Toward Sustainable Soil Remediation: Progress and Perspectives on Biochar-Activated Persulfate Oxidation

Abstract

Organic soil pollution poses a persistent threat to environmental sustainability by disrupting nutrient cycling and ecosystem functioning. The biochar-activated persulfate (PS)-based advanced oxidation process (AOP) has emerged as a promising strategy for the sustainable remediation of organic-contaminated soils. This review provides a comprehensive overview of the recent progress in the PS-based degradation of organic pollutants, with a particular focus on the role of biochar as an efficient and environmental activator. This review further summarizes advancements in the design of modified biochars, including metal (Fe, Cu, Co, Mn, Zn, and La), non-metal (N, S, B, P), and functional group modifications, aimed at enhancing the PS activation efficiency while minimizing secondary environmental risks. Importantly, the overlooked contributions of soil microorganisms in PS/biochar systems are discussed, highlighting their potential to complement chemical oxidation and contribute to eco-compatible remediation pathways. This review emphasizes the sustainability-oriented evolution of PS/biochar technology, highlighting the importance of a cost-efficient implementation, ecological compatibility, and the rational engineering of smart, regenerable catalysts. These insights support the advancement of PS/biochar-based AOPs toward scalable, intelligent, and environmentally sustainable soil remediation.

1. Introduction

The soil environment is indispensable to the global ecosystem and plays a crucial role in nutrient cycling and energy flow [1]. The rapid urbanization and industrialization of modern society have resulted in a large number of pollutants in different types of soil. There are over five million contaminated soil sites (approximately 20 million hectares) worldwide at the current time [2]. These pollutants affect the functionality of the soil, inhibit the harvest of crops, and seriously accumulate in crops, thereby entering the food chain, posing a significant threat to the global ecosystem and further affecting human health [3,4]. For these reasons, the remediation of contaminated soil is now a globally common task and should be performed urgently. The existing research used common methods to treat contaminated soil, like ion exchange, precipitation, membrane filtration, and coagulation, and all of these methods are limited by exorbitant prices and are non-ecofriendly [5]. Thus, reliable alternative methods to address soil pollution efficiently in an environmentally sound manner are worth pursuing.

Persulfate (PS)-based advanced oxidation processes (AOPs) can generate sulfate radicals with a high redox potential (2.5–3.1 V), long half-life (30–40 µs), and wide operation pH (2–8), showing efficient removal efficiencies for refractory organic pollutants [6]. However, the oxidative ability of PS is relatively low. Interestingly, the peroxyl bonds within PS are easily cleaved by catalysts or external energy, thus generating abundant reactive radicals, including sulfate radicals (SO₄•⁻) and hydroxyl radicals (•OH). PS could be classified into peroxymonosulfate (PMS and HSO₅⁻) and peroxydisulfate (PDS and S₂O₈²⁻). Especially, the asymmetric PMS is easier to activate compared to the symmetric PDS [7]. Compared with PMS, PDS attracts attention due to its lower price and longer environmental retention time [8]. It is important to emphasize that the use of PS-based AOPs requires activators, such as biochar.

In recent years, researchers have tried to apply the PS-based AOP activated by biochar from wastewater treatment to soil remediation. For example, Wan et al. [9] synthesized a peanut shell-derived biochar loaded with nanoscale zero-valent iron (nZVI), which shows an obvious synergistic effect on the removal of nitrochlorobenzene (NCB) compared with the sole treatment of biochar or PDS. Under an optimal biochar dosage at 8 mg/kg, the removal of m-NCB, o-NCB, and p-NCB from soils was 98.71%, 95.25%, and 94.57% within 60 min, respectively. Ke et al. [10] also proved that a humic acid (HA)-modified biochar combined with PS could remove polycyclic aromatic hydrocarbons (PAHs) ranging from 60.7% to 100% during practical applications in soil remediation. Despite the progress, the PS-based AOP activated by biochar in soil has still not been well reviewed. In addition, huge advancements have been achieved during wastewater treatment using a PS-based AOP activated by biochar [11,12]. However, differently from wastewater treatment, the function of microorganisms plays a pivotal role during various organic degradation processes in soil remediation [13]. Thus, the role of microorganisms in the PS-based AOP activated by biochar during soil remediation requires a deep understanding, particularly regarding their responses to oxidative stress, shifts in the community composition, and their contribution to contaminant mineralization.

In this review, we systematically summarize the current research progress regarding PMS and PDS, as well as the PS activated by biochar during soil remediation. The advancement of a combined PS activated by biochar during wastewater treatment was also reviewed, which provided new insights for soil remediation. In addition, the functionality of microorganisms during soil remediation under the regulation of a combined PS activated by biochar was deeply discerned. This review offers new insights into soil remediation.

2. Soil Remediation by PS

PMS, also known as Caro’s acid, is commonly used with KHSO₅ (white powder). The PMS ion (HSO₅⁻) is a derivative of hydrogen peroxide (H₂O₂), where one hydrogen atom is replaced by a SO₃ group, thus sharing some features and characteristics of chemical reactions with H₂O₂. PMS is unstable at a pH < 6 and pH = 12, and it readily hydrolyzes to hydrogen peroxide at a pH < 1. In a non-activated state, the oxidizing ability of PMS is stronger than H₂O₂ but slightly weaker than PDS [14,15].

Both PMS and PDS contain O–O bonds in their molecular structures. However, PMS has an asymmetric molecular structure, while PDS is symmetric. This structural difference makes PDS more stable than PMS, thus requiring a higher input energy to break its O–O bonds and generate radicals. Compared to PDS, PMS is relatively easier to activate due to differences in its molecular structure and bond dissociation energy. After activation, PMS produces a more diverse range of reactive species. In the catalytic activation of PMS and PDS, PMS not only reduces the consumption of the activator but also lowers the environmental risks post-remediation [16].

2.1. Progress of PMS in Soil Remediation

As summarized in

Table 1. Research progress on soil remediation by PMS.

Pollutant	Pollutant Concentration (mg/kg)	Activator	Removal (%)	Process Descriptions	Reference
Oil	78–99	CoOOH	88.3	Rapid Co2+/Co3+ redox cycle and CoOH formation improved the continuous generation of ROS (SO4•−, O2•−, and 1O2).	Lyu et al. [17]
TPH	6625	None	32.8	Blank without activator.	Bajagain and Jeong [18]
		Fe0	43.3	Using reduced iron.
		Fe2+	39.7	Using FeSO4.
		Co2+	40.4	Using CoCl2.
		nZVI	61.1	The efficient catalyst of ZVI was attributed to its small size and large surface area, providing more reactive sites for the oxidation reaction.
		nZVI	>96.0	By five serial treatments of 3% PMS +0.2% nZVI.
TCE	100	None	95.3	Unactivated PMS can degrade TCE in soil, possessing a negligible impact on the particle size distribution and soil texture.	Oba et al. [19]
TCS	535	Trimetallic electrode	66.0	Co2+ in trimetallic oxidation electrode activated PMS to produce SO4•− coupled with an electrokinetic geo-oxidation system.	Yuan et al. [20]
DDT	7565	None	>95.0	Minerals in the soil participated in the soil remediation during the ball milling process, probably through non-radical ways rather than ROS oxidation.	Xu et al. [21]
CPF	100	Microwave	>90.0	The Fe(II)/Fe(III) oxidation–reduction cycle caused SO4•− generation, and reactive metastable heating pad waste-PMS caused electron transfer.	Shang et al. [22]
PAHs	692	None	24.5–82.8	Producing O2•−, SO4•−, •HO, 1O2, and SO5•− for soil remediation.	Zhou et al. [26]
	692	Electrokinetic	14.7–34.1	During the electrokinetic-enhanced process, the more rings the PAH had, the more difficult to remove.
	946.1	Amorphous FH	77.8–94.7	Radicals •OH and SO4•−, as well as non-radicals (1O2 and Fe(IV)=O), participated in the soil remediation.	Tang et al. [23]
	100	None	72.5	Blank without activator.	Zeng et al. [27]
		nZVI	79.9	HA and HA-like reductive compounds in soil play a vital role during Fe(II) and Fe(III) cycles, affecting the generation of ROS.
		nZVI+CA	96.8	CA could promote the desorption of PAH from the soil medium.
	69.4	N-CG	71.3–97.0	Pyridinic and graphitic N were speculated to be the reactive sites for PMS activation.	Liang et al. [24]
Bap	79.9	CG	17.2	Using coal gangue to calcine.	Li et al. [28]
		Ca–N	21.7	Using anhydrous calcium chloride and melamine to calcine.
		CG-N	27.9	Using coal gangue and melamine to calcine.
		CG-Ca	60.9	Using coal gangue and anhydrous calcium chloride to calcine.
		CG-Ca-N1/8	100	Using coal gangue, anhydrous calcium chloride, and melamine to calcine. And CG-Ca-N1/8 induced in the production of •OH, SO4•−, •O2− and 1O2.
STZ	50	None	96.5	1O2 was the predominant ROS that contributed to the soil remediation.	Zhang et al. [25]

, recent studies have demonstrated the versatility of PMS in degrading a wide range of organic pollutants in contaminated soils. For petroleum hydrocarbons, PMS activated by CoOOH [17] and nZVI [18] achieved removal efficiencies of 88.3% and 61.2%, respectively, primarily through the generation of reactive oxygen species (ROS). Halogenated organics, such as trichloroethylene (TCE) [19], triclosan (TCS) [20], Dichloro-Diphenyl-Trichloroethane (DDT) [21], and chlorpyrifos (CPF) [22], were effectively degraded via both radical (e.g., SO₄•⁻, •OH) and non-radical (e.g., ¹O₂) pathways, with degradation efficiencies exceeding 70–95% under optimized conditions. For polycyclic aromatic hydrocarbons (PAHs), amorphous porous iron material (FH) [23] and N-doped coal gangue (N-CG) [24] were shown to activate PMS and efficiently degrade compounds like pyrene and benzo[a]pyrene (BaP) via non-radical species such as Fe(IV)=O and ¹O₂. Additionally, for pharmaceutical pollutants such as sulfathiazole (STZ), PMS achieved a 96.5% removal within 60 min, with minimal disturbances to the soil structure [25]. These findings collectively highlight PMS’s broad applicability and diverse activation mechanisms for the remediation of various organic contaminants in soil.

2.2. Progress of PDS in Soil Remediation

As summarized in

Table 2. Research progress on soil remediation by PDS.

Pollutant	Pollutant Concentration (mg/kg)	Activator	Removal (%)	Process Descriptions	Reference
TPHP	50	TP	10.8–58.6	The interaction between Fe-minerals in soil and TP accelerated ROS generation for TPHP degradation, triggering the activation of PDS by accelerating the Fe(III) ↔ Fe(II)redox cycle.	Dong et al. [29]
		H2A	96.80	High Fe-minerals (e.g., α-Fe2O3) content enhanced PDS activation, while high SOM content inhibits TPHP degradation by consuming ROS.	Dong et al. [30]
NAP	48.4	None	46.5	Soil minerals and SOM could activate PDS to generate ROS (SO4•−) to degrade NAP.	Feng et al. [31]
		CAT	71.8	AA/CAT could promote the activation of PDS by Fe.
		AA	95.3	AA can activate PDS to produce ascorbate free radicals (AscH•−), which further transfer electrons to PDS and generate SO4•−.
		HA	>42.7	Modified HA with blocked Ar-OH and/or -COOH groups proved their importance in HA’s complexation and reduction capabilities.
SMX	20	nZVI	86.5–96.1	The reactive species during the degradation of SMX was radical •OH.	Zhou et al. [32]
CPF	100	None	20	Blank without activator.	Shang et al. [33]
		Microwave	36	Microwave at 60 °C.
		Microwave	85	Microwave at 80 °C. Radicals SO4•− and •OH were produced under microwave irradiation. The collisions between oxidants and soil facilitated the degradation of CPF through the thermal effects and non-thermal effects of microwaves.
MCB	95.6	Ball-milled pyrite	66.1–93.8	FeIV, SO4•−, and •OH were the main radicals produced in the ball-milled pyrite and PDS system. Ball milling restrained the formation of the passivation layer and increased the Fe utilization.	Qiu et al. [34]
HCB	200	CaO	80.0	Calculated radical •OH content was almost three times that of SO4•−, suggesting a dominant role of radical •OH.	Fan et al. [35]
PAHs	100	None	65.5	Blank without activator.	Zeng et al. [27]
		nZVI	72.6	HA and HA-like reductive compounds in soil play a vital role during Fe(II) and Fe(III) cycles, affecting the generation of ROS.
		nZVI+CA	93.5	CA could promote the desorption of PAH (naphthalene) from the soil medium.
	692	None	32.8–78.5	Producing O2•−, SO4•−, •HO, 1O2, and SO5•− for soil remediation.	Zhou et al. [26]
	40	None	2.5	Blank without activator.	Wang et al. [36]
		Fe3+	7.5	Weak activation of PDS by Fe3+.
		HA-1	30.0	Chemically modified HAs through ethylation (blocking -COOH and Ar-OH).
		HA-2	34.0	Chemically modified HAs through hydrolysis (blocking Ar-OH).
		HA-3	42.7	Chemically modified HAs through amination (blocking -COOH).
	17.0	nZVI	82.2	Micro/nanostructured ZVI (nZVI).	Song et al. [37]
		C-nZVI	62.8	Stearic-coated micro/nanostructured ZVI (C-nZVI).
		M-nZVI	69.1	Commercial micron-sized ZVI (mZVI).
DDT	7565	HA	>42.7	Modified HA with blocked Ar-OH and/or -COOH groups proved their importance in HA’s complexation and reduction capabilities.	Xu et al. [21]
STZ	50	None	22.9	PDS is not easily activated.	Zhang et al. [25]

, PDS has also shown great potential for degrading various organic pollutants in soil through diverse activation strategies. For flame retardants, such as triphenyl phosphate (TPHP), chelating and reducing agents (CRs), like tea polyphenols and ascorbic acid, significantly enhanced the PDS activation by promoting Fe(II) release and surface oxygen vacancies, leading to an efficient ROS generation [29,30]. Pharmaceuticals and pesticides including naproxen (NAP), sulfamethoxazole (SMX), and chlorpyrifos (CPF) were effectively degraded using PDS systems activated by L-ascorbic acid (AA), (+)-catechin hydrate (CAT), nZVI, or microwave irradiation, achieving removal efficiencies above 85% in most cases [31,32,33]. For halogenated organics like monochlorobenzene (MCB), an advanced activation via ball-milled pyrite or mechanochemical PDS treatments achieved a >85% removal, involving sulfur vacancies and Fe(IV) species [34,35]. In the case of PAH-contaminated soils, naphthalene was effectively removed by PDS activated with nZVI, citric acid, humic acid, or Fe³⁺, with reactive species (SO₄•⁻, •OH, ¹O₂) contributing dominantly to the degradation [27,36]. These studies collectively demonstrate the versatility of PDS in soil remediation, although the optimization of activators and the mitigation of the ROS competition remain important considerations.

2.3. Opinions on the Potential of PMS and PDS in Soil Remediation

PMS and PDS are both widely employed for various organic-polluted soil remediation processes (Table 1 and Table 2), offering distinct advantages and limitations [38]. Shang et al. [33] investigated the combination of microwave and PMS or PDS for the degradation of CPF, where the microwave/PMS system was more preferable. In the microwave/PDS system, the microwave provided a thermal effect and a special non-thermal effect to produce SO₄•⁻ and •OH. By comparison, the microwave mainly produced a thermal effect to promote the production of ¹O₂ and the degradation of CPF in the microwave/PMS system. PMS exhibits a high reactivity and a rapid degradation rate [39], making it particularly suitable for emergency scenarios involving complex contamination and the efficient removal of persistent organic pollutants. However, its lower environmental stability leads to a short-lived radical generation period [7]. Additionally, its higher cost constrains large-scale, long-term applications. PDS demonstrates a superior chemical stability, allowing for prolonged activity in environmental settings [7]. Its stability facilitates easier storage and transportation. In addition, the relatively slow reaction rate may limit its application in certain contexts. In summary, PMS is better suited for short-term and rapid remediation projects, especially for tackling complex contamination scenarios, whereas PDS is more appropriate for sustained, long-term remediation, including low-permeability soils.

Furthermore, the preferred activation methods are also different for PMS (electron-transfer-based activation, such as catalysis with transition metals and nanocarbons) and PDS (energy-transfer-based activation, such as thermolysis and photolysis) [38]. In practical soil remediation, the applicability of the PS activation by thermolysis and photolysis may be lower than the catalytic methods using transition metals and nanocarbon, potentially making PMS a more favorable option. Nevertheless, many studies have proven the visibility of PDS activation by various exogenous or endogenous activators, such as metals and carbons (Table 2). Importantly, the prices of PDS and PMS are 0.74 USD/kg and 2.2 USD/kg, making them a priority in large-scale remediation projects [40]. The selection between PMS and PDS should be based on an integrated assessment of contaminant characteristics, soil properties, remediation goals, and economic considerations. In some cases, the combined use of these oxidants or their integration with complementary remediation technologies may enhance the overall efficiency and environmental compatibility.

3. Soil Remediation by PS Activated by Biochar

3.1. Pristine Biochar

Biochar, with a large surface area and oxygen-containing functional groups (e.g., carboxyl, hydroxyl, and aromatic structures), can adsorb pollutants [41]. In addition, biochar has also been demonstrated to activate PS by the radical-initiated process to generate •OH [42] and SO₄•⁻ [43]. Biochar could further accelerate electron transfer or surface-bound free radicals [44] for PS activation to accelerate pollutant decomposition. In recent years, progress has been made in soil remediation by PS activated by biochar (Table 3). Biochar is rich in functional groups that can act as electron donors to activate PS via an electron transfer process [45]. Xue et al. [46] used sludge-derived biochar to activate PDS for atrazine (ATZ)-polluted soil’s remediation, in which 95.3% of the ATZ was removed. Dou et al. [47] supplemented low-molecular-weight organic acids (LMWOAs) in anaerobic biochar-activated PDS systems and found that the degradation of 12 γ-hexachlorocyclohexanes (HCH) reached 99%. An increase in α-hydroxyl groups of LMWOAs facilitated the formation of reductive carboxyl anion radicals (COO•⁻) via an electrophilic attack by SO₄•⁻/•OH. Liu et al. [48] used biochar to activate PS for the remediation of bisphenol A (BPA)-polluted soil, achieving a high removal efficiency of 98.4%. Chen et al. [49] investigated the effect of biochar additions on the activation of PDS on the remediation of SMX-polluted soil. The removal efficiency could reach 90.7%, in which SO₄•⁻ and •OH were the predominant reactive species. Masud et al. [50] modified biochar from kelp seaweed by a ball-milled process and used it for PMS activation for the degradation of ciprofloxacin (CIP), achieving an excellent performance at 96.1%. Zhao et al. [51] used biochar combined with MW to activate PDS for the remediation of ethyl-parathion (PTH)-polluted soil, achieving an 88.8% removal efficiency within 80 min. The decomposition of PDS into SO₄•⁻, •OH, O₂•⁻, and ¹O₂ contributed to the PTH degradation, which oxidized PTH into paraoxon, 4-ethylphenol, and hydroquinone.

3.2. Biochar Loaded with Fe

Fe, a widely distributed and studied metal, is commonly used for the activation of PS due to the advantages of its environmentally friendly reaction, low cost, and high reaction rate. ZVI, especially nZVI, could improve the efficiency of the PS activation by providing a higher surface area and a continuous source of Fe (Figure 1) [52]. However, as a nanoscale material, nZVI is prone to aggregation and oxidation during application. To improve the stability of nZVI, its loading on biochar can effectively address the above challenges [53]. In addition, FeS could effectively activate PS and reduce the surface passivation of particles, which is a potential alternative coating for the adhesion onto the surface of biochar [54].

ZVI: Zhang et al. [53] used biochar-supported nZVI (BC-nZVI) to activate PDS for TPHs-polluted soil’s remediation. The highest removal efficiency reached 62.6%, which was potentially related to the redox action of Fe²⁺ and Fe³⁺. Qu et al. [44] fabricated Fe–biochar composites (nZVI@BC) to activate PDS during PAHs-polluted soil’s remediation. A PAH (BaP) degradation of 71.8% within 5 min and the corresponding kobs of 0.265 min⁻¹ were achieved without adjusting the pH, in which ¹O₂ was the dominant ROS. The nZVI@BC with a graphitized structure and defects accelerated the electron transfer, thus improving its catalytic performance. Li et al. [55] developed a biochar-supported Fe nanoparticle catalyst (FeNPs@BC) to activate PDS to remediate ATZ-polluted soil, achieving a 100% removal of ATZ within 10 min with the participation of SO₄•⁻, •OH, and ¹O₂. FeNPs@BC still maintained a high performance after four cycles of recycling. Additionally, the catalyst proved to be effective in degrading three other triazine herbicides, simazine, terbuthylazine, and ametryn, for a broader implication. Guo et al. [56] adopted nanosized ZVI/biochar (B-nZVI/BC) with PDS for the remediation of p-chloroaniline (PCA)-contaminated soil, achieving a high degradation efficiency at 95.9%. Wan et al. [9] used biochar loaded with nZVI to activate PDS for NCB-polluted soil’s remediation, in which the removal of m-NCB, o-NCB, and p-NCB from the soils was 98.7%, 95.3%, and 94.6% within 60 min, respectively. The electron transfer between the biochar and nZVI was key to activating PDS, leading to the generation of active species such as SO₄•⁻, •OH, and ¹O₂ through both free radical and non-radical pathways. The effectiveness of the treatment technology was further validated through its successful application in actual NCB-contaminated soil remediation. Wang et al. [57] fabricated a N-doped biochar (NBC)-ZVI composite (NBC-ZVI) by ball milling to activate PDS for PAH (pyrene) degradation in the soil. The incorporation of N-doping and alloying heterojunctions led to a surface charge redistribution, optimizing the PDS oxidation and pyrene adsorption on NBC-ZVI. This modification created an efficient reaction center that significantly accelerated the overall reaction process. NBC-ZVI combined with PDS degraded 95.5% of pyrene within 7 days, which was 2.3 times higher than ball-milled ZVI and 1.5 times more effective than BC-ZVI. In addition, the NBC incorporation induced a direct electron transfer from ZVI to NBC for the PDS activation and SO₄•⁻ generation. Zeng et al. [58] successfully applied nZVI/BC to activate PDS for organic-polluted soil’s remediation on an in situ pilot scale. Specifically, the removal of organic pollutants, including 2-ethylnitrobenzene, 4-(methylsulfonyl) toluene, biphenyl, and 4-phenylphenol, ranged from 88.6 to 99.9%, providing a successful implementation case of remediation using nZVI/BC-PS and verifying its feasibility in real contaminated soil remediation.

FeS: Liu and Yang [54] synthesized the hybrids of biochar Fe0 and FeS (Fe0-FeS@BC) by ball milling to activate PDS for SMX-polluted soil’s remediation. SMX was efficiently removed with the participation of ¹O₂, O₂•⁻, •OH, SO₄•⁻, and persistent free radicals along with a high Fe³⁺/Fe²⁺ cycle under reducing conditions. Xia et al. [59] investigated the degradation of TPHs, PAHs, and n-alkanes in soil by an FeS@BC-activated PDS process and achieved removal rates of 61.8%, 78.2%, and 91.6%. These are higher than those using FeS as activators, which were 47.9%, 51.2%, and 79.3%, respectively. The reduction of Fe²⁺/Fe³⁺ and the activation by biochar acted as an electron transfer mediator to promote the generation of SO₄•⁻. Zhao et al. [60] developed a pyrite (FeS)–biochar composite to activate PMS for SMX-polluted soil’s remediation, achieving a superior activation performance with an SMX removal of 76% from the soil within 120 min. The ROS responsible for soil remediation was ¹O₂, which is derived from O₂•⁻.

3.3. Biochar Loaded with Other Metals

During the combination of the iron and manganese atoms, a promoted electron transfer occurred, which could accelerate the decomposition of pollutants, providing another coating adhesion onto the surface of biochar to activate PS [61]. Li et al. [62] developed biochar loaded with Fe and manganese oxides (FeMn@BC) to activate the PDS for Thiacloprid (THI)-polluted soil’s remediation, in which the removal rate of THI reached 92.5% within 60 days with a greater generation of HO• and SO₄•⁻ by the FeMn@BC addition. Zhu et al. [63] synthesized a biochar/geopolymer catalyst loaded with Fe and Cu nanoparticles (Fe–Cu@BC-GM), which could efficiently activate PMS owing to the redox cycling effects of iron and copper. The Fe–Cu@BC-GM and PMS were used to degrade PAH (naphthalene) in soil, achieving a 67.98% reduction within 120 min. The analysis of degradation products and phytotoxicity tests on plants demonstrated that the naphthalene was effectively transformed into low- or non-toxic substances, achieving the successful remediation of naphthalene-contaminated soil. Liu et al. [64] applied biochar loaded with ZnCl₂ (SSBC) to activate PDS for the crude-oil-polluted soil’s remediation, achieving a degradation rate reaching 34.2% in the first week. The SSBC is rich in surface -COOH groups (1.03 mM/g) and -OH groups (2.77 mM/g), which were responsible for the PS activation to generate SO₄•⁻ and •OH.

Although notable progress has been made in using biochar to activate PS for soil remediation, it is important to emphasize that the combined application of biochar and PS has achieved far greater advances in wastewater treatment [11,12]. A review of the current literature reveals that soil remediation studies have predominantly focused on pristine biochar or biochar modified with metals (e.g., Fe, Mn, and Cu) as PS activators. However, several other biochar modification strategies developed for aqueous systems, such as multi-element doping (e.g., N, S, B, and P) and surface functional group enhancement, have shown superior capabilities in enhancing the electron transfer, catalytic stability, and reactive species selectivity, as shown in the following section. Some of these advanced modifications may be transferable to soil matrices, particularly where the moisture content is limited and the soil chemistry is complex. Therefore, future research should explore and adapt these innovative biochar engineering approaches to improve the efficiency, adaptability, and sustainability of biochar–PS systems in contaminated soil remediation.

Table 3. Research progress on soil remediation by PS activated by biochar.

Biochar Descriptions	Pollutant	Pollutant Concentration (mg/kg)	Activator	Removal (%)	Activation Mechanism	Reference
Hydrochar from excess sludge	ATZ	101.8	PDS	95.3	The synergistic effect of hydrochar and PDS in soil remediation was observed.	Xue et al. [46]
Pyrochar from wheat straw	HCH	10.0	PDS	99.0	External LMWOAs addition enhanced the performance of the biochar-activated PDS system.	Dou et al. [47]
Pyrochar from the lychee branch	BPA	31.9	PDS	98.4	Biochar can activate PDS to generate SO4•− for BPA degradation and alleviate pH drop during soil remediation.	Liu et al. [48]
Pyrochar from peanut shells	SMX	20.0	PDS	68.4–90.7	SO4•− and •OH were the predominant reactive species. Iron minerals in the soil exert a facilitating effect, whereas organic matter exists as an inhibitor.	Chen et al. [49]
Pyrochar from seaweed with a ball-milled modification	CIP	126.0	PMS	96.1	1O2 is the more dominant ROS, and the non-radical pathway is dominant.	Masud et al. [50]
Pyrochar from wheat straw with microwave assistance	PTH	60.0	PDS	88.8	In biochar and MW systems, the activation of PDS into SO4•−, •OH, O2•−, and 1O2 contributed to the removal of PTH.	Zhao et al. [51]
Pyrochar from corn straw loaded with Fe (nZVI@BC)	PAHs	27.0	PDS	71.8	nZVI@BC activated PDS and enhanced non-radical pathways (1O2). Biochar can also act as an electron shuttle and accelerate electron transfer from BaP to PDS.	Qu et al. [44]
Pyrochar from peanut shells supported nanoscale nZVI (nZVI/p-BC)	NCB	13.0–15.2	PDS	64.0–82.4	The cooperation of the non-free radical (1O2) and the free radical (SO4•− and •OH) contributed to the high degradation, owing to nZVI and p-BC collaboratively activating PDS.	Wan et al. [9]
Pyrochar from bamboo waste supported nano iron (BC-nZVI)	TPHs	13,259	PDS	62.6	The degradation of TPHs was potentially related to the redox action of Fe2+ and Fe3+.	Zhang et al. [53]
Pyrochar from corn stalks supported Fe nanoparticle (FeNPs@BC)	ATZ	20.6	PDS	100	The participation of SO4•−, •OH, and 1O2 degraded ATZ.	Li et al. [55]
Pyrochar from rice straw loaded with Fe (B-nZVI/BC)	PCA	3.6	PDS	95.9	SO4•−, •OH, and O2•− radicals were responsible for PCA degradation.	Guo et al. [56]
Pyrochar from wood pulp N-doped biochar-modified ZVI (NBC-ZVI)	PAHs	98.3	PDS	95.5	NBC induced direct electron transfer from ZVI to NBC to activate PDS for SO4•− generation.	Wang et al. [57]
Biochar-supported nZVI (nZVI/BC)	2-ethylnitrobenzene Biphenyl 4-(methylsulfonyl) toluene 4-phenylphenol	1.5–1.6 0.02–0.2 0.3–0.4 1.7–2.5	PDS	88.6–99.9	In situ pilot-scale study.	Zeng et al. [58]
Pyrochar from corn stalks loaded with Fe0 and FeS (Fe0-FeS@BC)	SMX	20	PDS	97.5	SMX was efficiently removed with the participation of 1O2, O2•−, •OH, and SO4•−.	Liu and Yang [54]
Pyrochar loaded with FeS (FeS@BC)	TPHs	4186.4	PDS	61.8	The reduction of Fe2+/Fe3+ and the activation by biochar acted as an electron transfer mediator to promote the generation of SO4•−.	Xia et al. [59]
Pyrochar from corn straw and pyrite (pyrite-biochar composites)	SMX	10.0	PMS	76.0	The introduction of pyrite into biochar significantly increased the release of Fe(II), further enhancing the activation of PMS and the generation of ROS. •OH, SO4•−, and 1O2 (dominant ROS) were produced.	Zhao et al. [60]
Pyrochar from lignin loaded with Cu and Fe (Fe–Cu@BC-GM)	PAHs	20.0	PMS	68.0	Fe–Cu@BC-GM activated PMS to generate a lot of free radicals, such as O2•−, SO4•−, •HO, and 1O2, through electron transfer.	Zhu et al. [63]
Pyrochar from wheat straw loaded with Fe and Mn (FeMn@BC)	THI	5.0	PDS	92.5	FeMn@BC produced more HO• and SO4•−.	Li et al. [62]
Pyrochar from sludge and rice straw loaded with ZnCl2 (SSBC)	Crude oil	10	PDS	34.2	The SSBC is rich in surface -COOH groups (1.03 mM/g) and -OH groups (2.77 mM/g), which were responsible for PS activation to generate SO4•− and •OH.	Liu et al. [64]

Table 3. Research progress on soil remediation by PS activated by biochar.

Biochar Descriptions	Pollutant	Pollutant Concentration (mg/kg)	Activator	Removal (%)	Activation Mechanism	Reference
Hydrochar from excess sludge	ATZ	101.8	PDS	95.3	The synergistic effect of hydrochar and PDS in soil remediation was observed.	Xue et al. [46]
Pyrochar from wheat straw	HCH	10.0	PDS	99.0	External LMWOAs addition enhanced the performance of the biochar-activated PDS system.	Dou et al. [47]
Pyrochar from the lychee branch	BPA	31.9	PDS	98.4	Biochar can activate PDS to generate SO4•− for BPA degradation and alleviate pH drop during soil remediation.	Liu et al. [48]
Pyrochar from peanut shells	SMX	20.0	PDS	68.4–90.7	SO4•− and •OH were the predominant reactive species. Iron minerals in the soil exert a facilitating effect, whereas organic matter exists as an inhibitor.	Chen et al. [49]
Pyrochar from seaweed with a ball-milled modification	CIP	126.0	PMS	96.1	1O2 is the more dominant ROS, and the non-radical pathway is dominant.	Masud et al. [50]
Pyrochar from wheat straw with microwave assistance	PTH	60.0	PDS	88.8	In biochar and MW systems, the activation of PDS into SO4•−, •OH, O2•−, and 1O2 contributed to the removal of PTH.	Zhao et al. [51]
Pyrochar from corn straw loaded with Fe (nZVI@BC)	PAHs	27.0	PDS	71.8	nZVI@BC activated PDS and enhanced non-radical pathways (1O2). Biochar can also act as an electron shuttle and accelerate electron transfer from BaP to PDS.	Qu et al. [44]
Pyrochar from peanut shells supported nanoscale nZVI (nZVI/p-BC)	NCB	13.0–15.2	PDS	64.0–82.4	The cooperation of the non-free radical (1O2) and the free radical (SO4•− and •OH) contributed to the high degradation, owing to nZVI and p-BC collaboratively activating PDS.	Wan et al. [9]
Pyrochar from bamboo waste supported nano iron (BC-nZVI)	TPHs	13,259	PDS	62.6	The degradation of TPHs was potentially related to the redox action of Fe2+ and Fe3+.	Zhang et al. [53]
Pyrochar from corn stalks supported Fe nanoparticle (FeNPs@BC)	ATZ	20.6	PDS	100	The participation of SO4•−, •OH, and 1O2 degraded ATZ.	Li et al. [55]
Pyrochar from rice straw loaded with Fe (B-nZVI/BC)	PCA	3.6	PDS	95.9	SO4•−, •OH, and O2•− radicals were responsible for PCA degradation.	Guo et al. [56]
Pyrochar from wood pulp N-doped biochar-modified ZVI (NBC-ZVI)	PAHs	98.3	PDS	95.5	NBC induced direct electron transfer from ZVI to NBC to activate PDS for SO4•− generation.	Wang et al. [57]
Biochar-supported nZVI (nZVI/BC)	2-ethylnitrobenzene Biphenyl 4-(methylsulfonyl) toluene 4-phenylphenol	1.5–1.6 0.02–0.2 0.3–0.4 1.7–2.5	PDS	88.6–99.9	In situ pilot-scale study.	Zeng et al. [58]
Pyrochar from corn stalks loaded with Fe0 and FeS (Fe0-FeS@BC)	SMX	20	PDS	97.5	SMX was efficiently removed with the participation of 1O2, O2•−, •OH, and SO4•−.	Liu and Yang [54]
Pyrochar loaded with FeS (FeS@BC)	TPHs	4186.4	PDS	61.8	The reduction of Fe2+/Fe3+ and the activation by biochar acted as an electron transfer mediator to promote the generation of SO4•−.	Xia et al. [59]
Pyrochar from corn straw and pyrite (pyrite-biochar composites)	SMX	10.0	PMS	76.0	The introduction of pyrite into biochar significantly increased the release of Fe(II), further enhancing the activation of PMS and the generation of ROS. •OH, SO4•−, and 1O2 (dominant ROS) were produced.	Zhao et al. [60]
Pyrochar from lignin loaded with Cu and Fe (Fe–Cu@BC-GM)	PAHs	20.0	PMS	68.0	Fe–Cu@BC-GM activated PMS to generate a lot of free radicals, such as O2•−, SO4•−, •HO, and 1O2, through electron transfer.	Zhu et al. [63]
Pyrochar from wheat straw loaded with Fe and Mn (FeMn@BC)	THI	5.0	PDS	92.5	FeMn@BC produced more HO• and SO4•−.	Li et al. [62]
Pyrochar from sludge and rice straw loaded with ZnCl2 (SSBC)	Crude oil	10	PDS	34.2	The SSBC is rich in surface -COOH groups (1.03 mM/g) and -OH groups (2.77 mM/g), which were responsible for PS activation to generate SO4•− and •OH.	Liu et al. [64]

4. Biochar Modification Enhanced Activation of PS

In recent years, many PS activation technologies have been reported during wastewater treatment processes, which potentially inspired soil remediation. These technologies included the activation by heat, UV, microwave, ultrasound irradiation, gamma-ray, electrochemistry, nanosecond pulsed gas–liquid discharge plasma, carbon materials, metals, etc. [65]. Unlike the traditional wastewater treatment, soil remediation processes require more engineering and operational feasibility considerations. Consequently, biochar and modified biochar-activated PS technologies show great potential for application [66].

4.1. Metal-Modified Biochar

The loading of metals onto biochar can significantly modulate its surface properties. The covalent bonding interactions between the biochar and the metal can enhance the stability of the catalyst and further boost the catalytic activity of biochar-based catalysts [67].

4.1.1. Iron

Iron (Fe) is widely used for PS activation due to its low cost, accessibility, and non-toxic nature. In particular, Fe-loaded catalysts, such as Fe-doped biochar, provide a large specific surface area for uniform dispersion, exposing more metal sites and synergistically enhancing the catalytic performance of the PS-based process. As summarized by Cao et al. [68], during the soil remediation under acidic conditions, Fe-doped biochar activated PS to generate SO₄•⁻ (S₂O₈²⁻ + Fe²⁺ → Fe³⁺ + SO₄²⁻ + SO₄•⁻). With the increase in the pH, the SO₄•⁻ radicals converted into •HO (SO₄•⁻ + OH⁻ → •HO + SO₄²⁻). In addition, O₂ could act as an acceptor, capable of receiving electrons generated by Fe-doped biochar to produce more O₂•⁻ (Fe²⁺ + O₂ → Fe³⁺ + O₂•⁻), which is crucial in facilitating the cycling of Fe²⁺/Fe³⁺ redox pairs (Fe³⁺ + O₂•⁻ → Fe²⁺ + O₂). This could further enhance the PS activation to generate more SO₄²⁻ and •HO (S₂O₈²⁻ + H₂O → HSO₅⁻ + HSO₄⁻, HSO₅⁻ + O₂•⁻ → SO₄•⁻ + O₂ + OH⁻, and S₂O₈²⁻ + O₂•⁻ → SO₄²⁻ + SO₄•⁻ + O₂). Moreover, O₂•⁻ could also further generate ¹O₂ (O₂•⁻ + HO•₂ → ¹O₂ + HO₂⁻, O₂•⁻ + O₂•⁻ + H⁺ → ¹O₂ + 2H₂O₂). The generated ROS is responsible for the degradation of pollutants in soil. The research progress on the soil remediation by Fe-doped biochar has been well discussed in Section 3.2 (Table 3).

4.1.2. Copper

As a common transition metal, copper (Cu) is non-toxic, inexpensive, readily available, and possesses the potential for PS activation. Zhou et al. [69] utilized a CuO/PDS system to degrade phenol at a removal rate of 84.0% through a non-radical pathway (¹O₂) and a radical pathway (SO₄•⁻) mediated by Cu(I)/Cu(II). Qian et al. [70] developed a graphene-based copper complex and used it to activate PS for the removal of TCS. Wang et al. [71] prepared Cu-based hydroxyapatite with biochar to activate PS for the removal of 2,4,6-trichlorophenol, reaching 100% within 60 min under broad pH ranges (3–10). Song et al. [72] adopted the molten metal salt technique to prepare Cu-loaded biochar from lotus leaf biomass and used it for the PDS activation for tetracycline antibiotic degradation. The C-O-Cu structure created electron-rich centers, triggering Cu(I)/Cu(II) cycling and enabling Cu-HBC to activate PDS, generating ¹O₂ to oxidize tetracycline. However, Cu is limited by its high cost [65], which requires further exploration.

4.1.3. Cobalt

Compared to the metals Fe, Mn, and Cu, cobalt (Co) possesses a higher catalytic activity in the metal-based catalysts, which is related to its high redox potential of Co³⁺/Co²⁺ (E₀ = 1.92 V), while it was 0.77 V in Fe³⁺/Fe²⁺, 1.54 V in Mn³⁺/Mn²⁺, and 0.15 V in Cu²⁺/Cu⁺. Furthermore, the biochar carrier and Co active sites work synergistically by enhancing electron transfer, which boosts the activity and stability of the catalyst, efficiently activating PS, and lowering costs [65,67]. Hu et al. [67] adopted Co-loaded magnetic biochar (Co-MBC) to activate PMS to degrade metronidazole. A high removal rate of more than 90% was achieved, and it still showed an excellent catalytic performance of 75% after four cycles of reuse. Wang et al. [73] reported that the Co modification could alter the morphology of hydrochar from a granular to rose-shaped lamellar structure and further to helical sheet structures. It could also increase the specific surface area, defect degree, and oxygen-containing groups of biochar, showing an excellent activation of PS for TCS removal (98%). Yi et al. [74] reported that Co-loaded biochar (Co-BC) was also effective in orfloxacin degradation (97.7%), in which Co mainly existed in the form of Co0 and Co₃O₄. However, Co is limited in practical applications due to its high toxicity [65], requiring measurements to overcome this.

4.1.4. Manganese

Besides Fe and Co, manganese (Mn) also has strong catalytic properties [75]. Luo et al. [76] utilized Mn to modify the magnetic biochar-based catalysts for PS activation, achieving a removal of metronidazole of 95.6%. Yan et al. [77] prepared Mn-doped biochar from waste sludge through the hydrothermal–calcination co-pyrolysis method and used it to activate PDS for the removal of phenol, reaching 100% within 180 min. Liu et al. [78] proved that Mn-Mg loading granted biochar with enhanced surface areas and adsorption sites, thus effectively activating PS. The target pollutant CIP was degraded by 90.8% using the Mn-Mg-doped biochar/PS system. Gao et al. [75] synthesized Mn-N-S co-doped biochar to activate PMS to degrade CIP, and over 99% of the CIP was removed after 60 min. The participation of Mn greatly improved the degradation performance due to the reaction between the PMS and manganese dioxide, which accelerated the electron transfer.

4.1.5. Zinc

Zn modifications can enlarge the pore size and volume and increase the surface area, thereby improving its performance in PS activation. Yang et al. [79] investigated the effectiveness of ZnCl₂-modified biochar on PS activation in the removal of acetaminophen, in which the acetaminophen removal reached 96% within 5 min. Demarema et al. [80] reported that nitrogen-doped ZnO is effective in PS activation for the removal of methylene blue (95.7%), congo red (65.4%), and methyl orange (59.2%). Zhang et al. [81] prepared porous pie-like nitrogen-doped biochar with ZnCl₂ loading and applied it for PMS activation to degrade SMX, reaching 99.8% within 5 min, in which the main active reaction sites were graphitic N, pyridinic N, and C=O groups.

4.1.6. Lanthanum

Lanthanum (La) is a rare earth element with a high biocompatibility, low toxicity, and minimal environmental hazards. La-modified biochar could enhance the adsorption capacity and promote surface reactions, which also shows great potential in PS activation. Peng et al. [82] used La-doped magnetic biochar from bagasse to activate PS for the degradation of antibiotic florfenicol, in which the removal efficiency reached 99.5%. Jun et al. [83] also suggested a critical role of La in assisting the PMS activation for the degradation of RhB. Razmi et al. [84] also activated PDS by La-modified biochar to degrade phenol, reaching 97.7%. Li et al. [85] loaded La and Co onto rice straw biochar and then used it for PMS activation for the degradation of 4-aminophenylarsonic acid (100% within 30 min), simultaneously achieving the adsorption of arsenic (nearly complete removal) along with a superior reusability and stability. Koba-Ucuna et al. [86] reported that La was effective in PS activation for the removal of black 5 dye.

4.1.7. Multiple Metals

During the metal-activated PS process, the process from M⁽ⁿ⁺¹⁾⁺ to Mn+ determines the reaction rate, in which the accumulation of high-valence metal ions limits the process. The combination of different metals is a potential alternative to solve this problem. The introduction of another metal will change the original redox cycle. The rapid charge transfer between the different metals significantly facilitates the redox process and thus improves the catalytic performance of the PS activation [87].

As reported by Yue et al. [88], the presence of Fe enhanced the charge transfer kinetics for the reduction of Ni³⁺ to Ni²⁺, while also suppressing the formation of undesirable Ni components with higher oxidation states. The dual Ni-Fe composites demonstrated outstanding performances in PS activation, showing an exceptionally high activity for degrading phenolic compounds. In addition, PS can be efficiently activated by Fe-Co composites for BPA degradation [89], Fe-Cu composites for methyl violet degradation [90], Fe-Mn composites for acid orange 7 degradation [91], Cu-Co composites for SMX degradation [92], Mn-Co composites for metronidazole degradation [93], Al-Co composites for sulfadiazine degradation [94], Fe-Cu-Co composites for nitrobenzene degradation [95], Fe-Cu-Mg composites for ethylbenzene degradation [96], Fe-Mg-Co composites for carbamazepine degradation [97], Fe-Cu-Ni composites for lomefloxacin degradation [98], Fe-Co-Ni composites for congo red and rhodamine B degradation [99], Fe-Mn-Al composites for a wide pollutant degradation [100], Cu-Co-Zn composites for 4-chlorophenol degradation [101], Co-Mg-Al composites for ATZ degradation [102], Ni-Mg-Al composites for phenol degradation [103], and Mg-Mn-Cu composites for SMX degradation [104].

4.2. Non-Metallic-Modified Biochar Enhanced Activation of PS

4.2.1. Nitrogen Doping

N-doped sites are recognized as key active sites in carbon materials (Figure 2). The synthesis of N-doped carbon materials can be achieved through several methods, including the thermal polymerization of nitrogen-containing precursors, like urea, thiourea, melamine, dicyandiamide, and cyanamide. The structural defects and the extent of nitrogen doping can be finely tuned by varying the proportion of these nitrogen-containing precursors. The N-doped defective incorporated configurations mainly consisted of graphitic-N, pyridinic-N, pyrrolic-N, and oxidized-N [105]. Xia et al. [106] reported that biochar from pig bone waste was wrapped with nitrogen-doped additives, which exhibited active defect nitrogens (pyridinic and graphitic nitrogen), showing an effectiveness in tetracycline removal (93%) within 30 min. Among the nitrogen configurations at defect sites, pyridinic nitrogen notably impacted the electron cloud density of nearby carbon networks, thereby facilitating the formation of excited states in PS. Yang et al. [107] adopted biochar doped with nitrogen from Foxtail algae (N-rich wetland plant) to activate PS to degrade acid orange 7, reaching 97.4%. Cui et al. [108] used nitrogen-doped biochar from sewage sludge to activate PS for a 2,4-dichlorophenol removal (100%) within 120 min, in which graphitic N was identified as the main activation site. Zhao et al. [109] compared ball milling, roasting, and tempering for the nitrogen doping of biochar. Ball milling, a straightforward and cost-effective method, was effective in the nitrogen doping of the biochar preparation from sesame residue, which was featured by pyridine nitrogen as the main functional group along with an improved crystallinity and aromaticity. It showed an excellent performance in phenol degradation within 15 min (96.8%), with a satisfactory reusability within six cycles. Zeng et al. [110] used N-doped biochar from peanut hulls and soybean curd residue to activate PS for tetracycline degradation. The organic-N accelerated lattice disruption resulted in the formation of hierarchical structures and pyrrole-N. The removal of tetracycline reached 95%, which was related to C=O and pyrrole-N.

4.2.2. Sulfur Doping

Sulfur (S) doping could disturb the charge population and create more reactive sites, boosting the performance of biochar in the PS activation (Figure 2) [111]. He et al. [112] synthesized sulfur (S) and N co-doped biochar from sewage sludge and sulfonated lignin, in which graphitic nitrogen and thiophene sulfur were efficiently integrated into the biochar. The N-S-doped biochar could activate PMS for sulfadiazine degradation, achieving a 31.9% removal rate within 120 min. Wang et al. [113] developed S and N co-doped magnetic biochar using Eichhornia crassipes. Hydroxylamine was further added to enhance the PDS activation for an organic pollutant (AO7) degradation, reaching 95.4% within 10 min. Li et al. [114] prepared S and N co-doped biochar to activate PDS for phenol degradation, reaching 98.5%, in which S and N exist as thiophene sulfur and graphite nitrogen in the biochar. Wang et al. [111] synthesized sulfurized biochar from sewage sludge and used it for PS activation for the removal of BPA. The integrated S effectively activated the sp2-hybridized graphene lattice, resulting in the formation of additional Lewis acid and base sites in sulfurized biochar, which was beneficial to the PS activation.

4.2.3. Boron Doping

Boron (B), with a lower electronegativity than carbon, acts as a dopant by easily losing electrons and lowering the Fermi level. Its small radius allows it to replace sp2 and sp3 structures in the carbon lattice, enhancing graphitization. The electron-deficient nature of boron creates a positive charge density, making it an attractive adsorption site for negatively charged species like PS, thereby enhancing the PS activation process (Figure 2) [115]. Chen et al. [116] proved that B-doped biochar was effective in improving the PDS activation for the decomposition of practical dyeing wastewater. B-doped biochar enhances the stability and oxidation potential by forming a surface-bound complex, which facilitates the electron transfer from pollutants to the complex. In addition, Zhang et al. [117] prepared N-B co-doped biochar from Auricularia auricula, urea, diatomite, and boric acid and then used it to activate PDS for ATZ degradation (98%). Zhang et al. [118] adopted N-B co-doped biochar graphene from coconut shells for the activation of PDS, showing an efficient degradation of oxytetracycline (94.1%) within 15 min. In addition, pyrrolic N and pyridinic N played important roles in electron transfer within the radical pathway, contributing to the activation of PDS to produce radicals. Moreover, the introduction of N and B reduces the activation energy in the reaction process.

4.2.4. Phosphorus Doping

Phosphorus (P) is a favorable dopant because of its large atomic size, valence multiplicity, and electron–donor capacity. The difference in covalent radii between P and C can modulate the electron distribution and structural defects in the carbon lattice to produce more active sites, thus effectively activating PS (Figure 2) [119]. Wang et al. [120] found that N- and S-doped biochar failed to improve the PS activation, while P-doped biochar increased the reactivity by five times compared to pristine biochar during an antibiotic norfloxacin degradation, reaching 79.0% within 2 h. The mechanisms were attributed to large specific surface areas, a high adsorption, a high C-P-O content, graphitic P, and non-radical degradation pathways. Huang et al. [121] developed a P-doped biochar featuring abundant nanocracks for PS activation and gamma-hexachlorocyclohexane degradation (92.6% within 10 min). The nanocracked structure was attributed to electrostatic stress and new nucleation sites from P-doping. Xie et al. [119] anchored P on in situ N-doped biochar via ball milling, showing an excellent catalytic activity in the diclofenac sodium degradation (90.1%) within 10 min. This was attributed to the dual effect of the high-efficiency adsorption and electron transfer capabilities. Shi et al. [122] prepared P-doped biochar to activate PS for the degradation of acetaminophen, which was completely removed within 90 min. The free radicals and singlet oxygen were not the main reactive species, while the electron transfer mediated by biochar was responsible for the degradation of acetaminophen.

4.2.5. Multiple Elements

During the non-metal co-doping strategy for PS activation, the synergistic interaction among heteroatoms, such as N, S, P, and B, significantly influences the surface chemistry and electron distribution of biochar. These elements introduce abundant defects and modulate the electronic density of states, which in turn alter the redox behavior of the carbon matrix [123]. For example, N and S co-doping can enhance electron delocalization and promote the generation of reactive sites, such as thiazole, facilitating the electron transfer during the PS activation [124]. Similarly, the incorporation of P and B alters the local charge environment and improves the activation efficiency of PS [125]. The interaction between these non-metal dopants can break the electroneutrality of the carbon structure and create polarization zones, which are favorable for radical and non-radical generation pathways. Similar interactive effects were also found in the combinations of N-P [126], N-B [127], and N-S-P [123]. Moreover, the co-doping of these non-metallic elements in conjunction with metals (e.g., Fe, Co, Cu, and Mn) has emerged as a promising strategy to enhance the catalytic performance of biochar for PS activation [128,129,130].

4.3. Functional Group Modification

The content and type of oxygen-containing functional groups on the surface of carbonaceous materials affect the production of ROS [87,131]. Ketone-based functional groups, including quinones and carbonyl (C=O) groups on carbon materials, have been identified as active sites for PS activation [87] and selectively generate ¹O₂ for pollutant degradation [105,132]. Cheng et al. [133] reported that pollutant (2,4-dichlorophenol) degradation decreased when carbon materials (carbon nanotubes) lacking C=O groups were used, suggesting that the C=O groups function as active sites that are consumed during the process. This could be attributed to the high electron density and strong electron-donating properties of quinone intermediates. Xie et al. [119] reported a key role of the C=O group within the P-doped biochar/PS system for diclofenac sodium degradation. Yang et al. [134] reported that the C=O group was deemed to be the active site of the non-radical pathway of biochar during the aniline degradation in the biochar/PS system.

Metals in layered double hydroxides can increase the content of surface hydroxyl groups (-OH), which can react with PS and cleave the O–O bond to produce ROS [87]. Under the function of surface hydroxyl groups, a stable operational performance for the BPA removal over a wide pH range was achieved [89]. Demarema et al. [80] suggested that the presence of functional -OH groups could lead to additional radicals during the methylene blue degradation by a nitrogen-doped ZnO-activated PS process.

Besides the C=O and -OH groups, carboxyl groups (-COOH), ester groups (-COO-), nitrogen-containing groups (-NH₂ and -CONH₂), and sulfur-containing groups (-SH, -S-, SO₃H, and -SO₂) also showed potential to activate PS [135,136], while requiring further exploration in this field.

5. Impact of PS/Biochar Technology on Functional Microorganisms

Differently from wastewater treatment (dominated by chemical reactions of ROS), the biochar-activated PS technology for soil remediation is a synergetic function of chemical reactions and biological reactions [46]. As summarized in

Table 4. Impact of PS/biochar technology on microorganisms during soil remediation.

Pollutant	Biochar	Microorganisms and Functions	Reference
ATZ	Sludge-derived biochar	Comamonas and Cloacibacterium: The hydrolysis of nitrogenous heterocyclic compounds. Alicyclobacillus and Halomonas: Resistant to PDS stress. Bacillus: Positively correlated with the degradation of ATZ.	Xue et al. [46]
PAHs	nZVI@BC	Firmicutes: Main degradable bacteria. Bacillus: PAHs-degrading bacteria.	Zhu et al. [63]
THI	FeMn@BC	Actinobacteriota: Possibly converting the intermediate of THI into smaller molecules.	Li et al. [62]
Crude oil	SSBC	Bacteroidetes, Alcanivorax, Marinobacter, Aliifodinibius, and Salinisphaera: Crude-oil-degrading bacteria.	Liu et al. [64]
TPHs	BC-nZVI	Proteobacteria, Bacteroidetes, Firmicutes, and Actinobacteria: Related to TPHs degradation. Acinetobacter, Corynebacterium_1, and Staphylococcus: TPHs-degrading bacteria.	Zhang et al. [53]
PBDEs Metals	Commercial rice Husk biochar	Gammaproteobacteria, Alphaproteobacteria, Clostridia, and Acidobacteria: Associated with the degradation of PBDEs and their intermediates.	Ma et al. [140]

, during ATZ-polluted soil’s remediation by sludge-derived biochar and PDS [46], the abundance of bacteria showed negative relationships with the PDS dosage and positive relationships with the biochar dosage. Biochar could alleviate the negative impacts of PDS on soil functionality, in which the bacterial abundance recovered to 7.7 log gene copy number/g soil compared to the raw soil (8.4) under the proper amendment of the PDS and biochar. In raw soil, the main phyla were Actinobacteria, Proteobacteria, Chloroflexi, and Acidobacteria, accounting for 89.6%. Once the soil was polluted by ATZ, the dominant phyla shifted to Firmicutes and Proteobacteria, accounting for 87.7%. When the biochar was solely amended, the relative abundance of Comamonas and Cloacibacterium [137], contributing to the hydrolysis of nitrogenous heterocyclic compounds, increased from 0.05 and 19.2% (Ctrl) to 20.8% and 31.0%. However, they disappeared under the oxidative stress of PDS. PDS-resistant Alicyclobacillus was dominant in P1 groups (P1.5C0, P1.5C2.5, and P1.5C5.1). Genus Halomonas was enriched when further increasing the PDS additions. In the biochar and PDS system, the degradation of ATZ was positively correlated with Bacillus [46]. Li et al. [62] reported that THI significantly altered the structure of the bacteria at the genus level after a 15-day reaction based on an 16S rRNA analysis. FeMn@BC/PS amendments could alleviate such damage, with 658 more operational taxonomic units compared to the THI. After 60 days, the relative abundance of the genus Actinobacteriota increased to 51.5% with the amendment of FeMn@BC/PS, which was also reported to be enriched with the biochar and available Fe and Mn [138] with the function of decomposing complex organics [139].

During the PAH-polluted soil’s remediation by the BC-nZVI-activated PDS process [53], the degradation of TPHs was divided into a chemical process (0–6th day) and a biological process (6–60th day). The addition of BC-nZVI significantly affected the microbial structure and increased the abundance and microbial metabolic activities of bacteria in the soil. Dominant phyla were Proteobacteria, Bacteroidetes, Firmicutes, and Actinobacteria, which were potential TPH degraders. At the genus level, the TPH-degrading bacterium Acinetobacter increased while Corynebacterium_1 and Staphylococcus declined with the BC-nZVI amendment. Zhu et al. [63] reported that the synthesized nZVI@BC and PDS were effective in treating PAH (BaP)-polluted soil, and simultaneously increased the relative abundance of the functional bacteria. The dominant bacterial phyla in the control were Actinobacteriota (39.9%), Proteobacteria (16.9%), Firmicutes (16.1%), and Chloroflexi (11.8%). The relative abundance of the Firmicutes phylum, which was known as degradable bacteria, increased to 46.5% with the nZVI@BC addition. At the genus level, nZVI@BC improved the abundance of bacteria associated with the stress resistance of organic pollutants (Bacillus, Pacnibacillus, and Cohnella). Specially, the genus Bacillus could strongly tolerate extreme environmental conditions and is able to efficiently degrade PAHs [141,142].The nZVI@BC/PS system could enhance the PAH-degrading bacteria abundance and further improve the degradation of PAHs, changing soil physicochemical properties and favoring plant (Lettuce) growth.

During the crude-oil-polluted soil’s remediation by SSBC [64], potential crude oil degraders were enriched, including Alcanivorax (12.5–12.9%), Marinobacter (10.8–12.5%), Aliifodinibius (7.0–7.4%), and Salinisphaera (6.6–7.2%), which could degrade organics into CO2 and H2O via TCA. During the polybrominated diphenyl ethers (PBDEs) and toxic metals soil remediation [140], the utilization of PS reduced the soil organic content, leading to a dramatic decrease in the bacterial density. However, the microbial activity and numbers generally recovered after the 90-day bioremediation. In the initial soil, the dominant phylum is Proteobacteria (52.4%), followed by Bacteroidetes (12.5%), Firmicutes (7.7%), Synergistetes (4.9%), and Actinobacteria (2.2%). Compared to that of the oxidized soil (PS treatment), the abundance of Gammaproteobacteria, Alphaproteobacteria, Clostridia, and Acidobacteria increased after the bioremediation, which was associated with the degradation of PBDEs and their intermediates.

In general, the application of PS in soil remediation inevitably imposes oxidative stress on soil microorganisms due to the generation of ROS. In fact, this detrimental impact is not unique to PS. Other commonly used oxidants, such as hydrogen peroxide (H₂O₂) [143], potassium permanganate (KMnO₄) [144], and ozone (O₃) [145], also produce various ROS during remediation processes, which can adversely affect the microbial viability, diversity, and ecological function. For instance, H₂O₂ oxidation reduced the abundance of viable hydrocarbon-degrading microorganisms, possibly related to ROS, thereby inhibiting the microbial biodegradation of petroleum hydrocarbons [143]. Fortunately, biochar can effectively compensate for the negative impact [140]. The capillary structure and surface of biochar create an optimal environment for microorganisms, enabling the formation of biofilms on the biochar surface, which helps protect against adverse environmental factors. Additionally, biochar adsorbs nutrient cations through its functional groups, providing essential nutrients and ions to soil microorganisms, thereby promoting their growth and reproduction [5]. In addition, the stimulation of pollutants, biochar, and PS could facilitate the secretion of extracellular polymeric substances (EPSs) to protect the microorganisms themselves [146]. Using similar technology, the combination of biochar and PS increased the abundance of the genus Pseudomonas, which facilitated the hydrolysis of protein and promoted acetic acid production during anaerobic digestion [147]. As reported by Chen et al. [148], the genus Pseudomonas has shown great potential in PAHs-polluted soil’s remediation. It has been reported that EPS-bound active substances (e.g., flavins) impact electron transfer and potentially degrade pollutants [149,150]. Moreover, fungi, such as Pleurotus ostreatus, Aspergillus terreus, Trametes versicolor, Pleurotus ostreatus, white-rot fungi, etc., have shown great potential in organic pollutants during soil remediation. Fungi bioremediation also includes the following advantages: its adaptability to extreme conditions, enzyme diversity and efficiency, versatility in bioremediation applications, and symbiotic relationship with phytoremediation (Figure 3) [151]. Nevertheless, current studies only characterized the variation in the apparent microbial community (only in bacteria without an EPS analysis), overlooking the relationships between the function of microorganisms and organic pollutants. Future work should pay more attention to the interaction among pollutants, heavy metals, biochar, PS, bacteria, fungi, and microbial functionalities.

6. Perspectives and Outlooks

6.1. Sustainable Application: Cost, Scalability, and Environmental Safety

Although laboratory-scale investigations have demonstrated the potential of combined PS and biochar for soil remediation, several challenges must be addressed before a large-scale implementation. A major obstacle is the relatively high cost. PMS is currently priced approximately 2–3 times higher than PDS [152]. However, PMS exhibits high reactivity and a rapid degradation rate [39]. Future research should prioritize site-specific optimization strategies for selecting between PMS and PDS, taking into account practical conditions such as contaminant characteristics, soil properties, remediation goals, and cost constraints. Efforts should focus on reducing the oxidant dosage and enhancing utilization efficiency through approaches like slow-release oxidant formulations, precision injection techniques, and regenerable biochar-based catalysts. Such tailored strategies will substantially improve the economic viability and practical applicability of PS/biochar remediation technologies.

Additionally, the current scale of PS/biochar remediation studies remains primarily at bench or batch reactor levels [153,154], lacking sufficient validation under realistic, field-scale scenarios, especially in low-permeability or heterogeneous soils. Thus, more comprehensive field-scale trials across diverse global soil types, including sandy, clayey, loamy, peat, saline–alkaline, and anthropogenically altered urban soils, are required. These studies should be complemented by advanced transport and kinetic modeling to optimize the injection strategies, oxidant distribution, and interactions between oxidants, biochar activators, and soil matrices. Such comprehensive scalability evaluations will support the development of practical engineering solutions tailored for a wide range of contaminated soil environments worldwide.

Environmental safety and sustainability pose further challenges. Residual oxidants and their transformation intermediates, such as chlorinated derivatives or aromatic oxidation products, may exhibit a greater ecological toxicity or mobility compared to the original pollutants [155]. In addition, metal-doped biochar (e.g., Cu, Co, and Mn) may lead to a potential secondary contamination through metal leaching under prolonged oxidative conditions. Moreover, the impact of the combined biochar and PS treatment on soil microorganisms with diverse pollutant degradation remains unclear. It has been widely accepted that PS activation may pose phytotoxic risks. The co-application with biochar can mitigate these effects and even promote plant growth by improving soil physicochemical properties and reducing the bioavailability of certain contaminants. As reported by Zhang et al. [117], the combination of PS and biochar can effectively restore the growth status, the morphology of leaves, and the chlorophyll content of soybean seedlings to overcome the toxicity of ATZ. Nevertheless, it is uncertain whether residual oxidants adversely affect plant root systems. Thus, rigorous long-term ecological risk assessments employing advanced analytical techniques, such as high-resolution mass spectrometry, isotopic tracing, non-target screening, and microbial community profiling, are necessary to elucidate the fate, transport, and ecological toxicity of transformation products, as well as comprehensively evaluate the broader impacts on soil ecosystems.

Beyond pollutant degradation, the combined application of PS and biochar can significantly reshape the fundamental physicochemical properties of soil, which has critical implications for soil health and long-term remediation sustainability. PS tends to lower the soil pH due to the generation of acidic intermediates, such as sulfate and bisulfate ions, during oxidation processes [8]. Excessive acidification may deteriorate the soil structure and inhibit microbial activity. However, biochar, often possessing alkaline ash and rich in basic functional groups, can buffer this acidification effect, resulting in more stable or slightly elevated soil pH levels depending on its feedstock and pyrolysis conditions [156]. Moreover, biochar’s high porosity and large surface area enhance the soil water retention, which is especially beneficial in sandy or degraded soils [157]. Especially, the induced high moisture content can further enhance the performance of the PS activation [8]. In terms of soil organic matter, PS is a strong oxidant that may indiscriminately degrade humic substances and microbial-derived organic carbon, thus depleting soil fertility over time [8,54,140]. However, the application of biochar inevitably introduces exogenous organic matter into the soil, including low-molecular-weight compounds and persistent polyaromatic structures, depending on the feedstock and preparation conditions. These newly introduced organics may interact with native soil organic matter, forming complex matrices with distinct redox and sorption properties. Moreover, PS can oxidize heavy metals in soil, potentially increasing their mobility and environmental risk. Zhang et al. [158] evaluated the variation in the form of heavy metals in soil after using PMS. It was found that the employment of PMS has a strong ability to transfer heavy metals (such as Cd, Pb, and Zn) from a stable state to an active state in the soil. A further mechanical study indicates that the activation of heavy metals is mainly related to the decomplexation of organic chelators and the dissociation of inorganic macromolecules in soil. Importantly, the risk assessments showed that the reactive heavy metals after the PMS treatment in the soil will delay the growth of plants and increase the heavy metal content in plants and the risk of groundwater pollution. While biochar may counteract this effect through adsorption and stabilization mechanisms, systematic studies on their combined impact on metal speciation and long-term stability remain limited and warrant further investigation. Furthermore, the impact of such interactions on the PS activation and degradation efficiency remains insufficiently understood. Ignoring these synergistic or antagonistic effects could lead to the unintended degradation of soil functionality, undermining the sustainability of in situ treatment strategies.

The establishment of standardized evaluation protocols for PS- and biochar-based soil remediation technologies is important for ensuring their responsible, efficient, and sustainable application. These technologies, while promising in pollutant degradation, must be assessed through the lens of sustainability to ensure they contribute to long-term environmental, economic, and social well-being. Comprehensive life cycle assessments (LCAs) are particularly critical, as they provide systematic frameworks for quantifying environmental impacts, including carbon emissions, water resource consumption, energy demands, and alterations in soil ecological quality, associated with remediation processes. Specifically, LCA methodologies should encompass detailed evaluations of raw material sourcing, manufacturing, transportation, application processes, and the ultimate disposition of residual materials and transformation products post-treatment [159,160].

Moreover, standardized evaluation criteria should incorporate not only conventional environmental metrics but also economic indicators, such as cost–benefit analyses, technology scalability, and market feasibility. Such integrative assessments will empower policymakers and environmental managers to make informed decisions that balance remediation efficiency, economic viability, and environmental sustainability. Additionally, the development of region-specific decision-support tools, which consider local soil characteristics, climatic conditions, and socio-economic contexts, is imperative. These tools, in conjunction with risk management guidelines tailored to specific regional conditions, will facilitate the appropriate selection, design, and implementation of PS/biochar remediation systems. Ultimately, the establishment of comprehensive, standardized evaluation frameworks will significantly enhance the global adoption of PS- and biochar-based remediation technologies, ensuring environmental safety, economic feasibility, and ecological integrity across diverse application scenarios.

6.2. Toward Smart, Tunable, and Regenerable Biochar Catalysts

The design of next-generation biochar catalysts for PS activation should surpass traditional single-functionality approaches, shifting toward multifunctional, adaptive, and sustainable remediation technologies. Future research should prioritize fabricating smart biochar materials characterized by a precisely tunable surface chemistry, optimized pore structures, and tailored redox properties to enhance the catalytic performance under diverse environmental conditions. Specifically, innovative multi-doping strategies, such as metal–non-metal co-doping or hierarchical doping techniques, could provide synergistic active sites capable of significantly boosting the PS activation efficiency and selectivity [161]. The regenerability and recyclability of engineered biochar catalysts represent another critical frontier, which is currently underexplored yet essential for sustainable applications.

In addition, integrating machine learning (ML) algorithms and data-driven models into the catalyst design process represents a promising direction. ML approaches, including artificial neural networks, random forests, and deep learning algorithms, can efficiently process extensive datasets generated from material synthesis, characterization, and catalytic testing. Such computational strategies can predict and optimize biochar synthesis parameters, doping ratios, structural properties, and corresponding catalytic efficiencies, significantly accelerating the development of novel, high-performance biochar catalysts [162,163,164]. Ultimately, combining advanced materials engineering, intelligent stimuli-responsiveness, and data-driven optimization will promote the evolution of PS/biochar remediation technologies toward smarter, more sustainable, and environmentally adaptable solutions.

7. Conclusions

This review summarizes the research progress on different organic pollutant degradations during soil remediation using PMS and PDS, respectively. Subsequently, the activation of PS by biochar for the degradation of various pollutants during soil remediation was systematically reviewed, which could be classified into pristine biochar, Fe-loaded biochar, and other metal-loaded biochars. In addition, recent research advances in wastewater treatment and modified biochar for the activation of PS are deeply understood, which mainly consisted of metal modifications (iron, copper, cobalt, manganese, zinc, lanthanum, and multiple metals), non-metallic modifications (nitrogen, sulfur, boron, and phosphorus), and functional group modifications, which potentially inspired soil remediation. Finally, the overlooked functionality of microorganisms during soil remediation in the PS/biochar system was well described. This review provides a better understanding of organic soil remediation using the biochar-activated PS process.