An Experimental Investigation on the Barrier Performance of Complex-Modified Bentonite

Abstract

The barrier performance of containment liners against heavy metals and other contaminants is a critical element in ensuring environmental safety. However, the high concentration of multivalent cations in landfill leachate raises concerns about the effectiveness of conventional barriers (e.g., sodium bentonite). To address concerns regarding the high permeability and elevated heavy metal concentrations in effluents from sodium bentonite (Na-B) barriers, this study proposes the use of new complex-modified sorbent bentonite—specifically treated with disodium ethylenediaminetetraacetate (EDTA-2Na) and sodium tripolyphosphate (STPP). Batch adsorption and flexible-wall permeability tests in extreme synthetic leachate demonstrate that the complex-modified sodium bentonite not only maintains low permeability but also enhances contaminant adsorption capacity of barriers. When modified with 2% EDTA-2Na and 4% STPP (by mass), the maximum Zn(II) adsorption capacity of bentonite was measured at 43.22 and 48.22 μg/g, respectively. These values correspond to enhancements by a factor of 1.99 and 2.32 compared to the unmodified Na-B. Simultaneously, the hydraulic conductivity met the permeability requirements for engineering barrier systems (k < 1 × 10⁻⁷ cm/s) throughout the tested range of confining pressures. Microscopic analyses confirmed the successful incorporation of functional groups into bentonite by both EDTA-2Na and STPP. STPP-induced electrostatic repulsion, promoting ordered particle stacking and dense structure formation. EDTA-2Na physically filled pores to block ion migration pathways while electrochemically counteracting double-layer compression under high ionic strength. This effective strategy resolves the long-standing trade-off between permeability and adsorption capacity in conventional bentonite, providing a theoretical basis for designing barrier materials in complex contaminated sites.

1. Introduction

Landfill leachate acts as a primary vector for groundwater contamination, with its elevated concentrations of heavy metal constituents—such as Zn²⁺, Cu²⁺, and Mg²⁺—posing significant threats to ecosystems. International environmental organizations have established stringent standards for the impermeability of containment liner systems. A fundamental requirement within these standards is achieving a hydraulic conductivity lower than 1 × 10⁻⁷ cm/s to effectively prevent contaminant migration [1,2]. Sodium bentonite (Na-B) is widely adopted as a recommended material in landfill liner systems due to its exceptionally low hydraulic conductivity (ranging from 10⁻⁸ to 10⁻¹⁰ cm/s), pronounced swelling and self-sealing capacity upon hydration, and high cation exchange capacity [3,4,5]. However, the engineering performance of bentonite is highly susceptible to degradation in saline and acidic conditions. According to the literature, increased leachate concentration and decreased pH significantly compromise the barrier performance of bentonite. On one hand, elevated ionic strength compresses the bound water layer and diminishes the interlayer double-layer repulsion of montmorillonite, resulting in particle flocculation, expansion of pore channels, and consequently an increase in hydraulic conductivity [6,7]. On the other hand, multivalent cations such as Zn²⁺, Cu²⁺, and Mg²⁺ displace Na⁺ from exchange sites, promoting interlayer contraction, reducing cation exchange capacity, and consequently impairing selective adsorption of heavy metals [8,9,10,11]. Consequently, in acidic leachates derived from landfills, industrial effluents, and tailings—which are rich in multivalent heavy metal ions such as Zn²⁺ and Cu²⁺—bentonite is highly susceptible to performance degradation. Ion exchange-induced double-layer compression and reduced cation exchange capacity significantly increase the risk of failure in its hydraulic barrier function [12,13].

The hydraulic conductivity of bentonite liners is highly sensitive to both the concentration and type of cations present in the leachate. Studies indicate that leachates with high ionic strength and abundant divalent cations significantly deteriorate the hydraulic performance of liner systems. The chemical aggressiveness of leachate can be effectively characterized by key parameters such as ionic strength (I) and the relative abundance of multivalent cations (RMD) [14]. Empirical data demonstrate that bentonite can maintain excellent low hydraulic conductivity, in the order of 1 × 10⁻⁸ to 1 × 10⁻¹⁰ cm/s, in low-ionic-strength leachates such as those from typical municipal landfills or coal ash disposal facilities [15,16,17]. However, when exposed to high-ionic-strength leachates such as those from red mud or coal combustion products, the hydration and swelling capacity of bentonite is significantly suppressed. This suppression originates from cation-induced ion exchange, which promotes flocculation of bentonite particles, coarsens the pore structure, and consequently leads to an orders-of-magnitude increase in hydraulic conductivity [18,19]. Rainfall events alter soil hydro-chemical conditions, mobilizing heavy metals through infiltration-induced pH decrease and ion competition, while surface runoff transports and accumulates dissolved metals in depressions or permeability transition zones [20,21]. Therefore, enhancing the cation resistance of barrier materials has become a central challenge in the design of contaminant containment structures. Current modification strategies for enhancing bentonite performance primarily include activation methods (thermal, acid, and salt activation) and additive-based methods (inorganic, organic, and inorganic–organic composite modification). Among these, organic and polymer modifications have attracted significant attention in recent years. Polymer modification primarily enhances the chemical compatibility of bentonite in complex solutions through intercalation into the interlayer domains or formation of a protective coating on particle surfaces [22,23]. Commonly used polymeric modifiers include sodium polyacrylate, polyanionic cellulose, xanthan gum, sodium carboxymethyl cellulose, and propylene carbonate [24,25,26,27,28,29,30]. Although polymer modification significantly enhances the adsorption capacity of bentonite, its intricate synthesis procedures and high time cost considerably hinder large-scale engineering applicability. Moreover, existing research has predominantly focused on adsorption performance in single-pollutant systems, while investigation into the co-evolutionary behavior of adsorption and permeability in complex leachate environments remains scarce. Furthermore, the secondary pollution risks associated with degradation products of polymers have not yet been systematically evaluated, and their environmental compatibility under long-term service conditions requires urgent verification. In contrast, inorganic modification has become the preferred choice in practical engineering due to its straightforward methodology and well-established technology [31]. For natural sodium bentonite, the primary inorganic modification methods include sodium and inorganic phosphate treatment. Phosphate-based dispersants serve as effective stabilizers for natural clay by suppressing random edge-to-face aggregation of particles. This structural reorganization promotes a transition from disordered stacking to parallel alignment, consequently modifying the liquid limit and enhancing key engineering properties including strength, compressibility, and permeability [32]. Additionally, phosphate dispersants readily form complexes with heavy metal ions—such as Ca²⁺, Mg²⁺, Pb²⁺, Cd²⁺, and Zn²⁺—thereby synergistically enhancing the adsorption and immobilization capacity of clay toward target contaminants [33].

Zinc ions (Zn²⁺) represent a typical heavy metal pollutant widely present in industrial discharges and mining wastewater, with environmental concentration levels notably higher than those of most other heavy metal contaminants. Furthermore, heavy metal ions such as Zn²⁺ are key factors inducing health risks including anemia, renal dysfunction, and neurological damage [34,35,36,37,38,39]. Due to its high environmental abundance and potent toxicity, Zn²⁺ was selected as the model target pollutant in this study to evaluate the adsorption efficacy of anti-pollution barrier materials. Montmorillonite, the primary mineral in bentonite, possesses a typical layered structure that facilitates interlayer hydration. However, the restricted migration of ions into its interlayer domains constrains its natural adsorption capacity [40]. To enhance the adsorption capacity for heavy metal ions, physical and chemical modification techniques have been widely applied to tailor the structure of montmorillonite. Studies have shown that modification with cationic surfactants significantly improves the adsorption capacity of bentonite for Cu²⁺ and Zn²⁺ [41]. Furthermore, polymer modification techniques demonstrate superior adsorption characteristics. Studies by Rafiei et al. [42] and Chen et al. [43] have shown that polymer-functionalized bentonite can efficiently and simultaneously remove Cu²⁺ and Zn²⁺. It is noteworthy that most existing studies have focused primarily on adsorption behavior in single-ion systems. However, real landfill leachates exhibit complex and variable compositions. Whether coexisting anions and cations competitively inhibit or synergistically enhance the adsorption of target ions such as Zn²⁺ remains a critical scientific question requiring urgent clarification for engineering applications. Accordingly, this study compiled chemical composition data from landfill leachates and developed a representative formulation to simulate extreme leachate conditions. Using Zn²⁺ as the target adsorbate, the influence of multiple coexisting ions in complex synthetic leachate matrices on its adsorption behavior onto modified bentonite was thoroughly investigated.

Ethylenediaminetetraacetic acid disodium salt (EDTA-2Na) and sodium tripolyphosphate (STPP) are widely used in the industrial water treatment, pharmaceutical, and food industries owing to their cost-effectiveness and low secondary pollution risk [44,45,46,47]. These complexes form stable complexes with cations such as Ca²⁺, Mg²⁺, Fe²⁺, and Cu²⁺ through coordination sites composed of nitrogen and oxygen atoms [48,49,50]. In clay material modification, phosphate-based complexes such as STPP effectively induce a transition from disordered to parallel-aligned structures. This structural rearrangement significantly enhances the dispersibility of montmorillonite in bentonite, promoting the formation of a stable colloidal system. Simultaneously, these complexes readily form complexes with heavy metal ions including Zn²⁺, Cu²⁺, and Mg²⁺. Given this dual functionality—chelating competitive metal cations in leachate to reduce their ion exchange with montmorillonite while simultaneously enhancing the adsorption and immobilization of target heavy metals—integrating complexes into bentonite matrices is expected to synergistically optimize contaminant control efficiency. Accordingly, this study prepared modified materials by controlling the mass ratios of EDTA-2Na and STPP to sodium bentonite. Batch adsorption tests, flexible-wall permeability tests, and microstructural characterization were combined to systematically investigate the influence of complex dosage on Zn (II) adsorption capacity and hydraulic conductivity. Furthermore, the underlying mechanisms altering the structure and surface properties of bentonite were thoroughly revealed. This work aims to provide critical theoretical and practical guidance for developing advanced landfill liner systems with superior hydraulic sealing and heavy metal immobilization capabilities.

2. Materials

2.1. Preparation of Complex-Modified Bentonites

The natural sodium bentonite (Na-B) used in this study was supplied by Renzhong Industrial Co., Ltd, Shanghai, China. Key physical properties of the bentonite, measured in accordance with ASTM standards [51,52,53,54,55], are summarized in

Table 1. Basic physical properties of bentonite.

Specific Gravity (Gs)	Free Swell Index (mL/2g)	Free Swell Ratio (%)	Maximum Dry Density (g/cm3)	Natural Moisture Content (%)	Optimum Moisture Content (%)	Liquid Limit (%)	Plastic Limit (%)	Plasticity Index
2.70	31.02	365	1.583	24.65	12.52	255.77	32.45	223.29

. The complexes (analytical grade) for modifying bentonite (i.e., EDTA-2Na and STPP), were procured from Xilong Scientific Co., Ltd. (Guangzhou, China) and McLean Biochemical Technology Co., Ltd. (Shanghai, China), respectively, with their physicochemical properties listed in

Table 2. Basic physicochemical properties of amending complexes.

Complexes	Chemical Formula	Molecular Weight (g/mol)	Solubility (25 °C, g/L)
EDTA-2Na	C10H14N2Na2O8	336.21	100
STPP	Na5P3O10	367.86	140

Note: Solubility provided by the reagent manufacturer.

.

Referring to the wet processing method described by Norris et al. [56], modified dispersions were prepared by dissolving EDTA-2Na and STPP in 1 L of deionized water at mass fractions of 2%, 4%, 6%, 8%, and 10%, respectively [57]. Each dispersion was mixed with a predetermined amount of dried bentonite in a high-speed mixer and stirred for 1 h at room temperature. The resulting wet mixture was oven-dried at 105 °C for 8 h, then ground and sieved to ensure 100% passed through a 0.5 mm sieve. For clarity, the modified bentonite samples were designated as Na-B + n%X, where X represents the type of complex and n denotes the mass fraction.

2.2. Leachate

Representative extreme synthetic landfill leachate was prepared based on an analysis of leachate data from landfills. In landfill leachates, Al³⁺, Ca²⁺, Mg²⁺, K⁺, Na⁺, SO₄²⁻, and Cl⁻ account for a significant proportion of ionic species, with monovalent cations exhibiting higher average concentrations than multivalent cations. Based on the chemical indices influencing the hydraulic conductivity of Geosynthetic Clay Liners (GCLs)—ionic strength (I) and the relative abundance of cations (RMD), as defined by Kolstad et al. [17]—the relationship between I and RMD for landfill leachates was plotted, as shown in Figure 1. The measured values of I ranged from 4.51 to 1301.26 mM, while RMD varied between 0.0085 and 6.84 M¹ᐟ². The ionic strength (I) and the relative abundance of cations (RMD) can be calculated by Equation (1) and Equation (2), respectively.

$$I={\displaystyle \frac{1}{2}}{\sum }_{i=1}^n{c}_i{z}_i^2$$

(1)

$$RMD={\displaystyle \frac{M_M}{\sqrt{M_D}}}$$

(2)

where c_i is the concentration of ions in the leachate (mmol/L), and z_i is the corresponding valence of the ions; M_M is the total molar concentration of monovalent cations in the leachate (mol/L), and M_D is the total molar concentration of multivalent cations in the leachate (mol/L).

According to the theory proposed by Benson et al. [13], the worst-case scenario (extreme condition) for landfill leachate corresponds to the 90th percentile of ionic strength (I) and the 10th percentile of the relative abundance of multivalent cations (RMD) within the entire dataset. Extreme synthetic leachate can be formulated accordingly using these threshold values. Therefore, this study selected the 90th percentile value of I and the 10th percentile value of RMD from the entire dataset as the basis for formulating the extreme synthetic leachate. The resulting extreme synthetic leachate was prepared with an ionic strength (I) of 400 mM and a relative abundance of multivalent cations (RMD) of 0.08 M¹ᐟ² (The green dot in Figure 1). Additionally, the statistical analysis indicated that the pH of most landfill leachates in this study was approximately 5. Therefore, during the preparation of the extreme synthetic leachate, the pH was set to 5 and adjusted using hydrochloric acid and sodium hydroxide solutions. The reagents and their concentrations used in the extreme synthetic leachate formulation are summarized in

Table 3. Extreme synthetic leachate preparation scheme.

Leachate Properties	Na+	K+	Mg2+	Ca2+	Cu2+	Zn2+	Al3+	Cl-	SO42−	pH	I (mM)	RMD (M1/2)
Value	20.1	8.48	46.9	29.8	0.02	0.08	53.7	252.35	45.55	4.85	400	0.08

Note: The unit of ion concentration in this table is mM.

.

3. Methods

3.1. Batch Sorption Experiments

Batch adsorption tests were conducted in accordance with ASTM D4646-16 [58] to investigate the maximum adsorption capacity of modified bentonite by varying the Zn(II) concentration in the leachate (0.005–100 mg/L). The experimental design is outlined in

Table 4. Batch adsorption test programme.

Specimen Type	Complex Dosage (%)	Solid-to-Liquid Ratio (g/mL)	Zn(II) Solution Concentration (mg/L)
Na-B+EDTA-2Na	0, 2, 4, 6, 8, 10	1:100	0.005, 0.05, 0.5, 5, 50, 100
Na-B+STPP	0, 2, 4, 6, 8, 10	1:100	0.005, 0.05, 0.5, 5, 50, 100

Note: The “Zn(II) Solution Concentration” in the table refers to the concentration of Zn(II) in the synthetic leachate.

. Based on preliminary tests, a solid-to-liquid ratio of 1:100 was selected. For each test, 0.4 g of adsorbent and 40 mL of contaminated solution were placed in 50 mL polyethylene centrifuge tubes. The tubes were mounted on an automatic rotator and agitated at 50 rpm for 24 h at room temperature to ensure thorough interaction. Subsequently, the tubes were centrifuged at 5000 rpm for 10 min using a high-speed centrifuge. The supernatant was immediately filtered through a nylon membrane filter. The concentration of zinc ions in the filtrate was determined by 7900 ICP-MS generated by Agilent (Santa Clara, CA, USA). Two replicates were prepared for each treatment, and the results were averaged.

The equilibrium adsorption capacity q_e (mg/g) and equilibrium removal efficiency η were calculated using Equations (3) and (4), respectively:

$$q_e={\displaystyle \frac{(C_0-C_e)V}{m}}$$

(3)

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

(4)

where C₀: Initial concentration of the adsorbate (mg/L); C_e: Equilibrium concentration of the adsorbate (mg/L); m: Mass of the adsorbent (g); V: Volume of the solution (L).

3.2. Hydraulic Conductivity Tests

Flexible-wall permeability tests were performed following ASTM D5084-16a [59] to evaluate the hydraulic permeability characteristics of modified bentonite under varying confining pressures using deionized water. The experimental setup is detailed in

Table 5. Permeability test program.

Specimen Type	Complexes Dosage (%)	Permeant Solution	Confining Pressure (kPa)
Na-B+EDTA-2Na Na-B+STPP	0, 2, 4, 6, 8, 10	Extreme synthetic leachate	20 50 100 200

. As reported by Jo et al. [60], physical and chemical equilibrium between the modified bentonite and permeant was confirmed when three criteria were met: (1) the outflow-to-inflow ratio (Q_out/Q_in) remained within 1.00±0.25 during continuous measurement, (2) the average hydraulic conductivity (k) values stabilized within ±25% range, and (3) the fluid pore volume (PVF, calculated via Equation (5)) reached a minimum value of 2.

During the permeability test, the thickness of the soil specimen remains stable. After the test, the thickness and mass of the bentonite specimen are measured immediately upon removal. From these measurements, the density and volume of the specimen under saturated conditions can be calculated. Subsequently, the water content of a small portion of the soil is determined. Using this, along with the known specific gravity of soil particles (2.70) and other parameters, the porosity of the soil during the permeability test is comprehensively calculated based on three key indicators: saturated density, water content, and specific gravity. The total pore volume of the specimen is then obtained by multiplying the total volume by the porosity. Finally, by accumulating the volume of effluent and dividing it by the total pore volume, the cumulative value of PVF is determined.

$$PVF={\displaystyle \frac{Q_{out}}{V_w}}$$

(5)

where Q_out: Cumulative volume of effluent (mL); V_w: Total pore volume of the soil specimen (cm³).

Darcy’s Law (Equation (6)) serves as the fundamental principle for determining the hydraulic conductivity (k) of soil specimens, describing the relationship between flow velocity and hydraulic gradient (i) in saturated soils:

$$Q=kiAt$$

(6)

where Q: Permeation flux over time t (mL); k: Hydraulic conductivity (cm/s); i: Hydraulic gradient (dimensionless); A: Cross-sectional area of the specimen (cm²); t: Permeation duration (s).

The water head at the inflow and outflow surfaces of the specimen differs by the thickness of the specimen. This head is extremely small compared to the pressure acting on the specimen and can be neglected. Therefore, the hydraulic gradient (i) is directly related to the applied pressure difference (p) across the specimen:

$$i={\displaystyle \frac{p}{{\gamma }_{w}H}}$$

(7)

where p: Pressure difference between upstream and downstream (kPa); H: Thickness of the specimen (m); γ_w: Unit weight of water (9.8 kN/m³).

Substituting Equation (6) into Equation (5) yields the final expression for k:

$$k={\displaystyle \frac{QH{\gamma }_w}{pAt}}$$

(8)

The hydraulic conductivity of modified bentonite under extreme synthetic leachate exposure was determined using a flexible-wall permeameter (Figure 2). The testing procedure was conducted as follows: 150 g of dried and sieved (0.5 mm) soil (with a dry density of 1.583 g/cm³) was compacted into a flexible membrane (maintaining a soil sample thickness of 1 cm). Filter paper, a porous stone, and a top plate were sequentially placed and sealed. A confining pressure of 20 kPa was applied while venting air to ensure proper membrane-specimen contact. Subsequently, deionized water was injected into the soil specimen under an upstream pressure of 5 kPa to remove residual gas until bubble-free effluent was observed. All specimens were saturated under a back pressure of 5 kPa, followed by 48 h hydration using deionized water. The extreme synthetic leachate was then introduced and set the osmotic pressure to 10 kPa. (10 kPa upstream). Hydraulic conductivity was calculated by monitoring liquid level changes in the measuring tube. The test was repeated at varying confining pressures after stabilization was attained.

3.3. BET Specific Surface Area Testing

A larger specific surface area provides abundant active sites for processes such as ion exchange, physical adsorption, and surface complexation, thereby directly determining the adsorption capacity of bentonite. Therefore, Brunauer–Emmett–Teller (BET) nitrogen analysis was performed on the prepared samples using a surface area and porosity analyzer. Particles with a size of 3–5 mm were selected from the modified bentonite samples and pre-treated in a constant-temperature drying oven at 50 °C for 24 h. After cooling to room temperature, the specimens were transferred to a desiccator for storage. A sample weighing 0.7–1.0 g was then precisely measured using an analytical balance with 0.1 mg precision and loaded into a pre-cleaned sample tube. Degassing was performed at 150 °C for 2 h in accordance with ASTM D3663-03 [61] to completely remove physically adsorbed water and volatile impurities. The specific surface area of soil samples was tested using the ASAP 2460 produced by BeiShiDe Instrument-S&T. (Bejning, China) Co., Ltd., strictly following ASTM D1993-03 [62], through nitrogen adsorption–desorption isotherms measured at 77 K under liquid nitrogen temperature.

3.4. Fourier Transform Infrared Spectroscopy (FTIR) Testing

The successful intercalation of amendments into the interlayer domains of bentonite is a critical factor determining the effectiveness of modification. Therefore, the modified soil samples were characterized using an FTIR-850 Fourier transform infrared spectrometer manufactured by Gangdong Sci. & Tech. Development Co., Ltd., Tianjin, China. The successful grafting of modifier molecules into the bentonite matrix was identified by analyzing changes in the vibrational peaks of characteristic functional groups. Following ASTM D2216-19 [54], specimens from the Na-B reference group and optimal modified groups (Na-B+2%E, Na-B+4%S) were selected along with the adsorbents (EDTA-2Na, STPP) and dried at 60 °C for 24 h in a precision forced-air drying oven. The samples were then ground and passed through a 200-mesh standard sieve (aperture 0.075 mm). The collected powder was further dried in a vacuum oven at 105 °C for 10 h to ensure a moisture content below 0.5%. According to ASTM E168-16 [63], an appropriate amount of dried sample was mixed with KBr at a ratio ranging from 1:100 to 1:200 and thoroughly homogenized in an agate mortar. The mixture was compressed into a thin pellet using a hydraulic press and subsequently analyzed by the instrument.

3.5. Zeta Potential Tests

Zeta potential serves as a critical parameter for evaluating the stability of adsorbents in aqueous solutions and elucidating adsorption mechanisms, with lower negative potentials indicating stronger electrostatic attraction to metal cations. The zeta potential of soil samples was measured using a Brookhaven 90Plus PALS zeta potential analyzer. Na-bentonite (Na-B), EDTA-2Na-modified bentonite, and STPP-modified bentonite were oven-dried at 60 °C for 24 h. A sieved sample was transferred to a 500 mL glass beaker, mixed with approximately 250 mL of deionized water, and thoroughly stirred using a glass rod until a homogeneous suspension without visible particles or suspended matter was achieved. An appropriate aliquot of the suspension was then analyzed using the instrument to determine the zeta potential (Figure 3).

4. Results

4.1. Adsorption Characteristic

Figure 4 shows the relationship between the equilibrium adsorption capacity (qₑ) for Zn(II) and the equilibrium concentration (Cₑ) for various tested materials. As illustrated, variations in Cₑ had minimal influence on the adsorption capacity of unmodified bentonite, which exhibited a maximum qₑ of only 21.71 μg/g—significantly lower than that of the modified samples. Both STPP and EDTA-2Na modified materials showed increasing adsorption of Zn(II) with rising equilibrium concentration, exhibiting a non-linear growth trend. The maximum adsorption capacities reached 117.50 μg/g and 98.40 μg/g for the STPP and EDTA-2Na modified bentonites, respectively.

The qₑ–Cₑ relationship curves for bentonite modified with varying dosages of EDTA-2Na and STPP are presented in Figure 5. The adsorption curves exhibited a steep slope in the low equilibrium concentration region, which gradually decreased and eventually stabilized as concentration increased. This behavior is consistent with the “L-type” adsorption isotherm proposed by Giles et al. (1974) [64], indicating moderate affinity of the modified bentonite for Zn(II). At an initial Zn(II) concentration of 100 mg/L, the 2% EDTA-2Na modified bentonite achieved an optimal qₑ value of 43.22 μg/g, representing a 1.99-fold increase compared to the unmodified sample. Similarly, the 4% STPP modified bentonite reached a peak adsorption capacity of 48.22 μg/g, equivalent to 2.22 times that of the unmodified material.

The relationship between the removal efficiency (η) of Zn(II) and its initial concentration (C₀) for the modified bentonites is illustrated in Figure 6. At C₀ = 0.05 mg/L, the 2% EDTA-2Na modified bentonite achieved the highest removal efficiency of 42.10%. Under the same conditions, the 4% STPP modified bentonite reached a peak removal efficiency of 43.53%. Notably, when C₀ increased to 5 mg/L, all samples exhibited a sharp decline in removal efficiency, with values converging to nearly zero. This phenomenon is attributed to the gradual saturation of adsorption sites as the initial concentration rises.

To gain deeper insight into the adsorption behavior, the experimental data were fitted using Langmuir and Freundlich isotherm models (Figure 7 and Figure 8). The calculation formulas show in Equations (9) and (10) respectively. The Langmuir model [as shown by Equation (9)] exhibited a high degree of agreement with the experimental data, indicating that the adsorption of Zn(II) onto the modified bentonite follows the assumption of monolayer adsorption. Specifically, the maximum adsorption capacities (qₘ) of the 2% EDTA-2Na and 4% STPP modified sodium bentonite reached 42.76 μg/g and 46.64 μg/g, respectively, which are 2.13 and 2.32 times that of the unmodified bentonite (

Table 6. Parameters of isothermal adsorption models for Zn(II) on modified bentonite in extreme synthetic leachate.

	Langmuir Model			Freundlich Model
	KL (L/mg)	qm (mg/g)	R2	KF ((μg/g)/(mg/L))	n	R2
Na-B	0.2257	20.0966	0.9482	5.0932	3.2286	0.9647
Na-B+2%E	0.4008	42.7619	0.9940	13.6691	3.7901	0.9184
Na-B+4%E	0.3495	39.1588	0.9919	12.0715	3.6737	0.9211
Na-B+6%E	0.3385	35.6036	0.9958	11.5239	3.7796	0.8962
Na-B+8%E	0.3775	32.4175	0.9949	11.1437	3.9362	0.8688
Na-B+10%E	0.3443	30.1867	0.9908	10.3702	3.8775	0.8634
Na-B+2%S	0.4254	32.0048	0.9829	11.0546	3.9454	0.8823
Na-B+4%S	0.4349	46.6353	0.9920	16.1059	3.9538	0.8822
Na-B+6%S	0.3565	46.0409	0.9985	14.7489	3.7398	0.9060
Na-B+8%S	0.4919	35.8673	0.9956	13.2655	4.1727	0.8397
Na-B+10%S	0.2655	34.2763	0.9936	10.5678	3.6330	0.9041

Note: Freundlich parameter K_F depends on the units of the horizontal axes of the adsorption isotherm.

).

$$q_e=q_m{\displaystyle \frac{K_L{C}_e}{1-K_L{C}_e}}$$

(9)

where q_m is the maximum adsorption capacity of the adsorbate on the adsorbent surface (mg/g); K_L is the Langmuir constant related to the binding affinity, and C_e: Equilibrium concentration of the adsorbate (mg/L).

The Freundlich model [as shown by Equation (10)] fitting results showed that the heterogeneity coefficient (n) for all modified samples was greater than 1, further confirming that the adsorption process is favorable.

$$q_e=K_F{C}_e^{{\displaystyle \frac{1}{n}}}$$

(10)

where K_F is the Freundlich constant; n is the heterogeneity factor, and C_e: Equilibrium concentration of the adsorbate (mg/L).

4.2. Hydraulic Conductivity

During the deionized water permeation stage under a confining pressure of 20 kPa, the relationship between the effluent-to-influent volume ratio (Q_out/Q_in) and the pore volumes of flow (PVF) for modified bentonite is presented in Figure 9. As permeation progressed, most Q_out/Q_in values remained within the range of 1.00 ± 0.25. Concurrently, the hydraulic conductivity (k) of all samples gradually stabilized with increasing PVF (Figure 10), meeting the termination criteria specified in the standard test methods by Li et al. [65] and Norris et al. [56]. These observations confirm the establishment of hydraulic equilibrium and verify the reliability of the permeability coefficient measurements. Moreover, in engineering applications, a hydraulic conductivity value below 1 × 10⁻⁷ cm/s for barrier materials is considered acceptable [1,2].

The temporal evolution of hydraulic conductivity (k) for EDTA-2Na- and STPP-modified bentonite under varying confining pressures during permeation with extreme synthetic leachate is presented in Figure 11. During the initial stage of testing, the hydraulic conductivity of the EDTA-2Na-modified bentonite system exhibited a phased increase. This phenomenon originates from insufficient swelling development during the hydration process. When the extreme synthetic leachate permeates the system, the multivalent cations it carries (e.g., Ca²⁺, Mg²⁺) undergo cation exchange with sodium ions in the interlayer domains of the bentonite. This exchange process not only compresses the interlayer double layer but also causes charge imbalance within the montmorillonite lattice. As a result, the hydration and swelling capacity are suppressed, leading to the formation of a metastable pore network among the particles. Ultimately, these changes manifest as a non-linear increase in hydraulic conductivity over the duration of leachate interaction. Notably, the STPP-modified system exhibited no abnormal fluctuations in hydraulic conductivity under identical testing conditions. This stability is attributed to the electrostatic repulsion generated by the polyphosphate groups of STPP, which optimizes the spatial arrangement of soil particles and promotes the formation of a denser microstructure. Although multivalent cations in the leachate still initiate sodium ion exchange, the dispersive effect of STPP effectively compensates for the increased porosity resulting from suppressed swelling, thereby maintaining permeability stability.

Notably, the hydraulic conductivity (k) of EDTA-2Na-modified bentonite exhibited an increasing trend with higher dosage levels. However, even at a 2% dosage, the sample achieved a k value as low as 1.09 × 10⁻⁸ cm/s under a confining pressure of 200 kPa. In contrast, the STPP-modified system consistently enhanced the impermeability compared to unmodified bentonite. Its hydraulic conductivity (k) initially decreased and then increased with higher dosage. The 4% STPP-modified sample achieved the lowest k value of 5.74 × 10⁻⁹ cm/s under a confining pressure of 200 kPa. An increase in confining pressure reduced the hydraulic conductivity (k) of all modified bentonites. This effect is attributed to the synergistic action of interparticle pore compression and the constriction of flow paths under mechanical loading.

Figure 12 presents the correspondence between the equilibrium hydraulic conductivity (k) and the complexes dosage. Under identical confining pressure conditions, the STPP-modified bentonite exhibited significantly lower hydraulic conductivity (k) values compared to the EDTA-2Na-modified system. This enhanced performance is attributed to the superior dispersive properties of STPP. During hydration, STPP facilitates dense particle packing [28,66], effectively reducing pore volume and fluid transport capacity. Bentonite specimens prepared with both modifying complexes met the impermeability requirements for engineering barrier systems (k < 1 × 10⁻⁷ cm/s) throughout the tested range of confining pressures. The hydraulic conductivity (k) of EDTA-2Na-modified bentonite increased with higher additive dosage. This trend may be attributed primarily to the reduction in bentonite mass per unit area, which is a key factor influencing permeability [16,17]. Beyond the optimal dosage ratio, the bentonite mass per unit area decreases with increasing modifier content, thereby resulting in elevated hydraulic conductivity.

4.3. Morphological Analysis

4.3.1. Specific Surface Area

The BET specific surface area measurements for sodium bentonite (Na-B) along with EDTA-2Na- and STPP-modified bentonites are presented in Figure 13. The data indicate that the unmodified bentonite (Na-B) exhibits the largest specific surface area (39.38 m²/g). As the dosage of modifying complexes increases, the specific surface area of EDTA-2Na-modified bentonite decreases gradually from 12.15 m²/g to 5.10 m²/g. In contrast, the STPP-modified bentonite shows a more pronounced reduction, declining from 14.25 m²/g to 3.89 m²/g. Combined with the batch adsorption results, a key mechanism is revealed: The adsorption of Zn(II) onto modified bentonite is governed not primarily by physical adsorption mechanisms, but rather by a chemisorption process dominated by specific interactions between the grafted functional groups of the modifying agents and the metal ions. This conclusion is consistent with the characteristics of the adsorption isotherm models observed in the batch tests.

4.3.2. Fourier-Transform Infrared Spectroscopy (FTIR)

FTIR spectroscopic analysis in Figure 14 indicates that the EDTA-2Na-modified bentonite (Na-B+2%E) exhibits hydroxyl stretching vibrations v(OH) of Al–OH bonds at 3623 cm⁻¹ and 3439 cm⁻¹. The characteristic peak at 1030 cm⁻¹ was significantly sharpened due to vibrational coupling between v(COO⁻) and v(Si–O–Si), confirming the successful grafting of carboxylic functional groups into the interlayer domains of bentonite. The STPP-modified bentonite (Na-B+4%S) displayed Al–OH vibration peaks at 3617 cm⁻¹ and 3430 cm⁻¹. The peak at 1027 cm⁻¹ exhibited a blue shift (Δν = +3 cm⁻¹) along with broadening and sharpening, resulting from coupling between v(PO₄³⁻) and v(Si–O–Si). This confirms the successful incorporation of phosphate functional groups into the clay structure. The modified bentonite not only retained the inherent functional groups of the raw material but also incorporated highly electronegative groups—such as carboxyl (–COOH), hydroxyl (–OH), and phosphate (PO₄³⁻)—through the complexes. This significantly enhanced the particle surface zeta potential (Figure 15). These findings provide a microscopic explanation for the previously observed improvement in Zn(II) adsorption performance (Figure 7 and Figure 8).

4.3.3. Mechanisms of Action of EDTA and STPP

The findings of this study confirm that both EDTA and STPP have been successfully incorporated into the bentonite matrix, establishing a synergistic system of action. The adsorption of heavy metal ions onto natural bentonite is intrinsically a spontaneous physicochemical synergistic process. Bentonite captures heavy metal ions through physical adsorption, which relies on its extensive specific surface area. This process is reversible, allowing adsorbed ions to detach under appropriate conditions. In contrast, chemisorption stabilizes metal ions via ion-exchange reactions. These reactions occur between the metal ions and either structural hydroxyl groups or exchangeable interlayer cations (e.g., Na⁺), firmly immobilizing the heavy metals within the clay structure [67,68]. Although the modified soil specimens exhibited a reduced specific surface area, the introduced modifiers could form stable complexes with metal ions through their specific functional groups. This enhanced the chemisorption of heavy metal ions and substantially improved adsorption stability. Furthermore, the increased surface electronegativity of the modified soil promoted the enrichment of heavy metal ions, significantly raising the immobilization efficiency. Under extreme leachate conditions, the high I and relatively low RMD lead to an increase in the hydraulic conductivity of natural bentonite [69,70]. The EDTA and STPP in the modified soil specimens react with metal ions in the solution, preventing this situation. Furthermore, due to the unique mechanism of STPP, interlayer stability is effectively enhanced, promoting the formation of a denser microstructure in the soil specimens. As a result, the modified bentonite maintains a lower hydraulic conductivity than natural bentonite under the same extreme conditions, significantly improving the chemical compatibility and long-term stability of the barrier system. Last but not least, both EDTA and STPP are well-established industrial materials. Their ease of application, along with their widespread availability, low cost, and environmental friendliness, enables their rapid adoption in engineering applications.

5. Conclusions

Landfill leachate contains various pollutants, including heavy metal ions. Certain hazardous contaminants can pose significant environmental threats even at low concentrations after penetrating barrier systems. Therefore, the barrier performance of bentonite liners is critical to the long-term success of landfill projects. Recognizing the beneficial effects of enhancing the heavy metal retention capacity of barrier materials, this study presents an attempt to develop an adsorptive barrier by modifying conventional bentonite with complexes (i.e., EDTA-2Na and STPP). Using a representative extreme synthetic leachate and complexing agent-modified bentonite, batch adsorption tests and large-scale flexible-wall permeability tests were conducted using Zn(II) as a tracer contaminant. These experiments systematically evaluated the heavy metal adsorption capacity and hydraulic performance of the new barrier material. Furthermore, microstructural analyses including specific surface area measurements, FTIR, and zeta potential characterization were employed to investigate the mechanisms of action of EDTA-2Na and STPP in bentonite improvement. The following main conclusions can be drawn:

(1)

In Zn(II)-containing extreme synthetic leachate, the maximum Zn(II) adsorption capacities of 2% EDTA-2Na and 4% STPP modified bentonites reached 43.22 μg/g and 48.22 μg/g, respectively. These values represent a 1.99–2.32-fold enhancement compared to unmodified bentonite. Fourier transform infrared (FTIR) spectroscopy demonstrated that the enhanced adsorption originated from functional groups grafted by complexing agents. These included carboxyl (–COOH) and phosphate (PO₄³⁻) groups. Simultaneously, the significant negative shift in zeta potential demonstrates the enhanced capability of complexing agents to attract metal cations in bentonite.

(2)

BET specific surface area measurements demonstrated that the specific surface area of both EDTA- and STPP-modified sodium bentonites decreased with in-creasing additive content. The measured values decreased from 12.15 to 5.10 m²/g for EDTA-modified bentonite and from 14.25 to 3.89 m²/g for STPP-modified bentonite. This phenomenon results from the filling of montmorillonite interlayer domains and interparticle pores by complex molecules, which effectively blocks ion migration channels. These results confirm that the adsorption process is not governed by the material’s specific surface area.

(3)

Under a confining pressure of 200 kPa, the 4% STPP-modified sample exhibited an equilibrium hydraulic conductivity (k) as low as 5.74 × 10⁻⁹ cm·s⁻¹. This value represents a reduction of nearly one order of magnitude compared to unmodified bentonite. The hydraulic conductivity (k) of EDTA-2Na-modified bentonite increased with higher additive dosage, reaching its minimum value at a 2% dosage. This trend is attributed to the reduction in bentonite mass per unit area, resulting in the coarsening of the pore network. In contrast, STPP utilizes electrostatic repulsion from its phosphate groups to promote the transition from disordered aggregation to parallel-aligned stacking of clay particles. Notably, all modified specimens exhibited hydraulic conductivity (k) values significantly lower than the international impermeability standard (1 × 10⁻⁷ cm·s⁻¹) in extreme synthetic leachate. This confirms their suitability for engineering applications.