Competitive Adsorption Studies of Cd(II) and As(III) by Poly (Butylene Succinate) Microplastics: Based on Experimental and Theoretical Calculation

Abstract

Microplastics (MPs) can serve as vectors for heavy metals in aquatic environments; however, the adsorption behavior of MPs on multiple heavy metal systems is still unclear. This study investigated the adsorption characteristics of biodegradable poly (butylene succinate) (PBS) for cadmium (Cd(II)) and arsenic (As(III)) in both single and binary systems. Adsorption isotherms were studied using the Linear, Langmuir, and Freundlich models, and further analysis of MPs adsorption characteristics was conducted using site energy distribution theory and density functional theory. The results indicate that the maximum adsorption capacities of PBS for Cd(II) and As(III) are 2.997 mg/g and 2.606 mg/g, respectively, with the Freundlich model providing the best fit, suggesting multilayer adsorption on heterogeneous sites. As(III) has a higher adsorption affinity for PBS than Cd(II), with a binding energy of −11.219 kcal/mol. Additionally, the adsorption mechanisms of Cd(II) and As(III) on PBS include electrostatic interactions and surface complexation, with the primary adsorption sites at the C=O of the carboxyl group and the hydroxyl group. The comprehension of interfacial interactions between biodegradable plastics and heavy metals is facilitated by a combination of theoretical calculations and experimental investigations.

1. Introduction

Plastic products are widely used due to their low cost and versatility, but the disposal of their waste has caused severe environmental pollution. With the rapid growth of plastic production and consumption, approximately 700,000 metric tons of plastic waste are discharged into the environment annually [1]. Microplastics (MPs), which are smaller than 5 mm, form waste plastics and are broken down into tiny fragments by physical, chemical, and biological processes [2,3]. Conventional MPs may take decades or even centuries to degrade in the natural environment and can spread throughout ecosystems, accumulating in water bodies such as groundwater, lakes, and oceans [4,5]. As a result, numerous studies are dedicated to the development of biodegradable materials as sustainable ecological substitutes [4,6]. Common biodegradable plastics include polybutylene succinate (PBS), polylactic acid (PLA), and poly(butylene adipate-co-terephthalate) (PBAT) [7,8]. Studies have shown that in sunlit aquatic environments, biodegradable plastics have a higher microplastic generation rate in microbe-rich aquatic settings than non-biodegradable plastics, as their ester backbone is more prone to hydrolysis or enzymatic degradation [9]. PBS is a modern biopolymer made from petroleum-based materials or renewable agricultural products [10]. It is safe for direct contact with food and has excellent biodegradability; it is capable of being processed in existing composting systems [11]. Nowadays, PBS is widely used in fields such as food packaging, agricultural coverings, and tableware [12]. Notably, microplastic particles, with their large surface area and hydrophobic nature, have the ability to concentrate toxic chemicals, thereby influencing the distribution and bioavailability of pollutants in the environment [13,14].

Heavy metals pose a severe threat to ecosystems and human health due to their non-degradability, bioaccumulation, and toxicity [15]. Cadmium (Cd) and arsenic (As) pollution are major human health risks [16]. Cd exposure can cause cancer, birth defects, immune system damage, and liver dysfunction, while long-term As exposure is linked to skin cancer, neurological diseases, diabetes, cardiovascular issues, endocrine disruptions, and reproductive problems [17,18]. Cd and As often coexist in contaminated environments from metal-containing products and ore mining and smelting [19]. Wastewater discharge containing these pollutants contaminates water, soil, and rice, posing risks to water quality, food safety, and public health. Additionally, heavy metals and MPs often coexist in aquatic environments, with MPs adsorbing heavy metals, which may further alter their environmental behavior, bioavailability, and toxicity [20]. Research shows that polyamide (PA) microplastics, with C-O and N-H groups, have the strongest adsorption capacity for Cd compared to polystyrene (PS), polyvinyl chloride (PVC), acrylonitrile-butadiene-styrene (ABS), and polyethylene terephthalate (PET) [21]. Similarly, Guo et al. found that Cd adsorption capacity follows the order PVC > PS > polypropylene (PP) > polyethylene (PE), due to PVC’s glassy state enhancing interaction with chlorinated PVC groups [22]. Studies also indicate that MPs made from biodegradable plastics are more effective at adsorbing heavy metals than traditional MPs, likely due to their higher hydrophobicity and surface tension [23,24,25]. The adsorption of heavy metals by MPs depends on their unique physicochemical properties [14,26]. MPs in aquatic environments adsorb heavy metals through complexation, electrostatic attraction, cation-β bond interactions, and physical adsorption. However, reports on the interfacial behavior of heavy metals on biodegradable MPs remain limited. Therefore, exploring the adsorption properties of biodegradable MPs is particularly important.

Different heavy metals are present in the natural environment and can engage in competing adsorption or synergistic interactions [27,28]. The macroscopic approach currently emphasizes the isotherm, kinetics, and diffusion models in adsorption behaviors, but a comprehensive understanding of these aspects is lacking [29]. To analyze the competitive adsorption amongst contaminants, site energy distribution theory (SEDT) can reflect how various adsorption site types are occupied on heterogeneous adsorbents from an energy standpoint [30,31]. By using magnetic biochar (MBC-1), SEDT looked into the simultaneous adsorption of two common developing pollutants: bisphenol A (BPA) and sulfamethoxazole (SMX) [32]. The results demonstrated that BPA preferentially occupies high-energy locations in the co-adsorption system, such as π-π EDA interactions, ion exchange, and surface coordination. Nevertheless, the current approach of SEDT combined with characterization-based techniques for comprehensively assessing the adsorption mechanism is insufficient. Molecular simulations offer insights into adsorption behaviors at the molecular level, which is advantageous and practical for clarifying adsorption mechanisms, in contrast to experimental research [33]. For instance, Bao et al. [34] used density functional theory (DFT) calculations to elucidate the adsorption mechanism of PVC on 2-hydroxy naphthalene (2-OHN), naphthalene (NAP), phenanthrene (PHE), and pyrene (PYR). Feng et al. [35] investigated the adsorption of heavy metal ions (Cu (II)) and benzo[a]pyrene (BaP) by nanoplastics by molecular dynamics simulations (MD). Hence, using molecular simulation methods to elucidate the competitive adsorption mechanisms in binary competition systems of heavy metals and analyze their interactions is of paramount importance.

This study investigates the adsorption behavior of biodegradable microplastics on heavy metal coexistence systems through a combination of experimental and theoretical computations. The adsorbent was selected as PBS, and the adsorbates were set as Cd(II) and As(III). The principal aims of the study were as follows: (1) to examine the adsorption behavior of Cd(II) and As(III) on PBS microplastics; (2) to further analyze the adsorption characteristics of Cd(II) and As(III) on single and binary systems through theoretical calculations and the theory of site energy distribution; and (3) to examine the mechanisms of single and binary system adsorption on PBS through chemical calculations and physical phase analysis. The work done here can facilitate understanding the relationships between heavy metal coexistence systems and biodegradable MPs.

2. Materials and Methods

2.1. Materials and Instruments

The biodegradable PBS, with a particle size of 200 mesh, was acquired from Kunshan Kexin Polymer Materials Co., Ltd. (Kunshan, China). Cadmium and arsenic standard solutions (1 g/L) were sourced from Beijing Century Aoke Biotechnology Co., Ltd. (Beijing, China). The physical and chemical properties of PBS are shown in

Table 1. The chemical and physical properties of PBS.

Chemical Name	Molecular Formula	Structure	Molecular Configuration	Crystallinity (%)	${\mathit{T}}_{\mathbf{g}}$ (°C)	Elongation at Break (%)
Poly (butylene succinate)	H-(C8H12O4)n-OH			60.30	−34	>500

Note: Molecular formula from [36]; Crystallinity (%) from [12]; T_g (°C) means glass transition temperature, obtained from [23]; Elongation at break (%) from [10].

. Experimental settings were configured using a constant-temperature air bath oscillator (TH7-C, Taicang Experimental Equipment Factory, Suzhou, China) and an inductively coupled plasma emission spectrometer (ICPS-7510, Shimadzu, Kyoto, Japan) to measure heavy metal concentration infiltrates. The chemical and physical characteristics of PBS were analyzed with a scanning electron microscope (SEM, SU8010, Hitachi, Japan) and X-ray photoelectron spectroscopy (XPS, K-alpha, Thermofisher, Waltham, MA, USA). Furthermore, standard equipment like a vacuum drying oven (DZF-6050, Keran Instrument Co., Chongqing, China), an ultrasonic cleaner (AP-01P, Autosense Instrument Co., Tianjin, China), and an ultrapure water meter (Diect-Q5UV, Millipore, Burlington, MA, USA) were used.

2.2. Experimental Methods

2.2.1. Competitive Adsorption Experiments

The two primary forms of arsenic present in aqueous environments are known to be As(III) and As(V), with As(III) being more mobile, active, and dangerous than As(V) [37]. Based on investigations, the predominant component in the neutral range is As(III) [18]. In this study, competitive adsorption studies were carried out on the heavy metals Cd(II) and As(III).

All the experiments were repeated. Batch adsorption tests were performed under the following conditions: 0.1 g of MPs and 100 mL of 1 mg/L heavy metal solution were added to glass conical flasks. The samples were sealed and subjected to shaking in a constant-temperature air bath oscillator (200 rpm, 25 °C) for 72 h to assess the adsorption process. The experimental results of PBS adsorption capacity indicate that 48 h can be considered the equilibrium time (Figure S1). As described above, batch adsorption was conducted to test the adsorption isotherms of cadmium and arsenic solutions ranging from 1 to 25 mg/L and PBS. Before measurement, all the samples were filtered using a 0.22 μm syringe filter to eliminate MPs.

2.2.2. Chemical Calculation

The structure of PBS (C₂₄H₃₈O₁₃) was obtained using molecular dynamic simulations using the COMPASS force field within the Forcite module of Materials Studio (BIOVIA). The polymer underwent structural optimization, annealing, and minimization for 200 ps in the NVT mode at 500 K. The following simulations were conducted at the specified temperatures of 400 K and 300 K for a duration of 200 ps. The lowest energy conformations were re-optimized for adsorption experiments using DFT simulations. Adsorption tests were conducted via the adsorption locator module, utilizing the COMPASS force field and the Grand Canonical Monte Carlo (GCMC) method, which included ten annealing cycles and 10⁵ steps, with simulated annealing localizing the adsorption conformation between 100–1.0 × 10⁵ K [18,38]. To investigate the adsorption mechanisms of Cd(II) and As(III) on PBS, spin-polarized density functional theory was applied. The calculations for frontier molecular orbitals (FMOs), molecular electrostatic potentials (ESP), and binding energies were carried out using the Dmol3 module from Materials Studio, based on quantum chemistry principles. The simulations employed a conductor-like shielding model (COSMO) with a dielectric constant of 78.54, representing water. Electron exchange interactions were characterized using Perdew–Burke–Ernzerhof (PBE) functionals and generalized gradient corrections (GGA). Atomic basis functions were based on atom-centered lattices, the electronic basis set was dual numerically polarized (DNP4.4), and the core treatment was based on the “DFT semicore pseudopots” approach [39]. For geometry optimization in the Dmol₃ program, a self-consistent field convergence value of 1.0 × 10⁻⁷ was applied, along with the Grimme scheme for correct dispersion, a solvation model for aqueous solvents, and an orbit-occupation smearing parameter of 0.01 Ha [38]. Energy, force, and displacement tolerances were set at 1.0 × 10⁻⁵ Ha, 2 × 10⁻³ Ha Å, and 5 × 10⁻³ Å, respectively, to ensure precise fitting. The binding energy and energy gap were computed as indicated in [40,41]:

$$E_{\mathrm{ad}}=E_{\mathrm{pbs}\text{-}\mathrm{A}}-(E_{\mathrm{A}}+E_{\mathrm{pbs}})$$

(1)

$$\Delta E=E_{\mathrm{L}}-E_{\mathrm{H}}$$

(2)

where $\Delta E=E_{\mathrm{L}}-E_{\mathrm{H}}$, $E_{\mathrm{L}}$, and $E_{\mathrm{H}}$ represent the energy gap, LUMOs, and HOMOs energies, respectively. $E_{\mathrm{ad}}$ (kcal/mol) represents the binding energy between PBS and heavy metals; $E_{\mathrm{pbs}\text{-}\mathrm{A}}$ (Ha), $E_{\mathrm{A}}$ (Ha), and $E_{\mathrm{pbs}}$ (Ha) represent the energies of PBS and heavy metal complexes, and heavy metals and PBS, respectively. Hartree is the unit of atomic energy, which can be converted to eV and kcal/mol.

2.3. Data Analysis

(1)

The adsorption capacity

The PBS adsorption capacity at a specific time was calculated as follows:

$$q_{\mathrm{t}}={\displaystyle \frac{\left(C_0-C_{\mathrm{t}}\right)}{W}}V$$

(3)

The dynamic adsorption capacity of MPs for heavy metals at time t (h), represented as $q_{\mathrm{t}}$ (mg/g), was quantified using the equation. $C_0$ (mg/L) represents the initial heavy metal concentration, and $C_{\mathrm{t}}$ (mg/L) represents the concentration of heavy metals at time t. The volume of the aqueous solution is given by V (L), and the mass of MPs is denoted as W (g).

(2)

The sorption isotherm models

The adsorption data conformed to the Linear, Freundlich, and Langmuir models ((4)–(6)), and the fitted equations are shown below:

$$q_{\mathrm{e}}=K_{\mathrm{d}}{C}_{\mathrm{e}}$$

(4)

$$q_{\mathrm{e}}=K_{\mathrm{F}}{C}_{\mathrm{e}}^{\mathrm{n}}$$

(5)

$$q_{\mathrm{e}}=q_{\max }{\displaystyle \frac{K_{\mathrm{L}}{C}_{\mathrm{e}}}{1+K_{\mathrm{L}}{C}_{\mathrm{e}}}}$$

(6)

where $q_{\mathrm{e}}$ (mg/g) refers to the equilibrium capacity of MPs for the adsorption of heavy metals; $K_{\mathrm{d}}$ (L/g) represents the partition coefficient; $K_{\mathrm{F}}$ (mg^1−1/nL^1/ng⁻¹) is the Freundlich constant, with “n” as its value; $K_{\mathrm{L}}$ (L/mg), $q_{\max }$ (mg/g) are the adsorption parameters of the Langmuir model; and $C_{\mathrm{e}}$ (mg/L) represents the equilibrium concentration of the heavy metals.

(3)

The site energy distribution theory

To investigate the competitive behavior in the mixed system, the site energy distribution theory was applied for the analysis, with the corresponding expressions provided below. The correlation function for SEDT is represented in Equation (7) as follows:

$$q_{\mathrm{e}}(C_{\mathrm{e}})={\displaystyle {\int }_0^{+\infty }}{q}_{\mathrm{h}}(E,C_{\mathrm{e}})F(E)\mathrm{d}E$$

(7)

In this equation, $q_{\mathrm{e}}\left(C_{\mathrm{e}}\right)$ (mg/g) represents the equilibrium adsorption capacity of the PBS in the system; $q_{\mathrm{h}}(E,C_{\mathrm{e}})$ describes the local sorption isotherms at a sorption energy E (kJ/mol); and $F(E)$ (mg·mol/g·kJ) is the frequency distribution of site energies, providing insight into the heterogeneity of the sorption sites across the material. The energy E characterizes the difference in interaction energy between the solute and solvent at a given adsorption site. Using Cerofolini’s approximation, the equilibrium concentration of the adsorbate in the liquid phase ($C_{\mathrm{e}}$) is related to the adsorption energy (E) by:

$$C_{\mathrm{e}}=C_{\mathrm{s}}\exp (-{\displaystyle \frac{E-E_{\mathrm{S}}}{RT}})=C_{\mathrm{s}}\exp (-{\displaystyle \frac{E^{*}}{RT}})$$

(8)

where $C_{\mathrm{s}}$ represents the maximum absolute solubility in the solvent, $E_{\mathrm{s}}$ (kJ/mol) is the adsorption energy at the concentration of $C_{\mathrm{s}}$, R (8.314 × 10⁻³ kJ/(mol·K)) represents the universal gas constant, and T (K) denotes the temperature,.

The term $E^{*}$ represents the relative difference in adsorption energies between the adsorbate and solvent, normalized against the adsorbent surface. In this study, we assume that the Freundlich isotherm model accurately describes the adsorption behavior. Based on this assumption, an approximate site energy distribution function $F(E^{*})$ could be expressed as Carter et al. [42]:

$$F\left(E^{*}\right)={\displaystyle \frac{K_{\mathrm{F}}{C}_{\mathrm{s}}^{\mathrm{n}}\mathrm{n}}{RT}}\exp \left(-{\displaystyle \frac{n{E}^{*}}{RT}}\right)$$

(9)

The parameters in this expression have been well-established in previous studies. This function offers a way to model the energy distribution across heterogeneous adsorption sites, providing a key insight into the nature of the interaction between the adsorbate and adsorbent.

Further refinement of this model [43] allows for the quantification of the average energy ($E_{\mathrm{m}}$) and the degree of energy distribution inhomogeneity (${\sigma }_{\mathrm{e}}^{*}$), which can be derived from the mathematical expectation and standard deviation, respectively:

$$E_{\mathrm{m}}=E_{\mathrm{P}}\left(E^{*}\right)$$

(10)

$${\sigma }_{\mathrm{e}}^{*}=\sqrt{E_{\mathrm{p}}\left(E^{*2}\right)-E_{\mathrm{P}}{\left(E^{*}\right)}^2}=\sqrt{E_{\mathrm{p}}{\left(E^{*}\right)}^2-E_{\mathrm{m}}^2}$$

(11)

where $E_{\mathrm{P}}\left(E^{*}\right)$ and $E_{\mathrm{p}}\left(E^{*2}\right)$ are the statistical estimates of $E^{*}$ and $E^{*2}$ within the given energy range.

2.4. Software and Data Analysis

The X-ray photoelectron spectroscopy data were examined with Thermo Avantage 6.6 software to match the elemental composition and content in the material, examining both elemental fine spectra and whole spectra. The experimental data were analyzed for adsorption parameters in Microsoft Excel and then fitted to isotherm adsorption and site energy distribution isotherm trends in Origin 2024 software.

3. Results and Discussion

3.1. Competitive Adsorption Isotherms

The adsorption isotherms of both single and binary systems for Cd(II) and As(III) on PBS were examined; the results are displayed in Figure 1. As(III) and Cd(II) adsorption rose in both single and binary systems as their equilibrium concentrations rose, according to the data. Meanwhile, the adsorption isotherm of As(III) was higher than that of Cd(II), indicating that As(III) was more likely to be adsorbed by PBS.

To gain a deeper understanding of the adsorption characteristics, the adsorption data were evaluated using the Linear, Langmuir, and Freundlich isotherm models. The results are presented in Table S1. The Langmuir model focuses on a single adsorbent surface exhibiting unimolecular adsorption characteristics [34]. The interaction between pollutants and PBS in heterogeneous multilayer adsorption is described using the Freundlich model [30]. Based on the statistical parameter R², both the Freundlich and Langmuir models exhibit good fits to the adsorption data of Cd(II) and As(III) in the two systems [21,44]. The fitting parameter $q_{\max }$ derived from the Langmuir isotherm model indicates that the maximum adsorption capacities of PBS for Cd(II) and As(III) are 2.997 mg/g and 2.606 mg/g, respectively, which are relatively higher than the adsorption capacities of other reported conventional MPs (Table S2). However, in comparison to the Langmuir model, the Freundlich isotherm model yields the best fit with the lowest error value, suggesting that the adsorption process is primarily characterized by multilayer adsorption. The adsorption coefficient K_F indicates the ability of PBS to adsorb and bind to pollutants, whereas the parameter n reflects the adsorption site’s capacity or the adsorption process’s non-homogeneity [6,45]. In the Freundlich isotherm model, adsorption is primarily driven by chemical interactions when n < 1, while physical adsorption becomes the dominant process when n > 1 [46]. In this study, the n values for PBS adsorption of Cd(II) and As(III) were 1.460 and 1.652, corresponding to nonlinear adsorption, indicating that electrostatic attraction mainly governs the adsorption process [47]. Previous studies have shown that Cd is easily desorbed from the surface of MPs [48]. The interaction between MPs and Cd is weak, suggesting that chemical adsorption may not be the primary mechanism [20].

The K_F and n values of As(III) and Cd(II) in the binary system decreased compared to the single system, suggesting decreased adsorption ability and capacity. The results show that the two heavy metals in the binary system are adsorbed competitively and that As(III) inhibits the adsorption of Cd(II). This may be explained by differences in the two ions’ ionic radii and electronegativity. Greater electrostatic gravitational contact between As(III) and the microplastic surface results from As(III)’s higher electronegativity and lower ionic radius compared to Cd(II) [49,50]. Thus, the adsorption capacity of PBS for As(III) was greater than that of Cd(II). In binary systems, the presence of multivalent ions can lead to competing adsorption at surface sites, which may decrease the adsorption of metals in heterogeneous systems [51]. The discrepancy can also be caused by the fact that Cd(II) serves as a bridge between As(III) and PBS, but As(III) may have to compete with Cd(II) for adsorption sites on the PBS surface as well as with functional groups in PBS to form surface complexes [52]. In other words, the relationship between Cd(II) and As(III) can be both synergistic and competitive, depending on the number ratio between them and external variables [16,53].

3.2. Distribution of Competitive Adsorption Sites Energy

The site energy distribution theory provides essential insights into the distribution of low-energy and high-energy adsorption sites and surface inhomogeneities [54]. Figure 2 and Table S3 present the results of adapting the Freundlich model to PBS for Cd(II) and As(III) adsorption using SEDT simulations in both systems. The adsorption energy (E*) decreased significantly with an increase in $q_{\mathrm{e}}$, as shown in Figure 2a. Cd(II) and As(III) were shown to occupy the high-energy adsorption sites on PBS initially before moving to the low-energy sites [55]. Most of the Cd(II) and As(III) adsorption took place at locations with energies greater than 4.0 kJ/mol. Thermodynamically, this suggests that the adsorption of Cd(II) and As(III) on PBS is a physical process [56,57], which is consistent with the previous isotherm-fitting results.

Figure 2b,c shows the approximate potential energy distributions (F(E*)) of As(III) and Cd(II) adsorbed on PBS, derived using the Freundlich isotherm model. The absolute site frequency distribution value of As(III) was larger than that of Cd(II), indicating a concentrated distribution of high-energy sites with more adsorption sites having high adsorption energies for As(III) compared to that of Cd(II) [56]. Cd(II) exhibited lower E* values ranging from 4.11 to 12.75 kJ/mol in the single system compared to As(III) with a 4.09–12.76 kJ/mol range. In contrast, the E_m values of Cd(II) and As(III) decreased in the presence of both solutes, indicating that adsorption in the binary system affects the ability of Cd(II) and As(III) to occupy high-energy adsorption sites on PBS.

To assess the surface energy inhomogeneity of the adsorption system at the investigated temperatures, the average site energy (E_m) was employed to assess the interaction strength between Cd(II), As(III), and PBS microplastics. Additionally, the standard deviation (${\sigma }_{\mathrm{e}}^{*}$) was used to quantify the site energy inhomogeneity of the adsorbent materials [58]. The general consensus is that the adsorption affinity increases with E_m. Theoretically, the area beneath the site energy distribution curve correlates with the number of available adsorption sites. [59]. E_m values for Cd(II) and As(III) were 7.05 and 7.10 kJ/mol in the single system and 6.82 and 6.88 kJ/mol in the binary system, respectively, as Table S3 illustrates. PBS appears to have a stronger adsorption affinity for As(III), as indicated by the generally higher E_m values for As(III) compared to Cd(II) in both systems [60]. As a result, PBS absorbs As(III) in the binary system more easily than Cd(II), which is also reflected in the results of the competitive adsorption experiments. Furthermore, it was discovered that the number of adsorption sites and standard deviation (${\sigma }_{\mathrm{e}}^{*}$) for As(III) and Cd(II) in the binary system decreased by 29.11–51.22% and 0.38–0.54%, respectively. This suggested that the adsorption sites for As(III) were more evenly distributed in the binary system, whereas those for Cd(II) were more variably distributed. The results showed that the rivalry between heavy metals resulted in a considerable reduction in high-energy adsorption sites in the mixed system, and As(III) was significantly affected by Cd(II) in occupying the high-energy adsorption sites [61,62]. The SEDT effectively elucidates the alterations in single and binary systems during competitive adsorption, demonstrating that As(III) exhibits a higher competitive capacity.

3.3. Characteristics of PBS Microplastics and Point of Zero Charge

Microplastics demonstrate a variety of adsorption behaviors for heavy metals. The adsorption capacities of different MPs are influenced by factors such as their structure, hydrophobicity, functional groups, rubbery or glassy phases, particle size, and specific surface area [31]. PBS is a semi-crystalline polymer with amorphous and crystalline areas [12]. According to literature statistics, PBS has a specific surface area of 0.601 m²/g, a total pore volume of 0.00128 cm³/g, an average pore size of 10.57 nm, and a water contact angle of 74.09 [12,23]. Figure 3 displays the uneven shape of PBS, with a relatively flat and wrinkled surface that potentially offers lots of adsorption sites. PBS is an aliphatic copolyester with longer hydrocarbon repeating units and a more flexible molecular structure than the aliphatic homopolymer PLA [10]. Aliphatic polyesters can be broken down by the simple hydrolysis or enzymatic cleavage of their ester groups by enzyme activity. In addition, PBS is regarded as a rubber polymer [8]. Rubber polymers exhibit a pliable structure with numerous intermolecular cavities, and their molecular weight is closely linked to the active sites formed through molecular forces during adsorption [31].

The pH at the point of zero charge (pH_pzc) of an adsorbent refers to the pH at which its surface charge becomes neutral. Determining pH_pzc is essential in adsorption studies, as it provides insights into how variations in solution pH influence the interaction between the adsorbate and the adsorbent surface [63]. The adsorption of Cd(II) and As(III) onto MPs is pH-dependent, exhibiting opposite trends [52,64]. Zhou et al. [21] demonstrated that the adsorption capacity of various MPs, including PA, PVC, PS, ABS, and PET, for Cd(II) initially increases and then decreases within the pH range of 6.0–9.0. Conversely, studies on PTFE and PS microplastics for As(III) adsorption reveal a decrease in capacity with increasing pH [52,65]. This behavior is attributed to pH-induced changes in the ionization states and distribution of Cd(II) and As(III) in the solution [64]. Specifically, when pH < 6.0, Cd(II) predominantly exists in a simple ionic form, while at pH > 6.0, it is primarily found as Cd(OH)⁺ and Cd(OH)₂ precipitates [20]. Consequently, for MPs with a pH_pzc < pH, the negatively charged surface favors Cd(II) adsorption; however, as pH increases further, adsorption may decrease due to electrostatic repulsion and the formation of precipitates [21]. In contrast, the predominant species of As(III) shifts from neutral H₃AsO₃ to the negatively charged AsO₂₋ as pH increases, resulting in charge repulsion between negatively charged MPs and As(III), thereby reducing adsorption capacity [65,66]. Jiang et al. [23] reported that PBS consistently exhibits a negative surface charge across a pH range of 2–11, with the charge magnitude decreasing as the pH increases. Based on these findings, it can be inferred that the adsorption of Cd(II) and As(III) by PBS microplastics will vary with pH, and electrostatic interactions likely play a significant role in the adsorption process.

3.4. Adsorption Mechanism of Heavy Metals on PBS Microplastics

Heavy metals primarily adsorb onto MPs through various mechanisms, including electrostatic interactions, Van der Waals forces, hydrogen bonding, π-π bonding, complexation, and other mechanisms [44,50]. In addition, studies have confirmed the adsorption mechanisms of heavy metals, and adsorption behaviors may vary depending on the type of MPs and ambient conditions. Figure 4 presents XPS high-resolution spectra before and after PBS adsorption on Cd(II) and As(III) under various systems to explain the adsorption mechanism. After the adsorption of pollutants, there was a noticeable shift and reduction in the binding energy of O1s. This indicates a possible trapping process between the oxygen-containing functional groups of PBS and the pollutants, likely due to the π-π interactions and complexation between PBS and pollutants [17]. Figure 4a displays the original C1s spectra divided into four peaks at 248.80 eV, 286.07 eV, 287.44 eV, and 288.87 eV binding energies, representing C-C, C-O, C-O-C, and C=O groups, respectively [36,67]. Three distinct peaks located at 532.04 eV, 533.52 eV, and 534.97 eV were identified in the O1s spectra (Figure 4b), corresponding to the O=C-O, C-O-C, and C-O-H functional groups, respectively [12].

The ratios of elements in different bound states in C1s and O1s were computed using the area of the deconvoluted peaks to the total peak area (Table S4). The C1s spectra indicated a shift in the energy level of the oxygen-containing functional group C-O from 286.07 eV to a range of 286.17 to 286.32 eV. The increase in the percentage of C=O content (Table S4) suggests that the adsorption of Cd(II) and As(III) on MPs’ surfaces could initiate polymer reactions, leading to the formation of more C=O groups [21]. The O1s spectrum demonstrated a decrease in the concentration of O=C-O from 77.28% to a range of 67.42% to 52.55%, suggesting that certain carboxyl groups may form complexes with As(III) and Cd(II) during the adsorption process [12]. Fei et al. [68] investigated the adsorption of Cd(II) on biodegradable microplastic PLA and its co-migration in saturated porous media under various conditions. Electrostatic forces and the surface complexation process of oxygen functional groups were discovered to impact the adsorption of Cd(II) on PLA. On the other hand, Zhang et al. [44] researched the adsorption of As(III) on PS and PA alongside antibiotics. The study suggested that As(III) adsorption could be linked to hydrogen bonding, electrostatic interactions, surface complexation, and hydrophobic interactions. Likewise, similar to the findings of these studies, we propose that the Cd(II) and As(III) adsorption onto biodegradable PBS microplastics was predominantly governed by physical and chemical mechanisms, primarily involving electrostatic interactions and surface complexation.

3.5. Theoretical Computational Analysis

3.5.1. GCMC Method Simulation of Adsorption

Theoretical calculation techniques were employed to explain the adsorption discrepancies at the molecular level, providing a more comprehensive understanding of the adsorption process between MPs and heavy metals [69]. The study utilized the GCMC method to model the changes in the adsorption of Cd(II) and As(III) by PBS in both single and binary systems. Under neutral conditions, the predominant compounds present in water with a pH below 9.2 are arsenites (As(III)) and cadmium hydroxide (Cd(II)) [16,18].

Adsorption calculations were conducted using DFT-optimized PBS compounds with Cd(II) and As(III). The simulated energy distributions are depicted in Figure S2. The adsorption process is spontaneous because of the negative values of the interaction energies [70]. For both single and binary systems, the adsorption energy of As(III) ranges from −220 to −90 kcal/mol (−340 to −91 kcal/mol), whereas the adsorption energy of Cd(II) ranges from −140 to −50 kcal/mol (−260 to −50 kcal/mol). The findings demonstrated that As(III) showed higher adsorption affinity and held a prominent role in competitive adsorption systems [71]. The theoretical study of the site energy distribution and the experimental results further substantiate this.

3.5.2. DFT Calculations

DFT simulations explored the electronic structure and system energy relationship to assess the adsorption mechanism between PBS and heavy metals [41]. Figure S3 illustrates the optimal arrangement of PBS microplastics. Further optimization studies were conducted based on the original C=O binding site identified through Monte Carlo simulations and the potential adsorption sites (C-O-C and hydroxyl O-H) identified through XPS research.

Table S5 presents Cd(II) and As(III) binding energies at different sites in PBS microplastics. A greater absolute value of E_ad corresponds to a higher binding energy value, indicating increased stability of the adsorbent and adsorbate [38]. The highest binding energies for As(III) and Cd(II) were observed when they were located at the hydroxyl and C=O adsorption sites, respectively, after comparing their binding energies at different locations. PBS and As(III) exhibited the most significant binding impact, with a binding energy of −11.219 kcal/mol. This discovery aligns with the experimental data and suggests that As(III) exhibits a higher preferential adsorption status than Cd(II). Cd(II) has the greatest attraction for the adsorption site C=O, followed by OH and COC. Before structural optimization, the Cd-O and O-H bond lengths were 3.558, 4.431, and 2.945 Å, respectively. After structural optimization, the bond lengths decreased to 2.466, 3.843, and 2.298 Å, respectively, from their initial values. PBS and Cd(II) exhibited coordination and electrostatic interactions, as evidenced by the decreased Cd-O and O-H bond lengths [41]. In short, As(III) tends to dwell at the hydroxyl adsorption site, whereas Cd(II) tends to adsorb at the C=O site.

The HOMO–LUMO energy band gap indicates the molecule’s reactivity, stability, and electrical conductivity [72]. The frontier molecular orbitals, specifically the HOMO and LUMO orbitals, were extensively analyzed in various systems. Figure 5 displays the fitting findings. A decreased HOMO–LUMO energy gap between two reactants typically indicates poorer stability and increased chemical reactivity in the interaction, as the lower LUMO readily accepts electrons from the higher HOMO [73]. The reactivity parameters have been calculated and are displayed in Table S6. The calculated energy gap of PBS and Cd(II) was smaller than that of As (III), suggesting that the combination is more chemically reactive but less stable [74]. Figure 5 illustrates the complete transfer of the lowest unoccupied molecular orbital (LUMO) from PBS to Cd(II) in PBS-Cd (C=O) interactions. The partial transfer was also observed in the highest occupied molecular orbital (HOMO) of PBS-As (COC) but not towards the heavy metal. The findings suggest a chemical reaction occurs between PBS and Cd(II), while the interactions between PBS and As(III) are primarily physical [39].

Additionally, this study conducted an electrostatic potential (ESP) analysis of PBS and its adsorbates (Figure 6), which determines the distribution of reactive molecule charges and is utilized to analyze reaction sites [75]. The red and blue hues in the figure reflect positive and negative potentials, corresponding to electrophilic and nucleophilic regions, respectively. The color gradient corresponds to the intensity of the potentials. The negatively charged surfaces of MPs are expected to facilitate the adsorption of positively charged metal contaminants [76]. The oxygen-containing functional groups (C=O, C-O-C, and -OH) on the structure of pristine PBS are mainly located in blue regions, as shown in Figure S3b. These areas were more nucleophilic, suggesting that the oxygen-containing functional groups of PBS could serve as appropriate adsorption sites for Cd(II) and As(III) [39]. As(III) exhibited a higher positive electrostatic potential than Cd(II), as seen in Figure 6c. The difference in electronegativities between As (1.69) and Cd (2.18) atoms results in a directional shift of electron density [77]. Arsenic, in its trivalent form, displayed superior adsorption properties and the most significant affinity for the hydroxyl functional group in competitive adsorption trials with heavy metal ions. Meanwhile, Cd(II) demonstrated the highest affinity for C=O. The ESP plot shows the significance of the differences between red and blue. Significant variations in positive and negative charges were observed after adsorption at various sites, especially with As-(OH) and Cd-(C=O) exhibiting the most notable differences, indicating a more potent binding effect, consistent with the findings of the binding energy analysis. The ESP pictures of PBS show a more prominent red area following heavy metal adsorption, indicating that PBS is affected by the electrophilicity of Cd(II) and As(III).

4. Conclusions

This study examined the adsorption properties of biodegradable microplastic PBS for Cd(II) and As(III) in both single and binary systems. The results of isotherm data fitting indicated that the experimental data were more consistent with the Freundlich adsorption isotherm model. The adsorption capacity of PBS microplastics for Cd(II) in a single system increased due to electrostatic interactions, with the Langmuir model showing the maximum adsorption amount, reaching 2.977 mg/g. In the competition experiment involving the coexistence of heavy metals, the adsorption order of PBS was As(III) > Cd(II), and the presence of As(III) inhibited the adsorption of Cd(II). As(III) tended to occupy high-energy sites, while Cd(II) was more inclined to occupy low-energy sites. The results of model fitting and theoretical analyses of site energy distribution indicated that the physical mechanism primarily affects the adsorption process of the two heavy metals. The results of DFT calculations were also consistent with the experimental findings. FMOs analyses showed that electron transfer and chemical reactions occurred between the adsorbent and the heavy metal ions. The results of binding energy and electrostatic potential analyses, including the major adsorption sites and binding strength analyses of As(III) and Cd(II), further confirmed the competitive adsorption differences observed experimentally.