Nitrate Absorption and Desorption by Biochar

Abstract

Biochar is a potential solution for addressing environmental problems related to excessive nitrogen (N). However, there is still some debate about the absorption and desorption of nitrate nitrogen (NO₃⁻-N). Therefore, this study investigated the NO₃⁻-N adsorption and desorption performance onto biochar and biochar-soil mixture to address this gap. The results showed that the biochar produced from apple branches had the ability to absorb NO₃⁻-N with an absorption capacity of 3.51 mg·g⁻¹. The absorption data fitted well with the pseudo-second-order kinetic model and Langmuir model. The application of biochar significantly improved soil absorption capacity and slow release of NO₃⁻-N. While higher NO₃⁻-N concentrations had better NO₃⁻-N supply capacity and poorer slow-release effect. Integrating nutrient supply and slow-release effect, it is recommended to control the application ratio of biochar to NO₃⁻-N at 34–42.75 g·g⁻¹. Although the unoptimized biochar application rate cannot be directly applied to the soil as a slow-release fertilizer carrier to meet commercial standards, biochar modification provides new possibilities for this purpose. Moreover, compared with traditional slow-release fertilizer, biochar had good stability and regeneration performance, alleviating the high cost due to the biochar price. In general, biochar still has potential and prospects as a slow-release material. This study provides support for biochar in mitigating environmental problems associated with excess N.

1. Introduction

N fertilizer serves as the primary source of nourishment for approximately 48% of the global population [1], and the demand for maximizing land productivity has led to a further increase in N fertilizer. According to the National Bureau of Statistics of China, the amount of N fertilizer used in China has increased from 9.3 million t in 1980 to 18.3 million t in 2021. The high input of N fertilizer and unequal N balance lead to N leaching [2], weakening the ability of soil to provide nutrients for plant growth and increasing the N content in the receiving water environments [3]. These contribute to reducing N fertilizer utilization efficiency, groundwater pollution, and eutrophication of lake water bodies [1,4,5].

The inorganic N in the soil mainly exists as ammonium (NH₄⁺) and nitrate (NO₃⁻), with soil colloids having a stronger affinity for NH₄⁺ [6]. In comparison, the lower soil anion exchange capacity (AEC) and anion competition (phosphate and sulfate anions) inhibit NO₃⁻ binding [7,8]. Consequently, soil N leaching is largely dominated by the level of NO₃⁻-N [9,10]. To mitigate the adverse environmental impact of NO₃⁻-N leaching, NO₃⁻-N immobilization is considered a viable method [11,12], and the use of absorbents to bind NO₃⁻ holds promise for achieving this goal [13]. However, most absorbents are more suitable for the removal of NO₃⁻-containing wastewater than for application in soil [3,14,15]. Additionally, the implementation of slow-release fertilizers represents a useful intervention to reduce NO₃⁻-N leaching by releasing nutrients in a controlled manner [16]. Overall, seeking NO₃⁻-N absorbents and slow-release materials suitable for soil is a hopeful method to reduce N leaching.

Recently, the utilization of modern waste technology has drawn attention to the preparation of biochar from agricultural waste as an absorbent [17,18,19]. Biochar, a solid product synthesized using pyrolysis and carbonization in an anaerobic environment [20], exhibits rich pores, a large specific surface area, abundant functional groups, and exchangeable cations, all of which contribute significantly to its excellent absorption capacity [21,22,23]. Previous studies have proposed biochar as an effective absorbent for NO₃⁻- immobilization [24]. Kameyama et al. [25] reported that bagasse biochar absorbs NO₃⁻ through interaction with functional groups. It is also found that the NO₃⁻ absorption capacity of modified bagasse biochar is as high as 28.21 mg·g⁻¹ due to the carbon skeleton and honeycomb structure [26]. Additionally, biochar derived from rice husk can absorb NO₃⁻ at 2.1 mg·g⁻¹ [27], and this property has enabled its successful application in reducing NO₃⁻ leaching from soil [28]. However, not all biochar can be used for NO₃⁻ absorption [29,30]. This may be related to the raw material and pyrolysis temperature. In other words, the NO₃⁻ absorption characteristics of biochar need to be carefully verified before using it to reduce N leaching. Furthermore, biochar shows promise as a slow-release fertilizer material [16,31,32] owing to its high aromatic stability and porous structure [33,34]. Lateef et al. [35] reported that biochar derived from corncob prolongs the nutrient release period. Biochar-based fertilizers are steadier and more consistent than mineral fertilizers, thus reducing the N losses [34]. Compared with pure urea, molten urea-impregnated biochar also has slow-release performance and reduces the N release rate [36]. Similar results were observed after biochar was mixed with urea and minerals [37]. However, despite biochar’s potential in absorption and slow-release performance for N, practical application of its sorption and desorption properties remains challenging and faces certain limitations, necessitating further research [16].

The favorable effects of biochar on soil fertility, crop yield, and N loss are closely related to the biochar application rate [1,25,38]. However, there is ongoing debate about the optimal biochar application amount, and no consensus has been reached on the recommended biochar application rate. The range of biochar application varies widely in agricultural production systems, from 1 to 200 t·ha⁻¹ [39], while from an agronomic and economic standpoint, the appropriate amount is suggested to be 15 t·ha⁻¹ [40]. Considering the effects of biochar on photosynthesis, an appropriate biochar application rate is suggested to be within the range of 10.1–20 t·ha⁻¹ or 2.01–4% [41]. These may be partly due to differences in biochar performance, N fertilizer application rate, and crop species. For a given biochar, the optimal biochar application rate may depend more on the fertilization regime and soil nutrients [28,32]. Therefore, it is more theoretically significant to determine the optimal biochar application amount based on the biochar absorption and desorption behavior at different nutrient levels.

Under the support of multitudinous advantages of biochar made from different feedstocks [1,28,42,43], high cost is the main factor limiting the large-scale application of biochar. It is reported that the biochar price varies between $80 t⁻¹ and $13,480 t⁻¹. Despite substantial reductions in production costs owing to technological advancements [44], as an independent farming input, biochar still proves uneconomical [45]. Several investigations have shown that biochar has excellent stability and reusability performance, offering a potential solution to mitigate its high cost [40,46]. Li et al. [47] reported that the lignin-based biochar exhibited excellent reusability. Due to the reversible ion exchange, the absorption efficiency of TI by biochar is as high as 95% after five absorption/desorption cycles [48]. Even after 12 months of biochar application, soil still had high NO₃⁻-N content and showed positive effects on crop growth and yield during the second and third year [40]. These imply that biochar has a long-term absorption effect. However, there are few studies on the long-term effects and stability of biochar on NO₃⁻-N absorption, which necessitates further exploration.

According to these, compared with previous studies, this study was designed to investigate the absorption and desorption characteristics of NO₃⁻-N onto biochar and address the key gaps in the current knowledge by optimizing the biochar dosage and NO₃⁻-N concentration to improve the application effectiveness. In addition, by evaluating the regeneration ability, the potential benefits of biochar for nutrient management were explored, providing theoretical support for the long-term economic feasibility of sustainable slow-release fertilizer carriers. The specific objectives are as follows: (i) explore the absorption and desorption behavior of NO₃⁻-N onto biochar; (ii) optimize the application rate based on absorption and desorption performance, and (iii) evaluate the recyclability of biochar and its possibility as NO₃⁻-N slow-release fertilizer carrier. In summary, this study provides valuable insights into the practical application of biochar in commercial fertilizer product development.

2. Material and Methods

2.1. Soil Preparation and Biochar Production

The soil used in the experiment was taken from a maize field near the Northwest Agriculture and Forestry University (34°15′ N, 108°01′ E) and consisted of 27.8% clay, 37.7% silt, and 34.5% sand, with a bulk density of 1.35 g·cm⁻³. The soil was taken from 0–20 cm of the tillage layer and rinsed with plenty of deionized water, air dried, and sieved through a 2 mm sieve. After air drying, the soil was found to contain 2.3 mg·kg⁻¹ of NO₃⁻-N and 1.9 mg·kg⁻¹ of NH₄⁺-N.

The biochar was purchased from Shaanxi Yixin Bioenergy Technology Development Co., LTD. The feedstock was apple tree branch pyrolyzed at 450 °C for 30 min in the short supply of oxygen. The prepared biochar was manually ground into powder, cleaned with deionized water, and air-dried [49,50,51]. The soil and biochar were accurately weighed using an electronic weigher (one ten-thousandth). A blender was used to thoroughly mix the soil with the biochar.

2.2. Characterization of the Biochar

The physical and chemical properties of the biochar are shown in

Table 1. Physicochemical properties of biochar.

Feedstock	Apple Branch
pHpzc	3.4
pH	9.03
Mean pore size (nm)	5
Surface area (m2·g−1)	192
CEC (mol·kg−1)	50.11
Element content (%)	C	79.22
	H	2.21
	O	14.01
	N	1.03
	Mg	0.97
	Ca	0.83
	K	0.81

. The pH of the biochar was determined using a pH meter in a 1:5 ratio of biochar to deionized water. The element contents were measured using an elemental analyzer (Elementar Analysensysteme GmbH, Berlin, Germany) coupled with an energy-dispersive X-ray spectroscopy (EDS, Oxford Instruments, Concord, MA, USA). Cation exchange capacity (CEC) is measured using the barium chloride-sulfate forced exchange method [52]. Fourier Transform infrared spectroscopy (FTIR) was obtained using KBr pellets and the Nicolet iS50 FT-IR (Thermo Fisher Scientific Co., Ltd., Waltham, MA, USA) with 16 scans over 400–4000 cm⁻¹ at a resolution of 2 cm⁻¹ to identify the functional groups present on the biochar surface [53,54]. The specifical surface area and mean pore diameter of biochar were determined using nitrogen (N₂) sorption–desorption isotherms at −77 K (Belsorp max II instrument, Osaka, Japan). The apparent point of zero charge (pH_pzc) was determined using Zetasizer Nano ZS90 (Malvern Instruments, Worcestershire, UK).

2.3. Absorption Experiment on Biochar

The biochar absorption experiment included two parts: absorption isotherm and kinetic experiments. Various NO₃⁻-N standard solutions were prepared using KNO₃. In the absorption isotherm experiment, 1 g of biochar was placed in a 250 mL conical flask and mixed with a 50 mL solution containing varying initial concentrations of NO₃⁻-N (20–180 mg·L⁻¹). Due to the biochar pH being 9.03, the initial pH of the mixed solution was 8.8 to 9.0. To mitigate the effect of pH on the absorption process, the solution pH was adjusted to 7.0 using 1 mol·L⁻¹ HCl [55]. Use 0.01 mol·L⁻¹ KCl as the background electrolyte [56]. To prevent evaporation of water, the conical flasks were sealed with plastic wrap and stirred at 120 rpm for 4 h at 25 °C. After the absorption process, the solution pH was 7.1–7.2. The suspension samples were then filtered through a 0.45 μm syringe to prevent the influence of biochar on absorbance. The NO₃⁻-N concentrations of the filtrates were measured using the sulfamic acid UV spectrophotometry method (an error of 2% and a measurement range of 0–4 mg·L⁻¹) [57,58]. The specific measurement process was as follows: (1) Transferred the filtrate to a sealed centrifuge tube to prevent evaporation. (2) Prepared standard NO₃⁻-N solutions (0, 0.25, 0.50, 1.00, 1.50, 2.00, 2.50, 3.00, 3.50, and 4.00 mg·L⁻¹) by dissolving KNO₃. Then, 50 mL of each standard solution was mixed with 1 mL of HCL (1 mol·L⁻¹) and 0.1 mL of sulfamic acid (0.8%). Absorbance was measured at 220 nm and 275 nm, and the calibration curve was generated by plotting the NO₃⁻-N concentration (mg·L⁻¹) against the calculated absorbance values at 220 nm minus twice the absorbance values at 275 nm (220 nm–2 × 275 nm) for a series of standard solutions. (3) Determined the absorbance of the filtrate using the same method and calculated the NO₃⁻-N concentration from the calibration curve. (4) If the obtained NO₃⁻-N concentration was outside the measurement range, the filtrate was diluted and then re-measured to improve the accuracy of the measurement.

Similarly, in the absorption kinetic experiment, 1 g of biochar was mixed with 100 mL of a 100 mg·L⁻¹ NO₃⁻-N solution, and the contact time was varied from 0 to 480 min. The experimental conditions, including the pH, stirring speed, temperature, and NO₃⁻-N measurements, were kept identical to those of the absorption isotherm experiment.

2.4. Absorption, Desorption and Regeneration Experiments on Biochar-Soil Mixture

A few drops of chloroform were added during the biochar-soil mixture absorption and desorption experiments to prevent the influence of microorganisms [59,60]. No denitrification in the biochar-soil mixture systems after the use of chloroform. The absorption experiment for the biochar-soil mixture was conducted at varying biochar application rates (0–8%) and initial NO₃⁻-N concentrations (20–240 mg·L⁻¹). Specifically, 13.5 g of the biochar-soil mixture was placed in a PE pipe with a diameter of 1 cm, which was sealed with 100 mesh nylon at the top and bottom. PE pipe and nylon mesh did not absorb NO₃⁻-N. The sealed PE pipe was then immersed in 250 mL of NO₃⁻-N solution with different initial concentrations. Moreover, the background electrolyte was 0.01 mol·L⁻¹ KCl, and the solution pH was adjusted to 7.0 with 1 mol·L⁻¹ HCl. After 48 h, the solution had a pH of 7.04–7.2, and the NO₃⁻-N concentration in the solution was measured to calculate the absorption capacity of the mixture, following the procedure described previously.

After the absorption experiment, the spent biochar-soil sample was separated from the solution and dried in a vacuum oven at 30 °C. The dried spent biochar-soil mixture sample was then immersed in 1000 mL of deionized water, and the deionized water was replaced every 24 h by transferring a PE pipe containing the soil-biochar mixture with tweezers. Over a duration of 168 h, the NO₃⁻-N concentration in the filtrate was measured to calculate the desorption capacity and efficiency of the mixture. These procedures constituted the biochar-soil mixture desorption experiment. Moreover, the absorption-desorption experiment of 168 h (4% biochar application rate and 80 mg·L⁻¹ NO₃⁻-N concentration) was repeated seven times as the regeneration experiment.

2.5. Data Analysis

The absorption capacity Q_e (mg·g⁻¹) and the absorption rate (%) were calculated as follows [56]:

Q_e = V(C_o − C_e)/M

(1)

absorption rate = 1 − C_e/C_o × 100%

(2)

where C_o and C_e (mg·L⁻¹) are the initial and equilibrium concentrations of NO₃⁻-N, respectively. V (L) and M (g) are the solution volume and mass of dry biochar, respectively. The meanings of Q_e, C_o, and C_e are the same in the following. Considering the difference in absorption capacity between treatments, the relatively low absorption rate can meet the requirement of experimental accuracy [61,62,63].

Langmuir (Equation (3)) and Freundlich (Equation (4)) isotherm models were employed to describe absorption isotherms of NO₃⁻-N by biochar.

$$Q_e{=(Q}_{max}{K}_L{C}_e{)/(1+K}_L{C}_e)$$

(3)

$$Q_{e }{=K}_F{ C}_e{}^{1/n}$$

(4)

where Q_max (mg·g⁻¹) is the maximum absorption capacity. K_F and K_L are the Freundlich and Langmuir balance constants, respectively. A dimensionless constant separation factor R_L in Langmuir isotherm models is defined as [64]:

$$R_L=1/({1+K}_L{C}_o)$$

(5)

The absorption kinetics of NO₃⁻-N onto biochar was investigated using the pseudo-first and second-order kinetic models (Equations (6) and (7)), Elovich model (Equation (8)) and Weber-morris intraparticle diffusion model (Equation (9)).

$${log(Q}_e-{ Q}_t{)=logQ}_e-k_{1}t/2.303$$

(6)

$${t/Q}_t{=1/(k}_2{Q}_e{}^2{)+t/Q}_e$$

(7)

$$Q_t=\left(ln\right(ab)+lnt)/b$$

(8)

$$Q_t{=k}_{pi}{t}^{1/2}{+C}_i$$

(9)

where k₁ (min⁻¹) and k₂ (mg·(g·min)⁻¹) are the pseudo-first and second-order kinetic model absorption rate constant, respectively. Q_t (mg·g⁻¹) is the absorption capacity at time t, respectively. a (mg·(g·min)⁻¹) is the initial absorption rate, and b (mg·g⁻¹) is the desorption constant. k_pi (mg·(g·min^1/2)⁻¹) and C_i are the intraparticle diffusion rate constant and intercept of stage i, respectively.

The biochar desorption efficiency and capacity were obtained by:

$$D_t={{\displaystyle \sum }}_{24}^t{V}_t{C}_t$$

(10)

$${\eta }_t{=(D}_t-D_{t - 24}{)/Q}_e$$

(11)

where D_t (mg·g⁻¹) is the cumulation desorption capacity at time t. V_t (L) and C_t (mg·L⁻¹) are the filtrate volume and NO₃⁻-N concentration at time t. η_t (%) is the desorption efficiency at time t.

The sustainable index (SI) and Coefficient of variation (CV) were calculated as follows [65]:

SI = (Ῡ − σ)/Y_max

(12)

CV = (σ/Ῡ)

(13)

where Ῡ is 24 h average desorption efficiency/capacity, and Y_max is the maximum desorption efficiency/capacity in 24 h. σ is the standard deviation,

Data analysis and statistical tests were performed using SPSS 17.0 (SPSS Inc., Madison, WI, USA) and MATLAB R2020b software (MathWorks Inc., Natick, MA, USA). Diagrams and graphs were drawn using OriginPro 2016 (OriginLab Corporation, Northampton, MA, USA) software.

3. Results

3.1. Characterization

The FTIR spectra exhibited distinguishable differences between the pre- and post-sorption stages of NO₃⁻-N onto apple branch biochar (Figure 1). The band around 3420 cm⁻¹ corresponded to the valence vibrations of the -OH bands. The intensity of absorption peak at 1585 cm⁻¹ corresponds to -COOH, confirming the existence of chemical interactions of NO₃⁻-N with biochar. Moreover, apple branch biochar possessed a specific surface area of 192 m²·g⁻¹ and an average pore size of 5 nm (Table 1), establishing a theoretical foundation for the physisorption of NO₃⁻-N by biochar [54,66]. The cations in the biochar also contributed to the absorption of NO₃⁻-N. However, it is worth noting that the pH_pzc and pH of biochar were measured at 3.4 and 9.03 (Table 1), respectively, which may not be favorable for the absorption process. In general, these varied physicochemical properties provide different performances for the absorption of NO₃⁻-N onto biochar.

3.2. NO₃⁻-N Absorption by Biochar

3.2.1. Absorption Isotherms of NO₃⁻-N by Biochar

Figure 2 shows the biochar absorption capacity for NO₃⁻-N at different initial NO₃⁻-N concentrations. Initial NO₃⁻-N concentrations significantly affected the biochar absorption capacity. The absorption capacity first increased with the addition of NO₃⁻-N concentration and then gradually stabilized. Obviously, the absorption capacity did not increase linearly with the initial NO₃⁻-N concentration. An 800% increase in the initial NO₃⁻-N concentration resulted in only a 327% increase in absorption capacity. This is because the high initial NO₃⁻-N concentration can enhance the absorption process by overcoming all mass transfer resistances between the liquid and solid phases [67]. However, at elevated NO₃⁻-N concentrations, the active biochar sites were the main factor affecting absorption capacity. These findings imply that biochar absorption capacity is governed by both the biochar application amount and NO₃⁻-N concentration. Moreover, according to Equation (2), the absorption rate of NO₃⁻-N was 39–82% at different initial NO₃⁻-N concentrations.

The absorption isotherm of NO₃⁻-N on biochar was simulated using the Freundlich and Langmuir models, and differences in the accuracy of the simulation results were observed in Table 1. The R² values for the Langmuir and Freundlich models were 0.977 and 0.946, respectively, indicating a better description of Langmuir model. Moreover, the maximum absorption capacity observed in the Langmuir model was closer to the measured value. However, the value of R² for the Freundlich model suggests that the biochar surface is heterogeneous [68]. The high R² value of the Langmuir model indicates that biochar absorption sites are energetically identical, and the absorption occurs on a structurally similar binding site [68,69].

The absorption process properties were estimated using equilibrium constants. The value of 1/n is a benchmark to assess the absorbate interaction [70]. In the Freundlich model, the absorption constant 1/n was determined to be 0.580 (less than 1), indicating a favorable process [68]. The difficulty of the absorption reaction can be reflected by the value of R_L, calculated using the Langmuir model absorption equilibrium constant K_L (Equation (4)). The value of R_L was in the range of 0.390–0.852, less than 1. It is generally considered that the absorption reaction is easy to proceed. The values of 1/n and R_L indicate that the absorption of NO₃⁻-N on the biochar surface was favorable.

3.2.2. Absorption Kinetic of NO₃⁻-N by Biochar

The effect of contact time on NO₃⁻-N uptake by biochar was investigated for a period of 480 min. The absorption capacity of NO₃⁻-N by biochar significantly increased with the rising contact time (Figure 3). The biochar absorption capacity was 1.86 and 2.45 mg·g⁻¹ at the contact time of 5 and 20 min, respectively. As the contact time increased from 20 to 120 min, the biochar absorption capacity accordingly increased from 2.45 to 3.39 mg·g⁻¹. After 120 min, the absorption capacity was almost constant (increased from 3.39 to 3.51 mg·g⁻¹), so it can be considered as the equilibrium time. Similarly, the absorption rate of NO₃⁻-N showed the same trend, ranging from 68% to 70% after 120 min. From this behavior, it is evident that the absorption process occurred in three steps: the first step was a fast step (0–20 min), followed by a slower second phase (20–120 min), leading to equilibrium.

The Weber-Moriss intraparticle diffusion model was used to distinguish the possible rate control step affecting the absorption kinetics [71]. NO₃⁻-N absorption capacity was plotted against T^0.5 (Figure 3b). Three steps were observed in the pots, and the curve gradually entered a horizontal state from the rising trend at the initial stage. It was also noticed that the diffusion pots did not pass through the origin, which indicated that both film and intraparticle diffusion control and limit the absorption system.

Kinetic models were performed to investigate the mechanism of NO₃⁻-N absorption onto biochar. The kinetic absorption data were subjected to the pseudo-first and second-order models, the Elovich model, and the Weber-Moriss intraparticle diffusion model, and model parameters were summarized in

Table 2. Absorption constants and parameters for the isotherm and kinetic models.

Isotherm models	Langmuir model constants	Qmax	6.09
		RL	0.390–0.852
		R2	0.977
	Freundlich model constants	K	0.189
		1/n	0.580
		R2	0.946
Kinetic models	Pseudo-first-order model parameters	Qe	3.219
		k1	0.258
		R2	0.657
	Pseudo-second-order model parameters	Qe	3.449
		k2	0.048
		R2	0.903
	Elovich model parameter	R2	0.947
	Weber-Moriss intraparticle diffusion model parameters	Kip1	0.256
		R2	0.912
		Kip2	0.145
		R2	0.969
		Kip3	0.006
		R2	0.273

. The R² value and calculated Q_e value were used to select the most suitable model for the absorption process. The higher R² value suggested that the pseudo-second-order model described the absorption process well as compared to others. Furthermore, the closely calculated Q_e and experimental absorption capacity values provided further support. The R² value of 0.947 for the Elovich model proposes a chemisorption process [72].

3.3. NO₃⁻-N Absorption by Biochar-Soil Mixture

3.3.1. Biochar Application Rate

The NO₃⁻-N absorption capacity onto biochar-soil mixture at different biochar application rates is shown in Figure 4. As expected, the application of biochar significantly improved the uptake of NO₃⁻-N by soil. At 1% biochar application rate, NO₃⁻-N absorption capacity by biochar-soil mixture was about 0.32 mg·g⁻¹, which was not significantly different from the CK (no added biochar). However, the biochar-soil mixtures had a significantly higher absorption capacity than CK at both the 2% and 4% application rates, with a notable difference between the 2% and 4% application rates. Compared with CK, the absorption capacity increased by 10%, 106%, 162%, and 177% at the biochar application rate of 1%, 2%, 4%, and 8%, respectively. Obviously, the absorption capacity did not proportionally correspond to biochar application rates. Moreover, there was no remarkable difference in absorption capacity between 4% and 8% biochar application rates. This is because when the initial NO₃⁻-N concentration is ascertained, a high biochar application rate is prone to competitive absorption, which leads to a decrease in absorption capacity [26]. Absorption rate increased with increasing biochar application rate, with 20%, 22%, 40%, 51%, and 54% for CK, 1%, 2%, 4% and 8% treatment, respectively. Overall, high biochar application rates increased the soil absorption capacity but reduced the biochar application efficiency.

3.3.2. NO₃⁻-N Concentration

The effect of NO₃⁻-N concentration on biochar-soil mixture absorption capacity was investigated at a 4% biochar application rate with NO₃⁻-N concentration ranging from 20 to 240 mg·L⁻¹ (Figure 5). The response of biochar-soil mixture absorption capacity to NO₃⁻-N concentration was consistent with that of biochar. Initial NO₃⁻-N concentration greatly affected the absorption capacity. The absorption capacity was 0.284, 0.408, 0.755, 0.801, and 0.837 mg·g⁻¹ for the NO₃⁻-N concentrations of 20, 40, 80, 160, and 240 mg·L⁻¹, respectively, corresponding to absorption rate of 77%, 55%, 51%, 27%, and 19%. The high absorption capacity occurred for high initial NO₃⁻-N concentrations. This increase is due to the fact that higher NO₃⁻-N concentration strengthens the driving force between absorbate and improves collision probability between biochar and NO₃⁻-N. However, there was no significant difference in the absorption capacity of 80, 160, and 180 mg·L⁻¹ NO₃⁻-N concentrations, which can be explained by the limited number of absorption sites.

3.4. Desorption of NO₃⁻-N from Biochar-Soil Mixture

3.4.1. Biochar Application Rate

The desorption efficiency of NO₃⁻-N from the soil is significantly affected by biochar addition and desorption time (Figure 6). Regardless of the biochar application rates, the desorption behavior of NO₃⁻-N followed a consistent decreasing trend over time. According to the desorption curves, the desorption process exhibited two distinct phases: the initial 24 h and the period after that. Within the first 24 h, all treatments showed the highest NO₃⁻-N desorption efficiency, likely due to variations in the binding between NO₃⁻-N and biochar, leading to different degrees of adhesion [73]. The most significant difference in desorption efficiency among treatments was also observed during the initial 24 h. Notably, lower biochar application rates resulted in higher desorption efficiency, with the CK treatment showing the highest desorption efficiency. However, beyond the initial 24 h, higher biochar application rates demonstrated a larger desorption efficiency, attributed to varying declining patterns of desorption efficiency. The smooth desorption efficiency curves at high biochar application rates demonstrated that the biochar could obviously lead to a slower release process of NO₃⁻-N in soil, and greater biochar application rates provided better slow-release effects.

Cumulative desorption capacity is a vital index for evaluating nutrient supply, and it was significantly influenced by the biochar application rate (Figure 7). In the first 24 h, the desorption capacity of NO₃⁻-N ranged from 0.164 to 0.256 mg·g⁻¹ for different biochar application rates. Although the 8% biochar application rate showed the smallest desorption efficiency in the first 24 h (Figure 6), the desorption capacity was still significantly higher than that in CK and the 1% biochar application rate in the first 24 h. This is because the absorption capacity of NO₃⁻-N in the 8% biochar application rate was significantly higher than that of CK and 1%. The results also showed that the highest NO₃⁻-N desorption capacity was observed at the 2% biochar application rate in the first 24 h, which was attributed to both the high desorption efficiency and NO₃⁻-N content in the biochar-soil mixture (Figure 4). This indicates that the desorption capacity of NO₃⁻-N in the soil was controlled by the desorption efficiency and absorption capacity, both closely related to the biochar application rate.

The cumulative desorption capacity of NO₃⁻-N over 168 h was 0.283, 0.299, 0.542, 0.702, and 0.739 mg·g⁻¹ at CK, 1%, 2%, 4%, and 8% biochar application rate, respectively. The cumulative desorption capacity of 2%, 4%, and 8% biochar application rates were significantly higher than 1% and CK. However, there was no significant difference between 1% and CK. Similarly, no significant difference was observed between 4% and 8% biochar application rates. This result could be attributed to insignificant differences in their cumulative desorption efficiency and absorption capacity. Overall, soil with higher biochar application rates exhibited better NO₃⁻-N availability, but the cumulative desorption capacity did not consistently increase significantly with biochar addition.

3.4.2. NO₃⁻-N Concentration

The desorption behavior at different initial NO₃⁻-N concentrations was also investigated (Figure 8). The initial NO₃⁻-N concentration had a significant impact on the desorption efficiency and capacity. Similar to different biochar application rates, cumulative desorption capacity increased significantly with increasing NO₃⁻-N concentrations. The high NO₃⁻-N concentrations could provide more NO₃⁻-N (Figure 8). The cumulative desorption capacity of 80, 160, and 240 mg·L⁻¹ NO₃⁻-N concentrations was significantly higher than 20 and 40 mg·L⁻¹. However, the steeper desorption efficiency curve was observed for the high NO₃⁻-N concentration treatment, indicating a poorer slow-release effect. In general, high NO₃⁻-N concentrations had better NO₃⁻-N supply capacity, while low NO₃⁻-N concentrations favored the slow release of NO₃⁻-N.

3.4.3. Stability and Sustainability of the Desorption Process

The CV and SI were employed to assess the stability and sustainability of the NO₃⁻-N desorption process. As shown in Figure 9, both biochar application rate and NO₃⁻-N concentration significantly affected the stability and sustainability of desorption efficiency. At different biochar application rates, the maximum (1.68) and minimum (0.80) CV were observed in the CK and 8% biochar application rates, respectively. The lowest and highest CV at different NO₃⁻-N concentrations were observed in the 20 and 240 mg·L⁻¹ NO₃⁻-N concentration treatments, respectively. With increasing biochar application rate and NO₃⁻-N concentration, the corresponding CV values decreased and increased, respectively, indicating that the high biochar application rate and low NO₃⁻-N concentration enhanced the desorption stability. Similar to CV, the high biochar application rate and low NO₃⁻-N concentration were conducive to the sustainability of NO₃⁻-N desorption due to the higher SI values. This reiterates that the slow-release effect is enhanced by the high biochar application rate and low NO₃⁻-N concentration.

3.5. Optimization of Biochar Application Rate and NO₃⁻-N Concentration Management

The cumulative desorption efficiency did not differ significantly between treatments, suggesting that desorption capacity is closely related to absorption capacity. Therefore, the absorption capacity can be used to assess both the absorption and desorption capacity. The difference in desorption efficiency reached its peak in the first 24 h and can represent the slow-release effects of NO₃⁻-N. According to the above results, although increasing the biochar application rate and NO₃⁻-N concentration can improve the soil absorption capacity, a mismatched combination of biochar application rate and NO₃⁻-N concentration does not consistently deliver satisfactory results. For instance, a high biochar application rate with deficient NO₃⁻-N generated low absorption capacity (8% biochar application rate with a NO₃⁻-N concentration of 80 mg·L⁻¹ had a lower absorption capacity than 4% biochar application rate with a NO₃⁻-N concentration of 240 mg·L⁻¹; Figure 4 and Figure 5). Similarly, for a given NO₃⁻-N supply, the absorption capacity varied considerably (a biochar supply of around 4%, absorption capacity ranged from 0.284 to 0.837 mg·g⁻¹ in Figure 4), depending on the biochar application rate. This confirms that the absorption capacity would be efficiently amplified at the appropriate biochar application rate and NO₃⁻-N concentration. More importantly, high NO₃⁻-N concentration significantly increased the desorption efficiency in the first 24 h, which conflicted with the goal of a slow-release effect. Therefore, it is necessary to optimize the management of biochar and NO₃⁻-N to maximize the input efficiency while satisfying both the absorption capacity and the slow-release effect.

The response function (the biochar application rate and initial concentration of NO₃⁻-N as independent variables and the absorption capacity/desorption efficiency in the first 24 h as response variables) was employed to optimize the biochar application rate and initial NO₃⁻-N concentration (

Table 3. Regression relationships of biochar application rate and NO₃⁻-N concentration with absorption capacity and desorption efficiency in the first 24 h.

Response Variable	Regression Equation	R2
Absorption capacity (Z1)	Z1 = −0.0189734 + 0.113089x − 0.00948437x2 + 0.00483395y + 0.00019284xy − 0.0000146292y2	0.9194
Desorption efficiency in the first 24 h (Z2)	Z2 = 46.54 − 7.709x + 0.1883y − 0.005531xy + 0.5297x2 − 0.0002832y2	0.9427

Note: x and y represent the biochar application rate (%) and initial NO₃⁻-N concentration (mg·L⁻¹), respectively.

). The coupling effects of biochar application rate and NO₃⁻-N concentration on absorption capacity exhibited a downward convex shape (Figure 10a). The maximum absorption capacity (0.974 mg·g⁻¹) was achieved when 8.19% of biochar and 219 mg·L⁻¹ NO₃⁻-N concentration was applied. However, the increasing biochar application rate and decreasing NO₃⁻-N concentration caused a continuous decrease in desorption efficiency in the first 24 h. The biochar application rate and NO₃⁻-N concentration corresponding to the theoretical minimum of desorption efficiency in the first 24 h were not practically significant (Figure 10b). To satisfy both the absorption capacity and the slow-release effect, 80% of the maximum absorption capacity (0.779 mg·g⁻¹) and 120% of the measured minimum desorption efficiency in the first 24 h (30%) as the boundary conditions to define the optimal biochar and NO₃⁻-N management. The results showed that the absorption capacity of 3.5–12.5% biochar application rate was 0.779 mg·g⁻¹ at NO₃⁻-N concentration ranging from 100 to 330 mg·L⁻¹. Converting NO₃⁻-N concentration to weight, the optimal ratio of biochar to NO₃⁻-N varied from 9 to 42.75 g·g⁻¹. Similarly, the calculation results for the desorption efficiency in the first 24 h were achieved at 1.5–20% biochar application rate with the NO₃⁻-N concentration ranging from 10 to 400 mg·L⁻¹, and the optimal ratio of biochar to NO₃⁻-N varied from 27 to 1080 g·g⁻¹. With careful consideration of various results, when the ratio of biochar to NO₃⁻-N was 34–42.75 g·g⁻¹, the maximum input efficiency, better nutrient supply, and slow-release effects can be obtained.

3.6. Regeneration

To investigate the reusability of biochar-soil mixture for NO₃⁻-N, the absorption and desorption behaviors were examined for seven consecutive cycles. The repetition cycle significantly affected the absorption capacity and desorption efficiency in the first 24 h (Figure 11). As the repetition cycle increased, the absorption capacity showed a downward trend, mainly due to the incomplete desorption sites. The decreasing rate gradually lessened with each repetition cycle, and stability in the absorption capacity was achieved after five cycles of NO₃⁻-N absorption and desorption. Compared with the first cycle, the absorption capacity of NO₃⁻-N dropped by about 18% in the seventh cycle. However, the desorption efficiency in the first 24 h was increased significantly with the rise of the repetition cycle. In the first, second, third, fourth, and fifth repetition cycles, the desorption efficiency in the first 24 h was 33%, 39.3%, 41%, 42.9%, and 43.3%, respectively. Similar to the absorption capacity, the desorption efficiency in the first 24 h essentially remained stable after the fifth cycle, ranging from 43% to 44%. Compared with the first cycle, the desorption efficiency in the first 24 h increased by about 10% in the seventh cycle. Generally, the absorption and desorption performance demonstrated that part of the absorption process was irreversible. After the fifth cycle, only reversible absorption was repeated in the NO₃⁻-N absorption and desorption process, and the absorption capacity and slow-released effect were not greatly reduced in the subsequent repetition cycles. These indicate that the reversible absorption is relatively stable, and biochar has excellent cycling properties for the NO₃⁻-N absorption.

4. Discussion

Numerous studies have explored the absorption and desorption of NO₃⁻-N by biochar [74], and the absorption capacity of NO₃⁻-N onto biochar varies greatly. Due to the net negative surface charge and insufficient AEC, some biochar has almost no NO₃⁻-N absorption capacity [29,75]. However, other research findings show contrasting results. For instance, Wang et al. [76] reported that the oak sawdust biochar had an absorption capacity of 2.8 mg·g⁻¹ for NO₃⁻-N and can reduce NO₃⁻-N leaching in soil [27]. Similarly, Han et al. [65] also claimed that the absorption capacity of date palm biochar for NO₃⁻-N ranged from 0.12 to 7.73 mg·g⁻¹. In our research, the absorption capacity of NO₃⁻-N onto biochar was 3.51 mg·g⁻¹. This may be due to a considerable specific surface area of the biochar used in our study and an average pore size of 5 nm, exceeding the Stokes ionic radius (0.129 nm) of nitrate and the hydration ionic radius (0.335 nm) [66], suggesting that NO₃⁻-N can penetrate into biochar pores [54]. In addition, the presence of metal ions in the biochar provided a positively charged absorption site, contributing to the absorption of NO₃⁻-N by electrostatic interaction [54,77]. However, according to Langmuir, the maximum absorption capacity of 6.09 mg·g⁻¹ for biochar in this study was not very impressive compared to some studies. Biochar made from palm leaf residues had a maximum NO₃⁻-N absorption capacity of 42.55 mg·g⁻¹ [54]. Alsewaileh et al. [70] showed that biochar made from date palm had a maximum NO₃⁻-N absorption capacity of 8.37 mg·g⁻¹. In contrast, the maximum NO₃⁻-N absorption capacity of biochar made from mustard straw and wheat straw was 1.3 mg·g⁻¹ and 1.1 mg·g⁻¹, respectively. Variations in maximum absorption capacity can be attributed to the diverse physicochemical properties arising from distinct raw materials and production temperatures [43,74,78]. In this study, the small maximum absorption capacity might be attributed to biochar’s alkaline pH and pH_zpc of 3.4 [54]. Generally, when the solution pH was lower than pH_zpc, the surface of the absorbent was protonated and more positively charged [79,80], which was more favorable for the absorption. However, in this study, the solution pH was higher than the pH_zpc, and the electrostatic repulsion between the negatively charged surface sites and NO₃⁻-N resulted in lower absorption [81]. In addition, the absorption capacity was also affected by absorption conditions. The absorption capacity was increased with the increasing contact time and initial NO₃⁻-N concentration (Figure 2 and Figure 3), which might relate to different mechanisms and limiting factors.

According to the R² and the calculated Q_e value from the pseudo-first and second-order models, both physical and chemical absorption were the mechanisms for NO₃⁻-N absorption [82]. Moreover, the intensity of absorption peak at 3420 cm⁻¹ and 1585 cm⁻¹ indicating that biochar formed hydrogen bonds or exchanged ions with NO₃⁻-N [81]. The absorption of NO₃⁻-N onto biochar involved three consecutive processes [83,84]: (i) film diffusion of the NO₃⁻-N to the biochar surface, (ii) intra-particular diffusion of NO₃⁻-N into biochar, and (iii) diffusion from macro to micropores. The last process was found to be very rapid in comparison to the other two [85]; therefore, the overall rate of the absorption process depended on the film and intra-particular diffusion processes [85]. Additionally, according to the curve of NO₃⁻-N absorption capacity plotted against T^0.5 did not pass through the origin (Figure 3b), both film and intraparticle diffusion control and limit the absorption system [85]. In summary, the mechanisms and controlling factors of NO₃⁻-N absorption by biochar include Van der Waals’ force, ion exchange, hydrogen-bond interaction, electrostatic interaction, film diffusion, and intraparticle diffusion (Figure 12).

Due to the ability to absorb NO₃⁻-N from solution, placing biochar into soil has also been shown to affect the leaching of NO₃⁻-N [24,78,86,87]. In this study, the results showed that biochar increased the soil absorption capacity (Figure 4), which was consistent with previous findings [25]. There might be two reasons for the increased absorption capacity [88]. On the one hand, biochar has a special composition structure, water-ion hydrogen bonding, physical properties, and higher AEC [86,89], which explains a strong absorption capacity for NO₃⁻-N [24,90]. In this study, the biochar had 12.1 times the absorption capacity of the soil for NO₃⁻-N (Figure 2 and Figure 4). On the other hand, biochar increased the soil porosity and promoted the formation of soil aggregates, thereby improving the ability to absorb and retain NO₃⁻-N, resulting in a slow-release desorption process of NO₃⁻-N from the soil and biochar mixture [11]. As NO₃⁻-N absorption by biochar was controlled by multiple processes [91,92], its desorption process was also expected to be similarly governed [10,93]. Different control mechanisms differed in the intensity of NO₃⁻-N absorption, leading to distinct desorption performance. In our study, as reported in earlier studies [94], the rapid desorption efficiency in the first 24 h was observed for NO₃⁻-N. This might be related to the weaker bindings and rapid diffusion into the aqueous solutions. Thereafter, the NO₃⁻-N desorption was probably governed by the solute diffusion from the porous biochar solid phase into the aqueous phase [95]. This diffusion-limited phase was expected to be slow, probably due to a tortuous porous flow path [94,96].

Notably, similar to pure biochar, the high NO₃⁻-N concentration could also increase the biochar-soil mixture absorption capacity (Figure 5). However, the high absorption capacity at a high concentration of NO₃⁻-N was not conducive to a slow-release effect. Treatments of 8% biochar application rate with 80 mg·L⁻¹ NO₃⁻-N concentration (Figure 4) had similar NO₃⁻-N absorption capacity (0.8 mg·g⁻¹) and desorption trends to the 4% biochar application rate with 160 mg·L⁻¹ NO₃⁻-N concentration (Figure 5). However, their desorption efficiency and capacity in the first 24 h were obviously different. The NO₃⁻-N desorption efficiency and capacity at 8% biochar application rate with 80 mg·L⁻¹ NO₃⁻-N concentration was 28% and 0.225 mg·g⁻¹ in the first 24 h, respectively (Figure 6 and Figure 7), while the 4% biochar application rate and 160 mg·L⁻¹ NO₃⁻-N concentration treatment was 48% and 0.384 mg·g⁻¹, respectively (Figure 8). This result enforced the hypothesis of NO₃⁻-N desorption from biochar limiting by multiple processes. The stronger binding sites were occupied first, and the binding strength decreased with the increasing degree of site occupation [43,97]. At higher NO₃⁻-N concentrations, the site occupancy was higher, and the absorption strength decreased, which explained the higher NO₃⁻-N desorption efficiency at high NO₃⁻-N concentrations in the first 24 h. Similarly, it was also observed that the absorption and desorption performance could be controlled by biochar application rate and NO₃⁻-N concentration (Figure 4 and Figure 6). Overall, these all affect biochar’s effectiveness and ultimately determine the optimal biochar amount.

The optimal biochar application rate was crucial for adopting biochar as a practice due to its availability and price constraints. Existing debates on biochar application rates in agricultural production systems recommended a range from 1 t·ha⁻¹ to 200 t·ha⁻¹ [39,98,99]. Positive responses to increasing yields in wheat, maize, rice, and soybeans have been observed for biochar application rates in the range of 1–10 t·ha⁻¹. However, excessive biochar could diminish the positive effects [39]. In the absence of N fertilizer, the application of biochar did not increase the radish yield. However, the radish yield evidently increased with biochar application in the presence of N fertilizer [100]. This implies that the effectiveness of biochar was also related to the amount of fertilizer used. In this study, integrating nutrient supply and slow-release effect, it was recommended to control the application ratio of biochar to NO₃⁻-N at 34–42.75 g·g⁻¹. On this basis, it maximized soil nutrient availability and enhanced the slow-release effect with reduced biochar inputs. Moreover, the perspective of this study was similar to soil testing and fertilizer recommendation; the biochar application rate was only related to the fertilization schedule and soil nutrients. It could greatly accelerate the promotion and application of biochar in different crops and soil conditions.

Due to its excellent absorption and desorption characteristics, biochar was considered a potential sustainable supporting material for the formulation of slow-release fertilizers [42,77]. It has been shown that biochar had a similar nutrient release pattern to Agroblen (conventional slow-release fertilizer) and could improve nutrient loss [92,101,102]. Moreover, biochar impregnated with fertilizer has been used to synthesize or manufacture slow-release fertilizer [37,103,104,105]. Dominguez et al. [102] reported that the impregnation of N and P fertilizer on oil palm kernel shell biochar improved its nutrient release performance. However, the slow-release effect of biochar-based fertilizer was still facing limitations [16]. Decreased performance in terms of preventing N release was demonstrated by biochar compared to the developed slow-release fertilizers. In this study, NO₃⁻-N desorption efficiency in the first 24 h ranged from 28% to 90% in all treatments, which was higher than the developed formulation [11]. This suggests that although biochar can slow down nutrient desorption efficiency, the unoptimized raw biochar application rate cannot be directly applied to the soil as a slow-release fertilizer material to meet commercial standards [106]. In addition, the absorption capacity of biochar for NO₃⁻-N was 3.51 mg·g⁻¹, lower than that of traditional slow-release fertilizers. In other words, when using biochar as a slow-release fertilizer, a large amount of biochar needed to ensure the growth of plants. Therefore, low N content and high input costs due to low N content need to be focused and solved when developing and applying biochar as slow-release fertilizer.

The nutrient release performance from biochar was related to ion binding capacity and absorption level [16]. Therefore, in addition to optimizing the biochar application rate, biochar modification for absorption capacity and strength has become an inevitable trend to enhance its slow-release effect [107]. Commonly used methods include chemical and physical modifications, with chemical modification being the most extensively employed method. It mainly included acid modification, oxidizing agent modification, carbonaceous materials modification, metal salts, and alkalinity modification. Acid modification could remove impurities such as metals, introduce acid functional groups on the surface, and change the biochar surface area [108]. Modification with oxidizing agents could increase the content of oxygen-containing functional groups on biochar [109]. Biochar modified with montmorillonite had an absorption capacity of up to 9 mg·g⁻¹ [110]. In addition, encapsulation technology combined with nutrient-impregnated biochar has been introduced to improve release performance [111]. Compared to original biochar-based fertilizer, encapsulated biochar-based fertilizer could further improve the slow-release performance [73,94]. Biochar developed by an integrated co-pyrolysis and co-polymerization process could also be used as a potential slow-release fertilizer [112]. These provide new approaches and possibilities for the development and utilization of biochar as a slow-release fertilizer carrier.

Compared with traditional slow-release fertilizers, the high price of biochar is another important factor restricting its application. While biochar has long-term effectiveness compared to traditional slow-release fertilizers, which might reduce its use costs [113]. MnFe-LDO-biochar maintained 80% of antibiotic removal after three consecutive cycles [114]. After seven repeated absorption and desorption of NO₃⁻-N, biochar still showed excellent absorption and slow release of NO₃⁻-N (Figure 11). Even as biochar becomes more hydrophilic during aging, aged biochar is expected to absorb more NO₃⁻-N compared to fresh biochar [115]. Overall, the long-term effectiveness of biochar might be an important breakthrough in the use of biochar as a slow-release carrier, with both economic and environmental benefits.

Our results highlight the potential benefits of biochar application in improving soil fertility, particularly in the application of slow-release material. However, considering the absence of plant limitations in this study, further field trial studies are needed to systematically understand the long-term interaction between biochar and fertilizer. Of particular interest is the stability and temporal effectiveness of biochar in field environments.

5. Conclusions

In this study, we investigated the kinetics study, isotherms study, and effect of biochar dosage and the NO₃⁻-N concentration on the absorption and desorption performance to optimize the biochar application dosage. The biochar from apple branch pyrolysis had an absorption capacity of 3.51 mg·g⁻¹ for NO₃⁻-N. The absorption process could be described by the pseudo-second-order model and Langmuir model. Both biochar application rate and NO₃⁻-N concentration influenced the absorption and desorption performance. High biochar rates contributed to a favorable slow-release effect, while low biochar rates contributed to a higher absorption capacity. Satisfactory nutrient supply and slow-release performance were obtained when the ratio of biochar to NO₃⁻-N was 34–42.75 g·g⁻¹. Compared to traditional slow-release fertilizers, biochar was more expensive, had a lower N content of 3.5%, and needed to be applied in large quantities for optimal plant growth, thus requiring significant supplementation costs. Notably, these increased expenses might be mitigated using biochar modification and its sustained effectiveness over time. Overall, combined with optimized biochar absorption and desorption techniques, biochar still has potential and promise as a carrier for slow-release fertilizers in commercial production.