Magnetic Activated Carbon Functionalized with Polyaniline for Efficient Pb (II) Adsorption from Aqueous Solutions

Abstract

Lead (Pb) contamination in water poses a significant threat to both human health and the environment as it is toxic even at very minimal concentrations. In the scope of this study, a novel magnetic composite material, AC/Fe₃O₄/PANI-SDS, was synthesized to efficiently eliminate Pb²⁺ ions from polluted water. Each component of the composite has a significant impact: the activated carbon provides a large surface area for adsorption, the magnetic iron oxide (Fe₃O₄) allows easy magnetic recovery from water systems using a magnet, and the polyaniline (PANI) and sodium dodecyl sulfate (SDS) improve the capability of the material to attract and hold onto Pb²⁺ ions. To assess the surface, magnetic, and structural properties of the prepared material, several characterization techniques were applied, such as FTIR, XRD, SEM-EDS, BET analysis, VSM, and zeta potential measurements. These tests confirmed that the composite has the right structure and functional groups to perform as a capable and efficient adsorbent. Batch adsorption studies were used to evaluate the effects of pH, interaction time, initial Pb²⁺ ion concentration, and temperature on removal efficiency. The findings highlight the composite’s remarkable adsorption efficiency after 220 min under optimal conditions, specifically at pH 6. Adsorption kinetic studies demonstrated strong agreement with the pseudo-second-order model, while isotherm analysis showed that the Langmuir model provided the highest correlation coefficient within the investigated concentration range. This fitting suggested apparent Langmuir-type adsorption behavior, with a maximum adsorption capacity of 348.39 mg/g. Thermodynamic assessment demonstrated that the elimination of Pb²⁺ ions is an endothermic and spontaneous process. In addition, the composite can be reused and recycled repeatedly without significantly reducing its effectiveness, offering an economical and ecologically sustainable approach. The findings of this research highlight the potential of the AC/Fe₃O₄/PANI-SDS composite as a new, efficient, and eco-friendly adsorbent for the elimination of Pb²⁺ ions from solutions. In real-world applications, its high capacity for adsorption, ease of separation, and reusability make it a promising treatment for heavy metal contamination.

1. Introduction

Heavy metals have long been a concern to scientists and engineers because of their increased discharge, toxic nature, and other negative effects on receiving waters [1,2,3]. Numerous heavy metals are both non-biodegradable and toxic, posing significant threats to human health and the economy. Toxic elements of interest include arsenic, cadmium, chromium, copper, lead, manganese, mercury, nickel, and zinc. Lead, in particular, is widely utilized in several industries, such as battery manufacturing, metal plating, printing, photography, explosives production, tetraethyl lead synthesis, and the manufacture of glass and ceramics. The main sources of lead in water are the effluents of processing industries [4,5]. Lead exposure in humans can result in serious harm to vital organs such as the kidneys, brain, and liver, as well as reproductive and nervous systems. Several studies have linked severe lead exposure to sterility, abortion, stillbirths, and neonatal death [6,7]. According to the World Health Organization (WHO), the guideline value for lead (Pb) in drinking water is 0.01 mg/L, while the Council of the European Union establishes a parametric value of 0.01 mg/L. By contrast, the United States Environmental Protection Agency (US EPA) has set an action level of 0.015 mg/L for lead in drinking water, with a maximum contaminant level goal (MCLG) of zero due to the absence of a safe exposure threshold for lead. The toxicity of lead in water is very high even at minimal concentrations. Therefore, the removal of toxic heavy metals from contaminated water has been attempted by several scientists, employing various effective and sustainable methods, including chemical precipitation, ion exchange, electro flotation, membrane filtration, and reverse osmosis [8,9,10,11,12]. Among the various technologies developed for heavy metal removal, adsorption has attracted considerable attention as a cost-effective and operationally simple method, making it a practical choice for wastewater treatment owing to its high removal efficiency and flexibility in material design [13,14].

In recent years, research has focused on modifying activated carbon-based materials to overcome limitations associated with metal ion adsorption. Although activated carbon possesses a high surface area and well-developed porosity, its surface chemistry is often dominated by non-specific functional groups, which can result in limited affinity and selectivity, particularly in complex aqueous matrices containing competing species [15]. Moreover, the microporous structures of conventional activated carbon can cause diffusion limitations, leading to relatively slow adsorption kinetics under certain conditions. These limitations have motivated the development of functionalized and composite activated carbon materials to enhance metal–adsorbent interactions and adsorption performance [16,17,18,19,20]. Nevertheless, regenerating carbon-based adsorbents remains a persistent challenge. Therefore, the innovation of novel materials with both high capacity and efficient regeneration potential is essential for advancing adsorption-based treatment technologies. Magnetic nanoparticles, particularly Fe₃O₄ (magnetite), have garnered considerable interest as functional elements in adsorbent composites. Fe₃O₄ nanoparticles exhibit superior magnetic properties that facilitate the straightforward recovery of the adsorbent via an external magnetic field, thereby reducing operational expenses and secondary waste production [21]. Moreover, Fe₃O₄ offers numerous surface sites for functionalization, thereby improving the material’s adsorption efficiency and selectivity for particular contaminants [22,23]. Conductive polymers, including polyaniline, polypyrrole, polyethylenimine, as well as polyacrylonitrile, constitute a distinct category of materials that offer significant efficiency in heavy metal remediation. Among these polymers, PANI is the most frequently applied in various technologies owing to its multifunctional properties, including its porous structure, excellent electrical conductivity, great thermal and environmental stability, low toxicity, ease of synthesis, non-localized π–electron conjugate system, and remarkable redox characteristics [18]. Additionally, PANI possesses various functional groups, including amine and imine, that engage with heavy metal ions via mechanisms such as electrostatic interactions, hydrogen bonding, and chelation [24,25]. Doping PANI with surfactants such as sodium dodecyl sulfate (SDS) enhances its conductivity, dispersion stability, and mechanical strength, rendering it very appropriate for composite synthesis [26].

Based on the above considerations, the present work goes beyond a simple combination of activated carbon, Fe₃O₄, and polyaniline by introducing an SDS-doped PANI layer that is uniformly integrated onto magnetic activated carbon. While activated carbon provides a high surface area and porous framework and Fe₃O₄ enables efficient magnetic separation, the incorporation of SDS plays a crucial role in improving the dispersion of PANI, increasing the accessibility of amine and imine functional groups, and introducing additional sulfonate sites. This synergistic AC/Fe₃O₄/PANI-SDS architecture enhances electrostatic interactions and chelation with Pb²⁺ ions, leading to improved adsorption kinetics, higher adsorption capacity, and stable reusability compared to conventional magnetic carbon/PANI systems. Therefore, the innovation of this study lies in the rational design of a multifunctional, magnetic activated carbon composite in which SDS-assisted PANI functionalization uniquely enhances adsorption performance, achieving efficient Pb²⁺ ion removal, facile magnetic recovery, and repeated reuse, highlighting its potential as a resource-efficient adsorbent for water remediation.

2. Chemicals and Experimental Methods

Sigma-Aldrich (St.Louis, MO, USA) provided the sodium dodecyl sulfate (SDS). Acros Organics (Geel, Belgium) provided the iron (III) chloride hexahydrate (FeCl₃·6H₂O) and iron (II) sulphate heptahydrate (FeSO₄·7H₂O). Chempur (Piekary Śląskie, Poland) provided the aniline, ammonium hydroxide, and ammonium persulfate, which were used for the synthesis of the polymer. Norit Nederland B.V. company (3824 MJ Amersfoort, The Netherlands) provided the activated carbon (Norit ROX 0.8), which was subsequently milled into a fine powder. Common laboratory acids and bases, including sodium hydroxide, hydrochloric acid, and sulphuric acid, were provided by Avantor Performance Materials (Gliwice, Poland) and Chemland Materials (Stargard, Poland), respectively. HPLC-grade ethanol was used as a solvent throughout the study. The lead (II) chloride (PbCl₂) was of analytical grade and purchased from Chempur (Piekary Śląskie, Poland). All of the reagents and chemicals used in this investigation were of analytical grade and did not require any additional purification.

2.1. Synthesizing AC/Fe₃O₄/PANI-SDS Composite

The preparation of the magnetic iron oxide on activated carbon was carried out using an alkaline co-precipitation method [27,28]. In the first step, 0.5 g of powdered activated carbon was dispersed in 250 mL of aqueous solution composed of ferric chloride hexahydrate (0.09 mol/L) and ferrous sulfate tetrahydrate (0.045 mol/L) at a molar ratio of 2:1. The mixture was then stirred at 500 rpm and room temperature to promote the homogeneous dispersion of the carbon support. Precipitation was initiated by slowly adding 0.5 M sodium hydroxide until a strongly alkaline environment (pH 11–12) was reached. N₂ gas was continuously supplied into the reaction flask to prevent oxidization in the process of magnetite generation. Subsequently, the mixture was heated to 80 °C under stirring for 3 h to promote the growth and anchoring of Fe₃O₄ particles on the activated carbon surface (Figure 1). After the reaction, the magnetic carbon solid was collected and washed three times with deionized water and then twice with ethanol to remove impurities and byproducts. The purified material was then dried in an oven at 80 °C for 12 h to yield the final product.

The resultant AC/Fe₃O₄ composite was further modified with PANI via in situ oxidative polymerization in the presence of sodium dodecyl sulfate (SDS). In a typical process, 3 g of AC/Fe₃O₄ was added into the mixture solution of aniline (7 mL) and hydrochloric acid (15 mL, 1 M), and ultrasonicated for 15 min to promote interfacial contact. Separately, SDS (4 g) was dissolved in 100 mL of distilled water and the obtained solution was then added slowly to the reaction mixture. Polymerization was initiated by adding 100 mL of ammonium persulfate solution (0.5 M), corresponding to an APS-to-aniline molar ratio of approximately 0.65:1. The reaction was stirred at room temperature overnight. The obtained dark green AC/Fe₃O₄/PANI-SDS composite was separated, washed with three times with 1 M HCl, three times with distilled water, and twice with absolute ethanol to remove unreacted materials before drying at 60 °C for 12 h.

2.2. Characterization Instruments

The concentrations of metal ions in solution during adsorption experiments were measured using a microwave plasma atomic emission spectrometer (MP-AES, Agilent 4200, Santa Clara, CA, USA). Elemental quantification was carried out at the selected analytical emission wavelengths of Pb: 405.78 nm, Co: 340.51 nm, Cd: 228.80 nm, Ni: 352.45 nm, and Zn: 213.86 nm. An automatic background correction procedure was applied to compensate for non-specific plasma emission and matrix-related effects. The instrument operated with nitrogen (from a generator) as the carrier gas. The sampling time was 30 s, the rinsing time with the tested sample was 30 s, and the stabilization time was 30 s. Prior to measurement, the instrument was calibrated using a series of standard solutions prepared from certified stock solutions. Calibration curves with correlation coefficients (R²) above 0.999 were obtained for all analyzed metals. Quality control was ensured by including blank samples and replicate measurements. The limit of detection (LOD) and limit of quantification (LOQ) were calculated as 3σ/S and 10σ/S, respectively, where σ is the standard deviation of blank measurements and S is the slope of the calibration curve. Based on this approach, the LOD and LOQ were determined to be 0.012 mg/L and 0.041 mg/L, respectively. The MP-AES method was used to determine the total concentration of metal ions in solution, independent of their oxidation states. A potential spectral interference was identified only for cadmium under the applied MP-AES conditions. Although cobalt was quantified using its analytical emission line at 340.51 nm, cobalt also exhibited a weaker emission line at 228.615 nm, which lay in close proximity to the cadmium analytical line at 228.802 nm. According to the instrument spectral library, the intensity of the Co line at 228.615 nm (3099.1) was significantly lower than that of the Cd line at 228.802 nm (49,236.5), resulting in an estimated contribution of approximately 6.3% to the cadmium signal. This indicated that the interference effect of cobalt on cadmium quantification was limited. No significant spectral interferences were observed for cobalt, lead, or the other analyzed metals within the selected wavelength ranges.

Scanning electron microscopy (SEM, JEOL, JCM-6000 Plus, Tokyo, Japan) was utilized to characterize the morphology, such as the size, shape, and dispersion, of the composite. Elemental analysis was performed using energy-dispersive X-ray spectroscopy (EDS, Waltham, MA, USA) with an acceleration voltage of 15.00 keV. Fourier-transform infrared spectroscopy (FTIR, Nicolet, Waltham, MA, USA) and X-ray diffraction (XRD, Rigaku MiniFlex II, Tokyo, Japan) were utilized for functional group identification and to characterize the crystal structure, respectively. The composite and nanoparticles were tested for their magnetic behavior with the help of a vibrating sample magnetometer (VSM, LDJ 9600, LDJ Electronics Company, Troy, MI, USA). A Micromeritics ASAP 2010 was employed to determine the specific surface area of the nanocomposite. A Malvern Zeta Sizer Nano ZS (Malvern instrument Ltd., Malvern, UK) was employed to measure the surface charge and particle size of the colloidal form. A water bath shaker (ELPIN Type 357) was used for all experiments.

2.3. Adsorption Experiment

The stock solution of Pb²⁺ ions, having a concentration of 1000 mg/L, was prepared by dissolving 0.335 g of lead salt into distilled water, and the total volume was brought up to 250 mL using a volumetric flask. The solutions of different concentrations were prepared each day by diluting the stock solution with distilled water. To assess the effectiveness of the adsorbent, batch adsorption experiments were conducted to measure removal efficiency and adsorption capacity. A 50 mL of solution containing Pb²⁺ ions was used in each of the experiments and the initial Pb²⁺ range was between 10 and 110 mg/L. To obtain the solutions, the pH was varied between 2 and 10 with 0.1 M NaOH or HCl. A constant adsorbent dosage of 0.2 g/L (0.01 g) was injected into each sample, and mixing of the samples was performed at 150 rpm in a water bath shaker maintained at 298 K. The adsorbent was separated with the help of magnetic separation at a preset time and the remaining concentration of Pb²⁺ ions in the supernatant was examined via micro plasma atomic emission spectrometry (MP-AES, Agilent 4200, Santa Clara, CA, USA).

The efficiency of lead removal (R%), the capacity of the adsorption at a given time (q_t), and the equilibrium uptake (q_e) were estimated using the following equations:

$${\displaystyle \mathrm{R}\%=\frac{{\mathrm{C}}_{\mathrm{o}}-{\mathrm{C}}_{\mathrm{t}}}{{\mathrm{C}}_{\mathrm{o}}}} \times 100\%$$

(1)

$${\mathrm{q}}_{\mathrm{t}}={\displaystyle \frac{{(\mathrm{C}}_{\mathrm{o}}-{\mathrm{C}}_{\mathrm{t}})}{\mathrm{m}}\times \mathrm{V}}$$

(2)

$${\mathrm{q}}_{\mathrm{e}}={\displaystyle \frac{{(\mathrm{C}}_{\mathrm{o}}-{\mathrm{C}}_{\mathrm{e}})}{\mathrm{m}}\times \mathrm{V}}$$

(3)

where C₀ denotes the starting concentration of metal ions (mg/L), C_t indicates the concentration at a specific time (mg/L), and C_e refers to the equilibrium concentration (mg/L). The symbol m is the weight of the dry adsorbent (g), while V corresponds to the solution volume being tested (L).

Every experiment was carried out and the findings were presented. The optimal conditions for the batch experiments were determined as follows: pH 6, initial lead ion concentration of 25 mg/L, adsorbent dosage of 0.2 g/L (0.01 g), adsorption temperature of 298 K, and a contact time of 220 min.

2.4. Kinetic Evaluation of the Adsorption Process

Studying the adsorption kinetics helps in understanding the dynamics and rate-controlling steps of a process. These studies typically involve applying data to various models, such as pseudo-first-order, pseudo-second-order, and intraparticle diffusion models. By analyzing the fit of these models to experimental data, researchers may provide more detailed insight into the mechanisms and efficiency of the adsorption process. The linear models are represented by Equations (4), (5), and (6), respectively [29,30,31]:

$$\log \left({\mathrm{q}}_{\mathrm{e}}-{\mathrm{q}}_{\mathrm{t}}\right)=\log {\mathrm{q}}_{\mathrm{e}}-{\displaystyle \frac{{\mathrm{k}}_1}{2.303}}\mathrm{t}$$

(4)

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

(5)

$${\mathrm{q}}_{\mathrm{t}}={\mathrm{K}}_{\mathrm{p}}{\mathrm{t}}^{0.5}+\mathrm{C}$$

(6)

In this case, q_e and q_t are the capacities upon reaching the equilibrium and at a specific moment, respectively. The rate constants for the pseudo-first-order and pseudo-second-order models are K₁ and K₂ respectively, C is another constant, and K_p represents the intraparticle diffusion rate constant.

2.5. Evaluation of Adsorption Isotherms

The adsorption phenomena were explained using isotherm models. There are only a few isothermal models that take into account both the adsorbent material and adsorbed species characteristics, such as the Langmuir, Freundlich, and D–R models. These models can be used to describe the adsorption behavior in a qualitative and quantitative way. They are useful for predicting the highest adsorption capacity of an adsorbent material.

An adsorption model based on the Langmuir isotherm assumes a monolayer and homogeneous adsorption. It assumes that the adsorbate molecules are distributed uniformly and uniformly over the adsorbent surface [32]. This model further presumes that no interactions take place between adsorbate molecules and the adsorbent surface. Surface saturation occurs when an active site on the adsorbent material’s surface is full and cannot be used for further adsorption. The linear and nonlinear expressions of this model are given by Equations (7) and (8), respectively [33,34]:

$${\displaystyle \frac{{\mathrm{C}}_{\mathrm{e}}}{{\mathrm{q}}_{\mathrm{e}}}}={\displaystyle \frac{1}{{\mathrm{K}}_{\mathrm{L}}{\mathrm{q}}_{\mathrm{m}}}}+{\displaystyle \frac{{\mathrm{C}}_{\mathrm{e}}}{{\mathrm{q}}_{\mathrm{m}}}}$$

(7)

$${\mathrm{q}}_{\mathrm{e}}={\displaystyle \frac{{\mathrm{q}}_{\mathrm{m}}{\mathrm{K}}_{\mathrm{L}}{\mathrm{C}}_{\mathrm{e}}}{\left(1+{\mathrm{K}}_{\mathrm{L}}{\mathrm{C}}_{\mathrm{e}}\right)}}$$

(8)

To evaluate the possible adsorption of metal ions on the composite surface, the dimensionless factor R_L was determined using Equation (9). The calculated value of this parameter provides insight into the adsorption behavior: irreversible adsorption occurs when R_L ≥ 0, linear adsorption occurs when R_L = 1, unfavorable adsorption occurs if R_L > 1, and favorable adsorption occurs when 0 < R_L < 1.

$${\mathrm{R}}_{\mathrm{L}}={\displaystyle \frac{1}{1+{\mathrm{K}}_{\mathrm{L}}{\mathrm{C}}_{\mathrm{e}}}}$$

(9)

Ce (mg/L) refers to the metal ion concentration remaining in solution at equilibrium. The Langmuir constant K_L characterizes the ability of the binding sites, while q_m (mg/g) represents the theoretical maximal monolayer adsorption capacity. The parameter q_e (mg/g) gives the quantity of metal ions adsorbed at equilibrium.

According to the Freundlich isotherm, adsorption increases proportionally with concentration, indicating a linear correlation with surface saturation. The Freundlich isotherm is empirical and takes into account the effects of multilayer adsorption. The linear and nonlinear Freundlich equations are given by Equations (10) and (11), respectively, where 1/n represents the adsorption intensity and reflects the favorability of the adsorption process, and k_F is the Freundlich constant (mg/g) [35,36]:

$$\log {\mathrm{q}}_{\mathrm{e}}=\log {\mathrm{K}}_{\mathrm{F}}+{\displaystyle \frac{1}{\mathrm{n}}}\log {\mathrm{C}}_{\mathrm{e}}$$

(10)

$${\mathrm{q}}_{\mathrm{e}}={\mathrm{K}}_{\mathrm{F}}{\mathrm{C}}_{\mathrm{e}}^{1/\mathrm{n}}$$

(11)

The Dubinin–Radushkevich isotherm model is largely relevant to microporous adsorbents and does not adequately explain adsorption on mesoporous or macroporous materials. Additionally, it presupposes that the adsorption energy is uniform throughout the surface, which may not be the case in all instances. Chemical adsorption comprises the production of chemical bonds and is generally stronger and more specific, whereas physical adsorption is due to weaker van der Waals forces and is usually reversible [37]. Chemical and physical adsorption are separated by this paradigm. Equations (12) and (13) illustrates this model in its linear and nonlinear versions [38]:

$$\ln {\mathrm{q}}_{\mathrm{e}}=\ln {\mathrm{q}}_{\mathrm{s}}-\mathrm{\beta }{\mathrm{\epsilon }}^2$$

(12)

$${\mathrm{q}}_{\mathrm{e}}={\mathrm{q}}_{\mathrm{s}}\exp (-\mathrm{\beta }{\mathrm{\epsilon }}^2)$$

(13)

$$\mathrm{\epsilon }=\mathrm{R}\mathrm{T}\ln ({\displaystyle \frac{1}{{\mathrm{C}}_{\mathrm{e}}}}+1)$$

(14)

$$\mathrm{E}={\displaystyle \frac{1}{\surd \left(2\mathrm{\beta }\right)}}$$

(15)

The maximum theoretical saturation capacity is given by q_s (mg/g). The D–R constant, β, illustrates the mean adsorption energy (E, kJ/mol), and the Polanyi potential, ε, may be computed using Equation (14). Additionally, the nature of adsorption can be described using the mean adsorption energy (E): E values lower than 8 kJ/mol are characteristic of physical adsorption, E values in the range of 8–16 kJ/mol are generally associated with stronger electrostatic interactions or ion-exchange mechanisms, and values exceeding 16 kJ/mol may indicate chemisorption, depending on the system [39].

2.6. Thermodynamics Study

Temperature effects on adsorption were investigated over the range of 298 to 313 K. The variation in heat and spontaneity of the adsorption process were evaluated by estimating the value changes in apparent enthalpy (ΔH°), Gibbs free energy (ΔG°), and entropy (ΔS°) through the use of the following equations:

$$\ln {\mathrm{K}}_{\mathrm{d}}={\displaystyle \frac{\Delta \mathrm{S}°}{\mathrm{R}}}-{\displaystyle \frac{\Delta \mathrm{H}°}{\mathrm{R}\mathrm{T}}}$$

(16)

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

(17)

Here, the distribution coefficient (K_d) is determined using the formula K_d = (q_e/C_e), where q_e (mg/g) represents the quantity of adsorbed metal ions at equilibrium, and C_e (mg/L) is the corresponding equilibrium concentration of metal ions in solution. In Equations (16) and (17), R is the ideal gas constant (8.314 J/mol K), while T denotes the absolute temperature in Kelvin.

2.7. Regeneration Study

After the adsorption procedure, the magnetic composite was isolated from the cleansed water using a magnet. To remove the adsorbed metal ions, 50 mL of 0.1 M HCl was mixed with the recovered adsorbent and stirred at 298 °K for 2 h for each cycle. After removing the metal ions, distilled water was used to thoroughly clean the magnetic composite. For reuse, the material was dried at 333 °K for 3 h after washing. The regenerated magnetic composite was tested for five consecutive adsorption cycles under the previously optimal conditions (pH 6, initial lead ion concentration of 25 mg/L, adsorbent dosage of 0.2 g/L (0.01 g), temperature of 298 °K, and contact time of 220 min). This process demonstrated the stability of the material in preserving its adsorptive performance across several reuse cycles, highlighting its potential for sustainable and cost-effective applications in water treatment.

3. Results and Discussion

3.1. Characterization of Adsorbent

3.1.1. FT-IR Spectral Analysis

The FT-IR spectra of AC, AC/Fe₃O₄, PANI-SDS, and the AC/Fe₃O₄/PANI-SDS composite were recorded in the 400–4000 cm⁻¹ region and are presented in Figure 2a. The spectrum of activated carbon showed three major broad bands. The band detected at 3431 cm⁻¹ was related to O–H stretching vibrations, along with the tensile vibration of water molecules adsorbed onto the activated carbon surface. Meanwhile, the peak at 1634 cm⁻¹ was related to the stretching vibrations of functional groups such as C=O, C=C, or C–H present on the carbon surface. The wide peak at 1223 cm⁻¹ could be attributed to the stretching vibrations of C–O in ethers, esters, and phenols on the AC surface, as shown in Figure 2a [40,41]. The AC/Fe₃O₄ composite’s spectrum displayed a notable absorption peak at 590 cm⁻¹, related to Fe–O stretching vibrations in both the tetrahedral and octahedral positions of Fe₃O₄. Additionally, the broad band near 3435 cm⁻¹ was likely due to O–H stretching from surface hydroxyl groups and physically adsorbed water molecules. The peaks around 1629, 1383, and 1227 cm⁻¹ corresponded to the stretching vibrations of C=O, C=C, or C–O on the activated carbon surface, confirming the successful incorporation of Fe₃O₄ nanoparticles on the AC surface [42]. For the pure PANI–SDS, the spectrum showed significant absorption bands. The outstanding absorption peak at 3446 cm⁻¹ was attributed to the N–H stretching vibrations of secondary aromatic amines in PANI [43]. The bands at 2919 and 2884 cm⁻¹ confirmed the presence of the aliphatic C–H stretching mode of the alkyl tail of SDS and S=O stretching vibration of the SO₃ group in PANI–SDS [44]. The peaks at 1297 and 1481 cm⁻¹ were associated with the C=C stretching vibrations of the benzenoid and quinoid units, respectively [45]. Furthermore, the out-of-plane and in-plane bending vibrations of C–H were responsible for the peaks at 699 and 785 cm⁻¹, and the peak at 1239 cm⁻¹ corresponded to the C–N stretching vibration of an aromatic amine [46,47]. These results demonstrated that SDS was successfully integrated into the PANI matrix. The spectrum of the AC/Fe₃O₄/PANI-SDS composite was compared to those of AC, AC/Fe₃O₄, and PANI-SDS to verify its formation. The composite exhibited a characteristic peak at 1239 cm⁻¹, which was characteristic of the C–N stretching vibration of aromatic amine, and a peak at 3443 cm⁻¹, which was attributed to the N–H stretching frequency of amine in PANI. The bands observed at 570, 744, and 699 cm⁻¹ corresponded to Fe–O stretching, out-of-plane C–H bending, and C–C deformation in monosubstituted aromatic rings, respectively. Additionally, the peaks at 1301 and 1497 cm⁻¹ were attributed to the C=C stretching vibrations of benzenoid and quinoid structures, respectively. The peaks at around 1582 and 1392 cm⁻¹ corresponded to the stretching vibration bands of C=O, C=C, or C–H on the activated carbon surface, confirming the powerful incorporation of AC in the composite. Moreover, the bands at 2915 and 2845 cm⁻¹ indicated the aliphatic C–H stretching vibrations from the SDS alkyl chains. The band at 973 cm⁻¹ was attributed to the S=O stretching of the sulfonate (SO₃) group present in the composite. In summary, the differences observed in the FT-IR spectra of AC, AC/Fe₃O₄, PANI-SDS, and the AC/Fe₃O₄/PANI-SDS composite provided evidence for the effective synthesis of the newly developed AC/Fe₃O₄/PANI-SDS composite.

The FT-IR spectra of the AC/Fe₃O₄/PANI-SDS composite after adsorption provided further details on the interaction between the composite and metal ions, as shown in Figure 2b. The peak at 3443 cm⁻¹, attributed to the N–H stretching vibration of amine groups in polyaniline (PANI), shifted slightly to 3482 cm⁻¹ after adsorption. This suggested the formation of hydrogen bonds between the amine groups and the adsorbed metal ions (e.g., –NH···Pb²⁺), with a form of physical interaction. Similarly, the S=O stretching band of the sulfonate group (–SO₃⁻) in SDS, which appeared at 973 cm⁻¹, shifted to 953 cm⁻¹ post-adsorption. This shift was likely due to electrostatic attraction between the negatively charged sulfonate groups and the positively charged metal ions forming outer-sphere complexes (ion pairing) rather than inner-sphere (coordination) complexes. These types of interactions are typical of physical adsorption, where the adsorbate remains on the surface without penetrating into the electron cloud or bonding orbitals of the adsorbent. Moreover, slight shifts in other bands, such as the C–N stretching band shifting from 1239 to 1245 cm⁻¹, indicated that nitrogen atoms coordinated with metal ions through lone pair donation, forming metal–ligand complexes upon chemical adsorption through chelation. The shift in the C=C stretching band of benzenoid rings from 1301 to 1291 cm⁻¹ and that of quinoid rings from 1497 to 1484 cm⁻¹ could also be attributed to changes in local electronic environments caused by proximity to adsorbed ions due to electrostatic interactions and direct bonding. These shifts reflected strong interactions or slight polarization effects induced by the presence of metal ions near the polymer matrix. Furthermore, the peaks at 1571 and 1392 cm⁻¹, related to C=O, C=C, or C–H vibrations on the activated carbon surface, shifted to 1571 and 1385 cm⁻¹, respectively, implying that oxygen-containing groups participated in the interaction with metal ions. These spectral changes collectively indicated that chemical interactions, particularly chelation with nitrogen sites, were the dominant mechanism in Pb²⁺ adsorption, while physical mechanisms such as hydrogen bonding, electrostatic attraction, and van der Waals forces played supplementary roles. Although the SDS surfactant partially shielded some nitrogen sites, it did not entirely block the chelation process, allowing both types of interactions to proceed effectively without significant structural changes to the composite.

3.1.2. XRD Pattern Analysis

XRD spectroscopy was utilized to identify the crystal structures of activated carbon (AC), AC/Fe₃O₄, PANI-SDS, and the AC/Fe₃O₄/PANI-SDS composite, as illustrated in Figure 3. The two main wide peaks in the activated carbon XRD pattern were located at 2θ = 20.99° and 2θ = 43.64°, respectively, and were attributed to reflections from the (002) and (100) planes. The amorphous nature of activated carbon was confirmed by these peaks, which were typical of the graphitic hexagonal structure of carbon-based materials [48,49]. In the spectrum of the AC/Fe₃O₄ composite (Figure 3b), the characteristic peaks of AC disappeared, confirming that the stacking of AC sheets in the composite was almost disordered within the composite. Nevertheless, it displayed noticeable peaks at 2θ values of 30.04° (220), 35.25° (311), 43.06° (400), 57.05° (511), and 62.60° (440), which, according to the JCPDS database [50], indicated the cubic inverse spinel structure of magnetite. This confirmed the successful integration of Fe₃O₄ nanoparticles within the composite matrix. The XRD pattern of the PANI-SDS showed two large peaks at 2θ = 19.73° and 25.35° and a less noticeable Bragg diffraction peak at 16.03°. These peaks illustrated the semi-crystalline nature of PANI and correlated with the (020), (200), and (011) crystal faces of emeraldine, respectively [51]. Furthermore, as shown in Figure 3d, the AC/Fe₃O₄/PANI-SDS composite’s X-ray diffraction pattern exhibited the absence of the AC diffraction peak but the presence of distinctive peaks for PANI and Fe₃O₄. The absence of the AC peak indicated that the AC sheets were totally covered by the polymer and Fe₃O₄ nanoparticles, indicating the successful synthesis of the composite [52]. The composite’s overall crystallinity was often lowered by polyaniline’s (PANI) amorphous character, as seen by the diminished sharpness and intensity of the Fe₃O₄ peaks. However, as shown by the persisting unique peaks in the XRD pattern, the integration of PANI with AC/Fe₃O₄ did not alter the intrinsic crystalline structure or topological properties of the Fe₃O₄ nanoparticles. The well-dispersed AC/Fe₃O₄ composite provided surface sites for the deposition of PANI, enabling a more regular arrangement of the polymer and increasing the overall crystallinity of the composite. This showed substantial interactions between PANI and AC/Fe₃O₄, probably through π–π stacking and electrostatic forces, which permitted the effective combination of these components [53].

An XRD study was performed to assess the structural characteristics of the samples. The Scherrer formula was used for estimating the size of the crystallite:

$$\mathrm{D}= {\displaystyle \frac{\mathrm{K}\cdot \mathrm{\lambda }}{\mathrm{\beta }\cdot \mathrm{c}\mathrm{o}\mathrm{s}\mathrm{\theta }}}$$

(18)

In this equation, D is the average crystallite dimension, K is the shape factor (commonly taken as 0.9), λ is the wavelength of the X-rays used (0.154 nm for Cu Kα radiation), β is the full width at half maximum (FWHM) of the diffraction peak in radians, and θ denotes the Bragg angle corresponding to the peak position.

The pure Fe₃O₄ nanoparticles had an average size of 16.7 nm. After integrating with activated carbon and polyaniline (AC/Fe₃O₄/PANI-SDS), the size decreased to 11.04 nm. This change was likely because the polymer layer put pressure on the particles, causing strain and reducing their ability to grow, which led to smaller crystallite sizes.

3.1.3. SEM Analysis

Figure 4 illustrates the surface morphology of the AC, AC/Fe₃O₄, PANI-SDS, and AC/Fe₃O₄/PANI-SDS samples. The images provide important information about the surface architecture, texture, and modifications induced by the incorporation of various components. As seen in Figure 4a, the surface of the activated carbon (AC) displayed an irregular morphology characterized by uneven pores and a rough, porous texture. The SEM image of the AC/Fe₃O₄ sample (Figure 4b) showed the integration of Fe₃O₄ nanoparticles onto the surface of the activated carbon. The Fe₃O₄ nanoparticles appeared as small, spherical particles dispersed across the porous carbon matrix. This modification enhanced the material’s magnetic properties, making it easier to separate from aqueous solutions after adsorption. Figure 4c presents the structural features of PANI-SDS. As observed, the surface of polyaniline-based material displayed layers of flakes with a fibrous or granular morphology. The incorporation of SDS likely improved the material’s conductivity and stability, enhancing its adsorption performance. Figure 4d illustrates the morphological features of the AC/Fe₃O₄/PANI-SDS ternary composite, which exhibited a rough texture and a heterogeneous surface characterized by irregular pore sizes and distinct cavities. These pores were likely to contribute significantly to the adsorption of pollutants. The image of the AC/Fe₃O₄/PANI-SDS composite after Pb²⁺ ion adsorption shows significant changes in the surface morphology (Figure 4e). The pores and surface texture appeared less distinct, suggesting that the Pb²⁺ ions were successfully adsorbed onto the material. This was further evidence of the composite’s effectiveness in the elimination of hazardous metal ions from solution.

3.1.4. EDS Analysis

The elemental mass compositions of AC/Fe₃O₄, PANI-SDS, and the AC/Fe₃O₄/PANI-SDS composite were determined via EDS, as illustrated in Figure 5, and the results are collected in

Table 1. Elemental mass composition of the synthesized samples based on EDS analysis.

Sample	Fe (%)	O (%)	N (%)	C (%)	S (%)	Pb (%)
AC/Fe3O4	69.05	23.78	--	7.17	--	--
PANI-SDS	--	17.63	24.67	49.84	7.86	--
AC/Fe3O4/PANI-SDS	2.74	18.77	29.24	45.91	3.34	--
AC/Fe3O4/PANI-SDS after Pb2+ ion adsorption	4.68	21.64	16.77	33.22	3.83	20.31

. The EDS spectrum of AC/Fe₃O₄ (Figure 5a) exhibited the major elements of iron (Fe), oxygen (O), and carbon. As shown in Figure 5b, the composition of PANI-SDS was identified as carbon (C), nitrogen (N), oxygen (O), and sulfur (S), indicating that PANI was effectively functionalized using SDS. Additionally, Figure 5c displays the elements of the AC/Fe₃O₄/PANI-SDS composite, which includes iron (Fe), carbon (C), nitrogen (N), oxygen (O), and sulfur (S). These EDS results demonstrated the stepwise synthesis and functionalization of the composite, from AC/Fe₃O₄ to AC/Fe₃O₄/PANI-SDS. Furthermore, Figure 5d shows the composition of elements for the AC/Fe₃O₄/PANI-SDS composite after lead ion adsorption. The effective adsorption of metal ions on the surface of AC/Fe₃O₄/PANI-SDS was confirmed by the appearance of the Pb element in the EDS analysis.

3.1.5. VSM Analysis

Figure 6 illustrates the results of an additional analysis using a vibrating sample magnetometer (VSM) to examine the magnetic behavior of AC/Fe₃O₄ and the AC/Fe₃O₄/PANI-SDS composite. Because there was no hysteresis loop where the coercivity and remanence were zero, the magnetization curves for these samples were superparamagnetic based on the reported results [54]. Furthermore, the AC/Fe₃O₄/PANI-SDS composite’s magnetic saturation (18.34 emu/g) was significantly lower than that of AC/Fe₃O₄ (55.63 emu/g). This decrease may have been the result of the influence of both AC and the non-magnetic PANI-SDS matrix. However, it should be highlighted that the AC/Fe₃O₄/PANI-SDS composite had sufficient magnetization to enable efficient separation from solution using an external magnetic field [55,56].

3.1.6. Surface Area and Pore Structure Analysis

The texture properties of the synthesized materials were tested by means of nitrogen sorption–desorption isotherms. From Figure 7, we may note that the isotherms for all of the prepared materials are type IV and the presence of emphatic hysteresis loops for the samples indicates the pore characteristic of the prepared materials. The BET surface areas, micropore volumes, and BJH average pore widths are presented in

Table 2. Textural parameters of AC, AC/Fe₃O₄, and AC/Fe₃O₄/PANI-SDS composite.

Sample	AC	AC/Fe3O4	AC/Fe3O4/PANI-SDS
BET surface area (m2/g)	1844.52	444.92	284.63
Total pore volume (cm3/g)	0.7475	0.1639	0.1062
Pore diameter (nm)	5.41	15.94	13.35

for the unmodified and modified materials. The unmodified activated carbon (AC) had the highest surface area of 1844.52 m²/g and a rather high micropore volume of 0.7475 cm³/g, all of which reflected the highly porous characteristic of AC and its capacity to offer abundant surface sites for adsorption. The BJH average pore diameter was 5.41 nm, indicating the presence of a well-developed porous framework with relatively narrow internal voids. After the incorporation of Fe₃O₄ nanoparticles (AC/Fe₃O₄), the BET surface area decreased to 444.92 m²/g and the micropore volume decreased to 0.1639 cm³/g. These decreases were explained by the partial blockage or partial occupation of micropores by Fe₃O₄ nanoparticles. But the pore diameter increased to an average of 15.94 nm, which indicated that larger pores were formed, or rearrangement of the pore system took place during the incorporation of the nanoparticles. Additionally, the surface area (284.63 m²/g) and micropore volume (0.1062 cm³/g) decreased upon further functionalization with polyaniline and SDS (AC/Fe₃O₄/PANI-SDS) due to additional coverage of the pore walls and entrances by the polymer and surfactant layers. The mean pore diameter also became smaller by a small margin to 13.35 nm but was still much larger than that of pure AC, meaning that despite the reduction in the surface area, a portion of the larger pores were still available. Also, due to the existence of the functional groups of PANI and the dispersing effect of SDS, specific interactions with metal ions, such as electrostatic attraction and coordination, were likely to be enhanced. These results indicated that while the surface area and micropore volume were reduced with each modification step, the porous architecture remained open to a considerable extent, which was advantageous for applications involving the adsorption of metal ions.

3.2. Evaluation of Adsorption Behavior

3.2.1. Influence of Adsorbent Type

The elimination of Pb²⁺ ions from water was examined using AC, PANI-SDS, and the AC/Fe₃O₄/PANI-SDS composite. The AC/Fe₃O₄/PANI-SDS composite demonstrated significantly higher removal efficiency than pure polyaniline and activated carbon under identical conditions, with an initial Pb²⁺ concentration of 25 mg/L. This enhanced performance resulted from the high surface density of functional adsorption sites within the composite, which facilitated stronger interactions with Pb²⁺ ions. Furthermore, the introduction of Fe₃O₄ nanoparticles in the composite improved Pb²⁺ ion adsorption in solution, which could be linked to stronger electrostatic attractions and hydrogen bonding between the AC/Fe₃O₄/PANI-SDS composite and metal ions in solution. The AC/Fe₃O₄/PANI-SDS composite was chosen as the main adsorbent in this study because of its high efficiency, which suggested that it has the potential to be a material that efficiently removes heavy metals.

3.2.2. Interaction Time

The interaction duration between adsorbent and solution is a key factor in effective contaminant removal. It offers an estimate of how long metal ions may contact the adsorbent surface during adsorption to reach equilibrium. It is only via sufficient contact time that maximal adsorption capacity may be obtained. When the contact duration is insufficient, the adsorption process is incomplete, and metal ions cannot be removed from solutions efficiently. Batch studies were conducted to comprehend the adsorption process’s equilibrium behavior. The effect of contact duration on Pb²⁺ ion adsorption was investigated using 0.2 g/L (0.01 g) of the AC/Fe₃O₄/PANI-SDS adsorbent. Figure 8a illustrates the adsorptive effectiveness of varying Pb²⁺ concentrations with time (0–220 min) using the AC/Fe₃O₄/PANI-SDS composite. The data clearly showed that removal efficiency increased rapidly with extended contact time, eventually reaching a plateau where further time had little impact on performance. Initially, the composite offered abundant active sites for ion interaction, resulting in a sharp increase in Pb²⁺ removal. This continued until equilibrium was established, beyond which the adsorption rate stabilized [57,58]. The interaction of Pb²⁺ ions with the surface of the AC/Fe₃O₄/PANI-SDS composite likely caused the quick initial removal of Pb²⁺ ions. There were several factors that contributed to the efficient adsorption of Pb²⁺ ions, including the presence of functional groups on the surface of the AC/Fe₃O₄/PANI-SDS composite [59], as well as the availability of unoccupied adsorption sites [60]. This behavior was consistent with that previously observed for the removal of various pollutants, such as organic compounds [61], Pb (II), and Cd (II) [62], as well as Cr (IV) [63].

3.2.3. Adsorbent Dosage

A crucial feature of an efficient adsorbent is its ability to achieve high removal efficiency quickly and with minimal material usage. To investigate how the weight of the adsorbent affected the elimination of Pb²⁺ ions, 50 mL of solution containing 25 mg/L of Pb²⁺ ions was used in the experiments, which was maintained at pH 6. The dosage of the AC/Fe₃O₄/PANI-SDS composite varied between 0.05 and 0.4 g/L, with a set adsorption time of 220 min. As seen in Figure 8b, the data revealed that when the adsorbent dosage increased from 0.05 to 0.2 g/L, the Pb²⁺ ion elimination efficacy grew dramatically from 45.12% to 95.12%. This enhancement resulted from the composite surface’s enhanced availability of active adsorption sites, which favored Pb²⁺ ion binding [64]. Higher dosages provided a larger surface area for the AC/Fe₃O₄/PANI-SDS composite, improving the probability of adsorption and accelerating the removal process [65]. However, the effectiveness of elimination reached its maximum at a dosage of 0.2 g/L (0.01 g). At this stage, all available adsorption sites were occupied by Pb²⁺ ions and further increases in dosage did not yield additional improvements [66]. It is essential to optimize the dose of the adsorbent in order to decrease material consumption and increase efficiency. The best dosage for removing the most Pb²⁺ ions while reducing adsorbent waste was found to be 0.2 g/L (0.01 g) in this investigation.

3.2.4. Initial Concentration

The effect of starting metal ion concentration on the removal elimination efficacy of the AC/Fe₃O₄/PANI-SDS composite was examined throughout a range of 10–110 mg/L. As illustrated in Figure 9a, the adsorption percentage of lead ions by the adsorbent exhibited a distinct trend with rise initial concentration. At lower Pb²⁺ concentrations, a high adsorption percentage was observed, which could be associated with the higher availability of functional sites on the composite surface compared to the number of metal ions present. However, when the initial lead ion concentration grew, the adsorption percentage rapidly fell, despite the increasing absolute quantity of ions being adsorbed. The lower effectiveness of removal may have been due to the complete occupation of available sites on the composite surface at increasing concentrations, which lowered the availability of functional sites for additional ions. Under the experimental conditions of 50 mL of solution volume, 25 mg/L lead ion concentration, 0.2 g/L (0.01 g) adsorbent dosage, pH 6, and an equilibrium time of 220 min, the removal percentage exceeded 95.12%. The adsorption percentages decreased to a lower value when the lead ion concentration was raised from 10 to 110 mg/L. As a result, there was clearly a negative relationship between the starting concentration of lead ions and the adsorption percentage, emphasizing the importance of optimizing concentrations to maximum removal efficiency. Comparing the lead ion adsorption percentages noticed in this investigation and those documented in earlier studies showed that the AC/Fe₃O₄/PANI-SDS composite performed better (Table 6).

3.2.5. Solution pH and Zeta Potential

As a crucial operating parameter in the adsorption process, the solution pH impacts the surface characteristics, shape, and chemical aspects of heavy metal ions in solution [67]. The influence of pH (ranging from 2 to 10) on Pb²⁺ ion elimination efficiency was evaluated in the present work, as seen in Figure 9b. The lowest adsorption performance (17.74%) by the AC/Fe₃O₄/PANI-SDS composite was observed at pH 2. As the pH increased, the elimination efficiency rose sharply, reaching a maximum of 95.12% at pH 6, after which it began to decline. Based on this finding, pH 6 was selected as the best-performing condition for adsorption. The reduced efficiency in the acidic pH range could be associated with the high concentration of H⁺ ions, which competed with Pb²⁺ ions for active sites on the adsorbent surface, thereby limiting lead ion uptake. Additionally, it was easy to determine that the zero-charge point of the adsorbent was 5.09, implying that the positively charged surface of the adsorbent caused electrostatic repulsion with Pb²⁺ ions in solution when pH < 5.09 [68]. As a consequence, the optimum adsorption efficacy was found to be at pH 6. With an increase in pH, the hydrogen ion concentration in the solution declined, which in turn lessened competition for available adsorption sites and gave metal ions a greater advantage in binding to the adsorbent. The nitrogen-containing and hydroxyl functional groups on the adsorbent acted as soft bases, making them particularly effective in interacting with soft metal ions, which likely contributed to the highest removal efficiency observed at pH 6. These amine and hydroxyl groups were capable of forming strong bonds with metal ions through their available lone electron pairs, enabling both physical and chemical interactions. Such interactions facilitated the effective extraction of metal ions from solution [69]. As the pH increased, the slight decrease in Pb²⁺ removal efficiency at pH 6–8 could be explained by the changing speciation of lead. In this pH range, Pb²⁺ and Pb(OH)⁺ were the dominant species, both positively charged, and could still interact electrostatically with the negatively charged composite surface. Above pH 8.0, the adsorption of lead was mainly attributable to the mixed effect of adsorption and formation of insoluble lead hydroxide (Pb(OH)₂). Therefore, adsorption experiments conducted at pH > 6 were not used for isothermal or kinetic modeling to avoid interference from precipitation effects. All adsorption parameters reported in this study were derived from experiments performed at pH = 6, where Pb²⁺ remained predominantly soluble and adsorption was the dominant removal mechanism. Similar trends have been observed in other studies [70,71,72,73]. The point of zero charge (pH_PZC) for the AC/Fe₃O₄/PANI-SDS adsorbent was identified between pH 2 and 12 utilizing the zeta potential technique, as depicted in Figure 10a. This suggested that the adsorbent’s surface might be somewhat negatively or positively charged depending on the medium. As the pH increased, the surface charge moved from positive to negative as the degree of protonation of its functional groups decreased. Additionally, at a pH of 5.09, AC/Fe₃O₄/PANI-SDS revealed a point of zero charge, demonstrating that the material was negatively charged under alkaline and neutral circumstances. As illustrated in Figure 10b, the main Pb²⁺ species in aqueous solutions varied with pH, including Pb²⁺, Pb (OH)⁺, Pb (OH)₂, and Pb (OH)₃⁻. Within the pH range of 1 to 6, Pb²⁺ was the predominant form. Between pH 6 and 8, both Pb²⁺ and Pb(OH)⁺ were prevalent. In this range, lead removal was primarily driven by electrostatic attraction between the positively charged lead species and the negatively charged surface of the nanocomposite. At pH values above 8.0, lead adsorption resulted from a combination of surface adsorption and the precipitation of insoluble lead hydroxide (Pb(OH)₂).

3.3. Adsorption Kinetics Analysis

Adsorption of Pb²⁺ ions occurred mainly at the surface of the AC/Fe₃O₄/PANI-SDS composite, where interactions dictated the adsorption mechanism. The adsorption kinetics were examined using three models, including pseudo-first-order, pseudo-second-order, and intraparticle diffusion, to examine the relationship between adsorption rate and concentration [74,75]. The adsorption kinetics of the AC/Fe₃O₄/PANI-SDS composite were examined at various initial concentrations of Pb²⁺ ions. A quick uptake of Pb²⁺ ions occurred within the first 20 min, then there was a slow decline in the adsorption rate until an equilibrium was reached. This initial fast adsorption could be attributed to the composite’s favorable structural features. Compared to conventional adsorbents, the combination of a well-developed internal structure, high specific surface area, and functionalization with PANI-SDS offered an increased number of active and stronger binding sites, facilitating quicker and more efficient Pb²⁺ ion capture. Linearized and nonlinearized versions of the pseudo-first-order and pseudo-second-order models for 25 mg/L Pb²⁺ ion adsorption on the composite are illustrated in Figure 11a,b. According to the regression coefficients (R²) and error analysis in

Table 3. Adsorption kinetics data, including model parameters for Pb²⁺ ions.

Pb2+ Ions (mg/L)		Pseudo-First-Order			Pseudo-Second-Order			Intraparticle Diffusion
Linear Form
	qe·exp (mg/g)	K1 (min−1)	R2	qe·cal (mg/g)	K2 × 10−4 (g/mg min)	R2	qe·cal (mg/g)	Kp1 (mg/g/min)	R2	Kp2 (mg/g/min)	R2
10	48.31	0.086	0.964	33.45	21.06	0.999	50.86	7.23	0.924	2.71	0.945
25	118.90	0.090	0.958	89.39	8.29	0.999	125.47	15.16	0.918	4.98	0.996
50	222.06	0.108	0.988	182.13	4.93	0.999	233.64	42.25	0.975	12.13	0.992
80	291.61	0.093	0.939	327.16	2.54	0.999	310.55	50.09	0.988	14.83	0.940
110	332.95	0.083	0.989	374.79	4.37	0.999	344.82	57.37	0.978	18.95	0.995
Nonlinear form
10	48.31	0.066	0.968	46.98	19.30	0.992	51.33	-	-	-	-
25	118.90	0.065	0.970	115.71	7.56	0.993	126.66	-	-	-	-
50	222.06	0.066	0.989	218.42	4.06	0.994	236.04	-	-	-	-
80	291.61	0.058	0.968	278.87	2.65	0.996	308.81	-	-	-	-
110	332.95	0.096	0.939	320.01	4.64	0.988	342.48	-	-	-	-

, the pseudo-second-order model demonstrated high R² values, indicating the closest match to the experimental data. This was further substantiated by the good match between the experimentally measured equilibrium adsorption capacity (q_e) values and those computed using the pseudo-second-order model. Consequently, this model was found to be the best suited for modeling the adsorption of Pb²⁺ ions onto the composite. To better understand the adsorption kinetics, the intraparticle diffusion model was employed. The plot of q_t versus t^0.5 (Figure 11c) was used to determine the intraparticle diffusion rate constant and the boundary layer thickness constant, as summarized in Table 3. If the plot had intersected the origin, it would have indicated that intraparticle diffusion was the primary controlling mechanism. However, the curve did not cross the origin, showing that while intraparticle diffusion was important, further steps in the adsorption process also affected the total adsorption rate. The adsorption of Pb²⁺ ions onto the composite surface proceeded through three distinct stages. Initially, the process was governed by boundary layer diffusion. This was followed by a step where intraparticle diffusion became dominant as the ions migrated toward active sites within the adsorbent, representing the rate-limiting step. Finally, the system reached an equilibrium, where the adsorption sites became fully saturated. As demonstrated in Table 3, the first step had the largest diffusion rate constant (K_p), whereas the second step had the lowest. Initially, the abundance of accessible adsorption sites on the composite surface enabled fast external diffusion of metal ions. However, once these surface sites were saturated, the metal ions began to diffuse into the pores, resulting in a decreased adsorption rate and reduced adsorption force. In the end, the adsorption process found an equilibrium [76]. Adsorption was a multifaceted process that involved both surface interactions and intraparticle diffusion, each playing a role in the overall kinetic behavior. The findings emphasized the importance of considering multiple factors when modeling the adsorption of metal ions onto composite materials.

3.4. Adsorption Isotherm

Lead ion (Pb²⁺) adsorption on the AC/Fe₃O₄/PANI-SDS composite was analyzed using linear and nonlinear approaches using Langmuir, Freundlich, and Dubinin–Radushkevich (D–R) isotherm models, as seen in Figure 12a–c.

Table 4. Isotherm parameters of Pb²⁺ ions on the surface of AC/Fe₃O₄/PANI-SDS composite.

Isotherm	Form	Parameter	Value
Langmuir	Linear form	KL (L mg−1)	0.360
		qmax (mg g−1)	348.39
		R2	0.996
	Nonlinear form	KL (L mg−1)	0.402
		qmax (mg g−1)	336.80
		R2	0.986
Freundlich	Linear form	1/n	0.383
		KF (mg g−1)	91.81
		R2	0.915
	Nonlinear form	1/n	0.297
		KF (mg g−1)	113.60
		R2	0.940
Dubinin–Radushkevich (D–R)	Linear form	qs (mg g−1)	250.76
		E (KJ mol−1)	18.36
		R2	0.843
	Nonlinear form	qs (mg g−1)	291.20
		E (KJ mol−1)	13.87
		R2	0.886

presents a summary of the findings, which includes the correlation coefficients (R²). Among the tested isotherm models, the Langmuir model showed the highest correlation coefficient (R² = 0.996) within the investigated concentration range. However, considering the composite nature of the adsorbent, this result should be interpreted as an apparent mathematical fit rather than definitive evidence for a strictly homogeneous monolayer adsorption mechanism. The fitting suggested predominant Langmuir-type behavior under the studied experimental conditions, with a maximum adsorption capacity of 348.39 mg/g. In addition, the calculated value of 1/n for the adsorption of Pb²⁺ ions on the AC/Fe₃O₄/PANI-SDS surface was 0.383, which confirmed that Pb²⁺ ions were favorably adsorbed on the surface. Furthermore, the Dubinin–Radushkevich (D–R) isotherm model, predicated on the mean free energy of adsorption (E), was employed to assess the nature of the adsorption process as either physical or chemical. According to Table 4, the calculated E value of 18.36 kJ/mol was consistent with standard values for chemical adsorption. The results indicated that Pb²⁺ ions bonded to the AC/Fe₃O₄/PANI-SDS composite mainly through chelation interactions.

The proposed adsorption mechanisms of the AC/Fe₃O₄/PANI-SDS composite for Pb²⁺ ions showed multiple synergistic pathways, likely involving chemical and physical interactions, as illustrated in Figure 13. Chemical interactions included coordination interactions, ion exchange, hydrogen bonding, and electrostatic attractions, which led to stronger and more specific binding. The chemical attraction between the negatively charged surface groups (amine, –SO₃⁻, –OH, and –COOH) and Pb²⁺ ions played a primary role, enhancing the initial binding affinity. Additionally, hydrogen bonding formed between electronegative hydroxyl or amine groups on the composite surface and the electropositive hydrogen in metal hydroxide species, providing strong, reversible intermolecular forces that significantly increased adsorption capacity and kinetics [77]. Ion exchange also contributed to Pb²⁺ removal, likely involving the displacement of H⁺ ions from functional groups by Pb²⁺, particularly under acidic to neutral conditions [78]. Coordination (chelation) between Pb²⁺ ions and nitrogen atoms within the polyaniline (PANI) rings further stabilized the adsorption process. According to Dubinin–Radushkevich (D–R) isotherm analysis, the calculated adsorption energy (E_D-R) for Pb²⁺ ions was higher than 8 kJ/mol, demonstrating that the process followed an electrostatic and chemical adsorption mechanism. Physical interactions also contributed notably to the adsorption process. These included electrostatic attraction between charged functional groups and Pb²⁺, van der Waals forces, pore sorption, and physical entrapment (pore filling). While van der Waals interactions are weaker than hydrogen bonds, they supported overall adsorption, especially at high temperatures. The porous structure of the composite facilitated physical adsorption through pore diffusion and filling. Hydrogen bonding and pore-filling effects played dominant roles among the physical mechanisms, while electrostatic interactions enhanced surface-level binding. As shown in Figure 2b, many peaks in the AC/Fe₃O₄/PANI-SDS composite shifted after Pb²⁺ adsorption, suggesting structural interaction but without major disruption, consistent with both reversible physical adsorption and stable chemical binding. Although the SDS surfactant partially shielded the nitrogen sites, it did not fully prevent chelation, allowing both chemical and physical interactions to proceed. Finally, the high density of accessible adsorption sites and the pH-dependent protonation–deprotonation behavior of surface functional groups further improved Pb²⁺ ion accessibility and selectivity, consistent with the cooperative role of both chemical and physical mechanisms in achieving high adsorption capacity.

3.5. Thermodynamic Parameters and Temperature Effect

To evaluate the impact of temperature and related parameters on the thermodynamics on the adsorption of Pb²⁺ ions by the composite, experiments were conducted with a 50 mL volume of Pb²⁺ ion solution (25 mg/L) at pH 6 and a duration of 220 min. The efficiency of removal and the capacity of adsorption increased with increasing temperature. These results indicated that endothermic interplay was the predominant process of adsorption because stronger adsorbent–adsorbate connections increased with an increase in temperature. The thermodynamic parameters, including the apparent changes in enthalpy (ΔH°), Gibbs free energy (ΔG°), and entropy (ΔS°), were determined using the distribution coefficient (K_d = q_e/C_e) and are summarized in

Table 5. Apparent thermodynamic parameters of Pb²⁺ ion adsorption on AC/Fe₃O₄/PANI-SDS composite.

Temperature (°k)	ΔG° (kJ mol−1)	ΔH° (kJ mol−1)	ΔS° (J mol−1 K−1)
298	−11.34	47.94	198.65
303	−12.13
308	−13.21
313	−14.30

. The positive apparent ∆H° value (47.94 kJ /mol) indicated that the adsorption enthalpy of the Pb²⁺ ions on the surface of the AC/Fe₃O₄/PANI-SDS composite was chemical and endothermic. Additionally, the negative ΔG° values across different temperatures indicated the spontaneity of the adsorption process. The high negative ΔG° at higher temperatures suggested that adsorption was more favorable at high temperatures. Furthermore, the positive ΔS° value signified an increase in system randomness at the solid–liquid interface during adsorption. Overall, the results confirmed that the process was chemical, endothermic, and becomes more efficient as temperature increased.

3.6. A Comparative Adsorption Study Toward Various Heavy Metal Ions

We systematically assessed the adsorption performance of the AC/Fe₃O₄/PANI-SDS composite toward toxic heavy metal ions, including Pb²⁺, Cd²⁺, Co²⁺, Ni²⁺, and Zn²⁺, in a single-metal system under controlled conditions (Figure 14a). All metal ion solutions were prepared at an initial concentration of 25 mg/L, with the pH adjusted to 6 and the temperature maintained at 298 K. A fixed dose of 0.2 g/L (0.01 g) of the adsorbent was introduced into 50 mL of each solution. As presented in Figure 14a, the composite exhibited remarkable selectivity toward Pb²⁺, achieving an adsorption efficiency of 95.12%, which was considerably higher than those for Cd²⁺ (33.97%), Co²⁺ (21.72%), Ni²⁺ (12.86%), and Zn²⁺ (8.68%). This preferential adsorption behavior could be attributed to several key physical and chemical factors, as supported by previous studies on similar adsorbents [79,80,81]. Chemically, Pb²⁺ had a strong tendency to form stable inner-sphere complexes with functional groups present on the adsorbent surface, such as –OH, –COOH, –SO₃⁻, and nitrogen-containing sites from PANI. Its moderate electronegativity (2.33) and high polarizability enhanced its ability to accept electron pairs, leading to partial covalent bonding (chelation) and surface complexation, even in the presence of SDS, which may have partially shielded some active sites. Physically, Pb²⁺ possesses a relatively large ionic radius (1.19 Å) and low hydration energy (−1481 kJ/mol), which reduced the energy barrier for dehydration and facilitated diffusion toward the adsorbent surface. These properties promoted stronger electrostatic attractions, van der Waals forces, and ion–dipole interactions with the negatively charged and porous surface of the composite. By contrast, metal ions such as Ni²⁺, Zn²⁺, and Co²⁺ exhibit higher hydration energies, smaller radii, and lower polarizability, which hindered both their chemical coordination and physical adsorption [82,83,84,85,86]. The porous structure of AC and PANI also favored Pb²⁺ due to its lower hydration competition, while ions like Ni²⁺ and Zn²⁺ remained strongly hydrated and less accessible to adsorption sites. Consequently, while physical interactions including electrostatic attraction, hydration competition, and ion–dipole forces played a predominant role in the overall adsorption process, chemical mechanisms such as chelation also contributed significantly. Therefore, the superior uptake of Pb²⁺ was attributed to its strong coordination chemistry combined with enhanced physical accessibility to active adsorption sites [87].

Additionally, in multi-metal systems (Figure 14b), competitive adsorption was observed, with reduced overall removal efficiency of lead ions compared to the single-metal system due to competitive effects among coexisting ions. Notably, Cd²⁺ exhibited a stronger inhibitory effect on Pb²⁺ adsorption than Co²⁺. Nevertheless, AC/Fe₃O₄/PANI-SDS maintained superior selectivity toward Pb²⁺, with the observed order of Pb²⁺ > Cd²⁺ > Co²⁺. This behavior was consistent with the previously discussed factors governing chemical and physical adsorption, particularly the dominant influence of hydration radius, ionic size, and polarizability, which enhanced its access to active sites over more hydrated ions. Overall, these findings confirmed that the selectivity of AC/Fe₃O₄/PANI-SDS toward Pb²⁺ was governed by a combination of physical and chemical adsorption mechanisms, with chelation adsorption being predominant yet complemented by potential physical interactions. Considering Pb²⁺’s high toxicity and environmental relevance, its selective removal remains critical. The achieved removal efficiency (95.12%) highlighted AC/ Fe₃O₄/PANI-SDS as a strong candidate for Pb²⁺ elimination from contaminated complex water systems.

3.7. Reusability Study of the Adsorbent

The long-term application of low-cost and effective adsorbents in wastewater treatment depends heavily on their reusability, which ensures both economic and practical feasibility. In this study, the adsorbent demonstrated successful regeneration and was reused over five consecutive cycles. Its adsorption performance was assessed through five successive adsorption–desorption cycles, as illustrated in Figure 15. The findings indicated that the removal efficiency of Pb²⁺ exhibited a slight decrease to 86.47% throughout five cycles. The relatively stable performance may have been related to the composite’s structure and porosity under the investigated conditions. These results suggested that the composite could serve as a potential adsorbent for wastewater treatment under the investigated conditions. The observed reduction in the composite’s performance could be associated with the incomplete elimination of residual metal ions from its active sites during the regeneration process. These remaining ions may block available binding sites, limiting the composite’s ability to adsorb additional metal ions in subsequent cycles.

Besides exhibiting adsorption performance, the AC/Fe₃O₄/PANI-SDS composite also exhibited acceptable operational regeneration and reusability potential. The incorporation of Fe₃O₄ enabled quick and efficient adsorbent recovery from aqueous solutions, reducing the need for filtration or centrifugation. The magnetic separation capability adds significant practical value during post-treatment operations due to its ability to simplify operational processes in large-scale or continuous water purification systems. Furthermore, Fe₃O₄ offers both magnetic functionality and efficient dispersion in aqueous solutions as it integrated well with the composite while facilitating efficient separation without obvious performance loss during the tested cycles. The composite showed acceptable capacity retention throughout several regeneration processes, suggesting preliminary reusability and potential operational longevity in lead removal from water; however, its long-term stability under repeated regeneration remains to be further investigated.

3.8. Comparison Studies with Another Research

The adsorption performance of the AC/Fe₃O₄/PANI-SDS adsorbent was compared with the capacity of other adsorbents described in the literature, and the findings are provided in Table 6. The results indicated that the adsorption capacity of the AC/Fe₃O₄/PANI-SDS composite toward Pb²⁺ ions was as high as 348.39 mg/g, which effectively highlighted its superiority compared to most of the other reported adsorbents under various conditions. It is noteworthy that this high capacity was achieved with a significantly lower adsorbent dose (0.01 g) and at a near-neutral pH (6) compared to many adsorbents in the table, which may suggest advantages in material efficiency and practical applicability. The strong uptake indicated the high affinity of the composite’s active sites for Pb²⁺ ions, supporting its potential as an effective material for water purification. The increased performance of the AC/Fe₃O₄/PANI-SDS adsorbent can be connected to special attributes, e.g., unique structural features, high surface area, and functional groups. Overall, while direct ranking is constrained by differing experimental parameters, these results demonstrate that the AC/Fe₃O₄/PANI-SDS composite is a highly efficient and promising candidate for real-world wastewater treatment applications, particularly for lead ion removal [88,89,90,91,92,93].

Table 6. Different types of adsorbent materials reported for the elimination of Pb²⁺ ions from wastewater.

Adsorbent Materials	Dose (g)	pH	Temp. (°C)	qmax (mg/g)	Removal%	Reference
Fe3O4/AC nanocomposite	0.04	5–6	20	144.92	97.66	[88]
Chitosan Poly (acrylic acid) porous polymer Nanocomposite	--	5	25	138.9	--	[89]
MCGO composite material	--	5	27	112.35	92.00	[90]
Polyaniline/Fe3O4	0.03	9	25	111.11	--	[91]
Magnetic carbon (Fe3O4@Carbon) nanocomposites	0.04	5.5	27	151.5	96.30	[92]
Carbon magnetic nanocomposites	1.0	5.9	25	83.54	84.50	[93]
AC/Fe3O4/PANI-SDS	0.01	6	25	348.39	95.12	This work

Table 6. Different types of adsorbent materials reported for the elimination of Pb²⁺ ions from wastewater.

Adsorbent Materials	Dose (g)	pH	Temp. (°C)	qmax (mg/g)	Removal%	Reference
Fe3O4/AC nanocomposite	0.04	5–6	20	144.92	97.66	[88]
Chitosan Poly (acrylic acid) porous polymer Nanocomposite	--	5	25	138.9	--	[89]
MCGO composite material	--	5	27	112.35	92.00	[90]
Polyaniline/Fe3O4	0.03	9	25	111.11	--	[91]
Magnetic carbon (Fe3O4@Carbon) nanocomposites	0.04	5.5	27	151.5	96.30	[92]
Carbon magnetic nanocomposites	1.0	5.9	25	83.54	84.50	[93]
AC/Fe3O4/PANI-SDS	0.01	6	25	348.39	95.12	This work

4. Conclusions

In this study, activated carbon (AC), magnetic iron oxide (Fe₃O₄), polyaniline (PANI), and sodium dodecyl sulphate (SDS) were successfully incorporated to generate a novel AC/Fe₃O₄/PANI-SDS composite. The composite was found to be a highly effective adsorbent for the elimination of Pb²⁺ ions from polluted water. Several techniques, namely FT-IR, XRD, SEM-EDS, BET, VSM, and zeta potential, were employed to characterize the AC/Fe₃O₄/PANI-SDS composite and estimate its structural, magnetic, and surface properties. These analyses ensured the specific structure of the composite, which combined a magnetic part to facilitate separation, a large surface area for adsorption, and the existence of functional groups to enhance interactions with Pb²⁺ ions. Under varying conditions (pH, dosage of adsorbent, duration time, and initial Pb²⁺ concentration), the adsorption effectiveness of the AC/Fe₃O₄/PANI-SDS composite was tested. The findings revealed that the composite has a great capability of removing Pb²⁺ ions from wastewater at a slightly acidic pH (around 6). The architecture of the composite material, which consists of PANI and SDS functional groups, enables it to interact strongly with Pb²⁺ ions, which facilitates high-efficiency adsorption. The pseudo-second-order kinetic model was found to illustrate the adsorption process, and isothermal analysis showed that the Langmuir model provided the highest correlation coefficient within the investigated concentration range. This fitting suggested apparent Langmuir-type adsorption behavior with a maximum adsorption capacity of 348.39 mg/g. Additionally, thermodynamic parameters showed that the adsorption is thermodynamically spontaneous and endothermic, hence proving the effectiveness of the composite. The relatively easy separation of the Fe₃O₄ component, based on its magnetic properties, enables easy magnetic isolation of the composite from the medium, allowing it to be regenerated and used over successive cycles with minimal loss in performance. A possible adsorption mechanism was proposed, with chemical chelation, pore filling, hydrogen bonds, and electrostatic interaction being involved in the elimination of Pb²⁺ ions. To conclude, the AC/Fe₃O₄/PANI-SDS composite is suitable for use as a convenient, cost-effective, and eco-friendly substitute for the elimination of Pb²⁺ ions that contaminate water. It has a high adsorption capacity and is easily recoverable and reusable, making it a suitable candidate that could be applied in the real world for treating wastewater and removing heavy metal contamination.