Performance Evaluation of Modified Biochar as a Polycyclic Aromatic Hydrocarbon Adsorbent and Microbial-Immobilized Carrier

Abstract

Herein, biochars derived from corn stalks, rice husks, and bamboo powder were modified by nitric acid oxidation and sodium hydroxide alkali activation to identify efficient and cost-effective polycyclic aromatic hydrocarbon-adsorbent and microbial-immobilized carriers. The surface characterization and adsorption investigation results suggested that acid/alkali modification promoted the phenanthrene removal ability in an aqueous solution of biochars via facilitating π–π/n–π electron donor–acceptor interactions, electrostatic interactions, hydrogen bonds, and hydrophobic interactions. Subsequently, the degrading bacteria Rhodococcus sp. DG1 was successfully immobilized on the rice husk-derived biochar with nitric acid oxidation (RBO), which exhibited the maximum phenanthrene adsorption efficiency (3818.99 µg·g⁻¹), abundant surface functional groups, and a larger specific surface area (182.6 m²·g⁻¹) and pore volume (0.141 m³·g⁻¹). Degradation studies revealed that the microorganisms immobilized on RBO by the adsorption method yielded a significant phenanthrene removal rate of 80.15% after 30 days, which was 38.78% higher than that of the control. Conversely, the polymer gel network-based microenvironment in the microorganism-immobilized RBO by the combined adsorption–embedding method restricted the migration and diffusion of nutrients and pollutants in the reaction system. This study thus introduces an innovative modified biochar-based microbial immobilization technology characterized by a simple design, convenient operation, and high adsorption efficiency, offering valuable insights into material selection for PAH contamination bioremediation.

1. Introduction

Polycyclic aromatic hydrocarbons (PAHs) are representative persistent organic pollutants, which are widely recognized for their ubiquitous presence in the environment, being generally detected in aquatic environments (0.03–8310 μg·L⁻¹) [1], sediment (11–66 μg·g⁻¹) [2], and soil (0.05–543.36 μg·g⁻¹) [3]. PAHs exert serious threats to both natural ecosystems and human health because of their recalcitrant and toxic properties, including carcinogenicity, teratogenicity, and mutagenicity, which have attracted considerable attention [4]. Currently, there are various PAH remediation techniques (e.g., physical repair, chemical oxidation, bioremediation, and integrated approaches) available for removing PAHs. Physicochemical approaches are efficient, but they are typically expensive and can cause secondary pollution, which poses environmental health risks. In contrast, bioremediation is regarded as an eco-friendly and cost-effective alternative [5]. Among these strategies, immobilized microorganism (IM) technology has emerged as an effective and promising approach for remediating PAH contamination by enabling microorganisms to withstand harsh environments while mitigating competition from indigenous species, thereby enhancing the survival rates and metabolic activities of introduced degrading bacteria. In recent years, biochar-based microbial immobilization technology has become a global research hotspot in PAH remediation [6,7].

Biochar (BC) acts as a provider, supporter, and cooperator in biochar-immobilized microorganism (BM) applications [6]. It exhibits an exceptional pollutant adsorption capacity, thereby reducing the spatial distance between microorganisms and contaminants [8,9]. Factors such as the surface area, pore size distribution, surface functional groups of biochar, molecular structure of adsorbate, and specific interactions between the biochar surface and adsorbate can significantly influence the adsorption process of pollutants [10]. Key mechanisms governing this behavior include the hydrophobic distribution, pore filling, electron donor–acceptor interactions, hydrogen bonding, and hydrophobic interactions, etc. [9,11,12]. As both an adsorbent of pollutants and carrier of immobilized microorganisms, the selection of an appropriate biochar becomes paramount. The physiochemical properties of biochar vary with the types of feedstocks and preparation methods (pyrolysis, gasification, and hydrothermal carbonization, etc.). The modification of biochar can optimize its physiochemical properties, such as increasing the active functional group, degree of aromatization, and the specific surface area, as well as enriching the surface and pore structure, etc., ultimately improving the adsorption performance. Chemical modification (e.g., acid/alkali modification, oxidizing agent modification, metal salts or oxidizing agent modification, and carbonaceous material modification) is available to improve the adsorption performance of biochar [13]. Among them, acid oxidation effectively removes impurities such as metals on the surface of biochar and introduces carboxyl, phenol, carbonyl, and other oxygen-containing functions on the surface of biochar [14]. It can also change the microstructure of biochar and increase its surface area, thus improving the adsorption performance of microorganisms and PAHs [11,15]. In addition, alkali-activated biochar has better surface features, which can promote hydrogen bonding, hydrophobic, π–π, and electrostatic interactions between organic pollutants and biochar [16,17]. Acid/alkali modification is most widely used at present, which has the advantages of a simple design, convenient operation, and high adsorption efficiency and has practical application and popularization value [17,18,19]. Zhou et al. prepared oxygen-rich biochar derived from waste wood by sulfuric acid treatment and found that the biochar had a targeted adsorption effect on PAH removal, especially symmetrical PAHs [15]. Shin et al. suggested that the reinforced aromatic structure of NaOH-activated spent coffee waste biochar facilitated the π-π interaction [16]. Despite the importance of biochar modification, its application on immobilizing microorganisms has rarely been reported. Based on this, the fundamental knowledge of the adsorption behaviors of PAHs on modified biochar and its degradation behaviors after the immobilization of microorganisms remains incomplete.

Biochar has been reported to be a suitable carrier for microbes [8]. Biochar’s porous structure coupled with its large specific surface area and abundant functional groups promotes microorganism adhesion and growth while providing protective habitats conducive to bacterial survival. Adsorption and embedding are the most widely used methods for microbial immobilization in current research. The immobilized microorganisms produced via adsorption methods exist in powder form, which offers advantages such as a low cost combined with simplicity of preparation. Due to its small steric hindrance, it can maintain the microorganisms’ high activity, making them suitable for application in non-fluid media such as soil environments. However, microbial displacement could occur because the bond between the microbes and the carrier in adsorption-based immobilization is weak. Combining biochar adsorption with embedding can address this limitation since the grid structure of embedding can prevent the displacement of the microbes from the biochar [8]. Sodium alginate (SA) and polyvinyl alcohol (PVA) are frequently utilized crosslinking materials; notably, SA-PVA co-immobilization has garnered significant interest owing to its high mechanical strength and strong stability [20].

To sum up, the remediation of PAH contamination with biochar-immobilized microorganisms has received significant public interest. However, there remains a substantial gap regarding the use of modified biochar for microorganism immobilization. We hypothesize that the acid/alkali modification of biochar can improve its physicochemical properties, thereby enhancing its PAH removal performance and producing an excellent microbial immobilization carrier. Therefore, the primary objective of this study was to evaluate the feasibility of utilizing modified biochar as an adsorbent material and microbial immobilized carrier. The specific aims include the following: (i) to investigate the adsorption behaviors and mechanisms of pristine and acid/alkali-modified biochars derived from various raw materials toward PAHs, and (ii) to identify the degradation performance and mechanisms of modified biochar-immobilized degradation bacteria targeting PAHs.

2. Materials and Methods

2.1. Biochar Preparation and Chemical Reagents

Corn stalk (CS), rice husk (RH), and bamboo powder (BP) were utilized as raw stocks to produce biochar. With reference to the relevant literature reports and the pyrolysis conditions of commercial biochar production, the raw stocks were pyrolyzed at 500 °C for 4 h in a muffle furnace under a N₂ environment to produce biochar [5,21]. The produced biochar was referred to as CB, RB, and BB, respectively. The biochar was hereafter oven-dried at 60 °C for 24 h and passed through a 60-mesh sieve. Subsequently, all biochars were treated with 2 M nitric acid (HNO₃) or hydroxide sodium (NaOH) at a ratio of 20:1 (v/w) and stirred at 75 °C for 8 h. Then, the supernatant was removed by centrifugation. All modified biochars were then rinsed several times with deionized distilled water until neutral pH, oven-dried, and milled to <0.25 mm. The acid-oxidized biochars produced from CS, RH, and BP were named as CBO, RBO, and BBO, respectively. The alkali-activated biochars produced from CS, RH, and BP were named as CBA, RBA, and BBA, respectively. All biochar samples were stored in a desiccator prior to subsequent experiments.

The characterization methods of biochar samples were similar to the methods described in our other work [22]. Briefly, the surface morphology features of the biochars were investigated via a scanning electron microscope (SEM; Gemini 300, Zeiss, Oberkochen, Germany). The presence and abundance of surface functional groups on the biochars were identified by Fourier transform infrared spectroscopy (FTIR; Affinity-1, Shimadzu, Kyoto, Japan). The crystallographic structures of the biochars were determined using a powder X-ray diffractometer (XRD; X′ pert-Pro MPD, Panalytical, Almelo, The Netherlands). The total pore volume and specific surface areas of the biochars were measured from N₂ adsorption–desorption isotherms using an automatic analyzer (ASAP2460, Micromeritics, Atlanta, GA, USA). The elemental composition (i.e., C, H, and N) of the biochars was determined by an element analyzer (Vario EL cube, Hanau, Germany). The C/H ratios were used as surrogate indices to predict the aromaticity and polarity of the biochars.

Organic solvents used for the extraction and analysis of PAHs, including acetone, n-hexane, and dichloromethane, were of chromatographic grade and purchased from J.T. Baker Chemical Company, Phillipsburg, NJ, USA. Other chemicals, such as HNO₃, NaOH, and CaCl₂, were of analytical grade and were obtained from Beihua Fine Chemicals Co., Ltd., Shenzhen, China.

2.2. Bacterial Strain and Biochar-Immobilized Microorganism Preparation

The bacterial strain Rhodococcus sp. DG1 (China General Microbiological Culture Collection Center No. 18981) preserved in our lab was employed. It was isolated from PAH-contaminated soil via a classical shaken liquid medium enrichment method as described elsewhere [23]. The strain exhibited higher PAH tolerance and better PAH removal efficiency. Before use, the DG1 strain was incubated in LB medium for 12 h. The cells were collected by centrifugation at 4 °C and then washed with 0.85% saline three times and finally re-suspended in the same solution with an OD₆₀₀ value of 1.0.

Biochar having maximum adsorption efficiency of PAHs was selected as a carrier for bacteria immobilization [24]. Afterward, the biochar-immobilized microorganisms were synthesized by physical adsorption and combined adsorption–embedding methods. First, the biochar was mixed with the above bacterial suspension at a ratio of 1:10 (w/v) and shaken at 120 rpm under 30 °C for 24 h. After centrifuging the mixture, the precipitate was freeze-dried and hereafter referred to as ARB. Second, the biochar was mixed with the above bacterial suspension at a ratio of 1:10 (w/v) and shaken at 120 rpm under 30 °C for 2 h. Subsequently, the mixed solution of biochar and bacteria was mixed with polyvinyl alcohol (PVA, 8%, w/v) and sodium alginate (SA, 1%, w/v). The colloidal solution containing the bacteria was added to a 5% (w/v) sterile CaCl₂ and saturated boric acid solution by dripping slowly and then left for 12 h at 4 °C to complete gelation. The immobilized bacteria in biochar-gel beads were washed three times with normal saline and hereafter referred to as ERB. The fabrication procedure of the biochar-immobilized microorganisms is shown in Figure 1.

All operations were performed under strict aseptic conditions. The surface morphology of the BC and BM were investigated using SEM. According to the method proposed by Chen et al. [20], the pelletizing ratio was calculated for the convenience of later calculating the amount of free-living bacteria and immobilized bacterial agents. The count of viable bacteria in the BM was detected using the traditional plate counting method.

2.3. Batch Adsorption Experiments of Biochar in Aqueous Solution

2.3.1. Kinetic Adsorption Experiments

Phenanthrene (PHE) was selected as the model PAH compound in this study. Considering the concentrations of the wastewater and facilitating the study in the laboratory [25,26], a PHE concentration of 2 mg·L⁻¹ was chosen. Batch adsorption kinetics experiments of PAH onto various biochars were performed. In each test, 10 mg of each biochar was mixed with 20 mL of PHE solution in a 100 mL brown glass vial. The background solution contained 0.01 M CaCl₂ and 200 mg·L⁻¹ NaN₃ to maintain a constant ionic strength and to inhibit microbial activity, respectively [27]. After sealing, the samples were kept in the dark, shaken at 200 rpm under 25 °C, and then sampled at predetermined times (0.5, 1, 2, 4, 6, 9, 21.5, 24, 48, 72, 96, 120, 144, and 168 h) to ensure adsorption equilibrium. Subsequently, after centrifugation (4000 rpm, 10 min), the supernatant liquid was filtered through a 0.22 μm nylon millipore membrane, and the filtered solutions were pre-treated and then analyzed via high-performance liquid chromatography (HPLC, DIONEX, UltiMate 3000) for the residual PAH content. Compound separation was accomplished on an Eclipse XDB-C18 column using methanol–ultrapure water (90/10, v/v, 25 min, 1 mL/min) as the mobile phase. Peak detection was at 254 nm, and chromatography was performed at 25 °C. All experiments were conducted in triplicate. Controlled trials were conducted without biochar under otherwise identical conditions.

The amount of adsorbed contaminant was calculated using the following formula:

$$Q_t={\displaystyle \frac{{(C}_0-C_t)V}{m}},$$

(1)

where Q_t (mg·g⁻¹) is the content of PAH in biochar at time t (h), C₀ (mg·L⁻¹) and C_t (mg·L⁻¹) are the PAH concentration in liquid at initial and time t (h), respectively, V (L) is the volume of the solution, and m (g) is the mass of biochar.

The adsorption kinetics of PAHs were expressed in terms of the pseudo-first-order (PFO), pseudo-second-order (PSO), and intra-particle diffusion (IPD) equations, respectively:

$$Q_t=Q_e\left(1-e^{-k_{1}t}\right),(\mathrm{P}\mathrm{s}\mathrm{e}\mathrm{u}\mathrm{d}\mathrm{o}–\mathrm{f}\mathrm{i}\mathrm{r}\mathrm{s}\mathrm{t}–\mathrm{o}\mathrm{r}\mathrm{d}\mathrm{e}\mathrm{r})$$

(2)

$$Q_t={\displaystyle \frac{k_2{Q}_e^{2}t}{1+k_2{Q}_{e}t}},(\mathrm{P}\mathrm{s}\mathrm{e}\mathrm{u}\mathrm{d}\mathrm{o}–\mathrm{s}\mathrm{e}\mathrm{c}\mathrm{o}\mathrm{n}\mathrm{d}–\mathrm{o}\mathrm{r}\mathrm{d}\mathrm{e}\mathrm{r})$$

(3)

$$Q_t=k_3{t}^{0.5}+b,(\mathrm{I}\mathrm{n}\mathrm{t}\mathrm{r}\mathrm{a}–\mathrm{p}\mathrm{a}\mathrm{r}\mathrm{t}\mathrm{i}\mathrm{c}\mathrm{l}\mathrm{e}\mathrm{d}\mathrm{i}\mathrm{f}\mathrm{f}\mathrm{u}\mathrm{s}\mathrm{i}\mathrm{o}\mathrm{n})$$

(4)

where Q_e (mg·g⁻¹) is the adsorbed PAH amount per mass unit of biochar at equilibrium; Q_t (mg·g⁻¹) has the same aforementioned meaning; k₁ (1·h⁻¹), k₂ (g·mg⁻¹·h⁻¹), and k₃ (mg·g⁻¹·min^−0.5) are the adsorption rate constants of the PFO, PSO, and IPD models, respectively.

2.3.2. Isotherm Adsorption Experiments

Isotherm adsorption experiments were conducted using six different concentration levels of PAH (0, 0.2, 0.5, 1.0, 1.5, 2.0 mg·L⁻¹). These experiments were carried out using the same procedures used for the adsorption kinetics experiments as described above. All samples were shaken for 3 days to achieve adsorption equilibration. The adsorption data were fitted by using the Freundlich equation:

$$Q_e={ k}_4{C}_e^n$$

(5)

where k₄ and n are the Freundlich constants indicating the relative adsorption capacity and intensity and C_e (mg·L⁻¹) is the solution concentration of PAH at equilibrium.

2.4. Incubation Experiment

The prepared biochar-immobilized microorganisms were used to amend PAHs in contaminated soil, and the degradation performance and kinetics were compared. To prepare the PAH-contaminated soil, an acetone solution of PHE was added and mixed thoroughly into the collected soil. The soil properties after balancing for 3 months were pH 7.1; organic matter 3.54 g·kg⁻¹; total N 1.62 g·kg⁻¹; total P 1.13 g·kg⁻¹; clay 4.59%, silt 35.57%, and sand 59.84%; and PHE: 42 mg kg⁻¹. Subsequently, culture dishes were filled with 1000 g of PHE-contaminated soil that was amended with ARB (3%), ERB, and DG1 inoculant, separately. The additional amount of free strain DG1 and ERB was equivalent to that of ARB according to the calculated immobilization rate [20,24]. The soil without inoculant was used as the control experiment (CK). All treatments were performed in triplicate and incubated in the dark at 25 °C for 30 days. Sterilized deionized water was added to keep the soil moisture at 60% of the maximum water-holding capacity. Soil samples were collected regularly for the determination of residual PHE, and then the degradation removal rate and first-order degradation kinetics of PHE were calculated.

3. Results and Discussion

3.1. Characterization of Biochar

3.1.1. Basic Characteristics

The basic properties of the nine biochars presented varying degrees of differences (

Table 1. The basic characteristics of the biochar.

Biochar	pH	Specific Surface Area (m2·g−1)	Pore Volume (m3·g−1)	Elements
				Carbon (C, %)	Hydrogen (H, %)	Nitrogen (N, %)	H/C	C/N
RB 1	8.11	70.69	0.052	46.23	0.99	1.35	0.02	34.24
RBO 2	6.34	182.6	0.141	39.51	1.34	1.54	0.03	25.66
RBA 3	8.52	81.18	0.061	41.36	2.26	1.42	0.05	29.13
BB 4	8.25	43.42	0.045	57.29	2.08	1.35	0.04	42.44
BBO 5	4.56	161.0	0.168	52.96	1.97	1.88	0.04	28.17
BBA 6	9.24	60.58	0.049	54.08	2.11	1.72	0.04	31.44
CB 7	9.16	22.46	0.037	65.86	1.82	1.97	0.03	33.43
CBO 8	6.12	69.88	0.053	62.59	2.45	2.03	0.04	30.83
CBA 9	10.03	50.11	0.050	60.55	1.53	2.54	0.03	23.84

¹ Rice husk-derived biochar. ² Rice husk-derived biochar modified by acid oxidation. ³ Rice husk-derived biochar modified by alkali activation. ⁴ Bamboo powder-derived biochar. ⁵ Bamboo powder-derived biochar modified by acid oxidation. ⁶ Bamboo powder-derived biochar modified by alkali activation. ⁷ Corn stalk-derived biochar. ⁸ Corn stalk-derived biochar modified by acid oxidation. ⁹ Corn stalk-derived biochar modified by alkali activation.

). The pH of the original biochars was weakly alkaline, which is probably due to the separation of alkaline mineral elements contained in the biomass material during the pyrolysis process, leading to an increase in the pH value. After being modified by acid or alkali, the biochar demonstrated correspondingly weak acidic or weak alkaline properties, among which the pH value of BBO was the lowest (4.56). In addition, acid/alkali modification can increase the specific surface area and pore volume of biochar, especially for nitric acid-oxidized biochar. After acid modification, the specific surface area of RB increased from 70.69 m²·g⁻¹ to 182.6 m²·g⁻¹, and the specific surface area of BB increased from 43.42 m²·g⁻¹ to 161.0 m²·g⁻¹. This was similar to reports in the relevant literature, because activation methods such as nitric acid can remove carbonate and other mineral components from biochar to a large extent, thereby improving the specific surface area and pore volume of biochar [28]. Moreover, the content of C element in biochar materials ranged from 39.51% to 65.86%, and the C/N ratio ranged from 23.84 to 42.44. The contents of C and N elements in biochar are closely related to the growth of microorganisms, and the C/N ratio of 25:1 is considered to be an ideal condition conducive to the growth and metabolism of microorganisms.

3.1.2. Structural Characteristics

The FTIR results for the nine biochars are depicted in Figure 2a. The characteristic peaks at approximately 3391 cm⁻¹ corresponded to the stretching vibration of hydroxyl (–OH) groups [29]. The presence of the C=C and C=O aromatic ring stretching vibration absorption peaks near 1610 cm⁻¹ indicates the existence of carboxyl (vCOOH) or carbonyl (C=O) oxygen-containing functional groups on the surface of biochar, implying that biochars possess a structure in which aromatic carbon compounds play a central role, which can provide π-electrons [26]. It is noteworthy that the absorption peaks of aromatic ring C=C and C=O functional groups also appeared in the acid-modified biochar at 1715 cm⁻¹ and 1530 cm⁻¹, especially RBO, which indicates that the nitro-modified biochar contains more characteristic functional groups. For RB, the band centered at 1445 cm⁻¹ was assigned to the CO₃²⁻ characteristic peak of the mineral component carbonates in biochar [26,30]. The peak at 1378 cm⁻¹ that appeared in BBO was ascribed to the NO₃⁻ stretching vibration, indicating that there are some residual nitrates in nitro-activated biochar materials, which belonged to the absorbed nitrates during the activation process, and this phenomenon has also appeared in other similar studies [10]. The bands at 1085 cm⁻¹ and 1025 cm⁻¹ were associated with the O-Si-O stretching vibration, indicating the presence of SiO₂ in biochar, which was similar to the results of other studies [28]. In addition, a stretching vibration peak of C-H appears in the region between 800 and 600 cm⁻¹, indicating the presence of aromatic compounds and heteroaromatic compounds in biochar [31]. The tensile vibration peak of the C-N-C bond existed near 400 cm⁻¹ [15]. The FTIR results suggested that the chemical activation of nitric acid or sodium hydroxide changed the surface functional group composition of the original biochar, and the acid-modified biochar was richer in functional groups that provided the specific adsorption sites for PAHs, which may affect the adsorption and removal performance of biochar for pollutants. Moreover, these abundant surface functional groups provided the biochar with good buffering performance [29].

As displayed in the XRD patterns of biochar (Figure 2b), a diffraction peak of the nine biochar materials appeared at around 2θ = 23°, which was consistent with the (002) graphene crystal surface [26] and signified that the biochars had a certain crystalline-order degree of carbon materials [30]. Meanwhile, some sharp crystal diffraction characteristic peaks of crystalline substances such as quartz (SiO₂) and calcite (CaCO₃) appeared near 2θ = 20.92°, 21.96°, 26.66°, 29.30°, 36.07°, and 50.14° successively in the XRD patterns of the biochars, which is consistent with previous studies’ findings [10,28,30]. However, the presence of alkaline base ions such as CaCO₃ is not conducive to adsorption. The CaCO₃ content of RB after acid modification was significantly reduced, which was in agreement with the FTIR results. This phenomenon is consistent with previous studies and is mainly due to the fact that protons (H⁺) in the acid dissolve the solid-phase carbonate/calcite in the biochar, forming dissolved alkalinity/HCO₃⁻ and Ca²⁺ [28]. There are a large number of impurities such as calcite (CaCO₃) in biochar prepared by direct pyrolysis of raw materials, and acid modification can remove these impurities to a large extent. Therefore, the surface functional groups and the structure of BC were changed significantly by the acid activation process, mainly through the elimination of carbonates [26].

3.1.3. Morphological Characteristics

The SEM results (Figure 2c) indicated that each biochar showed a porous structure, and acid/base modification can significantly alter the morphology and structure of the pristine biochar surface. The surface of CB was seriously damaged, while the surface of other biochars was relatively complete. Especially for RBO, the pore structure was evenly distributed on the surface. These pores are the main structure for biochar to adsorb pollutants, nutrients, and microorganisms. The fine debris on the surface of biochar is the ash produced by the pyrolysis of biomass to produce biochar. This ash blocks the surface pores, and the spatial collapse of biochar with different pore structures after pyrolysis will also cause pore blockage. The ideal carrier for the application of immobilized microorganisms requires that the material should be porous and have a high specific surface area, so as to ensure that the surface pores of the immobilized surface are large enough to maintain the normal physiological metabolism of the bacteria but also to avoid the occurrence of excessive pores resulting in the leakage of degrading bacteria cells. At the same time, the more abundant the pore structure of the biochar materials is, the better its mass transfer permeability, and the metabolites and nutrient substrates in the degradation system can easily enter and exit freely, while the space for the excretion and release of products will not be restricted. The SEM analysis indicated that the pores of RBO were clear and complete, with a more large and dense internal pore structure, which can provide more sufficient growth space for bacteria and can effectively accommodate and immobilize more microbial cells, suggesting that it is a suitable carrier for bacteria immobilization. Based on the above results, it was determined that the activated BC had a better appearance than BC, which may be an important factor in its adsorption capacity.

3.2. Adsorption Performance of Biochar

3.2.1. Adsorption Kinetics

The adsorption kinetic curves of biochar are presented in Figure 3. The results indicated that compared with the original biochar, the adsorption properties of the biochar modified by nitric acid or sodium hydroxide were improved to different degrees, and the biochar modified by acid was the most obvious. The adsorption capacity of PHE by different biochars from high to low was RBO, BBO, BBA, RBA, CBO, CBA, BB, RB, and CB. The adsorption capacity of RBO for PHE was the highest (3819 µg·g⁻¹), which was in accordance with the results of the characteristics discussed above, such as the large specific surface area and abundant oxygen functional groups on the surface. Other studies have also demonstrated that the adsorption performance of biochar after acid treatment can be improved [15]. As shown in

Table 2. The fitting parameters for PHE adsorbed by the biochars.

Biochar	PFO			PSO			IPD
	Qe (µg·g−1)	k1 (h−1)	R2	Qe (µg·g−1)	k2 (g·µg−1·h−1)	R2	b (µg·g−1)	k3 (µg·g−1·h−0.5)	R2
RB	2604	0.438	0.848	2782	2.375 × 10−4	0.973	1763	103.2	0.607
RBO	3621	0.664	0.862	3819	2.819 × 10−4	0.988	2788	106.2	0.556
RBA	2998	0.232	0.780	3201	1.137 × 10−4	0.924	1567	165.6	0.814
BB	2602	0.363	0.723	2796	1.929 × 10−4	0.919	1597	126.3	0.795
BBO	3521	0.473	0.864	3754	1.915 × 10−4	0.979	2442	133.4	0.596
BBA	3166	0.397	0.966	3358	1.878 × 10−4	0.987	2139	118.1	0.490
CB	2481	0.442	0.671	2648	2.582 × 10−4	0.905	1665	103.9	0.747
CBO	2716	0.150	0.927	2960	6.862 × 10−5	0.978	1037	182.6	0.776
CBA	2755	0.298	0.837	2932	1.625 × 10−4	0.959	1654	127.4	0.659

, by comparing the curves and correlation coefficients (R²) of the PFO and PSO models, it was found that the PSO model can better fit the adsorption process of PHE from different biochars; the R² was greater than 90.5%, and the equilibrium absorption capacity obtained by the fitting was relatively close to the experimental value. Therefore, the adsorption of PHE by biochar materials in this study mainly depended on chemical adsorption, indicating that the number of adsorption sites contained on the surface of the biochar was the main factor affecting the adsorption effect. In addition, in order to further determine the rate control steps of PHE adsorption by biochar, the IPD model fitting showed that the correlation coefficient R² (0.490–0.814) was relatively low, and all intercept (b) values were non-zero (1037 µg·g⁻¹–2788 µg·g⁻¹), which indicated that intra-particle diffusion was not the only factor limiting the adsorption rate during the adsorption process and that other processes such as outer film diffusion also affect the adsorption process.

3.2.2. Adsorption Isotherm

The adsorption equilibrium isotherm is helpful in describing how adsorbate molecules are distributed between the liquid and solid phases at equilibrium, which can provide insight into the homogeneity and heterogeneity of the adsorbent surface [27]. Freundlich isotherm models were performed to fit the adsorption data. As shown in Figure 4, with the increase in the initial dose of PHE, the adsorption amount by biochar gradually increased, but the increase range tended to be smaller because with the increase in the PHE equilibrium concentration, the adsorption site on the surface of biochar gradually decreased, and the adsorption amount gradually reached saturation [32]. The PHE adsorption capacity of the nine biochar materials in this study was greatly different, which was consistent with the equilibrium adsorption capacity in the adsorption kinetics experiment. As shown in

Table 3. Isothermal model constants for the adsorption of PHE on biochar.

	RB	RBO	RBA	BB	BBO	BBA	CB	CBO	CBA
k4	99.79	216.4	115.4	101.3	399.8	241.0	107.4	56.89	76.08
n	0.542	0.512	0.536	0.524	0.388	0.435	0.460	0.580	0.530
R2	0.937	0.894	0.968	0.951	0.975	0.968	0.946	0.972	0.957

, the result indicated a much better fitting of the experimental data by the Freundlich model with R² values all greater than 89.4%. The model can accurately describe the adsorption process under experimental conditions, which shows that the adsorption of PHE on biochar mainly occurs on non-uniform surfaces, indicating a complex physicochemical adsorption process [26]. The constant n value obtained by the Freundlich isotherm model was less than 1, indicating that the adsorption process easily occurred in the nine biochar species, and the surface heterogeneity is strong. The n value of modified biochar decreased and the nonlinearity increased, indicating that acid-modified biochar is more prone to PHE adsorption. In the modification process, the carbon in biochar was transformed from an indeterminate form to a transition state, and then from a transition state to an aromatic state, and finally a stable chaotic layer state was formed; thus, the overall structure of the biochar was constantly changing. The adsorption was mainly based on weak interactions such as partition and van der Waals (VDW) forces of the original biochar and was transformed into a common adsorption mechanism mainly based on strong interactions such as electrostatic, chemical, and π–π bonds, supplemented by pore retention. Therefore, when the simple distribution action is no longer the dominant factor of adsorption, the nonlinearity of the adsorption is gradually enhanced.

3.2.3. Adsorption Mechanism

The adsorption of biochar included multiple mechanisms [27]. Biochar is highly aromatic and rich in π−π electrons, while PHE contains aromatic π electrons with which the π−π bond can be formed. Thus, PHE could adsorb strongly onto biochar through π−π electron donor–receptor interactions [19]. In this study, the reinforced aromatic structure of the activated biochar facilitated the π−π interactions [16]. In addition, crystal SiO₂ was present in biochar, and according to relevant literature reports, the n−π interaction between the aromatic structure of PHE and the Si-O-Si of biochar could also contribute to the PAH adsorption [10]. Meanwhile, electrostatic adsorption is another relatively strong adsorption mechanism that can be achieved through the variation in the pH value. Nitric acid activation makes the surface of biochar positively charged, which strengthens the electrostatic adsorption with negatively charged microorganisms. As verified by other research, hydrogen bonds are also a pivotal mechanism responsible for the adsorption and immobilization behavior of the PAHs on biochars [29]. The surface O-containing functional groups such as –OH and –COOH of biochar provide crucial sites for the adsorption of PAHs, which could combine into a stable bond [19]. The acid modification method in this study increased the number of functional groups of biochar, promoted the formation of ionic bonds, and enhanced the hydrogen bond and electrostatic adsorption capacity [33]. Additionally, biochar with strong molecular polarity has a high PAH adsorption ability due to the intermolecular dipole–dipole interactions [13,34]. Acid/base modification could increase the surface hydrophobicity of the biochar, exposing more hydrophobic sorption sites and thus achieving an increase in the adsorption capacity for PAHs [17]. Van der Waals (VDW) forces are weak intermolecular forces that can facilitate the adsorption process and play an important role in stabilizing the adsorbed PAH molecules onto the adsorbent [34]. In addition, biochar has an organic phase available for the distribution of PAHs and a developed pore structure, which can inhibit the escape of PAHs after adsorption [19,33]. Previous studies have also found that the adsorption strength of biochar is positively correlated with the surface area. Acid/alkali activation can increase the surface area and porosity of biochar and provide more active adsorption sites for organic pollutants, which is conducive to improving the adsorption capacity of the biochar. Furthermore, it was found that the acid treatment reduced the ash content of the biochar, and related studies reported that the ash could bind to organic matter through specific interactions, which played a negative role in the sorption of biochar [27].

In conclusion, as elucidated in Figure 5, the PHE adsorbed on the surfaces of BC via the π−π/n−π electron donor–acceptor interactions, electrostatic interactions, hydrogen bonds, hydrophobic interactions, and VDW forces, and then it entered into the pores of the biochar through a pore-filling process [32].

3.3. Degradation Performance of Biochar-Immobilized Microorganisms

3.3.1. Degradation Performance

RBO was ultimately selected for loading with bacteria because of its maximum adsorption efficiency. In addition, a relatively larger specific surface area and pore volume were observed in this biochar, which can provide a suitable habitat for microbial colonization and proliferation, indicating a high bacterial adsorption capability [35]. In this study, biochar has a strong ability to fix microorganisms. For ARB, the immobilization rate of biochar to microorganisms was 53%, which was similar to the maximum immobilization rate noted by others [36]. The cell count of strain DG1 attached to biochar was determined by the plate counting method, which was about 0.8 × 10⁸ CFU·g⁻¹, and further indicated that the degrading bacteria were effectively fixed on the biochar [37]. The diameter of ERB was about 3.0 mm, and similar studies have reported that the degradation performance of PAHs was better when the particle size was about 3 mm [20]. ERB has a regular shape, full grain, good texture, and strong mechanical strength (97%), indicating that it has good stability and is conducive to maintaining the activity of degrading bacteria for a long time.

The incubation experiment results showed that the degradation rate of PHE by the immobilized bacteria was significantly higher than that by the free bacteria (Figure 6). After 30 days, the degradation rates of PHE by ARB, ERB, and DG1 were 80.15%, 63.92%, and 41.37%, respectively. The degradation process was in good fit (R² > 92.36%) with the first-order kinetic model (

Table 4. First-order kinetic parameters for PHE degradation.

	Regression Equation	Degradation Constant (k, d−1)	Half-Life Time (T1/2, d)	Regression Coefficient
ARB	C = 49.21e−0.0671t	0.0671	10.33	0.9236
ERB	C = 49.15e−0.0409t	0.0409	16.95	0.9613
DG1	C = 49.12e−0.0212t	0.0212	32.70	0.9660

). The degradation half-life of free bacteria to PHE was 32.70 days. When ARB or ERB was used, the utilization rate of PHE was improved, and the degradation half-lives were reduced to 10.33 and 16.95 days, respectively. Thus, when the same amount of degrading bacteria DG1 was added, the immobilization technology sped up the start-up rate of the microbial degradation of PHE and enhanced the efficiency of the microbial removal of PHE compared with free bacteria. This may be due to the microenvironment constructed between the immobilized degrading bacteria and the carrier, which can shield the harsh external environment. In addition, biochar with a high adsorbability can adsorb organic molecular PHE and increase the concentration of microorganisms, so that the bacteria can directly use the adsorbed PHE, which causes the reaction to start quickly, thus accelerating the degradation and removal of PHE.

Moreover, the degradation performance of ARB was higher than that of ERB, mainly because the mass transfer rate of PHE was lower than the biodegradation rate in the gel pellets, and the carbon source of microorganisms could not be replenished in a timely manner, which would inhibit the bio-utilization of PHE. Therefore, the polymer gel network used to provide a microenvironment in the ERB prepared by the embedding method limited the migration and diffusion of nutrients and pollutants in the reaction system, which was more suitable for the remediation of fluid environmental media with a faster mass transfer rate such as wastewater. The ARB in powdered form prepared by the adsorption method can effectively degrade PAHs in both solid degradation systems because of its good mass transfer. For ARB’s SEM (Figure 2c), a uniform and dense bacterial layer was found on the biochar surface, which filled the biochar surface and pore entrance, indicating that the degrading strain DG1 was successfully attached and immobilized on the biochar surface [24]. Similar to other studies, the cells of degraded strains clustered and even formed biofilms on the carrier, indicating that bacterial colonization on the biochar followed biofilm formation patterns, such as attachment, adhesion, and proliferation [24,35]. This also confirmed that biochar RBO was a good material for immobilizing strain DG1.

3.3.2. Degradation Mechanism

Biochar, serving as a carrier for immobilized degrading bacteria, plays a crucial role in the process of pollutant remediation. As depicted in Figure 5, biochar provides a protective habitat for the degrading bacteria DG1, shielding them from adverse soil conditions and competition with indigenous microorganisms. This facilitates the adaptability of functional strains to polluted environments, thereby enhancing their capacity to efficiently degrade pollutants in soil. Biochar can also function as a leavening agent, enhancing soil porosity and improving the mass transfer efficiency of water, oxygen, and nutrients within the soil [38]. Moreover, the degradation efficiency of biochar-immobilized microorganisms on pollutants is significantly influenced by the adsorption capacity of the immobilized carrier. Biochar facilitates the adsorption and accumulation of PHE in soil through various physicochemical interactions [34,39]. This process effectively reduces the spatial distance between the degrading bacteria and PHE, mitigates mass transfer limitations, and enhances contact between the degrading microorganisms and pollutants, thereby achieving superior degradation outcomes [40]. Additionally, the adsorption properties of biochar can diminish the bioaccessibility of highly toxic organic contaminants, thus safeguarding the microenvironment of indigenous microorganisms [41]. Furthermore, in the process of degrading pollutants, biochar provides C, N, P, and other nutrient elements for the growth of degrading bacteria DG1 and indigenous microorganisms, further promoting the degradation of PHE by microorganisms [42]. The porous structure of biochar provides a pathway for the diffusion of pollutants and degradation metabolites and provides enough space and oxygen for microorganisms to maintain their normal metabolism. Meanwhile, the enzymes secreted by the fixed degrading bacteria can penetrate into the pores of the biochar and interact with the adsorbed PHE to further degrade it. To sum up, biochar has a complex and diverse interaction with degrading microorganisms, which can affect the adsorption and degradation process of pollutants [43]. The adsorption–degradation process is an important mechanism for the enhanced remediation of pollutants by immobilized bacteria. Giving full play to the synergistic effect of adsorption and degradation with biochar and microorganisms can effectively improve the remediation effect of contaminated soil.

3.4. Implications

Biochars derived from CS, RH, and BP were chemically modified using HNO₃ and NaOH. The modified biochars demonstrated a superior adsorption affinity to the pristine biochars for PAH attributed to the facilitated π–π/n–π interactions, electrostatic interactions, hydrogen bonding, and hydrophobic interactions. Compared with biochar-based material for PHE removal in other studies (

Table 5. Removal efficiency of phenanthrene by different biochar-based adsorbents reported in previous studies.

Adsorbent	System	Removal Efficiency	References
Arundo donax-derived biochar-loading copper ions	Constructed wetlands with modified biochar as substrates, the influent phenanthrene was 12.8 ± 1.7 mg·L−1, hydraulic retention time lasted 3 days	94.09%	[44]
Magnetically modified rice husk biochar by a hydrothermal method	Phenanthrene concentration 5–70 mg·L−1, shaken at 150 r·min−1, under 25 °C for 24 h	97.6 mg·g−1/greater than 85%	[32]
Rice straw/wood/bamboo-derived biochar modified by NaOH and washed by HCl	Phenanthrene concentration 0.1–1 mg·L−1, shaken in dark at 130 r·min−1 overnight	42.9 mg·g−1	[17]
Willow-derived biochar based on volatile fatty acids	Phenanthrene concentration 0.21 mg·L−1, shaken at 120 r·min−1 under ambient temperature in dark	93%	[45]
Phragmites australis-derived biochar produced by ammoniation-hydrothermal method	Phenanthrene concentration 2 mg·L−1, shaken in dark at 25 °C under 125 ± 5 r·min−1	1.97 mg·g−1	[26]
Rice husk-derived biochar with nitric acid oxidation	Phenanthrene concentration 2 mg·L−1, shaken in dark at 200 r·min−1 under 25 °C	38.2 mg·g−1/95.47%	This study

), the modified biochar developed in this study exhibits better adsorption performance, higher cost-effectiveness, and can be popularized and applied. In addition, a modified biochar-immobilized microorganism system with high PHE removal capability was successfully synthesized. This immobilization technique accelerated the start-up rate of PAH biodegradation compared to free bacteria. The polymer gel network-based microenvironment in the combined adsorption–embedding immobilized bacteria restricted the migration and diffusion of nutrients and pollutants in the soil environment, making it particularly suitable for aquatic applications. Meanwhile, the powder form of adsorption-immobilized bacteria, with its superior mass transfer performance, is ideal for soil remediation. Compared with conventional PAH removal methods, the modified biochar-immobilized microorganism method is a more efficient, economical, and environmentally friendly technology. When applied to soil, biochar can have an impact on the soil microenvironment, including the soil physicochemical properties, enzyme activity, etc., as well as the activity of soil microorganisms. However, the introduction of degrading strains may inevitably disturb the indigenous microbial community, and the stability and reusability also need to be further investigated. These factors could potentially limit its sustainable development in practical applications within complex and diverse actual environments. In summary, this study presents an innovative modified biochar-based microbial immobilization technology characterized by its simplicity, ease of operation, and high adsorption efficiency, providing valuable insights for PAH remediation strategies.

4. Conclusions

In this study, the HNO₃ oxidation and NaOH activation treatment were used to chemically modify the biochars derived from CS, RH, and BP. The acid- and alkali-modified biochars exhibited enhanced adsorption capacities for PHE ranging from 2865.982 to 3818.99 µg·g⁻¹, compared to the pristine biochars that showed lower capacities of 2729.49 to 2978.058 µg·g⁻¹, attributed to significant improvements in their physicochemical properties such as increased specific surface area, pore volume, and aromaticity. Kinetic parameter calculations indicated that chemisorption was the predominant mechanism governing PHE adsorption onto the biochars. The adsorption mechanisms involved π–π/n–π electron donor–acceptor interactions, electrostatic forces, hydrogen bonding, hydrophobic interactions, VDW, and pore-filling effects. Notably, the RBO, which demonstrated the maximum adsorption efficiency (3818.99 µg·g⁻¹), a relatively larger specific surface area (182.6 m²·g⁻¹) and pore volume (0.141 m³·g⁻¹), and abundant surface functional groups, was ultimately selected as a carrier for bacterial immobilization and exhibited effective immobilization performance. The powdered ARB synthesized via adsorption achieved superior degradation efficacy with a PHE removal rate reaching 80.15% within 30 days; conversely, the spherical ERB struggled to attain satisfactory degradation performance in soil media due to restricted mass transfer capabilities. The combined adsorption–degradation process emerged as the primary mechanism facilitating enhanced remediation of pollutants. These findings provide compelling evidence that the nitric acid-oxidized biochar-based microbial immobilization technology represents a promising strategy for the effective removal of PHE contamination. Thus, a green and cost-effective technology for PAH contamination remediation is provided.