Robust Adsorption of Pb(II) and Cd(II) by GLDA-Intercalated ZnAl-LDH: Structural Engineering, Mechanistic Insights, and Environmental Applications

Abstract

The rapid pace of industrialization has led to widespread heavy metal contamination in water and soil, highlighting the need for efficient remediation strategies. Among various approaches, adsorption has proven to be an effective method for treating contaminated environments. Layered double hydroxide (LDH) is frequently used in such applications. However, its adsorption efficiency remains limited. In this study, glutamic acid diacetate tetrasodium salt (GLDA) was incorporated into ZnAl LDH via a straightforward co-precipitation and ion exchange method, yielding a modified material, GLDA-LDH, which was subsequently applied for the adsorption of Pb(II) and Cd(II). Adsorption behavior was investigated through kinetic and isothermal models, with results indicating that the process followed pseudo-second-order kinetics and fit well with the Langmuir isotherm, suggesting chemisorption onto monolayer surface. The maximum adsorption capacities reached 219.2 mg/g for Pb(II) and 121.9 mg/g for Cd(II). Furthermore, GLDA-LDH exhibited a strong retention capability for metal ions with minimal desorption and remained effective in the presence of hard water and contaminated soils. XPS analysis revealed distinct interaction mechanisms; surface oxygen and carboxyl groups played a key role in Pb(II) adsorption, whereas nitrogen coordination was involved in Cd(II) uptake. These results point to the potential of GLDA-LDH as a reliable material for addressing heavy metal pollution and provide insights into the design of enhanced LDH-based adsorbents.

1. Introduction

Heavy metal contamination in the environment has become a critical concern due to the persistence, toxicity, and bioaccumulative nature of metal ions such as lead (Pb(II)) and cadmium (Cd(II)). Once released into ecosystems, heavy metal ions pose severe threats to human health, aquatic life, and soil quality [1]. Lead is a well-documented neurotoxin, posing significant risks to children’s development. Cadmium exposure, meanwhile, is associated with renal dysfunction and long-term carcinogenic effects. Therefore, the development of effective methods to remove or immobilize heavy metal ions is a critical environmental priority. Nowadays, various remediation strategies have been investigated in wastewater treatment, like chemical precipitation [2], ion exchange [3], membrane filtration [4], electrochemical treatments [5], and adsorption. Among these, adsoption is widely recognized for its simplicity, cost-effectiveness, and efficiency, especially at low metal concentrations [6]. Numerous materials have been studied as adsorbents, including activated carbon [7], zeolites [8], biochar [9], synthetic nanomaterials [10], and metal-organic frameworks (MOFs) [11]. Despite their advancements, challenges such as poor selectivity, high production costs, and environmental risks are associated with some types of adsorbents. Recently, layered double hydroxide (LDH), also known as hydrotalcite-like compounds, have gained significant attention as promising candidates due to their unique structural characteristics, high anion exchange capacities, and tunable composition [12,13,14]. These features make them attractive for wastewater and soil remediation applications.

LDH is a two-dimensional nanostructured material composed of positively charged brucite-like layers and intercalated anions. The general chemical formula of LDH could be defined as ${\left[{\mathrm{M}}_{1-\mathrm{x}}^{2+}{\mathrm{M}}_{\mathrm{x}}^{3+}{\left(\mathrm{OH}\right)}_2\right]}^{\mathrm{x}+}{\mathrm{A}}_{\mathrm{x}/\mathrm{n}}^{\mathrm{n}-}·\mathrm{m}{\mathrm{H}}_2\mathrm{O}$, where M(II) and M(III) are divalent and trivalent metal cations, respectively, and Aⁿ⁻ represents the interlayer anions. This brucite-like layered structure enables LDH to achieve ion exchange and surface complexation, effectively removing contaminants. Modification strategies for LDH include anion intercalation, surface functionalization, and calcination-rehydration, as well as composites with biochar [15,16,17,18], graphitic carbon [19,20,21], or magnetic nanoparticles [22,23]. For anion intercalation, Li et al. developed a biochar-LDH composite using acetate intercalation, achieving maximum adsorption capacities of 402.70 mg/g (Pb(II)), 68.50 mg/g (Cu(II)), and 21.68 mg/g (As(V)) [15]. Su et al. developed LDHs intercalated by 2,4,6-trimercapto-s-triazine for removing Cu(II) and Pb(II) from contaminated soil, and the adsorption capacity reached 102.8 mg/g for Cu(II) and 105.8 mg/g for Pb(II) [24]. Bi et al. prepared Mg-Al LDH-intercalated mercaptocarboxylic acid, which exhibited 250.30, 122.101, and 105.33 mg/g maximal adsorption capacities for Hg(II), Pb(II), and Cu(II) at 0.2 g/L and a pH of 5.5 [25]. These findings demonstrate that LDH modification, especially anion intercalation, plays a crucial role in enhancing the adsorption performance of LDH.

Tetrasodium glutamate diacetate (GLDA), synthesized from bio-based L-glutamic acid, is an environmentally friendly chelating agent. It exhibits excellent metal ion binding capabilities, high solubility in water, and robust stability across a broad pH spectrum. Owing to its strong ability to complex and remove metal contaminants, GLDA serves as a promising and sustainable alternative to conventional chelators such as EDTA [26]. In this work, a green adsorbent (GLDA-LDH) was successfully prepared by intercalating GLDA into ZnAl-LDH. The synthesis routes and schematic diagram are shown in Figure 1. The structure of modified LHD was confirmed by X-ray diffraction (XRD), Fourier transform infrared spectroscopy (FT-IR), and scanning electron microscopy (SEM) analyses. Batch adsorption experiments were conducted to investigate the effects of the adsorption time, temperature, initial pH, and coexisting ions on the adsorption capacity of Pb(II) and Cd(II) using GLDA-LDH. Additionally, the adsorption behavior of GLDA-LDH in both hard water and contaminated soil environments was simulated. This work reveals the application potential of GLDA-LDH for heavy metal uptake.

2. Materials and Methods

2.1. Materials

GLDA was bought from Shanghai Yuanye Bio-Technology Co., Ltd., Shanghai, China. Zinc nitrate hexahydrate (Zn(NO₃)₂·6H₂O) was purchased from Guangdong Guangshi Regent Science and Technology Co., Ltd., Zhaoqing, China. Nitric acid (HNO₃), sodium hydroxide (NaOH), and aluminum nitrate nonahydrate (Al(NO₃)₃·9H₂O) were bought from Shanghai Macklin Biochemical Co., Ltd., Shanghai, China. Cadmium nitrate (Cd(NO₃)₂) was acquired from Sun Chemical Technology Shanghai Co., Ltd., Shanghai, China and lead nitrate (Pb(NO₃)₂) was bought from Tianjin Damao Chemical Reagent Factory, Tianjin, China.

2.2. Synthesis of GLDA-LDH

Initially, a salt solution was prepared with a Zn/Al molar ratio of 2:1, with NaOH serving as the precipitating agent. The pH was adjusted to 9, and the solution was continually stirred at 30 °C for 2 h. Subsequently, the temperature was maintained at 100 °C for 24 h. Upon the completion of crystallization, the zinc-aluminum-nitrate-type LDH precursor (NO₃-LDH) was obtained through filtration and drying.

The GLDA-LDH was prepared using an ion exchange method. Specifically, a dispersion of NO₃-LDH in distilled water was prepared, and the process temperature was controlled at 40 °C under high-speed stirring. The pH was adjusted to 7 during the dropwise addition of GLDA. Subsequently, stirring was continued to facilitate the ion exchange reaction. Afterward, the mixture was heated at 100 °C for 24 h. The GLDA-LDH was collected by filtering the mixture and drying at 80 °C.

2.3. Kinetics

Five 250 mL conical flasks were prepared, each containing 100 mL of a 100 mg/L solution of lead ions (Pb(II)) and cadmium ions (Cd(II)). The flasks were incubated in a thermostatic shaker at 25 °C. Subsequently, 100 mg of GLDA-LDH adsorbent was added to each flask. Adsorption was carried out under oscillation for varying durations. After oscillation, the supernatant was collected and diluted. Finally, the concentrations of Pb(II) and Cd(II) were measured using an atomic absorption spectrophotometer.

2.4. Factors Affecting Adsorption

The pH range for this experiment was set to 2–6. Four 250 mL conical flasks were prepared, each filled with 100 mL of 100 mg/L Pb(II) or Cd(II) solution. The pH was adjusted using nitric acid followed by incubation in a thermostatic shaker at 25 °C. Subsequently, 100 mg of GLDA-LDH was added to each flask, and the mixtures were oscillated for 4–6 h. After oscillation, the supernatant was collected, diluted, and analyzed for its ion concentrations using an atomic absorption spectrophotometer.

Additionally, five 250 mL conical flasks containing 50 mL of 100 mg/L Pb(II) or Cd(II) ion solution. The flasks were incubated in a thermostatic shaker at 25 °C. Graded doses (0.01 g/L, 0.015 g/L, 0.02 g/L, 0.025 g/L, 0.03 g/L, 0.04 g/L, and 0.1 g/L) of GLDA-LDH were added to the respective flasks. After 5 h of shaking, the supernatant was collected, diluted, and subjected to ion concentration analysis.

2.5. Mixing Dynamics

A competitive adsorption stock solution containing Pb(II) and Cd(II) was prepared in ultrapure water (pH = 5) with a final concentration of 0.01 mol/L. Afterward, 1 mL, 5 mL, and 10 mL of the stock solution (500 mg/L) were poured into separate 500 mL volumetric flasks. Each flask was diluted to the mark with ultrapure water, resulting in initial concentrations of 1, 5, and 10 mg/L. Nitric acid was employed to set the pH of all solutions to 5.

2.6. Soil Adsorption

The precursor soil was obtained from Panyu district, Guangzhou, China, whose pH value was 6.3. Simulated contaminated soil was prepared by thoroughly mixing 100 g of precursor soil with 200 mL of a solution containing 0.4 g/L Pb(II) and 0.3 g/L Cd(II), followed by drying and grounding. Then, 50 g of dry soil contaminated with Pb(II) and Cd(II) was added to three 250 mL conical flasks. Among these, two flasks were supplemented with 0.5 g of GLDA-LDH adsorbent and 100 mL of deionized water, while the third flask served as an untreated control. The adsorbent-containing flasks were divided into two treatment groups; one was shaken in a 298 K incubator for 6 h to simulate dynamic environmental conditions, and the other remained static for the same duration to represent quiescent natural settings. After 6 h, all three samples were separately filtered. The filtrates were centrifuged for 10 min, and the supernatant was collected, diluted, and analyzed for its Pb(II) and Cd(II) concentrations using an atomic absorption spectrophotometer. Each experiment was conducted in triplicate, with final values calculated as the means of three independent measurements. The untreated sample served as a control, representing the maximum concentration of contaminants that could be released from the polluted soil.

2.7. Characterization Methods

Detailed information can be found in the Supplementary Materials.

3. Results and Discussion

3.1. Characterization of GLDA-LDH

The XRD curves of NO₃-LDH and GLDA-LDH are shown in Figure 2a. It can be seen that the NO₃-LDH precursor showed diffraction peaks at 9.9°, 19.9°, 30.8°, and 60.4°, corresponding to the (003), (006), (009), and (110) crystal planes, respectively, which proves that the NO₃-LDH precursor was successfully prepared [27]. After intercalation, the peak corresponding to the (003) crystal plane exhibited a notable shift to a lower angle of 7.46°, accompanied by a significant reduction in its intensity. This observation suggests a decrease in crystallinity following intercalation modification, which can be attributed to the disruption of regular lattice arrangements due to guest anions’ insertion into the interlayer spacing. According to Bragg’s formula, the lattice distance of d₍₀₀₃₎ is shown in

Table 1. Lattice parameters of NO₃-LDH and GLDA-LDH.

Samples	(003)
	2θ (°)	d (nm)
NO3-LDH	9.90	0.89
GLDA-LDH	7.46	1.18

. The d₍₀₀₃₎ of NO₃-LDH is similar to that in the reported literature [28,29]. The increase in d₍₀₀₃₎ indicates that the interlayer spacing of GLDA-LDH increased after intercalation, which can be preliminary proof that GLDA has successfully entered into the interlayer spaces.

FT-IR was employed to characterize functional groups and chemical bonding configurations in the LDH samples. The broad band observed at 3448 cm⁻¹ in Figure 2b corresponds to O-H stretching vibrations from water molecules and hydroxyl groups in the host layers of LDH [30,31], and a strong absorption peak at 1385 cm⁻¹ was identified as the nitrate ions presented in the interlayer spaces of NO₃-LDH [32]. The peaks below 800 cm⁻¹ were assigned to the M-OH structure [33]. Compared with the NO₃-LDH precursor, GLDA-LDH exhibited new absorption peaks. The absorption peaks at 1618 cm⁻¹ and 1400 cm⁻¹ represent the asymmetric and symmetric stretching vibration of carboxyl group, respectively [34]. Meanwhile, the absorption peaks at 1296 cm⁻¹ represent the stretching vibration of the C-N group presented in GLDA [35]. The FT-IR results further indicate that GLDA has successfully intercalated into the interlayer spaces of GLDA-LDH.

Thermogravimetric analysis (TGA) of GLDA-LDH revealed its thermal stability. It is clear that GLDA-LDH gradually decomposes with increasing temperatures. As illustrated in Figure 2c, the first decomposition stage below 200 °C corresponds to the liberation of interlayer water molecules [36]. The second decomposition stage below 320 °C is associated with the interlayer GLDA degradation and M-OH dehydroxylation in the host layers, indicating initial structural collapse of the LDH backbone. The third decomposition stage below 620 °C showed a pronounced mass loss (14.87%), attributable to the complete dehydration and structural collapse of the layered framework, accompanied by metal oxide formation.

The microstructure and morphological dimensions as well as the elemental distribution of the LDH samples were characterized by using SEM combined with energy-dispersive X-ray spectroscopy (EDS). The NO₃-LDH precursor displayed well-defined lamellar structures with noticeable particle aggregation (Figure S1) [35]. Following intercalation, GLDA-LDH maintained the characteristic layered structure of LDH while showing partial lamellar disintegration and reduced particle dimensions (Figure 2d), consistent with XRD results. As presented in Figure 2e, the EDS mappings of GLDA-LDH demonstrated uniform distributions of Zn, Al, O, and N elements. The EDS elemental analysis of GLDA-LDH is presented in Figure S2, and the corresponding elemental composition is summarized in Table S1. The detected Zn and Al originated from the LDH host layers, whereas the signal of the N element was attributed to GLDA introduction.

The textural properties of GLDA-LDH, including the specific surface area (S_BET), pore diameter (D_P), and pore volume (V_P), were quantified via BET analysis and nitrogen isotherm analysis with detailed parameters listed in

Table 2. The textual properties of GLDA-LDH.

Sample	SBET (m2/g)	Vp (cm3/g)	Dp (nm)
GLDA-LDH	39	0.05603	5.571

. As shown in Figure S3, GLDA-LDH exhibited a type IV nitrogen adsorption-desorption isotherm with a distinct H3 hysteresis loop, demonstrating the mesoporous structure of GLDA-LDH. This reveals the potential of GLDA-LDH as an effective adsorbent.

3.2. Adsorption Tests

3.2.1. Effect of Adsorbent Dose and Initial pH

Optimizing the adsorbent dosage proves critical for achieving the target adsorption efficiency while maintaining process cost-effectiveness [37]. Adsorbent dosage effects (0.01–0.1 g/L) of GLDA-LDH on Pb(II) and Cd(II) adsorption capacity and removal efficiency were systematically investigated. Figure 3a,b clearly demonstrates a proportional relationship between the adsorbent dosage and removal efficiency while showing an inverse correlation with the adsorption capacity. With adsorbent dose escalation from 0.01 to 0.1 g/L, the Pb(II) adsorption capacity underwent a sharp decline (from 141.16 mg/g to 38.51 mg/g), whereas the Cd(II) capacity experienced a notable reduction (from 303.79 mg/g to 49.84 mg/g). Concomitantly, the Pb(II) removal efficiency progressively enhanced from 16.44% to 38.52%, paralleled by Cd(II) efficiency elevation (35.38% → 49.59%). At elevated adsorbent dosages, heavy metal ions were rapidly adsorbed from the solution due to abundant adsorption sites, resulting in a diminished concentration gradient between the bulk solution and adsorbent surfaces. This reduction in the primary driving force decreased the adsorption capacity per unit mass. However, the higher dosage increased the active site availability, enhancing contaminant removal efficiency [37]. Another contributing factor to the diminished adsorption capacity is the aggregation of LDH, particularly at high dosages [38]. This agglomeration effect reduces the accessible active sites and effective surface area of the adsorbent.

The adsorption performance of GLDA-LDH exhibited a strong pH dependency for two reasons: (1) pH-influenced surface charge distribution of LDH material and (2) basic conditions causing precipitation of heavy metal ions [39]. The pH range of 2.4–5.5 was focused on, and this was determined by two constraints; structural dissolution of LDH occurs below pH 2.0 under strong acidic conditions, while Pb(II) undergoes hydrolysis precipitation above pH 5.5. Figure 3c,d depicts the pH-dependent heavy metal adsorption profiles of GLDA-LDH. The minimum Pb(II) and Cd(II) adsorption capacities occured at a pH of 2.4, and the maximum capacities occurred at pH levels of 4.80 and 5.55, respectively. As the solution’s pH increased to weakly acidic conditions (4.6–5.2), the adsorbent’s surface charge underwent significant transformation; strong acidity (pH < 4.6) induced excessive protonation, creating a highly positive surface charge that hindered cation (Pb(II) and Cd(II)) adsorption through electrostatic repulsion and H⁺ competition for active sites [40]. With the pH elevation, the gradual deprotonation reduced the surface positive charge density, thereby improving electrostatic attraction with metal cations and enhancing complexation between the metal ions and GLDA. Another point worth mentioning is that optimal adsorption capacities were achieved within distinct pH values for each metal species. Pb(II) removal peaked at a pH of 4.80 (tested in 0.1 g/L solution), while Cd(II) adsorption attained its maximum efficiency at a slightly elevated pH value of 5.55. This differential pH dependence likely originated from the contrasting hydration properties and ionic radii of the two metal ions, which governed their specific interactions with the active sites of GLDA-LDH.

3.2.2. Effect of Contact Time and Kinetic Analysis

The heavy metal adsorption performance of GLDA-LDH was analyzed across varying contact times. The adsorption capacities of NO₃-LDH after 240 min of uptake (as shown in Figure S4) were 15.5 mg/g for Pb(II) and 19.0 mg/g for Cd(II). Following intercalation with GLDA, the adsorption capacities of GLDA-LDH increased significantly to 61.9 mg/g for Pb(II) and 37.4 mg/g for Cd(II), as presented in Figure 4a. These results clearly demonstrate the facilitating effect of GLDA on heavy metal ion uptake. Furthermore, Figure 4a reveals that the adsorption process of GLDA-LDH exhibited distinct kinetic behavior, characterized by an initial rapid adsorption phase. This phase was primarily driven by a high concentration gradient and strong mass transfer driving forces. The subsequent gradual deceleration of adsorption kinetics resulted from progressive occupancy of active sites by metal cations, ultimately reaching equilibrium when the material’s adsorption capacity became fully exhausted [41]. The adsorption system demonstrated a higher binding affinity for Pb(II) compared with Cd(II), governed by distinct hydration properties. Cd(II) possessed elevated hydration-free energy and an expanded hydrated radius compared with Pb(II), establishing substantial energy barriers for dehydration during interfacial migration. This strong hydration stabilization thermodynamically favors Cd(II) ions remaining solvated in the aqueous phase. In contrast, Pb(II) exhibited reduced hydration-free energy demand and a smaller radius, facilitating partial dehydration with lower energy consumption [42]. These characteristics enable Pb(II) ions to effectively approach adsorption sites.

To investigate the adsorption kinetic mechanisms between adsorbents and metal ions, researchers have developed various kinetic models, primarily including Lagergren’s pseudo-first-order equation (Equation (1)) and Ho’s pseudo-second-order equation (Equation (2)) [31]:

$$\lg \left(Q_{\mathrm{e}}-Q_{\mathrm{t}}\right)=\lg Q_{\mathrm{e}}-{\displaystyle \frac{{\mathrm{K}}_{1}t}{2.303}}$$

(1)

$${\displaystyle \frac{\mathrm{t}}{Q_{\mathrm{t}}}}={\displaystyle \frac{1}{{\mathrm{K}}_2{Q}_{\mathrm{e}}}}+{\displaystyle \frac{\mathrm{t}}{Q_{\mathrm{e}}}}$$

(2)

where Q_e denotes the equilibrium adsorption amount (mg/g); Q_t denotes the instantaneous adsorption amount (mg/g); K₁ represents the primary rate constant (h⁻¹); and K₂ represents the secondary rate constant (g/(mg h)). The acquired kinetic data are displayed in

Table 3. Parameters of heavy metal ion adsorption kinetics.

Ion	Pseudo First Order		Pseudo Second Order
	K1	R12	K2	R22
Pb(II)	0.06629	0.9613	0.0015	0.9811
Cd(II)	0.05524	0.9652	0.0019	0.9839

and depicted in Figure 4b,c. The pseudo-second-order kinetic model yielded higher linear correlation coefficients (0.9811 for Pb(II) and 0.9839 for Cd(II)) compared with the pseudo-first-order models, confirming a chemisorption-controlled adsorption mechanism.

3.2.3. Adsorption Isotherm

Elevated metal ion concentrations enhance the concentration gradient between the liquid phase and adsorbent surface, thereby increasing the mass transfer driving force. Adsorption isotherm studies provide conclusive evidence for evaluating the adsorbent capacity, serving as a critical criterion for performance assessment. The adsorption equilibrium of Pb(II) and Cd(II) on GLDA-LDH was analyzed using the Langmuir (Equation (3)) and Freundlich (Equation (4)) isotherm models under varied initial concentration regimes:

$${\displaystyle \frac{C_{\mathrm{e}}}{Q_t}}={\displaystyle \frac{1}{Q_{\mathrm{m}}\mathrm{b}}}+{\displaystyle \frac{C_{\mathrm{e}}}{Q_{\mathrm{m}}}}$$

(3)

$$\ln Q_{\mathrm{e}}=\ln {\mathrm{K}}_{\mathrm{f}}+{\displaystyle \frac{1}{\mathrm{n}}}\ln C^{\mathrm{e}}$$

(4)

where C_e denotes the equilibrium mass concentration of heavy metal ions in a solution (mg/L); Q_e denotes the equilibrium adsorption amount of heavy metal ions (mg/g); Q_m denotes the maximum adsorption capacity of heavy metal ions (mg/g); b is a constant (L/mg); and K_f and n denote constants related to the adsorption capacity ((mg^1−1/n L^1/n/g) and 8.314 (J/(mol K)).

The Langmuir model demonstrated high fitting correlation coefficients (R² > 0.99), indicating that the adsorption of heavy metals by GLDA-LDH predominantly follows a monolayer chemisorption mechanism (Figure 5a,b). However, the Langmuir model showed a significantly reduced fitting accuracy at 318 K. This may be attributed to relatively high temperatures affecting the monolayer adsorption mechanism. The Freundlich model yielded slightly lower correlation coefficients than the Langmuir model, with R² values of 0.98, 0.99, and 0.94 for the Pb(II) adsorption isotherms at three temperatures and 0.98, 0.99, and 0.99 for the Cd(II) adsorption isotherms (Figure 5c,d). The Freundlich parameter n reflects the adsorption affinity between GLDA-LDH and the target metal ions. As shown in

Table 4. Parameters of heavy metal ion adsorption isothermal adsorption.

T (K)	Ion	Langmuir			Freundlich
		Qm	b	RL2	Kf	n	RF2
298	Pb(II)	219.2	0.017	0.997	10.057	1.43	0.984
308		167.2	0.006	0.999	4.391	1.32	0.989
318		91.4	−0.00045	0.992	0.706	1.06	0.943
298	Cd(II)	121.9	0.0018	0.992	1.667	1.19	0.985
308		61.7	0.0077	0.999	1.585	1.57	0.993
318		58.9	0.0015	0.994	0.504	1.15	0.991

, the n values suggest moderate-strength interactions between GLDA-LDH and both Pb(II) and Cd(II), with comparable interaction magnitudes between the two metal ions.

The maximal adsorption capacities of GLDA-LDH for Pb(II) and Cd(II) were 219.2 mg/g and 121.9 mg/g, respectively.

Table 5. Maximal adsorption capability comparison across materials for Pb(II) and Cd(II).

Adsorbent	Adsorption Capacity (mg/g)		Reference
	Pb(II)	Cd(II)
GLDA-LDH	219.2	121.9	This work
Potato starch phosphate polymer	106.25	91.84	[43]
ZnAl-LDH-BC600	128.4	99.8	[18]
M-AGS	200.0	/	[44]
LDH-200	125.9	69.11	[45]
Magnetized activated carbons	253.2	73.3	[46]
LDH@GO-SH	/	102.77	[47]
LDH-DTC600	181.7	125.5	[48]

compares the Pb(II) and Cd(II) adsorption capacities of GLDA-LDH with other adsorbents, demonstrating that GLDA-LDH exhibited superior adsorption performance for both metal ions in aqueous solutions.

3.2.4. Competitive Adsorption and Soil Adsorption

Under consistent adsorption conditions, GLDA-LDH was employed to treat individual and mixed solutions containing Pb(II) and Cd(II) ions, with the adsorption time systematically varying from 0 to 240 min. Figure 6a illustrates the temporal evolution of metal removal efficiencies and competitive adsorption patterns observed during these experiments. As predicted, the adsorption contact time had a positive effect on the adsorption of Pb(II) and Cd(II) mixed solutions with GLDA-LDH. In Figure 6a, Pb(II) and Cd(II) both exhibited lower adsorption capacities and reached their equilibrium adsorption capacities more rapidly compared with testing in a single solution. This behavior is attributable to the simultaneous adsorption of Pb(II) and Cd(II) by GLDA-LDH, where the finite adsorption sites become occupied early, thereby preventing further adsorption processes. This demonstrates that Cd(II) competed with Pb(II) for adsorption sites on GLDA-LDH, which is similar to the findings in the reported literature [49]. Meanwhile, a stronger hydration-free energy and a larger hydrated radius of Pb(II) led to its predominant adsorption in competitive systems, enabling GLDA-LDH to adsorb greater Pb(II) uptake than Cd(II) under identical contact durations [50]. The competitive adsorption curves were fitted using Equations (1) and (2), and the results are shown in Figure 6b and

Table 6. Parameters of adsorption kinetics of mixed heavy metal ions.

Ion	Pseudo First Order		Pseudo Second Order
	K1	R12	K2	R22
Pb(II)	0.2129	0.9357	0.0097	0.9866
Cd(II)	0.3017	0.8713	0.0353	0.9421

. Based on the contact time and adsorption capacity model, the trend of the fitting curves was similar in the adsorption of metal ions in single and mixed solutions onto GLDA-LDH. The adsorption capacity demonstrated a rapid initial uptake followed by gradual stabilization over time, approaching equilibrium with minimal variation after 90 min. The secondary kinetic model revealed superior linear fitting correlation coefficients (R² = 0.9866 for Pb(II) and 0.9421 for Cd(II)) compared with the primary kinetic model in competitive heavy metal adsorption. These high correlation values further confirm that chemisorption serves as the primary mechanism governing the adsorption behavior.

EDS analysis confirmed the adsorption of heavy metal ions by GLDA-LDH. In Figure S5a, no distinct Pb or Cd signal peaks were detected in GLDA-LDH prior to adsorption. Following adsorption, as presented in Figure S5b–d, EDS elemental mapping revealed the more intensive signals of Pb(II) and Cd(II) across the GLDA-LDH surface, indicating that Pb(II) and Cd(II) were successfully adsorbed by GLDA-LDH. X-ray photoelectron spectroscopy (XPS) was utilized to compare the information of GLDA-LDH before and after adsorption. The XPS survey spectra of GLDA-LDH before and after adsorbing Pb(II) or Cd(II) are shown in Figure S6. The peaks observed at 143.58 eV (Pb 4f_5/2) and 139.08 eV (Pb 4f_7/2) are displayed in Figure 6c, which may have been due to the presence of a surf-O-Pb complex in the LDH surface and COO-Pb structure in the interlayer spaces [51]. In Figure 6d, the peaks at 412.28 eV (Cd 3d_3/2) and 405.58 eV (Cd 3d_5/2) were detected, revealing that the Cd(II) was adsorbed onto the GLDA-LDH, which aligned with the EDS results. According to the N 1s spectra of GLDA-LDH before and after adsorbing Pb(II) (Figure 6e), it can be seen that the peak at 400.00 eV corresponding to C-N barely exhibited distinct change, indicating that Pb(II) does not chelate with N in GLDA-LDH. In contrast, after adsorbing Cd(II), GLDA-LDH presented a new peak of 405.58 eV assigned to the Cd-N bond, accompanied by a noticeable decrease in the signal intensity at 400.00 eV. This behavior suggests that GLDA can effectively bind Cd(II) through chelation involving its functional group.

Meanwhile, the GLDA-LDH after the competitive adsorption experiments was subjected to release experiments in deionized water, as illustrated in Figure 6f. After 180 min of slow release, the amounts of Pb(II) and Cd(II) desorbed from GLDA-LDH were 10.3 mg/g and 1.8 mg/g, respectively. Both values were significantly lower than the equilibrium adsorption capacities of GLDA-LDH (35.1 mg/g for Pb(II) and 13.3 mg/g for Cd(II)), demonstrating its strong retention capability for heavy metal ions (70.66% for Pb(II) and 86.47% for Cd(II)) with minor desorption after adsorption. The higher value can be attributed to the chelation effect of GLDA, as confirmed by the XPS results, which enabled GLDA-LDH to enhance the retention rate of Cd(II).

Under consistent adsorption conditions, GLDA-LDH was employed to treat Pb(II) and Cd(II) in both individual and mixed solutions. The experimental parameters were maintained while exclusively varying the calcium carbonate (CaCO₃) concentrations in hard water systems. As depicted in Figure 7a, the increasing CaCO₃ concentrations in hard water progressively reduced the Q_m of GLDA-LDH for both the individual Pb(II) and Cd(II) systems. A parallel decline in co-adsorption performance was observed for the mixed Pb(II)-Cd(II) solutions. This observation indicates that Ca(II) competes with heavy metal ions (Pb(II) and Cd(II)) for adsorption sites under hard water conditions. In fact, Ca(II) is identified as the interfering ion because of its similar structures and hydration energies in aquation solutions [52]. It is evident that the declines in adsorption capacity were more pronounced in the single heavy metal ion solutions compared with the Pb(II) and Cd(II) mixed solutions. However, the decrease remained gradual in all cases. Even when the CaCO₃ concentration increased to 675 mg/L, GLDA-LDH still maintained adsorption capacities of 59.6 mg/g for Pb(II) and 26.3 mg/g for Cd(II) in single heavy metal ion solutions and maintained adsorption capacities of 57.0 mg/g for Pb(II) and 20.1 mg/g for Cd(II) in the Pb(II) and Cd(II) mixed solutions.

Similarly, GLDA-LDH was applied to adsorb Pb(II) and Cd(II) in contaminated soils over a 6 h treatment period, with systematic evaluation of static versus oscillating environments. As can be seen in Figure 7b, GLDA-LDH exhibited a decent heavy metal adsorption capacity in contaminated soil. Notably, the oscillating environment significantly enhanced the adsorption of heavy metal. The oscillating environment demonstrated equilibrium adsorption capacities of 20.80 mg/g (Pb(II)) and 25.70 mg/g (Cd(II)), representing 43% and 301% increases, respectively, compared with static conditions. Meanwhile, the metal removal rates reached 50.00% (Pb(II)) and 41.79% (Cd(II)) under oscillation, having 43% and 301% improvements over static operation, respectively. This augment confirms the critical role of fluid dynamics in enhancing the adsorption capacity and removal rate of GLDA-LDH.

4. Conclusions

This study developed a novel GLDA-intercalated ZnAl layered double hydroxide (GLDA-LDH) with robust performance in removing Pb(II) and Cd(II) from aqueous and soil environments. Structural characterization via XRD, FT-IR, TGA, and SEM-EDS confirmed the intercalation of GLDA into LDH interlayers. Adsorption studies demonstrated strong pH-dependent and dosage-sensitive behavior for GLDA-LDH, which achieved maximum capacities of 219.2 mg/g for Pb(II) and 121.9 mg/g for Cd(II). Competitive adsorption studies further underscored GLDA-LDH’s higher affinity for Pb(II) over Cd(II). XPS analysis elucidated distinct binding mechanisms, where surface oxygen and carboxyl groups dominated Pb(II) uptake, whereas Cd(II) adsorption involved nitrogen coordination. Desorption tests indicated strong retention, highlighting the stability of metal–LDH interactions. Remarkably, GLDA-LDH maintained its performance in hard water and contaminated soils. The adsorption performance, structural tunability, and environmental compatibility make it a strong candidate for practical water treatment applications.