Investigation of the Influence of Pyrolysis Temperature on the Adsorption Performance of Municipal Sludge-Derived Biochar Toward Metal Ions

Abstract

In response to the growing issue of iron and manganese pollution in water bodies, this study systematically investigated the adsorption performance of municipal sludge-derived biochar prepared at pyrolysis temperatures ranging from 300 to 700 °C for the removal of Fe²⁺ and Mn²⁺. Among the series of adsorbents (BC300–BC700), BC600—with its well-developed pore structure and high specific surface area—exhibited the best adsorption performance for both metal ions. Kinetic and isothermal adsorption experiments, in combination with XPS characterization, collectively revealed that (1) the adsorption mechanisms of Fe and Mn differ markedly, with Fe adsorption primarily governed by physical interactions, whereas Mn adsorption is largely controlled by chemical processes; (2) Fe²⁺ adsorption occurs mainly via electrostatic interactions and hydrogen bonding; and (3) Mn²⁺ forms carbonate precipitates with C=O groups during redox reactions. Thermodynamic analysis further indicated that the adsorption process was spontaneous and endothermic. Moreover, BC600 demonstrated excellent reusability for Fe adsorption across different water matrices, maintaining efficiencies above 95% after five cycles, although the adsorption performance for Mn declined. This study provides theoretical support for the application of sludge-derived biochar as a cost-effective and efficient adsorbent for metal ion remediation.

1. Introduction

With the continuous advancement of industrialization and social development, water pollution has become an increasingly critical issue. Among various contaminants, metal ions are of concern for their toxicity and bioaccumulation, which may pose severe threats to both human health and aquatic ecosystems upon long-term exposure [1,2]. It has been reported that the typical discharge standards for coal mine wastewater worldwide are 2 mg/L for both Fe and Mn [3]. However, recent studies have predominantly focused on the removal of individual metal ions, while comparatively little attention has been devoted to the simultaneous removal of Fe and Mn. Previous studies have indicated that excessive iron intake may cause organ damage, while manganese exposure has been linked to neurotoxicity and brain tissue degeneration [4]. Redox processes have been reported to enable rapid removal of Fe and Mn. Nevertheless, this approach exhibits certain limitations, including substantial fluctuations in effluent pH and the preferential oxidation of Fe²⁺ (0.77 V) over Mn²⁺ (1.23 V), which hinders the efficient removal of Mn [3].Therefore, the development of efficient, cost-effective, and sustainable strategies for Fe and Mn removal is of great significance.

Adsorption has been widely recognized as an effective strategy for metal ion removal owing to its simple design, low operational cost, and flexible applicability [5]. Traditional adsorbents such as activated carbon, alumina, polymeric resins, and zeolites have been extensively employed in wastewater treatment; however, their practical utility is hindered by factors including high cost, limited selectivity, complicated regeneration procedures, and poor recyclability [5,6]. In recent years, biochar has attracted increasing attention in the field of metal removal owing to its high specific surface area, well-developed porous structure, and abundant oxygen-containing functional groups [7,8,9]. Biochar is a carbonaceous material derived from the pyrolysis of biomass wastes and biodegradable residues, such as crop straw, garden waste, animal manure, and municipal sludge [10]. Due to its inherent electronegativity, biochar exhibits considerable potential for the adsorption of metal ions from wastewater [11]. Considerable research efforts have been devoted to enhancing biochar’s adsorption performance through various surface modification techniques; nevertheless, potential environmental risks associated with such modifications warrant cautious consideration [12]. Recent studies have reported that tuning the pyrolysis temperature can substantially improve the adsorption performance of pristine biochar [4]. For example, Wang et al. prepared a series of turtle-shell-derived biochars (BC300, BC400, and BC500) under different pyrolysis temperatures and found that their Cd²⁺ adsorption capacities followed the order BC500 > BC650 > BC300 [13]. Nevertheless, to date, there have been no systematic investigations on the effect of pyrolysis temperature on the adsorption behavior of unmodified biochar toward binary metal ion systems.

In this study, a series of biochars (BC300, BC400, BC500, BC600, and BC700) were synthesized at varying pyrolysis temperatures. Their adsorption performances toward Fe and Mn were comparatively evaluated to elucidate the influence of pyrolysis temperature on metal ion removal efficiency. The findings are anticipated to provide theoretical insights for the application of sludge-derived biochar in the remediation of metal-contaminated wastewater, thereby advancing the sustainable and circular utilization of municipal sludge resources.

2. Materials and Methods

2.1. Materials and Chemicals

Municipal sludge was collected from a wastewater treatment plant located in Chongqing, China. All chemicals used in this study, including manganese (II) chloride tetrahydrate (MnCl₂·4H₂O), iron (II) sulfate heptahydrate (FeSO₄·7H₂O), hydrochloric acid (HCl), sodium hydroxide (NaOH), potassium pyrophosphate (K₄P₂O₇), sodium acetate (CH₃COONa), potassium periodate (KI₃O₈), 1,10-phenanthroline (C₁₂H₈N₂), and hydroxylamine hydrochloride (NH₂OH·HCl), were of analytical grade and used without further purification.

2.2. Preparation of the Adsorbents

The collected sludge was air-dried in the dark, sieved through a 100-mesh screen, and then oven-dried at 85 °C for 24 h. The dried sludge was subsequently pyrolyzed by heating to 300 °C, 400 °C, 500 °C, 600 °C, and 700 °C at a ramp rate of 10 °C/min, with a holding time of 1 h at the target temperature. The resulting biochars were thoroughly washed with deionized water until the pH became neutral and were designated as BC300, BC400, BC500, BC600, and BC700, respectively. All samples were stored in a desiccator for further use. The preparation process is shown in Figure 1.

2.3. Calculation Formulas and Fitting Models

In this study, the adsorption kinetics were analyzed by fitting the pseudo-first-order, pseudo-second-order, and Elovich models using the adsorption capacity at time t (Q_t, mg/g) as a function of time (t, min). For the four isothermal adsorption models, the equilibrium adsorption capacity (Q_e) was correlated with the equilibrium solution concentration (C_e). All the models employed are summarized below:

(1) The phosphate removal efficiency and adsorption capacity of biochar at equilibrium are calculated as follows [Equations (1) and (2)].

$$\mathrm{R}={\displaystyle \frac{\left({\mathrm{C}}_0-{\mathrm{C}}_{\mathrm{e}}\right)}{{\mathrm{C}}_0}}\times 100\%$$

(1)

$${\mathrm{q}}_{\mathrm{e}}={\displaystyle \frac{\left({\mathrm{C}}_0-{\mathrm{C}}_{\mathrm{e}}\right)\times \mathrm{V}}{\mathrm{m}}}$$

(2)

Here, C₀ and C_e are the concentrations of phosphorus (mg⋅L⁻¹) at 0 and equilibrium, respectively; q_e is the amount adsorbed at equilibrium, mg⋅g⁻¹; V is the volume of the solution, mL; and m is the mass of biochar, g.

(2) The kinetic experimental data were fitted using pseudo-first-order and pseudo-second-order kinetics equations, as well as the Elovich model and class intra-particle diffusion (IPD) model [Equations (3)–(6)].

$${\mathrm{Q}}_{\mathrm{t}}={ \mathrm{Q}}_{\mathrm{e}}\left(1-{\mathrm{e}}^{-{\mathrm{k}}_1\mathrm{t}}\right)$$

(3)

$${\mathrm{Q}}_{\mathrm{t}}={\displaystyle \frac{{\mathrm{k}}_2{\mathrm{Q}}_{\mathrm{e}}^2\mathrm{t}}{1+{\mathrm{k}}_2{\mathrm{Q}}_{\mathrm{e}}\mathrm{t}}}$$

(4)

$${\mathrm{Q}}_{\mathrm{t}}={\displaystyle \frac{1}{\mathsf{\beta }}}\ln (\mathsf{\alpha }\mathsf{\beta }\mathrm{t}+1)$$

(5)

$${\mathrm{Q}}_{\mathrm{t}}={\mathrm{k}}_3{\mathrm{t}}^{0.5}+\mathrm{C}$$

(6)

Here, t is the adsorption time, h; Q_t and Q_e are the amounts adsorbed at time t and equilibrium, respectively, mg⋅g⁻¹; K₁ is the rate constant for the kinetic model, min⁻¹; K₂ is the rate constant for the kinetic model, g⋅(mg⋅min)⁻¹; α is the rate of initial adsorption, mg⋅g⁻¹·min⁻¹; β is the Elovich rate constant; and C is intraparticle diffusion model desorption constant, mg⋅g⁻¹.

(3) The adsorption isotherms were fitted by the Langmuir, Freundlich, Temkin, and Sips models [Equations (7)–(10)].

$${\mathrm{q}}_{\mathrm{e}}={\displaystyle \frac{{\mathrm{Q}}_{\mathrm{m}}{\mathrm{K}}_{\mathrm{L}}{\mathrm{C}}_{\mathrm{e}}}{1+{\mathrm{K}}_{\mathrm{L}}{\mathrm{C}}_{\mathrm{e}}}}$$

(7)

$${\mathrm{q}}_{\mathrm{e}}={\mathrm{K}}_{\mathrm{F}}{\mathrm{C}}_{\mathrm{E}}^{{\displaystyle \frac{1}{\mathrm{n}}}}$$

(8)

$${\mathrm{q}}_{\mathrm{e}}={\displaystyle \frac{\mathrm{R}\mathrm{T}}{{\mathrm{B}}_{\mathrm{T}}}}\mathrm{l}\mathrm{n}\left({\mathrm{A}}_{\mathrm{T}}{\mathrm{C}}_{\mathrm{e}}\right)$$

(9)

$${\mathrm{q}}_{\mathrm{e}}={\displaystyle \frac{{\mathrm{Q}}_{\mathrm{m}}{\left({\mathrm{K}}_{\mathrm{S}}{\mathrm{C}}_{\mathrm{e}}\right)}^{\mathsf{\gamma }}}{1+{\left({\mathrm{K}}_{\mathrm{S}}{\mathrm{C}}_{\mathrm{e}}\right)}^{\mathsf{\gamma }}}}$$

(10)

Here C_e (mg/L) is the equilibrium concentration; Q_m (mg/ g) is the predicted maximum adsorption capacity; K_L (L/mg), K_F ((mg·g⁻¹) (L/mg)^1/n), and K_S (L/mg) are the Langmuir, Freundlich, and Sips constants; and n and γ are empirical parameters. B_T (J/mol) and A_T (mL/mg), respectively, represent the Temkin coefficient and equilibrium binding coefficient.

(4) The thermodynamics parameters of the adsorption processes were estimated using the following equations [Equations (11)–(14)].

$${\mathrm{K}}_{\mathrm{d}}={\displaystyle \frac{{\mathrm{q}}_{\mathrm{e}}}{{\mathrm{C}}_{\mathrm{e}}}}$$

(11)

$$∆{\mathrm{G}}^0=-\mathrm{R}\mathrm{T}\mathrm{I}\mathrm{n}{\mathrm{K}}_{\mathrm{d}}$$

(12)

$$∆{\mathrm{G}}^0=∆{\mathrm{H}}^0-\mathrm{T}∆{\mathrm{S}}^0$$

(13)

$$\mathrm{In}{\mathrm{K}}_{\mathrm{d}}={\displaystyle \frac{∆{\mathrm{S}}^0}{\mathrm{R}}}-{\displaystyle \frac{∆{\mathrm{H}}^0}{\mathrm{RT}}}$$

(14)

Here ΔG⁰ (kJ/mol) is the change in Gibbs free energy, R is the gas constant (8.314 J/(mol·K)), T is the adsorption temperature (K), Kc is the equilibrium constant for the adsorption process (L/g), qe is the adsorption quantity at equilibrium (mg·g⁻¹), Ce is the equilibrium concentration (mg/L), and ΔS⁰ (J/mol/K) and ΔH⁰ (kJ/mol) are the standard entropy change and standard enthalpy change.

2.4. Determination of Fe and Mn Concentrations

The concentrations of Fe²⁺ and Mn²⁺ were determined using the 1,10-phenanthroline spectrophotometric method and potassium periodate spectrophotometric method, respectively. Measurements were performed with a UV–vis spectrophotometer at wavelengths of 508 nm for Fe²⁺ and 525 nm for Mn²⁺.

2.5. Regeneration and Practical Application Experiments

A 0.1 M HCl solution was employed as the desorption reagent for the biochar regeneration experiments. The biochar underwent four regeneration cycles, after which its adsorption performance was evaluated. Practical application experiments were conducted using tap water and secondary sedimentation tank effluent from a wastewater treatment plant as background matrices. In contrast to the deionized water experiments, the Fe and Mn ion solutions were prepared using tap water and secondary sedimentation tank effluent, respectively. Both water samples were collected via the five-point sampling method to ensure stable water quality. All other experimental conditions were maintained identically to those employed in the deionized water background experiments.

3. Results and Discussion

3.1. Characterization of Adsorbents

The surface morphologies of biochars pyrolyzed at different temperatures are shown in Figure 2a–f, intuitively illustrating the influence of pyrolysis temperature on the morphological characteristics of the materials. BC300 and BC400 exhibited relatively dense surfaces, while BC500, BC600, and BC700 developed well-defined pore structures. In particular, the sample prepared at 600 °C also presented a distinct layered structure (Figure 2d). N₂ adsorption–desorption isotherms revealed that all samples displayed typical type IV isotherms (Figure 2g,h), accompanied by pronounced H3-type hysteresis loops in the relative pressure range of P/P₀ = 0.4–1.0. This feature is characteristic of mesoporous materials with irregular pore structures [14,15], which can facilitate the diffusion of metal ions from solution to the adsorbent surface and thereby enhance removal efficiency. Among all samples, BC600 exhibited the most developed pore structure (0.172 cm³/g) and the highest specific surface area (102.87 m²/g) (

Table 1. Morphological parameters of the adsorbents.

Adsorbents	Surface Area (m2/g)	Pore Diameter (nm)	Pore Volume (cm3/g)
BC300	69.67	1.913	0.120
BC400	70.53	1.516	0.116
BC500	97.08	1.768	0.043
BC600	102.87	1.757	0.172
BC700	95.74	1.775	0.166

). According to the correlation analysis between pyrolysis temperature and specific surface area, pore size, and pore volume (Figure 2k–m), pyrolysis temperature exhibited a positive correlation with specific surface area and pore volume, but a negative correlation with pore diameter [16]. To more clearly describe the linear (monotonic) relationships between pyrolysis temperature and specific surface area, pore size, and pore volume, the Spearman model was applied for analysis, and the results are shown in

Table 2. Spearman correlation analysis results.

		Surface Area (m2/g)	Pore Diameter (nm)	Pore Volume (cm3/g)
Pyrolysis temperature	coefficient	0.700	−0.100	0.500
	p	0.188	0.873	0.391
	sample size	5	5	5

. The results showed that the pyrolysis temperature was moderately correlated with the specific surface area and pore volume but was essentially uncorrelated with pore diameter. However, all p-values were greater than 0.05, indicating that the correlations were not significant. The results indicated that the pyrolysis temperature was not a significant factor determining the pore structure of the biochar.

A well-developed pore structure and high specific surface area provide favorable conditions for adsorption, while surface functional groups of the adsorbent also play a critical role in determining its adsorption performance. The FTIR spectra of all samples (Figure 2i) revealed similar functional group compositions. The characteristic peak observed in the range of 1600–1670 cm⁻¹ was attributed to the stretching vibrations of C=O and C=C bonds. The peak at 1011 cm⁻¹ corresponded to the C–O–C stretching vibrations on the biochar surface, while the peak at 462 cm⁻¹ was ascribed to the stretching vibration of Si–O–Si bonds [17,18,19,20]. In addition, XRD analysis was conducted for all samples, and the results are presented in Figure 2j. The diffraction peaks at 20.9° and 50.1° were assigned to the (100) and (112) crystal planes of SiO₂ (PDF#99-0088), respectively [11]. The peak at 26.1° was attributed to (002) crystal planes of graphitic carbon (PDF#99-0057), indicating that higher pyrolysis temperatures promoted the formation of graphitic structures within the biochar matrix [21]. In addition, the characteristic peak at 27.9° corresponded to the (040) crystal plane of Na(AlSi₃O₈). As shown in Figure 2j, the XRD patterns of all samples were nearly identical, suggesting that pyrolysis temperature had no significant effect on the crystalline phase structure of the sludge-derived biochar.

3.2. Effect of Pyrolysis Temperature on the Adsorption of Fe and Mn by Biochar

To investigate the adsorption performance of biochar prepared at different pyrolysis temperatures, a series of adsorption experiments were conducted using biochar synthesized at 300 °C, 400 °C, 500 °C, 600 °C, and 700 °C. The initial concentrations of Fe and Mn were set at 10 mg/L and 6 mg/L, respectively, with an initial pH = 6. The dosage of each biochar adsorbent was 5 g/L, and the contact time was 2 h. The experimental results are shown in Figure 3a. For Mn removal, the removal efficiencies of BC300, BC400, BC500, BC600, and BC700 followed a trend of initial increase followed by a decrease. Specifically, the Mn removal efficiencies were 33.32%, 46.03%, 70.59%, 95.16%, and 55.35%, respectively. This may be attributed to the collapse of the biochar’s porous structure caused by excessively high temperatures, as clearly evidenced in Figure 2d–f. Notably, all biochars prepared at the five pyrolysis temperatures exhibited removal efficiencies for Fe above 98%.

As shown in Figure 3b, all adsorbents exhibited a rapid adsorption phase for Fe within the first 0–30 min, followed by a slower adsorption phase from 30 to 90 min, and finally reached equilibrium between 90 and 120 min. This behavior was primarily attributed to the abundance of available adsorption sites on the biochar surface during the initial stage, which facilitated the rapid uptake of Fe ions driven by the concentration gradient [22]. As illustrated in Figure 3c, the adsorption rate of Mn by BC600 was significantly higher than that of the other biochars, which may be ascribed to the stronger electrostatic interactions between BC600 and Mn ions. Based on the comparative results, BC600 demonstrated superior adsorption performance toward both Fe and Mn and was therefore selected as the target material for subsequent studies.

To investigate the impact of pH on the removal rate of Fe and Mn, the removal experiments were conducted under different pH conditions. As shown in Figure 3e, the removal efficiencies of both Fe and Mn initially increased and then plateaued with increasing pH. Previous studies have indicated that the zeta potential of an adsorbent is pH-dependent [21]. Within the pH range of 2~10, the zeta potential of BC600 remained negative (Figure 3d), suggesting that electrostatic interaction was one of the primary driving forces for Fe and Mn adsorption. Considering practical application conditions, all subsequent experiments in this study were conducted at an initial solution pH of 6. Figure 3f showed that the removal efficiencies of Fe and Mn increased with increasing BC600 dosage, and higher dosages led to shorter times required to reach adsorption equilibrium. This could be attributed to the greater number of available adsorption sites provided by higher dosages of BC600, thereby accelerating the approach to equilibrium. When the BC600 dosages were 1, 2, 3, 5, and 7 g/L, the corresponding Fe removal efficiencies were 95.76%, 96.81%, 98.39%, 99.97%, and 99.91%, while the Mn removal efficiencies were 70.59%, 81.60%, 88.38%, 95.16%, and 96.00%, respectively.

Considering the combined influences of pyrolysis temperature, adsorption time, initial pH, and adsorbent dosage on the removal efficiencies of Fe and Mn, subsequent experiments in this study were primarily conducted using BC600 under the following conditions: adsorption time of 2 h, initial pH of 6, and adsorbent dosage of 5 g/L.

3.3. Kinetic and Isotherm Analysis of Fe and Mn Adsorption

To elucidate the adsorption behavior of Fe and Mn on BC600 and to further identify the rate-limiting steps, adsorption kinetic experiments were conducted. The adsorption capacities at various time intervals were fitted using pseudo-first-order, pseudo-second-order, Elovich, and intra-particle diffusion models, as shown in Figure 4a. The corresponding kinetic parameters are listed in

Table 3. Kinetic parameters of BC600 adsorption of Fe and Mn.

	Pseudo First Order			Pseudo Second Order			Elovich
	Qe (mg·g1)	K1 (min−1)	R2	Qe (mg·g−1)	K2 (g/mg·min)	R2	α (mg·g−1·min)	β (g/mg)	R2
Fe	1.988	1.593	0.999	2.015	1.862	0.999	1.719	13.886	0.984
Mn	1.040	0.082	0.798	1.136	0.118	0.869	1.071	6.079	0.948

. Notably, both the pseudo-first-order and pseudo-second-order models exhibited excellent fits for Fe adsorption (R² = 0.999), while Mn adsorption was best described by the Elovich model (R² = 0.954). According to the FTIR results, no functional groups capable of directly interacting with Fe were detected on BC600. Under different pH conditions, the zeta potential of BC600 remained negative, and the Fe removal efficiency exceeded 90% in all cases. These findings indicate that Fe adsorption was primarily driven by physical adsorption. Therefore, the pseudo-first-order model was more suitable for describing the adsorption process of Fe, whereas Mn adsorption was dominated by chemisorption occurring on a heterogeneous surface, likely involving multilayer chemical interactions [22]. As illustrated in Figure 4b,c, the intra-particle diffusion model indicated that the dominant rate-limiting stage for both Fe and Mn adsorption occurred in the third phase—an equilibrium phase—with corresponding rate constants of k₃ = 0.0003 mg·g⁻¹·min^−1/2 for Fe and k₃ = 0.031 mg·g⁻¹·min^−1/2 for Mn [23].

To further investigate the effect of temperature on Fe and Mn adsorption by BC600, isothermal adsorption experiments were carried out at varying initial concentrations. The experimental data were fitted to four isotherm models: Langmuir, Freundlich, Temkin, and Sips, as presented in

Table 4. The relevant parameters of BC600 for Fe adsorption.

Isotherm Models	Parameters	T (K)
		288	298	308
Langmuir model	Qm (mg·g−1)	12.368	14.189	15.212
	KL (L·mg−1)	91.266	107.788	143.355
	R2	0.815	0.950	0.914
Freundlich model	n	8.292	9.231	8.995
	1/n	0.121	0.108	0.111
	KF ((mg·g−1)(L·mg−1)1/n)	10.176	11.444	12.269
	R2	0.829	0.766	0.820
Temkin model	AT (mL·mg−1)	7666.475	23,909.592	27,457.243
	BT (J·mol−1)	1969.408	2042.707	1989.158
	R2	0.913	0.846	0.878
Sips model	Qm (mg·g−1)	15.253	14.144	15.887
	Ks (L·mg−1)	16.369	109.633	96.512
	γ	0.418	1.055	0.619
	R2	0.932	0.943	0.908

and

Table 5. The relevant parameters of BC600 for Mn adsorption.

Isotherm Models	Parameters	T (K)
		288	298	308
Langmuir model	Qm(mg·g−1)	1.314	1.633	1.701
	KL (L·mg−1)	0.577	3.597	3.838
	R2	0.971	0.877	0.858
Freundlich model	n	3.403	5.747	5.574
	1/n	0.294	0.174	0.179
	KF ((mg·g−1)(L·mg−1)1/n)	0.558	1.094	1.139
	R2	0.852	0.741	0.766
Temkin model	AT (mL·mg−1)	6.161	141.479	141.543
	BT (J·mol−1)	8846.145	10,968.638	10,792.987
	R2	0.932	0.797	0.812
Sips model	Qm (mg·g−1)	1.178	1.532	1.595
	Ks (L·mg−1)	0.695	4.187	4.825
	γ	1.505	5.204	6.562
	R2	0.989	0.955	0.887

and Figure 4d–g. Notably, the Sips model is a modified version of the Langmuir model [24]. Isotherm models can provide insights into adsorption mechanisms, layer structure, and the macroscopic surface characteristics of adsorbents [25]. The results showed that the adsorption capacities for both Fe and Mn increased with increasing initial concentrations until equilibrium was reached. Moreover, adsorption capacity also increased with temperature. At 308 K, the correlation coefficients (R²) for the Fe adsorption models followed the following order: Sips (R² = 0.908) > Langmuir (R² = 0.914) > Temkin (R² = 0.878) > Freundlich (R² = 0.820), and for Mn: Sips (R² = 0.887) > Langmuir (R² = 0.858) > Temkin (R² = 0.812) > Freundlich (R² = 0.766). The maximum adsorption capacities (Q_m) predicted by the Sips model were 15.887 mg·g⁻¹ for Fe and 1.595 mg·g⁻¹ for Mn, which closely matched the experimental values. These results confirmed that the Sips model provided the best fit, indicating that the improved Langmuir model effectively described the monolayer adsorption process. In addition, the Freundlich constant K_F for both Fe and Mn was greater than 1 (Fe: 10.176, 11.444, and 12.269 at 15, 25, and 35 °C, respectively; Mn: 1.094 and 1.139 at 25 and 35 °C), suggesting that Fe adsorption was primarily a monolayer chemisorption process [26]. These findings were consistent with the results derived from the kinetic analysis and the Langmuir model.

3.4. Thermodynamic Analysis of Fe and Mn Adsorption

As shown in

Table 6. Thermodynamic parameters of BC600 adsorption for Fe and Mn.

	T (K)	Kd	${\mathbf{\Delta }\mathbf{G}}^0$ (KJ/mol)	${\mathbf{\Delta }\mathbf{H}}^0$ (kJ/mol)	${\mathbf{\Delta }\mathbf{S}}^0$ (kJ/mol K)
Fe	288	1.217	−0.470
	298	2.218	−1.974	33.117	0.117
	308	2.977	−2.794
Mn	288	0.397	2.318
	298	3.930	−3.391	94.422	0.320
	308	4.806	−4.020

, the standard Gibbs free energy changes (ΔG⁰) for the adsorption of Fe and Mn by BC600 were negative at 288 K, 298 K, and 308 K, indicating spontaneous adsorption processes. Notably, a positive ΔG⁰ value was observed for Mn adsorption at 288 K, suggesting that the adsorption process of Mn under this specific condition was non-spontaneous [27]. The standard enthalpy changes (ΔH⁰) were calculated to be 33.117 kJ/mol for Fe and 94.422 kJ/mol for Mn, both of which were positive, indicating that the adsorption processes were endothermic [28,29]. Notably, the adsorption of Fe was primarily attributed to hydrogen bonding, whereas Mn adsorption was dominated by chemical interactions (the ΔH⁰ was caused by various forces with various ranges, including van der Waals forces = 4 to 10 kJ/mol, dipole forces = 2 to 29 kJ/mol, hydrogen bonding = 2 to 40 kJ/mol, and chemical bonds with bond energies greater than 60 kJ/mol) [30]. Additionally, the positive standard entropy changes (ΔS⁰) of 0.117 kJ/mol·K for Fe and 0.320 kJ/mol·K for Mn implied an increase in randomness at the solid–liquid interface, favoring the spontaneity of adsorption at higher temperatures [31,32,33]. The relatively high ΔS⁰ can be attributed to the greater complexity of the binary metal ion system investigated in this study, compared with a single-ion system [21]. These results also suggest that structural changes may have occurred between the adsorbent and the adsorbates during the adsorption process [34].

3.5. Adsorption Mechanism of Fe and Mn

To elucidate the adsorption mechanism of Fe and Mn ions onto BC600, XRD, FT-IR, and XPS characterizations were performed on BC600 samples after adsorption (Figure 5). As shown, no significant differences in crystal phase structure or surface functional groups were observed between BC600 before (Figure 2i,j) and after adsorption (Figure 5a,b). However, the characteristic peak at 27.9° disappeared after adsorption, indicating that ion exchange might have occurred during the adsorption process. As presented in Figure 4c, a strong linear correlation was found between specific surface area and the adsorption capacities of Fe (R² = 0.887) and Mn (R² = 0.679), suggesting that a larger surface area facilitates Fe and Mn ion uptake [35], which highlights the importance of pore filling in the adsorption process. This conclusion is further supported by Figure 2d,f. In addition, Figure 5d shows a pronounced increase in the Fe signal after adsorption, while no clear Mn peak was detected. The C 1s spectra of BC600 before and after adsorption (Figure 5e,f) were deconvoluted into four components: C–C/C=C, C–O, C=O, and π–π* [21,36]. After adsorption, the relative proportions of C–C/C=C, C–O, and C=O decreased from 52.66%, 31.41%, and 10.48% to 52.45%, 31.17%, and 9.56%, respectively. These changes suggest that the C–C/C=C, C–O, and C=O groups in biochar were involved in interactions with Fe and Mn ions [21]. Notably, the appearance of a carbonate C=O peak at 532.7 eV in the O 1s spectrum after adsorption further supports the participation of C=O groups in binding with Fe and Mn [37]. Meanwhile, based on the Mn 2p spectrum deconvolution results (Figure 5i, j), the appearance of the Mn (II) characteristic peak after adsorption [38] further confirmed that carbonate C=O groups combined with Mn (II) to form MnCO₃ precipitates. Before adsorption, the relative contents of Fe²⁺ and Fe³⁺ were 32.01% and 67.99%, respectively [21,39]. After adsorption, the Fe³⁺ proportion increased from 67.99% to 68.52%, indicating that oxidation occurred during the adsorption process, which again verified the participation of C–O and C=O groups in the adsorption. Together with the adsorption kinetics results, it was deduced that the adsorption processes of Fe and Mn mainly involved (1) Fe²⁺ being oxidized to Fe³⁺ through an oxidation process involving C–O and C=O groups, followed by removal via electrostatic interaction with surface electrons of BC600; (2) Mn²⁺ reacting with C=O groups to form manganese carbonate precipitates.

3.6. Reusability Analysis of BC600 for Fe and Mn Adsorption

To assess the reusability of BC600 for Fe and Mn adsorption, adsorption–desorption cycling experiments were performed using 0.1 M HCl as the desorbing agent. The desorption procedure was conducted as follows: the adsorbed biochar samples were first dried overnight at 60 °C. Subsequently, 0.5 g of the dried biochar was added to 100 mL of 0.1 M HCl and shaken at 200 rpm for 2 h. The desorbed samples were then filtered, thoroughly rinsed with deionized water until neutral pH was achieved, and finally dried overnight. This procedure was applied consistently across all desorption cycles. The adsorption performance of the regenerated biochar toward Fe and Mn is presented in Figure 6. The Fe removal efficiency exhibited only a slight decrease from 99.97% to 96.61% with increasing cycle numbers, whereas the Mn removal efficiency declined markedly from 95.16% to 60.43%. The pronounced decrease in Mn removal efficiency could be ascribed to the progressive occupation or depletion of active adsorption sites. These results demonstrated that BC600 possessed excellent reusability for Fe adsorption, while its reusability for Mn adsorption was relatively limited. Furthermore, when tap water (Figure 6b) and secondary effluent (Figure 6c) were employed as background matrices, the adsorption and reusability trends of BC600 toward Fe and Mn were consistent with those observed in ultrapure water (Figure 6a). Notably, during five adsorption–desorption cycles in both tap water and secondary effluent, the Fe removal efficiency remained above 95.49% and 94.83%, respectively, whereas the Mn removal efficiency decreased from 90.07% and 91.77% to 48.57% and 44.33%, respectively. Although BC600 exhibited poor reusability for Mn in cyclic tests, its initial Mn removal efficiency in all three water matrices exceeded 90%, indicating strong adsorption affinity. Overall, these findings suggest that BC600 holds considerable potential for metal ion removal, and its cyclic stability could be further enhanced through surface functionalization or composite modification in future applications.

4. Conclusions

This study systematically investigated the influence of pyrolysis temperature on the removal of Fe and Mn by sludge-derived biochar. Among the five biochars evaluated, BC600 exhibited the highest removal efficiencies for both Fe and Mn, attributable to its optimized pore structure and specific surface area. Kinetic and isothermal adsorption experiments, in combination with XPS characterization, collectively revealed that (1) the adsorption mechanisms of Fe and Mn differ markedly, with Fe adsorption primarily governed by physical interactions, whereas Mn adsorption is largely controlled by chemical processes; (2) Fe²⁺ adsorption occurs mainly via electrostatic interactions and hydrogen bonding; and (3) Mn²⁺ forms carbonate precipitates with C=O groups during redox reactions. Thermodynamic analysis further indicated that the adsorption process was spontaneous and endothermic. Moreover, BC600 demonstrated excellent reusability for Fe adsorption across different water matrices, maintaining efficiencies above 95% after five cycles, although the adsorption performance for Mn declined. This work focused specifically on the effect of final pyrolysis temperature on the adsorption behavior of unmodified biochar in a binary metal ion system. Future studies will examine the influence of pyrolysis parameters, including heating rate, atmosphere, and holding time, on the adsorption performance of biochar. In addition, the physicochemical and thermal properties of biochar during regeneration will be investigated in detail. Overall, biochar represents a low-cost and environmentally benign material with considerable potential for the remediation of Fe- and Mn-contaminated wastewater.