High-performance magnetic biochar prepared via acid and mg/Fe Co-modification for ultraefficient Pb (II) adsorption

Summary

In this study, MgO-containing magnetic composite biochar (MBC) was prepared from activated corn stover for the efficient removal of Pb²⁺. Through the introduction of magnesium and iron ions, the surface and pore structures of the acid-treated corn stover biochar adsorbent were optimized, with its adsorption capacity being enhanced to 253.6 mg g⁻¹. MBC showed excellent adsorption performance and monolayer chemisorption, as evidenced by its pseudo-second-order kinetics and conformance with the Langmuir isotherm model. Moreover, it could maintain >85% efficiency even after five regeneration cycles. Mechanistic studies revealed that physical adsorption, electrostatic attraction, hydrogen bonding, and ion exchange were the key pathways of MBC-mediated Pb²⁺ removal. Finally, XRD and XPS analyses confirmed the occurrence of Pb (OH)₂ precipitation and Pb-O coordination during removal. This study represents an advance in sustainable remediation technologies, providing an eco-friendly solution for managing water pollution and agricultural waste.

Graphical abstract

Highlights

• •Magnetic biochar (MBC) synthesized via acid/Mg/Fe co-modification for Pb (II) adsorption
• •Ultrahigh Pb (II) adsorption capacity of 253.6 mg g⁻¹ achieved with optimized MBC
• •Adsorption follows pseudo-second-order kinetics and Langmuir isotherm
• •Agricultural waste resourceful use for water pollution control fits green development

Environmental science; Materials synthesis; Materials characterization

Introduction

In recent years, human activities such as industrial production, smelting, and the operation of landfills have significantly contributed to the contamination of water resources by heavy metals, posing a serious threat to public health.^1,2,3,4,5 One such heavy metal, Pb (II), is one of the most prevalent and toxic pollutants found in water bodies.⁶ Even trace amounts of Pb (II) can lead to a variety of health issues, as it is not readily eliminated through physiological processes and tends to exhibit bioaccumulation. Both prolonged exposure and high levels of ingestion can cause symptoms such as headaches, diarrhea, liver dysfunction, and, in severe cases, damage to the central nervous system.^7,8 Consequently, the development of environmentally friendly and sustainable methods for removing Pb (II) from aquatic systems is of critical importance.

The primary techniques for removing heavy metals from water include evaporation, adsorption, membrane separation, biological treatment, chemical precipitation, coagulation, and ion exchange.^9,10 However, most of these methods are cost-intensive and often provide low removal efficiencies. Overall, adsorption is favored due to its cost-effectiveness, selectivity, environmental friendliness, and simplicity of operation.¹¹ Consequently, substantial research efforts have been directed toward developing more efficient adsorbents and optimizing their preparation to enhance both safety and environmental sustainability. Biochar, a carbon-rich material produced via the thermochemical conversion of biomass under hypoxic conditions, is commonly used as an adsorbent for water treatment.¹² Biochar has gained considerable attention due to its environmental benefits, widespread availability, and low production costs. In particular, biochar derived from animal manure is known to demonstrate strong adsorption capacity for heavy metals, particularly Pb (II), due to its high polarity and ash content. The primary adsorption mechanisms of manure-derived biochar include the formation of precipitates such as lead phosphate and lead carbonate.^13,14,15 However, certain types of livestock manure contain high concentrations of antibiotics, necessitating extensive pretreatment before biochar production. This additional step increases the overall cost of production.¹⁶

Lignin, a complex polymer rich in functional groups, can chelate heavy metals to form lignin–metal complexes.¹⁷ However, the adsorption capacity of pure lignin for heavy metals is relatively low. For instance, alkali lignin adsorbs only 2.6 mg g⁻¹ of Pb (II), while lignosulfonate adsorbs just 27.1 mg g⁻¹ of Pb (II). Moreover, the poor separation of lignin in aqueous mediums can lead to secondary pollution, limiting its industrial application.¹⁸ Nevertheless, studies have highlighted the immense potential of agricultural wastes in the large-scale treatment of wastewater, especially for sequestering heavy metals.¹⁹ The physicochemical properties of these wastes vary depending on their composition, resulting in differing heavy metal adsorption capacities. In particular, biochar adsorbents produced from materials such as hazelnut shells, coffee grounds, rice husks, tea leaves, and corn stover have all shown promising adsorption capabilities for heavy metals.^20,21,22,23 Among these, corn stover is particularly favored due to its widespread availability and low cost. Composed primarily of cellulose, hemicellulose, and lignin,²⁴ corn stover is considered well-suited for the removal of heavy metal ions. Functionalization (for instance, with CaFe-layered double hydroxides) has been found to considerably improve the heavy metal adsorption capacity of biochar.²⁵ Moreover, one study showed that pyrolyzing corn stover at 500°C facilitates the high-value transformation of waste materials, reducing raw material costs and minimizing the environmental impact typically associated with open-field incineration.²⁶ When compared to equivalent dosages of other agricultural wastes such as bagasse, corn stover exhibits superior performance in heavy metal removal.²⁷ Therefore, corn stover biochar could serve as a highly efficient, cost-effective, and environmentally benign agent for heavy metal adsorption and the remediation of contaminated water.

Despite the potential of biochar as an adsorbent, its underdeveloped pore structure and low specific surface area hinder the effective elimination of heavy-metal pollutants, and biochar often has insufficient surface functional groups and adsorption sites.²⁸ To overcome these limitations, researchers have modified biochar to enhance its surface chemistry, specifically by introducing functional groups such as hydroxyl, aldehyde, ketone, and phenolic groups. These modifications promote the complexation of heavy metals on the surface of biochar, thereby increasing its adsorption capacity.²⁹ This study focused on enhancing the Pb (II) adsorption efficiency of corn stalk-derived biochar through acid treatment and the incorporation of magnesium and iron ions. These modifications sought to address the aforementioned limitations related to the physical and chemical properties of biochar.³⁰ Key operational parameters, including adsorbent dosage, reaction temperature, and solution pH, were systematically examined to understand their specific roles in the adsorption process. Additionally, the reusability of the modified biochar was evaluated. This study offers an efficient and low-cost solution for the management of heavy metal pollution.

Experimental methods

Preparation of modified biochar

Corn stalks were crushed and washed to remove impurities before drying at 80°C. The corn stover was ground and filtered through a 60-mesh sieve. The washed and ground corn stover was placed in a muffle furnace and heated for 2 h (temperature increased at a rate of 10°C·min⁻¹) to obtain pyrolyzed corn stover (BC). This BC was then washed and dried.

Meanwhile, some washed and ground corn stover was soaked in 1.2 mol L⁻¹ citric acid for 24 h, dried, and then washed to neutrality. It was then pyrolyzed in a muffle furnace at 500°C for 2 h (temperature increased at a rate of 10°C·min⁻¹) to produce acid-modified biochar (HBC). Subsequently, 1 g of HBC was mixed with MgCO₃ and 0.25 g of Fe₃O₄, ultrasonicated for 30 min, and transferred to a high-pressure reactor to react at 150°C for 5 h. The product was washed with anhydrous ethanol and deionized water, dried at 80°C, and stored as magnetic composite biochar (MBC) for future use.³¹ Figure 1 shows the adsorption agent synthesis diagram.

Figure 1.

Schematic illustration of material synthesis

Analysis of adsorption properties

Intermittent adsorption tests

Adsorption studies were conducted by introducing 20 mL of a Pb (II) solution (100 mg L⁻¹) into a 50 mL conical flask along with 0.02 g of the adsorbent. The pH was adjusted to 6, and the mixture was shaken at 303 K and 180 r·min⁻¹ using a thermostatic shaker. Subsequently, experiments were designed to assess the influence of adsorbent dosage (ranging from 5 to 50 g L⁻¹), temperature (between 303 and 323 K), and pH (from 2 to 7) on the rate of adsorption. Upon the completion of the adsorption process, samples were collected to determine the Pb (II) concentration. Each test was repeated thrice to ensure accuracy, and the average of three measurements was calculated for analysis. The adsorption capacity (Equation 1) and removal rate (Equation 2) were calculated using the following equations:

$$\begin{array}{c}\mathrm{q}=(C_0-C)V/m\end{array}$$ (Equation 1)

$$\begin{array}{c}\mathrm{R}=\left[\frac{C_0-C}{C_0}\right]\times 100\%\end{array}$$ (Equation 2)

Here, R represents the removal rate (%); C₀, the initial concentration of the heavy metal (mg·L⁻¹); C, the concentration of the heavy metal remaining in the solution after adsorption equilibrium is reached (mg·L⁻¹); q, the amount of adsorption (mg·g⁻¹); V, the volume of the solution (mL); and m, the mass of the adsorbent (mg).

Adsorption kinetics

In a 50 mL conical flask, 0.02 g of MBC was mixed with 20 mL of a 100 m L⁻¹ Pb(II) solution. The flask was then placed in a thermostatic oscillator at a shaking speed of 180 r·min⁻¹, with the temperature maintained at 303 K and the pH maintained at 6. The Pb(II) levels were analyzed using atomic absorption spectrometry at intervals of 5, 10, 30, 60, 90, 120, and 150 min. Three replicate measurements were obtained, and the average was calculated for subsequent analysis. The quasi-primary kinetic model (Equation 3), quasi-secondary kinetic model (Equation 4), and intraparticle diffusion model (Equation 5) were used to determine the Pb (II) adsorption mechanisms of modified biochar.

$$\begin{array}{c}{q}_t=q\left(1-e^{-k_{1}t}\right)\end{array}$$ (Equation 3)

$$\begin{array}{c}{q}_t=q^2{k}_{2}t/(1+q{k}_{2}t)\end{array}$$ (Equation 4)

$$\begin{array}{c}{q}_t=K_{di}{t}^{\frac{1}{2}}+C_i\end{array}$$ (Equation 5)

Here, q_t is the amount of heavy metals adsorbed by modified biochar at moment t (mg·g⁻¹); t is the adsorption time (min); k₁ and k₂ are adsorption rate constants (min⁻¹ and g·mg⁻¹·min⁻¹, respectively); K_di is the intraparticle diffusion constant (mg·g⁻¹·min^−1/2); and C_i is the boundary layer constant.

Isothermal adsorption experiments

First, 0.02 g of the adsorbent was mixed with 20 mL of a Pb (II) solution (concentration ranging from 20 to 450 mg L⁻¹). The mixtures were then agitated at a pH of 6 and speed of 180 r·min⁻¹ for a duration of 120 min, with the temperature maintained at a constant value of 303 K. At the end of the experiment, samples were removed to measure the concentration of Pb(II), and three measurements were averaged before fitting with Langmuir’s model (Equation 6) and Freundlich’s model (Equation 7).

$$\begin{array}{c}{q}_e=q_m[K_{L}C/(1+K_{L}C)\end{array}$$ (Equation 6)

$$\begin{array}{c}{q}_e=K_F{C}^{\frac{1}{n}}\end{array}$$ (Equation 7)

Here, q_m is the maximum adsorption amount (mg·g⁻¹); K_L is the Langmuir adsorption equilibrium constant; K_F is the Freundlich adsorption equilibrium constant; and n is the Freundlich constant.

Adsorption thermodynamics

The effect of temperature on adsorption was explained based on Gibbs free energy (ΔG), enthalpy changes (ΔH), and entropy changes (ΔS). The following equations were utilized.

$$\begin{array}{c}{k}_d=q/C_0\end{array}$$ (Equation 8)

$$\begin{array}{c}\Delta \mathrm{G}=\Delta \mathrm{H}-\mathrm{T}\Delta \mathrm{S}\end{array}$$ (Equation 9)

$$\begin{array}{c}\mathit{ln}{k}_d=\Delta \mathrm{H}/\left(\text{RT}\right)+\Delta S/R\end{array}$$ (Equation 10)

Here, k_d is the equilibrium constant (mL·g⁻¹), R = 8.314 J mol⁻¹ K⁻¹, and T is the reaction temperature (K).

Regeneration experiments

To study the reusability of the adsorbent, we reacted the dried adsorbate with 0.1 mol L⁻¹ HCl on a mixer for 30 min, washed with distilled water until neutral, and dried. Subsequently, its adsorption capacity was tested again. The adsorption–regeneration test was repeated ten times in three parallel replicates.

Results and discussion

Characterization of biochar

The morphology and elemental distribution of BC, HBC, and MBC were analyzed using SEM-EDS (Figures 2A–2G). Compared with the original corn stover (Figure 2A), the citric acid-treated corn stover exhibited enlarged pores and a rougher surface (Figures 2B and 2C). Moreover, HBC had the highest specific surface area (164.3 m² g⁻¹), followed by MBC (42.6 m² g⁻¹) and BC (5.9 m² g⁻¹). The specific surface area of HBC was higher than that of MBC because Mg and Fe filled the pores of MBC. Subsequently, elemental mapping revealed the presence of Fe and Mg (Figures 2D and 2E) in MBC, and distinct MgO particles could also be detected (Figure 2C). The results demonstrated that magnesium and iron were effectively incorporated onto the surface of corn stover, which not only increased the specific surface area but also generated new active sites. These improvements significantly amplified the ability of MBC to adsorb Pb (II) ions. Figure 2H presents the VSM magnetization curve of MBC at room temperature, revealing a saturation magnetization value of 24.2 emu·g⁻¹. These findings indicated that the synthesized biochar possesses magnetic properties and can thus be separated easily from aqueous solutions using a magnet.

Figure 2.

Characterization and Analysis of BC, HBC, and MBC

SEM images of BC (A), HBC (B), and MBC (C). Elemental maps of Mg (D), Fe (E), C (F), and O (G). Magnetization curve of MBC (H). Thermal decomposition curves for BC, HBC, and MBC (I).

The results of TGA are shown in Figure 2I. Notably, the weight loss rates of BC, HBC, and MBC were approximately 23%, 45%, and 45%, respectively, and the weight loss process could be divided into three stages. The first stage occurred between 30°C and 150°C and corresponded to mass losses of approximately 9.05%, 8.72%, and 11.96% in BC, HBC, and MBC, respectively. The mass loss in this stage primarily occurred due to volatilization and moisture loss. The second weight loss stage occurred at 150°C–640°C and 150°C–680°C, respectively, with HBC experiencing a mass loss of approximately 20.13% and BC and MBC exhibiting mass losses of 15.95% and 22.97%, respectively. The mass loss in this stage was primarily due to the decomposition of functional groups and organic matter.³² In the final stage, the weight loss rates for BC, HBC, and MBC were 7.80%, 10.21%, and 13.5%, respectively. The mass loss in this stage indicated that the materials exhibit good high-temperature stability.³³

Figure 3A displays the pre-reaction XRD patterns of BC, HBC, and MBC. In MBC, a diffraction peak corresponding to MgO crystals appeared at 2θ = 38.1°. Additionally, MBC exhibited diffraction peaks at 2θ = 44.4° and 64.7°, consistent with the standard pattern of Fe₃O₄. This indicated the successful loading of Mg and Fe onto MBC. In contrast, BC exhibited the characteristic peaks of SiO₂ crystals at 2θ = 20.7°, 26.6°, 27.7°, and 59.8°. After acid leaching treatment, some of the SiO₂ in HBC was decomposed, leading to the disappearance of certain SiO₂ characteristic peaks and the emergence of new peaks.³⁴ A comparison of the XRD patterns of BC, HBC, and MBC revealed significant alterations in the crystal structure of MBC, particularly due to the successful incorporation of MgO and Fe₃O₄. Meanwhile, HBC showed the partial decomposition of SiO₂ and the appearance of new characteristic peaks.

Figure 3.

Pre-reaction Structural and Spectral Analysis of BC, HBC, and MBC

Pre-reaction XRD analysis of BC, HBC, and MBC (A). Pre-reaction gross XPS spectra of BC, HBC, and MBC (B). C1s (C), O1s (D), Mg1s (E), and Fe2p (F) splits.

The full-range XPS spectra of BC, HBC, and MBC before the reaction are shown in Figure 3B. A comparative analysis revealed the presence of new peaks corresponding to Mg1s and Fe2p in MBC, indicating the successful incorporation of MgO onto corn stover via magnetic iron oxide. Figures 3C–3F present the detailed high-resolution spectra for C1s, O1s, Mg1s, and Fe2p, respectively. In the XPS analysis of the C1s region, the C―C bonds in BC, HBC, and MBC consistently registered a binding energy of 284.7 eV. Meanwhile, the C―O―C bond exhibited binding energies of 285.8 eV, 285.4 eV, and 285.7 eV, while the O―C = = O bond showed values of 289.5 eV, 289.4 eV, and 289.1 eV, respectively. In the XPS curve of the O1s region, the binding energies for the C―O bond were 531.5 eV, 532.2 eV, and 531.9 eV for BC, HBC, and MBC, while the binding energies for the C = = O bond were 533.1 eV, 533.9 eV, and 532.9 eV, respectively. Additionally, the binding energies for the Mg―O and Fe―O bonds in MBC were 530.5 eV and 530.0 eV, respectively.³⁵ The Mg1s spectrum revealed the formation of MgO in MBC,³⁶ while the Fe2p orbitals confirmed the presence of Fe in various oxidation states.³⁷ A comparison of the XPS data for BC, HBC, and MBC clearly demonstrated significant variations in the chemical bond pattern and elemental composition of MBC.

Adsorption efficiency of MBC prepared under different conditions

To examine how the mass ratio of biochar to Fe₃O₄ affects the Pb (II) adsorption performance, biochar was modified with different mass ratios of Fe₃O₄ (1:3, 1:2, 1:1, 2:1, and 3:1). The biochar samples obtained (0.02 g each) were added to 20 mL of a Pb (II) solution (100 mg L⁻¹), reacted at room temperature with shaking (180 r·min⁻¹) for 2 h, and then tested (Figure 4A). When the mass ratio of HBC to Fe₃O₄ was 2:1, the adsorption capacity reached its maximum value (94.77 mg g⁻¹). However, as the mass ratio of Fe₃O₄ increased, the adsorption capacity decreased to 89.62 mg g⁻¹, potentially because excessive Fe₃O₄ covered the surface of the biochar, thus decreasing its specific surface area. Accordingly, modified biochar with a Fe₃O₄ mass ratio of 2:1 was selected for subsequent experiments.

Figure 4.

Effect of Fe3O4 and MBC on Pb(II) Adsorption

The effect of different proportions of Fe3O4 and MBC on the adsorption of Pb (Ⅱ) (A); Effect of different pyrolysis temperatures and times on the adsorption of Pb (Ⅱ) (B and C).

The MBC samples were placed in a muffle furnace for pyrolysis treatment at temperatures of 200°C, 300°C, 400°C, 500°C, and 600°C for 2 h each. Then, 0.02 g of the MBC sample was added to 20 mL of a 100 mg L⁻¹ Pb (II) solution before shaking at room temperature and 180 r·min⁻¹ for 2 h. Finally, the supernatant was collected to determine the Pb (II) concentration of the solution (Figure 4B). The adsorption capacity of Pb (II) was found to increase as the pyrolysis temperature increased, with adsorption capacities of 43.36 mg g⁻¹, 56.61 mg g⁻¹, 80.77 mg g⁻¹, 94.32 mg g⁻¹, and 92.30 mg·g⁻¹ detected at pyrolysis temperatures of 200°C, 300°C, 400°C, 500°C, and 600°C, respectively. As the temperature rises, the organic components in corn stalks, such as cellulose and lignin, break down, resulting in the formation of more porous structures. However, there was no significant change in the adsorption capacity between the 500°C and 600°C groups. Therefore, considering factors such as yield and energy consumption, 500°C was selected as the optimal pyrolysis temperature.

Subsequently, pyrolysis durations of 60, 90, 120, 150, and 180 min were used to prepare MBC samples, and 0.02 g of each sample was added to 20 mL of a 100 mg L⁻¹ Pb (II) solution. After shaking at room temperature and 180 r·min⁻¹ for 2 h, the supernatant was collected for concentration measurements (Figure 4C). The pyrolysis time was too short, resulting in incomplete reactions and low adsorption capacity of the adsorbent for Pb (II). As the pyrolysis time increased, the structure of the corn stover exhibited higher stability and adsorption performance. Notably, the Pb (II) adsorption capacity of MBC first increased and then decreased with an increasing pyrolysis duration. The adsorption capacity of Pb (II) peaked at 96.7 mg g⁻¹ when the pyrolysis time was 120 min.

Factors affecting adsorption

pH levels have a critical impact on heavy metal ion speciation and can modify the surface charge of adsorbents, influencing adsorption dynamics. In this study, the Pb (II) removal efficiency and zeta potential of BC, HBC, and MBC were examined within the pH range of 2–7, as shown in Figures 5A and 5B. The pH_pzc of BC was 5.9, while that of HBC and MBC was 6.1. MBC achieved a maximum adsorption capacity of 93.91 mg g⁻¹ at pH 6, while HBC and BC reached their maximum adsorption capacities at pH 7 (72.45 mg g⁻¹ and 60.93 mg g⁻¹, respectively). At pH 2, the Pb (II) adsorption capacity of biochar was weak. This could be attributed to H⁺ ions, which occupied the adsorption sites and induced the protonation of functional groups on the surface of biochar. This created a positively charged surface, leading to electrostatic repulsion.^38,39,40 When the pH value was greater than 6.1, the surfaces of all three adsorbents carried a negative charge. As a result, adsorption was enhanced through electrostatic attraction, leading to a gradual increase in adsorption capacity. Even when the zeta potential was greater than zero, the adsorbents exhibited strong adsorption capacity for Pb (II).

Figure 5.

Factors Influencing Pb(II) Adsorption by Biochar and MBC

Effect of pH (A) and the zeta potential (B); dosage (C) on the adsorption capacity of biochar; The effect of different coexisting ion concentrations on the adsorption of Pb (Ⅱ) by MBC(D); Regeneration cycle of biochar(E); Effectiveness of MBC in removing Pb from real water body(F); quasi-primary kinetic model (G); quasi-secondary kinetic model (H); intraparticle diffusion model fitting curves (I); isothermal fitting curves for Pb adsorption (J and K).

The effect of biochar dosage (5–50 mg L⁻¹) on Pb (II) adsorption was also explored (Figure 5C). When the adsorbent dosage increased to 10 mg L⁻¹, the adsorption of Pb (II) by MBC, HBC, and BC reached a maximum of 94.07 mg g⁻¹, 65.67 mg g⁻¹, and 58.97 mg g⁻¹, respectively. This enhancement was attributed to the increase in available active sites at higher adsorbent dosages, which significantly accelerated the Pb (II) removal rate.⁴¹ However, as the adsorbent dosage continued to increase to 50 mg L⁻¹, the aggregation of adsorbent particles reduced the available surface area, causing a gradual decline in adsorption capacity.⁴² Considering both the adsorption capacity and economic cost, a dosage of 10 mg L⁻¹ was selected as the optimal concentration for Pb (II) removal.

To examine the effects of common wastewater ions on the adsorption process, a mixed solution containing 100 mg L⁻¹ of Pb (II), Mg (II), Na(I), K(I), and Ca (II) was prepared. Then, 0.02 g of MBC was mixed with 20 mL of this solution before shaking at room temperature and 180 r·min⁻¹ for 2 h. Then, the concentration of each ion in the solution was measured (Figure 5D). The results indicated that the presence of other ions had no significant effect on the adsorption of Pb (II) by MBC, as Pb ions showed stronger adsorption affinity than other cations and could preferentially occupy adsorption sites.⁴³

The regeneration cycle of an adsorbent is crucial for cost efficiency and environmental sustainability.⁴⁴ In this study, biochar was subjected to treatment using 0.1 M HCl to extract Pb (II); the outcomes are depicted in Figure 5E. After the reaction, the adsorbent material was recovered using a magnet. Across 10 cycles, the Pb (II) adsorption efficiency of MBC decreased from 98.76 mg g⁻¹ to 71.52 mg g⁻¹, and the recovery rate decreased from 98.78% to 89.23%. The results showed that MBC has good recyclability.

Finally, surface water samples from Ordos Saline Lake —obtained through 7 consecutive days of sampling —were used to evaluate the Pb (Ⅱ) removal efficacy of MBC. Table 1 shows the chemical characteristics of water samples collected from Ordos Saline Lake, where the concentration of Pb (Ⅱ) exceeds the Class V threshold (0.1 mg L⁻¹) specified in the Surface Water Environmental Quality Standards of China. At an initial pH of 7.8 and temperature of 303 K, 0.1 g MBC adsorbent was added to 100 mL of each water sample, and adsorption was carried out by shaking at 180 r·min⁻¹ for 30 min in a constant temperature oscillator. Figure 5F shows the ability of MBC to remove Pb (Ⅱ) from the water samples. At the end of the reaction, the concentration of Pb (Ⅱ) in the water samples was less than 0.003 mg L⁻¹, within the quality threshold (0.01 mg L⁻¹) for Class I water bodies in China. This showed that MBC could maintain excellent adsorption performance for Pb (II) in environments with high salinity and alkalinity, without any need to adjust the physicochemical parameters of the water body. Therefore, MBC could be used as an adsorbent for water purification.⁴⁵

Table 1.

Chemical characterization of water samples from the Ordos Saline Lake

Date of Sampling	Pb (mg·L−1)	Cd (mg·L−1)	Cu (mg·L−1)	DO (mg·L−1)	TN (mg·L−1)	TP (mg·L−1)
3.30	0.175	0.064	0.09	9.5	39.7	3.6
3.31	0.163	0.055	0.08	9.5	40.2	3.4
4.1	0.203	0.059	0.09	9.7	38.4	3.6
4.2	0.184	0.042	0.07	9.4	39.2	3.8
4.3	0.177	0.061	0.07	9.5	38.6	3.8
4.4	0.172	0.043	0.04	9.6	40.1	3.9
4.5	0.173	0.052	0.10	9.5	39.5	3.5

Adsorption kinetics

The physicochemical properties of adsorbents significantly influence the adsorption process. The adsorption process of all types of biochar follows a typical trend. Specifically, it is characterized by an initial rapid adsorption phase, followed by a slower adsorption phase and then an equilibrium stage. This behavior can be attributed to the large abundance of adsorption sites in the initial stages, enabling the quick uptake of Pb(II) from the aqueous phase onto the biochar surface.⁴⁶ As adsorption sites are gradually occupied and the structure of biochar becomes more compact, the uptake efficiency for metal ions shows a steady decline.⁴⁷ In this study, the adsorption of Pb (II) by MBC, HBC, and BC reached equilibrium within 150 min, with adsorption capacities of 98.83 mg g⁻¹, 71.64 mg g⁻¹, and 56.23 mg g⁻¹, respectively. The adsorption efficiency was the highest for MBC, followed by HBC and BC. To investigate the Pb (II) adsorption behavior of various types of biochar, the pseudo-first-order, pseudo-second-order, and intraparticle diffusion models were applied. As depicted in Figures 5G and 5H and Table 2, kinetic fitting outcomes consistent with the pseudo-first-order model indicate the predominance of physical adsorption. Meanwhile, in the pseudo-second-order model, chemical interactions primarily drive the adsorption process.^48,49 As illustrated in Table 1, the R² values associated with the pseudo-second-order model were consistently higher than those of the pseudo-first-order model in this study. This disparity strongly indicated that the adsorption of Pb (II) onto various types of biochar was largely mediated by chemical adsorption processes.¹⁰

Table 2.

Kinetic fitting parameters

Model	Parameters	Fitting results
		BC	HBC	MBC
Actual adsorption capacity	Qe,exp/(mg·g−1)	56.23	71.64	98.83
Quasi-primary kinetic model	Qe,cal/(mg·g−1)	55.109	66.114	97.991
	k1/(mg·g−1)	0.067	0.063	0.0893
	R2	0.902	0.764	0.776
Quasi-secondary kinetic model	Qe,cal/(mg·g−1)	56.783	74.606	104.415
	K2/(mg·g−1)	0.002	0.001	0.001
	R2	0.939	0.937	0.929
Intra-particle diffusion model	C1	14.770	19.253	36.255
	Kd1/(mg·g−1·min−1/2)	3.777	5.354	6.680
	R2	0.954	0.968	0.981
	C2	27.181	51.360	81.763
	Kd2/(mg·g−1·min−1/2)	2.463	1.715	1.541
	R2	0.906	0.841	0.724

The intraparticle diffusion model was also applied to analyze the diffusion processes and investigate the adsorption kinetics of the biochar. Based on this model, adsorption can be classified into three phases: external diffusion, internal diffusion, and equilibrium.⁵⁰ The curve in Figure 5I shows two linear segments. The curve indicated that in the first stage, Pb (II) was rapidly absorbed onto the surface of biochar through external diffusion. Hence, the surface area and pore characteristics of biochar significantly influenced its adsorption process. The second stage, representing the intraparticle diffusion phase, was characterized by the gradual adsorption of Pb (II) into the interior of the biochar particles. The slower adsorption rates suggested that Pb (II) diffusion within the porous structure of the biochar was more restricted than the initial external diffusion process. The intraparticle diffusion rate constants for Pb (II) followed the order K_d1 > K_d2, indicating that in the initial stages, the surface adsorption sites of biochar were more easily occupied by Pb (II), while the diffusion of Pb (II) into deeper pores was slower and more restricted. The non-zero Ci values across all stages indicated that intraparticle diffusion was not the sole rate-limiting factor in the kinetic process.⁵¹

Adsorption isotherms and thermodynamics

The experimental data were also fitted using the Langmuir and Freundlich isotherm models to further elucidate the Pb (II) adsorption mechanisms of MBC, HBC, and BC. The Langmuir isotherm model describes the monolayer adsorption of pollutants onto a homogeneous adsorbent surface,⁵² while the Freundlich isotherm model generally describes multilayer adsorption on non-uniform absorbents.⁵³ The fitting outcomes are displayed in Figures 5J and 5K and Table 3. Both the Langmuir and Freundlich models exhibited good fitting results. The Langmuir model consistently outperformed the Freundlich model in terms of R² values for all types of biochar, suggesting that it provides a more accurate representation of how Pb (II) is removed from groundwater with biochar. This finding underscored the superior explanatory power of the Langmuir model in this context, suggesting that the removal process followed a mechanism of homogeneous monolayer chemisorption. Based on the Langmuir fitting results, the Pb (II) adsorption capacities were found to peak at 633.531 mg g⁻¹ for MBC, 465.144 mg g⁻¹ for HBC, and 396.057 mg g⁻¹ for BC. The Freundlich model n⁻¹ fitted data revealed that the n⁻¹ values of all three adsorbents exceeded 1, suggesting a favorable adsorption process.⁵⁴

Table 3.

Adsorption isothermal fitting parameters

Model	Parameters	Fitting results
		BC	HBC	MBC
Langmuir	Qm/(mg·g−1)	396.057	465.144	633.531
	KL(L·mg−1)	0.002	0.002	0.002
	R2	0.987	0.979	0.991
Freundlich	KF	1.57	2.896	3.070
	n	1.274	1.289	1.355
	R2	0.975	0.965	0.980

Notably, for all types of biochar, Pb (II) adsorption increased with an increase in the initial concentration, eventually stabilizing. The highest P b(II) uptake values for MBC, HBC, and BC were 253.6 mg g⁻¹, 191.2 mg g⁻¹, and 167.9 mg g⁻¹, respectively. In comparison with previous results from the relevant literature (shown in Table 4), the Pb (II) adsorption capacity of the biochar developed in the present study appeared to be superior. Furthermore, our biochar was low-cost, showed good recyclability, and could effectively avoid secondary pollution.

Table 4.

Adsorption capacity of various adsorbents reported in the literature

Adsorbents	Optimal pH	Qm (mg·g−1)	Optimal dosage (g·L−1)	Reference
Raw corn stalk	7.0	0.029	5.00	Šćiban et al.55
Ash+Fe nanoparticles	6.0	30.00	0.250	Ghasemi et al.56
Oxidized multi-walled carbon nanotubes	6.0	124.02	0.625	Zhang et al.57
Titanium isopropoxide + Zirconia chloride	5.0	175.00	1.000	Chen et al.58
Sugarcane bagasse	6.0	5.82	0.010	Phaenark et al.27
Clay	6.0	172.00	0.040	Teğin et al.59
Calcium alginate-nZVI-biochar	5.0–6.0	247.99	2.500	Zhao et al.60
Fe3O4 +MNPs	6.0	64.50	0.100	Kothavale et al.61
Paulownia leaves	7.35	110.90	0.002	Koprivica et al.62

Finally, the thermodynamic characteristics of Pb (II) adsorption by BC, HBC, and MBC were investigated at temperatures of 303 K, 313 K, and 323 K (Table 5). The Gibbs free energy (ΔG) values were negative, indicating that the adsorption process was spontaneous. As the temperature increased, the ΔG values gradually decreased, indicating that the spontaneity of the process rose with increasing temperature. The enthalpy change (ΔH) was greater than 0, demonstrating that the reaction was endothermic, and higher temperatures facilitated the adsorption process.⁶³ The entropy change (ΔS) was greater than 0, reflecting an increase in randomness at the solid–liquid interface.

Table 5.

Thermodynamic parameters of adsorption of heavy metals byBC, HBC and MBC

Biochar	T (K)	ΔG (kJ·mol−1)	ΔH (kJ·mol−1)	ΔS (J·mol−1·K−1)
BC	303	−5.03	13.33	438.53
	313	−7.54
	323	−13.88
HBC	303	−2.04	1.19	39.49
	313	−2.43
	323	−2.83
MBC	303	−0.72	1.24	37.40
	313	−1.16
	323	−1.46

Adsorption mechanism of MBC

The FTIR, XPS, and XRD characterization results were further employed to investigate the mechanism of Pb (II) adsorption by MBC. The XRD pattern after Pb (II) adsorption, shown in Figure 6A, revealed diffraction peaks at 2θ = 23.3°, 25.3°, 37.2°, and 43.5°, which are characteristic of PbO. This indicated that the mineral elements on the surface of MBC could fix Pb (II) through precipitation as PbO.⁶⁴

Figure 6.

Characterization of MBC After Pb(II) Adsorption

XRD pattern after MBC adsorption (A). FTIR spectra before and after Pb(Ⅱ) adsorption by MBC (B). Gross XPS spectra of the reacted MBC (C). C1s (D), O1s (E), Mg1s (F), Fe2p (G), and Pb4f (H) splits.

As shown in Figure 6B, the changes in the functional groups of MBC after the adsorption of Pb (II) were comparatively analyzed using FTIR. In the unadsorbed state, MBC showed a broad absorption peak corresponding to the –OH stretching vibration at 3240 cm⁻¹, indicating the presence of abundant hydroxyl groups on the material’s surface.⁶⁵ Meanwhile, the characteristic peak at 1774 cm⁻¹ could be attributed to the carbonyl (–C=O) stretching vibration,⁶⁰ while those at 1438 cm⁻¹ and 1088 cm⁻¹ corresponded to the bending vibrational mode of methylene (C―H).^66,67 After Pb (II) adsorption, the intensities of the –OH and C―H vibrational bands decreased significantly, and the intensity of the –C=O peak increased. These data suggested that –OH and C―H are involved in the chemisorption of Pb (II) on MBC. Further, the enhancement effect provided by carbonyl functional groups may involve structural reorganization after metal coordination.

The XPS spectra depicted in Figures 6C–6H demonstrated notable shifts in the C1s region following the adsorption of Pb (II). Specifically, the abundance of C―O―C, O―C=O, and C―C groups dropped from 13.26%, 3.79%, and 41.39%–12.22%, 3.4%, and 40.5%, respectively. Meanwhile, in the O1s region, the abundance of O―C=O, Mg―O, and Fe―O declined from 21.5%, 3.65%, and 1.30%–3.76%, 1.96%, and 1.22%, whereas the C―O content surged from 8.81% to 35.81%. These alterations highlighted the involvement of C―O―C, O―C=O, and C―C groups in the complexation of Pb (II), with both Mg and Fe playing a role in the ion exchange process.⁶⁸ After Pb (II) adsorption, the Pb4f spectrum of the biosorbent became deconvoluted into two distinct peaks with binding energies of 139.1 eV and 143.9–144.1 eV. The 139.1 eV peak reflected the interaction of Pb (II) with electron-dense areas of the biosorbent, highlighting the role of physical adsorption and electrostatic forces in the adsorption process. Meanwhile, the peak at 143.9–144.1 eV reflected the binding of Pb (II) to functional groups, implying that the generation of Pb (II) complexes was part of the adsorption mechanism.⁶⁹

Based on the aforementioned analysis and considering the adsorption kinetics models, we inferred that MBC adsorbs Pb (II) through various processes, including physical adsorption, electrostatic Attraction, hydrogen bonding, chemical adsorption, surface complexation, and ion exchange.

Conclusion

In this study, highly efficient adsorbent (MBC) was successfully synthesized for removing Pb (II) from aqueous solutions. Our experimental findings demonstrated that the surface of MBC possessed a coarse, porous, and uneven texture, with magnesium and iron oxides anchored onto the corn stover-derived biochar. The Pb (II) adsorption capacity peaked at an impressive 253.6 mg g⁻¹. Kinetic analysis showed that the adsorption process adhered to pseudo-second-order kinetics, involved intraparticle diffusion, and was consistent with the Langmuir model. This indicated that the dominant mechanism of adsorption was chemical adsorption occurring on a uniform monolayer. Furthermore, regeneration tests demonstrated that MBC is easily separable and reusable, maintaining its effectiveness even after multiple cycles of reuse. We also discovered that Pb (II) adsorption by MBC occurs via a combination of physical adsorption, hydrogen bonding, chemical adsorption, surface complexation, and ion exchange. Overall, MBC exhibits a high adsorption capacity and excellent regeneration potential, making it a promising, cost-effective, and environmentally friendly adsorbent for Pb (II) removal during water treatment.

Limitations of the study

This study focused solely on the adsorption performance of Pb (II) and did not verify the simultaneous removal effect of the material on other heavy metals, nor did it examine the competitive adsorption mechanism when multiple heavy metals coexist.

Resource availability

Lead contact

Further information and requests for resources and reagents should be directed to and will be fulfilled by the lead contact, Jiaqi Pang(pangjiaqi@emails.imau.edu.cn).

Materials availability

This study did not generate new unique reagents.

Data and code availability

• •All data reported in this paper will be shared by the lead contact upon request.
• •This paper does not report original code.
• •Any additional information necessary to reanalyze the data reported in this article is available upon request from the primary contact, Ms. Jiaqi Pang (pangjiaqi@emails.imau.edu.cn).

Acknowledgments

This work was supported by the National Key R&D Program (2023YFC3206504); Inner Mongolia Key R&D and Achievement Transformation Plan Project (2023YFDZ0022); National Natural Science Foundation of China (52369014, 52060022, and U244320048).

Author contributions

Writing – original draft, J.Q.P. and B.S.; conceptualization, J.Q.P., B.S., and A.J.S.; formal analysis, J.Q.P., A.J.S., Y.L.Z., and M.C.; investigation, Y.Y.G., Y.L.Z., and X.Z.G.; project administration, B.S. and X.H.S.

Declaration of interests

The authors declare no competing interests.

STAR★Methods

Key resources table

REAGENT or RESOURCE	SOURCE	IDENTIFIER
Software and algorithms
Origin2024	Origin Lab	https://www.originlab.com
Other
Fe3O4	Macklin	1317-61-9; Cat#I811693
anhydrous ethanol	Sinopharm Chemical Reagent Co., Ltd	64-17-5; Cat#10009218
Citric acid	Macklin	5949-29-1; Cat#C805019
NaOH	Sinopharm Chemical Reagent Co., Ltd.	1310-73-2; Cat#10019718
HCl	Sinopharm Group Chemical Reagent Co., Ltd	7647-01-0; Cat#10011018
PbCl2	Shanghai Rinne Science and Technology Development Co., Ltd.	7758-95-4; Cat#R005529
MgCO3	Shanghai Rinne Science and Technology Development Co., Ltd.	39409-82-0; Cat#R019676
Atomic absorption spectrophotometer	Thermo Scientific	https://www.thermofisher.cn
X-ray photoelectron spectroscopy	Shimadzu Corporation	https://www.shimadzu.com.cn
Scanning electron microscopy	COXEM	https://www.coxemchina.com
Surface area analyzer	JWGB	https://www.jwgb.net
X-ray powder diffraction	Malvern Panalytical	https://www.malvernpanalytical.com.cn
Vibrating sample magnetometer	DEXIN MAG	https://www.ydxcd.com
Fourier transform infrared spectroscopy	Thermo Scientific	https://www.thermofisher.cn
Thermal Analysis System	METTLER TOLEDO	https://www.mt.com

Method details

The material synthesis methods, experimental procedures, and calculation methods are described in detail in the experimental procedures section. SEM, XRD, XPS, VSM, BET, and TGA are described in detail in the characterization of biochar.

Quantification and statistical analysis

All statistical analyses and software used in this section were primarily performed using Origin. No methods were used in this section to determine whether the data met the assumptions of the statistical methods.

Published: August 5, 2025