Removal of Tetracycline from Water by Biochar: Mechanisms, Challenges, and Future Perspectives

Abstract

Tetracycline (TC) is widely used in medicine and livestock farming. TC is difficult to degrade and tends to persist and accumulate in aquatic environments, and it has gradually become an emerging pollutant. Biochar (BC) has strong potential for removing TC from water. This potential arises from its excellent surface properties, low-cost raw materials, and renewable nature. However, raw biomass materials are highly diverse, and their preparation conditions vary significantly. Modification methods differ in specificity and the application scenarios are complex. These factors collectively cause unstable TC removal efficiency by biochar. The chemical activation process using KOH/H₃PO₄ significantly enhanced porosity and surface functionality, transforming raw biochar into an activated carbon material with targeted adsorption capacity. Adjusting the application dosage and environmental factors (particularly pH) further enhanced the removal performance. Solution pH critically governs the adsorption efficiency: optimal conditions (pH 5–7) increased removal by 35–40% through strengthened electrostatic attraction, whereas acidic/alkaline extremes disrupted ionizable functional groups. The dominant adsorption mechanisms of biochar involved π–π interactions, pore filling, hydrophobic interactions, hydrogen bonding, electrostatic interactions, and surface complexation. In addition, the main challenges currently hindering the large-scale application of biochar for the removal of TC from water are highlighted: (i) secondary pollution risks of biochar application from heavy metals, persistent free radicals, and toxic organic leaching; (ii) economic–environmental conflicts due to high preparation/modification costs; and (iii) performance gaps between laboratory studies and real water applications.

1. Introduction

Tetracycline (TC) is a low-cost antibiotic with extensive global use. Its residues create serious ecological risks in aquatic environments. Medical wastewater serves as a primary TC contamination source. TC concentrations exhibit substantial regional variations. For example, Beijing hospital wastewater contains 0.02081–0.07633 μg/L TC [1] while Portuguese wastewater contains 23.2–54.7 μg/L TC [2]. Lower levels have been recorded in Vietnam (0.1 ± 0.0 μg/L) [3] and Iran (0.068–0.144 μg/L) [4]. In addition, the excessive use of antibiotics in the livestock industry is also a significant contributor to TC contamination in the aquatic environment. TC consumption in Asia is projected to surge from 25,260 tons/year (2010) to 51,851 tons/year by 2030 [5]. The EU and USA currently use 2575 and 3230 tons/year, respectively [6]. Poor TC absorption leads to environmental release through animal manure. Chinese farm manure contains 290.5 μg/kg TC [7], and pig operations produce 6500–9600 μg/kg TC [8]. Aquaculture contributes significantly to TC pollution. Global aquaculture used 4242 tons of total TC in 2018 [9]. The detection rates reached 64% in Vietnamese aquaculture systems and 10.8% in Korean fish samples [10,11]. Thai aquaculture wastewater contains 180 ng/L TC, which exceeds China’s Pearl River Basin level (80 ng/L) [12]. Approximately 23% of residual antibiotics (1563 tons) enter freshwater systems annually, with 18% (1247 tons) reaching oceans. China’s Liao River Basin has a TC concentration of 160 ng/L, which is 1.6 times higher than that of Nigerian rivers [13]. Zimbabwe’s surface water contains extreme concentrations of 150 μg/L [14].

The global TC consumption has a strong geographical concentration. Russia, China, and the United States are the top three consumers. Russia uses 13,579 tons annually [15], followed by China at 6950 tons [16], and the U.S. at 3230 tons [17]. Europe represents a major regional consumer, with an annual TC use of 2575 tons [5]. However, significant disparities exist among European countries: France uses 185 tons/year [18], the UK 308 tons, and Germany 542 tons [19]. In contrast, South Korea’s annual TC consumption has reached 1278 tons [20], exceeding that of most European nations.

Table 1. Global TC concentrations in major urban rivers.

S. No		Region	Unit	TC	Refs.
1	River	Yangtze River in China	ng/L	11.16	[21]
2	River	Tonghui River, China	ng/L	ND–6800	[22]
3	River	Wenyu River, China	ng/L	ND–9500	[22]
4	River	Qinghe River, China	ng/L	ND–6800	[22]
5	Lake	Poyang Lake, China	ng/L	ND–10.8	[23]
6	River	River Ganges, India	ng/L	98.62	[24]
7	River	Soeste River, Germany	ng/L	21	[25]
8	River	Euphrates River, Turkey	ng/L	50.0 ± 2.5	[26]
9	River	Drwęca River, Poland	ng/L	54	[27]
10	Lake	Lake Victoria, Uganda	ng/L	3–70	[28]
11	River	Bang Pakong River, Thailand	ng/L	15.1	[29]
12	River	Han River, Korea	ng/L	2093	[30]
13	River	Tigris River, Iraq	ng/L	43.57	[31]
14	River	Malaysia, Kajang	ng/L	132	[32]
15	Rivers	Hong Kong, China	ng/L	ND–31.5	[33]
16	Rivers	Austria, Vienna	ng/L	122	[32]
17	Lake	Hanoi, Vietnam	ng/L	101	[34]
18	Lake	Taihu Lake, China	ng/L	ND–125	[35]
19	Lake	Dongting Lake, China	ng/L	ND–21.5	[35]

Note: ND represents not determined.

summarizes the global TC concentrations in major rivers.

TC demonstrates chemical stability and environmental persistence and remains in water environments for long periods. This pollutant accumulates in food chains through biomagnification. The accumulation creates a cross-media transport pathway of “water–soil–organisms”. Toxicological studies have shown that TC inhibits aquatic organism development. Exposure to 20 μg/L TC significantly inhibited zebrafish embryo development, which led to delayed hatching, reduced body length, impaired yolk sac absorption, and the underdevelopment of swim bladder cavities [36]. TC enters agricultural systems through irrigation. Gudda et al. reported TC accumulation in vegetables irrigated with contaminated water. The TC concentrations in leafy vegetables were 4.2 times greater than those in the soil. The enrichment factor reached 6.8 for these vegetables [37]. Recent studies revealed hepatocyte membrane damage at TC concentrations above 50 μg/L. TC disrupts hepatocyte membrane lipid bilayers. This disruption increases membrane permeability. As a result, mitochondria and endoplasmic reticulum dysfunction occur [38].

In recent years, major advances in TC removal from water have been reported. Primary technologies include biodegradation [39], photocatalytic degradation [40], adsorption [41,42], membrane filtration, and advanced oxidation processes [43]. Adsorption is a key method for removing TC and offers low operational costs. This process generates minimal secondary pollution. The adsorbents demonstrated strong resistance to degradation. Biochar is widely regarded as the most promising green adsorbent. It uses renewable biomass feedstocks. Biochar production supports carbon sequestration. Its surface properties allow for chemical modifications [44]. TC adsorption efficiency depends on two physicochemical characteristics: (1) multiple pore structures provide mass-transfer channels [45], and (2) oxygen-containing functional groups (carboxyl and hydroxyl groups) and aromatic carbon skeletons enable molecular trapping. Key mechanisms include π–π electron donor–acceptor interactions, hydrogen bonding, and electrostatic interactions [46]. Current biochar research focuses on three areas: the optimization of production parameters (pyrolysis temperature, duration, and atmosphere); surface modification techniques (magnetic particle incorporation, heteroatom doping, microbial loading); and adsorption performance evaluation under various conditions (pH, temperature, coexisting ions). Practical applications face unresolved challenges. Environmental interference reduces field effectiveness while secondary pollution risks create implementation barriers. Therefore, a significant knowledge gap persists between laboratory studies and real-world deployment.

This manuscript provides a comprehensive review of biochar modification schemes as well as the effectiveness and removal mechanisms of biochar for TC in water. The scope, aims, and objectives of the manuscript’s exploration include (i) biochar surface characteristics; (ii) methods for modifying biochar to remove TC; (iii) analysis of the effects of biochar on the removal of TC from water; (iv) mechanisms of modified BC adsorption to TC; and (v) present knowledge gaps. These results emphasize the great potential of biochar for the deep purification of TC-contaminated water. The manuscript fills the key gaps in the field of TC removal by biochar, from basic mechanisms to engineering translation. In addition, this work highlights current challenges in scaling up biochar applications. These challenges include secondary pollution risks (e.g., heavy metal leaching, toxic persistent free radicals), conflicts between economic and environmental benefits (high energy costs for pyrolysis), and performance gaps between laboratory tests and real-world water treatment. We propose targeted solutions to address these limitations. These recommendations aim to advance biochar technology development and practical implementation.

2. Biochar Surface Characteristics

The surface of biochar serves as the critical zone for adsorption and catalytic reactions. Surface properties directly determine the pollutant removal efficiency. The main characteristics are specific surface area, pore arrangement, types of functional groups, and how charges are distributed. The active site distribution depends on the specific surface area and pore count. A larger surface area and more developed mesopores expose more catalytic sites. These properties increase organic pollutant adsorption by enhancing π–π electron donor–acceptor (EDA) interactions [47,48]. Multilayered pore structures (micropore combinations) provide additional adsorption sites and diffusion pathways [49]. For example, the surface area of ZnCl₂-treated wheat straw activated carbon increased from 515 to 2944 m²/g. The pore volume of biochar expands to 1.33 cm³/g, creating a synergistic micromesoporous structure that increases its adsorption capacity [50]. The degree of the graphitization of biochar (as measured by Raman ID/IG) regulates π–π EDA interactions. Pyrolysis above 500 °C strengthens the effects of EDA. High temperatures promote aromatic ring condensation and graphite microcrystal growth. This process lowers the ID/IG and enhances π-electron delocalization [51]. Chen et al. reported that pyrolysis at 800 °C reduced the number of oxygen-containing functional groups in woodchip biochar by 40%. The dominant adsorption mechanism shifted from hydrogen bonding to π–π interactions [52]. The surface charge influences pollutant binding via electrostatic forces. Positively charged surfaces attract negatively charged contaminants when the solution pH falls below the zero-charge point (pHpzc) of biochar [53]. Deprotonation enhances the electron transfer efficiency. The generation of PFRs also improves catalytic oxidation processes [54].

The surface properties of biochar depend on the type of raw material and the conditions of preparation. The composition of the feedstocks determines the initial pore structure and chemical characteristics of biochar. Feedstocks with high lignin contents (e.g., wood) produce biochar with highly aromatic structures and low ID/IG values, indicating a high degree of graphitization, which enhances electron donor–acceptor (EDA) interactions [55]. The hydrophobicity of biochar is correlated with the O/C ratio and the (N+O)/C ratio of the material. A lower ratio corresponds to greater hydrophobicity, which favors the adsorption of nonpolar pollutants [56]. The specific surface area of biochar varies with the type of feedstock. Woody biomass-derived biochar achieves specific surface areas of 2.35 to 1081 m²/g under optimized pyrolysis conditions. Feedstocks rich in cellulose and lignin (e.g., macadamia nut shells, pine sawdust) yield biochar with specific surface areas exceeding 300 m²/g under suitable preparation protocols [57]. The functional group composition of biochar reflects the nature of the raw materials. Cellulose-rich feedstocks favor the formation of oxygen-containing functional groups, whereas lignin-rich feedstocks retain aromatic functional groups [58]. The pyrolysis temperature controls the surface properties of biochar. Temperatures below 500 °C retain oxygen-containing functional groups but result in low specific surface areas. Temperatures above 500 °C promote graphitization at the expense of functional group loss [59]. For example, the activation of poplar biochar-derived carbon with KOH at 300 °C increases the content of oxygenated functional groups (SOFGs). The adsorption capacity of tetracycline increased from 4.30 to 21.17 mg/g. At 700 °C, the specific surface area increases from 106.80 to 337.81 m²/g, but the adsorption performance decreases [60]. Extended residence times during pyrolysis increase the development of micropores and increase specific surface areas [61]. The preparation of biochar involves environmental risks. These include the leaching of heavy metals, the generation of persistent free radicals (PFRs), and the formation of toxic compounds such as polycyclic aromatic hydrocarbons (PAHs) [62].

3. Methods for Modifying Biochar to Remove TC

3.1. Physical Modification

3.1.1. Heat-Treated Micropores

Heat treatment is a key method used to modify biochar surface properties. This process improves critical physicochemical properties including the specific surface area and pore volume. It also adjusts the distribution of surface functional groups. Studies indicate that biochar prepared in a CO₂ atmosphere has better pore connectivity than biochar produced in a N₂ atmosphere. CO₂-treated biochar also has increased hydrophilicity and surface acidic functional group density [63]. For example, Zhu et al. treated biochar with a N₂/air mixed atmosphere at 600–800 °C. This increased its mesopore specific surface area to 316 m²/g (3.8 times greater than that of untreated biochar). The mesopore volume also increased to 0.284 cm³/g. These changes increased the TC adsorption capacity by 5.5–9.2 times [64]. In another study, softwood sawdust-derived biochar was pyrolyzed at 700 °C and its adsorption capacity reached 96.1 mg/g. This high performance was attributed to two factors: a well-developed pore structure and abundant oxygen-containing functional groups. The pore structure and functional groups work together to enhance the adsorption mechanism. These interactions include hydrogen bonding and π–π electron donor–acceptor interactions [65].

3.1.2. Ball Milling

The ball milling technique reduces the biochar particle size to the nanometer scale. This process uses mechanical energy to break the chemical bonds, optimizing the pore structure and surface properties. It enhances the specific surface area, pore size distribution, and density of active sites on the particle surfaces and edges. For example, Zhang et al. studied crayfish shell-derived biochar. When the pyrolysis temperature increased to 800 °C, the specific surface area increased from 127.9 m²/g to 289.7 m²/g (a 2.26-fold increase). The modification increased the TC adsorption capacity by approximately 50% (39.1 mg/g to 60.7 mg/g) [66]. Yu et al. demonstrated that ball milling increased the number of oxygen-containing functional groups. The hydroxyl and carboxyl group contents were 3–5 times greater than those of the untreated biochar. This led to an increase in the TC adsorption capacity from 20.1 mg/g to 103.7 mg/g (a 4.2-fold increase) [67]. Jiang et al. further confirmed these findings. Ball milling improved the sawdust biochar adsorption capacity by approximately four times. This enhancement was correlated with an optimized surface charge distribution and improved particle dispersion stability [68].

3.1.3. Steam Activation

Steam activation improves biochar adsorption performance by modifying the surface properties. This method focuses on enhancing the SSA and pore structure [69]. For example, bamboo-based biochar treated with steam activation at 500 °C presented an increase in SSA from 156.2 m²/g to 312.5 m²/g. Its TC adsorption capacity reached 98.66 mg/g [70]. The improved adsorption arose from synergistic pore structure optimization. Micropores trap pollutants via size-sieving effects, whereas mesopores increase the mass transfer efficiency [71,72]. However, steam activation faces challenges. High energy consumption and pore clogging in humid environments limit its efficiency. Future studies could balance the adsorption performance and energy costs by combining chemical activation (e.g., KOH-assisted methods) or adjusting the steam partial pressure. Additionally, integrating steam activation with other techniques, such as metal loading or heteroatom doping, may lead to the development of biochar adsorbents with greater practical potential.

3.2. Chemical Modification

3.2.1. Acid–Base Modification

Acid–base modification enhances carbon material properties through ion exchange and chemical etching. This method improves the SSA, pore volume, and surface chemistry. Acid treatments (e.g., H₃PO₄ and H₂SO₄) selectively remove ash and introduce oxygen-rich functional groups. For example, H₃PO₄-activated oily tea husk carbon (pyrolyzed at 600 °C) presented an SSA increase from 384.2 m²/g to 1392.4 m²/g. Its pore volume increased 3.2-fold, and the TC adsorption capacity increased 6.8-fold to 451.5 mg/g [73]. Similarly, H₂SO₄-activated rice husk carbon presented a 36% increase in SSA and a 137% increase in adsorption capacity due to enhanced microporosity [74]. Acid modification also alters the surface chemical properties. The hydroxyl group density increases by 3–5 times, and the carboxyl group content increases by 2.8–4.6 times. These changes increase the surface positivity and pollutant affinity [75]. For example, KOH/MnCl₂ activation of banana peels resulted in 339-fold SSA expansion (from <1 m²/g to 338.6 m²/g) and a TC adsorption capacity of 295 mg/g [76]. KOH-activated rice husk carbon presented a 3.5-fold increase in adsorption (Q_max = 58.82 mg/g). The increased number of oxygen-rich surface groups and improved mesopore proportions drove this increase. The surface carbon content of activated biochar increased from 71.6% to 83.2% [74]. These improvements result from synergistic tuning of the physical structure (SSA, pore size) and chemical traits (functional groups, surface charge) of activated carbons, enabling precise property optimization for adsorption applications.

3.2.2. Heteroatom Codoping Modification

Heteroatom codoping modifies biochar by introducing two or more heteroatoms (e.g., N, S, B, P, or Fe). This method enhances the surface functional groups and active site density. The added heteroatoms increase biochar alkalinity and reduce electron density on its carbon surface, improving adsorption performance [77]. For example, N/S-codoped biochar presented a TC adsorption capacity of 156 mg/g, which was 5.4-fold greater than that of pristine biochar (29 mg/g). This improvement resulted from enhanced surface alkalinity and reduced electron density [73]. Similarly, the Fe/S-codoped biochar achieved an equilibrium adsorption capacity of 174.06 mg/g because the Fe–S and Fe–O bonds formed uniform metal oxide clusters. This capacity was 5.3 times greater than that of unmodified biochar [78].

3.2.3. Magnetic Modification

Magnetic biochar adsorbents enhance the biochar properties by loading iron-based nanoparticles (e.g., Fe₃O₄). These modifications improve the SSA, pore volume, and surface characteristics. For example, FeCl-modified pine sawdust biochar resulted in a 51-fold increase in SSA, increasing from 13.08 m²/g to 666.22 m²/g [79]. Similarly, magnetically modified wood chip biochar increased the pore volume from 0.003 cm³/g to 1.301 cm³/g, with an adsorption capacity of 423.7 mg/g [52]. Iron-based nanoparticles also increase the accessibility of adsorption sites. Fe–O bonding modulates the surface charge, improving pollutant interactions [80]. However, excessive iron oxide loading can clog pores and reduce performance. For example, FeCl₃-modified biochar experienced an SSA reduction due to Fe³⁺-induced pore structure damage [81]. Overloading may also lower the adsorption capacity [82,83]. Therefore, it is necessary to balance the quality of the iron source with the retention of the pore structure during the modification process to achieve synergistic optimization of the adsorption performance and separation efficiency.

The adsorption performance of various biochars for tetracycline removal is systematically compared in

Table 2. The removal effects of different modification methods on TC.

Biochar Modification Methods		Biochar Raw Material	Modifying Reagents	Pyrolysis Temperature	SSA (m2/g)		Qmax (mg/g)		Initial Concentration of TC mg/L	Removal Rate of TC (%)	Refs.
					Before Modification	After Modification	Before Modification	After Modification
Raw biochar		Pine bark	/	800	683.33	/	58.47	/	50	89.5	[84]
		Rice straw	/	400	6.72	/	8.246	/	100	/	[85]
		Rice straw		600	21.69		14.185		100	/	[85]
		Swine manure	/	400	5.17	/	5.886	/	100	/	[85]
		Swine manure	/	600	10.56	/	8.169	/	100	/	[85]
		Cow manure	/	700	31.23	/	11.793	/	50	/	[86]
		Grape leaves	/	900	234.201	/	52.858	/	30	/	[87]
		Waste Auricularia auricula dregs	/	300	2.55	/	2.02	/	10	41.22	[88]
		Waste Auricularia auricula dregs	/	500	18.95	/	3.27	/	10	80.81	[88]
		Waste Auricularia auricula dregs	/	700	46.56	/	4.13	/	10	92.00	[88]
		Food and garden waste	/	300	20	/	2.98	/	20	/	[89]
		Food and garden waste	/	300	/	/	4.68	/	40	/	[89]
		Food and garden waste	/	300	/	/	5.78	/	60	/	[89]
		Food and garden waste	/	300	/	/	8.23	/	80	/	[89]
		Food and garden waste	/	300	/	/	9.45	/	100	/	[89]
Physical modification	Ball milling	Crayfish shell	/	800	127.9	289.7	39.1	60.7	100	88.0	[90]
		Soybean residue	/	800	398.81	863.1	/	/	50	77.35	[91]
		Mulberry	/	500	0.81	12.62	20.1	103.7	50	80.71	[67]
		Coconut shell biochar	/	900	1263	1057	/	/	20	69.3	[66]
		Iron-rich paper sludge	/	700	/	/	/	/	20	96.4	[92]
		Iron-rich paper sludge	/	700	/	/	/	/	200	86.1	[92]
		Wheat straw	/	700	/	/	/	/	20	82.47	[93]
		Sawdust	/	600	166.95	219.24	/	41.08	50	82.64	[81]
	Heat-treated	Softwood (Pinus radiata) sawdust	/	400	363	586	10.4	96.1	200	/	[64]
		The waste hydrochar	/	700	7.16	316	/	28.77	50	/	[94]
		Poplar saw dust	300 °C air	220	/	/	6.29	33.32	100	/	[65]
		Poplar saw dust	500 °C air	220	/	/	6.29	196.7	250	/	[65]
		Poplar saw dust	700 °C air	220	/	/	6.29	60.9	100	/	[65]
		Poplar saw dust	300 °C N2	220	/	/	6.29	6.22	100	/	[65]
		Poplar saw dust	500 °C N2	220	/	/	6.29	14.00	250	/	[65]
		Poplar saw dust	700 °C N2	220	/	/	6.29	22.21	100	/	[65]
	Steam activation	Bamboo	/	500	1.22	2.12	/	98.66	/	95.75	[70]
		Beetroot residue	/	850	/	821	/	288	200	/	[95]
		Ectoparasitic shell	/	850	/	829	/	28	200	/	[95]
		Coconut shell	/	850		1138	/	133.1	200		[95]
Chemical modification	Acid–base modification	Swine manure	H3PO4	700	227.56	319.04	109.5	141.5	120	/	[96]
		Rice straw biochar	H3PO4	700	369.26	372.2	132.7	167.5	120	/	[96]
		Camellia oleifera shell	H3PO4	600	384.2	1392.4	33.4	451.5	25	99.5	[97]
		Rice straw biochar	H2SO4	500	34.4	46.8	16.95	23.26	800	/	[74]
		Rice straw biochar	KOH	500	34.4	117.8	16.95	58.82	800	/	[74]
		Chicken feather	KOH	800	412.5	1838.86	228.6	288.33	50	99.65	[98]
		Pomelo peel	KOH	600	27.50	2457.37	12.33	476.19	10	85.04	[99]
		Poplar	KOH	300	12.32	111.39	4.30	21.17	30	/	[60]
		Walnut shells	KOH	900	14.66	1713.87	/	607.00	20	95	[100]
		Alfalfa	NaOH	800	0.68	796.5	30.7	302.37	100	/	[101]
		Pine wood	NaOH	800	1.4	959.9	29.42	274.81	100	/	[102]
		Macadamia nut shells	NaOH	700	/	1524	/	455.33	600	80	[103]
		Coffee grounds	NaOH	500	3.644	116.591	39.22	113.64	100	/	[104]
		Banana skin	KOH	700	3.77	1276.63	51	295	50	98	[76]
		Waste Auricularia auricula dregs biochar	NaOH	700	/	46.56	/	11.9	10	91.10	[105]
	Heteroatom codoping	Corn stover	Fe, S	400	/	/	12.5	505.68	120	/	[106]
		Waste sludge	Fe, S	500	10.68	62.25	32.58	174.06	100	/	[78]
		Bagasse	N, S	300	2.0710	10.0525	16.23	51.1	50	64.2	[107]
		Rice straw	Fe, N	700	177.7154	606.6225	29	156	50	69	[73]
		Waste sludge	Fe, S	500	/	/	51.78	183.01	400	80	[108]
		Poplar tree	N, O, S	900	477.1	542.5	17.4	41.7	40	71.84	[109]
		Rice straw	Mo, S	500	/	176.8	/	249.45	100	85	[110]
		Cabbage leaf	N, P	500	462	1400	/	247	50	97.5	[111]
		Bagasse	La2S3, S, N	600	/	/	/	419.18	200	98.45	[112]
		Tea residue	S, N	220	/	11.689	/	271	100	94.16	[113]
	Magnetically modified	Pine sawdust	FeCl	500	13.08	666.22	33.76	100.74	100	80.6	[79]
		Rice straw	FeCl3	300	28.13	115.5	37.803	98.334	50	69	[88]
		Poplar	NH4Fe(SO4)2	500	165.18	122.12	/	81.13	60	80.1	[114]
		Sawdust	FeCl3	600	166.94	17.19	4.22	66.91	50	78.43	[81]
		Suaeda	Fe	800	35.98	248.8	54.88	70.17	80	83.28	[68]
		Spent mushroom substrate	Fe	500	90.61	100.29	11.06	37.95	100	42.20	[115]
		Spent mushroom substrate	FeSO4·7H2O	500	90.61	296.89	11.06	31.89	100	35.00	[115]
		Spent mushroom substrate	FeCl3·6H2O	500	90.61	160.91	11.06	25.36	100	27.80	[115]
		Spent mushroom substrate	K2FeO4	500	90.61	373.29	11.06	19.63	100	23.82	[115]
		Microalgae	Fe	170	/	128.3	/	95.86	50	93	[116]
		Wood chip	FeCl3	800	1.7	1710	/	423.7	100	96	[52]
		Cotton dust	FeSO4·7H2O and FeCl3·6H2O	450	/	/	/	146.70	100	96.92	[117]
		Water hyacinth	FeCl3⋅6H2O,	700	/	/	/	202.62	50	90.91	[118]
		Waste chicken bones	FeCl3·6H2O, FeSO4·7H2O	500	316.05	328.06	/	98.89	100	96	[119]
		Straw	FeCl3, FeCl2·4H2O	450	7.467	504.994	2.10	15.45	20	/	[120]
		Coconut, pinenut, and walnut shells	Fe, Fe2O3, Fe3O4	500	30.9	365	/	94.2	50	99	[121]

Notes: SSA: specific surface area; Q_max: maximum adsorption capacity; the removal rate (R, %) along with its corresponding adsorption capacity (Q_e, mg g⁻¹) for the adsorbent can be calculated (the formula is shown in the Supplementary Materials).

, which summarizes key parameters including feedstock type, modification method, pyrolysis temperature, specific surface area, maximum adsorption capacity, initial TC concentration, and removal efficiency.

4. Analysis of the Effects of Biochar on the Removal of TC from Water

4.1. Influence of Different Biomass Feedstocks, Preparation Conditions, and Modification Methods on the TC Removal Effect

The TC adsorption performance of biochar depends on the synergistic effects of the type of biomass feedstock, the pyrolysis temperature, and the modification methods. An analysis of 76 studies (Figure 1) revealed a prevalence of chemical modification methods (58% of studies), followed by physical modification (32%) and unmodified biochar (10%), highlighting chemical modification as the primary strategy for optimizing TC adsorption. Agricultural waste (AFW) accounts for 80% of feedstocks in these studies, whereas organic waste (OW) represents 20%. Organic waste (OW) dominated 60% of the studies under high initial TC concentrations (≥100 mg/L). This reflects the high ash content of OW, which is rich in mineral elements such as P, K, and Ca (Figure 1), demonstrating its suitability for high-pollution scenarios.

The chemical composition of raw materials determines the surface properties of biochar. Agricultural waste-derived biochar (e.g., rice husk and straw) has a high carbon content (>60% C) and porous structures with an SSA exceeding 300 m²/g. In contrast, sludge- or manure-based biochar contains abundant metal oxides (e.g., CaO, Fe₂O₃, and Al₂O₃), enhancing TC fixation through complexation and coprecipitation [122]. The pyrolysis temperature influences adsorption by regulating carbonization and functional group abundance [123]. At temperatures > 500 °C, biochar achieved higher values of Q_max and SSA under low TC concentrations (C₀ < 100 mg/L) than the ≤500 °C groups (Figure 2a). This indicates enhanced pore development and active site exposure via high-temperature chemical modification. High-temperature biochar (>500 °C) relies primarily on micropore filling and van der Waals forces but also stabilizes TC through mineral phases (e.g., CaCO₃ and Fe₃O₄) via chemical bonding [124]. The initial TC concentration significantly affects adsorption (Figure 2a–d). At low concentrations (C₀ < 100 mg/L), chemically modified biochar at >500 °C presented a 42% increase in Q_max (p < 0.01) versus unmodified biochar, driven by pore optimization and surface charge modulation. At C₀ ≥ 100 mg/L, the modified biochar maintained superior Q_max and SSA values (p < 0.01), confirming its stable performance in high-concentration environments. For example, H₂SO₄ modification introduces -COOH and sulfur groups, increasing the negative surface charge and enhancing TC adsorption by 2.2-fold at pH 5 via electrostatic interactions [125].

4.2. Influence of the Environmental Factors of the Application Scenarios on the TC Removal Effect

The removal efficiency of TC by biochar is influenced by environmental factors. The solution pH has the most critical effect. The solution pH controls the surface charge of biochar. The solution pH determines the ionization state of TC. The solution pH regulates the surface functional group activity. These groups include carboxyl and hydroxyl groups. This control governs the adsorption equilibrium. Coexisting cations and anions reduce the TC removal capacity, compete for adsorption sites, and change the ionic strength. Coexisting soluble organic matter (DOM) forms complexes with TC. DOM blocks biochar pores, which decreases the adsorption efficiency. These factors are frequently ignored in real water systems. Ionic interference and DOM interact synergistically. This interaction diverges from laboratory predictions under ideal conditions.

4.2.1. pH

Solution pH significantly affects the biochar TC adsorption performance by regulating the biochar surface charge and TC speciation. The mechanism involves three stages. (1) Under acidic conditions (pH < 4), biochar becomes protonated with positive charges. TCs exist mainly as TCH₃⁺ cations (>85%). Electrostatic repulsion between them reduces the adsorption efficiency. (2) At neutral pH values (5–7), biochar becomes deprotonated and has negative charges. TCs primarily exist in neutral (TCH₂⁰) or monoanionic (TCH⁻) forms. Electrostatic attraction and hydrogen bonding jointly increase the adsorption capacity to peak levels. (3) Under alkaline conditions (pH > 8), TC fully deprotonates to TC²⁻. Strong Coulomb repulsion occurs between negatively charged biochar and TC²⁻. This significantly increases the adsorption energy barrier [126]. Experimental evidence confirms this pH dependence. Wang et al. reported that the adsorption capacity increased from 52 mg/g (pH = 2) to 128 mg/g (pH = 5) and then decreased to 85 mg/g (pH > 7) due to electrostatic repulsion [70]. Meng et al. reported the maximum adsorption (243.76 mg/g) at pH = 4.41 for BM-S/ZVI@BC, which was maintained at >200 mg/g at pH = 7.08. Under alkaline conditions (pH 10.77), adsorption decreased to 105.71 mg/g because Fe⁰ oxidation formed Fe(OH)₃ passivation layers [106]. Zhang et al. reported that the TC removal efficiency improved from 55.3% (pH = 2.0) to 69.3% (pH = 6.0) and then decreased to 65.6% (pH = 10.0) as the TC²⁻ proportion increased. These findings demonstrate that pH optimization is crucial for maximizing the biochar adsorption efficiency [66].

4.2.2. Coexisting Cations/Anions

Coexisting cations (e.g., Ca²⁺ and Mg²⁺) and anions (e.g., Cl⁻ and SO₄²⁻) in water reduce the biochar TC removal efficiency through multiple mechanisms [127]. Cations primarily compete with TC for adsorption sites on biochar. This competition weakens electrostatic attraction and cation–π interactions. Chen et al. demonstrated this effect. Increasing the concentrations of Na⁺, K⁺, Mg²⁺, and Ca²⁺ from 0 to 0.1 M reduced the TC adsorption by 12.5%, 13%, 25%, and 37.5%, respectively. The inhibition strength was correlated with the cation properties [128]. A larger hydrated ion radius (Ca²⁺ > Mg²⁺ > K⁺ > Na⁺) and higher charge density (Ca²⁺ = Mg²⁺ > K⁺ = Na⁺) enhance interference. Anions affect adsorption through complexation and conductivity changes. Li et al. reported greater TC removal (99.72%) in mixed SO₄²⁻/Cl⁻ systems (0.1–0.5 M) than in single-anion solutions. This improvement stems from the optimized biochar surface charge distribution through anion synergy [98]. However, HCO₃⁻ has distinct inhibitory effects. Meng et al. reported a 59% decrease in TC adsorption by BM-S/ZVI@BC as the HCO₃⁻ concentration increased from 0 to 0.1 M (128–52 mg/g). The mechanism involves HCO₃⁻ forming passivation layers on iron surfaces. These layers block electron transfer, suppressing Fe²⁺/Fe³⁺-TC complexation [106].

4.2.3. Dissolved Organic Matter

Dissolved organic matter (DOM) widely exists in environments and interferes with biochar TC adsorption through multiple pathways. DOM components (e.g., humic acid, fulvic acid, and polysaccharides) compete with TC for biochar active sites. DOM also reduces the adsorption efficiency via pore blockage and surface charge neutralization [129,130]. Xiong and Bi demonstrated the structural impact of DOM. Corn stover biochar pyrolyzed at 300 °C released biodegradable DOM (BDOM) during adsorption. This caused a 54% decrease in surface area (4.02 to 1.83 m²/g). The H/C ratio increased from 0.80 to 0.91, indicating reduced aromaticity and weakened π–π interactions. DOM effects vary by concentration [131]. Wang et al. reported that low humic acid concentrations (0–10 mg/L) enhanced TC adsorption on Zn-LBC. The adsorption capacity increased by 14.6% (157.36–180.37 mg/g) due to DOM-TC complex bridging. Conversely, BDOM removal significantly improved performance [132]. Yin et al. reported that residual biochar (RMBC) after BDOM removal achieved a Q_max = 328.030 mg/g, which was 25.01% greater than that of the KMnO₄-modified biochar (262.392 mg/g). This enhancement stems from the exposed Mn⁴⁺ sites on RMBC, which form complexes with the phenolic hydroxyl groups of TC [133].

5. Mechanisms of Modified BC Adsorption to TC

The adsorption mechanisms of TC onto biochar include pore filling, π–π interactions, hydrophobic interactions, hydrogen bonding, electrostatic interactions, and surface complexation (Figure 3). Among these interactions, π–π interactions dominate due to strong electron coupling between the aromatic carbon structures of biochar and the benzene rings of TC. This mechanism contributes most significantly under neutral to weakly alkaline conditions. Pore filling plays a key role in high-surface-area biochar but depends on pore size matching with TC molecules. The adsorption efficiency decreases sharply when the biochar pores are smaller than the TC molecules [134]. Electrostatic interactions and hydrogen bonding have limited effects and rely heavily on the solution pH and TC ionization state. At pH > 8, TC deprotonation increases the electrostatic repulsion, reducing adsorption by 20–35%. Hydrophobic interactions and surface complexation act as secondary mechanisms. Hydrophobic interactions contribute up to 15% of the total adsorption in high-ionic-strength environments. Iron-modified biochar enhances adsorption via Fe-O-TC complexation, resulting in a 1.8-fold greater capacity than that of unmodified biochar [120].

5.1. π–π Interactions

π–π interactions serve as a primary driver in TC adsorption mechanisms. The surface functional groups (carboxyl, nitro, and ketone groups) of biochar act as electron acceptors. These groups form π–π electron donor–acceptor (EDA) interactions with the aromatic rings of TC [135]. Experimental evidence confirms this dominance. Jiang et al. analyzed Fe@MBC via Raman spectroscopy. After TC adsorption, the D-band (1307 cm⁻¹) and G-band (1587 cm⁻¹) intensities decreased significantly. The ID/IG ratio increased by 8.14, confirming the major role of π–π interactions. Iron-modified biochar also showed spectral changes, indicating cation–π interactions between the biochar’s π-bonds and Fe²⁺/Fe³⁺ [81]. Wang et al. further validated this mechanism. The C=C characteristic peak (1588 cm⁻¹) of the alkali-activated carbon was reduced by 68% after TC adsorption. This strongly supports π–π EDA interactions as the key pathway [56].

5.2. Electrostatic Interactions

The surface charge of biochar governs its TC adsorption mechanism. Biochar surfaces typically carry negative charges due to the deprotonation of oxygen-containing groups (e.g., carboxyl and hydroxyl groups), enabling the electrostatic adsorption of positively charged pollutants [136]. The zero-charge pH (pHpzc) directly controls the adsorption performance. At pH < pHpzc, biochar surfaces become positively charged, preferentially adsorbing deprotonated TC⁻/TC²⁻. At pH > pHpzc, negatively charged biochar surfaces favor protonated TC⁺ adsorption [137]. Experimental studies have confirmed these principles. Oladipo and Ifebajo reported reduced adsorption under acidic conditions (pH < pHpzc) due to electrostatic repulsion between the positively charged MCB and cationic TC. At pH 8.0, MCB achieved 93.2% TC removal through electrostatic attraction with anionic TCH⁻. Adsorption decreased at pH > pHpzc, as negatively charged MCB repelled TC²⁻ [119]. Yu et al. identified pH 5–9 as the optimal range for electrostatic-driven TC adsorption. Direct evidence comes from surface analysis [113]. Liu et al. detected significant changes in the C=O and C-OH/O-C=O peaks via XPS O 1 s spectra after TC adsorption, confirming electrostatic interactions between TC and these functional groups on IKBCH surfaces [134].

5.3. Hydrogen Bonding

Hydrogen bonding significantly contributes to TC adsorption on biochar. The oxygen-containing functional groups (-OH, -COOH) introduced during biochar modification enable the formation of hydrogen bonds. These bonds provide stable and persistent adsorption due to their high bond energy [138]. Iron-modified biochar (Fe-BC) demonstrates this mechanism. Its hydroxyl and carboxyl groups act as hydrogen bond donors/acceptors through oxygen atoms, enhancing the TC adsorption capacity [115]. Nitrogen-doped biochar also utilizes hydrogen bonding. Mei et al. confirmed that nitrogen functional groups (e.g., -NH₂) participate in TC adsorption via hydrogen bonds [73]. Meng et al. revealed hydrogen-bonding interactions at multiple sites. The hydroxyl (-OH), methyl (-CH₃), and amino (-NH₂) groups of TC function as hydrogen bond acceptors. Interactions with the surface oxygen groups and iron oxides of BM-S/ZVI@BC highlight the essential role of hydrogen bonding in adsorption [106].

5.4. Pore Filling

Pore filling is a critical mechanism in TC adsorption by biochar, with the efficiency governed by the SSA and hierarchical pore structure. The diffusion of TC molecules within modified biochar pores significantly influences the adsorption kinetics [139]. Zhang et al. demonstrated this via diffusion models: TC exhibited a 3.2-fold higher diffusion coefficient in mesoporous biochar (MPBC) than in conventional biochar, which was attributed to rapid mass transfer through mesoporous channels. Pore structure optimization further enhances adsorption [114]. Zhu et al. adjusted the pyrolysis atmosphere (air/N₂ mixture replacing pure N₂) to develop mesoporous carbons (A050, A070), with the TC adsorption capacities exceeding 90–5.5–9.2 times greater than those of microporous biochars [64]. Three mesoporous advantages explain this: (1) the pore size (2–50 nm) matches the TC dimensions (1.2 × 0.6 nm) for efficient capture; (2) open channels minimize the diffusion resistance; and (3) a high SSA ensures dense active sites for stable adsorption. Microporous structures (<2 nm) exhibit limited TC adsorption due to steric hindrance, underscoring the importance of hierarchical porous designs [140,141]. These findings highlight pore engineering as a strategic approach to maximize the biochar TC removal performance.

5.5. Surface Complexation

Surface complexation significantly contributes to TC adsorption on biochar and is strongly related to the density and distribution of oxygen-containing functional groups [142]. Zhang et al. provided direct evidence through C1s XPS analysis. Post-adsorption analysis revealed reductions in C-O-C groups (16.61% to 11.77%) and O-C=O groups (12.6% to 5.65%), with a 0.18 decrease in O/C. These changes confirm the dominance of surface coordination mechanisms [114]. Meng et al. further quantified this mechanism via the use of sulfur-doped ZVI biochar (S-ZVI@BC). The C=O and -OH group contents decreased by 23.4% (37.6–28.5%) and 12.8% (42.1–36.7%), respectively. These reductions were strongly negatively correlated (R² = 0.9949–0.9167) with the maximum TC adsorption capacity (Q_max = 505.68 mg/g), highlighting hydroxyl complexation as a key pathway [106]. Wei et al. identified Fe-O-TC bonds via Fe 2p XPS analysis. A characteristic peak at 531.2 eV confirmed stable coordination between TC and iron oxides (FeO/FeOOH) on the biochar, with a coordination number of 4.2 [143].

5.6. Hydrophobic Interactions

Hydrophobic interactions enhance TC adsorption on modified biochar through multiscale mechanisms. These interactions originate from van der Waals forces between hydrophobic groups (e.g., methyl groups, aromatic carbon clusters) on biochar and nonpolar TC moieties (e.g., benzene rings, methyl groups). Carbon skeleton topology optimization further strengthens the performance. High-temperature pyrolysis promotes the graphitization and aromatization of carbon structures. This process significantly increases material hydrophobicity, increasing the TC adsorption capacity by 3–5 times. Zou et al. demonstrated this enhancement. Compared with raw biochar, biochar pyrolyzed at 700 °C presented a 12.23-fold greater TC adsorption capacity (49.92 to 610.72 mg/g). This improvement correlated with reduced structural defects (ID/IG = 1.06) and increased graphitic carbon content. This behavior can be explained by hydrophobicity-driven phase partitioning [144]. TC migrates from polar aqueous phases to nonpolar biochar surfaces. The partition coefficient (Kd) was strongly negatively correlated with the biochar O/C ratio. Ma et al. validated this relationship. Vinyl chloride-modified biochar (CDBC) reduced the O/C ratio from 0.704 to 0.532. This decreased surface polarity and enhanced hydrophobicity significantly improved the TC removal efficiency. The synergy between hydrophobic and π-π electron donor–acceptor (EDA) effects further optimizes adsorption [145]. Tang et al. achieved 95% TC removal within 9 h via NaOH-modified biochar (NBC). Its maximum adsorption capacity reached 101.01 mg/g, demonstrating that hydrophobic–π interactions were the dominant mechanism [146].

6. Present Knowledge Gaps

6.1. Risk of Secondary Contamination of Raw Charcoal

Despite the potential of biochar for TC removal, its large-scale use faces risks of secondary pollution. Biochar from sludge or industrial waste may retain heavy metals (e.g., Cd and Pb) if the feedstocks are incompletely decomposed during pyrolysis, causing dual contamination during TC adsorption [147]. Wang et al. reported that the Cd content in textile dye sludge biochar increased from 0.793 mg·kg⁻¹ (300 °C) to 0.928 mg·kg⁻¹ (500 °C) and then decreased to 0.139 mg·kg⁻¹ at 600 °C [148]. The presence of persistent free radicals (PFRs) in biochar poses environmental hazards. Light or heat activates PFRs to produce reactive oxygen species (ROS). These ROS directly damage DNA [149] and convert pollutants into toxic by-products (e.g., nitro-PAHs) [150]. Zhang et al. reported plant membrane damage from biochar-borne PFRs: catalase activity increased by 38.1% at 20 g/kg biochar but decreased by 25.2% at 80 g/kg [151]. Lieke et al. demonstrated neurotoxicity in Caenorhabditis elegans exposed to straw-derived biochar (500 °C, 6.37 × 10⁵ C/O-centered radicals), resulting in impaired excretion and chemotaxis [152]. Physical hazards include inhalable dust from biochar grinding and crystalline silica in rice husk biochar pyrolyzed at >550 °C [153]. These risks underscore the need for safety assessments before large-scale biochar deployment.

6.2. Economic–Environmental Efficiency Conflicts

Biochar has potential for pollutant adsorption but faces economic and practical limitations. Compared with activated carbon, its adsorption processes often require longer contact times. This necessitates either increased biochar dosages (increasing costs) or extended treatment durations (reducing efficiency). High production costs stem from energy-intensive pyrolysis (≥500 °C) and chemical activation requirements. For example, U.S.-produced wood-derived biochar costs up to USD 17.80 per kg, typically exceeding the activated carbon prices [154]. Snyder quantified cost disparities: rapid pyrolysis at 500 °C yielded triple the net present value of slow pyrolysis at 400 °C, confirming performance–cost trade-offs. Low-cost agricultural wastes (e.g., straw and nut shells) are affordable but suffer from inconsistent adsorption capacities [155]. Katiyar et al. reported 15–40% performance variations in real wastewater treatments due to fluctuating pollutant concentrations. Additional barriers include limited regeneration efficiency (≤65% capacity retention after 3 cycles) and pH-dependent stability, which results in 15–40% performance variations in real wastewater treatments due to fluctuating pollutant concentrations [156]. Addressing these challenges requires multistrategy optimization: (1) developing low-temperature activation methods (<400 °C), (2) standardizing waste-derived feedstocks through pretreatment protocols, and (3) engineering hybrid biochar composites for enhanced regeneration. These improvements could reduce the production costs by 30–50% while maintaining a ≥80% adsorption efficiency.

6.3. Laboratory–Actual Water Efficiency Deviation

Compared with laboratory conditions, practical water matrices significantly alter the contaminant removal efficiency of modified biochar. For example, Wang et al. reported 84.5% TC and 84.0% Cu²⁺ simultaneous removal by SBC in Yangtze River water. However, the TC removal efficiency at low Cu²⁺ concentrations (0.2 mmol/L) was lower than that at ultrapure water concentrations because of ion competition. At high Cu²⁺ concentrations (2.0 mmol/L), the TC removal efficiency surpassed the laboratory results through complex bridging effects between Cu²⁺ and adsorbents [70]. Similarly, Meng et al. reported reduced TC adsorption capacities for BM-S/ZVI@BC in artificial lake water (266.09 mg/g) and river water (246.29 mg/g) compared with those of deionized water (387.22 mg/g). This reduction stems from three factors: competitive ions, microbial interference, and pore blockage by suspended particles [106]. Wang et al. reported a 50% decrease in TC adsorption (159.64 → 85.64 mg/g) for Zn-LBC in hospital wastewater versus controlled settings. The performance gap likely arises from complex pollutant competition for active sites in real water samples [132].

7. Conclusions and Prospects

This study investigated the surface properties of biochar. These properties affect TC adsorption in water. This study investigated the adsorption performance, mechanisms, and environmental factors. Pristine biochar has a limited TC adsorption capacity. This limitation comes from constraints. For example, pristine biochar has a small particle size and also causes electrostatic repulsion. Physical and chemical modifications significantly improve biochar performance. Compared with the other methods, modified biochar better removed TC. Both physical and chemical modifications enhanced performance. Treatment with H₃PO₄ at high temperatures (>500 °C) was particularly effective and increased the adsorption capacity by 13.5 times, reaching 451.5 mg/g. This showed superior TC removal efficiency. The adsorption mechanisms involved pore filling, π–π interactions (dominant), electrostatic interactions, hydrogen bonding, hydrophobic effects, and surface complexation. π–π interactions primarily originated from strong electronic coupling between the aromatic carbon structures of biochar and the benzene rings of TC. The solution pH critically affects the adsorption efficiency. pH controls the surface charge of biochar as well as the ionization state of the TC. It also controls functional group activity. Coexisting cations such as Ca²⁺ and Mg²⁺ reduce TC removal. Coexisting anions such as Cl⁻ and SO₄²⁻ also reduce removal and cause competitive adsorption. These compounds alter the ionic strength. The DOM indirectly lowers the efficiency. DOM causes complexation and blocks pores. For practical use, medium- to high-temperature (>500 °C) biochar works best. Chemical modifiers such as KOH or H₃PO₄ improve this effect, resulting in optimal TC removal. However, high-temperature pyrolysis uses more energy, increasing consumption by approximately 30% compared with that at low temperatures. Field trials have shown only a 15–20% capacity increase in real wastewater, which is caused by ion interference. This highlights a gap between the laboratory results and real water performance. Scale-up also faces challenges, and secondary contamination risks exist. Examples include heavy metal leaching and toxic radicals. Economic and environmental trade-offs are difficult. High-cost pyrolysis results in only marginal gains. Therefore, there is a need for matrix-specific modification protocols. These aspects need urgent attention and will help bridge the lab-to-industry gap. These aspects need to be considered in the near future to address the challenges being faced.

(i).

Secondary pollution risks of biochar

The secondary pollution risk of biochar requires control across the production and application stages. Agricultural residues such as corncobs and rice husks have lower heavy metal contents than sludge or industrial waste, making them safer feedstocks. Staged pyrolysis can reduce risks: low-temperature treatment (300–400 °C) removes volatile pollutants, followed by brief high-temperature pulses (e.g., 600 °C) to break down free radicals. Surface coatings with humic acid or zeolite seal heavy metals and reduce radical activity through mineral adsorption.

(ii).

Economic–environmental trade-offs

High production costs and high energy consumption limit the large-scale use of biochar. Solutions include developing low-cost, low-energy modification methods. Multifunctional applications, such as combining water treatment with soil remediation or carbon sequestration, improve the economic viability.

(iii).

Lab-to-field performance gaps

The real-world performance of biochar often falls below the laboratory results because of ion competition, microbial interference, and pore blockage. Addressing this requires testing under realistic water conditions and optimizing biochar for complex environments. Smart monitoring technologies could adjust biochar dosing in real-time to maintain stable efficiency.