Enhanced Adsorption of Aqueous Pb(II) by Acidic Group-Modified Biochar Derived from Peanut Shells

Abstract

Using peanut shells, a sustainable agricultural waste product, as its raw material, the acid group-modified biochar (AMBC) was prepared through phosphoric acid activation, partial carbonization, and concentrated sulfuric acid sulfonation for efficient removal of lead ion from aqueous solutions. Characterization techniques such as N₂ isothermal adsorption–desorption, SEM, XRD, FT-IR, TG-DTA, and acid–base titration were utilized to fully understand the properties of the AMBC. It was found that there were high densities of acidic oxygen-containing functional groups (-SO₃H, -COOH, Ph-OH) on the surface of the AMBC. The optimal adsorption performance of the AMBC for Pb(II) in water occurred when the initial concentration of Pb(II) was 100 mg/L, the pH was 5, the dosage of the adsorbent was 0.5 g/L, and the contact time was 120 min. Under the optimal conditions, the removal ratio of Pb(II) was 76.0%, with an adsorption capacity of 148.6 mg/g. This performance far surpassed that of its activated carbon precursor, which achieved a removal ratio of 39.7% and an adsorption capacity of 83.1 mg/g. The superior adsorption performance of AMBC can be caused by the high content of acidic oxygen-containing functional groups on its surface. These functional groups facilitate the strong binding between AMBC and Pb(II), enabling effective removal from water solutions.

1. Introduction

With improvements in social industrialization, heavy metal pollution incidents have occurred frequently in recent years. Given their high toxicity, carcinogenicity, and non-biodegradability, countries have implemented strict limits on the discharge of heavy metals in water [1]. Lead, a common heavy metal ion contaminant, is typically found in water as a divalent ion. Pb(II) is non-degradable and can persist in the environment for extended periods, causing significant damage to the human cardiovascular system, liver, and the natural environment [2]. The primary sources of lead-containing wastewater discharge are the electroplating industry, textile industry, and metal processing industries [3].

The treatment of lead-containing wastewater has always been a challenge in the field of sewage treatment. Many techniques have been employed for the treatment of Pb(II) from aqueous environments, including chemical precipitation [4], adsorption [5], membrane filtration [6], ion exchange [7], bioremediation [8], and the electrochemical method [9]. However, these methods exhibit drawbacks in the form of high costs and low efficiencies [10]. Adsorption has emerged as a promising approach in these methods due to its simplicity, high efficiency, and minimal secondary pollution [11]. Adsorbents should be highly selective and efficient, easy to process and apply, should not induce the generation of secondary chemicals, and should have a low cost [12]. Biochar, a low-cost adsorbent derived from biological waste, has garnered significant attention as a sustainable and environmentally friendly adsorbent [5]. It offers an adjustable surface area and controllable porous structure [13], making it highly effective for Pb(II) removal. Carbon nanotubes [14] and clay [15] are also used as adsorbents, but biochar’s renewability and cost-effectiveness make it particularly attractive for wastewater treatment applications [16].

There are many kinds of oxygen-containing functional groups on the surface of biochar, primarily in the form of -CHO, -OH, -COOH, and -C=O, which are usually the active centers of adsorption on activated carbon [17]. To further improve its adsorption performance, biochar can be modified to increase the number and diversity of surface functional groups. Momcilovic et al. [18] modified biochar with phosphoric acid, resulting in a surface containing not only -COOH and -OH, but also P-O-P, P=O, and P=OOH groups. This modification promoted the adsorption capacity for Pb(II). Similarly, Ahmed et al. [5] prepared H₂O₂-modified watermelon seed biochar with more oxygen-containing functional groups. After activation, the adsorption capacity for Pb(II) increased significantly from 44.32 mg/g to 60.87 mg/g.

Carbon-based solid acid is a kind of amorphous carbon material [19] with acidic groups modified on its surface. It can be prepared from polycyclic aromatic hydrocarbons [20], carbohydrates [21], biomass [22], and mesoporous carbon materials [23]. Carbon-based solid acid has acidic groups such as sulfonic acid and phosphoric acid on the surface and has the advantages of a high acid density, simple post-treatment, easy reusability, low cost, and ecological friendliness. It is commonly used in the field of catalytic research [19,24,25] and serves as a strong acid material that can replace liquid protonic acids. Among these carbon-based solid acids, those derived from biomass sources such as potato peels [22], coconut shells [26], and wakame [25] are essentially an acidic group modified biochar. These acidic groups, such as -SO₃H, are achieved through a carbonization and sulfonation process. Yu et al. [27] prepared carbon-based solid acid containing -OH, -COOH, and -SO₃H groups via a one-step solvothermal method using axonopus compressus as the raw material. This material demonstrated an exceptional Pb(II) adsorption capacity, achieving 191.07 mg/g within just 5 min. Although the application of carbon-based solid acids for adsorbing heavy metal ions in water is still in its early stages, it holds significant promise for Pb(II) sewage treatment. This is due to its high density of acidic oxygen-containing functional groups on its surface [22], which is expected to improve its adsorption capabilities compared to biochar alone.

Peanut shells, a common form of agricultural waste, can be used to prepare biochar [28]. In this study, an acidic group-modified biochar (AMBC) was prepared from peanut shells through phosphoric acid activation, partial carbonation, and sulfonation with concentrated sulfuric acid. This material was then used for Pb(II) adsorption in water. The effects of the contact time, pH, dosage of adsorbent, and initial Pb(II) concentration on the adsorption performance were studied. Additionally, the thermodynamics and kinetics of the AMBC adsorption for Pb(II) were investigated. Lastly, the adsorption performance of the AMBC was compared with that of its activated carbon precursor, and the adsorption mechanism was discussed.

2. Materials and Methods

2.1. Materials

Analytical grade (AR) chemicals, including H₂SO₄ (98%, Macklin Biochemical Technology Co., Ltd., Shanghai, China), NaCl (Macklin Biochemical Technology Co., Ltd., Shanghai, China), NaHCO₃ (Fuchen Chemical Reagent Co., Ltd., Tianjin, China), HCl (KMO Reagent Co., Ltd., Tianjin, China), Pb(NO₃)₂ (Sinopharm Chemical Reagent Co., Ltd., Shanghai, China), HNO₃ (65%, Macklin Biochemical Technology Co., Ltd., Shanghai, China), NaOH (Wind Ship Chemical Reagent Co., Ltd., Shanghai, China), and H₃PO₄ (Sinopharm Chemical Reagent Co., Ltd., Shanghai, China), were used in this work without further purification. The peanut shells were obtained from peanuts purchased from a market in Tianjin, China.

2.2. AMBC Preparation

The peanut shells were thoroughly washed and dried at 80 °C, then crushed with a crusher and sieved to obtain 80–100 mesh powder. The peanut shells were then macerated in a 1.22 g/mL H₃PO₄ solution for 16 h, following a mass ratio of 3:1, and then dried at 110 °C. The resulting solid was carbonized at 300 °C for 2 h under a N₂ flow. After this, the sample was washed with 0.1 mol/L HCl solution at 100 °C for 1 h, cooled and filtered, and then cleaned repeatedly with deionized water (≥80 °C) until free of Cl⁻. Finally, the pre-AMBC was obtained after drying at 80 °C.

To sulfonate the pre-AMBC, 1 g of pre-AMBC was mixed with 25 mL of H₂SO₄ in a flask. The mixture was then heated at 160 °C for 10 h under a N₂ flow. After the sulfonation process, the carbon material was cooled to room temperature, then filtered and washed with deionized water (≥80 °C) until no SO₄²⁻ was present. Finally, the AMBC was obtained after drying at 80 °C.

2.3. Acid–Base Titration

Acid–base neutralization titration was used to determine the density of -SO₃H groups on the surface of AMBC [29]. In this process, 0.25 g of AMBC and 30 mL of NaCl solution (0.05 mol/L) were ultrasonically shaken at room temperature for 1 h to ensure that the H⁺ ions in the -SO₃H on the adsorbent surface were fully exchanged by Na⁺. The solution was then filtered to obtain the filtrate. Finally, phenolphthalein was added as an indicator, and the filtrate was titrated with a 0.05 mol/L NaOH standard solution until the filtrate changed from colorless to a slightly reddish hue. According to the dosage of NaOH, the density of the -SO₃H groups in the adsorbent was calculated using Equation (1):

$$F=\frac{cV}{m},$$

(1)

where F is the density of -SO₃H, mmol/g; c is the concentration of NaOH standard solution, mol/L; V is the volume of NaOH solution consumed during titration, mL; and m is the mass of absorbent weighed for the acid–base titration, g.

The density of -COOH and -PhOH groups on the adsorbent surface was determined using Boehm’s method [30]. Firstly, 0.5 g of adsorbent was transferred into two 150 mL conical flasks. Then, NaHCO₃ and NaOH solutions (25 mL, 0.05 mol/L) were added into each flask, respectively. The samples were ultrasonically shaken for 60 min and left to stand for 24 h at room temperature before being filtered. Subsequently, HCl solution (50 mL, 0.05 mol/L) was added to the filtrate containing NaHCO₃. In order to remove the CO₂ produced by the reaction of HCl and NaHCO₃, the conical flasks were heated for 30 min in a water bath at 80 °C. Phenolphthalein was then added dropwise to the flask, and the mixture was titrated with a 0.05 mol/L NaOH standard solution until the solution turned slightly red, indicating the endpoint of the titration. The adsorbent-based NaOH or NaHCO₃ consumption was calculated using Equation (2):

$$a=\frac{c_{NaOH}V+25c_0-50C_{HCl}}{m},$$

(2)

where a is the NaOH or NaHCO₃ consumption, mmol/g; m is the quality of adsorbent, g; c_NaOH is the concentration of NaOH standard solution, mol/L; V is the volume of NaOH standard solution used in the titration process, mL; 25 is the volume of NaOH standard solution used, mL; c₀ is the concentration of NaOH standard solution used, mol/L; 50 is the volume of excess HCl added, mL; and c_HCl is the concentration of HCl solution used, mol/L.

The density of the sulfonic acid groups (F-_SO3H) and the consumption of NaHCO₃ can be calculated using Equations (1) and (2), respectively. The value of (a_NaHCO3-F_SO3H) represents the density of carboxylic acid groups (-COOH) on the adsorbent surface, expressed in units of mmol/g.

The consumption of NaOH and NaHCO₃ can be determined using Equation (2). The value of (a_NaOH-a_NaHCO3) represents the density of phenolic hydroxyl groups (-PhOH) on the surface of the adsorbent, expressed in units of mmol/g.

2.4. Characterization

The surface area and porosity of the adsorbents were measured using a N₂ adsorption analyzer (Micromeritics ASAP 2020 PLUS, Norcross, GA, USA). The morphology of the adsorbents was characterized using a field emission scanning electron microscope (SEM, FEINova Nano 450, Waltham, MA, USA). X-ray diffractometer spectroscopy (XRD, Rigaku D/Max-2500, Tokyo, Japan) was employed to identify the crystal phase of the AMBC. Various chemical properties of the organic functional groups of the adsorbents were analyzed using an FT-IR instrument (Thermo-Nicolet Nexus 470, Waltham, MA, USA). The thermostability of the AMBC was investigated using a thermogravimetric analyzer (TGA, Rigaku PTC-10A Instrument, Tokyo, Japan). The chemical state analysis of the adsorbents before and after adsorption was measured using X-ray photoelectron spectroscopy (XPS, Thermo Scientific 250Xi, Waltham, MA, USA). In addition, an inductively coupled plasma optical emission spectrometer (ICP-OES, Perkin Elmer Optima 7300v, Waltham, MA, USA) was used to determine the concentrations of Pb(II).

2.5. Adsorption Test

The simulated lead wastewater (100 mg/L) was obtained by dissolving Pb(NO₃)₂ in deionized water and diluting it to a specified concentration. Before the adsorption experiments, the pH of the solvent was adjusted using NaOH (0.1 mol/L) and HNO₃ (0.1 mol/L).

Adsorption experiments for Pb(II) using the AMBC and pre-AMBC were carried out using a shaker. A total of 0.05 g of adsorbent and 100 mL of lead-containing simulated wastewater were sequentially added to a conical flask and shaken at a constant temperature with a stirring rate of 200 rpm. The effects of the contact time (0–200 min), solution pH (3–6), dosage of adsorbent (0–2 g/L), and initial Pb(II) concentration (50–200 mg/L) on the adsorption performance were investigated. After the adsorption experiment, the solution was filtered, and the concentration of Pb(II) in the filtrate was determined using an ICP.

The efficiency of adsorbents is typically assessed based on the removal ratio and adsorption capacity. These parameters can be calculated using Equations (3) and (4), respectively:

$$R=\frac{(c_0-c_1)}{c_0}\times 100\%,$$

(3)

$$Q=\frac{{(c}_0-c_1)\times V}{m},$$

(4)

where R is the removal ratio, %; Q is the adsorption capacity at the time t, mg/g; c₀ and c₁ are the concentrations of Pb(II) in the solution before and after adsorption, respectively, mg/L; m is the adsorbent dosage, g; and V is the volume of solution, L.

3. Results and Discussion

3.1. Characterization of Adsorbents

The density of acidic functional groups on the AMBC and pre-AMBC was determined through acid–base neutralization titration. The total acid quantity on the AMBC was 2.28 mmol/g, which was higher than that of the pre-AMBC (1.26 mmol/g). It is worth noting that the density of acidic oxygen-containing groups increased to varying degrees after sulfonation. AMBC contained 0.52 mmol/g -SO₃H groups, indicating that that concentrated H₂SO₄ serves as an effective sulfonating agent, successfully introducing -SO₃H groups onto the surface of the adsorbent.

The surface areas of AMBC and pre-AMBC were analyzed, and the results are shown in

Table 1. Pore structure and surface acidic functional group density of adsorbents.

Adsorbent	SBET a (m2/g)	VP b (cm3/g)	DP c (nm)	Acid Density (mmol/g)
				Total	Ph-OH	-SO3H	-COOH
Pre-AMBC	832.6	3.99 × 10−1	2.2	1.26	0.05	-	1.21
AMBC	329.3	2.67 × 10−2	2.1	2.28	0.33	0.52	1.43

Notes: ^a S_BET was measured using the multi-point BET method. ^b Average pore size was calculated based on the desorption branch of the isotherm using the BJH method. ^c Total pore volume was measured at P/P₀ = 0.99.

. The surface area of AMBC, prepared via phosphoric acid activation, was found to be 329.3 m²/g. H₃PO₄ serves as an effective activator, facilitating the transformation of the inorganic matter in peanut shells into phosphate. This transformation results in an expansion and increase in the distance between carbon microcrystals. During the washing process, the removal of phosphate leads to more developed pores in the activated carbon materials, ultimately increasing the specific surface area. H₃PO₄ pretreatment has been reported to produce biochar with a large surface area and abundant porous structures [31]. Compared to the carbon precursors pre-AMBC, the specific surface area and pore volume of AMBC decreased significantly after sulfonation. This may be due to the structural shrinkage caused by strong acid treatment [32].

The surface morphology of the pre-AMBC and AMBC was characterized using a scanning electron microscope, as shown in Figure 1, revealing that they exhibited lamellar structures composed of polyaromatic carbon layers. Compared with pre-AMBC, the lamella of AMBC was smaller and the surface was rougher, which may have led to the increase in the specific surface area and porosity of the AMBC.

Figure 2 shows the XRD spectrum of pre-AMBC and AMBC, which exhibits a broad diffraction peak between 20° and 30°. The peak correspond to the diffraction of C(002) [33] indicates that the carbon-based solid acids consisted of aromatic carbon sheets. Therefore, it can be inferred that the prepared carbon material was an amorphous carbon adsorbent [34]. In addition, the XRD spectrum of pre-AMBC showed a weak diffraction peak between 40° and 50°, which can be attributed to C(004). Xue et al. [24] showed that the latter peak was due to the α axis of the graphite structure. The peak could hardly be observed in the XRD spectrum of the AMBC, indicating that the graphitization degree of the AMBC was lower and the carbon layer size was smaller, which is consistent with the results of the SEM.

The functional groups on the AMBC and pre-AMBC were characterized based on the FT-IR spectra, and the results are shown in Figure 3. Both spectra showed a peak around 1580 cm⁻¹, which was attributed to the stretching of the C=C bonds formed during the carbonization of the carbon-based solid acid. This suggests the formation of polycyclic aromatic carbon due to incomplete carbonization of the peanut shells [35,36]. The peaks near 1703 cm⁻¹ and 2923 cm^–1 can be devoted to the C=O and O-H stretching modes of the -COOH groups, respectively [24]. It illustrated that -COOH groups were present on the surfaces of both the pre-AMBC and AMBC. The peak at 3422 cm⁻¹ may be ascribed to the result of the combination of the stretching modes of the Ph-OH group [37] and the symmetric and asymmetric stretching vibration of the adsorbing -OH group in the water molecule. Compared to the pre-AMBC, the AMBC exhibited two additional peaks at 1168 cm⁻¹ and 1029 cm^–1, which were attributed to the O=S=O symmetric and asymmetric stretching modes, respectively [13]. This result suggests that sulfonic acid groups are loaded onto the surface of AMBC through the sulfonation process. Moreover, the peak at 1076 cm⁻¹ in the pre-AMBC spectrum could be attributed to the ionized linkage of P⁺-O⁻ in the acid phosphate esters or P-O-P symmetry vibrations in the polyphosphate chains [28]. However, this peak did not appear in the AMBC spectrum, indicating that acid phosphate esters and phosphoric acid are removed during the sulfonation process.

The thermal stability of the pre-AMBC and AMBC was investigated using TG-DTA, and the results are shown in Figure 4. During the heating process from 28 °C to 700 °C, the total weight loss of AMBC was 49.8%. Specifically, AMBC lost 6.0% of its weight from 28 °C to 100 °C, and the corresponding DTA temperature was 70.7 °C. This weight loss was primarily due to the removal of physically adsorbed water in the solid acid. Water molecules may be physically adsorbed or bound to the surface of AMBC adsorbents through hydrogen bonds. A slow weight loss occurred between 100 °C and 227 °C, and the mass change was not obvious. As the temperature rose above 227 °C, the thermogravimetric curve began to decrease. The 4.4% weight loss from 227 °C to 330 °C was attributed to the thermal decomposition of -SO₃H groups as well as the condensation and structural reforming of carbon-based materials [38]. The DTA curve showed a decreasing trend after 328.5 °C, indicating that the sample was undergoing endothermic processes as well. Subsequently, beyond 330 °C, there was a notable weight loss of 36.6%, which was due to the continuous carbonization of the AMBC at high temperatures.

Based on the TG-DTA curve of the pre-AMBC in Figure 4a, it can be seen that the trend of the TG curve of the pre-AMBC is consistent with that of the AMBC. Throughout the entire temperature range (28–700 °C), the weight loss rate of the adsorbent reached 54.0%. The corresponding DTA temperature was 70.4 °C, which was consistent with the endothermic peak of AMBC, indicating that it also contributed to the removal of physically adsorbed water from the biochar.

3.2. Pb(II) Adsorption by AMBC

3.2.1. Effects of Contact Time

The effects of contact time between the adsorbent and Pb(II) on the adsorption performance were investigated, and the results are shown in Figure 5.

As the time increased, both the removal ratio and adsorption capacity gradually increased. A significant growth in the adsorption capacity was observed when the contact time was extended from 10 min to 90 min, with the removal ratio rapidly increasing from 40.8% to 57.9%. This increase can be due to the high concentration of Pb(II)-containing solution and the presence of a large number of active sites on the surface of the solid acid at the initial stage of adsorption [39]. Subsequently, as the amount of adsorbed Pb(II) accumulates, most of the adsorption sites become occupied, leaving fewer unoccupied sites [40].

Moreover, the adsorbed Pb(II) repels the free Pb(II) in the solution via electrostatic repulsion, leading to a gradual decrease in the adsorption rate until an equilibrium is reached. At a contact time of 120 min, the removal ratio was 59.4%. Prolonging the adsorption time beyond this point did not significantly alter the removal ratio. At this point, the rate of adsorption and desorption were the same, indicating that the adsorption had reached a dynamic equilibrium [41]. In summary, during the adsorption process, both the removal ratio and adsorption capacity increased rapidly with contact time and then gradually reached an equilibrium. The adsorption capacity reached a maximum after 120 min (114.4 mg/g) and remained essentially constant thereafter.

3.2.2. Effects of pH Value

The adsorption of Pb(II) in solution by the adsorbents is strongly influenced by the pH value, as it determines the chemistry properties of the adsorbent surface and the types of metal ions. In general, the adsorption of cations is favored at pH > pH_PZC [42]. If the pH value is lower than pH_PZC, the zeta potential of the biochar expresses a net positive charge, which is responsible for the electrostatic repulsion of Pb(II) [43]. As the pH increases, this repulsive force becomes less and less intense. When the solution pH exceeds 6, Pb(II) gradually forms a Pb(OH)₂ precipitate [44], which is unfavorable for adsorption. Therefore, considering the influence of PH_PZC and avoiding precipitation, the pH was adjusted within the range of 3–6 for this experiment, and the experimental results are shown in Figure 6.

When the pH value was 3, the removal ratio of AMBC to Pb(II) was fairly low, at 59.4%. As the pH value increased, the removal ratio gradually increased. When the pH value reached 5, the adsorption capacity and removal rate reached the maximum values of 148.6 mg/g and 76.0%, respectively. However, as the pH continued to increase to 6, the adsorption performance decreased slightly.

In the lower pH range, the removal ratio and adsorption capacity are limited. This is due to two factors: competitive adsorption of active sites by H⁺ and Pb(II) [43] and the presence of a high concentration of H⁺ inhibiting the dissociation of weakly acidic groups such as -COOH, thus reducing the availability of active sites on the AMBC surface [27]. With the increase in pH value, the competition between Pb(II) and H⁺ decreases, while the number of active sites on the AMBC surface increases, leading to a gradual increase in the removal ratio until it reaches an adsorption equilibrium.

3.2.3. Effects of the Dosage of AMBC

In the adsorption process, the adsorbent dosage is a significant factor that affects the adsorption performance. Additionally, it is also a crucial indicator of the economic cost associated with the adsorbent. The effects of AMBC dosage on its adsorption of Pb(II) are shown in Figure 7.

The removal ratio of Pb(II) increased with higher AMBC dosages. With the dosage of AMBC increased from 0.1 g/L to 0.5 g/L, the removal ratio of Pb(II) rapidly increased from 25.5% to 76.0%. As the dosage of AMBC reached 0.9 g/L, the removal ratio plateaued at 94.9%, followed by a slight increase as the AMBCA dosage further increased. At lower AMBC dosages, there were fewer active adsorption sites on its surface, leading to lower removal ratios. With increasing dosages, the number of adsorption active sites increased, and the removal ratio gradually increased until it reached saturation.

The adsorption capacity of AMBC decreased as its dosage increased. When the dosage was small, the relative amount of Pb(II) per unit mass of AMBC was higher, leading to a large mass transfer driving force. However, when the amount of AMBC was excessive, a significant number of adsorption sites remained vacant, resulting in unnecessary waste. In addition, higher dosages could cause agglomeration of the AMBC, reducing the total adsorption surface area and leading to a lower adsorption capacity [45,46].

3.2.4. Effects of Initial Pb(II) Concentration

Figure 8 displays the effects of the initial concentration of Pb(II) on the adsorption performance of the AMBC.

In the entire concentration range studied, as the initial concentration increased, the adsorption capacity of AMBC exhibited almost linear growth. The adsorption capacity of AMBC increased rapidly from 100.8 mg/g to 195.6 mg/g as the initial concentration of Pb(II) was increased from 50 mg/L to 200 mg/L. However, the removal ratio of Pb(II) decreased as the initial Pb(II) concentration increased. This can be attributed to the enhanced migration of ions to the AMBC surface due to the increased concentration, allowing Pb(II) to more fully diffuse into the vicinity of active sites and thereby increasing the adsorption capacity. However, as the active sites on the surface of AMBC became gradually occupied, the excess Pb(II) in the solution could not be adsorbed and remained free. Due to the limited adsorption sites on the surface of the adsorbent and electrostatic repulsion, the removal ratio of the AMBC decreased significantly with increasing concentrations [43].

3.3. Comparison of Pre-AMBC and AMBC with Other Adsorbents

The adsorption experiments were carried out using AMBC and pre-AMBC as the adsorbents, and the experimental data are shown in

Table 2. Comparison of the adsorption of Pb(II) by AMBC and its precursor.

Adsorbents	Q (mg/g)	R (%)
Pre-AMBC	83.1	39.7
AMBC	148.6	76.0

Note: Conditions: initial concentration of Pb(II) = 100 mg/L, pH = 5, dosage of AMBC = 0.5 g/L, contact time = 120 min, shaking rate = 200 rpm.

. Under the optimal adsorption conditions, the Pb(II) removal ratio and adsorption capacity of the pre-AMBC reached 39.7% and 83.1 mg/g, respectively, which were much lower than the values for AMBC. After being modified by concentrated H₂SO₄, the adsorption capacity of AMBC was greatly increased.

The results of the acid–base titration showed that there were -SO₃H groups (0.52 mmol/g) on the surface of the AMBC, and the density of -COOH and -OH was higher than that of the pre-AMBC. The presence of acidic oxygen-containing functional groups enhanced the affinity of the adsorbent for metals and the ion-exchange capacity [27]. These groups resulted in very low zero-point charge values, greatly increasing the surface charge, which improved the accessibility of Pb(II) to enter the internal pores [28].

Table 3. Comparisons of the removal efficiencies of Pb(II) by different adsorbents.

Adsorbents	Qmax (mg/g)	Activator	Experiment Conditions
			pH	T (°C)	Dosage (g/L)	Equilibration Time (min)	Initial Concentration (mg/g)	Ref.
AMBC	148.6	H2SO4	5	25	0.5	120	100	This study
C-KOH	57.5	KOH	5	25	2	35	100	[47]
SBC	191.1	H2SO4	4.5	-	2	5	200	[27]
HP-BC	60.9	H2O2	5	25	-	500	100	[5]
M-RH-AC	134.9	HNO3	5.5	-	1	90	180	[43]
PLAC	98.4	H3PO4	5.0	25	0.6	30	80	[48]

Notes: C-KOH, KOH-modified biochar; SBC, sulfonated axonopus compressus biochar; HP-BC, H₂O₂-modified watermelon seed biochar; M-RH-AC, nitric acid-modified rice husk biochar; PLAC, H₃PO₄-modified polygonum orientale linn biochar.

shows a comparison of the adsorption performance of AMBC with various modified biochars for Pb(II). The results indicate that AMBC exhibits a superior adsorption efficiency at lower adsorbent dosages, making it a hopeful candidate to remove Pb(II).

3.4. Pb(II) Adsorption Isotherms

Adsorption experiments were carried out by mixing AMBC with varying concentrations of Pb(II) solution (50–250 mg/L) at different temperatures (25–45 °C) and shaking the mixture on a shaker at 130 rpm for 180 min. Figure 9 shows the isothermal adsorption line of Pb(II) adsorption by the AMBC.

As the temperature rose, the adsorption of Pb(II) on the adsorbent also increased, indicating that the adsorption process was an endothermic reaction [49]. Then, Langmuir and Freundlich isothermal models were used to fit the experimental data, and the results are shown in Figure 10 and Figure 11. The adsorption modelling parameters are listed in Table 3.

The linearization equations of the two adsorption models are as shown in Equations (5) and (6):

Langmuir model:

$$\frac{C_e}{q_e}=\frac{C_e}{q_m}+\frac{1}{{K_{L}q}_m},$$

(5)

Freundlich model:

$$ln{q}_e=\frac{1}{n}ln{C}_e+ln{K}_F,$$

(6)

where Ce is the Pb(II) equilibrium concentration, mg/L; q_e is the adsorption capacity at the adsorption equilibrium of AMBC, mg/g; q_m is the maximum adsorption capacity of AMBC, mg/g; K_L and K_F are the constants of the Langmuir and Freundlich isotherms respectively, L/mg; and n is an empirical constant that indicates the adsorption intensity. It varies depending on the heterogeneity of the materials used as adsorbents.

Based on the data in

Table 4. Parameters of adsorption isotherm models for the adsorption of Pb(II) on AMBC.

Temperature (°C)	Langmuir Model				Freundlich Model
	KL (L/mg)	qm (mg/g)	RL	R2	KF (mg/g)	1/n	R2
25	0.0716	210.1	0.046~0.195	0.98947	68.117	0.211	0.97800
35	0.0786	220.3	0.048~0.203	0.98606	69.611	0.215	0.98978
45	0.0826	318.5	0.053~0.218	0.96706	75.708	0.280	0.97686

, it can be seen that the maximum adsorption capacity q_m and Langmuir adsorption constant K_L of AMBC for Pb(II) rise with increasing temperature, indicating that this process is endothermic. The factorless separation factor (R_L) listed in Table 4, which is calculated using Equation (7), reflects the basic features of the Langmuir isotherm equation.

$$R_L=\frac{1}{1+K_L{c}_0},$$

(7)

For the Langmuir model, when R_L = 0, the adsorption is considered irreversible; when 0 < R_L < 1, adsorption occurs relatively easily; when R_L > 1, the adsorption process is challenging to achieve; while R_L = 1 indicates linear adsorption. The R_L values obtained indicate that AMBC can easily adsorb Pb(II).

For the Freundlich model, a favorable adsorption process occurs when 0 < 1/n < 1. In Table 4, all the values of 1/n are less than 1, indicating that this process is favorable and consistent with the analysis of the Langmuir model.

3.5. Thermodynamic Study of the Adsorption

The thermodynamic activation parameters of Pb(II) on AMBC were determined using the equilibrium adsorption data obtained at different temperatures. These parameters include entropy change (ΔS), enthalpy change (ΔH), and free energy change (ΔG) for the adsorption of Pb(II).

These parameters were calculated using the following Equations (8) and (9):

$$\Delta G=-RTln{K}_d,$$

(8)

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

(9)

In these equations, K_d denotes the adsorption equilibrium constant. K_d can be obtained by performing a linear fit of q_e against ln(q_e/C_e), and the resulting linear intercept will yield K_d.

Table 5. Thermodynamic parameters for the adsorption of Pb(II) on AMBC.

Temperature (°C)	Kd	ΔG (kJ/mol)	ΔH (kJ/mol)	ΔS (kJ/mol.K)
25	5.304	−4.14	3.248	0.025
35	5.334	−4.29
45	4.764	−4.13

shows that the ΔG is negative for each temperature under investigation, indicating that the adsorption of Pb(II) by AMBC is a spontaneous process. The positive value of ΔH demonstrates the endothermic properties of the Pb(II) adsorption, which may be due to an increase in the internal diffusion rate of Pb(II) in the AMBC particles with increases in temperature. Moreover, the ΔS value of the AMBC is positive, indicating a higher molecular disorder for the whole adsorption process due to mass transfer on the solid–solution interface.

3.6. Kinetics Study of the Adsorption

To evaluate the activation energy and rate of the adsorption of Pb(II) by AMBC, sorption kinetic experiments were carried out. The effects of temperature on the adsorption of Pb(II) was investigated through time-based analyses. The adsorption experiments were performed under the following conditions: a dosage of AMBC of 0.5 g/L, solution pH of 5, initial Pb(II) concentration of 100 mg/L, and oscillation rate of 130 rpm.

The effects of temperature and contact time on the adsorption by AMBC are shown in Figure 12. It is evident from the adsorption curves that there was a steep increase within the first 40 min for AMBC, indicating that an increase in temperature is beneficial for adsorption. As time progressed, the adsorption curves tended to rise slowly, ultimately reaching equilibrium after 60 min.

In order to gain insight into the rate-limiting steps of Pb(II) adsorption by AMBC, linearized pseudo-second-order kinetic models were employed. The equations for the pseudo-second-order kinetic model are shown in Equation (10).

$$\frac{{dq}_t}{d_t}=K_{ads}{(q_e-q_t)}^2,$$

(10)

Its linearized equation is given in Equation (12):

$$\frac{t}{d_t}=\frac{1}{K_{ads}{q_e}^2}+\frac{1}{q_e}t,$$

(11)

where q_t is the adsorption capacity at time t, mg/g; q_e is the adsorption capacity at adsorption equilibrium, mg/g; and K_ads is the constant of the pseudo-second-order kinetic model, g/mg.min.

The experimental data on the adsorption of Pb(II) by AMBC at different temperatures were fitted by using the pseudo-second-order model (Equation (10)), and the results are shown in Figure 13 and

Table 6. Kinetic parameters from the pseudo-second-order model.

Temperature °C	qe mg/g	kads g/(min.mg)	R2
25	147.0	5.53 × 10−3	0.99989
35	155.3	6.92 × 10−3	0.99991
55	157.5	8.71 × 10−3	0.99992

. The pseudo-second-order model provides the best fit to the experimental data for the AMBC (R² = 0.99). The constant of the pseudo-second-order kinetic model, k_ads increased with increasing adsorption temperatures, and the equilibrium adsorption quantity, q_e exhibited a similar pattern of change. This suggests that higher temperatures favor the adsorption of Pb(II) from aqueous solutions by AMBC, which is in agreement with the results of the thermodynamic analysis. In addition, the pseudo-second-order model also indicates that the adsorption rate of Pb(II) by AMBC is affected by both adsorbents and adsorbates [50].

The Arrhenius equation (Equation (12)) can express the relationship between k_ads and the temperature T:

$$ln{K}_{ads}=lnA-\frac{E_a}{R}(\frac{1}{T})$$

(12)

where A is the collision frequency factor, g/mg·min; R is the ideal gas constant, 8.314 J/mol·K; and Ea is the activation energy of adsorption, J/mol.

Using the data in Table 6, lnK_ads and 1/T were plotted and fitted linearly, and the results are shown in Figure 14.

The linear fitting equation in Figure 14 is y = −0.3317 − 1442.91x. By utilizing this equation, the activation energy for the adsorption by AMBC is calculated to be 11.99 kJ/mol. Typically, the activation energy for physical adsorption is less than 4 kJ/mol [51], indicating that the rate-controlling step of AMBC adsorption for Pb(II) is chemical adsorption.

3.7. Possible Adsorption Mechanisms

XPS can be used to determine the surface functional groups of the adsorbent as well as the state of the adsorbed metals, making it an important technique for adsorbent surface analysis. In order to better understand the possible adsorption mechanism, the XPS before and after Pb(II) adsorption by AMBC was analyzed, and the results are shown in Figure 15. The respective peaks of C 1s were deconvoluted into several possible spectra, having varied contents of oxygen-containing functional groups such as C=O/O-C=O (283.35 eV), C-O (286.36 eV), C-O-C (285.08 eV), C-C (284.69 eV), and C-S (284.23 eV) [52]. The O 1s peak spectrum can be deconvolved into four peaks located at 531.16 eV, 531.94 eV, 532.85 eV, and 533.68 eV, which are attributable to O-H, C=O, C-O, and O=C–O [27]. As shown in Figure 15d, there were three peaks at 169.07 eV, 168.44 eV, and 167.76 eV in the S 2p spectrum, corresponding to S=O, S-O, and S-C [53]. The XPS spectra of AMBC further confirmed that the sulfonated biochar formed -SO₃H groups on the carbon framework after sulfonation. In summary, acidic oxygen-containing functional groups such as -OH, -COOH, and -SO₃H were successfully connected on the surface of AMBC, which was consistent with the results of the FT-IR.

It can be seen from Figure 15a that a new peak corresponding to Pb 4f appeared on the AMBC spectrum after Pb(II) adsorption, which indicates that Pb(II) ions were successfully adsorbed to the surface of the biochar [8]. The characteristic peak of Pb demonstrated a nonstoichiometric adsorption behavior, which could be generally resolved into Pb 4f_7/2 and Pb 4f_5/2, with clear bands around 138.72 eV and 143.57 eV [5]. After the adsorption process, the intensity of the O 1s and S 2p peaks decreased obviously, which may be the result of the interaction between -SO₃H functional groups and Pb(II). In addition, after the adsorption of Pb(II), the intensity of the C-S and O-H functional groups decreased obviously, which indicated that the acidic oxygen-containing functional groups played a leading role in the adsorption of Pb(II).

Comparing the XPS spectra before and after the adsorption of Pb(II), it can be inferred that the acidic oxygen-containing functional groups could promote the binding of AMBC and Pb(II), thus effectively removing the Pb(II) from the water.

4. Conclusions

A biochar precursor (pre-AMBC) was prepared from agricultural waste peanut shells, then treated with concentrated H₂SO₄ to give an acidic group-modified biochar (AMBC). These materials were subsequently used as an adsorbent for the removal of Pb(II) from aqueous solutions. The following conclusions were reached:

(1)

The optimal adsorption performance of AMBC for heavy metal ions in water occurred when the initial concentration of Pb(II) was 100 mg/L, the pH was 5, the dosage of the adsorbent was 0.5 g/L, and the contact time was 120 min. Under these optimal conditions, the Pb(II) removal ratio was 76.0% and the adsorption capacity was 148.6 mg/g, which was much better than the pre-AMBC.

(2)

It is believed that the higher acid content of the AMBC, along with the introduction of the -SO₃H groups, imparts greater electronegativity, stronger complexation, and more active adsorption sites to the AMBC, thereby enhancing its ability to adsorb Pb(II) in aqueous solution.

(3)

The adsorption system followed a pseudo-second-order kinetic model and reached an equilibrium after 90 min. Chemisorption was the rate-limiting step. The adsorption isothermal data showed that the equilibrium adsorption of Pb(II) by AMBC increased with increasing adsorption temperatures, indicating a heat-absorbing process.