Facile Fabrication of Calcium-Doped Carbon for Efficient Phosphorus Adsorption.

High phosphorus concentrations mainly result in environmental problems such as agricultural pollution and eutrophication, which have great negative influence on many natural water bodies. In this work, calcium lignosulfonate was employed to produce calcium-doped char at 400 and 800 °C. To compare the phosphorus adsorption behaviors of the two carbon materials, batch adsorption experiments were conducted in a phosphorus microenvironment. The factors including the initial solution pH, phosphorus concentration, and adsorbent amount were considered, and the main characteristics of calcium-doped chars before and after adsorption were assessed. The results revealed that the phosphorus removal processes fitted both the Freundlich and pseudo-second-order-kinetic models. According to the Langmuir model, the maximum adsorption capacities of the two adsorbents obtained at 400 and 800 °C toward phosphorus (50 °C) were 53.22 and 17.77 mg/g adsorbent, respectively. The former was rich in calcium carbonate (CaCO3) and hydroxyl and carboxyl groups, and it mainly served as a precipitant and a chelating agent, while the latter with a high surface area was dominant in P adsorption.

Introduction

Although phosphorus (P) is an important nutrient for plant growth regulation, its eutrophication has caused a major pollution problem, with unprecedented amounts of P originating from agricultural sources.[1] Considerable amounts of P translocated to the surface and ground water bodies come from agricultural sources that include urine and agricultural organic fertilizers.[1,2] The Chinese Government has promulgated and implemented the strict first-class (A) wastewater discharge standard for P (0.5 mg/g), but a small amount of P (>0.02 mg/L) in lakes and rivers is likely responsible for severe eutrophication.[3] No living organism or ecosystem can survive without clean water, which is the most indispensable resource on Earth.[4] On the other hand, P is an essential nutrient, but its accumulation can result in eutrophication in many aquatic environments.[4] However, humans must rely only on 0.6% of global freshwater resources such as groundwater, rivers, and lakes.[5] Moreover, it has been proven that there is a positive correlation between the release rate of P into aqueous ecosystems and the continuous increase in human activities.[4] The severe environmental concerns related to P pollution have become an essential target of economic, social, and sustainable developments.[6]

Therefore, it is necessary to develop a sustainable technology to recycle P and avoid environmental pollution. Many techniques have been developed to control and treat P-rich water bodies, including the physical, chemical, and biological methods.[7−9] Chemical methods mainly include chemical precipitation, ion-exchange, and electrolysis processes, while membrane separation and adsorption are generally considered to be physical processes.[10] In all methods, P is removed by converting the phosphorus ions in aqueous solutions into a solid fraction.[7,8] This fraction can be an insoluble salt precipitate or a microbial mass in activated sludge.7,8 Physical P removal methods are expensive and have a low adsorption efficiency. The biological process may have certain advantages over chemical precipitation because they do not require the addition of chemicals; an improved biological process yielded a high total P removal level (97%).11 However, the technique is highly dependent on the external conditions, revealing poor practicability. Moreover, biological P removal methods are usually limited by the operational difficulties in removing P at a low dosage in water bodies.[12] Although chemical P removal techniques are facile, with high removal efficiency, they require a large amount of chemical reagents and easily cause secondary pollution such as chloride and sulfate ions.[11] The addition of chemicals, such as magnesium (Mg), aluminum (Al), calcium (Ca), and iron (Fe) salts, to wastewater is regarded a simple P removal procedure that separates P from aqueous solutions via chemical precipitation.[8,13] Biological and physical processes cannot effectively recycle P, which limits their potential for industrial applications. Moreover, chemical and biological removal techniques suffer high costs and environmental risks related to P-rich sludge disposal.[11]

Among these techniques, adsorption is considered to be a promising process because of its simple operation, cost-effectiveness, high efficiency, and less potential for secondary pollution.[4,14,15] Furthermore, the chemical adsorption process is a highly selective recycling-based strategy for P removal, which can compensate for the disadvantage of the above techniques to a certain extent. It has been proven to be an efficient and feasible control technology for agricultural pollution. Some carbon-based materials, such as activated carbon (AC), biochar (BC), graphite, graphene, magnetite carbon, and modified versions thereof, can act as excellent adsorbents that are capable of adsorbing P.[15,16] Unmodified carbons often have lower P removal than modified carbons. This is likely due to the negatively charged surfaces and limited functional groups of unmodified carbon materials.[4] Micháleková-Richveisová observed that the maximum P adsorption capacity of three unmodified BCs was very low,[17] with values of 0.036 mg/g for BC corn cobs, 0.132 mg/g for BC from garden waste, and 0.296 mg/g for BC from wood chips. To enhance the P adsorption performance of BC, Ca, Mg, Fe, and lanthanum (La) modifications are usually employed.[18−21] Physical and chemical activation processes are more frequently applied, likely because such treatment methods greatly benefit some functional improvements in the carbon surface area and porosity. Compared to physical activation/modification techniques, chemical modification can be less expensive and more time efficient.[22] Antunes also reported that although the P adsorption capacity of BC was similar to that of feedstock (biosolids) (17.1 mg/g),[15] employing BC for P removal from aqueous environments has many advantages over biosolids. Heavy metals from biosolids can leach via soil, causing serious harm to the food chain.[15] As concluded in previous reports, the heavy metals present in BC are not bioavailable.[23] Thus, employing BC for P adsorption and subsequently utilizing this loaded P material for land application should be a sustainable strategy for biowaste management and P reuse. Furthermore, BC land application has many advantages: enhanced water retention, improved soil structure, and soil fertility, which can increase crop yields.[6] Ca is also an essential element for plant growth. Despite a few reports on CaO- and MgO-codoped BC composites,[24] Mg-doped BC,[25] and Ca-decorated sludge BC for P removal,[26] the fabrication processes of adsorbents were complex and expensive. Therefore, it is absolutely essential to develop a facile process to produce Ca-doped BC with low costs. In addition, calcium lignosulfonate (CL) is one of the main byproducts of neutral sulfite and acid sulfite during pulping processes. CL contains a hydrophilic sulfonic acid group and forms a spherical three-dimensional network.[27] It can be feasible to recover CL from waste liquor, but CL removal is usually low.[28] Currently, there are few reports on BC production with low-molecular-weight lignin (e.g., CL).[28] Zhao et al.[28] investigated improved biohydrogen production with BC produced at 250 °C, but they did not further study high-temperature (over 400 °C) BC products from CL for other applications.

Therefore, there are some interesting studies indicating that CL-derived char (CLDC) can be produced to recover P from aqueous solutions and the combined P/Ca-C is a slow and controlled release fertilizer. The aim of this work is to (a) produce CLDC at 400 °C (CLDC400) and 800 °C (CLDC800); (b) evaluate their chemical compositions and physical characteristics and CaCO3-doped BC formation mechanisms; (c) compare their potential for P adsorption or removal as P-doped carbon to illustrate the process feasibility; (d) investigate the P adsorption kinetics and isotherms of CLDC400 and CLDC800; (e) clarify the P adsorption mechanisms of CLDC samples; and (f) demonstrate the potential of P/Ca-doped carbon as a slow-release fertilizer.

Materials and Methods

CLDC Production

CL was purchased from Tianjin Yeatschem Group, China. Purified water was produced in the Engineering Laboratory of Cleaner Energy for Light Industrial Wastes of Shandong (Qilu University of Technology). Some main characteristics of CL are as follows: molecular formula, C20H24CaO10S2; molecular weight, 528.61 g/mol; and CL content, 96%. Other chemicals (e.g., KH2PO4, NaOH, and HCl) were of analytical reagent grade and obtained from Beijing Sinopharm Chemical Reagent Group. Fifty grams of CL was carbonized in a vacuum tube furnace (OTF-1200X, Jingke, China) under nitrogen conditions (400 mL/min N2) at 400 or 800 °C for 2 h with a heating rate of 5 °C/min. After cooling to approximately 40 °C, the carbonized product was ground into powder, sieved to different sizes (80–100 mesh), and then washed with deionized water three times. Subsequently, the washed carbon was dried at 80 °C for 48 h in a vacuum drying oven, labeled CLDC400 or CLDC800, and stored in a sealable bag for future use.

Adsorption Process Design

A 50 mg/L P stock solution was prepared by dissolving KH2PO4 (0.2197 g) in 1000 mL of deionized water. Subsequently, various experimental parameters that may affect the P adsorption behaviors were applied in 150 mL conical flasks. Fifty milliliters of P solution (5.1–50 mg/L) was added to each conical flask, which was vibrated in a constant temperature shaker. When the set time was reached, the supernatant was filtered with a syringe-type filter (0.45 μm pore size) to determine the residual P concentration of aqueous solution. The influencing factors of adsorption performance of CLDC400 and CLDC800 included the adsorbent concentration (0.6–1.6 g/L CLDC400 or 1.0–6.0 g/L CLDC800), P solution (50 mL, 5.1 mg/L), contact time (0–360 min), contact temperature (30 and 50 °C), and initial solution pH (2.0–12). Each adsorption experiment was conducted in triplicate. The average value and standard deviation were calculated.

Test of P-Loaded CLDC as a P-Based Fertilizer

CLDC400 (0.12 g) and CLDC800 (0.3 g) were mixed with 100 mL of P (5.1 mg/L) and shaken for 24 h. After 24 h of shaking, the mixed samples were filtered and tested. The P-loaded CLDC samples after the adsorption process were collected, washed with deionized water, dried at 100 °C for 12 h before characterization, and utilized in the desorption experiment. In the P release test, 0.05 g of P-loaded CLDC samples was added to a centrifuge tube with 40 mL of deionized water with a pH of 5.0 or 7.0.[24] After 24 h of shaking (120 rpm), the mixed samples were filtered and the P concentrations in the supernatants were determined. The P release rates were measured by dividing the release dosage by the adsorbed content.

Model Description

The removal efficiency (R, %) and sorption capacities at time t (q, mg/g) and equilibrium (qe, mg/g) of P were calculated according to eqs –3, respectively[29]Here, m is the adsorbent mass (g); V represents the solution volume (L); C0 (mg/L) is the initial P concentration; C (mg/L) is the P concentration at time t; and Ce (mg/L) denotes P concentration at equilibrium.

Two isotherm models were employed to clarify the adsorption performance of P on CLDC materials, and the two models are presented as eqs and 5(30)where qe (mg/g) represents the sorption capacity at equilibrium; qm (mg/g) is the Langmuir maximum capacity; Ce (mg/L) denotes the P concentration at equilibrium; KL (L/mg) is the Langmuir constant; and KF (mg(1–1/ L1//g) and n describe the Freundlich sorption capacity and intensity, respectively.

To further understand the P sorption kinetic mechanisms, three common models were employed to evaluate the experimental data, which are shown in eqs –8[31]where qe (mg/g) and q (mg/g) are the P sorption capacities of adsorbent at equilibrium and time t, respectively; and k1 (min–1), k2 (g/mg min), and kp [g/(mg min0.5)] represent the kinetic constants.

Analytical Methods

The as-prepared CLDC400 and CLDC800 samples were characterized as follows: (1) scanning electron microscopy combined with energy-dispersive spectroscopy (SEM/EDS) (Regulus 8220, Japan) was employed to investigate the surface structure morphology and elemental compositions of CLDC materials. (2) The specific surface area (SSA) pore size distribution of the tested CLDC samples was analyzed using a Brunauer–Emmett–Teller (BET) analyzer (Mac 2020 hd88, USA) under the conditions of 77 K and relative pressures (P/P0) between 0.0 and 1.0 and calculated using the BET equation.[32] (3) To obtain X-ray diffraction (XRD) patterns of CLDC materials, samples were scanned from 10 to 80° at 5 °C/min under the conditions of 40 kV and 200 mA using a powder diffractometer (D8-Advance, Germany) with Ni-filtered Cu Kα radiation (AXS, Germany). (4) The CLDC samples were dried at 60 °C for 24 h. Fourier-transform infrared (FTIR) spectroscopy (IR Prestige-21, Shimadzu, Japan) was used to determine the chemical functional groups of CLDC materials; analysis was conducted from 4000 to 400 cm–1 with a resolution of 4 cm–1 by the KBr method. (5) The zeta potentials of CLDC400 and CLDC800 dispersed in Milli-Q water were determined using a Malvern Zetasizer (Nano ZS90, Malvern, UK). (6) CL thermal behavior was further evaluated by thermogravimetric analysis (TGA) with differential scanning calorimetry (DSC) (STA 449F3-QMS403C, Germany). CL was heated from room temperature (30 °C) to 1000 °C at a heating rate of 5 °C/min, mimicking the CL pyrolysis process. Then, the TGA/DSC curve of CL was obtained. In addition, the total P content before and after adsorption were measured using an ammonium molybdate spectrophotometric method. (7) The P-loaded chars were recovered by centrifugation, dried at room temperature (25–30 °C) for 24 h, and further examined by SEM/EDS. (8) X-ray photoelectron spectroscopy (XPS) of CLDC before and after adsorption was performed on an X-ray photoelectron spectrometer (Thermo Fisher, ESCALAB250Xi, USA).

Statistical Analysis

All results are presented as an average of two replicates. The adsorption effect was evaluated by one-way analysis of variance and differences were statistically significant at a level of P < 0.05. Isotherm and kinetic models were used to describe the P sorption data using Origin 8.5.

Results and Discussion

Physicochemical Characteristics of CLDC Samples

As described in the Supporting Information (Figure S1), the thermochemical behavior of CL was recorded and evaluated through a TGA/DSC system. It is vital to uncover the thermochemical characteristics of CL after carbonization, as these characteristics are particularly important when CLDC is used as a potential adsorbent to recover or remove P from aqueous solutions. Figure S1 shows that the effect of pyrolysis temperature on mass evolution revealed a two-stage process. The first stage of pyrolysis process was accompanied by water loss from the CL sample, which occurred at approximately 100 °C. However, the pyrolysis stages mainly depended on the species or physicochemical characteristics of lignin. The CL mass loss was negligible when the carbonization temperature was less than 200 °C, and an obvious mass loss was observed from 200 to 500 °C (Supporting Information, Figure S1). The molecular weights of CL were easily modified over 130 °C.[33] This further revealed that CL is thermally unstable above 250 °C. The second stage, which was observed between 250 and 580 °C, was exothermic. This stage played an important role in sulfur release and porous CLDC formation. After 400 °C, the carbonization of CL was mainly finished, and the weight loss was slow (Supporting Information, Figure S1). High temperature promoted sulfur gas release and increased carbon, while a small amount of S was retained and embedded within the CLDC samples (Supporting Information, Figure S2). Through SEM, more porous surface was found for CLDC800 than for CLDC400 (Supporting Information, Figure S2). A thorough understanding of the mechanism of the CL carbonization process will be conducive to improving the structural properties of CLDC, product selectivity, and application level.

In addition, higher temperatures caused increases in the SSA and pore volume of CLDC. The surface area of CLDC800 was 268.81 m2/g, which was significantly higher than that of CLDC400 (8.95 m2/g). The corresponding pore volumes were 0.118 and 0.008 cm3/g, respectively. During the CL pyrolysis process, volatile organic matter released at 800 °C was considered the main reason for the increased porosity and surface area. The decomposition and release of volatile organic compounds produced more micropores and mesopores, increasing the surface area of CLDC (Figure ). The pore size distribution of CLDC materials from the sorption branch reveals that CLDC800 had much smaller average pore diameter (1.94 nm) than that of CLDC400 (4.47 nm) (Figure bd). According to the International Union of Pure and Applied Chemistry (IUPAC) classification, the amount of nitrogen adsorbed onto CLDC400 was quite low, indicating that the adsorption capacity of CLDC400 depends on the accessible microporous volume rather than on the surface area (Figure a). CLDC800 (Figure c) exhibited type Ib adsorption, suggesting that multilayer adsorption gradually formed on the uniform nonporous surface of the graphitized carbon, and each layer provided a different adsorption amount. This was in line with the morphological characteristics obtained from SEM analysis. Moreover, to further demonstrate the crystalline structures of samples, XRD patterns and FTIR spectra were obtained for each CLDC and are described in Figure . A broad adsorption peak was observed for CLDC400 at approximately 3460 cm–1, which was closely related to −OH stretching vibrations. The −OH groups facilitate P adsorption because of the interaction between −OH groups and P.[34] The peaks at 1396 and 1620 cm–1 were attributed to CO32– and −C=O (−COOH), respectively, found in CLDC400. However, for CLDC800, the above peaks weakened or disappeared. With increasing carbonization temperature, the spacing of cellulose graphite microcrystalline layers decreased, and the superposition density and crystallinity improved, which could be an important reason for the strong stability of CLDC800. During the CL pyrolysis process, CaCO3 within CLDC800 partly decomposed into CaO at 800 °C, producing a peak at 53.8° associated with CaO.[24] The peak near 28° (3.36 nm) further confirmed the presence of CaCO3 in CLDC400, as also confirmed by Wang et al.[35] and Zhao et al.[28] During the CL carbonization process, the two groups of CO32– and Ca2+ were produced, and they could interact together and form CaCO3. Interestingly, CLDC400 contained the −OH and −C=O (−COOH) substituents necessary for P adsorption.[36] Although the XRD curves of CLDC400 and CLDC800 showed the existence of inorganic impurities (Figure ), the graphitization degree of CLDC800 was much higher than that of CLDC400 (Supporting Information, S2). In addition, Figure a shows the zeta potential measurements for CLDC400 and CLDC800 at different pH values. The zero point of charge (pHzpc) values of CLDC400 and CLDC800 were 2.1 and 2.9 (Figure a), respectively, which indicated that electrostatic attraction was negligible during the adsorption experiment conducted at pH 2.0–12.

Figure 1

Phosphorus adsorption–desorption isotherms of CLDC400 (a) and CLDC800. (c) Pore size distributions of CLDC400 (b) and CLDC800 (d).

Figure 2

FTIR spectra (a) and XRD patterns (b) of CLDC400 and CLDC800.

Figure 3

Zeta potential of CLDC samples at different pH values (a), effects of initial pH values on final pH values (b), effects of initial pH for P adsorption on CLDC samples (c), and P release rates in waters of different pH values (d).

In addition, samples of CLDC400 and CLDC800 before and after sorption tests at the initial solution pH 7.0 and 5.1 mg/L P were also characterized by XPS (Figures and 5). The XPS spectra of O 1s in CLDC400 before and after sorption (Figure a,b) reveal that O appears in two forms: −OH (83.48%) and C–O (16.52%); while the −OH content decreased to 64.6%, the C–O content increased to 29.59%. This indicated that P sorption could cause the changes in the chemical state of oxygen. The form of Ca before and after sorption was obviously changed (Figure c,d). Ca exists in the main form of CaCO3 before sorption while it is present in the main form of Ca3(PO4)2 in CLDC400 after sorption, which indicates that CaCO3 in CLDC400 could participate in the P removal process. The XPS spectra of C 1s in CLDC400 before and after sorption (Figure e,f) describe that C–H/C–C, C–O, and O–C=O peaks appear at 284.8, 286.3, and 289.2 eV, respectively, in both samples, revealing no obvious differences in the samples. It could be verified that the carbon skeleton of CLDC400 mainly played a physical role in P sorption. Figure g shows that S mainly exists in the forms of R–SH (51.81%) and SO42– (48.19%) before adsorption. Figure h shows that the P content in CLDC400 before adsorption is lower than that in XPS detection line, and a large amount of P in the form of PO43– is detected in CLDC400 after adsorption. For CLDC800, O is mainly present in the form of −OH (63.26%), C–O (30.27%), and metal oxides (6.46%); while the −OH content increased to 67.53%, C–O and metal oxide contents decreased to 28.62 and 3.85%, respectively (Figure a,b). Figure c,d shows that Ca exists in the main form of CaCO3 before sorption, whereas it is present in the main form of CaHPO4 in CLDC800 after sorption, thus suggesting that CaCO3 in CLDC800 could take part in the P sorption process. Figure g reveals that S mainly exists in the forms of R–SH (34.85%) and SO42– (65.15%) before adsorption. CLDC800 samples before and after sorption contained similar peaks at 284.8 eV (C–H/C–C), 286.4 eV (C–O), and 289.3 eV (O–C=O) in both samples, showing no obvious differences in the samples (Figure e,f). This further confirmed that the carbon skeleton of CLDC800 mainly played a physical role in the P removal process. Figure h indicates that the P content in CLDC800 before adsorption is lower than that in the XPS detection line, and a large amount of P in the form of PO43– is determined in CLDC800 after adsorption.

Figure 4

O 1s (a), Ca 2p (c), C 1s (e), and S 2p (g) spectra of CLDC400 before sorption; O 1s (b), Ca 2p (d), C 1s (f), and P 2p (h) spectra of CLDC400 after sorption.

Figure 5

O 1s (a), Ca 2p (c), C 1s (e), and S 2p (g) spectra of CLDC800 before sorption; O 1s (b), Ca 2p (d), C 1s (f), and P 2p (h) spectra of CLDC800 after sorption.

Effect of Initial pH on Phosphorus Removal

The system pH is the main factor affecting P sorption onto carbon materials. The pH affects not only the ionization morphology of phosphate in aqueous solutions but also the active components and surface charges on the surface of carbon-based adsorbent.[37] The effects of initial solution pH on the final solution pH and P adsorption onto CLDC400 and CLDC800 samples are described in Figure b,c. The results revealed that the P adsorption capacity of CLDC400 increased from 0.63 to 4.20 mg/g, while the capacity of CLDC800 increased from 0.14 to 1.65 mg/g (Figure c). A further increase in solution pH did not significantly change the P adsorption capacities of the two carbon materials (Figure c). Karaca et al.[38] also observed that the P adsorption capacity of calcinated dolomite powder was not significantly changed (45.7–47.8 mg/g) with the solution pH increasing from 1.0 to 11.0. Dai et al.[39] found that the P adsorption capacity increased from 58.23 to 60.64 mg for P/g Ln-doped BC when the initial pH was increased from 2.5 to 4.5. A similar report described that the sorption of P onto BC particulates depended on the initial solution pH.[40] P adsorption was the lowest at pH 2.0. When the pH value increased from 2.0 to 4.1, the adsorption of P by BC exhibited an increasing trend. Further increases in the pH value from 4.1 to 10.4, however, lowered the adsorption of P onto BC.[40] The highest P recovery capacity was obtained at pH 4 because the acid solution dissolved calcium compounds leaving Ca2+ free for complexation with P species.[31] Cai et al.[41] also found that P sorption onto BC significantly decreased at pH > 10 because of enhanced electrostatic repulsion between the elevated negative surface charge and multivalent P oxyanions (pKa 7.21, pKa 12.31). Yin concluded that Ca-doped BC had the best sorption capacity for K3PO4,[42] and the sorption amounts were on the order of K3PO4 > K2HPO4 > KH2PO4. The presence of H+ obstructed the complexation of P with Ca, lowering the sorption ability for K2HPO4 and KH2PO4.[42] Considering that most natural eutrophic waters have pH 6–9, BC produced at 450 °C had broad applicability for P recycling from a majority of eutrophic water.[41] Thus, solution pH 7.0 was employed to investigate the effects of other factors on the P sorption behaviors of two carbons materials in the following sections.

Effect of Adsorbent Dosage on Phosphorus Removal

It is necessary to investigate the between adsorbent dosage and P removal level. Theoretically, the removal efficiency of P is directly proportional to the adsorbent dosage. The greater the adsorbent amount is, the higher the removal efficiency of P is. However, if the adsorbent is overused, the cost increases, which is not beneficial to practical applications. Similarly, if the amount of adsorbent is insufficient, the P removal/sorption effect is insignificant. Therefore, it is very important to select the appropriate amount of adsorbent. As described in the Supporting Information (Figure S3), at adsorbent concentrations of 0.6–1.6 g of CLDC400/L or 1.0–6.0 g of CLDC800/L, the initial P concentration of 5.1 mg/L, the initial pH 7.0, the contact temperature of 30 °C, and a time of 24 h, the P removal rate increased with increasing CLDC concentration. The P removal with 0.8 g of CLDC400/L was 78.97%, whereas the P removal rate with 2.0 g of CLDC800/L was 71.79% (Supporting Information, Figure S3a). When the adsorbent concentration reached 1.2 g for CLDC400/L and 3.0 g for CLDC800/L, the P removal rates were 99.49 and 97.44%, respectively (Supporting Information, Figure S3a). With a further increase in the adsorbent concentration, the P removal rates had little increase. These phenomena likely occurred because the contact surfaces between P and the adsorbent were close to saturation, which caused a decrease in the utilization rate of sorption sites. Therefore, considering the P adsorption effect and the cost of adsorbent, CLDC400 and CLDC800 concentrations of 1.2 and 3.0 g/L, respectively, were selected for use in the following sections.

Effect of Contact Time on Phosphorus Removal

P adsorption processes become complicated with prolonged contact time and contain fast adsorption reactions, followed by slow phases.[43]Figure S3b shows that at an adsorbent concentration of 1.2 g for CLDC400/L or 3.0 g for CLDC800/L, the initial P concentration of 5.1 mg/L (50 mL), the initial pH of 7.0, and the contact temperature of 30 °C, the P adsorption capacities of two adsorbents increased gradually with increasing contact time until reaching the adsorption equilibrium. As revealed in Figure S3a, the P removal efficiency of CLDC400 was higher than that of CLDC800 (99.49%), which was attributed to the surface chemical characteristics of CLDC400. Violante and Pigna also observed that P adsorption increased with contact time.[44] The contact time should be long enough to enable the adsorption process to reach equilibrium.[43] When the initial phosphate concentration was 5.1 mg/L, it took 4 h for CLDC400 and 5 h for CLDC800 to reach sorption equilibrium (Supporting Information, Figure S3b). Therefore, the equilibrium time mainly depended on the sorbent type and dosage, initial P concentration, adsorption temperature, and agitation level.[45]

Effects of Adsorption Temperature and Initial P Concentration on P Sorption

The initial P concentration and contact temperature are two important factors that affect the adsorption capacity of CLDC. They affect not only the adsorption speed but also the adsorption capacity of the adsorbent. With increasing initial P concentration, the adsorption capacities of CLDC400 and CLDC800 obviously improved (Supporting Information, Figure S3c). The phenomena likely occurred because an increase in the P concentration in the system would increase the coordination probability between P and the adsorbent, resulting in an improvement in adsorption performance. Surprisingly, CLDC400 had great potential in the P adsorption process. The sorption capacity of CLDC400 (1.2 g/L) for P (5.1–50 mg/L) increased from 0.43 to 37.49 mg/g at 30 °C and pH 7 for 24 h. CLDC400 showed great potential in recovering P, with adsorption rates from 0.42 to 38.18 mg/g under the same conditions, except that 50 °C was used (Supporting Information, Figure S3c). Although the adsorption capacity increased with increasing solution temperature, the overall improvement was insignificant when the initial P concentration was between 5.1 and 50 mg/L. The highest sorption capacity (13.44 mg/g) of CLDC800 was obtained at 50 °C, which is slightly higher than that (11.72 mg/g) obtained at 30 °C under the conditions of pH 7, initial P concentration of 50 mg/L, and 3.0 g of CLDC800/L (Supporting Information, Figure S3c). The Langmuir parameter KL showed a positive correlation with solution temperature, which indicated that an increase in temperature improved the affinity between the CLDC adsorbent and P. This also indicated that the adsorption temperature slightly affected P removal by CLDC400 and CLDC800, which is also beneficial to the CLDC practical application. Interestingly, comparing CLDC materials with previously reported adsorbents for P adsorption indicated that CLDC is better than most other adsorbents (Table ).[3,9,15,26,46−49] However, there was some difficulty in comparing the P removal capacities with those reported in previous studies because of the differences in the initial P concentration, pH, adsorbent type, and dosage. Thus, CaCO3-doped CLDC400 produced by in situ and low-temperature pyrolysis is a promising adsorbent for P recovery.

Table 1

Comparison of the Sorption Abilities of Carbon-Based Adsorbents for Aqueous P Removal

BC-based materials	surface area (m2/g)	initial P con. (mg/L)	carbon dosage (g/L)	initial pH	Qmax (mg/g)	references
Ce/Fe3O4-BC	279.16	75	1.0	6.12	18.75	(3)
La/Fe3O4-BC	236.02	75	1.0	6.08	25	(3)
Biowaste BC	53.0	90	2.0		23.9	(9)
Biowaste BC	64.67	1500	10		16.4	(15)
CaO–MgO BC	169.33	450	2.0	7.0	201.23	(24)
Ca-doped BC		200	3.0		116.82	(26)
Mg-doped BC	180	125	1.0	7.0	24.08	(46)
MgO-magnetic BC	27.22	200	2.5	4.0	121.25	(47)
magnetic BC	92.54	200	2.5	4.0	2.47	(47)
magnetic BC	19.4	12	6.25		1.24	(48)
Fe-doped AC	442.23	50	3.0	3.0	2.692	(49)
CLDC400	8.95	50	1.2	7.0	38.18	this work
CLDC800	268.81	50	3.0	7.0	13.44	this work

Characteristics of Phosphorus Adsorption Isotherms

P adsorption isotherm experiments were performed with P concentrations from 0–50 mg/L (50 mL). The adsorbent dosage was 1.2 g of CLDC400/L or 3.0 g of CLDC800/L, while the contact time was 10 h and the solution pH was 7 for all experiments. To investigate the adsorption isotherms, Freundlich and Langmuir models were employed to fit the experimental data, and the described results are revealed in the Supporting Information (Figure S4) and Table . The model constants KF (mg(1–1/ L1//g) and KL (L/g) describe the sorption capacity, qm (mg/g) is associated with the monolayer adsorption capacity, and the Freundlich exponent (n) is the sorption intensity. The model parameters for Langmuir and Freundlich equations are shown in Table , and the values indicated that the two models fit the experimental data well (R2 > 0.98). The results also showed that the Langmuir model was very suitable for P adsorption on the surfaces of CLDC400 and CLDC800. Similarly, the Freundlich isothermal equation is an empirical model that describes multilayer adsorption on heterogeneous surfaces with sites of different energies. The isotherm results indicated that the amount of P adsorbed by BC was the sum of all adsorption sites on the surface of BC, and the strong binding sites were occupied first.[50] With an increase in the initial P concentration, the Langmuir maximum P adsorption capacity (qm) of CLDC400 at 30 °C increased from 0.50 to 50.05 mg/g (Supporting Information, Figure S4a), while the qm of CLDC800 at 30 °C increased from 0.42 to 15.49 mg/g (Supporting Information, Figure S4c). Similarly, the high adsorption revealed similar trends in the qm values for CLDC400 and CLDC800 (Supporting Information, Figure S4b,d). Interestingly, the n values of CLDC400 at 30 and 50 °C were approximately 1.84 and lower than those of CLDC800 (Table ). This finding further confirmed that the improvement in the adsorption intensity of CLDC400 attributed to the chemical reaction between Ca2+ released from CLDC400 and solution phosphate ions.[31]

Table 2

Adsorption Isotherm Parameters

		Langmuir model			Freundlich model
temperature	Adsorbent	KL (L/mg)	qm (mg/g)	R2	KF (mg(1–1/n) L1/n/g)	n	R2
30 °C	CLDC400	0.471	51.05	0.971	15.51	1.841	0.991
	CLDC800	0.190	15.49	0.964	3.116	2.100	0.987
50 °C	CLDC400	0.557	53.22	0.979	17.83	1.846	0.992
	CLDC800	0.303	17.77	0.966	4.573	2.059	0.982

Characteristics of Phosphorus Adsorption Kinetics

The adsorption kinetics is the most important factor in evaluating the sorption efficiency of adsorbents. To compare the P adsorption kinetics of CLDC400 and CLDC800, kinetic adsorption experiments were carried out under the conditions of 1.0 g/L adsorbent, 5.1 mg/L P (50 mL), P solution pH of 7, adsorption temperature of 30 °C, and rotational speed of 100 rpm. The three common models, the pseudo-first-order model (eq ), pseudo-second-order model (eq ), and intraparticle diffusion model (eq ), were employed to evaluate the experimental data, and the fitted parameters are described in the Supporting Information (Figure S5) and Table . The initial adsorption was rapid between 0 and 30 min, followed by slow adsorption (Figure S5a), which was also observed by Wu et al.[36] The initial rapid period adsorption process was likely due to the electrostatic attraction between the adsorbent and P ions.[36] The subsequent slow adsorption phase indicated intraparticle diffusion. Fitting the experimental data to the three kinetic models suggested that P adsorption processes for CLDC400 and CLDC800 were best fitted by the pseudo-second-order model, which had the highest R2 values (over 99%) (Figure S5b, Table ). The sorption reactions were similar to the chemical adsorption process revealed for other metal oxides (e.g., Fe3O4, MgO)-doped carbons.[36,51] In addition, in terms of eq , a pattern of qe versus t0.5 should be a straight line with the intercept C and slope kp when the adsorption step follows the intraparticle diffusion model. Moreover, Ho concluded that the plot of qe versus t0.5 must cross the origin if the intraparticle diffusion is the sole limiting step.[52]Figure S5c indicates that although intraparticle diffusion was involved in the adsorption process, it was not the sole rate-controlling step. To some extent, the boundary layer diffusion likely controlled the P adsorption behavior. The intraparticle diffusion was more obviously involved in P adsorption onto CLDC800 than that onto CLDC400 (Table and Figure S5c). In addition, the above adsorption characteristics of CLDC composites are closely related to their textural properties.

Table 3

Kinetic Adsorption Parameters Fitted by Three Common Models

	pseudo-first-order			pseudo-second-order			intra-particle-diffusion
adsorbent	k1 (min–1)	qe (mg/g)	R2	k2 (min–1)	qe (mg/g)	R2	kp [mg (g min0.5)−1]	Ci	R2
CLDC400	0.018	2.148	0.85	0.04	4.247	0.999	0.092	2.856	0.981
CLDC800	0.014	1.734	0.963	0.007	1.994	0.995	0.103	0.013	0.982

Adsorption Mechanisms

The statistical analysis results are described in Tables and 5, including the differences in average values of the indicators for any two groups under the same conditions, as well as the related p-values. All indicators of P removal have a p value less than 0.001, suggesting that the type and dosage of CLDC, solution pH, and initial P concentration significantly affected the P removal and adsorption capacities of the two adsorbents during the P adsorption process. However, there were some differences between this work and previous report by Ngatia et al.[1] who found that P adsorption increased with increasing BC pyrolysis temperature and that the optimum P adsorption was strongly associated with the feedstock property. They also observed that the high thermally stable carbon predominated by aromatic carbon and alkaline BC facilitated P sorption.[1] In this work, the possible mechanisms of P removal by the CLDC materials are described in Figure .[24]

Table 4

One-Way ANOVA on Different Indicators in P Adsorption with CLDC400

independent variable		DF	MS	F	p-value
Time	between groups	7	0.550	68.702	<0.001
	residual	16	0.008
	total	23
CLDC400 concentration	between groups	5	997.750	125.381	<0.001
	residual	12	7.958
	total	17
initial pH	between groups	10	3.354	248.866	<0.001
	residual	22	0.013
	total	32
phosphorus concentration (30 °C)	between groups	8	520.051	350.580	<0.001
	residual	18	1.483
	total	26
phosphorus concentration (50 °C)	between groups	8	542.935	344.519	<0.001
	residual	18	1.576
	total	26

Table 5

One-Way ANOVA on Different Indicators in P Adsorption with CLDC800

independent variable		DF	MS	F	p-value
time	between groups	7	0.835	110.332	<0.001
	residual	16	0.008
	total	23
CLDC800 concentration	between groups	5	3910.433	529.031	<0.001
	residual	12	7.392
	total	17
initial pH	between groups	10	0.597	86.963	<0.001
	residual	22	0.007
	total	32
phosphorus concentration (30 °C)	between groups	8	51.538	225.842	<0.001
	residual	18	0.228
	total	26
phosphorus concentration (50 °C)	between groups	8	68.347	316.379	<0.001
	residual	18	0.216
	total	26

Figure 6

Possible mechanisms of P removal by CLDC materials, adapted with permission.[24]

As revealed in Figure a, the XRD results showed the precipitation of P with Ca on CLDC samples, which agreed with the EDS characterization in Figure . To intuitively compare the P adsorption capacities of CLDC400 and CLDC800, the morphologies and chemical compositions of P-adsorbed carbon materials were analyzed by SEM–EDS, as shown in Figure . After P adsorption, the P content of CLDC400 was higher than that of CLDC800 (Figure ). CLDC400 and CLDC800 were doped with CaCO3, but there was an obvious difference in their surface structures and chemical functional groups (Figures and S2), leading to significantly different adsorption mechanisms. Figure b shows obvious changes in the adsorption peaks after P removal. Peaks appeared at 1039 and 1300 cm–1, which were associated with P–O stretching vibration and adsorption.[25] The mechanisms of CLDC400 and CLDC800 before and after P adsorption were further investigated. The highest adsorption capacity of CLDC400 at 30 °C was 37.49 mg/g, which was 3.2 times higher than that of CLDC800. Similar results were observed in which the MgO-modified BC promoted P removal.[50] Chemisorption is the main force attributed to the sorption of P on Ca-doped BC, and the doped Ca played an important role in adsorbing P from aqueous solutions. Interestingly, Ca2+ was readily released from CaCO3 on the surface of CLDC400 under acidic conditions, which promoted P removal and it is likely due to the formation of amorphous calcium phosphate.[53] In addition, CLDC400 was rich in −COOH (1620 cm–1) and −OH (3460 cm–1) substituents necessary for P adsorption behavior, where CLDC400 mainly served as a chelating agent for removing P (Figures b and 6). Under acidic conditions (pH ≤ 6.0), the HPO42– adsorption mechanism was as follows: CLDC400 could release −OH into the liquid due to HPO42– substituents and form CaHPO4–C complexes on the surface of CLDC400. The peak at 935 cm–1 associating with the stretching of P–O[54] was observed in CaHPO4–C (Figures b, 5d,h). A similar report showed that at pH 3.5, the stoichiometry of −OH release (−OH/P ratio) increased with increasing surface coverage.[55] The Langmuir model and the pseudo-second-order-kinetic model showed better fits, further confirming the chemisorption of P onto CLDC400. The yield of CLDC800 was lower than that of CLDC400, and CLDC800 possessed a larger pore volume and a higher SSA. The aromaticity increased, whereas the polarity decreased. The P in solution exists in the species of HPO4– and HPO42– when the pH of P solution is between 5.2 and 7.9, and these P species interact with Ca to produce CaHPO4 and Ca(H2PO4)2 through hydrogen bonding or chemical precipitation (Figure ).[24]

Figure 7

CLDC surface characteristics and elemental components after adsorption under the conditions of initial 5.1 mg/L P, 30 °C, contact time: 24 h, pH 7, and 1.2 g of CLDC400 (a) or 3.0 g of CLDC800/L (b).

However, the high SSA allowed a small amount of dissolved inorganics inside CLDC800; these compounds were not readily extracted with hydrochloric acid and deionized water because of the dense graphite structure of CLDC800. Although CLDC800 and CLDC400 contained similar contents of Ca (Supporting Information, Figure S2), Ca2+ could not be readily released from CLDC800, thereby lowering P removal from aqueous solutions. In addition, the initial P solution pH not only affects calcium compound solubility but also determines the P species in solution. The primary species in solution at pH 3.0 was H2PO4– (98.5%), which most favored the adsorption process.[36] An increase in the P solution pH changed the positively charged surface because of deprotonation, which caused repulsive electrostatic interactions toward P. Kong et al. identified hydroxylapatite [Ca5(PO4)3 (OH)] as the main P crystalline phase after the sorption of P on Ca-loaded BC.[26] This indicated that the high sorption capacity of Ca-doped BC for P was attributed to the crystallization of P and Ca(OH)2. Thus, the BC rich in Ca exhibited an obvious advantage in recycling P from wastewater through sorption-induced crystallization.[26] They also found that the Ca-doped BC after P adsorption contained C (8.07%), O (45.15%), Ca (30.90%), and P (14.32%).[26] Therefore, for CLDC400, the P removal mechanisms were mainly related to electrostatic interactions, chemical reactions (chemical precipitation), and surface complexation (e.g., interaction between −OH and P). The P sorption behavior kinetics followed the pseudo-second-order model, which indicated that the P adsorption behavior was mainly controlled by surface chemical reactions. Similar mechanisms were found in chemical reactions, and the slowest step was controlled by the internal diffusion process.[56] For CLDC800, physicochemical adsorption played a dominant role in the P removal process. The P adsorption capacities with CLDC800 were associated with the increase in its SSA and porosity. In addition, the zeta potential of the CLDC leaching solutions was negative (Figure a), making it difficult for the CLDC materials to adsorb negatively charged PO43– because of electrostatic interactions. In addition, the qualitative and quantitative analyses and structure identification of CLDC material before and after sorption was performed by XPS method. The XPS results further revealed the main mechanisms of P removal by CLDC400 and CLDC800, which were described in Section . The results indicated that the CLDC composition, surface structure, SSA, and solution pH affected the P sorption, which included physical and chemical sorption, precipitation, and complexation (Figure ).[24]

P-Loaded CLDC Materials as a P-Based Fertilizer

The P mass ratio of CLDC was approximately 38 mg/g, suggesting that CLDC400 adsorbed P could serve as a phosphorus fertilizer.[26] P could be dissolved back into aqueous solutions by subsequent control of the solution pH value. The release of P from P-loaded CLDC samples by deionized water with different pH values was conducted and is described in Figure d. P desorption highly depended on the solution and soil pH value. The P release of P-loaded CLDC400 and CLDC800 in deionized water with a pH of 7.0 was 3.96 and 36.87%, respectively (Figure d). However, the P release increased to 44.2% for P-loaded CLDC400 and 40.4% for P-loaded CLDC800 when the solution pH decreased to 5.0 (Figure d). The different release rates of CLDC400 and CLDC800 also indicated their different adsorption mechanisms. CLDC400 and CLDC800 rich in, P and C, are capable of improving the soil properties such as fertility and plant nutrient contents, causing enhanced crop productivity through soil pH regulation, macrotrace elemental addition, and microbial activity improvement.[57] Thus, the P-loaded CLDC materials could be regarded as a slow-release and eco-friendly P-fertilizer.

Conclusions

An in situ approach for the fabrication of CaCO3-doped BC using CL has been developed. The P adsorption behaviors and mechanisms of CLDC materials were compared. The highest adsorption capacity of CLDC400 at 30 °C was 37.49 mg/g, and 3.2 times higher than that of CLDC800. CaCO3 particles on the BC surface made the CLDC material more favorable to P precipitation reactions and formed Ca3(PO4)2 and CaHPO4. Moreover, the two functional groups (−OH and −COOH) on the surface of CLDC400 played important roles in P complexation reactions. The P adsorption process onto CLDC800 was controlled by physicochemical adsorption due to the porosity and high specific surface. Considering the sustainable adsorbent fabrication cost and adsorption performance, CLDC400 exhibited great potential for P adsorption and soil fertility applications.