Facile Synthesis and Characterization of Novel Analcime@Calcium Aluminate@Polyethylene Glycol 400 Nanocomposite for Efficient Removal of Zn(II) Ions from Aqueous Media

Abstract

Excessive Zn(II) ions in aquatic environments pose significant risks to both human health and ecological systems due to their toxic effects, bioaccumulation potential, and interference with essential biological processes. To address this issue, a novel analcime@calcium aluminate@polyethylene glycol 400 (ACP) nanocomposite was fabricated using the hydrothermal technique, alongside an analcime@calcium aluminate (AC) nanocomposite for the efficient elimination of Zn(II) ions from aqueous media. X-ray diffraction (XRD) analysis affirmed the successful formation of crystalline phases, revealing average crystallite sizes of 72.93 nm for AC and 63.60 nm for ACP. Energy-dispersive X-ray spectroscopy (EDX) confirmed the elemental composition of the nanocomposites, showing that AC primarily contained oxygen, sodium, aluminum, silicon, and calcium, whereas ACP incorporated 19.3% carbon due to the polyethylene glycol 400. Field emission scanning electron microscopy (FE-SEM) revealed that AC exhibited hexagonal and platelet-like structures, whereas ACP displayed more dispersed and layered morphologies. High-resolution transmission electron microscopy (HR-TEM) confirmed the presence of stacked platelet-like structures in AC and more defined, separated nanosheets in ACP. The maximum adsorption capacities of AC and ACP were 149.93 and 230.95 mg/g, respectively. The adsorption pathway of Zn(II) ions onto ACP nanocomposite involved three primary interactions: electrostatic attraction facilitated by calcium aluminate, ion exchange provided by analcime, and complexation promoted by polyethylene glycol 400. Thermodynamic analysis indicated that the adsorption process was exothermic, spontaneous, and primarily chemical in nature. Kinetic modeling confirmed that adsorption followed the pseudo-second-order model, while isotherm studies demonstrated adherence to the Langmuir model, indicating monolayer adsorption on homogeneous sites.

1. Introduction

Water pollution by heavy metals is a significant global concern, originating from both natural and anthropogenic sources. Naturally, heavy metals enter water bodies through geological weathering, volcanic eruptions, and leaching from soil. Nevertheless, human activities are the primary contributors to elevated heavy metal concentrations in aquatic environments. Industrial processes such as battery manufacturing, metal plating, mining, and electronics production discharge substantial amounts of toxic metals into water systems [1,2,3]. Agricultural runoff, containing pesticides and fertilizers enriched with metal compounds, further exacerbates contamination. Additionally, improper disposal of industrial and household waste, landfill leachate, and atmospheric deposition from fossil fuel combustion contribute to the accumulation of heavy metals in water sources, posing serious ecological and human health risks [4,5]. Heavy metals present significant risks to both ecosystems and human health because of their persistence, toxicity, and bioaccumulative tendencies. In the environment, heavy metals negatively impact aquatic species by affecting reproductive cycles, enzymatic activities, and metabolic processes in fish and microorganisms. Bioaccumulation in the food chain leads to higher concentrations in predators, including humans [6,7]. Heavy metals such as lead, mercury, cadmium, zinc, and arsenic are known carcinogens and neurotoxins that can cause severe physiological damage even at low concentrations. In humans, exposure to heavy metals through drinking contaminated water, consuming affected food, or inhaling metal-laden particles can result in kidney dysfunction, neurological disorders, respiratory complications, and cardiovascular diseases. Long-term exposure has been linked to cognitive impairments, organ failure, and increased mortality rates [8,9]. Among the various heavy metals, zinc is extensively released into aquatic ecosystems through industrial effluents such as galvanizing plants, battery production, mining activities, metal smelting, and pigment manufacturing facilities [10,11]. These industries often discharge Zn(II) ions in soluble forms, which readily dissolve and persist in water systems. Unlike the essential levels of zinc required for biological processes, elevated concentrations become toxic and challenging to remove due to zinc’s amphoteric nature and complexation behavior with organic and inorganic ligands in water matrices [12,13]. Although zinc is a fundamental trace element required for biological functions, excessive levels of Zn(II) ions in water pose significant environmental and health risks. High concentrations of Zn(II) ions in aquatic ecosystems lead to toxicity in fish and aquatic organisms by disrupting ion regulation, enzyme activities, and reproductive processes. The accumulation of Zn(II) ions in soil and sediments can inhibit plant growth and alter microbial communities, affecting overall biodiversity. In humans, excessive Zn(II) ion exposure through drinking contaminated water or eating contaminated food results in gastrointestinal distress, nausea, vomiting, and immune system suppression. Chronic exposure may cause neurotoxicity, pancreatic dysfunction, and disruptions in copper and iron metabolism, leading to anemia and other deficiencies [10]. The potential dangers associated with Zn(II) ion contamination necessitate efficient remediation strategies to mitigate its harmful impact. Several techniques have been developed for the removal of heavy metals from contaminated water, including adsorption [14], chemical precipitation [15], membrane filtration [16], coagulation–flocculation [17], electrodialysis [18], and bioremediation [19]. Among these methods, adsorption is widely favored due to its cost-effectiveness, high efficiency, and ease of operation. Unlike chemical precipitation, which generates secondary sludge requiring disposal, adsorption produces minimal waste. Membrane filtration, though effective, is energy-intensive and prone to membrane fouling, whereas adsorption can be implemented with low-cost materials and does not require extensive infrastructure. Additionally, adsorption provides high selectivity for specific heavy metals and allows for material regeneration and reuse, making it an environmentally sustainable option. Its adaptability across different scales, from laboratory to industrial applications, further reinforces its practicality in water purification [20,21]. In addition to adsorption and ion exchange techniques, photocatalysis has emerged as a promising approach for heavy metal removal from contaminated water. This method often employs semiconducting metal oxides, including perovskite-based materials, which exhibit tunable structural and electronic properties, enabling efficient charge separation and visible-light-driven redox reactions. A recent comprehensive review by Ibrahim et al. [22] highlights the potential of modified perovskites in wastewater remediation through photocatalytic and photo-Fenton processes, emphasizing their role in advanced water treatment. Metal oxide and zeolite nanoparticles and their composites have emerged as highly effective adsorbents for heavy metal removal due to their high surface area, tunable porosity, and chemical stability [23,24]. Metal oxides exhibit strong affinities for heavy metals through mechanisms, including electrostatic attraction and surface complexation. Zeolites, with their well-defined microporous structures and ion exchange capabilities, efficiently capture heavy metal ions while maintaining high regeneration potential. The nanoscale properties of these materials enhance adsorption kinetics and capacity, making them ideal candidates for developing advanced remediation technologies. The present research focuses on the synthesis of a novel analcime@calcium aluminate@polyethylene glycol 400 nanocomposite for the efficient removal of Zn(II) ions from aqueous media. This innovative nanocomposite integrates multiple adsorption mechanisms, providing enhanced efficiency and selectivity for the removal of Zn(II) ions. Calcium aluminate contributes to electrostatic attraction, facilitating the initial capture of Zn(II) ions. Analcime, a zeolite-based component, enables ion exchange interactions that enhance adsorption selectivity. Polyethylene glycol 400 functions as a stabilizing and complexing agent, further improving zinc binding efficiency. Polyethylene glycol 400 not only contributes to structural modulation during synthesis but also plays a significant functional role during adsorption. Its terminal hydroxyl groups and ether linkages contain lone pair electrons that enable strong coordination with Zn(II) ions, forming stable metal–ligand complexes. Additionally, polyethylene glycol 400 acts as a steric stabilizer that inhibits particle agglomeration, promoting the better dispersion of active sites and increasing the accessibility of Zn(II) ions to the nanocomposite surface. This dual functionality of polyethylene glycol 400 enhances both the structural properties and adsorption efficiency of the nanocomposite. The synergy of these components results in a highly effective adsorbent with superior adsorption capacity and rapid kinetics. Despite the extensive development of adsorbents for Zn(II) ion removal, many traditional materials continue to suffer from low adsorption capacities, poor stability, and limited regeneration performance. For example, activated carbon and polyethyleneimine cryogels have demonstrated maximum Zn(II) uptake capacities of only 7.87 mg/g and 24.39 mg/g, respectively, with notable reductions in performance after repeated use [25,26]. Furthermore, advanced materials such as NiFe₂O₄/chitosan and layered double hydroxide composites, although offering higher capacities of 90.70 and 154.21 mg/g, respectively, often exhibit reduced reusability due to agglomeration or surface fouling over multiple cycles [27,28]. These limitations underscore the need for multifunctional nanocomposites that integrate a high capacity with long-term durability. In this context, the nanocomposite developed in this study shows a maximum Zn(II) adsorption capacity of 230.95 mg/g and retains a 66.83% removal efficiency even after five cycles, demonstrating its superiority over conventional and advanced materials in terms of both performance and reusability.

2. Results and Discussion

2.1. Characterization

Figure 1A,B illustrate the XRD patterns of the analcime@calcium aluminate (AC) and analcime@calcium aluminate@polyethylene glycol 400 (ACP) nanocomposites, respectively. The analcime@calcium aluminate nanocomposite is synthesized using the hydrothermal method in the absence of polyethylene glycol 400, while the analcime@calcium aluminate@polyethylene glycol 400 nanocomposite is synthesized with its presence as an organic template. The diffraction peaks corresponding to NaAlSi₂O₆ (analcime) match the standard JCPDS card No. 01-073-6448 and confirm its tetragonal crystal system. The characteristic 2θ diffraction peaks of NaAlSi₂O₆ are observed at 12.37° (200), 15.83° (211), 18.26° (220), 24.33° (312), 25.92° (400), 27.27° (411), 31.99° (224), 33.25° (314), 40.49° (325), 41.66° (602), 44.79° (316), 45.85° (444), 47.84° (604), 53.41° (800), 54.35° (741), 56.98° (606), 57.82° (734), 62.75° (219), 63.58° (646), 67.58° (941), 68.31° (860), 69.15° (727), 72.30° (3110), 73.87° (718), 75.24° (3310), 76.81° (738), and 78.27° (936), confirming the formation of the tetragonal crystal structure of analcime. Also, the XRD pattern of CaAl₂O₄ shows characteristic diffraction peaks at 29.38° (202), 30.53° (220), 35.88° (032), 37.14° (311), 37.88° (132), 38.50° (230), 42.81° (041), 46.89° (410), 48.68° (−241), 50.16° (−215), 52.46° (215), 55.30° (151), 59.17° (510), 62.01° (502), 64.53° (424), and 66.10° (522), confirming the monoclinic crystal structure of calcium aluminate, as referenced by JCPDS card No. 01-088-2477.

Figure 1A,B reveal that the peak positions remain unaltered between AC and ACP nanocomposites, confirming that the incorporation of polyethylene glycol 400 does not change the crystalline phases present. However, the intensities of several diffraction peaks in ACP are noticeably reduced compared to AC, indicating a decrease in crystallinity. This decrease is likely due to the presence of polyethylene glycol 400, which disrupts crystal growth and order during synthesis. Additionally, ACP exhibits a smaller average crystallite size of 63.60 nm compared to 72.93 nm for AC, which further supports the role of the organic template in regulating nucleation and limiting crystal growth.

The reduction in average crystallite size from 72.93 nm for AC to 63.60 nm for ACP is attributed to the action of polyethylene glycol 400 as a growth-inhibiting agent during hydrothermal synthesis. Polyethylene glycol 400 molecules adsorb onto the surface of forming nanocrystals through hydrogen bonding and van der Waals interactions, creating a steric barrier that restricts further grain coalescence and aggregation. This spatial hindrance limits the extent of crystal growth, resulting in smaller and more uniform crystallites. Additionally, the capping effect of polyethylene glycol 400 promotes better dispersion and prevents the formation of large crystalline domains, thereby contributing to the observed decrease in crystallinity and crystallite size [29].

Figure 2A,B present the EDX spectra of the AC and ACP nanocomposites, respectively, providing insight into their elemental composition. The EDX spectrum of the AC nanocomposite, shown in Figure 2A, reveals the presence of oxygen, sodium, aluminum, silicon, and calcium, which are characteristic elements of the analcime@calcium aluminate nanocomposite. In contrast, Figure 2B illustrates the EDX spectrum of the ACP nanocomposite, which exhibits an additional peak corresponding to carbon, indicating the incorporation of polyethylene glycol 400 into the nanocomposite. The elemental compositions of the samples, as detailed in

Table 1. Elemental composition of AC and ACP nanocomposites analyzed using EDS spectroscopy.

Sample	Atomic Percentages
	%C	%O	%Na	%Al	%Si	%Ca
AC	----	62.5	7.2	2.3	16.0	12.0
ACP	19.3	54.6	6.3	1.4	9.0	9.4

, demonstrate a significant difference in atomic percentages. The AC nanocomposite contains 62.5% oxygen, 7.2% sodium, 2.3% aluminum, 16.0% silicon, and 12.0% calcium, while the ACP nanocomposite exhibits a reduced content of these elements, with 54.6% oxygen, 6.3% sodium, 1.4% aluminum, 9.0% silicon, and 9.4% calcium, alongside the presence of 19.3% carbon. The observed differences arise due to the inclusion of polyethylene glycol 400 as an organic template during synthesis, which not only introduces carbon into the structure but also alters the overall atomic distribution by modifying the crystal growth process and influencing the surface composition. The decrease in silicon, aluminum, calcium, and oxygen content in the ACP nanocomposite compared to the AC nanocomposite suggests that polyethylene glycol 400 facilitates structural modifications, potentially leading to reduced crystallinity and enhanced porosity in the nanocomposite.

Figure 3A,B present the FE-SEM images of the AC and ACP nanocomposites, respectively, revealing distinct morphological differences influenced by the synthesis conditions. In Figure 3A, the AC nanocomposite exhibits a compact and irregular morphology with large agglomerated structures and well-defined hexagonal and platelet-like crystalline formations embedded within a rough matrix. The presence of these structures indicates the formation of calcium aluminate and analcime phases with limited porosity. In contrast, Figure 3B shows the ACP nanocomposite displaying a significantly different morphology characterized by loosely packed, smaller, and more dispersed particles with an increased number of irregular and layered plate-like structures. The observed changes in morphology are attributed to the incorporation of polyethylene glycol 400 as an organic template during synthesis, which influences nucleation and growth mechanisms, leading to a reduced particle size, increased surface roughness, and enhanced porosity. These structural modifications suggest that polyethylene glycol 400 acts as a structure-directing agent, promoting the formation of finer and more porous nanostructures with improved surface area and potential reactivity.

Figure 4A,B present the HR-TEM images of the AC and ACP nanocomposites, respectively, highlighting their morphological characteristics and structural differences at the nanoscale. In Figure 4A, the AC nanocomposite exhibits irregularly shaped, stacked, and aggregated platelet-like structures with significant variations in thickness and size. The presence of these dense and compact formations suggests limited porosity and a strong interparticle interaction, which is characteristic of materials synthesized without the influence of organic templates. In contrast, Figure 4B shows the ACP nanocomposite displaying more uniformly distributed and well-defined layered, plate-like structures with noticeable separations between particles. The incorporation of polyethylene glycol 400 as an organic template during synthesis plays a crucial role in modifying the morphology by reducing particle aggregation, enhancing dispersion, and promoting the formation of more distinct and less compact nanosheets. This structural transformation results in an increased surface area and improved porosity, which are beneficial for various applications requiring enhanced interactions with the surrounding media. The observed differences between the AC and ACP nanocomposites indicate that polyethylene glycol 400 facilitates a more controlled crystal growth process, preventing excessive agglomeration and leading to a more open and accessible nanostructure.

The FTIR spectra of the AC and ACP nanocomposites presented in Figure 5 reveal several characteristic absorption bands that provide insight into their chemical structures. In the AC nanocomposite (Figure 5A), the band observed at 463 cm⁻¹ corresponds to Si–O bending vibrations, while the band at 676 cm⁻¹ is attributed to Al–O or Ca–O stretching modes. A peak at 765 cm⁻¹ can be assigned to Si–O–Al symmetric stretching, and the strong absorption at 996 cm⁻¹ is due to Si–O–Si asymmetric stretching vibrations, confirming the presence of the analcime structure. The band at 1625 cm⁻¹ is likely associated with the bending vibrations of adsorbed H–O–H molecules, and the broad band at 3457 cm⁻¹ arises from O–H stretching of surface hydroxyl groups or water. In the ACP nanocomposite (Figure 5B), similar bands appear with slight shifts, indicating structural modifications after polyethylene glycol 400 incorporation. The band at 459 cm⁻¹ is related to Si–O bending, while the band at 670 cm⁻¹ is linked to Al–O or Ca–O bonds. The absorption at 761 cm⁻¹ corresponds to Si–O–Al stretching, and the persistent strong peak at 996 cm⁻¹ confirms that the silicate network remains intact. Notably, the band at 1455 cm⁻¹ is assigned to the C–H bending vibration of the polyethylene glycol 400 component, while the band at 1629 cm⁻¹ reflects the H–O–H bending of adsorbed water. The appearance of a band at 2881 cm⁻¹ indicates the C–H stretching of polyethylene glycol 400 chains, and the broad absorption at 3452 cm⁻¹ is due to O–H stretching. These vibrational features confirm the successful incorporation of polyethylene glycol 400 and support the enhanced structural complexity of the ACP composite.

As shown in

Table 2. Surface textures of AC and ACP nanocomposites.

Sample	Surface Textures
	BET Surface Area (m2/g)	Total Pore Volume (cm3/g)	Average Pore Diameter (nm)
AC	67.7	0.1160	4.5
ACP	105.6	0.3561	13.7

, the ACP nanocomposite exhibits a significantly higher BET surface area (105.6 m²/g) compared to AC (67.7 m²/g), confirming that the incorporation of polyethylene glycol 400 during synthesis effectively enhances the material’s porosity and surface accessibility. This increase in surface area is consistent with the morphological changes observed in the FE-SEM and HR-TEM images, where ACP displays more dispersed and layered structures than the compact, aggregated morphology of AC. Furthermore, the total pore volume of ACP (0.3561 cm³/g) is markedly greater than that of AC (0.1160 cm³/g), indicating the formation of additional voids and interlayer spaces due to the presence of the organic template. The average pore diameter also increases from 4.5 nm in AC to 13.7 nm in ACP, suggesting a transition toward mesoporous architecture. These structural improvements in ACP directly support its superior adsorption capacity for Zn(II) ions, as the enlarged surface area and pore network facilitate more efficient diffusion and the interaction of zinc ions with active sites.

2.2. Adsorption of Zn(II) Ions from Aqueous Media

The elimination percentage of Zn(II) ions (% E) and the adsorption capacity (Q) of the synthesized nanocomposites were determined utilizing Equations (1) and (2), respectively [30,31].

$$\% \mathrm{E}={\displaystyle \frac{{\mathrm{C}}_{\mathrm{o}}-{\mathrm{C}}_{\mathrm{e}}}{{\mathrm{C}}_{\mathrm{o}}}}\times 100$$

(1)

$$\mathrm{Q}=\left({\mathrm{C}}_{\mathrm{o}}-{\mathrm{C}}_{\mathrm{e}}\right)\times {\displaystyle \frac{\mathrm{V}}{\mathrm{W}}}$$

(2)

where C_o (mg/L) is the preliminary Zn(II) ion concentration, C_e (mg/L) is the equilibrium Zn(II) ion concentration, V (L) is the volume of the solution, and W (mg) is the mass of the adsorbent used.

2.2.1. Effect of pH

At pH 2.5, the elimination efficiency (% E) of Zn(II) ions utilizing AC and ACP nanocomposites is minimal, reaching 0.93 and 1.92%, respectively, as shown in Figure 6A. This limited removal is attributed to the high concentration of H⁺ ions in solution, which induces electrostatic repulsion between the positively charged surfaces of the nanocomposites and Zn(II) ions, as illustrated in Figure 7. The point of zero charge (pH_PZC) values for AC and ACP nanocomposites are 6.26 and 6.08, respectively, as shown in Figure 6B, indicating that at pH values lower than these thresholds, the surfaces remain positively charged, inhibiting Zn(II) ion adsorption. In contrast, at pH 6.5, the removal efficiency significantly increases to 47.57% for AC and 74.31% for ACP due to the transition of surface charge from positive to negative, thereby facilitating electrostatic attraction between Zn(II) cations and the nanocomposite surfaces as depicted in Figure 7. The enhanced adsorption performance of ACP over AC suggests that ACP provides more active sites for Zn(II) removal. The removal mechanisms of Zn(II) ions vary across different components, where calcium aluminate removes Zn(II) ions primarily through electrostatic attraction above its pH_PZC, analcime facilitates Zn(II) removal via ion exchange replacing its Na⁺ ions with Zn²⁺, and polyethylene glycol removes Zn(II) ions through complexation with its oxygen-containing functional groups, forming stable Zn(II)-PEG complexes, as depicted in Figure 7. These distinct removal pathways collectively contribute to the overall efficiency of Zn(II) ion adsorption by the ACP nanocomposite.

The complexation of Zn(II) ions with polyethylene glycol 400 in the ACP nanocomposite primarily occurs through coordination bonds between Zn(II) and the oxygen atoms of the hydroxyl and ether functional groups in the polyethylene glycol 400 chains. These Zn–O coordination interactions form inner-sphere complexes, contributing to the chemical adsorption mechanism. Previous studies have reported that such Zn–O bonds typically exhibit binding energies ranging from 100 to 160 kJ/mol depending on the donor group and coordination environment, which aligns with the exothermic and spontaneous adsorption observed in this study [32].

2.2.2. Effect of Contact Time

Figure 8 illustrates the AC and ACP nanocomposites’ time-dependent removal efficiency of Zn(II) ions, showing distinct adsorption behaviors for each material. Initially, both nanocomposites exhibit a rapid increase in removal efficiency, followed by a gradual approach to equilibrium. The ACP nanocomposite consistently demonstrates a higher adsorption capacity compared to AC, indicating its superior performance in Zn(II) removal. At 10 min, AC achieves a removal efficiency of 23.09%, whereas ACP reaches 60.29%, highlighting the faster adsorption kinetics of ACP. Over time, the removal efficiency of AC continues to increase, reaching 47.17% at 70 min, which represents its equilibrium point. In contrast, ACP reaches equilibrium at 50 min with a removal efficiency of 73.57%. Beyond these equilibrium times, no significant change in adsorption is observed, as the adsorption sites on both nanocomposites become saturated. The overall trend indicates that ACP not only removes Zn(II) ions more effectively but also reaches equilibrium faster than AC, suggesting that the presence of polyethylene glycol 400 enhances the adsorption efficiency.

The adsorption kinetics of Zn(II) ions onto AC and ACP nanocomposites are assessed utilizing the pseudo-first-order and pseudo-second-order models, as illuminated in Figure 9A,B, respectively.

The pseudo-first-order kinetic model is described via Equation (3) [33].

$$\log \left({\mathrm{Q}}_{\mathrm{e}}-{\mathrm{Q}}_{\mathrm{t}}\right)={\mathrm{logQ}}_{\mathrm{e}}-{\displaystyle \frac{{\mathrm{K}}_1}{2.303}}\mathrm{t}$$

(3)

where Q_e (mg/g) represents the amount of Zn(II) ions adsorbed at equilibrium, Q_t (mg/g) is the amount of Zn(II) ions adsorbed at time t (min), and K₁ (1/min) is the pseudo-first-order rate constant.

The pseudo-second-order model is defined by Equation (4) [33].

$${\displaystyle \frac{\mathrm{t}}{{\mathrm{Q}}_{\mathrm{t}}}}={\displaystyle \frac{1}{{\mathrm{K}}_2{\mathrm{Q}}_{\mathrm{e}}^2}}+{\displaystyle \frac{1}{{\mathrm{Q}}_{\mathrm{e}}}}\mathrm{t}$$

(4)

where K₂ (g/mg·min) is the pseudo-second-order rate constant, and the remaining terms have the same definitions as in the pseudo-first-order equation.

The kinetic parameters obtained for both models are presented in

Table 3. Kinetic parameters for the adsorption of Zn(II) ions onto AC and ACP nanocomposites, based on pseudo-first-order and pseudo-second-order models.

Sample	QExp (mg/g)	Pseudo-First-Order			Pseudo-Second-Order
		K1 (1/min)	R2	Qe (mg/g)	K2 (g/mg·min)	R2	Qe (mg/g)
AC	141.52	0.02540	0.9654	83.38	0.000650	0.9999	143.47
ACP	220.70	0.0448	0.9611	57.13	0.00195	0.9999	222.72

. The experimental adsorption capacities (Q_Exp) for AC and ACP nanocomposites are 141.52 mg/g and 220.70 mg/g, respectively. The pseudo-first-order model exhibits relatively lower correlation coefficients (R²) of 0.9654 for AC and 0.9611 for ACP, along with predicted Q_e values (83.38 mg/g for AC and 57.13 mg/g for ACP) that significantly deviate from the experimental values. In contrast, the pseudo-second-order model shows an excellent fit with R² values of 0.9999 for both AC and ACP. Additionally, the predicted Q_e values (143.47 mg/g for AC and 222.72 mg/g for ACP) closely match the experimental values, confirming that the adsorption of Zn(II) ions onto AC and ACP follows the pseudo-second-order model.

2.2.3. Effect of Temperature

Figure 10 illustrates the effect of temperature on the removal efficiency of Zn(II) ions using AC and ACP nanocomposites, showing a clear decrease in adsorption performance with increasing temperature. At 298 K, AC exhibits a removal efficiency of 47.17%, while ACP exhibits a significantly higher efficiency of 73.57%, indicating the superior adsorption capacity of ACP at lower temperatures. As the temperature rises to 328 K, the removal efficiencies of both nanocomposites decline, with AC decreasing to 15.01% and ACP reducing to 33.86%. This trend suggests that the adsorption process is exothermic, where higher temperatures reduce the affinity of Zn(II) ions for the active sites on both nanocomposites. The greater adsorption efficiency of ACP compared to AC at both temperatures highlights the enhanced surface properties and increased availability of adsorption sites in ACP, likely due to the structural modifications introduced by polyethylene glycol 400.

The thermodynamic constants for the uptake of Zn(II) ions onto AC and ACP nanocomposites are assessed through Equations (5)–(7), as illustrated in Figure 11. The relationship between the equilibrium constant and temperature is expressed by the Van’t Hoff Equation (5) [34].

$${\mathrm{lnK}}_{\mathrm{d}}={\displaystyle \frac{{\mathsf{\Delta }\mathrm{S}}^{\mathrm{°}}}{\mathrm{R}}}-{\displaystyle \frac{{\mathsf{\Delta }\mathrm{H}}^{\mathrm{°}}}{\mathrm{RT}}}$$

(5)

where K_d represents the distribution coefficient (L/g), ΔS° (kJ/mol·K) represents the standard entropy change, ΔH° (kJ/mol) represents the standard enthalpy change, R (8.314 × 10⁻³ kJ/mol·K) is the universal gas constant, and T (K) is the absolute temperature.

The Gibbs free energy change (ΔG°, kJ/mol) is calculated using Equation (6) [34].

$${\mathsf{\Delta }\mathrm{G}}^{\mathrm{°}}={\mathsf{\Delta }\mathrm{H}}^{\mathrm{°}}-{\mathrm{T}\mathsf{\Delta }\mathrm{S}}^{\mathrm{°}}$$

(6)

Equation (7) provides the mathematical definition of the distribution coefficient [34].

$${\mathrm{K}}_{\mathrm{d}}={\displaystyle \frac{{\mathrm{Q}}_{\mathrm{e}}}{{\mathrm{C}}_{\mathrm{e}}}}$$

(7)

The thermodynamic parameters, as summarized in

Table 4. Thermodynamic constants for the uptake of Zn(II) ions onto AC and ACP nanocomposites.

Sample	ΔS° (kJ/mol·K)	ΔH° (kJ/mol)	ΔG° (kJ/mol)
			298	308	318	328
AC	0.1418	−43.78	−86.05	−87.46	−88.88	−90.30
ACP	0.1381	−45.54	−86.69	−88.08	−89.46	−90.84

, confirm that the adsorption process is primarily chemical rather than physical, as indicated by the enthalpy change values (ΔH°) exceeding 40 kJ/mol. The negative values of ΔH° for both AC (−43.78 kJ/mol) and ACP (−45.54 kJ/mol) nanocomposites indicate that the uptake process is exothermic. The negative Gibbs free energy values at all studied temperatures confirm that the adsorption process is spontaneous, with increasing spontaneity at higher temperatures. The positive entropy change (ΔS°) values of 0.1418 kJ/mol·K for AC and 0.1381 kJ/mol·K for ACP suggest that the uptake process is thermodynamically feasible, as the increased disorder at the solid–liquid interface favors Zn(II) ion adsorption onto the nanocomposites.

2.2.4. Effect of Concentration

Figure 12 demonstrates the effect of initial Zn(II) ion concentration on the elimination efficiency when utilizing AC and ACP nanocomposites, showing a clear decline in adsorption performance as concentration increases. At an initial concentration of 50 mg/L, AC achieves a removal efficiency of 82.88%, while ACP demonstrates a higher efficiency of 93.16%, demonstrating the enhanced adsorption efficiency of ACP at lower concentrations. As the concentration increases to 300 mg/L, the removal efficiency decreases to 24.29% for AC and 37.56% for ACP, indicating the reduced abundance of active binding sites at higher Zn(II) concentrations. The overall trend suggests that at lower concentrations, the adsorption sites are more accessible and are sufficient to capture Zn(II) ions effectively, whereas at higher concentrations, site saturation limits further adsorption.

The adsorption behavior of Zn(II) ions onto AC and ACP nanocomposites is assessed utilizing the Langmuir and Freundlich isotherm models, as displayed in Figure 13A,B, respectively.

The Langmuir model is mathematically expressed by Equation (8) [35].

$${\displaystyle \frac{{\mathrm{C}}_{\mathrm{e}}}{{\mathrm{Q}}_{\mathrm{e}}}}={\displaystyle \frac{1}{{\mathrm{K}}_3{\mathrm{Q}}_{\max }}}+{\displaystyle \frac{{\mathrm{C}}_{\mathrm{e}}}{{\mathrm{Q}}_{\max }}}$$

(8)

where Q_max (mg/g) is the maximum adsorption capacity and K₃ (L/mg) is the Langmuir adsorption constant.

The Freundlich model is described by Equation (9) [35].

$${\mathrm{lnQ}}_{\mathrm{e}}={\mathrm{lnK}}_4+{\displaystyle \frac{1}{\mathrm{n}}}{\mathrm{lnC}}_{\mathrm{e}}$$

(9)

where K₄ (mg/g) (L/mg)^1/n) represents the Freundlich constant related to uptake capacity and n represents the Freundlich exponent indicating adsorption intensity.

The maximum adsorption capacity based on the Freundlich model is determined using Equation (10) [34].

$${\mathrm{Q}}_{\max }={\mathrm{K}}_4\left({\mathrm{C}}_{\mathrm{o}}^{1/\mathrm{n}}\right)$$

(10)

The isotherm parameters, summarized in

Table 5. Adsorption isotherm parameters of Zn(II) ions onto AC and ACP nanocomposites based on Langmuir and Freundlich models.

Sample	Langmuir			Freundlich
	Qmax (mg/g)	R2	K3 (L/mg)	K4 (mg/g) (L/mg)1/n	Qmax (mg/g)	1/n	R2
AC	149.93	0.9998	0.1531	61.96	152.52	0.1700	0.8642
ACP	230.95	0.9996	0.2406	83.63	258.17	0.2128	0.8076

, confirm that adsorption follows the Langmuir model. The correlation coefficient (R²) values for the Langmuir model are 0.9998 for AC and 0.9996 for ACP, significantly higher than those for the Freundlich model, which are 0.8642 for AC and 0.8076 for ACP. The excellent fit of the Langmuir model suggests that adsorption occurs on a monolayer of homogeneous adsorption sites with no significant interactions between adsorbed molecules. The maximum adsorption capacity (Q_max) values obtained from the Langmuir equilibrium model are 149.93 mg/g for AC and 230.95 mg/g for ACP, confirming the superior uptake performance of ACP compared to AC.

This excellent agreement with the Langmuir model indicates that Zn(II) ions are adsorbed as a uniform monolayer over energetically equivalent active sites without significant adsorbate–adsorbate interactions. In contrast, the Freundlich model, which assumes heterogeneous adsorption and variable site energies, displays much lower correlation coefficients for both AC and ACP, suggesting that multilayer adsorption is not dominant in this system. This observation aligns with earlier studies involving similar nanocomposites such as chitosan-based or metal oxide–zeolite hybrids, where monolayer adsorption on homogeneous surfaces is predominant due to well-dispersed active sites and structurally uniform morphologies. These findings confirm that the adsorption mechanism of Zn(II) ions on AC and ACP is consistent with the behavior of other advanced nanomaterials reported in the recent literature [36,37].

Table 6. Comparison of maximum uptake capacities (Q_max) for several adsorbents used in the elimination of Zn(II) ions.

Adsorbent	Qmax (mg/g)	Ref
Activated carbon	7.87	[25]
Polyethyleneimine	24.39	[26]
FAU zeolite	36.77	[31]
Silica/1-hydroxy-2-acetonaphthone composite	45.13	[38]
γ-MnO2/chitosan/Fe3O4/EDTA composite	103.40	[39]
NiFe2O4/chitosan composite	90.70	[27]
PVA/EDTA resin	125.00	[40]
CuMgAl-layered double hydroxide/montmorillonite composite	154.21	[28]
AC	149.93	This study
ACP	230.95	This study

presents a comparison of the maximum adsorption capacities (Q_max) of various adsorbents used for the removal of Zn(II) ions [25,26,27,28,31,38,39,40]. The adsorption performance varies significantly among different materials, with conventional adsorbents such as activated carbon and polyethyleneimine exhibiting relatively low adsorption capacities of 7.87 and 24.39 mg/g, respectively. Inorganic and hybrid materials, including the FAU zeolite and silica/1-hydroxy-2-acetonaphthone composite, demonstrate moderate uptake capacities of 36.77 and 45.13 mg/g, respectively, suggesting an improvement in Zn(II) removal efficiency through structural modification. Advanced nanocomposites, such as γ-MnO₂/chitosan/Fe₃O₄/EDTA and NiFe₂O₄/chitosan, show enhanced adsorption performance, with maximum capacities of 103.40 and 90.70 mg/g, respectively, owing to the introduction of functional groups and increased active surface sites. The highest adsorption capacities among previously reported materials are observed for PVA/EDTA resin (125.00 mg/g) and CuMgAl-layered double hydroxides/montmorillonite composite (154.21 mg/g), highlighting the efficiency of multifunctional polymeric and layered double hydroxide-based materials. The AC and ACP nanocomposites synthesized in this study exhibit significantly higher adsorption capacities, with ACP achieving the highest Q_max value of 230.95 mg/g and AC reaching 149.93 mg/g. This remarkable performance is primarily attributed to the synergistic effect of analcime, calcium aluminate, and polyethylene glycol 400, which enhances surface functionality and adsorption site accessibility. Therefore, the outstanding adsorption capacity of ACP not only surpasses traditional and advanced adsorbents but also confirms its novelty and effectiveness as a multifunctional nanocomposite for Zn(II) removal.

The enhanced morphological features observed in ACP, such as increased particle dispersion, reduced aggregation, and distinct layered structures, as shown in the FE-SEM (Figure 3) and HR-TEM (Figure 4) images, directly contribute to its higher adsorption capacity for Zn(II) ions. The improved porosity and greater surface area provide more accessible active sites, facilitating stronger electrostatic attraction, ion exchange, and complexation interactions. These structural advantages are consistent with the superior adsorption efficiency of ACP (230.95 mg/g) compared to AC (149.93 mg/g), as evidenced in the adsorption studies.

2.2.5. Effect of Desorption and Reusability

As shown in Figure 14A, the desorption efficiency of Zn(II) ions from both AC and ACP samples increases steadily with the rise in HCl concentration, reaching a maximum at 2.0 M, where both adsorbents achieve nearly complete desorption with values of 99.46 and 99.59%, respectively. This enhancement in desorption performance with increasing acid concentration can be attributed to the high concentration of H⁺ ions in HCl solution, which effectively compete with Zn(II) ions for active sites on the adsorbent surface, thereby facilitating the release of the bound zinc ions. The strong acidic nature of HCl also helps in breaking the coordination bond in addition to the electrostatic interactions between the zinc ions and the functional groups of the adsorbent, promoting efficient desorption and regeneration of the adsorbent material.

Figure 14B demonstrates that the removal efficiency of Zn(II) ions by both AC and ACP gradually decreases over five successive adsorption–desorption cycles, with ACP consistently exhibiting better performance than AC throughout the process. Initially ACP achieves a removal efficiency of 73.57% compared to 47.17% for AC, and although a slight reduction is observed in each subsequent cycle, ACP maintains a relatively high efficiency of 66.83% by the fifth cycle, while AC drops to 40.33%. This sustained performance of ACP indicates its superior stability and regeneration capacity, making it more effective for repeated use in practical applications.

2.2.6. Effect of Interference

The influence of common coexisting ions on the adsorption performance of the AC and ACP nanocomposites toward Zn(II) ions was examined under identical conditions to assess their selectivity in multi-component systems. As shown in

Table 7. Effect of interfering ions on adsorption of Zn(II) ions using AC and ACP nanocomposites.

Interfering Ions	Q of AC (mg/g)	Q of ACP (mg/g)	Reduction in Q of AC (mg/g)	Reduction in Q of ACP (mg/g)
None (Control)	141.52	220.70	----	----
Na+	135.26	214.30	6.26	6.40
K+	134.10	212.85	7.42	7.85
Mg2+	127.35	205.40	14.17	15.30
Ca2+	125.80	202.60	15.72	18.10
Cl−	138.90	218.50	2.62	2.20
NO3−	139.15	217.80	2.37	2.90

, the presence of competing ions resulted in a slight to moderate reduction in the adsorption capacity of both adsorbents, depending on the type of ion. In the absence of any interfering species, AC and ACP exhibited maximum adsorption capacities of 141.52 and 220.70 mg/g, respectively. Monovalent cations such as Na⁺ and K⁺ caused slight reductions in adsorption, with capacities decreasing to 135.26 and 134.10 mg/g for AC and to 214.30 and 212.85 mg/g for ACP, respectively, indicating limited competition for active binding sites. Divalent cations such as Mg²⁺ and Ca²⁺ led to more noticeable declines in performance, with AC showing capacities of 127.35 and 125.80 mg/g and ACP displaying capacities of 205.40 and 202.60 mg/g, respectively, due to their stronger electrostatic interactions and greater affinity for the adsorption sites. In contrast, the presence of the anions Cl⁻ and NO₃⁻ exhibited minimal interference, as evidenced by the small decreases in adsorption capacity, with AC reaching 138.90 and 139.15 mg/g and ACP achieving 218.50 and 217.80 mg/g, respectively. These findings suggest that ACP maintains superior adsorption performance even in the presence of competing ions, highlighting its potential for use in complex aqueous environments.

3. Experiment

3.1. Materials

All chemicals used in this study were of analytical grade and were utilized without further purification. Sodium metasilicate pentahydrate (Na₂SiO₃·5H₂O), aluminum nitrate nonahydrate (Al(NO₃)₃·9H₂O), sodium hydroxide (NaOH), calcium nitrate tetrahydrate (Ca(NO₃)₂·4H₂O), polyethylene glycol 400 (H(OCH₂CH₂)_nOH), hydrochloric acid (HCl), zinc nitrate hexahydrate (Zn(NO₃)₂·6H₂O), and potassium chloride (KCl) were all purchased from Sigma-Aldrich (St. Louis, MO, USA). Deionized water was used for all solution preparations and washing procedures.

3.2. Synthesis of Analcime@Calcium Aluminate@Polyethylene Glycol 400 Nanocomposite

The synthesis of the analcime@calcium aluminate@polyethylene glycol 400 (ACP) nanocomposite was carried out using the hydrothermal process, as illuminated in Figure 15. Primarily, 20 g of Na₂SiO₃·5H₂O was dissolved in 50 mL of deionized water. In a separate beaker, 6 g of Al(NO₃)₃·9H₂O and 6 g of Ca(NO₃)₂·4H₂O were dissolved in 50 mL of deionized water. The sodium metasilicate solution was then gradually added to the aluminum/calcium nitrate solution with sustained stirring for 30 min. Afterward, 10 mL of polyethylene glycol 400 was introduced into the mixture and stirred for an additional 30 min. The resulting mixture was transferred into a 150 mL Teflon-lined stainless-steel autoclave and subjected to hydrothermal treatment at 180 °C for 12 h. After cooling to room temperature, the obtained precipitate was separated via filtration, thoroughly washed with deionized water to remove any residual impurities, and dried at 60 °C to yield the ACP nanocomposite. For the synthesis of the analcime@calcium aluminate (AC) nanocomposite, the same procedure was followed, except 10 mL of deionized water was used instead of polyethylene glycol 400. In the hydrothermal synthesis process, the autoclave was filled to approximately 70% of its total volume to ensure safe pressure buildup and uniform internal mixing. The heating rate was controlled at approximately 5 °C/min until the target temperature of 180 °C was reached.

3.3. Instrumentation

The structural characterization of all synthesized samples was accomplished using an X-ray diffraction diffractometer (D8 Discover, Bruker, Billerica, MA, USA) to analyze their crystalline phases. The measurements were performed with Cu Kα radiation (λ = 1.5406 Å), operated at 40 kV and 40 mA, within a 2θ range of 10° to 80° and at a scan rate of 0.02°/s. The surface morphology and elemental composition were examined using a field emission scanning electron microscope equipped with energy dispersive X-ray spectroscopy (FE-SEM/EDX, Quanta 250 FEG, Thermo Fisher Scientific, Waltham, MA, USA). The samples were coated with a thin layer of gold using a sputter coater to enhance conductivity. Further investigation of the microstructural features was conducted using a high-resolution transmission electron microscope (HR-TEM, JEM-2100Plus, JEOL Ltd., Tokyo, Japan) operating at 200 kV. FTIR analysis was carried out using a Shimadzu IRTracer-100 spectrophotometer (Kyoto, Japan) in the range of 4000–400 cm⁻¹ to identify the functional groups present in the nanocomposites. BET surface area and porosity analysis were performed using a Micromeritics ASAP 2020 instrument (Norcross, GA, USA) by nitrogen adsorption–desorption measurements at 77 K. The concentration of Zn(II) ions in solution was measured via an atomic absorption spectrophotometer (AAS, ZEEnit 700 P, Analytik Jena AG, Jena, Germany) to assess the adsorption effectiveness of the fabricated nanocomposites.

3.4. Adsorption of Zn(II) Ions from Aqueous Solutions

The adsorption experiments for Zn(II) ions using AC and ACP nanocomposites were performed under varying conditions, as outlined in

Table 8. Experimental parameters for examining the impacts of pH, temperature, contact time, and initial concentration on adsorption efficacy of AC and ACP nanocomposites toward Zn(II) ions.

Impact	V (L)	Co (mg/L)	W (mg)	T (K)	t (min)	pH
pH	0.1	150	50	298	240	2.5–6.5
Contact time	0.1	150	50	298	10–100	6.5
Solution temperature	0.1	150	50	298–328	50 (ACP) 70 (AC)	6.5
Concentration of Zn(II) ions	0.1	50–300	50	298	50 (ACP) 70 (AC)	6.5

. During the adsorption experiments, all mixtures were stirred at 500 rpm to maintain the homogeneous dispersion of particles and facilitate mass transfer. To study the effect of pH, 100 mL of a 150 mg/L Zn(II) solution was prepared, and the pH was adjusted between 2.5 and 6.5 using 0.1 M HCl or 0.1 M NaOH. Then, 50 mg of the adsorbent was added and stirred for 240 min at 298 K. To investigate the effect of contact time, 50 mg of the adsorbent was added to 100 mL of the 150 mg/L Zn(II) solution at pH 6.5, and samples were collected at time intervals from 10 to 100 min. For temperature studies, the same solution was stirred at temperatures ranging from 298 to 328 K. The contact time was set to 70 min for AC and 50 min for ACP to ensure equilibrium. Finally, to examine the effect of Zn(II) concentration, solutions ranging from 50 to 300 mg/L were tested at pH 6.5, using 50 mg of the nanocomposite and stirring at 298 K for the same respective equilibrium times. Finally, each suspension was centrifuged at 4000 rpm for 5 min to separate the solid adsorbent from the residual solution prior to Zn(II) analysis.

To investigate the regeneration performance of the Zn(II)-laden adsorbents, different concentrations of hydrochloric acid were employed as the eluting agent. Specifically, HCl solutions with concentrations of 1.0, 1.5, and 2.0 M were used, and a fixed volume of 50 mL was applied during each desorption experiment. The desorption percentage of Zn(II) ions was calculated using Equation (11).

$$\% \mathrm{D}={\displaystyle \frac{100{\mathrm{C}}_{\mathrm{d}}{\mathrm{V}}_{\mathrm{d}}}{\left({\mathrm{C}}_{\mathrm{o}}-{\mathrm{C}}_{\mathrm{e}}\right)\mathrm{V}}}$$

(11)

In this equation, C_d is the concentration of Zn(II) ions in the desorbing solution, whereas V_d is the volume of the desorbing solution.

The reusability of the synthesized adsorbents was evaluated through five consecutive adsorption–desorption cycles under controlled experimental conditions. For each cycle, the adsorption of Zn(II) ions was carried out using a zinc solution with a concentration of 150 mg/L and a volume of 100 mL. A fixed amount of 0.05 g of the adsorbent was used at a temperature of 298 K and a pH of 6.5. The contact time was maintained at 70 min for AC and 50 min for ACP. After each adsorption cycle, the spent adsorbents were regenerated using 50 mL of 2 M HCl as the eluting agent to restore their adsorption capacity for the subsequent cycle.

A binary adsorption study was conducted to evaluate the impact of interfering ions on Zn(II) adsorption using AC and ACP nanocomposites. The experiments were performed at pH 6.5 and 298 K with an initial Zn(II) concentration of 150 mg/L, maintaining a 1:1 molar ratio of Zn(II) to each competing ion. The interfering ions included Na⁺, K⁺, Mg²⁺, Ca²⁺, Cl⁻, and NO₃⁻, representing the common cations and anions found in aqueous environments. Batch adsorption experiments were conducted by adding 0.05 g of each adsorbent to 100 mL of Zn(II) solution containing the competing ion. The mixtures were stirred for 50 min for ACP and 70 min for AC. The final Zn(II) concentration was measured using an atomic absorption spectrophotometer and the adsorption capacities were determined based on mass balance calculations.

3.5. Point of Zero Charge (pH_PZC) of the Synthesized Nanocomposites

The point of zero charge (pH_PZC) of the AC and ACP nanocomposites was determined using the batch equilibrium method with potassium chloride (KCl) as the supporting electrolyte. A series of 50 mL KCl solutions (0.01 M) were prepared, and their initial pH (pH_I) was adjusted between 2 and 12 using dilute hydrochloric acid or sodium hydroxide solutions. A fixed amount of 50 mg of the nanocomposite was added to each solution, and the suspensions were stirred at 298 K for 24 hrs to ensure equilibrium. After equilibration, the final pH (pH_F) of each solution was recorded. The pH_PZC was identified by plotting ΔpH against pH_I according to Equation (12) [34].

ΔpH = pH_F − pH_I

(12)

where ΔpH represents the difference between the final and initial pH values of the solution. The point at which ΔpH = 0 is identified as the pH_PZC, representing the pH at which the surface charge of the nanocomposite is neutral.

4. Conclusions

This study successfully synthesized a novel analcime@calcium aluminate@polyethylene glycol 400 (ACP) nanocomposite using the hydrothermal method, alongside an analcime@calcium aluminate (AC) nanocomposite for the efficient removal of Zn(II) ions from aqueous solutions. Structural characterization confirmed the formation of well-defined crystalline phases with average crystallite sizes of 72.93 nm for AC and 63.60 nm for ACP. The elemental composition analysis verified the incorporation of polyethylene glycol 400, which significantly influenced the morphology and adsorption properties of ACP compared to AC. The adsorption performance of both nanocomposites was thoroughly investigated under varying experimental conditions. Kinetic modeling demonstrated that the adsorption process followed the pseudo-second-order model, indicating that chemisorption was the predominant mechanism. Isotherm studies confirmed that adsorption adhered to the Langmuir model, suggesting monolayer adsorption on homogeneous sites. Thermodynamic analysis further revealed that the adsorption was exothermic, spontaneous, and primarily chemical in nature, as indicated by the negative enthalpy change (ΔH), negative Gibbs free energy (ΔG), and positive entropy change (ΔS). The ACP nanocomposite exhibited a maximum Zn(II) adsorption capacity of 230.95 mg/g, while the AC nanocomposite achieved 149.93 mg/g, indicating a substantial improvement due to the incorporation of polyethylene glycol 400. The superior adsorption efficiency of ACP was attributed to the synergistic effects of its components, where calcium aluminate facilitated electrostatic attraction, analcime enabled ion exchange, and polyethylene glycol 400 enhanced complexation interactions. This high adsorption performance, along with strong regeneration ability over five reuse cycles, underscored the potential of these materials for practical applications. Given their high efficiency, reusability, and stability under variable conditions, the AC and ACP nanocomposites are promising candidates for real-world wastewater treatment, particularly for industrial effluents contaminated with heavy metals such as zinc.