Evaluating Phosphorus Sorption and Desorption in Agricultural Wastewater Using Designer Biochar Pellets
Agnes Millimouno, Jorge A. Guzman, Wei Zheng, Richard A. Cooke, Maria L. Chu
TL;DR
This study shows that designer biochar pellets can effectively remove phosphorus from agricultural wastewater, reducing runoff that causes algal blooms.
Contribution
The novelty lies in the development and evaluation of designer biochar pellets for phosphorus removal in agricultural effluents.
Findings
DBPs removed 18 to 155 mg kg−1 of phosphorus from different effluents.
SEM and ICP confirmed phosphorus sorption and the role of elements like iron and calcium.
DBPs can lead to soil pore clogging due to phosphorus precipitation.
Abstract
Tile drains enhance crop productivity but also increase phosphorus (P) runoff into nearby water bodies, contributing to harmful algal blooms. This study examines the effectiveness of designer biochar pellets (DBPs) in removing or releasing P from agricultural effluents, soils, or deionized water, respectively. The DBPs are composed of pine sawdust biomass and bentonite clay, pretreated with lime sludge prior to pyrolysis, and subsequently exposed to various wastewater effluents and field conditions. DBP treatment in P removal varied across effluent types, ranging from 18 to 155 mg kg−1. In contrast, P desorption in deionized water ranged from 0.1 to 8.9 mg L−1. DBP extracted from the field after the trial showed contrasting soil phosphorus extraction results, ranging from 0.45 to 0.6 mg L−1 for new and 0.3 to 1.2 mg L−1 for spent, respectively. Furthermore, P extracted from soil before…
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FIGURE 14| Acronyms | Definition |
|---|---|
| SEM | Scanning electron microscopy |
| ICP | Inductively coupled plasma |
| DBPs | Designer biochar pellets |
| N | New pellets |
| U | Spent pellets |
| V1 | Vertical monitoring Well_1 |
| H1 | Horizontal monitoring Well_1 |
| BPS | Batch phosphate sorption |
| N‐BSP | New pellet in batch phosphate sorption |
| EMW | Effluent from the monitoring well |
| WBE | Woodchip bioreactor effluent |
| DFE | Dairy farm experiment |
| N‐DFE | New pellet in dairy farm experiment |
| U‐DFE | Spent pellets in dairy farm experiment |
| N‐WBE | New pellet in woodchip bioreactor effluent |
| U‐WBE | Spent pellet in woodchip bioreactor effluent |
| CME | Cow manure effluent |
| U‐CME | Spent pellet in cow manure effluent |
| N‐EMW_V1 | New pellet in effluent from vertical monitoring Well_1 |
| N‐EMW_H1 | New pellet in effluent from horizontal monitoring Well_1 |
| U‐EMW_V1 | Spent pellet in effluent from vertical monitoring Well_1 |
| U‐EMW_H1 | Spent pellet in effluent from horizontal monitoring Well_1 |
| SSBP | Soil sample before planting |
| SSAP | Soil sample after planting |
| SSAH | Soil sample after harvest |
| EDS | Energy‐dispersive X‐ray spectroscopy |
| Designer biochar pellets analytes following digestion (A) | |||||||
|---|---|---|---|---|---|---|---|
| Sample ID |
|
| N_WBE | U_WBE | U_CME | ||
| Element | mg kg−1 | ||||||
| P | 180 | 81 | 230 | 248 | 194 | ||
| Al | 31,068 | 32,096 | 29,356 | 34,923 | 37,044 | ||
| Ca | 323,343 | 271,309 | 288,657 | 301,993 | 298,728 | ||
| Cu | 6.4 | 131 | 6 | 179 | 97 | ||
| Fe | 14,593 | 16,083 | 12,977 | 17,613 | 16,464 | ||
| K | 1697 | 1249 | 1085 | 2684 | 1797 | ||
| Mg | 31,937 | 33,82 | 29,390 | 37,759 | 37,16 | ||
- —U.S. Environmental Protectional Agency (EPA)
- —Agricultural and Biological Engineering Watershed‐Ecosystem Research Laboratory (WERL)
- —Illinois State Water Survey’s Health and Environment Application Laboratory (ISWS‐HEAL)
- —Materials Research Laboratory (MRL) at the University of Illinois
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Taxonomy
TopicsPhosphorus and nutrient management · Soil and Water Nutrient Dynamics · Constructed Wetlands for Wastewater Treatment
Introduction
1
Biochar has attracted attention for its potential to enhance soil health, reduce pollution, and support ecosystem restoration. Its benefits are particularly evident in nutrient‐poor tropical soils, where biochar produced from nutrient‐rich feedstocks, such as manure, has significantly increased crop yields (Jeffery et al. 2011; Dai et al. 2020). Biochar is typically produced via controlled pyrolysis, in which biomass is carbonized under low‐oxygen conditions at high temperatures (Zhao et al. 2017). This process, often enhanced by incorporating external elements, produces designer biochar with porous carbon structures and high surface areas. Incorporating metal‐rich materials during pyrolysis or post‐pyrolysis can help tailor the properties of designer biochar. For instance, Yang et al. (2021) investigated biochar derived from lime‐slag‐pretreated sawdust for the effective capture and recovery of dissolved phosphorus from water, a major contributor to harmful algal blooms.
Furthermore, Katuwal et al. (2023) developed a novel designer biochar from lime‐sludge‐treated sawdust. They investigated its sorption kinetics and mechanisms for dissolved phosphorus. Their work examined the sorption mechanisms and performance of this designer biochar, compared it with other materials, and considered its potential as a nutrient‐rich fertilizer, demonstrating that agricultural effluent treatment can be integrated into a sustainable nutrient management approach.
Although metal‐incorporated biochar can adsorb and recycle phosphorus (Dai et al. 2017), the practical implementation of biochar‐based systems remains challenging. These include nutrient retention and availability issues, as biochar's physical and chemical properties directly influence the accessibility of plant nutrients (Zhao et al. 2021). Additionally, biochar–microbe interactions play a crucial role in modulating nutrient cycling, soil health, and the controlled release of nutrients from nutrient‐enriched biochar (Wang et al. 2020). Recent research highlights the need for a deeper understanding of these processes, particularly regarding the long‐term behavior of biochar in soil environments. Key knowledge gaps include its stability, decomposition dynamics, and potential leaching of organic compounds, all of which are critical for a comprehensive evaluation of the environmental risks and benefits of biochar application (Zhang et al. 2014). Notably, one promising area for biochar utilization is the mitigation of phosphorus (P) pollution from agricultural wastewater, including tile drain effluents.
Surface runoff and subsurface (tile) drainage are major drivers of water pollution from agricultural lands (Melland et al. 2008; King et al. 2014; Smith et al. 2015). Phosphorus, whether in dissolved form or bound to soil particles, primarily contributes to water quality deterioration, fueling eutrophication and triggering harmful algal blooms in aquatic systems (Kleinman et al. 2015). In southern Australia, surface runoff contributed up to 0.25‐kg P ha^−1^ yr^−1^, while tile drainage contributed substantially less, peaking at 0.027‐kg P ha^−1^ yr^−1^ (Melland et al. 2008). Long‐term data from central Ohio indicated average total phosphorus loads of 0.98 kg ha^−1^ yr^−1^ at the watershed outlet and 0.48 kg ha^−1^ yr^−1^ from tile drains (King et al. 2014), highlighting the significant role of subsurface tile flow. Recent studies have further demonstrated that tile drainage is a persistent source of dissolved reactive phosphorus (DRP). Williams et al. (2022) reported DRP concentrations of 0.01–5.62 mg L^−1^ in northeastern Indiana, with daily DRP loads reaching up to 0.1 kg ha^−1^ and cumulative loads exceeding 1.0 kg ha^−1^. Therefore, developing effective management strategies to mitigate nutrient losses from agricultural lands is critical.
Previous studies have highlighted the potential of biochar for phosphorus removal, with both batch and continuous‐flow evaluations examining the sorption mechanisms. MgO‐modified biochar shows high efficiency and stability (Chen et al. 2023), while comparisons of metal oxide‐modified biochars (Al, Ca, Fe, La, and Mg) reveal factors affecting sorption capacity and kinetics (Wang, Liao, et al. 2021). Iron‐modified biochar, in particular, benefits from increased surface area and porosity, enhancing nitrogen and phosphorus adsorption (Wu et al. 2024). However, most research relies on synthetic solutions under controlled conditions, overlooking competing ions and environmental variability, which limits their real‐world applicability. Furthermore, it is common to find experimental biochar setups that use P concentrations outside realistic ranges for drainage effluents, thereby obscuring their real‐world applications due to nonlinear sorption dynamics occurring at high P concentrations (Penn 2021).
This study integrated laboratory experiments and field‐based evaluations to assess the performance of designer biochar pellets (DBPs) under more realistic environmental conditions. By employing various wastewater effluents that simulate irrigation and rainfall scenarios, we investigated the effects of a key environmental factor, the dynamics of pH, on the P sorption capacity of the DBPs. Furthermore, we evaluated the potential for P recovery and the effectiveness of DBPs as nutrient‐rich soil amendments in agricultural applications. This approach provides critical insights into the practical applicability of DBPs and highlights their role in sustainable P management. The primary objective of this research is to evaluate the effectiveness of DBPs in capturing dissolved P from tile drain outflows and other agricultural wastewater effluents, and to explore their potential for reuse as fertilizer amendment. Specifically, this study aimed to (1) characterize the physicochemical properties and elemental composition of DBPs to support a comprehensive evaluation of their performance, (2) assess the P sorption capacity and kinetics of DBPs in different agricultural wastewaters, and (3) investigate P desorption under varying pH conditions to explain pH‐dependent P‐removal mechanisms and their implications for surface water quality. By quantifying the potential of DBPs as a sustainable material for P capture and recovery, this research advances the integration of biochar‐based solutions into circular agricultural practices, thereby improving water quality and promoting nutrient recycling.
Materials and Methods
2
Biochar Preparation
2.1
DBP fabrication follows the procedure described on Yang et al. (2021): A biochar matrix was prepared using dried and milled lime sludge (105°C and < 75 μm) from a drinking water treatment facility in Champaign, IL, mixed with air‐dried and sieved to < 1.0‐mm activated carbon (Darco G‐60, Aldrich Chemical Co.) and pine sawdust (Pinus spp., University of Illinois Energy Farm). Pine sawdust and lime sludge were mixed at a 1:4 (w/w) ratio and pyrolyzed to 450°C at 15°C min^−1^ under a nitrogen gas stream, with a 2‐h holding time. The biochar was then cooled to room temperature overnight and stored in airtight containers. A laboratory‐scale pyrolysis system equipped with a tube reactor, a temperature controller (1200°C), a cooling system for the bio‐oil condensation unit, and nitrogen flow control was used. Finally, pelletization was carried out using a MILL‐10 Pellet Mill (10 HP; Colorado Mill Equipment, USA) with bentonite clay as the binding agent.
Experimental Layout
2.2
DBPs with an approximate diameter of 0.5 cm and 1–3‐cm length were used in batch and field adsorption–desorption experiments and categorized into two groups based on their evaluation stage: (1) new pellets that had not been exposed to any phosphorus source after manufacturing, indicated with N in Figure 1, and (2) spent pellets that had been exposed to phosphorus‐rich effluents for approximately 6 months at the outlet of a tile drain system, indicated with U in Figure 1.
Setup the sorption experiment for designer biochar pellets (DBPs), pure phosphate solution (BPS), in agricultural effluent (CME, EMW, and WBE), and field trial (DFE).
Experiments with new DBPs were designed to assess P sorption and desorption under two scenarios: (1) ideal conditions without competing ions, using a phosphate solution prepared from deionized water, and (2) agricultural effluents from a tile drain system. Experiments with spent DBPs focused on characterizing the slow release of P upon exposure to collected agricultural effluents, thereby mimicking DBPs exposed to soil solutions typically encountered under field conditions during rainfall or irrigation events.
Furthermore, new and spent DBPs were used to conduct field experiments in an agricultural field located in Champaign County, Illinois (Figure 2b).
Dairy farm experimental site at the University of Illinois, Urbana‐Champaign, Champaign County (CH): (a) regional location, (b) treated and control plots with new (N), spent (U), and not DBPs, and (c) cross‐section indicating location of vertical (V) and horizontal (H) monitoring wells for water samples and groundwater level.
Spent Pellets: New
2.2.1
DBPs were initially exposed to tile drain effluents in a farm located in Fulton County (i.e., spent), then further used to evaluate their properties when exposed to (a) monitoring wells (U‐EMW) effluents, (b) effluents after a woodchip bioreactor (U‐WBE), and (c) fresh cow manure effluent (U‐CME).
Field Experiment
2.2.2
A field experiment was conducted during the duration of corn production. Nine 4 × 3‐m plots were arranged before the spring 2023 corn planting season (Figure 2b) into three treatment groups: DFE (Figure 2) to indicate new and spent pellets, and control (C) for no pellet application, to assess the P sorption and release efficiency of the DBPs. An application rate of 5 tons ha^−1^ was used on replicated treatments. Hence, plots treated with new pellets were used to evaluate P sorption efficiency under field conditions. In contrast, those with spent pellets were assessed for the potential to release P. Control plots provided a baseline for comparing the effects of pellet application on soil.
DBPs and Effluents Characterization
2.3
Different analytical techniques were used on new and spent pellets or effluents for elementary characterization based on energy‐dispersive X‐ray spectroscopy (EDS), X‐ray diffraction (XRD), scanning electron microscopy (SEM), inductively coupled plasma (ICP), and Fourier transform infrared spectroscopy (FTIR).
Surface Morphology Analysis
2.3.1
Surface morphology, porosity, elemental distribution, and microstructural features of the biochar were analyzed using EDS and SEM (Thermo Scientific Axia ChemiSEM). Analysis of the biochar pellet samples was conducted in high‐vacuum mode, using an accelerating voltage of 15 kV, working distance of 9.5, 9.6, and 10.2 mm, and spot size of 3.0 and 5.0 to achieve high‐resolution imaging of the biochar surface.
ICP and Dissolved Organic Carbon (DOC) Analyses
2.3.2
The PerkinElmer NexION 2000 ICP‐MS was used to quantify total element concentrations from new and spent DBPs after acid digestion using nitric acid (HNO_3_) following the U.S. Environmental Protection Agency (U.S. EPA 1986) procedure. The Agilent 5100 ICP instrument was used to determine metal elements in agricultural effluents, without digestion, following the EPA method 200.7 SM3120B, while DOC concentrations were quantified following the SM5310 B 21st Edition 2000 method.
FTIR Analysis
2.3.3
The Thermo Nicolet iS50 FTIR Spectrometer was used to measure functional group composition on the surface of the new and spent DBPs after the initial exposure to tile drain effluent.
XRD Analysis
2.3.4
The Bruker D8 Advance Powder X‐ray Diffractometer was used to analyze the crystal structures of the new and spent DBPs before and after the initial exposure to tile drain effluent.
Orthophosphate Solution and Effluent Collection
2.4
Three different agricultural effluents were collected: (i) 2 L of effluent from each horizontal and vertical monitoring well at the dairy farm research site 1 week before planting, (ii) 2 L of effluent collected from the outlet of a woodchip bioreactor at the Nutrien Ag Solution farm during the growing season, and (iii) 60 L of fresh cow manure effluent (CME) from the dairy farm research site during early winter (December 2023) (Figure 2).
Dairy Farm Monitoring Wells Effluents (EMW)
2.4.1
Water samples were collected from vertical (V) and horizontal (H) monitoring wells on a 0.41‐ha experimental field at the University of Illinois, planted with corn and soybeans on a yearly rotation (Figure 2c). Water aliquots were sampled monthly for orthophosphate analysis, and larger samples were collected for sorption experiments using a peristaltic pump. Vertical monitoring wells were positioned 1.5 m from the cowshed fence, extended 1.5 m into the ground, and instrumented with 10‐m TD‐Diver pressure sensors (van Essen Instrument) for continuous groundwater monitoring (i.e., temperature and water level). Horizontal monitoring wells were installed at a shallower depth of 0.91 m, at the same depth, and between tile drains in the field, starting approximately 1.5 m from the cowshed fence and extending 40 m to the midway point. These wells were designed to capture subsurface water samples that represent the theoretical water source for tile drainage, providing a broader perspective on nutrient movement within the field system.
Woodchip Bioreactor Effluent (WBE)
2.4.2
Effluents were collected at the outlet of a bioreactor that treats the outflow of a 30‐ha drainage system on the Nutrien Ag Solution experimental Farm located in Savoy, Illinois, for corn and soybean row crop production.
Cow Manure Effluent (CME)
2.4.3
Sixty liters of fresh CME was collected from the biogas fermentation at the University of Illinois beef farm. The effluent was divided into four 13‐L containers, with one serving as the control (no pellets). Seven kilograms of spent DBP was added to the three treatment containers. The experiment was conducted in a controlled room at a constant temperature for 43 days. Before aliquot collection, each container was thoroughly mixed with a long wooden stick, and 30‐mL aliquots were collected at predetermined time points (1, 2, 3, 5, 8, 15, 22, 29, 36, and 43 days).
Phosphate Solution (N‐BPS)
2.4.4
A 10‐g L^−1^ stock solution of potassium dihydrogen phosphate (99.0% KH2PO4; Sigma‐Aldrich, St. Louis) was prepared on deionized water (> 18.0 MΩ·cm; Labconco Water Pro Plus, Kansas City, MO) without further purification for batch sorption experiments at different concentrations, following the EPA/600/B‐07/001 method.
Sorption Experiments
2.5
Adsorption experiments were conducted in four setups: duplicate analyses for orthophosphate (non‐ion competing solution), no replications for woodchip bioreactor and monitoring well effluents (ion‐competing solution), triplicate analyses for cow manure, and field experiment. Orthophosphate concentrations were measured using the Astoria‐Pacific Analyzer (Model 411 XYZ Sampler) in accordance with USEPA Method 365.3.
- Batch Adsorption in Non‐Ion‐Competing Solution: A KH_2_PO_4_ stock solution was used to investigate DBPs' sorption capacity at ambient temperature, diluted to target phosphorus concentrations of 0.2, 0.5, 1.0, 10.0, 25.0, 50.0, 75.0, and 100.0 mg L^−1^. Sorption experiments for N‐BPS were conducted to assess the theoretical equilibrium condition, with absorbent mass and solution volume adjusted according to predefined equations (see Supporting Information for more details). The maximum sorption capacity of DBP was evaluated using both the Freundlich and Langmuir isotherm models via nonlinear regression analysis in OriginPro V. 2005 (OriginLab Corporation, MA, USA).
- Adsorption in Ion‐Competing Solution: This setup examined new and spent DBP sorption at ambient temperature in contact with agricultural effluents, including tile drainage effluents and monitoring well effluents from agricultural fields. Samples were exposed to different effluents under shaker conditions, and aliquots were analyzed for orthophosphate at specific intervals. The Supporting Information provides full experimental details, including time intervals and sampling procedures.
- Adsorption in CME: This experiment evaluated the sorption of spent DBPs in organic‐rich conditions using fresh CME from an agricultural farm. The Supporting Information provides detailed procedures for sampling and analysis.
- Field Sorption: These experiments evaluate DBS sorption under natural and farming conditions. Six field plots were treated with new and spent DBPs mixed into the 15‐cm topsoil layer during the 2023 corn planting season. Soil phosphorus concentrations were measured before planting, after planting, and after harvest. DBPs were collected after harvest and analyzed for phosphate release efficiency across varying pH conditions. Please see the Supporting Information for detailed experimental procedures.
Triplicate desorption experiments were conducted at ambient temperature, with observed pH used both as an indicator and as a reaction driver, yielding orthophosphate (P) desorption efficiency for the DBPs. After experiments, the collected aliquots were stored at −10°C until P analysis. For P analysis, aliquots were thawed for 2 h and centrifuged at 4000 rpm for 20 min. Orthophosphate concentrations were measured using the Astoria‐Pacific Analyzer (Model 411 XYZ Sampler) according to USEPA Method 365.3.
Desorption Experiments
2.6
Desorption experiments were conducted in one setup each with no replications for DBPs previously spiked in pure phosphate solution, wastewater effluent, field trial, and soil sample from the field into the deionized water. Orthophosphate concentrations were measured using the Astoria‐Pacific Analyzer (Model 411 XYZ Sampler) according to USEPA Method 365.3.
- Batch phosphate desorption (BPS): Two desorption sets involving pellets from different phosphate concentrations during sorption were performed: (1) Pellets from N‐BPS on 150 mL of DIW were shaken and aliquot samples collected at 1‐, 3‐, 7‐, and 10‐h intervals. (2) Pellets from U‐WBE sorption on 150 mL of DIW were shaken and aliquot samples collected at 0.5‐, 1‐, 2‐, 4‐, 6‐, 12‐, 24‐, 48‐, and 72‐h intervals, and pH measurements were not taken.
- Field soil sample desorption: 20 g of soil samples collected before planting (SSBP), after planting (SSAP), and after harvest (SSAH) were added to separate flasks containing 150 mL of deionized water, shaken, and aliquot samples were collected daily for 4 days. Aliquots were sampled for P analysis following the same method as other DBP desorption experiments.
- DBPs desorption: 25 g of new and spent DBPs (treated with tile drain effluents and pellets retrieved from field plots) were placed in beakers with 150 mL of deionized water. The solution pH was adjusted daily from 7.0 to 9.5 by adding H_2_SO_4_ or HCl. After pH adjustment, 13 mL of solution was collected and stored at 4°C for analysis.
Results and Discussion
3
DBPs Characterization
3.1
Surface Morphology Analysis
3.1.1
Elementally, the spent DBPs show a noticeable difference in surface composition, enrichment of elements such as P, C, Na, Mg, and Fe, and a decrease in N, Si, Ca, and K. SEM results confirmed low phosphate retention (0.2%) in exposure to tile drain agricultural effluents, whereas during cow manure exposure, Mg, Al, P, and Fe concentrations decreased, and most other elements increased in surface concentration (Figure 3). These contrasting behaviors were explained by the complex composition of agricultural effluents, including DOC (see Section 3.1.2) and microbial activity, particularly in experiments associated with cow manure.
SEM micrograph of biochar samples (a) before phosphate adsorption, (b) after adsorption in tile drain effluent (spent pellets), and (c) spent pellets in cow manure effluent. The negative differences in the percentage weight are represented as a loss.
For instance, when DBPs were exposed to agricultural effluents, it was observed that, in general, the solution pH increased over time, ranging from near neutral to 8 to 9, and, in some cases, reaching above 12. These elevated alkaline conditions alter the presence of the P anion speciation: (1) decreasing hydrogen phosphate, H2PO4− (from pH 7 to around 10), (2) increasing and decreasing hydrogen phosphate, H O42− (from pH 7 to around 10, and pH around 10 to 14, respectively), and (3) increase phosphate PO43− (from pH around 10 to 14). This P speciation, as a function of changing pH, promotes different reactions across salts (e.g., Ca, Mg, K, and Na), other metals (e.g., Al and Fe), and DOC concentration. Further, the presence of metal oxides in DBPs (e.g., MgO and CaO), and anionic phosphates (H2PO4− and H2PO42−) likely from the dissolution of salts from agricultural practices (e.g., fertilizer application) supports the observed increase in pH. Hence, it was hypothesized that under increased alkalinity conditions, one path of orthophosphate removal was driven by the formation of suspended particles or precipitates of ionic metals with phosphate (e.g., Ca3PO42 and Mg3PO42). These reactions are likely to occur in the presence of elevated phosphate concentrations (e.g., 2 mg L^−1^ and greater) under rich Ca and Mg concentrations commonly present in agricultural effluents and the DBPs, thereby reducing P availability in solution. In contrast, DOC likely buffers P availability in solution by binding to Ca and Mg, thereby increasing orthophosphate concentrations. Further, under dissolved oxygen (note that this was not measured) and alkaline conditions, Al‐ and Fe‐amorphous oxides (e.g., FeOH3 and AlOH3) promote phosphate binding. However, in the presence of organic matter, organic acids can chelate Fe and Al, reducing their ability to bind P.
SEM on new to spent DBPs indicates that P adsorption may be constrained by three distinct mechanisms: (i) Site blockage occurs as inorganic and carbonaceous deposits accumulate during exposure to tile effluent for the duration of 6 months, covering active sites that are readily accessible in new DBPs and reducing the effective reactive surface area. (ii) Pore constriction and transport limitations further degrade performance, as initially open pores in new DBPs become narrowed, increasing internal diffusion resistance and shifting control from surface to mass transport. (iii) Finally, chemical transformation of the DBP's surface is evident by altered elemental compositions, indicating the formation of less reactive surface phases, suppressing adsorption.
ICP and DOC Analysis
3.1.2
Metal concentrations from ICP analysis of DBPs after digestion and exposure to agricultural effluents are summarized in Table 2, as well as the DOC in the effluents. Analysis of new DBP was used as a control, and spent DBP resulted in notable changes in elemental composition, particularly increased concentrations of trace metals such as Cu and Zn, as well as Al, Mg, and Fe in U‐based treatments. Ca and K concentrations generally declined, indicating leaching or Fe exchange during the exposure. The N_WBE treatment remained most like the control, indicating minimal alteration due to the duration of pellets in the effluent. At the same time, U_WBE and U_CME exhibited the greatest elemental enrichment. For agricultural effluents, using tile drain effluent as the control, differences in elemental concentrations highlight the influence of subsurface flow paths, manure inputs, and treatment processes (Table 2B). Relative to tile drainage, vertical and horizontal well effluents exhibited substantially higher concentrations of major cations (Ca, Mg, and K) and nutrients (P), indicating enhanced water–soil–sediment interactions beyond the shallow tile system. The horizontal well showed the greatest enrichment, particularly for K, Fe, and P, suggesting longer residence times and more reduced conditions that promote mineral dissolution and Fe‐associated P mobilization. In contrast, exposure to the vertical well effluent displayed elevated Ca and Mg but comparatively lower Fe, reflecting differences in redox conditions and depth‐dependent flow paths.
Dairy manure effluents differed most strongly from tile drainage, with higher K concentrations and consistently elevated P. Compared with tile drain values, manure effluent showed increasing K concentration over time and declining Ca and Mg concentrations, indicating nutrient release from organic matter decomposition during manure aging. These differences highlight manure as a dominant and dynamic nutrient source compared with background tile drain effluent, with a high risk of nutrient loading if manure enters tile systems.
WBE, on the other hand, showed marked reductions in elemental concentrations relative to the influent and, for several elements, fell below tile drain levels after 6 h. P, K, Ca, Mg, and Fe were substantially lower than under tile drainage, indicating effective nutrient reduction. This contrast highlights the bioreactor's capacity to counteract elevated nutrient inputs from manure or deeper subsurface flow before it is discharged. A complete elemental analysis of the wastewater effluents used by ICP is shown in the Supporting Information (Table 1). Fe is widely recognized as a key driver of phosphate removal under acidic conditions but is expected to play a lesser role than carbonates under the alkaline conditions observed in the effluents and study samples. Wang, Zhi, et al. (2021) describe Fe‐mediated phosphate immobilization on iron‐based materials, noting that, in addition to acidity, high dissolved oxygen levels promote Fe oxidation and enhance phosphate binding. In alkaline systems, DOC can further modulate phosphorus behavior by complexing Ca and Mg ions, thereby limiting carbonate–phosphate reactions and potentially increasing dissolved phosphate concentrations. In Table 2c, the DOC concentration exhibited substantial variability across the wastewater effluents used in the study, spanning four orders of magnitude from 12 to 256,200 mg L^−1^. Tile drain effluent exhibits a lower DOC concentration, suggesting limited organic input, which may be favorable for phosphorus removal. In contrast, effluent from horizontal wells and dairy manure (0 day) exhibited extremely high DOC concentrations, indicating elevated organic matter and pointing to strong inputs from waste‐derived sources, such as sludge or leachate, which may hinder phosphorus sorption due to competition. Dairy manure effluent (40 days) and effluent from woodchip bioreactor (0–6 h) exhibited intermediate DOC concentrations, reflecting organic loading that is associated with wastewater or agricultural effluent sources.
In general, pronounced variability in DOC concentration highlights organic matter inputs among the effluents. It plays a critical role in controlling phosphorus removal efficiency across the tested effluents.
Mg and Mo levels remained largely stable, while elevated Cu and Zn levels in some treatments may pose risks to plants, depending on soil conditions. In general, these results highlight the influence of effluent exposure on DBPs and underscore the need to evaluate potential impacts on long‐term interactions with soils, microbes, and plant growth before land application. To enhance crop resilience, a study by Millaleo et al. (2010) discussed the importance of understanding plant uptake of Mn levels and tolerance mechanisms in acidic soils, where the risk increases. Although Zn and Mn can pose harmful risks under certain soil conditions, such considerations are beyond the scope of this study.
Results in Table 2B illustrate clear contrasts between subsurface flow paths, manure inputs, and bioreactor treatment. While wells and manure represent substantial nutrient reservoirs, woodchip bioreactors show strong potential as a mitigation strategy to reduce nutrient loading and protect downstream water quality.
In this study, pH plays a dual role: (1) as an indicator of underlying mechanistic reactions and (2) as a driver of secondary reactions that regulate phosphorus (P) solubility, acting as a feedback mechanism. Given that effluent concentrations of Ca, Mg, and K were substantially higher than those of other elements (Table 2B), elevated pH conditions were expected to favor Ca‐ and Mg‐driven P fixation. As alkalinity increases (DBPs), Ca and Mg readily form suspensions or precipitates of Ca and Mg‐phosphate minerals, thereby reducing orthophosphate solubility. As described by Brunno da Silva Cerozi and Fitzsimmons (2016), alkaline systems typically progress through the formation of dibasic calcium phosphate dihydrate, octacalcium phosphate, and ultimately hydroxyapatite.
Overall, the combined presence of Ca, Mg, and K with phosphate and dissolved carbon raises pH by displacing exchangeable H^+^ and forming stable mineral and aqueous complexes that buffer acidity and limit phosphorus mobility.
FTIR Analysis
3.1.3
FTIR spectra of the new and spent DBP samples in Figure 4 reveal the presence of oxygen‐containing functional groups, aromatic carbon structures, and mineral components characteristic of thermochemically converted biomass. A weak but distinct absorption band at ~3754 cm^−1^ in both samples is attributed to O–H stretching vibrations of free hydroxyl groups, indicating a limited presence of surface hydroxyl functionalities. The broad, weak bands near 2622–2623 cm^−1^ are assigned to hydrogen‐bonded O–H stretching of carboxylic acid groups, suggesting residual acidic surface functionalities. Weak absorption bands in the range of 2273–2274 and 2036–2043 cm^−1^ are associated with C≡N or C≡C stretching vibrations, which may likely originate from nitrile or alkyne structures formed during biomass devolatilization. These features are minor in biochar and indicate advanced carbonization. Strong absorption bands at 1582 cm^−1^ for the new DBP sample and 1555 cm^−1^ for the spent DBP sample correspond to aromatic C=C skeletal vibrations, confirming the dominance of condensed aromatic structures. The slight shift to lower wavenumbers in the spent DBP sample suggests increased aromatic structural ordering. Bands in the region of 1191–1251 cm^−1^ are attributed to C–O stretching vibrations of phenolic, ether, or ester groups, with lower intensity indicating partial decomposition of oxygenated functional groups during pyrolysis. Out‐of‐plane aromatic C–H bending vibrations are evident at 883, 757, and 745 cm^−1^, further supporting the presence of substituted aromatic rings. A weak band near 511 cm^−1^ observed in the new DBP sample is attributed to Si–O or metal–oxygen bending vibrations, indicating inorganic mineral components derived from biomass ash.
FTIR spectra of new and spent DBP.
Though both DBP samples exhibit similar functional group profiles, noticeable differences are observed in peak positions and intensities. The spent DBP shows a shift in the aromatic C=C stretching band (from 1582 to 1555 cm^−1^) and a reduction in the intensity of the C=O stretching bands. These changes indicate a greater degree of aromatic condensation and a lower abundance of oxygen‐containing functional groups than in the new DBP samples. Additionally, the absence of lower‐wavenumber mineral bands in the spent DBP sample suggests either reduced ash content or masking by increased carbon ordering.
XRD Analysis
3.1.4
The XRD patterns of the new and spent DBP samples in Figure 5 reveal their largely amorphous carbon structure with the presence of crystalline mineral phases. The new pellets showed a broad diffraction feature at ~11.8°, a characteristic of amorphous carbon and disordered organic structures commonly found in biochar. Distinct peaks at ~22.4° and 29.4° are attributed to silica (SiO_2_) and carbonaceous aromatic layers, indicating the presence of mineral ash and partially ordered carbon domains. The peaks at 36.5° may correspond to metal oxides or carbonate phases, such as CaCO_3_ or MgO, mostly from biomass minerals. A diffraction peak at ~47.1° is associated with crystalline inorganic phases, further confirming the mineral content of the new pellets. Broader peaks at ~54.5° and ~62.9° are assigned to silicate and oxide minerals, respectively. In comparison, the weak peak at ~74.4° corresponds to high‐angle crystalline reflections from mineral impurities.
XRD analysis of new and spent DBPs, indicating major compounds.
After exposure to tile effluent, the XRD pattern of the spent pellets shows noticeable changes in peak intensities. The broad peak at ~10.8° confirms that the amorphous carbon framework remains intact after application. Newly visible peaks at 25.9° and 29.7° suggest mineral crystallization, likely due to sorption or precipitation of inorganic species during exposure. A broader reflection around ~47.5° suggests partial reorganization of crystalline phases or surface‐bound compounds. Additional peaks at ~59.4°, 68.5°, and 79.9° indicate the presence of new or transformed mineral phases, supporting the occurrence of surface interactions and possible ion deposition during exposure to tile effluent.
New and spent DBP samples exhibit a predominantly amorphous nature, typical of biochar materials. However, the spent DBP showed sharper, more numerous diffraction peaks, suggesting mineral accumulation, crystallization, or adsorption of inorganic species during application. The preservation of broad amorphous peaks confirms that the carbon matrix remains structurally stable. At the same time, changes in crystalline reflections highlight the biochar's active role in surface interactions.
DBP Phosphate Sorption
3.2
Batch Phosphate Sorption by DBPs
3.2.1
The effect of competing ions on phosphate sorption is illustrated in Figure 6. The Langmuir model provided the best overall fit to the experimental data (R ^2^ = 0.88 and 0.81) for estimating P sorption by N‐DBPs. In contrast, the Freundlich model yielded an R ^2^ of 0.40. Hence, the Langmuir model was used, which assumes monolayer sorption on a homogeneous surface with a finite number of sorption sites. These experiments evaluated theoretical DBP sorption under ideal conditions, in which reactive and competing ions are absent. Phosphate concentrations in tile drains and agricultural runoff typically range from 0 to 0.5 mg L^−1^, and in rare cases, exceed 5 mg L^−1^. However, elevated phosphate concentrations (indicated as Ce in Figure 6) were used to estimate the maximum sorption capacity of DBP. The Langmuir isotherm model revealed distinct sorption capacities of 3058 and 158 mg kg^−1^ for the two experimental conditions (Figure 6a,b). This contrast suggests that the presence of competing ions influences the outcome, and isotherms derived from plain phosphate solutions should not be used to assess agricultural effluents with complex matrices. The observed differences indicate that distinct phosphate solution removal mechanisms predominate as the mass of Ce and DBPs varies.
Batch sorption using a stock phosphate solution at various concentrations (Ce). (a) Pellets from phosphate solution without competing ions and (b) pellets from phosphate solution with competing ions.
The Langmuir model fits the experimental data at C _ e _ concentrations below 2.5 mg L^−1^, suggesting monolayer sorption on a homogeneous surface. This behavior was expected since the adsorption sites on the pellets were considered available for binding in this context. However, as C _ e _ concentration increased to approximately 2.5 mg L^−1^, the reduction in P concentration in solution is characteristic of P binding to Ca and Mg rather than being adsorbed by the DBP.
The increase in C _ e _ concentrations likely increases the concentration of available phosphate ions, which form phosphoric acid and potassium hydroxide, favoring reaction with carbonates (i.e., CaCO_3_ or MgCO_3_, available as lime sludge in the DBP) to form Ca‐ or Mg‐phosphate (Equations (1), (2), (3) to 4).
In this context, the removal of phosphate (H_2_PO_4_) from water was primarily driven by chemical reactions with Ca or Mg, forming insoluble Ca‐ or Mg‐phosphate rather than adsorption on the DBP's surface. Depending on pH, concentration of competing ions in the effluent (e.g., cations and nitrate), and temperature, the resulting phosphate species may vary and commonly form compounds such as CaHPO_4_, Ca(H_2_PO_4_)2, or Ca_5_(PO_4_)_3_OH, as discussed by Penn and Camberalo (2019), as well as changes in the alkalinity of the solution due to denitrification processes. Note that these dynamics were not investigated, and further investigation is necessary.
N‐BPS sorption, as shown in Figure 6a, resulted in a notable reduction in C _ e _ concentrations relative to Figure 4b. The desorption experiments in Section 3.3 support the proposed two stages of phosphate reduction: (1) monolayer adsorption at C _ e _ concentrations below approximately 2.5 mg L^−1^ and (2) a P concentration reduction driven by insoluble Ca‐ or Mg‐phosphate crystals (CaHPO_4_, Ca_3_(PO_4_)2, or Mg_3_(PO_4_)2). Yang et al. (2021) confirmed that precipitation is the primary mechanism for phosphorus removal by the DBPs. The results indicate that lime sludge‐amended pellets increased phosphorus removal efficiency via pH‐mediated sorption and precipitation processes in solution. Equally, Penn (2021) emphasizes the importance of understanding the practical implications in real‐world scenarios, particularly in pollutant management, where complex solutions in soil and runoff exist.
Woodchip Bioreactor Effluent (WBE)
3.2.2
Figure 7 summarizes the batch experiments from new and spent DBP exposed to a WBE (i.e., N‐WBE and U‐WBE), with initial effluent concentrations of 6.48 mg L^−1^. The results showed an initial increase in P removal, followed by a gradual decrease, then a final rise after 5 days. The maximum P removal from the solution was estimated at 140 mg kg^−1^ for the N‐WBE and 90 mg kg^−1^ for the U‐WBE following 1 day of exposure. In both cases, the lowest P removal occurred around the fifth day, at approximately 40 mg kg^−1^. A study by Sø et al. (2011) on rapid, reversible phosphate adsorption onto calcite materials highlights the influence of pH and Ca‐ or Mg‐carbonate activity. Their experimental data were modeled using a constant‐capacitance model with two types of sorption sites, underscoring the complexity of phosphate interactions with the calcite surface. It is important to note that the U‐WBEs were previously exposed to a tile drain effluent, air‐dried, and then re‐exposed to the effluent in the batch experiment, consuming calcium and magnesium. This is further supported by the pH measurements, which show substantial differences between the N‐WBE (mean pH = 12.65) and U‐WBE (mean pH = 8.43) experiments, most likely due to highly alkaline potassium carbonate (K_2_CO_3_) resulting from phosphate ion lime reactions (see Equations 1, 2, and 3), and phosphate ions hydrolyzation forming a strong base (see Equations (5), (6) to 7). Hence, the initial high reduction in phosphate in the WBE was likely due to carbonate formation as the pH increased. On the other hand, the decrease in phosphate (Figure 4) was explained by a buildup of phosphate hydrolysis products, which, in turn, led to the insoluble formation of carbonates in solution or precipitation after the fifth day (see Equation 8). Also, note that the surfaces of new DBPs were rich in alkaline functional groups. In contrast, the U‐WBE, as a spent sorbent, has depleted these groups due to prior sorption processes.
Comparison of the adsorption kinetics of new (N) and spent (U) pellets in contact with woodchip bioreactor effluent (WBE). Dashed lines indicate the 95% confidence limits of the observed data.
Effluent From the Monitoring Wells (EMW)
3.2.3
A summary of the DBP experiments with effluents collected at 0.91‐ and 1.5‐m depths from monitoring wells near a cattle barn is presented in Figure 8. The initial phosphate effluent concentrations for vertical (i.e., V) wells in Figure 8a,b and horizontal (i.e., H) wells in Figure 8c,d were estimated at 6.19 and 2.31 mg L^−1^, respectively. This initial P effluent concentration was attributed to contamination from manure runoff from the cow barn (see Figure 1c) and rapid macropore transport of surface runoff, while the differences were due to the mixing effect of subsurface water on the horizontal well (with dilution). The highest phosphate removal following DBP exposure was observed with the effluent collected from the horizontal well, despite its lower initial phosphate concentration (2.31 mg L^−1^). In contrast, when new and spent DBPs were in contact with effluents from the vertical wells, a decrease in phosphate concentration in solution was observed for the new DBP, attributed to phosphorus reduction via Ca‐ and Mg‐binding.
Kinetic adsorption of new and spent pellets using effluents collected at the Dairy Farm from the vertical monitoring well (V), upper figure (a and b), horizontal (H1) monitoring wells, lower figures (c and d). Dashed lines indicate 95% confidence limits of the observed data.
Following the first day of DBP sorption, phosphate removal reached equilibrium in both effluents, estimated at 50–55 and 10–20 mg kg^−1^ for the horizontal and vertical effluents, respectively. The differences in P removal from the two effluents indicated that an essential component of accumulated P in the DBPs, which was present in the new pellets and modulated by Ca‐ or Mg‐based oxides and carbonates binding, as well as surface sorption. Additionally, the initial pH of the effluent accounted for the differences in the final pH, which was approximately 12.6 and 9.5 for the horizontal and vertical samples, respectively.
Cow Manure Effluent (CME)
3.2.4
Figure 9 summarizes the 43‐day batch experiment with spent DBPs in contact with fresh CME having an initial P concentration of 1.57 mg L^−1^. During the first 15 days of contact, the phosphate concentration in the solution increased substantially, indicating phosphorus decomposition. Concurrently, the pH remained relatively low, starting at around 7.73 and gradually rising to 8.63 by Day 15. This pH range corresponds to phosphate desorption, as indicated by the negative sorption in Figure 8. Between Days 15 and 25, phosphate removal increased sharply, reaching approximately 13 mg kg^−1^ by Day 25. This upward trend continued, peaking at nearly 18 mg kg^−1^ on Day 31, indicating substantial phosphate uptake. During this period, pH steadily rose from approximately 8.63 to 8.92. The strong correlation between increasing pH and phosphate removal highlights the influence of alkaline conditions, in which pH likely affects phosphate hydrolysis and insoluble reactions between Ca and Mg ions and anionic P species. Following the peak at Day 31, phosphate removal declined sharply to approximately 8 mg kg^−1^ by Day 37. It decreased gradually, stabilizing at around 7 mg kg^−1^ by Day 43, indicating a release of excess phosphate into the solution as the pH increased to approximately 8.92.
Experimental kinetic adsorption of spent pellets in cow manure effluent with the measured solution pH. Dashed lines indicate 95% confidence limits. Dashed lines indicate 95% confidence limits of the observed data.
The results highlight a dynamic relationship between phosphate sorption processes and pH. During the initial phase (Days 0–13), lower pH values (~7.73–8.4) facilitated phosphate desorption. In contrast, during the phosphate removal phase (Days 19–31), higher pH levels (~8.63–8.93) likely promoted phosphate removal via Ca‐ and Mg‐binding. Dairy manure effluent exhibits a complex matrix rich in organic compounds that influences DBP processes and controls of P removal in solution. Nevertheless, the effluent's water quality parameters were not assessed, leaving gaps in understanding its composition and its impact on sorption dynamics. These findings, as shown in Figure 9, align with those of Shimabukuro et al. (2021), who demonstrated that the dynamics of CaCO_3_ and Ca_3_(PO_4_)2 are strongly influenced by pH, particularly at pH values above 8. Also, Fe and Al are important drivers at this pH in the presence of organic matter if oxygen is available.
Sorption experiments confirm that while phosphate sorption can be accurately assessed in controlled settings, agricultural effluents introduce complexities, including microbial activity, pH fluctuations, and competing ions. The effectiveness of DBPs depends on their physicochemical properties and wastewater composition; factors such as ionic competition and interactions with organic matter influence sorption efficiency. Understanding these dynamics is crucial for optimizing biochar design and ensuring long‐term effectiveness in wastewater treatment applications.
Desorption of DBPs in Deionized Water (DIW)
3.3
Batch Phosphate Desorption
3.3.1
This section summarizes two batch experiments designed to evaluate the effectiveness of DBP pellets in releasing phosphorus into solution: (1) exposure of new DBPs to phosphate solutions in the absence of competing ions and (2) exposure of both new and spent DBPs to phosphate‐rich effluent containing competing ions.
Effect of Effluent P Concentration (U‐BSP)
3.3.1.1
Figure 10 illustrates that phosphorus concentrations in deionized water (DIW) increase when DBPs are previously exposed to higher initial phosphate concentrations (e.g., 100 mg L^−1^). In contrast, pellets exposed to lower initial concentrations (e.g., 0.2 mg L^−1^) exhibited reduced phosphate desorption in DIW compared to those exposed to 1 mg L^−1^ or higher. This trend suggests that greater initial phosphate loading enhances the subsequent release of phosphorus into DIW. Additionally, phosphorus concentrations remained relatively stable over time, indicating minimal temporal variation. These findings highlight the role of initial solution conditions in influencing phosphate release behavior from the pellets (see Section 3.2.1).
Pseudo‐first‐order model for the desorption of spent pellet from batch phosphate sorption (U‐BPS) in solutions of varying initial concentrations (a–h). Dashed lines indicate 95% confidence limits of the observed data.
Pellets Reuse and Slow‐Release Effect (U‐WBE)
3.3.1.2
Figure 11 shows that phosphate desorption gradually increased, peaking at 5 h before declining to 0.05 mg L^−1^ after 72 h. A kinetic model (R ^2^ = 0.80) estimates a desorption capacity of 0.52 mg L^−1^. Before the sorption experiment with the new pellets, the WBE exhibited an initial phosphorus concentration of 6.48 mg L^−1^, providing a baseline for evaluating phosphorus removal efficiency by the new designer pellets.
Pseudo‐first‐order model for the desorption of spent pellet from woodchip bioreactor effluent (U‐WBE).
During precipitation, calcium from CaCO_3_ is thought to react with phosphate to create insoluble calcium phosphate molecules. The most frequent product at higher pH values is tricalcium phosphate (Ca_3_(PO_4_)2), while the dominating reaction is pH‐dependent (see Equation 8).
Effect of pH on Phosphate Desorption
3.3.1.3
Figure 12 illustrates the impact of pH on phosphate desorption from both new and spent DBPs. The new pellets (Figure 12a) served as a control, exhibiting no desorption across varying pH levels; however, the new pellet exposed during the field trial at the dairy farm experimental (N‐DFE) exhibited a change in the phosphate desorption compared to the new (N) and the new pellet from woodchip bioreactor effluent (N‐WBE) with an increase in the solution pH. In contrast, the P concentration in DIW increases with pellets exposed to P effluents (Figure 12b), with minimal temporal and pH variation, except for the spent (U) pellet, whose concentration remains low with increasing pH. The study highlights that pH levels substantially affect phosphorus desorption. While variations in effluent type influence overall phosphate desorption levels, the trend of increasing phosphate desorption with increasing pH remains clear. Additionally, ion competition is a critical factor influencing phosphate desorption. Silva and Souza (2005), in their study of Ultisols from Bahia, Brazil, reported that phosphate desorption increased with rising soil pH, as indicated by both the total phosphorus desorbed and the desorbed‐to‐initially‐sorbed phosphorus ratio. Similarly, Bai et al. (2017), working in the Yellow River Delta of China, found that pH and ion competition significantly affected phosphate desorption. They noted that competing anions, such as silicates, carbonates, sulfates, arsenate, and molybdate, can displace phosphate from sorption sites, enhancing desorption. Pellets that had been in contact with tile drainage effluent for 3 months (i.e., field‐aged) and were later exposed to a phosphorus‐rich effluent exhibited a more dynamic desorption response to increasing pH. Phosphate concentrations peaked at 9 mg L^−1^ in pellets previously deployed in the field. A summary analysis of the background soil from Field 8 at the study site was conducted in 2021 using the Mehlich 3 extraction method. These soil sample results provide baseline information on soil nutrient status for a field treated with dairy manure waste and lime yearly and are presented in Figure 13. The soil property described in Figure 13 is a well‐buffered Ca‐rich soil with substantial nutrient reserves. The chemical context is critical for interpreting DBP performances in the field. The pH of 7.9 and high Ca, K, and Mg saturation favor P association with Ca/Mg bearing phases, so introduction of DBP in the soil brought changes in P availability, which may reflect shifts in sorption–desorption balance rather than increases in overall P in solution, particularly given the high background soil P levels. High cation exchange capacity (CEC) and organic matter further support substantial P interaction capacity without a specific binding mechanism. Though iron (Fe) was abundant, the alkaline pH reduces the likelihood of Fe‐dominated P control.
Effect of pH on phosphate desorption in deionized water after exposure to different effluents: (a) new pellets and (b) spent pellets.
Soil analysis for Field 8, study site at the dairy farm experimental research.
In summary, the soil chemistry in Figure 13 is consistent with the observed DBP trends, indicating a moderate response in P availability within a Ca‐rich, P‐saturated system. Still, it does not provide mechanisms beyond the scope of the study's measurements.
Effect of pH on Phosphate Desorption From Soil Sample
3.3.1.4
Figure 14 presents phosphate desorption from soil samples collected at different intervals before planting (no pellets) and after planting (with new and spent pellets). For soil samples collected before planting (SSBP), phosphate desorption remained stable across pH values, indicating consistent behavior. However, in the soil sample collected after harvest, phosphate desorption increased, peaking at ~60 mg L^−1^ at pH 8.5 in the plot treated with new pellets (N). Though it declined over time, it remained higher than in other plots. In contrast, control (C) and spent pellet‐treated (U) soils showed a gradual increase in phosphate with rising pH, suggesting consistent desorption. Overall, pH strongly influences phosphate sorption and desorption, highlighting its role in evaluating DBPs for environmental and agricultural applications.
The effect of pH on the desorption of phosphate from field soil samples was collected at specific times during the planting season with pellets from SSAP and SSBP (a), and SSAH 4_N, SSAH, and 6_U, while SSBP and SSAH 2 had no pellets (b).
The desorption experiments evaluated DBPs as slow‐release P fertilizers, mimicking nutrient release during rainfall or irrigation. Prior wastewater exposure and pH fluctuations influenced the stability and extent of phosphate desorption. These findings highlight the DBPs' ability to support controlled nutrient recycling. This research emphasizes the role of biochar in circular nutrient management.
However, the study did not examine the effects of variations in water quality parameters, such as ionic strength or competing ions, on sorption. This leaves a gap in understanding the impact of these factors on the efficacy of DBPs in real‐world applications. The phosphate solution used in the laboratory (Section 3.2.1) was assumed to be free of competing ions, thereby providing a controlled comparison. In contrast, the field‐collected effluents (Sections 3.2.2, 3.2.3, and 3.2.4) were rich in ions due to environmental interactions. Furthermore, pellets exposed to field conditions may encounter complex ecological factors, including field activities and exposure to microorganisms, which could affect their stability and performance.
Summary and Conclusions
4
This study examined the effectiveness of new and spent DBPs in removing phosphorus from wastewater and their behavior in soil‐like conditions. New DBPs exhibited phosphorus removal via sorption and precipitation, with capacities exceeding 140 mg kg^−1^, although this decreased with shorter contact times. Spent pellets retained lower capacities (55–100 mg kg^−1^). Performance was influenced by pH and contact time; DBP composition raised the solution pH to alkaline levels, reducing sorption efficiency but promoting phosphate removal from solution via precipitation.
Effluents from synthetic solutions (WBE, CME, and EMW) were assumed to contain more competing ions, exhibit higher microbial activity, and present more complex sorption dynamics, thereby reducing the performance of DBP phosphate removal in real‐world applications. The Langmuir isotherm model fits the laboratory conditions in the absence of competing ions, but it is not recommended for use with complex agricultural effluents. SEM and ICP analyses confirmed the presence of phosphorus adsorption within the pellets following exposure to effluents. Metal analyses and DOC concentration, along with FTIR and XRD analyses, revealed the presence of Ca, Mg, and K, as well as P, which increases pH by supplying hydrogen ions and forming stable compounds that buffer acidity. On the other hand, soil application studies suggest that DBPs may promote phosphate removal in solution driven by phosphorus precipitation. DBP lime sludge can alter the soil solution to alkaline conditions, increasing the risk of phosphorus leaching and the accumulation of oxides and carbonates, potentially resulting in legacy phosphorus buildup and drainage clogging. Given the dual role of DBPs in sorption and desorption, assessing soil phosphorus levels prior to use is essential. Additionally, pellets with high Cu and Zn should be limited or blended with other materials to reduce metal toxicity risks.
Overall, DBPs present both opportunities and risks in phosphorus management. Further research is needed to evaluate their long‐term environmental impact, degradation, and practicality as sustainable fertilizer alternatives. Investigating phosphorus behavior across various soil types will be vital in determining whether DBPs could inadvertently increase phosphorus mobility in agricultural systems.
Author Contributions
Agnes Millimouno: conceptualization, original draft, methodology, investigation, formal analysis, visualization, data curation, writing review and editing. Jorge A. Guzman: conceptualization, methodology, supervision, data curation, formal analysis, review and editing. Richard A. Cooke: visualization, methodology, data curation, supervision, resources, formal analysis, review and editing. Wei Zheng: conceptualization, visualization, methodology, data curation, funding, resource, review and editing. Maria L. Chu: supervision, methodology, review and editing.
Conflicts of Interest
The authors declare no conflicts of interest.
Supporting information
Table S1: Elemental composition on the agricultural effluents using EPA 200.7 SM3120B method. Table S2: Dissolved organic carbon concentration of wastewater effluent. Table S3: Metal analysis on the designer biochar pellets using the EPA: 3050 method. Table S4: ICP‐OES total P analysis on the designer biochar pellets.
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