Preparation of CS/PVA/POP Nanofiber Membranes and Adsorption Behavior for Hg(II) Ions

Abstract

Chitosan (CS) and polyvinyl alcohol (PVA) nanofiber membranes were synthesized via electrospinning and used as supporting materials for powdered porous organic polymer (POP). These membranes were then crosslinked with glutaraldehyde, resulting in nanofiber membranes (CS/PVA/POP) as an efficient adsorbent for Hg(II) ions. Characterization using Fourier transform infrared spectroscopy, X-ray diffraction, and scanning electron microscopy showed that the membranes effectively removed up to 92.9% of mercury ions at optimal conditions, with an adsorption capacity of 116.1 mg/g. The adsorption data fit well with the Langmuir isotherm and pseudo-second-order kinetic models. The efficient uptake of mercury ions was attributed to chemisorption involving active groups (C=S, -NH₂, -OH), facilitated by mechanisms such as chelation, complexation, or electron exchange. The CS/PVA/POP nanofiber membranes demonstrated significant advantages in adsorption capacity, economic viability, and recyclability, providing an effective solution to mercury pollution in water.

1. Introduction

Mercury and its derivatives have the characteristics of permanence, easy migration and difficult degradation, and are currently acknowledged as one of the most serious heavy metal pollutants in the world by the United Nations Environment Programme [1,2]. Mercury easily enters the human body or other organisms through the food chain, causing birth defects, brain damage, cardiovascular diseases, and other irreversible damage to people and other species, posing a critical hazard to public health and the ecological environment [3,4,5]. In 2017, the Minamata Convention on Mercury was adopted internationally, strictly binding emissions of highly toxic mercury, which is of great significance for reducing and ultimately eliminating the harm of mercury [6]. Therefore, it is urgent to develop new materials and approaches for the efficient decontamination of mercury from water. Compared with other Hg(II) treatment technologies, such as chemical precipitation [7], microbial remediation [8], and ion exchange [9], the adsorption method was proved to be an economical and environmentally friendly method with advantages such as a high utilization rate, simple operation, and high efficiency [10,11,12,13].

The natural polymer chitosan (CS) exhibits strong adsorption for mercury ions because its structure is rich in active groups (-OH, -NH₂), which can be chelated with metal ions. At the same time, chitosan has a broad spectrum of sources, low cost, easy biodegradation, and non-toxicity, so it has become the first-choice environmental protection heavy metal adsorbent for researchers [14,15]. However, CS has low mechanical strength and poor thermal stability; in particular, CS is easily dissolved under acidic conditions, resulting in its structural destruction, and then forms a gel, which affects the chelation ability for heavy metal ions [16]. Many researchers commonly use chemical crosslinking to enhance the stability of chitosan [17,18]. The introduction of glutaraldehyde (GA) as a crosslinking agent can increase the pore size of the chitosan membrane, thus providing a greater active surface and more accessible adsorption active sites, enhancing the acid and water resistance of chitosan, and reducing the detrimental effect of crosslinking on the adsorption capacity. Vieira et al. [19] compared the adsorption properties of the crosslinked chitosan membranes prepared by crosslinking agents glutaraldehyde and epichlorohydrin (ECH), respectively, for mercury ions. The data indicate that crosslinking enhanced the adsorption capacity of chitosan at low pH; furthermore, the adsorption capacity of GA-CS was greater than that of ECH-CS.

In addition, the specific surface area of natural chitosan is small, and the unmodified chitosan powder itself may not provide sufficient internal adsorption binding sites, which will affect its adsorption effect on Hg(II) ions and may cause adsorbent waste. The chitosan nanofiber membrane prepared by electrospinning technology has a relatively large surface area and porosity, which can greatly improve the adsorption capacity of chitosan [20]. However, due to the instability and inadequate mechanical strength of chitosan, it is challenging to prepare nanofibrous films by electrostatic spinning using chitosan alone. It has been reported that the addition of an easily filamentable polymer (such as polyvinyl alcohol PVA) to the chitosan solution is beneficial to improving the spinnability of chitosan and make it more mechanically strong and durable [21,22,23].

Porous organic polymer (POP) is a kind of porous mesh material that is formed by strong covalent bonds between various organic structural units with different geometries and topologies. POP has the characteristics of a large specific surface area, adjustable porosity, high chemical stability, and easy post-modification etc. POP-based adsorbents have been extensively applied in environmental remediation [24,25,26]. When POP is used as an adsorbent to adsorb Hg(II) in water, it is also necessary to introduce functional groups (such as nitrogen, sulfur, oxygen groups) to increase its active sites. Gorginpour et al. [27] stated that they had prepared a POP by Schiff base reaction, with the corresponding adsorption capacity reaching up to 833 mg/g for Hg(II) ions. Recently, we reported a novel POP by reacting thiourea with trialdehyde phloroglucinol as an adsorbent of Hg(II) ions [28], which exhibited an adsorption capacity of 1250 mg/g and a removal rate of 99.3% within 120 min over a broad pH range from 4 to 10. Although the reported POP adsorption materials show high adsorption capacity, high removal rate, and excellent selectivity and reusability, they are not only expensive but also exist in powder form, which is difficult to collect and recycle effectively in practical applications. Therefore, it is very important and necessary to study nanofiber membranes containing powdered POP for practical applications.

In this study, a nitrogen- and sulfur-rich POPs were synthesized using a solvothermal method [28]. CS/PVA was used as a support carrier of POP, and a series of nanofiber membranes were prepared through electrospinning, followed by crosslinking with glutaraldehyde. By varying the glutaraldehyde crosslinking time and POP loading, we identified the optimal formulation of the CS/PVA/POP adsorbent with enhanced adsorption properties. The membrane not only facilitates the efficient recovery of powdered POP but also enhances the adsorption capacity of CS/PVA for heavy metal ions through POP doping. The CS/PVA/POP can provide numerous adsorption active sites responsible for the chemical adsorption of Hg(II) ions. This unique physicochemical dual-action mechanism endows the adsorbent with superior adsorption performance, rendering it one of the most promising adsorbent materials. Characterization techniques such as FT-IR, XRD, and SEM were employed, and we investigated factors affecting adsorption, including pH, contact time, and initial Hg(II) ion concentration. Typical adsorption isotherm and kinetic models were used to analyze the experimental data and explore the adsorption mechanisms of CS/PVA/POP for Hg(II) ions. This research highlights the potential of using CS/PVA nanofiber films as substrates for loading POP, offering a novel approach for utilizing powdered porous organic polymers while enhancing their adsorption capacity.

2. Materials and Methods

2.1. Materials

The chemical reagents used were all analytical grade. 2,4,6-Trihydroxybenzene-1,3,5-tricarbaldehyde, glutaric dialdehyde, mesitylene, and hydrochloric acid were all purchased from Shanghai Aladdin Reagent Co., Ltd. (Shanghai, China). Chitosan with 85% deacetylation was obtained from Beijing Inokai Technology Co., Ltd. (Beijing, China). Polyvinyl alcohol (Type 1799) was obtained from Shanghai Macklin Biochemical Technology Co., Ltd. (Shanghai, China). Mercury dichloride was purchased from Guizhou Tongren Tailuier chemical plant (Tongren, China). Glacial acetic acid was purchased from Sinopharm Chemical Reagent Co., Ltd. (Shanghai, China).

2.2. Characterization

FT-IR spectra were used to characterize the functional groups of the samples on a PerkinElmer Spectrum One (B) spectrometer (PerkinElmer, Foster City, CA, USA), where the wavelength ranged from 500 cm⁻¹ to 4000 cm⁻¹, and the samples were tablet pressed with KBr. The morphology of the samples was observed by electron scanning microscope (SEM, JEOL 6500F, Tokyo, Japan). The X-ray diffraction (XRD) pattern of the sample was obtained by an X-ray diffractometer (DMAX-3A, Tokyo, Japan) using Cu-Kα radiation (λ = 1.54 Å) between 0° and 25° (2θ) at 40 kV and 40 mA. The concentration of Hg(II) ions was determined by a precision cold atomic absorption mercury meter (Type CG-1C, Changzhou, China).

2.3. Preparation of Adsorbents

2.3.1. Preparation of POP

The POP was synthesized according to the literature reported by Hu’s group [28]. 2,4,6-Trihydroxybenzene-1,3,5-tricarbaldehyde (126 mg), thiourea (90 mg), 1,4-dioxane (3.5 mL), and mesitylene (3.5 mL) were added into a Schlenk tube, and the mixture was sonicated. Then, 0.25 mL glacial acetic acid and 0.5 mL water were added to the tube and mixed well by stirring. After a freeze–pump–thaw cycle, the Schlenk tube was closed under nitrogen. Thereafter, the tube was heated in an oil bath at 120 °C for 24 h. Upon completion of the reaction, the Schlenk tube was removed and allowed to reach room temperature, and the product was collected by centrifugation. The product was washed with anhydrous tetrahydrofuran and acetone three times each, and then dried in a vacuum oven at 60 °C. The purplish-red powder POP was prepared.

2.3.2. Preparation of CS/PVA/POP Nanofiber Membranes

Firstly, the spinning solution was prepared. A 10 wt.% PVA solution was prepared by dissolving PVA (5.0 g) in deionized water (50 mL) and stirring at 120 °C for 0.5 h. CS (2.0 g) was dissolved in 50 mL concentrated acetic acid (90 wt.%), and stirred for 24 h at 75 °C, and the 4 wt.% of CS solution was prepared. Then, a 10 wt.% PVA solution and 4 wt.% chitosan solution were prepared in a mixed solution with a mass ratio of 1:1. Finally, POP of different qualities was added to the above mixed solutions, respectively, and ultrasonically stirred; thus, a series of CS/PVA spinning solutions with different POP loadings were obtained. Next, the electrospinning process was carried out. A stationary plate was covered with release paper, and the spinning solutions with varying POP loads were loaded into syringes equipped with 21-gauge needles (inner diameter 0.5 mm). Under the conditions of a 19.7 ± 1.6 °C temperature and a 28 ± 5% humidity, a high voltage of 25 ± 2 kV was applied between the needle and the collector, with a needle-to-collector distance of 10 cm, and the spinning solution flow rate was adjusted to 0.10 mm/min. These syringes were placed in a liquid supply pump tank, with different spinning parameters set based on the POP load. Once the liquid in the pump was fully dispensed, the release paper with the composite films was removed and placed in an oven at 40 °C for drying. This produced the CS/PVA/POP nanofiber membranes. The dried membranes were then crosslinked with glutaraldehyde to enhance their acid and water resistance. The overall preparation process of the CS/PVA/POP nanofiber membranes via electrospinning is illustrated in Figure 1.

2.4. Adsorption Experiments

CS/PVA/POP was used to adsorb mercury ions, and the effects of solution PH, adsorption time, initial mercury ion concentration, and other factors on the adsorption effect were investigated. A 500 mg/L Hg(II) ion standard solution (HgCl₂) was prepared, and other initial concentrations of Hg(II) ion solutions were prepared by dilution. The pH values of the standard solution of Hg(II) ions were adjusted with HCl solution(0.1 mol/L) and NaOH solution (0.1 mol/L). CS/PVA/POP nanofiber films (10 mg) with different POP loads were, respectively, added into 50 mg/L Hg(II) ion standard solution (25 mL) and adsorbed for a certain time under magnetic stirring at 600 rpm. Then, the concentration of Hg(II) ions after adsorption was measured by cold atomic absorption mercury meter. According to the Equations (1) and (2), the adsorption capacity and removal rate of CS/PVA/POP can be determined as follows:

Q_e = ((C₀ − C_e) × V)/m

(1)

R = (C₀ − C_e)/C₀ × 100%

(2)

where C₀ and C_e (mg/L) denote the initial and equilibrium concentrations of Hg(II) ions. V (mL) and m (mg) are the volume of solution and mass of adsorbent. Q_e (mg/g) and R (%) represent the adsorption capacity and the removal rate.

3. Results and Discussion

3.1. Effect of Glutaraldehyde Crosslinking Time

Glutaraldehyde can interact with active groups in CS/PVA, such as -NH₂ and -OH, which reduces the adsorption performance for mercury ions [29]. In order to take into account the higher stability of CS/PVA nanofiber membranes and adsorption capacity for Hg(II) ions, it is essential to explore the optimal crosslinking time of glutaraldehyde. The CS/PVA nanofiber films were crosslinked in a 25 wt.% glutaraldehyde vapor at 40 °C for varying durations of 0, 0.5, 1, 2, and 3 h. Residual glutaraldehyde was removed by drying the crosslinked membranes in an oven at 45 °C for 1 h.

The water resistance of the crosslinked fiber membranes was measured. The fiber films were immersed in deionized water for 12 h, and then removed and dried in the oven until the quality of the films was constant. The insolubility fraction was obtained by Equation (3):

N = m_i/m₀ × 100%

(3)

where N is the insolubility fraction (%), and m₀ and m_i are the mass of CS/PVA nanofiber films before and after water immersion, respectively.

The adsorption capacity of crosslinked fiber membranes was measured. A 25 mL standard Hg(II) ion solution (10 mg/L) at pH = 4 was adsorbed by the crosslinked CS/PVA fiber membrane, and the Hg(II) ion concentration in the solution was measured after adsorption for 2 h. The changes in the insoluble fraction and adsorption capacity of the CS/PVA nanofiber films under different crosslinking times are presented in Figure 2a.

As shown in Figure 2a, increasing the crosslinking time led to a gradual rise in the insoluble fraction of the CS/PVA nanofiber membrane, which increased from 60% to 99.7%. The adsorption capacity exhibited a trend of first increasing and then decreasing, reaching a peak of 23.75 mg/g from an initial 17.5 mg/g, before falling to 9.8 mg/g. Consequently, the optimal crosslinking time for glutaraldehyde was found to be 1 h. Importantly, the crosslinking of glutaraldehyde did not affect the structure of the POP, indicating that the optimal crosslinking time for the CS/PVA/POP electrospun nanofiber film is also 1 h.

3.2. Effect of POP Loadings on Adsorption

To ensure economic viability, it is essential to create composite membranes that have a low load of materials while maintaining effective adsorption capabilities. In this study, CS/PVA/POP nanofiber membranes with POP loadings of 0, 0.1, 0.2, 0.5, 0.7, 1, and 3 wt.% were prepared, respectively. At room temperature, 10 mg samples of each CS/PVA/POP membrane with the various POP loadings were tested for their ability to adsorb Hg(II) ions from solution. The adsorption properties of the adsorbents were evaluated after a 2 h adsorption period. The results of the CS/PVA/POP nanofiber membranes with different POP loadings for Hg(II) ions, including the adsorption capacity, removal rate, and physical representation, are illustrated in Figure 2b. Figure 2b shows that the adsorption capacity of CS/PVA/POP membrane loaded with only 0.1 wt.% POP increased from 54.38 mg/g to 64.58 mg/g, indicating that the doping of POP in the fiber membrane can improve the adsorption for mercury ions. With the increase in POP loading capacity, the adsorption capacity and removal rate of CS/PVA/POP for Hg(II) ions increased rapidly. The membrane with 3 wt.% POP achieved the highest removal rate at 92.9%. This trend indicates that a higher POP loading correlates with improved adsorption capacity and removal efficiency. The improved performance is attributed to the abundant active adsorption sites (such as -C-N- and C=S) in the POP-doped membranes, which chemisorb with Hg(II) ions. In conclusion, the optimal preparation conditions of CS/PVA/CS nanofiber membrane are a glutaraldehyde crosslinking time of 1 h and a POP loading of 3 wt.%.

3.3. Characterization of CS/PVA/POP

3.3.1. FT-IR Analysis

The FT-IR spectra of CS/PVA, POP and CS/PVA/POP are shown in Figure 3a. In the FT-IR spectra of CS/PVA, a sharp and broad absorption peak at 3413 cm⁻¹ indicates the presence of hydrogen bonding between -OH and -NH₂ groups formed during the electrospinning process. In addition, the strength of the deformation vibration peak at nearly 1601 cm⁻¹ was significantly reduced, indicating that -NH₂ may form -NH₃⁺ under the action of hydrogen bond, and thus the influence of -NH₂ on spinning was weakened [30,31]. The FT-IR spectral analysis of POP reveals an N-H stretching vibration peak at 3415 cm⁻¹, while the absorption bands detected at 1600 cm⁻¹ and 1199 cm⁻¹ correspond to the characteristic vibrations of C=O (ketone group) and C-N, which proved that POP was successfully synthesized [32,33]. Meanwhile, the N-C=S absorption peak at 1418 cm⁻¹ indicates that the sulfur successfully coalesced into the POP molecules [34]. Compared with CS/PVA, CS/PVA/POP exhibits new absorption peaks at 1634 cm⁻¹ and 1429 cm⁻¹, ascribed to the characteristic C=O and N-C=S peaks of POP, individually, which confirmed that POP was successfully doped into CS/PVA.

3.3.2. XRD Analysis

As can be seen from Figure 3b, the XRD patterns of the POP and CS/PVA/POP nanofiber films show no obvious diffraction peaks. The XRD patterns of amorphous polymers often show a flat baseline without obvious crystal diffraction peaks. Therefore, it is proved that the POP and CS/PVA/POP nanofiber films are amorphous polymers. Compared to POP, CS/PVA/POP has a weak diffraction peak at 19.2°, which corresponds to the (020) peak of the CS/PVA basement membrane at 19.8° [30,31]. The slight shift to a lower angle indicates the successful doping of POP onto the composite membrane. Although the amount of POP in CS/PVA/POP is low, it still produces observable changes in the XRD signal.

3.3.3. SEM Analysis

Figure 4 shows the SEM images of POP, CS/PVA, and CS/PVA/POP. Figure 4a reveals its spherical morphology and smooth surface. Figure 4b presents the SEM image of CS/PVA, and Figure 4e shows the results of its statistical diameter distribution. The results reveal that CS/PVA presents a uniform and continuous fiber shape with a smooth surface and a mean diameter of 94.14 nm, indicating that the one-dimensional nanomaterial CS/PVA was successfully prepared by electrospinning technology.

Figure 4c,d shows the SEM image of CS/PVA/POP, and Figure 4f shows the results of its statistical diameter distribution. POP particles can be observed in Figure 4b,d, indicating that the POP particles were embedded in the fibers. Because the doping of POP particles increased the concentration of the spinning solution, the average diameter of CS/PVA/POP also increased from 94.14 nm to 124 nm. Meanwhile, because the synthesized POP shows a purplish red color, the color of the nanofiber film changed from white to purple after the addition of POP, so the appearance of the fiber film also proves that POP was successfully doped into the CS/PVA nanofiber film. CS/PVA/POP after glutaraldehyde crosslinking reveals a crosslinked network structure, which enhances the fiber breaking strength and improves the membrane’s mechanical strength and stability in water. Moreover, crosslinking created 3D porous fiber structures with tighter interconnection. This configuration allows for the rapid fixation of Hg(II) ions on the pore surfaces via van der Waals forces or electrostatic interaction between the porous channels and Hg(II) ions.

3.4. Adsorption Properties

3.4.1. Effect of pH

The pH value of the solution has a great influence on the adsorption performance of the adsorbent, as it affects the protonation of functional groups at the binding sites and the form of the target metal ions being adsorbed [35]. The effect of pH on the adsorption capacity and removal rate of CS/PVA is shown in Figure 5a. With the increasing pH from 1 to 3, the corresponding adsorption capacity increases from 145.2 mg/g to 155.3 and the removal rate increases from 72.3% to 76.8%, respectively. Figure 5c shows the relationship between the removal rate of Hg(II) ions in water by CS/PVA/POP and the pH values of the solution (from 1 to 7). According to Figure 5a, when the pH value is 4, the removal rate of Hg(II) ions by the fiber membrane reaches the maximum of 92.9%. Under the same conditions, the removal rate of Hg(II) ions using CS/PVA/POP is greater than that of CS/PVA.

At a lower pH (pH < 4), the nitrogen-containing adsorption active sites of the fiber membrane are easily protonated, and they find difficulty binding with Hg(II) ions. Meanwhile, the Zeta potential of POP is positive at pH = 1–3 [28], creating significant electrostatic repulsion between POP and Hg(II) ions, which hinders effective complexation. As the pH increases, the Zeta potential of POP decreases sharply, reducing the electrostatic repulsion and protonation of the adsorption sites, inducing a steep increment in the removal rate of the Hg(II) ions by CS/PVA/POP. When the pH > 3 for CS/PVA and pH > 4 for CS/PVA/POP, Hg(II) ions gradually form Hg(OH)₂ [36], which may cover the surface adsorption active sites of CS/PVA and CS/PVA/POP, thus blocking the pore channels and preventing further adsorption for Hg(II) ions. Therefore, the corresponding adsorption capacity or removal rate decreases.

3.4.2. Effect of Contact Time

The effect of the contact time on the adsorption capacity and removal rate of CS/PVA and CS/PVA/POP is presented in Figure 5b and Figure 6a. During the initial phase (0–20 min for CS/PVA and 0–10 min for CS/PVA/POP), there was a rapid increase in both the adsorption capacity and the removal rate of mercury ions by the composite membrane as the adsorption time elapsed. This is because of the van der Waals forces or electrostatic interaction between the 3D porous structure of the composite membrane and Hg(II) ions, combined with the large specific surface area of the composite membrane and numerous adsorption active sites (such as -NH₂, -OH, C=S, C-N, C=O). After 10 min or 20 min of adsorption, both the adsorption amount and removal rate of the Hg(II) ions began to increase slowly and gradually, and leveled off within 90 min. This is attributed to most active sites being occupied by Hg(II) ions, leading to significant charge repulsion between the Hg(II) ions in solution and those already adsorbed on the surface of CS/PVA/POP, making it difficult for the remaining active sites to be occupied. Moreover, as the Hg(II) ion concentration progressively declined, the driving force for mass transfer between the Hg(II) ions and CS/PVA/POP decreased. Finally, within 120 min, the adsorption of mercury ions by the composite membrane reached equilibrium, and the removal rate was 92.9%.

3.4.3. Effect of Initial Hg(II) Ion Concentration

The effect of the initial Hg(II) ion concentration (C₀, 10–500 mg/L) on the adsorption capacity of CS/PVA/POP is shown in Figure 6b, comprising two main stages. Phase I: the initial Hg(II) ion concentration was below 300 mg/L. The adsorption capacity escalated significantly with rising Hg(II) ion levels, from 24.8 mg/g to 443.8 mg/g. This originated from the large ratio of adsorbed active sites on the surface of CS/PVA/POP to the initial Hg(II) ion concentration, thereby increasing the opportunity for interaction between the active sites and Hg(II) ions. Phase II: the initial Hg(II) ion concentration varied between 300 and 500 mg/L. The adsorption capacity increased slowly with the rising Hg(II) ion concentration until it reached equilibrium. At this stage, most of the adsorption sites on CS/PVA/POP were occupied by Hg(II) ions, leading to competition for the remaining effective binding sites and causing a gradual rise in the composite membrane’s adsorption capacity for Hg(II) ions. When the initial Hg(II) ion concentration was 500 mg/L, the adsorption active sites of the composite membrane showed a tendency to reach saturation, and the adsorption capacity of the composite membrane for Hg(II) ions was 493.8 mg/g.

3.4.4. Sorption Isotherm Models

Freundlich [37] and Langmuir [38] isotherm models were employed to characterize the interaction mechanisms between the adsorbate and adsorbent at equilibrium. The isothermal equations can be expressed as follows:

Langmuir: (Q_m K_L C_e)/(1 + K_L C_e)

(4)

Freundlich: Q_e = K_F⋅C_e^(1/n)

(5)

where C_e and Q_e are the Hg(II) ion concentration (mg/L) and adsorption capacity (mg/g) at equilibrium, respectively; Q_m is the theoretical maximum adsorption capacity of the adsorbent (mg/g); K_L is the Langmuir adsorption constant (L/mg); K_F is the Freundlich adsorption constant ((mg/g)(L/mg)^1/n); and n reflects the adsorption intensity.

Figure 6c presents the fitting results of Langmuir and Freundlich adsorption isotherms, while the corresponding model parameters are listed in

Table 1. Langmuir and Freundlich isotherm parameters.

Langmuir				Freundlich
Qm (mg/g)	KL (L/mg)	R2	n	KF ((mg/g)(L/mg)1/n)	R2
723.30	0.0049	0.9893	19.24	1.8635	0.9532

. The Langmuir model exhibits a higher R² value (0.9893) compared to the Freundlich model (0.9532). Therefore, the Langmuir adsorption isotherm model is preferable for describing the adsorption behavior of CS/PVA/POP on Hg(II) ions. The results show that the adsorption of CS/PVA/POP is a single-layer adsorption, and that the adsorption active sites are evenly distributed.

The Langmuir isotherm model uses the separation factor R_L to assess the feasibility of adsorption. The calculation formula of R_L is presented below:

R_L = 1/(1 + K_L C₀)

(6)

The calculation results are shown in

Table 2. R_L values at different initial Hg(II) ion concentration.

C0 (mg/L)	10	50	100	200	300	400	500
RL	0.5308	0.1845	0.1016	0.0535	0.0363	0.0275	0.0221

. It can be observed that the R_L values are all less than 1 across different initial concentrations of mercury ions (10–500 mg/g). This indicates that Hg(II) ions can be effectively adsorbed by CS/PVA/POP, demonstrating that CS/PVA/POP is an effective adsorbent for Hg(II) ions.

3.4.5. Sorption Kinetics

Fitting of Pseudo-First-Order and Pseudo-Second-Order Kinetic Models

To measure the adsorption efficiency of CS/PVA/POP on Hg(II) ions and to speculate the adsorption mechanism, a pseudo-first-order kinetic model [39] and a pseudo-second-order kinetic model [39,40] were fitted. The equations of the two models ((7) and (8)) are as follows, respectively:

Q_t = Q_e (1 − exp(−k₁t)

(7)

Q_t = (k₂ Q_e² t)/(1 + k₂ Q_e t)

(8)

where Q_t (mg/g) and Q_e (mg/g) denote adsorption capacity at time (t) and at equilibrium, respectively, while k₁ and k₂ correspond to the rate constants of the pseudo-first and pseudo-second order kinetic equations.

The fitting results of adsorption kinetics are presented in Figure 6d, with the corresponding model parameters summarized in

Table 3. Fitting parameters of pseudo-first-order and pseudo-second-order kinetic models.

Pseudo-First-Order			Pseudo-Second-Order
k1 (min−1)	Qe (mg/g)	R2	k2 [g/(mg·min)]	Qe (mg/g)	R2
0.158	102.03	0.8349	0.00192	111.87	0.9326

. The data reveal that the fitting constant R² of the pseudo-second-order kinetic (0.9326) is significantly higher than that of the pseudo-first-order kinetic (0.8349). In addition, the theoretical adsorption quantity Q_e (111.87 mg/g) derived from the pseudo-second-order kinetic model is very close to the experimental value Q_t (116.1 mg/g); therefore, this model is better suited to depicting the adsorption kinetic of the CS/PVA/POP adsorbent. It is suggested that the adsorption mechanism might involve complexation, chelation, electron sharing, or exchange transfer between the active groups of CS/PVA/POP and the Hg(II) ions, rather than solely relying on physical adsorption.

Intraparticle Diffusion Model

According to the intraparticle diffusion model, adsorption is governed by multiple diffusion mechanisms [41], and its equation is show in Formula (9):

Q_t = kt^0.5 + C

(9)

In this equation, k (mg/(g·min^0.5)) represents the rate constant of intraparticle diffusion adsorption, and C is the boundary layer constant thickness.

The relevant fitting outcomes and parameters of the intraparticle diffusion model are detailed in Figure 6e and

Table 4. Adsorption parameters of intraparticle diffusion model (units of k (mg/(g·min^0.5)).

Phase I (Film Diffusion)			Phase II (Pore Diffusion)			Phase III (Equilibrium)
k1	C1	R12	k2	C2	R22	k3	C3	R32
22.47	4.88	0.9200	12.48	19.17	1	3.90	72.97	0.8934

. The adsorption of Hg(II) ions by CS/PVA/POP occurs in three stages: film diffusion, pore diffusion, and equilibrium stage. Table 4 indicates that the adsorption rate constant (k) in the film diffusion stage is 22.47, which is much higher than the other stages, proving that film diffusion is a decisive factor in the adsorption process. When the surface of CS/PVA/POP reached adsorption saturation, the Hg(II) ions diffused into the 3D porous channels of the CS/PVA base membrane and the pores within the POP particles, and then gradually moved into the inside of the adsorbent. Then, the Hg(II) ions bonded to the adsorption active sites until the active sites were overwhelmingly occupied, reaching equilibrium stage.

3.5. Reusability

Taking into account the practical application and cost factors, reusability is an important technical index in evaluating excellent adsorbents. In this study, a mixture of HCl and thiourea was selected as the eluent for the regeneration of CS/PVA/POP. Using this desorption agent, the adsorbent loaded with Hg(II) could be successfully removed. Figure 6f shows the removal rate of CS/PVA/POP with 5 cycles. After 5 sorption–desorption cycles, the final removal rate was 91.4% of the initial removal rate, which may show the loss of adsorbent that led to the decrease in removal rate through the repeated sorption–desorption experiments. This indicates that CS/PVA/POP has good reusability and stability.

3.6. Adsorption Mechanism

Zuo et al. [28] showed that the adsorption mechanism of POP primarily involves chemical chelation and electrostatic interaction by XRD and density functional theory (DFT) calculations. Moreover, the molecular electrostatic potential and adsorption binding energy near the sulfur-containing functional groups are the largest, presenting the optimal location for Hg(II) binding. Sharma et al. [42] analyzed the FT-IR spectra of germanium (Ge)-functionalized CS/PVA electrospinning film prior to and following the adsorption of Hg(II) ions and proved that -NH-, -NHCO-, and -OH indicate all the adsorption binding sites of Hg(II) ions. Kyzas et al. [43] analyzed the FT-IR spectra of the glutaraldehyde crosslinked CS pre- and post-adsorption of Hg(II) ions and found that the newly formed -C=N- and the existing free amino groups after crosslinking were still complexed with Hg(II) ions.

According to the above analyses, combined with adsorption isotherm models and kinetic models in this work, it is concluded that the adsorption mechanism of CS/PVA/POP for removing Hg(II) ions may be a single-layer adsorption dominated by chemisorption, with uniform adsorption active sites. Moreover, physical adsorption is also involved.

Firstly, physical adsorption. CS/PVA/POP possesses a large specific surface area and 3D porous fiber structure. Hg(II) ions are first adsorbed onto the surface of the nanofiber membrane during the film adsorption stage, and van der Waals force or electrostatic interaction between the composite film and mercury ions is generated, which helps to fix Hg(II). Moreover, POP has a rich porous structure, and Hg(II) ions slowly enter the interior of the adsorbent through the pore channels during the pore diffusion stage. Secondly, chemical adsorption. Hg(II) ions may preferentially undergo soft-acid–soft-base complexation and electron exchange with sulfur-containing adsorption sites (C=S) in POP; then, Hg(II) ions chelate C-N and complex with C=O. In addition, some active groups in CS/PVA, such as -NH₂, C=N, -NHCO-, and -OH, may be used as adsorption active sites and participate in chemisorption.

3.7. Comparison with the Related Adsorbents

Table 5. Comparison of the maximum adsorption capacity (Q_max) of Hg(II) ions by different adsorbents.

Absorbent	Qmax (mg/g)	Reference
Germanium-modified CS/PVA nanofiber	31.4	[42]
CS–cotton fibers	104.3	[44]
Sulfur-modified CS Hydrogel	186.9	[45]
Mesoporous silica/CS composite	478.5	[46]
Amidinothiourea-modified CS particles	322.5	[47]
Diatomite-CS/PVA composite	195.7	[16]
Magnetic polydopamine–CS NPs	245.2	[48]
Monodisperse magnetic functional CS	246.1	[49]
Thiourea based POP	1250	[28]
POP/CS	249.2	[50]
CS/PVA/POP nanofibre membrane	493.8	This work

presents a comparison of the maximum adsorption capacity of Hg(II) ions between CS/PVA/POP and other related adsorbents reported in the literature. It can be seen that the adsorption capacity of physicochemical CS/PVA/POP co-adsorbent is higher than that of other adsorbents, except for the POP adsorbent.

The CS/PVA/POP composite, created by using an electrospun nanofiber film of CS/PVA as a support for POP, enhances the specific surface area of chitosan while generating numerous pores between composite nanofiber fibers. At the same time, doping POP rich in nitrogen and sulfur heteroatom adsorption active sites can more effectively improve the adsorption performance of Hg(II) on CS/PVA. In addition, CS/PVA electrospinning fiber film doped with only 3 wt.% POP can reach 39.5% of the maximum adsorption capacity of POP alone. However, due to the higher cost of POP and the fact that CS/PVA/POP fiber membrane is easier to recover and recycle than powdered POP in practical application, CS/PVA/POP is more economical than POP in removing Hg(II) ions from wastewater. According to the above analysis, CS/PVA/POP is superior to other related adsorbents in terms of adsorption performance, economic cost, and recycling.

4. Conclusions

This study successfully developed CS/PVA/POP nanofiber membranes by incorporating nitrogen-rich and sulfur-rich POP powder into the electrospinning solution of CS/PVA, followed by crosslinking with GA. The optimal conditions were identified as 1 h of crosslinking time and a 3 wt.% loading of POP. The adsorption experiments revealed that at pH 4, after 2 h of contact time, the removal rate and adsorption capacity reached 92.9% and 116.1 mg/g, respectively. The synthesized membranes exhibited Hg(II) adsorption behavior consistent with the Langmuir isotherm model, and the kinetics followed a pseudo-second-order model, indicating that the process is primarily governed by uniform chemical adsorption sites. Utilizing chitosan-based nanofiber membranes as support facilitates the recovery of Hg(II) ions, with the inclusion of POP significantly enhancing their adsorption performance. This research offers a promising approach to improve the recyclability of porous organic polymer adsorbents and effectively remove Hg(II) ions from water. Future studies could explore chitosan-based electrospun membranes doped with other porous adsorbents for selective mercury ion adsorption and develop crosslinking agents rich in heteroatoms to further enhance the removal efficiency of chitosan-based adsorbents.