Innovative Valorization of Wood Panel Waste into Activated Biochar for Efficient Phenol Adsorption

Abstract

Construction and demolition byproducts include substantial amounts of wood panel waste (WPW) that pose environmental challenges. They also create opportunities for sustainable resource recovery. This study investigates the potential of WPW-derived biochar as an efficient adsorbent for phenol removal from aqueous solutions. Biochar was produced via pyrolysis at 450 °C and subsequent activation at 750, 850, and 950 °C. The biochar’s physicochemical properties, including surface area, pore volume, and elemental composition, were characterized using advanced methods, including BET analysis, elemental analysis, and adsorption isotherm analysis. Activated biochar demonstrated up to nine times higher adsorption capacity than raw biochar, with a maximum of 171.9 mg/g at 950 °C under optimal conditions: pH of 6 at 25 °C, initial phenol concentration of 200 mg/L, and biochar dosage of 1 g/L of solution for 48 h. Kinetic and isotherm studies revealed that phenol adsorption followed a pseudo-second-order model and fit the Langmuir isotherm, indicating chemisorption and monolayer adsorption mechanisms. Leaching tests confirmed the biochar’s environmental safety, with heavy metal concentrations well below regulatory limits. Based on these findings, WPW biochar offers a promising, eco-friendly solution for wastewater treatment in line with circular economy and green chemistry principles.

1. Introduction

Accelerated industrialization and urbanization have dramatically increased quantities of lignocellulosic wastes, with a significant portion originating from end-of-life wood panels. Wood waste from construction, renovation, and demolition (CRD) contributes substantially to the growing waste stream. According to Statistics Canada, approximately 4 million tonnes of CRD waste were generated in Canada in 2020, and possibly more depending on tracking methods [1]. Of particular concern are wood panels bonded with phenol-formaldehyde (PF), urea-formaldehyde (UF), melamine formaldehyde (MF), or isocyanate resins, which leach toxic chemicals when dumped in landfills, for additional environmental hazards [2].

Thus, end-of-life wood panels, often laden with toxic compounds such as formaldehyde and ammonia, pose significant environmental and health risks. These compounds leach into soil and water systems, resulting in long-term contamination and bioaccumulation that threaten ecosystems. Moreover, when incinerated, they release hazardous gases, including hydrogen chloride (HCl), nitrogen oxides (NO_x), and ammonia (NH₃), exacerbating air pollution. Despite recently implemented energy recovery strategies, these emissions continue to substantially deteriorate air quality. Additionally, when combusted, these materials release gaseous pollutants such as NO_x and carbon monoxide (CO), which contribute to smog formation and soil acidification [3,4,5].

Simultaneously, phenol, an organic contaminant frequently found in effluents from petrochemical and pharmaceutical industries, is classified as a priority pollutant due to its acute toxicity and environmental persistence [6]. Even at low concentrations, phenol poses significant risks: ingestion can cause severe gastrointestinal complications, and in extreme cases, death [7].

Accordingly, national and international regulations enforce stringent limits on phenol concentrations in drinking water. The U.S. Environmental Protection Agency (EPA) and the European Union (EU) have set a maximum allowable concentration of 0.001 mg/L and 0.0005 mg/L, respectively [8]. Health Canada has established a slightly higher threshold of 0.002 mg/L for drinking water safety [9]. In the Canadian province of Québec, municipal regulations limit phenolic compound concentrations in treated wastewater to 0.5 mg/L or less [10]. These strict rules underscore the urgent need to develop treatments to efficiently and effectively remove phenol from contaminated water sources.

Various approaches to phenol treatment in industrial effluents are possible. The viable options include biological, chemical, and physical methods, each with its advantages and limitations. Biological processes based on phenol biodegradation by specialized microorganisms such as Rhodococcus pyridinivorans and Acinetobacter towneri are economically feasible and generate little secondary waste. However, their effectiveness is generally limited by the toxicity and concentration of phenol in wastewater [11,12]. Chemical methods, including advanced oxidation processes (AOPs) and the Sono-Fenton reaction, promote phenol degradation into less noxious compounds. At the same time, they involve substantial energy inputs and costly reagents (e.g., hydrogen peroxide, iron salts, UV light) [13,14]. Membrane separation technology offers efficient alternatives for phenolic compound separation, but the high cost and risk of membrane fouling limit large-scale applications [15]. In comparison, adsorption emerges as a particularly promising solution due to its simplicity, low-cost efficiency, and ability to treat large effluent volumes. Innovative adsorbents such as modified biomaterials and nanomaterials have demonstrated considerable potential for enhanced phenol removal [16,17,18]. Thus, adsorption offers an attractive and sustainable approach for phenolic water remediation, spurring the development of novel high-performance adsorbent materials.

Biochar, a byproduct of biomass pyrolysis in an oxygen-free environment, is a potent adsorbent thanks to its exceptional physicochemical properties. Its high porosity (with a specific surface area of up to 1500 m²/g), modular carbon structure, and ability to trap organic contaminants via π-π interactions and hydrogen bonds make it an excellent choice for water decontamination [19,20,21]. Recent studies have demonstrated its effectiveness for phenol removal, particularly after chemical activation (e.g., with KOH) or after physical activation via CO₂ gas treatment. Under optimized conditions (pH 6, temperature 25 °C, initial phenol concentration 100 mg/L, biochar dosage 0.5 g/L of solution, contact time 480 min), adsorption capacities up to 458.9 mg/g were obtained [22]. Recently, CO₂-activated wheat straw biochar pretreated with acid washing achieved 471.2 mg/g adsorption capacity under optimized conditions: pH 6.55, temperature 25 °C, initial phenol concentration 2500 mg/L, contact time 240 min, and adsorbent dosage 1 g/L [23]. Bamboo-derived nitrogen-doped magnetic porous hydrochar co-activated with K₂FeO₄ and CaCO₃ efficiently removed phenol from aqueous media. Under optimized conditions (initial phenol concentration 200 mg/L, pH 6, temperature 25 °C, contact time 480 min, adsorbent dosage 0.5 g/L), 211.7 mg/g adsorption capacity was achieved [24]. Chen et al. [25] prepared almond-shell biochar modified with both KOH and EDTA-4Na (mass ratio 1:1:3, pyrolysis at 750 °C). Under key experimental conditions (initial phenol 400 mg/L, 1 g/L adsorbent dosage, 25 °C, 60 min), 161 mg/g adsorption capacity for synthetic solutions was obtained, with 149 mg/g for actual coal-chemical wastewater. With performance comparable to commercial activated carbon, these biochars offer more sustainable and less expensive alternatives. However, the biochar research focuses mainly on agricultural biomass (e.g., rice or wheat straw residues) or forest biomass (e.g., virgin wood chips) while overlooking the potential of WPW [26,27]. Meanwhile, despite its high cellulose (40–60%) and lignin (~30%) content and its abundance, WPW remains an under-exploited source for synthesized high-performance biochar [28]. Notably, while agricultural residues such as rice husks typically contain only 15–20% lignin [29,30], WPW exhibits a substantially higher lignin content at around 30% [31]. Therefore, beyond conventional agricultural biomass, WPW could be used to produce biochar with enhanced carbon yield, aromaticity, and structural stability. Thus, WPW valorization presents a promising pathway for producing advanced adsorbent materials for environmental applications.

This study aims to address the knowledge gap in this area by evaluating the effectiveness of biochar derived from wood panel waste (WPW) to adsorb phenol at variable concentrations. The synthesis parameters of particular interest are the pyrolysis temperature and its impact on the adsorption mechanisms, including Langmuir or Freundlich isotherms and pseudo-second-order kinetics. This study adopts a circular economy framework for a twofold benefit: less WPW deposited in landfills and economical and sustainable water decontamination solutions. This approach aligns with the objectives of Québec’s green economy, particularly the 2030 Québec Green Hydrogen and Bioenergy Strategy (Stratégie québécoise sur l’hydrogène vert et les bioénergies 2030) [32], which aims to ensure optimum use of local natural and energy resources and to develop new energy sectors, including “diversified residual biomass distribution” to address environmental concerns.

2. Materials and Methods

2.1. Preparation of Biochar and Activated Biochar

Wood panel waste (WPW) was sourced from sawmills in the Abitibi-Témiscamingue region, QC, Canada. According to the manufacturer and the literature [33], the WPW contained approximately 88–90% wood particles, 10–12% urea-formaldehyde (UF) resin, and less than 0.5% melamine-formaldehyde (MF) resin by dry weight. The collected residues were ground and milled to <1 mm particle size to ensure homogeneity. Further details on these processes are available in previous studies [34].

Biochar samples were prepared using CarbonFX fast pyrolysis (Airex Energy Inc., Bécancour, QC, Canada) at up to 250 kg/h process capacity. First, the biomass moisture content in the WPW particles was reduced to below 40%. They were then subjected to pyrolysis at 450 °C for 2 s in a cyclonic fluidized bed reactor under a low-oxygen atmosphere. The resulting biochar was then transferred to a custom-designed furnace for physical activation at temperatures of 750, 850, and 950 °C under continuous CO₂ and N₂ flow at a rate of 3 L/min. The biochar and its activated forms were labeled as BWP, BWP750, BWP850, and BWP950, respectively.

2.2. Physical and Chemical Characterization of Biochar and Activated Biochar

Pore structure was characterized by nitrogen (N₂) and krypton (Kr) adsorption at −196 °C and carbon dioxide (CO₂) adsorption at 0 °C using a Micromeritics ASAP 2460 surface area and porosimetry analyzer (Norcross, GA, USA). Prior to adsorption experiments, WPW, biochar, and activated biochar were degassed under vacuum conditions for 48 h at 80, 105, and 250 °C, respectively. Kr absorption was measured to assess materials with low surface area. Ultra microporosity was determined by CO₂ adsorption, and micro- and mesoporosity were determined by N₂ adsorption.

The adsorption isotherms for N₂ and CO₂ were processed mathematically to obtain key structural parameters. Specific surface area (SBET, m²/g) was determined using Brunauer–Emmett–Teller (BET) analysis applied within the appropriate relative pressure range (P/P₀ = 0.01–0.05) [35]. Micropore volume (Vμ, cm³/g) was estimated with the Dubinin–Radushkevich (DR) equation [36]. Total pore volume (Vt, cm³/g) was calculated based on the quantity of nitrogen adsorbed at a relative pressure of 0.97 [37], and mesopore volume (Vm, cm³/g) was determined as the difference between Vt and Vμ.

The elemental composition (C, H, N, S, O) of all materials was analyzed with a Perkin Elmer 2400 CHNS/O organic elemental analyzer (Waltham, MA, USA) using combustion in a pure oxygen stream. Oxygen content was determined by difference, using Equation (1):

O% = 100 − CHNS%

(1)

2.3. Phenol Quantification by GC-MS

The residual concentration of phenol was determined using gas chromatography coupled with mass spectrometry (GC-MS), following MA. 400—Phé 1.0 (Centre d’expertise en analyse environnementale du Québec—CEEAQ) [38]. Phenolic compounds were initially derivatized with acetic anhydride to increase the volatility, followed by liquid–liquid extraction with dichloromethane. Extracts were concentrated and analyzed using GC-MS operating in selected ion monitoring (SIM) mode to improve sensitivity and selectivity. The method detection limit (MDL) for phenol was 0.05 µg/L. Quality control measures included the use of a surrogate standard (Phenol-d5) to monitor recovery rates during sample preparation and a volumetric standard (Nitrobenzene-d5) for calibration correction during instrumental analysis. Data acquisition and quantification were based on the ion signal ratios between the analyte and the corresponding internal standard to ensure analytical robustness and data reliability.

2.4. Adsorption Kinetics and Isotherm Experiments

Kinetic studies were performed to determine the equilibrium time for phenol adsorption on the biochar. Experiments were conducted in 100 mL beakers containing 50 mL of synthetic phenol solution (200 mg/L) and 0.05 g of adsorbent, for a solid-to-liquid ratio of 0.1 g per 100 mL. The suspensions were agitated on a multi-position magnetic stirrer at 500 rpm and 25 °C, and supernatant samples were collected at predetermined intervals (2, 4, 6, 24, and 48 h).

For the adsorption isotherm experiments, 0.05 g of biochar was added to 50 mL of phenol solutions with initial concentrations ranging from 5 to 200 mg/L at a fixed dosage of 1 g/mL. The mixtures were agitated under the same conditions for an equilibrium contact time of 48 h, as determined from the kinetic results. After equilibration, the samples were filtered using 0.45 µm PTFE syringe filters to remove any remaining solid particles. The residual phenol concentration was then determined using GC-MS.

Kinetic and adsorption isotherms were studied to understand the mechanisms and dynamics of the adsorption processes, specifically for phenol adsorption using activated carbon. These studies help determine the time required to reach equilibrium between the adsorbent (activated carbon) and the synthetic effluent. Several models are used to describe these phenomena, each providing different insights into the interactions between activated carbon and phenol.

Table 1. Summary of kinetic and adsorption isotherm models.

Model	Type	Equation		Description
Pseudo-first-order	Kinetic	${\displaystyle \frac{dqt}{dt}}=k_1\left(q_e-q_t\right)$	(2)	A first-order rate equation for adsorption describing rapid adsorption processes.  Its linearized form is as follows:
				$\ln \left(q_e-q_t\right)=\ln \left(q_e\right)-{\displaystyle \frac{k_{1}t}{2.303}}$	(3)
Pseudo-second-order	Kinetic	${\displaystyle \frac{dqt}{dt}}=k_2{\left(q_e-q_t\right)}^2$	(4)	A second-order rate equation indicating a chemical or homogeneous adsorption mechanism.  Its linearized form is as follows:
				${\displaystyle \frac{1}{q_e-q_t}}={\displaystyle \frac{1}{k_2{q_e}^2}}+{\displaystyle \frac{t}{q_e}}$	(5)
Langmuir	Isotherm	${\displaystyle \frac{1}{q_e-q_t}}={\displaystyle \frac{1}{q_{max}}}+{\displaystyle \frac{t}{{bq}_{max}}}Ce$	(6)	A monolayer adsorption model that assumes localized adsorption without interaction between adsorbate molecules.
Freundlich	Isotherm	$Log\left(qe\right)=Log\left(k_f\right)+nLog\left(Ce\right)$	(7)	A model for heterogeneous adsorbent surfaces where Kf and n are coefficients describing the intensity and nonlinearity of adsorption.
Elovich	Kinetic/chemical mechanism	$\log \left(q_e-q_t\right)=\log \left(q_{max}\right)-\alpha .t$	(8)	An equation that describes chemical adsorption, suitable for systems with heterogeneous surfaces.
Temkin	Isotherm	$qe=k_1\ln (k_2)+k_1\mathrm{l}\mathrm{n}\left(c_e\right)$	(9)	A linear model in which the heat of adsorption decreases with molecular interactions, where k1 is related to the heat of adsorption.

Where q_e: adsorbed amount at equilibrium (mg/g); q_t: adsorbed amount at any time t (mg/g); Ce: concentration of adsorbate at equilibrium (mg/L); C₀: initial concentration of adsorbate (mg/L); k₁, k₂, k_f, n, α: specific constants for the models.

summarizes the main kinetic and isotherm models applied to phenol adsorption on activated carbon and their main interpretations [39,40,41,42,43].

2.5. Metal and Heavy Metal Leaching Analysis

Prior to analysis, the samples were digested according to the method described in [44]. The inorganic composition of the resulting filtrates of the WPW, BWP, BWP750, BWP850, and BWP950 samples was determined by microwave plasma atomic emission spectroscopy (Agilent 4200 MP-AES, Agilent Technologies, Santa Clara, CA, USA).

As reconstituted WPW can leach when deposited in landfills, the Toxicity Characteristic Leaching Procedure (TCLP) [45] was applied to melamine WPW, pyrolyzed biochar, and activated biochars to simulate potential leaching. The TCLP determines the mobility and toxicity of inorganic contaminants. Materials were first digested, and the pH was then measured according to two different leaching solutions: (i) leaching medium 1 at pH < 5.0 and (ii) leaching medium 2 at pH > 5.0. The WPW sample and its biochar had pH equal to 1.88 and 3.28, respectively, justifying the use of leaching solution 1, while biochars activated at 750, 850, and 950 °C had pH equal to 6.96, 8.01, and 8.81, respectively, justifying the use of leaching solution 2. The resulting leachate was filtered and analyzed by MP-AES.

3. Results and Discussion

3.1. Characterization of Reconstituted Wood Panels, Biochar, and Activated Biochar

The elemental, textural, and physicochemical analyses of wood panels and their components, as well as biochar and activated biochar, are comprehensively described in a recent study [46]. As illustrated in Figure 1, thermal activation of biochar results in substantial changes in the textural characteristics, such as increasing specific surface area (SBET) and total pore volume (Vt) with increasing temperature. Results show that raw biochar (BWP) has a low specific surface area (99 m²/g), which increases with temperature to values of 450, 682, and 866 m²/g after activation at 750, 850, and 950 °C, respectively. These increases can be attributed to porosity development due to thermal degradation of volatile compounds and structural reorganization of amorphous carbon. Furthermore, the increased volume of micropores (Vμ) and mesopores (Vm) indicates that thermal activation promotes the creation of additional adsorption sites, improving the material’s ability to capture organic and inorganic contaminants in solution [47,48]. This porosity development is essential for adsorption applications, as it increases surface interactions between the material and targeted molecules [49]. The strong correlation between increased specific surface area and improved adsorption performance has been confirmed in several studies of thermally or chemically activated biochars [50,51,52].

The elemental composition of activated biochar undergoes a series of changes in response to temperature variations. Figure 2 shows that carbonization increases with increasing temperature while volatile elements are eliminated. The carbon content increases from 69.2% for BWP to 81.2% for BWP950, reflecting increased condensation of carbon structures and aromatization of the graphite-like network [53,54]. Concurrently, the oxygen content decreases substantially from 23.4% to approximately 13.4% for BWP950, indicating loss of oxygenated functional groups under the influence of pyrolysis [55]. This transformation enhances the activated biochar’s hydrophobicity and chemical stability, which are critical parameters for effective adsorption [56,57]. Moreover, the hydrogen and nitrogen content are reduced due to the volatilization of non-refractory organic compounds and heteroatoms, which typically occurs during high-temperature activation of carbonaceous materials [58]. These chemical changes directly influence the adsorption performance of biochar by modifying its interactions with organic pollutants in aqueous systems.

To further investigate the thermal behavior and volatile emissions from the reconstituted panels, thermogravimetric analysis coupled with gas chromatography–mass spectrometry (TGA-GC/MS) was performed [46]. The results revealed the release of various nitrogen-containing and potentially toxic gases such as ammonia (NH₃), nitric oxide (NO), and trimethylamine (TMA). Notably, highly hazardous species like isocyanic acid (HNCO) and hydrogen cyanide (HCN), often reported in slow pyrolysis of wood-based composites, were not detected under the fast pyrolysis conditions applied in this study. This can be attributed to the short residence time and rapid heating rates associated with fast pyrolysis, which limit secondary reactions responsible for generating these toxic compounds.

Thus, the two-step thermochemical approach (fast pyrolysis followed by activation) not only produced biochars with favorable surface properties for gas adsorption but also mitigated the formation of highly toxic nitrogenous volatiles, thereby enhancing the environmental benefits of using fast pyrolysis to valorize reconstituted wood waste.

3.2. Kinetics and Isotherms of Phenol Adsorption

As shown in Figure 3, biochar (BWP) has a relatively modest phenol adsorption capacity, with increasing efficiency as a function of contact time and initial concentration up to a certain threshold. At low concentrations (5–20 mg/L), the maximum adsorption capacity after 48 h varies from 1.8 to 4.9 mg/g. At 100 mg/L, higher adsorption is observed (18.3 mg/g at 48 h), although a significant drop is recorded at 24 h, suggesting partial desorption or saturation of the active sites on its surface. At very high concentrations (200 mg/L), the efficiency becomes negligible or even zero. However, thermal activation of BWP at 750, 850, and 950 °C induces significant improvements in adsorption capacity, with an increasing trend as a function of activation temperature and contact time. The best-performing material, BWP950, shows 171.93 mg/g adsorption capacity after 48 h, followed by BWP850 (142.25 mg/g) and BWP750 (51.89 mg/g). These improvements correlate with significant increases in specific surface area, from 99 m²/g for raw BWP to 450, 682, and 866 m²/g for BWP750, BWP850, and BWP950, respectively, as well as increased total pore volume, including both micropores and mesopores. These significant improvements can be correlated with the higher specific surface area and pore volumes resulting from the increased volatilization of non-carbon organic constituents and the development of more microporous structures at high temperatures. Furthermore, the better adsorption yields as a function of time suggest that the process follows a dominant intraparticle diffusion mechanism, where electrostatic interactions and π-π bonds with the aromatic rings of the phenol play a critical role [59,60].

Figure 3 highlights the critical influence of thermal activation on the phenol removal efficiency of biochars. At low initial concentrations (5–20 mg/L), adsorption efficiency significantly increases with increasing pyrolysis temperature. For instance, while the raw biochar (BWP) achieves only 16.4% phenol removal after 2 h of contact at 5 mg/L, the biochars thermally activated at 850 °C and 950 °C (BWP850 and BWP950) exhibit efficiencies exceeding 97%, reaching up to 99.2% after 48 h. Similar trends are observed at 10 and 20 mg/L, where the activated biochars demonstrate near-complete adsorption within 24 h (99.8%).

At higher concentrations (100–200 mg/L), the performance of raw biochar remains limited. In contrast, the activated biochars maintain high adsorption capacities, with BWP950 achieving up to 93.3% at 100 mg/L and 86.3% at 200 mg/L after 48 h. These results underscore the pivotal role of activation temperature in enhancing porous structure, specific surface area, and active site availability, thereby improving affinity for aromatic compounds such as phenol. Increasing the biochar activation temperature leads to substantially greater phenol adsorption capacity, as evidenced by numerous recent studies [61,62]. Thus, high-temperature activation promotes the development of optimal porous structures in biochar, thereby enhancing phenol adsorption.

The kinetic modeling results (

Table 2. Kinetic parameters obtained from phenol adsorption on biochars.

Biochar	Pseudo-First-Order			Pseudo-Second-Order			Elovich’s Equation
	qe mg/g	K1 h−1	R2	qe mg/g	K2 mg/(gh)	R2	α mg/(gh)	β	R2
BWP750	27.63	0.581	0.67	49.751	4.63 × 10−5	0.80	7.275 × 103	0.282	0.17
BWP850	35.34	−0.071	0.67	142.86	1.16 × 10−6	0.99	5.742 × 105	0.105	0.98
BWP950	34.27	−0.071	0.99	172.41	1.55 × 10−7	1	3.836 × 107	0.112	0.94

) reveal that the pseudo-second-order model best describes the phenol adsorption behavior for all biochars, as evidenced by the high correlation coefficients (R² > 0.99 for BWP850 and BWP950). This suggests that the adsorption mechanism is predominantly governed by chemisorption involving specific chemical interactions such as hydrogen bonding or π-π interactions between phenol molecules and the aromatic structures on the biochar surface. The adsorption rate constant (K₂) significantly increases with pyrolysis temperature, ranging from 4.63 × 10⁻⁵ mg/(g·h) for BWP750 to 1.55 × 10⁻⁷ mg/(g·h) for BWP950. This indicates higher diffusion rates of phenol molecules into the mesopores of the biochar, consistent with the increased mesopore volume from 0.028 cm³/g (BWP750) to 0.139 cm³/g (BWP950).

The adsorption isotherm data (

Table 3. Parameters of phenol adsorption isotherms on biochars.

Biochar	Freundlich			Langmuir			Temkin
	Kf	n	R2	Qmax (mg/g)	RL	R2	K1 L/mg	K2	R2
BWP750	40.28	2.28	0.89	178.57	5.94 × 10−5	0.99	4.78	14.77	0.68
BWP850	33.83	0.03	0.88	144.93	8.21 × 10−5	0.99	16.89	46.52	0.98
BWP950	40.28	2.28	0.88	178.57	5.94 × 10−5	0.99	21.58	38.89	0.94

) fit well with the Langmuir model (R² = 0.99), suggesting monolayer adsorption of phenol molecules on homogenous surface sites of the biochar. The maximum adsorption capacity (Q_max) increases proportionally with pyrolysis temperature, reaching 178.57 mg/g for BWP950 compared to 144.93 mg/g for BWP850. This increase is mainly attributed to the increased specific surface area (SBET = 866 m²/g) and higher carbon content (81.2%) obtained at 950 °C. Higher pyrolysis temperatures promote the carbonization process while reducing oxygen-containing functional groups (from 23.4% [BWP] to 13.4% [BWP950]), resulting in a more hydrophobic surface with a higher affinity for phenol adsorption. This process is explained by the increased graphitization degree and reduced polar functional groups on the biochar surface, which facilitate hydrophobic interactions and π-π electron donor-acceptor (EDA) interactions between phenol molecules and the biochar surface [63].

Table 4. Comparative summary of phenol adsorption by various biochar materials.

Material (Precursor, Activation)	SBET (m2/g)	Qmax (mg/g)	Conditions	Reference
Magnetic Fe–Zn biochar activated with KOH (co-pyrolysis)	1122	458.9	pH 6, 25 °C, 480 min, 0.5 g/L, 100 mg/L	[22]
Wheat straw biochar (HF wash + 10% CO2 activation at 900 °C)	492.6	471.2	pH 7, 25 °C, 240 min, 1 g/L, 2500 mg/L	[23]
Bamboo-derived nitrogen-doping magnetic porous hydrochar coactivated by K2FeO4 and CaCO3	610.5	211.7	pH 6, 25 °C, 480 min, 0.5 g/L, 200 mg/L	[24]
Almond shell-activated biochar (chemical activation with EDTA-4Na/KOH at 750 °C)	1050	161.0	pH 7, 25 °C, 60 min, 1 g/L, 400 mg/L	[25]
Black wattle bark-activated carbon (black wattle bark, ZnCl2-activated, pyrolysis)	414.1	98.6	pH 6.5, 55 °C, 120 min, 1 g/L, 500 mg/L	[64]
Sawdust-activated carbon (physical activation via steam at 900 °C)	1053	158.9	pH 4, 25 °C, 10 min, 0.5 g/L, 100 mg/L	[65]
Sunflower stalk-activated biochar (chemical activation with KOH)	452	333.0	pH 6, 25 °C, 30 min, 0.5 g/L, 100 mg/L	[66]
Rice husk biochar activated with KOH and modified by EDTA-4Na (M-AC)	1368	215.3	pH 5, 25 °C, 10 min, 0.5 g/L, 500 mg/L	[67]
Activated carbon (PET plastic, H3PO4-activated)	655.6	114.9	pH 7, 25 °C, 120 min, 0.4 g/L, 100 mg/L	[68]
Magnetic Fe3O4/ZIF-8 MOF composite (magnetic ZIF-8 adsorbent)	1120.7	129.8	pH 7, 25 °C, 20 min, 0.4 g/L, 70 mg/L	[69]
WPW-activated biochar at 950 °C (BWP950) (without adding a chemical)	866	171.9	pH 6, 25 °C, 48 h, 1 g/L, 200 mg/L	This study

presents a comparative overview of phenol adsorption capacities across various biochars and activated carbons. The adsorption capacities span a broad range, from 98.6 mg/g for ZnCl₂-activated black wattle bark biochar to 471.2 mg/g for CO₂-activated, acid-washed wheat straw biochar. These differences are mainly attributed to variations in precursor composition, surface area, pore size distribution, and activation methods (e.g., chemical vs. physical) as well as adsorption test conditions (pH, adsorbent: effluent dosage, initial concentration).

Notably, the WPW-derived biochar activated at 950 °C (BWP950) demonstrates 171.9 mg/g adsorption capacity and 866 m²/g surface area without the use of chemical agents such as KOH, ZnCl₂, or EDTA. Although this performance is moderate compared to highly engineered adsorbents, the use of reconstituted wood waste as a precursor offers significant economic and environmental advantages. Taken together, the results in Table 4 support the potential of industrial waste valorization via thermal activation to yield biochars with suitable properties for wastewater treatment applications.

Furthermore, whereas most studies of high-performance materials have applied shorter contact times (10–240 min), activated biochars in our study achieved competitive capacity even with a longer equilibrium time (48 h), suggesting strong surface interactions and the potential for continuous flow applications. The extended contact time of 48 h for BWP950 is attributable to the gradual diffusion within its predominant microporous structure, a common characteristic of physically activated biochars [47,48]. Although this characteristic may limit rapid adsorption in batch processes, BWP950 remains applicable in continuous flow systems given adequate residence time [61]. Despite its higher SBET (866 m²/g), BWP950 shows lower adsorption capacity compared to chemically activated biochars because adsorption performance depends on not only surface area but also surface chemistry. Chemically activated biochars possess more oxygen-containing functional groups (e.g., carboxyl, hydroxyl, carbonyl) that facilitate phenol adsorption via hydrogen bonding and π–π interactions [20,21]. The literature acknowledges the importance of these functional groups, as demonstrated by FTIR analyses in previous studies [20,21,25]. These findings confirm that thermal activation alone, when properly optimized, can produce biochars with effective adsorption properties while maintaining process sustainability and simplicity.

3.3. Analysis of Metal and Heavy Metal Leaching Behavior

Reconstituted WPW normally contains heavy metals, which are considered pollutants and which can significantly harm the environment and human health, depending on their concentration [70]. The metals present in the WPW are in raw, pyrolyzed, and activated states.

Table 5. Metal concentrations present in solid wood [71], wood waste furniture [72], and wood panel waste (WPW) and the derived biochar (BWP) and activated biochars (BWP750, BWP850, and BWP950) from this study.

	Concentration (mg/kg)
	Solid Wood [71]	Furniture Wood Waste [72]	WPW	BWP	BWP750	BWP850	BWP950
Al	30 ± 20	480 ± 10	199 ± 5.3	612 ± 39	1000 ± 44	1667 ± 129	2257 ± 127
As	0.01 ± 0.004	0.4 ± 0.03	<0.01	<0.01	<0.01	<0.01	<0.01
Ca	260 ± 5	1590 ± 240	2513 ± 147.4	14,000 ± 1153	14,233 ± 2316	21,233 ± 1021	32,733 ± 896
Cd	0.1 ± 0.02	1 ± 0.1	<5	<5	<5	<5	<5
Co	0.05 ±0.004	1.1 ± 0.1	<1	24 ± 4	87 ± 67	77 ± 3	110 ± 17
Cr	2 ± 1.3	6.7 ± 1.4	7 ± 2.2	64 ± 22	160 ± 67	229 ± 17	80 ± 18
Cu	1.4 ± 1.4	6.4 ± 0.7	<5	11 ± 2	18 ± 6	42 ± 1	44 ± 6
Fe	25 ± 4	280 ± 30	319 ± 4.2	4520 ± 321	9960 ± 750	13,233 ± 451	22,500 ± 3081
Hg	0.01 ± 0.003	0.03 ± 0.04	<0.01	<0.01	<0.01	<0.01	<0.01
K	160 ± 90	340 ± 150	811 ± 24.5	3810 ± 226	5473 ± 136	7063 ± 75	9733 ± 206
Mg	100 ± 30	320 ± 80	261 ± 2.9	2593 ± 704	2183 ±361	3440 ± 240	5640 ± 344
Mn	46.8 ± 7.1	40 ± 3.6	130 ± 5	419 ± 18	348 ± 31	683 ± 30	965 ± 30
Mo	0.02 ± 0.02	0.3 ± 0.03	<0.01	<0.01	<0.01	<0.01	<0.01
Na	20 ± 10	300 ± 120	684 ± 16.9	1957 ± 129	2237 ± 103	2743 ± 67	3393 ±70
P	20 ± 10	100 ± 20	<15	<15	<15	<15	<15
Pb	0.04 ± 0.06	6.3 ± 3.6	<1	<1	<1	<1	<1
Sb	0.01 ± 0.01	2 ± 0.4	<0.01	<0.01	<0.01	<0.01	<0.01
Si	109 ± 3	2150 ± 140	<1	<1	<1	<1	<1
Ti	2.1 ± 0.8	1600 ± 200	<1	<1	< 1	<1	<1
V	1 ± 0.002	0.9 ± 0.04	<0.01	<0.01	<0.01	<0.01	<0.01
Zn	7.7 ± 0.9	69.4 ± 2	19 ± 0.7	101 ± 12	25 ± 1	17 ± 1	7 ± 1

presents the metal analysis results for the different samples. The obtained values are the average of three measurements. The highest metal concentrations in the WPW are for Ca and K (2513 mg/kg and 811 mg/kg, respectively), followed by Na (684 mg/kg) and Fe (319 mg/kg). Concentrations of Cd, Co, Cu, and Pb are below the detection limit for the applied method.

The metal contents found in the WPW have high concentrations compared to those reported in the literature. For example, Moreno et al. [71] determined that uncontaminated solid wood presented lower metal concentrations compared to WPW, probably due to the presence of a melamine-formaldehyde coating bonded to both surfaces of the WPW. Indeed, Moreno and Font [72] found that Si and Ti were the two most abundant metals in furniture wood waste, followed by Ca, K, Na, and Fe, at lower concentrations than those found in the present study: 1590, 340, 300, and 280 mg/kg, respectively. The metal present in WPW could also come from the panel manufacturing process, which involves cutting and/or grinding tools [73].

Table 5 also shows the effects of pyrolysis and activation in terms of changes in metal concentrations. The concentrations of most metals increase after pyrolysis, and more so after activation. This is because endogenous, non-volatile heavy metals from biochar feedstocks remain in the biochar structure during the thermochemical processes [74]. Furthermore, the higher the activation temperature, the higher the concentration of metals in the activated biochar, because metals do not volatilize, whereas organic matter is degraded and removed during pyrolysis and activation [75]. Notably, Ca and Fe are the most abundant metals in the biochar and activated biochar. This could be due to contact between the biochar and activated biochars and the iron equipment during pyrolysis and activation.

Low concentrations of the toxic elements Cr, Cd, Co, Cu, Pb, Mn, and Zn were detected in the different biochars, indicating that they are safe to use as amendments to improve soil properties. In addition, high concentrations of the nutrients Fe, Al, Mn, Ca, Mg, K, and Na, which improve soil fertility, were detected. In general, biochar could improve soil quality and reduce soil ecotoxicity by adsorbing potentially toxic trace elements and organic contaminants from the soil [76]. Moreover, the concentrations of Pb and Cd are below the detection limit of the applied method. Additionally, the pH of the materials should be considered when assessing the environmental impact. While the raw biochar exhibits an acidic pH, the thermally activated biochar shows a more neutral pH, which is more suitable for agricultural applications and less likely to alter or acidify soil systems. This combination of neutral pH and low heavy metal availability further supports the use of activated biochars as sustainable soil amendments with minimal risk of secondary contamination.

Table 6. Concentrations of metals leached from WPW and the derived biochar (BWP) and activated biochars (BWP750, BWP850, and BWP950) using the Toxicity Characteristic Leaching Procedure (TCLP).

	Concentration (mg/kg)
	WPW	BWP	BWP750	BWP850	BWP950	Regulatory Limit (EPA)	Regulatory Limit (CEPA)
Ag	<0.00009	<0.00009	<0.00009	<0.00009	<0.00009	5	0.14
As	0.0047 ± 0.0001	0.0047 ± 0.0001	0.0182 ± 0.0006	0.0324 ± 0.0008	0.0025 ± 0.0002	5	2.5
Ba	0.5423 ± 0.0090	0.343 ± 0.007	0.407 ± 0.0123	0.1823 ± 0.0025	0.383 ± 0.0062	100	100
Cd	0.0044 ± 0.0003	0.0009 ±0.00002	0.0003 ± 0.00005	<0.00009	<0.00009	1	0.5
Cr	0.0060 ± 0.0003	0.0137 ± 0.0003	0.0013 ± 0.00005	0.0004 ± 0.0001	0.0005 ± 0.0001	5	5
Cu	0.0103 ± 0.0004	0.003 ± 0.0001	0.0002 ± 0.0001	<0.00007	<0.00007	5	5
Mn	5.1500 ± 0.0458	5.4367 ± 0.1069	9.1367 ± 0.3156	21.3667 ± 0.4726	34.900 ± 0	-	-
Pb	0.0069 ± 0.0003	0.0001 ± 0.00001	<0.00008	<0.00008	<0.00008	5	-
Se	<0.00044	0.0015 ± 0.0002	<0.00044	0.0022 ± 0.0002	0.0107 ± 0.0007	1	1
Zn	0.556 ± 0.0074	0.5037 ± 0.0115	0.292 ± 0.01	0.0179 ± 0.0011	0.0361 ± 0.0004	-	-

presents the leaching test results for the determination of heavy metals in the leachate along with the limiting concentrations. Very low levels of heavy metals were measured in the leachate, with some below detection limits. For WPW, Mn is the most leached metal, followed by Zn and Ba at concentrations of 5.147, 0.556, and 0.542 mg/kg, respectively. The detected concentrations are below the regulatory limits for hazardous levels in waste materials, as established by the United States Environmental Protection Agency (USEPA) [77] and the Canadian Environmental Protection Act (CEPA) [78]. Therefore, the WPW and the derived biochars would be considered as non-hazardous materials.

4. Conclusions

This study demonstrates the potential of biochar derived from wood panel waste (WPW) as a sustainable and efficient adsorbent for phenol removal from aqueous solutions. Biochar activation at high temperatures (750, 850, and 950 °C) significantly enhanced its adsorption capacity, with the highest experimental adsorption capacity observed for biochar activated at 950 °C (171.9 mg/g) and with a Langmuir predicted maximum adsorption capacity of 178.57 mg/g for BWP950. The adsorption process was best described by the pseudo-second-order kinetic model and the Langmuir isotherm, indicating chemisorption and monolayer adsorption mechanisms. The development of microporous and mesoporous structures during activation played a critical role in improving adsorption performance. Furthermore, leaching tests confirmed the environmental safety of the biochar, with heavy metal concentrations well below regulatory limits. These findings underscore the potential of biochar as a safe and sustainable material for environmental applications. The valorization of WPW into biochar not only addresses waste management challenges, but also provides a cost-effective solution for water treatment, in agreement with the principles of circular economy and green chemistry. Moreover, these findings align directly with the 2030 Québec Green Hydrogen and Bioenergy Strategy [32], which promotes the valorization of local biomass residues such as WPW to develop sustainable materials for environmental remediation and circular economy applications. This study contributes to the knowledge of sustainable waste management and water treatment technologies by offering a practical, leading-edge solution for the dual challenges of waste valorization and environmental remediation.