Synthesis of Poly(aniline-co-benzene)-Based Hypercrosslinked Polymer for Hg(II) Ions Removal from Polluted Water: Kinetic and Thermodynamic Studies

Abstract

The aim of this work was to investigate the adsorption performance of a highly crosslinked poly(aniline-co-benzene) (PAB) copolymeric network. This hypercrosslinked polymer (HCP) was obtained via the Friedel–Craft reaction in the presence of FeCl₃ as an alkylation catalyst. The HCP was characterized using FTIR, SEM, TGA-DTA-DSC thermograms, and BET surface area. The analysis revealed a major mesoporous (an average pore diameter of 4.96 nm) structure, a surface area of 987 m²/g, and adequate chemical and thermal stability, thus supporting its potential as an adsorbent. The PAB HCP capability as an adsorbent for removing mercury ions (Hg²⁺) from wastewater was examined, and the data obtained were kinetically and thermodynamically modeled. The data were found to fit PFO well (R² = 0.999), suggesting a physisorption process and a rate-limiting step involving the diffusion process, as proven with IPD and LFD models. The adsorption of Hg²⁺ on PAB was spontaneous (ΔG° is negative; −4.41 kJ/mol at 298 K), endothermic (ΔH° is positive; 32.39 kJ/mol), and random (ΔS° is positive; 123.48 J/mol·K) at the adsorption interface. The thermodynamic analysis also suggested a physical adsorption mechanism (ΔG° between −20 and 0 kJ/mol). These findings promote the potential application of PAB HCP as an efficient adsorbent for removing Hg²⁺ ions and other heavy metal ions from polluted environments.

1. Introduction

Water dirtiness has recently become a global issue—a major threat to human health and the environment. Pollutants could be chemicals, trash, or microbes, but chemical pollutants are considered the most harmful [1,2]. In this regard, heavy metal ions are highly toxic chemicals, even at low concentrations, due to their tendency to accumulate in living organisms, causing several health problems [3]. The primary concern is that heavy metals are nondegradable compared to organic pollutants; thus, they accumulate in living bodies and become highly dangerous over time. Mercury (II) (Hg²⁺) is one of the most poisonous metal ions ever known and is recognized for its acute and chronic toxicity to the central nervous system, kidneys, and lung tissues [3,4]. Several sources may contribute to mercury emission into the environment, including mining industries, power generation plants, paper manufacturing, rubber processing, and chemical and fertilizer industries, among others [5]. According to the United States Environmental Protection Agency (US EPA) and World Health Organization (WHO), the maximum allowable limit in drinking water was set to 1 and 0.03 μg/L [5,6], respectively. Therefore, removing heavy metals, including mercury, from a contaminated environment is essential and highly desirable for safe drinking water.

Different methods have been used to remove a heavy metal from polluted water and can be classified into membrane-, chemical-, electro-, photocatalytic-, and adsorption-based techniques [7,8]. The adsorption process is extensively used to remove heavy metals from wastewater owing to its low cost, availability, and eco-friendly nature. Adsorbents can be of various types, of which bio-adsorbents, carbon-based, carbon-nanotube, and polymeric-based adsorbents are the most common [9]. For the adsorptive elimination of Hg²⁺ from aquatic environments, various materials were tried [10,11]. It can be performed with natural materials and hybrid materials based on some metal oxide nanoparticles and their functionalized forms. Among all, activated carbon is the most popular and efficient [12]. On the other hand, it is stated that polymeric adsorbents are an important class of adsorbent materials, demonstrating high removal capability for capturing and removing dyes and heavy metal ions from wastewater [13,14]. Hence, various polymeric adsorbents such as polyaniline, polypyrrole and its copolymer and composites [15,16], poly-ε-caprolactone [17], amine-functionalized activated carbon [18], and hypercrosslinked polymers (HCPs) [19,20] have been prepared for the adsorption of heavy metal ions. The HCPs feature high uptake capacity, applicability for large-scale use, recyclability, and stability [20,21]. They are a class of porous polymers, advanced with a high surface area, well-distributed porosity, modest density, functionalized ability, eco-friendly characteristic, and easy fabrication [22].

An aniline-based HCP adsorbent has been synthesized and applied to remove radioactive iodine [23], reporting a highly efficient performance. A hyper-branched polypyrrole–polyethyleneimine nano-adsorbent was utilized for methylene blue (cationic dye) elimination from industrial effluent, resulting in an improved adsorption mechanism compared to its precursor [24]. Dichloroxylene- [25] and hydroquinone-based [26] HCPs have been fabricated and utilized for aniline adsorption, showing a high capacity of about 769 and 167 mg/g, respectively. Phenyl-based precursors of benzene, benzyl alcohol, aniline, biphenyl, and 1,3,5-triphenylbenzene were also prepared and their removal of benzene from cigarette smoke was reported with a great performance [27]. Polystyrene-based HCPs were prepared with the Friedel–Craft reaction and used to remove cadmium (II) Cd²⁺ ions and a binary system of nickel (II) Ni²⁺ and lead (II) Pb²⁺ ions from wasted water [20,28]. Under optimum conditions, the high uptake capacities (950, 147, and 137 mg/g) of HCPs toward Cd²⁺, Ni²⁺, and Pb²⁺, respectively, were attributed to their desirable surface charge (at pH > 4), substantial surface area, and suitable porous cavities for adsorbate accommodation.

Poly(aniline-co-benzene) HCPs were also synthesized [29] and have shown attractive surface properties. However, to our knowledge, no studies have been reported concerning their application as an adsorbent for capturing chemicals from polluted environments, including heavy metals in an aquatic environment. Therefore, this study aimed to synthesize, characterize, and examine PAB HCPs in the removal of Hg²⁺ as a model of heavy metal ions from polluted water. The performance of the PAB adsorbent and the adsorption mechanism was assessed in terms of kinetics and thermodynamics.

2. Materials and Methods

2.1. Materials

Aniline (An, 99%) and anhydrous ferric chloride (FeCl₃, 98%) (Loba Chemie Pvt. Ltd., Mumbai, India); mercury (II) nitrate monohydrate (Hg(NO₃)₂·H₂O, 98.5%), ethanol (EtOH), and methanol (MeOH) (Sigma-Aldrich, Taufkirchen, Germany); nitric acid (HNO₃, 68%), dichloromethane (DCM, 99%), and sodium hydroxide (NaOH, 98%) (BDH Chemicals, Poole, England, UK); benzene (Bz, 99%) (Qualikems Fine Chem Pvt Ltd., Vadodara, India); dimethoxymethane (DMM, 98%) (Alfa Aesar, Karlsruhe, Germany); hydrochloric acid (HCl, 37%) (Panreac, Barcelona, Spain); and sodium nitrate (NaNO₃, 99) (Winlab, Pontefract, London, UK) were used as received without further treatment. Deionized water was used wherever needed.

2.2. Synthesis of PAB Adsorbent

The target adsorbent of poly(aniline-co-benzene) was synthesized following a method described by Dawson et al. [29]. Briefly, 1 mmol of aniline, 9 mmol of benzene, 20 mmol of DMM, and 25 mL of DCM were mixed with stirring at room temperature. To this solution, 20 mmol of FeCl₃ was added under nitrogen conditions, and the reaction was continued with stirring at 80 °C for 18 h. After cooling, the solid phase was collected using filtration, repeatedly washed with ethanol, and purified using Soxhlet extraction in methanol for 18 h. The product was dried at 60 °C for 24 h to yield about 0.8 g (~80.2%) of brown solid material, termed the PAB network.

2.3. Characterization

The Fourier transform infrared (FTIR) spectrum of the PAB network was recorded on a Thermo Scientific Nicolet iS10 (Madison, WI, USA) using the KBr–disc method on the frequency range of 400–4000 cm⁻¹ with a 4 cm⁻¹ resolution and 32 runs per spectrum. The surface morphology was observed using a JSM-7610F LV scanning electron microscope (SEM) (JEOL, Tokyo, Japan). Energy dispersive X-ray spectroscopy (EDS) spectra were obtained using an SEM-coupled EDS, X-MaxN system, from Oxford instruments (Abingdon, UK). A thermogravimetric analysis (TGA and DTA) was carried out on an STA449F3 machine from NETZSCH (Hanau, Germany) for samples (5–10 mg) heated from 25–1000 °C at a heating rate of 10 °C/min and under a nitrogen flow of 20 mL/min. A differential scanning calorimetry (DSC) thermogram was obtained using Shimadzu DSC 60 for a 10 mg sample heated from 25 to 400 °C with a heating rate of 10 °C/min under a nitrogen environment of 50 mL/min. The surface area was measured using a NOVA 2200e surface area analyzer (Quantachrome Corp., Boynton Beach, FL, USA) based on nitrogen adsorption isotherms at 77 K. The machine was first vacuum-degassed at 180 °C for at least 2 h. The BET (Brunauer–Emmett–Teller) specific surface area was calculated at P/P₀ = 0.05–0.25. The concentration of the adsorbent Hg²⁺ ion was determined with ICP-MS using an iCAP^TM Q instrument from Thermo Scientific (Waltham, MA, USA). Samples were diluted with 5% HNO₃ before the analysis.

2.4. Adsorption Experiments

2.4.1. pH Drift

The pH at which the net surface charge of the PAB adsorbent is zero (pH_PZC) was determined using the pH drift approach described elsewhere [30]. Typically, a solution of NaNO₃ (0.1 M) was prepared from which a series of 10 mL of pHs 3–11 (denoted as pH initial, pHi) were prepared with the drop addition of 0.1 M of HCl or NaOH to the NaNO₃ solution. The pHs of the solutions were monitored using a bench-top pH-meter Orion 3-Star from Thermo Scientific (Beverly, MA, USA) previously calibrated for the test range. A total of 15 mg of the PAB adsorbent was added to each pHi solution. The solutions were agitated at 150 rpm for 24 h at room temperature (25 ± 2 °C). The samples were centrifuged, and the solutions’ final pHs (pHf) were measured. The pH_PZC was assessed by plotting pHi versus pHf as the average of two experiments.

2.4.2. Effect of Adsorption pH

The effect of pH on the adsorbent performance was examined at pHs 2, 4, 5, and 7 using 25 mL of a 50 mg/L Hg²⁺ adsorbent, 10 mg PAB adsorbate, 150 rpm agitation speed, 24 h contact time, and 25 °C adsorption temperature.

2.4.3. Adsorption Kinetic

The kinetic adsorption study was carried out in batches under the following conditions: a 10 mg PAB adsorbent dose, 25 mL Hg²⁺ adsorbate volume, 50 mg/L Hg²⁺ adsorbate initial concentration, pH of 5.5, 150 rpm agitation speed, and room temperature. At specific points in the given range of 5–120 min (and after 24 h), the adsorption process was stopped by centrifugation and the un-adsorbed adsorbate was analyzed using ICP-MS.

2.4.4. Adsorption Thermodynamics

The adsorption thermodynamic experiments were performed at 25, 35, 45, and 55 °C under the following constants: 10 mg of PAB, 25 mL of Hg²⁺, 50 mg/L of Hg²⁺, pH 5.5, 150 rpm, room temperature, and 24 h of contact time. After adsorption equilibrium, the adsorbent was separated with centrifugation and the supernatant was analyzed using ICP-MS. Each time, the metal samples to be analyzed using ICP-MS were subjected to digestion using 5% nitric acid followed by filtration, then storage at a low temperature to ensure accurate concentration.

2.5. Theoretical Calculations

The performance of the adsorption system was analyzed throughout the modeling of the adsorption rate, equilibrium, and adsorption thermodynamics.

The removal efficiency and capacity of the adsorbent were determined according to Equations (1)–(3):

$$Re\%=\left(\frac{C_0-C_e}{C_0}\right)100$$

(1)

$$q_t=\frac{\left(C_0-C_t\right)V}{m}$$

(2)

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

(3)

where C₀, C_t, and C_e (mg/L) are the adsorbate concentrations in the liquid phase at the adsorption times (min) at zero, t, and equilibrium. V (L) is the adsorbate volume, and m (g) is the dry mass of the adsorbent used.

The adsorption kinetic data were modeled for nonlinear pseudo-first-order (PFO), pseudo-second-order (PSO), intra-particle diffusion (IPD), and liquid film diffusion models, described by Lagergren–Svenska [31], Ho–McKay [32], Weber–Morris [33], and Boyed kinetic expressions [34], respectively, as given with Equations (4)–(7).

$$q_t=q_e\left(1-e^{-k_{1}t }\right)$$

(4)

$$q_t=\frac{k_2{q}_e^{2}t}{1+k_2{q}_{e}t}$$

(5)

$$q_t=k_{id}{t}^{0.5}+C$$

(6)

$$-\ln \left(1-F\right)=k_{fd}t+C; F=\frac{q_t}{q_e}$$

(7)

where q_t and q_e are the capacities at time (t) and at equilibrium, k₁ (min⁻¹), k₂ (g/(mg·min)), k_id (mg/(g·min^0.5)), and k_fd are the rate constants of the four models, respectively, and C (mg/g) is the y-intercept of the diffusion models. The value of the PSO initial rate (h) (mg/(g·min)) of the adsorption process is given in Equation (8).

$$h=k_2{q}_e^2$$

(8)

The thermodynamic parameters, including Gibbs free energy change (∆G°), enthalpy change (∆H°), and entropy change (∆S°), were studied following Equations (9)–(11) [35,36].

$$ln{K}_d=\frac{\Delta S°}{R}-\frac{\Delta H°}{RT}$$

(9)

$$K_d=\frac{q_e}{C_e}$$

(10)

$$\Delta G°=\Delta H°-T\Delta S°$$

(11)

where K_d (L/g) is the distribution constant, T is the adsorption temperature (Kelvin), and R (8.314 J/mol·K) is the universal gas constant.

3. Results and Discussion

3.1. Adsorbent Structure and Properties

Figure 1 illustrates the reaction condition for the synthesis of the PAB hypercrosslinked network. The reaction proceeded via the Friedel–Crafts alkylation of aniline and benzene with a methyl moiety as an alkyl crosslinker in the presence of FeCl₃ as a catalyst. The resulting brown, insoluble powder was subjected to characterization for its structural, morphological, and thermal properties. The 1:9 reactant ratio was selected based on a previous study [29], in which a series of aniline-to-benzene networks were investigated and it was found that the monomeric ratio of 1:9 provided a network with adequate surface area and porosity compared to others. However, aniline is more reactive toward Friedel–Crafts alkylation than benzene; therefore, the network is unlikely to incorporate the same rate of the monomers (Figure 1). Typically, the network consists of copolymerized aromatic rings of aniline and benzene and crosslinked with a methylene group. It is reported that by increasing the molar ratio of the crosslinker, the surface area decreases as expected. Hence, the utilized DMM/total monomer (2:1 mol ratio) could balance a strong network with the surface area necessary for an adequate adsorption capacity of the target adsorbate.

3.1.1. FTIR Analysis

The FTIR spectrum of the synthesized PAB network is presented in Figure 2. The spectra revealed a broad peak centered at 3404 cm⁻¹ assigned for the N-H stretching band of primary amine (-NH₂) [37], which overlapped with O-H from physisorbed water. The peaks around 2916 and 1435 cm⁻¹ are due to the aliphatic C-H stretching and scissor-like bending vibrations, indicating that the crosslinking reaction was successful [38]. The peak at 3009 cm⁻¹ is part of aromatic =C-H (sp²-C) stretching bands. The peak at 1660 cm⁻¹ can be assigned to the bending mode of amine groups (-NH₂) [39,40]; the ones at 1604 and 1500 cm⁻¹ (C=C stretching), 1435 cm⁻¹ (C-C stretching), and 1206 cm⁻¹ (C-H bending) are characterizing the aromatic ring [41]. The presence of amine functionality is supported with the existing band at 1304 cm⁻¹ for the C-N stretching mode, as previously reported [41]. The bands at 1000–567 cm⁻¹ are due to various substituted benzene rings [38,40,41] in the network.

3.1.2. SEM and EDS Analysis

The structural morphology of the synthesized network was monitored using the SEM technique. In Figure 3, the micrograph displays fiber-like networks with filaments as short as 100 nm. Furthermore, particles of a diameter as small as 30 nm could be detected. The general view indicates the presence of plenty of pores with a wrinkled surface, which supports the high surface area obtained and is explained below. As suggested with a BJH isotherm, mesopores are dominant, and while their average diameter is close to the edge of the micropores, the latter contribution to the material porosity cannot be ignored; obviously, micropores are too small to be measured under such SEM magnification. Nevertheless, surface pores and voids between particles contribute to the high surface area and support efficient adsorption sites.

Figure 4 shows the EDS spectral image and the corresponding elemental composition of PAB HCP. Theoretically, PAB HCP consists of carbon (C), hydrogen (H), and nitrogen (N). Technically, EDS is not a proper technique for detecting low Z elements due to X-ray absorption phenomena caused by the Be window of the detector unless the element fraction is high or the EDS equipment is modern [42]. Therefore, EDS cannot detect H atoms; nevertheless, elements with adequate low Z values (e.g., nitrogen with low content) can be traced. As expected, an elemental analysis (Figure 4 and

Table 1. BET surface area, pore volume, pore width, pore diameter, and energy dispersive X-ray spectroscopy (EDS) elemental analysis of PAB network.

Adsorbent	Nitrogen Adsorption–Desorption Isotherm				EDS Elemental Analysis
	BET Surface Area (m2·g−1)	Pore Volume (cm3·g−1)	BET Pore Width (nm)	BJH Pore Diameter (nm)	C	O	N	Cl
PAB	987	0.908	3.681	4.960	86.6	8.3	4.8	0.4

) revealed a high carbon content (86.6 at.%). The analysis also disclosed traces of chlorine (Cl) and oxygen (O) atoms, which were attributed to trapped catalysts and adsorbed water, respectively, in the pores. Indeed, water molecules could be affirmed using the FTIR adsorption peak at 3404 cm⁻¹ for OH stretching vibration. Nitrogen content was measured as 4.8 at.%, slightly higher than theoretically calculated values when a complete reaction is achieved (~1.4 at%). On the one hand, the reaction yield was quantitively averaged to 80.2%, suggesting an incomplete consumption of precursors. On the other hand, aniline is more reactive towards Friedel–Crafts alkylation than benzene and, therefore, it is unlikely for the products to incorporate the same rate of both monomers [29]. These may justify the found higher N/C ratio than the expected one. It is worth mentioning that a previous work [29] conducted under similar preparation conditions has reported a nitrogen content of 1.02%, which is lesser than the approximated theoretical value of 1.4 at.%. This variation may be a result of one or more of the following: (i) the complete consumption of reactants reported by the authors, and thus the observed and calculated values are much closer to each other; (ii) the trapped adsorbates such as gases and water may differ and affect the percentage of the end element’s ratio; (iii) the accuracy of the EDS analysis is low compared to other techniques such as a CHN-S elemental analysis; and (iv) the reactivity difference between aniline and benzene toward the Friedel–Crafts reaction may cause block polymerization, which may bring some challenges for area selection for an EDS analysis.

3.1.3. BET Analysis

Figure 5 shows the N₂ adsorption–desorption isotherm for the PAB adsorbent. The corresponding BET specific surface area, pore volume, and pore width as well as the BJH pore diameter are summarized in Table 1. The obtained surface area of 987 (m²·g⁻¹) was close to that of previously reported ones at 1097 [29] and 1270 (m²·g⁻¹) [43]. As can be seen, the pore volume and size were calculated to be 0.908 (cm³·g⁻¹) and 3.681 (nm), respectively. Furthermore, the instant high N₂ adsorption at low relative pressures indicates high uptake affinity, thus reflecting an abundant micropore structure [43]; however, the BJH average pore diameter (4.96 nm) supports a dominant mesopore structure, a case that the SEM micrograph could trace. According to the IUPAC classification of pore sizes, the following types can be distinguished based on the adsorption mechanism of N₂ at 77 K and 1 atm: macro- > 50 nm, 2 < meso- < 50 nm, and micro-pores < 2 nm [44]. The adsorption–desorption isotherm profile is of type IV with an H4 hysteresis loop. Typically, the shapes of a type-IV isotherm could vary and, thus, produce different hysteresis curves. Although H4 hysteresis reflects complex material with micropores and mesopores [45], its shape in Figure 5 supports the mesoporous structure of the PAB HCPs [46].

3.1.4. Thermal Analysis

The thermal stability of the PAB adsorbent was examined using a TGA analysis under an N₂ atmosphere, as shown in Figure 6A. The thermogram revealed a multistep decomposition; hence, at least three steps can be traced (

Table 2. Thermal properties of the PAB adsorbent.

Adsorbent	TGA			d-TGA	DTA	DSC
	Step	Temp. (°C)	Mass Loss (%)	Temp. (°C)	Temp. (°C)
PAB	1	25–125	2.9	74	61, endo	365 °C, exothermic
	2	126–280	4.4	195	293, exo
	3	281–780	28.3	420, 526	556, exo
	Residue	1000	43	-	-

). The first step of a 2.9 wt% mass loss, observed in the TGA range of 25 to 125 °C and centered at a d-TGA of 74 °C, is mainly due to adsorbed volatile substances, including moisture and gasses. This desorption can be confirmed with the endothermic peak at a DTA of 74 °C. The second step was calculated in the range 126–280 °C with a mass loss of 4.4 wt% centered at a d-TGA of 195 °C with a DTA exothermic peak at 293 °C, suggesting an initial decomposition of the weak functional group, e.g., amine. The third stage of decomposition revealed a significant mass loss of 28.3 wt% in the range 281–780 °C with two prominent peaks of a d-TGA of 420 and 526 °C, suggesting the stepwise decomposition of the network backbone. The mass loss was continued with a temperature increase, resulting in a residue of 43 wt% at 1000 °C as carbon content. According to DTA, a dominant exothermic peak for the decomposition of the network backbone is clear at 556 °C. The DSC curve in Figure 6B revealed no peaks for glass transition or melting temperature (T_g), supporting the DTA results. However, an exothermic peak at 365 °C was observed (Table 2), indicating that chemical reactions may occur and are of a decomposition type, as suggested with DTA.

3.1.5. Point of Zero Charge

The surface charge of the PAB adsorbent was assessed using the pH-drift method. As shown in Figure 7A, the pH at which the net surface charge is zero (pH_PZC) was estimated at pH 5.1. This indicates that the surface charge is positive below pH 5.1 and negative above it.

3.2. Effect of Solution pH

The effect of adsorption operating pH on the adsorption efficiency of the system was investigated in the pH range of 3–8. The result indicated a pH-dependent adsorption process, which decreased with a pH increase, as shown in Figure 7B. The removal efficiency revealed a slight decrease up to pH 5 while it became significant above pH 5 to pH 8. This result adequately agrees with the estimated surface charge (pH_PZC), which is negative above pH 5.1. Therefore, the pH of the adsorption media was set to 5.5 for the subsequent adsorption experiments, as the point was considered negative but close to the observed high adsorption efficiency shown in Figure 7B. The apparent, comparable high removal efficiency observed in the pH range of 3–5 may suggest an equal affinity for proton and mercury ions to adsorb onto PAB, supposedly via interaction with amine functional groups. However, above pH 5, the decline in the adsorption efficiency could be expressed in terms of porosity and pore size, which may accommodate Hg²⁺ ions. The BET analysis revealed a mesoporous scale (4.960 nm) of PAB pores. Hence, by increasing the pH, the OH⁻ ions increase, which may settle in the pores and thus hamper the adsorption of Hg²⁺. Indeed, such a mechanism needs experimental confirmation that may be carried out later.

3.3. Adsorption Kinetics—Effect of Contact Time

The results of adsorption kinetic studies are shown in Figure 8. The data were modeled using PFO, PSO, IPD, and LFD kinetic models and the values of the corresponding parameters are summarized in

Table 3. Kinetic parameters of PFO, PSO, IPD, and LFD models.

qe, exp. (mg/g)	PFO			PSO				IPD			LFD
	k1 (min−1)	qe, calc. (mg/g)	R2	k2 (min−1)	qe, calc. (mg/g)	h (mg/(g·min)	R2	kid (mg/g·min0.5)	Cid (mg/g)	R2	kfd (min−1)	Cfd (mg/g)	R2
99.75	0.579	99.04	0.999	0.030	98.84	293	0.999	6.484	46.71	0.468	0.032	3.19	0.836

. It is clear that the adsorption is fast in the first 10 min, over which more than 90% of the adsorbent capacity was used. After 10 min, the adsorption slowed and soon reached equilibrium (at about 20 min). To ensure adsorbent saturation, the operation continued for 120 min and further for 24 h. The early saturation of the adsorption sites may indicate easy access to the active sites on the adsorbent. The slow rate in the second phase (above 10 min) suggests the further filling of pores via intraparticle diffusion; however, this process is negligible, and no apparent obstacles in reaching the active sites of the pores could be detected.

According to the coefficient of determination (R²), the PFO (R² = 0.999) model fits the experimental data better than PSO (R² = 0.999). Furthermore, the experimental capacity q_e,exp (99.75 mg/g) is closer to PFO (99.04 mg/g); however, the difference between the R² and q_e values from both models is insignificant, supporting the possible use of both equations to describe the adsorption rate. The initial adsorption rate assessed with the PSO h constant was high (293 (mg/(g·min)), supporting the previously discussed observation. The adsorption mechanism was evaluated further by modeling the diffusion process using the IPD and LFD models. As given in Table 3, the data poorly fit the two models, confirming their incompetent contribution to the adsorption mechanism. According to R² (0.468, 0.836) and the boundary layer C (46.71, 3.19 mg/g) of IPD and LFD, respectively, LFD contribution seems to be higher than IPD. Therefore, the uptake is mainly film-diffusion-controlled.

3.4. Adsorption Thermodynamics

The temperature-dependent adsorption of Hg²⁺ using the PAB hypercrosslinked polymeric adsorbent is depicted in Figure 9 and the thermodynamic parameters were gathered in

Table 4. Thermodynamic data of the Hg²⁺ adsorption onto PAB.

Temp. (K)	ln Kd	ΔG° (kJ/mol)	ΔH° (kJ/mol)	ΔS° (J/mol·K)	R2
298	1.85	−4.41	32.39	123.48	0.977
308	2.09	−5.64
318	2.60	−6.88
328	3.02	−8.11

. The study was conducted at 25, 35, 45, and 55 °C. As can be seen, the sign of ΔG° was negative at all temperatures, confirming the feasibility and spontaneity of the sorption process. Furthermore, as temperature increased from 25 to 55 °C, the ΔG° values were slightly increased from −4.41 to −8.11 (kJ/mol). Accordingly, the adsorption performance is higher at high temperatures. The sign of both ΔH° and ΔS° is positive, indicating the endothermic nature of the adsorption process, and randomness increases at the solid–liquid interface.

The sign and extent of the thermodynamic parameters can be utilized to assess the adsorption mechanism. Hence, the adsorption process is favored at high temperatures. As the values of ΔG° and ΔH° were in the range of physical processes (i.e., ΔG° from −20 to 0 kJ/mol and ΔH° less than 40 kJ/mol), the adsorption is predominantly physical in nature [36].

3.5. Adsorption Mechanism

Figure 10 illustrates a proposed mechanism for the adsorption process of Hg²⁺ from wastewater onto PAB HCPs. Adsorption could occur via electrostatic interaction, possible complexation, and pore filling. The electrostatic attraction between metal ions and partially charged sites, like an electron pair on the nitrogen atom and electron density on the rings (cation-π), is considered and affirmed with FTIR. As aniline represents at least 10 mol% of the network, its contribution to the overall adsorption mechanism is unignorable. However, the benzene ring, with its high electron density that is polarizable in acidic and basic conditions, can also adhere to the ions throughout the electrostatic and complexation process, thus contributing to the overall adsorption mechanism. The PAB is negatively charged at the applied adsorption condition at pH 5.5 (the pH_PZC = 5.1). Hence, the expected mechanism dominantly involves complexation. At low acidic media (~pH 5–7), the amine group still protonated and thus formed ammonium hydroxide (or salt) with OH⁻ (or some other ions possibly existing in the system like NO₃⁻, etc.), which further complex Hg²⁺ ions. Another possible mechanism could be the mechanical trapping inside the pores. According to the kinetic models, LFD has a greater impact on the adsorption process than IPD. The large surface area indicates plenty of pores; however, the N₂ adsorption–desorption isotherm profile revealed a predominant mesopores structure followed by micropores. It is assumed that the abundance of Hg²⁺-philic sites on the PAB outer and pore surfaces are provided by electron donors involving nitrogen atoms and benzine rings, which works synergically with mechanical trapping pronounced by pore cavities. As the adsorption mechanism is physical, a weak interaction is expected. Therefore, the contribution from ion exchange is less effective at the applied adsorption pH, while interaction is poorly electrostatically bound. However, chemical interaction could occur in the first minutes, causing a fast adsorption rate, as observed in Figure 8. Such sorption may be due to the existence of some unprotonated amines, but the action is not part of the adsorption–controlling step. Finally, the amine and benzene rings play essential roles in chelating mercury ions to the PAB HCPs [23,28].

3.6. Relative Performance of PAB

To assess the position of the investigated PAB adsorbent among other reported HCPs in removing toxic metal ions from wastewater, a list of articles was screened [20,28,47,48,49,50], and the adsorbents’ performance is summarized in

Table 5. Adsorption performance of different hypercrosslinked polymers from the literature in the removal of metal ions including Hg²⁺.

Adsorbent				Adsorbate			Adsorption Conditions			Kinetic Model	Mechanism	Ref.
HCP Adsorbent	SSA	Pore Type	Dose (g/L)	Metal Ion	Conc. (mg/L)	Capacity (qe; mg/g)	Time (min)	pH	Temp. (°C)
Polystyrene- based HCP	854	Meso and micro	-	Cd2+	120	822	90	7	20	PFO	physical	[28]
Polystyrene- based HCP	824	Meso and micro	-	Ni2+	51	162	24 h	7	60	-	physical	[20]
				Pb2+	80	137	24 h	7
HTC-HCP	168	Meso and micro	1.0	Pb2+	100	178	80	7	25	-	-	[47]
				Cd2+		70	45			-
HCPMOS	813	micro	0.5	Fe3+	10	13	1000	7	25	PSO	physical	[48]
DTCP	3.9	-	1.5	Hg2+	40	30	40	5	50	PSO	physical	[49]
CHAP-SH	1.9	-	1.0	Hg2+	300	237	100	4.5	25	PSO	chemical	[50]
PAB HCPs	987	Meso and micro	0.4	Hg2+	50	99	120	5.5	25	PFO	physical	This work

Notes: Abbreviations. HTC-HCP: 2-Hydroxyterephthalic acid modified hypercrosslinked polymer; HCPMOS: Hypercrosslinked microporous functional polystyrene; DTCP: Polydithiocrabamates; CHAP-SH: Crosslinked hyperbranched sulfhydryl-modified polyamide–amines.

. For appropriate comparison, the listed adsorbents were all HCPs and, wherever applicable, presented with their surface area, pore type, adsorption condition, and adsorption mechanism. It is noteworthy that the reported capacities were essentially based on kinetic data. As can be seen, the adsorption capacity of PAB HCP is well-positioned in the list (capacity range: 822–13 mg/g; PAB capacity: 99 mg/g). However, the performance depends on the adsorbent intrinsic properties and many other variables; therefore, comparison using capacity is not always straightforward. For example, polystyrene-based HCPs [20,28,48] have shown a high surface area (813–854 m²/g) and their capacity toward divalent ions differs significantly, while meso- and micro-pores as well as physical adsorption mechanisms were reported for most (Table 5). Furthermore, for the same HCP, e.g., [20] and [47], and under similar adsorption conditions, the affinity of HCP toward different metal ions also varies. By comparing the adsorption capacity of various HCPs, e.g., DTCP and CHAP-SH, against Hg²⁺ [49,50] with the PAB adsorbent, it can be realized that the capacity of PAB is adequately placed in the range of 30–237 mg/g; however, despite its low surface area, the capacity of CHAP-SH was the highest (237 mg/g). This could be because CHAP-SH has been modified with a thiol-functional group that also supports chemical adsorption.

4. Conclusions

A hypercrosslinked poly(aniline-co-benzene) copolymer was synthesized and characterized with FTIR, BET, SEM, TGA/DTA, and DSC. The product featured a mesoporous structure and high surface area, which support its potential as an adsorbent. The adsorption performance of PAB HCP in removing Hg²⁺ ions from polluted water was assessed in terms of kinetics and thermodynamics. The data fit well with both PFO and PSO, but PFO performed better. This suggests that a physisorption process and rate-limiting step involves a diffusion process, as further confirmed with IPD and LFD models. The adsorption thermodynamic parameters confirmed the spontaneity nature (ΔG° ≤ −4.41) of the process and its endothermic (ΔH° is positive) nature, randomness increase (ΔS° is positive) at the adsorption interface, and physical adsorption mechanism. These findings suggest that the PAB HCPs might be used as an efficient adsorbent for removing Hg²⁺ ions and other heavy metal ions from industrial effluent.