U(VI) on the GMZ bentonite. The 1/n value (0.6352) falls within the range of 0 to 1 in the Freundlich model, which typically suggests that the adsorption process is relatively reversible and exhibits some heterogeneity. Meanwhile, in the Langmuir model, the K_L value (0.0186) also ranges from 0 to 1. These results suggest that GMZ bentonite exhibits good adsorption performance for the absorption of U(VI) [61,62].

3.2.6. Adsorption Mechanism

The characteristics of fresh and used bentonite are shown in Figure 7. A comparison of the TEM images before and after uranium adsorption (Figure 7a) indicates that most small pores and surface features are filled after adsorption, leading to a more compact internal structure. EDS analysis (Figure 7b) confirms the presence of uranium after adsorption. Moreover, the significant decrease in the sodium content suggests that ion exchange is involved in the adsorption process of GMZ bentonite [63].

Figure 7c displays the XRD spectra for both unaltered bentonite and bentonite that has been utilized. After interacting with uranyl ions, the interlayer spacing of the fresh bentonite changed, with the d (001) value increasing from 12.559 Å to 14.958 Å. This indicates that uranyl ions entered the interlayer and exchanged with the cations in the bentonite interlayers. In this experiment, the main interlayer cation of bentonite used is Na⁺, which has an ionic radius of 0.95 Å, smaller than that of U(IV) with an ionic radius of 0.97 Å. Moreover, U(IV) carries a greater charge than Na⁺. Therefore, the hydration capacity of U(IV) is larger than that of Na⁺, resulting in a larger hydrated cation and consequently an increased interlayer spacing after adsorption.

The FT-IR spectra of GMZ bentonite both prior to and following its reaction with U(VI) are shown in Figure 7d. The peak corresponding to the Al-O-H stretching vibration in the bentonite shifted from 3628 cm⁻¹ to 3622 cm⁻¹. Additionally, the peaks attributed to the H-OH group stretching vibrations at 3434 cm⁻¹ and 1642 cm⁻¹ shifted to 3439 cm⁻¹ and 1641 cm⁻¹, respectively. These peak shifts indicate that the Al-O-H and H-O-H groups in the bentonite may be involved in the adsorption process of uranium [37]. The bending vibration band of Si-O-Si moved from 1038 cm⁻¹ to 1036 cm⁻¹, suggesting an interaction between the metal ions and the bentonite [45]. Moreover, the bending vibration band of Al-O-Si at 694 cm⁻¹ shifted to 692 cm⁻¹, which can be attributed to the asymmetric stretching vibration of the uranyl ion and the stretching vibrations of oxygen ligands that are weakly bonded to uranium [64].

The XPS full-survey spectra of GMZ bentonite before and after uranium adsorption are shown in Figure 7e. After adsorption, the peak attributed to Na 1s at 1072.6 eV disappears, and the Ca 2p peak at 350.9 eV becomes weaker. This indicates that after adsorption, most of the sodium ions and some calcium ions were exchanged out from the montmorillonite interlayer through ion exchange. However, there are obvious peaks at 381.5 eV and 392.5 eV, which are attributed to the electron binding energies of U4f 7/2 and U4f 5/2, respectively, suggesting that hexavalent uranium is successfully adsorbed on bentonite [65,66]. Figure 7f shows the U 4f XPS spectrum of the bentonite after adsorption. The U4f_7/2 binding energy is influenced by the length of the U–O_eq bonds, which governs the redistribution of the electron density towards or away from U(VI) centers [67]. The two binding energy components of U 4f 7/2, at 380.4 eV and 381.8 eV, are attributed to UO₂²⁺. The lower binding energy is linked to the surface complexation of UO₂²⁺ with the bentonite, while the higher binding energy indicates ion exchange [68,69]. The analysis in Figure 7f indicates that 74.9% of UO₂²⁺ is adsorbed through ion exchange, while 25.1% is adsorbed through surface complexation [66,70,71].

In summary, the adsorption mechanism of uranium by GMZ bentonite (Figure 8) is mainly based on ion exchange, supplemented by surface complexation, which can be represented by the following reactions [26,53,57].

(1)

Surface complexation [72]

$$\equiv \mathrm{S}\mathrm{O}{\mathrm{H}}_2^{+}+\mathrm{U}{\mathrm{O}}_2^{2+}\to \equiv \mathrm{S}\mathrm{O}\mathrm{H}{\mathrm{U}\mathrm{O}}_2^{+}+{\mathrm{H}}^{+}$$

(3)

$$\equiv \mathrm{S}\mathrm{O}\mathrm{H}+\mathrm{U}{\mathrm{O}}_2^{2+}\to \equiv \mathrm{S}\mathrm{O}\mathrm{U}{\mathrm{O}}_2^{+}+{\mathrm{H}}^{+}$$

(4)

$$2\equiv \mathrm{S}\mathrm{O}{\mathrm{H}}_2^{+}+\mathrm{U}{\mathrm{O}}_2^{2+}\to \equiv (\mathrm{S}\mathrm{O}{{\mathrm{H})}_2\mathrm{U}\mathrm{O}}_2^{2+}+2{\mathrm{H}}^{+}$$

(5)

$$2\equiv \mathrm{S}\mathrm{O}\mathrm{H}+\mathrm{U}{\mathrm{O}}_2^{2+}\to (\equiv \mathrm{S}\mathrm{O}{)}_2\mathrm{U}{\mathrm{O}}_2+2{\mathrm{H}}^{+}$$

(6)

(2)

Ion exchange (taking sodium ion as an example)

$$\equiv \mathrm{S}\mathrm{O}\mathrm{N}\mathrm{a}+\mathrm{U}{\mathrm{O}}_2^{2+}\to \equiv \mathrm{S}\mathrm{O}{\mathrm{U}\mathrm{O}}_2^{+}+{\mathrm{N}\mathrm{a}}^{+}$$

(7)

$$2\equiv \mathrm{S}\mathrm{O}\mathrm{N}\mathrm{a}+\mathrm{U}{\mathrm{O}}_2^{2+}\to (\equiv \mathrm{S}\mathrm{O}{)}_2\mathrm{U}{\mathrm{O}}_2+{2\mathrm{N}\mathrm{a}}^{+}$$

(8)

(3)

The hydrolysis of U(VI) in solution

$$\mathrm{U}{\mathrm{O}}_2^{2+}+\mathrm{n}{\mathrm{H}}_2\mathrm{O}\leftrightarrow \left[\mathrm{U}{\mathrm{O}}_2\right({\mathrm{H}}_2\mathrm{O}{)}_{\mathrm{n}-\mathrm{m}}{]}^{2-\mathrm{m}}+\mathrm{m}{\mathrm{H}}^{+}$$

(9)

(4)

exchange with hydrolyzed species

$$\equiv \mathrm{S}\mathrm{O}\mathrm{H}+\mathrm{U}{\mathrm{O}}_2\left(\mathrm{O}\mathrm{H}{)}_{\mathrm{m}}\right({\mathrm{H}}_2\mathrm{O}{)}_{\mathrm{n}-\mathrm{m}}^{2-\mathrm{m}}\to \equiv \mathrm{S}\mathrm{O}\mathrm{U}{\mathrm{O}}_2(\mathrm{O}\mathrm{H}{)}_{\mathrm{m}}^{1-\mathrm{m}}+{\mathrm{H}}^{+}+(\mathrm{n}-\mathrm{m}\left)\right({\mathrm{H}}_2\mathrm{O})$$

(10)

4. Conclusions

This study examined the chemical composition and microstructure of GMZ bentonite before and after irradiation. The results indicate that irradiation induces slight structural changes in GMZ bentonite, including damage to the layered structure, loss of clarity at the boundaries, framework breakage, and dehydroxylation, which disrupt hydrogen bonds and lead to layer expansion. The batch adsorption technique was utilized to analyze the adsorption characteristics of bentonite concerning U(VI), revealing that U(VI) adsorption is greatly affected by the pH and ionic strength. The primary adsorption mechanism involves ion exchange between uranyl ions in the solution and ions in the interlayer of bentonite, while some uranyl ions are also removed from the aqueous phase through complexation with the surface functional groups of bentonites. Both gamma and electron beam irradiation reduce the uranium adsorption capacity of bentonite, with gamma irradiation having a more pronounced effect, likely due to structural damage. The adsorption isotherm for U(VI) on GMZ bentonite aligns with the Freundlich model, and the pseudo-second-order kinetic model provides a good fit, indicating that chemical adsorption is the dominant mechanism. These results are crucial for evaluating the safety performance of GMZ bentonite as a backfill material in high-level waste disposal sites.