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Article

Comparison of the Sorption of Cu(II) and Pb(II) by Bleached and Activated Biochars: Insight into Complexation and Cation–π Interaction

by
Jing Zhao
1,2,
Lin Wang
1,3 and
Gang Chu
1,4,*
1
Yunnan Provincial Key Laboratory of Soil Carbon Sequestration and Pollution Control, Faculty of Environmental Science and Engineering, Kunming University of Science and Technology, Kunming 650500, China
2
Faculty of Metallurgical and Energy Engineering, Kunming University of Science and Technology, Kunming 650093, China
3
Faculty of Material Science and Engineering, Kunming University of Science and Technology, Kunming 650093, China
4
Faculty of Resources and Environment, Anhui Agricultural University, Hefei 230036, China
*
Author to whom correspondence should be addressed.
Agronomy 2023, 13(5), 1282; https://doi.org/10.3390/agronomy13051282
Submission received: 8 April 2023 / Revised: 22 April 2023 / Accepted: 24 April 2023 / Published: 29 April 2023
(This article belongs to the Special Issue Environmental Ecological Remediation and Farming Sustainability—Volume II)
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Versions Notes

Abstract

:
Attention has been paid to the application of biochar in the remediation of heavy metal contamination in soils. In this study, two modification methods, bleaching and activation, were used to enhance the biochar sorption of Cu(II) and Pb(II). Multiple techniques, including XPS, FTIR and 13C NMR, were applied to investigate the properties of bleached and activated biochars. Combined with density functional theory (DFT) calculations, structural changes in the biochars and the main mechanism of Cu(II) and Pb(II) sorption were discussed after modification. The bleached biochar without non-condensed aromatic C possessed many oxygen-containing groups due to oxidation. In contrast, activation increased the specific surface area (SSA) and removed the carboxyl groups. Both modifications had an excellent effect on sorption by high-temperature biochars, especially the bleaching treatment. The bleached and activated biochars exhibited superior Pb(II) sorption compared with Cu(II) due to the larger electron cloud configuration of Pb(II). The significantly positive correlation of the Kd values with the COOH/C=O content indicated that the oxygen-containing groups were responsible for Cu(II) and Pb(II) sorption. The DFT calculation demonstrated a higher energy for the cation–π interaction than for the complexation for Cu(II) sorption, whereas the opposite case was observed for Pb(II) sorption. The complexation and cation–π interaction were the main mechanisms of Cu(II) and Pb(II) sorption. This study provides important guidance for the application of modified biochars in the sorption of heavy metals in the environment.

1. Introduction

As globally significant environmental contaminants, heavy metals have received significant attention due to their persistence and biotoxicity. A major obstacle in sustainable development is the global contamination of agricultural soil caused by heavy metals, which seriously threatens food safety and human health. Zhao et al. reported that 19% of the agricultural soil in China is contaminated by heavy metals, resulting in pollution losses totaling tens of billions of Chinese yuan [1]. In Russia, 10.8% of the land is moderately or heavily contaminated with heavy metals [2]. Pérez et al. estimated that 2.8 million soil sites in the European Union are polluted [3]. By 2050, the global agricultural output is expected to double to meet the estimated demand of a growing population with higher living standards [4]. Therefore, finding a solution for the heavy metal contamination of agricultural soils is vital.
Industrialization is a major cause of heavy metal soil contamination on agricultural land, with Cd, As, Cu, Hg, Pb, and Cr among the most prevalent heavy metals found in soil. Cu contamination in agricultural soil is also caused by mining, metallurgy, and the use of chemical fertilizers. Although Cu is an essential element for plant growth, it is also a priority pollutant to control [5] because excessive Cu can give rise to metabolic disorders and induce biological diseases in the environment. Pb precipitates in the soil through smelting, fuel burning, battery use, and transportation. The annual global production of Pb has doubled over the last 50 years to 11.3 million tons [4]. Meanwhile, as a heavy metal contaminant, Pb can harm the nervous and digestive systems [6].
Effective remediation strategies, including physical, chemical, and biological methods such as soil washing, hyperaccumulator phytoremediation, and sorption fixation, have been developed to reduce the dangers of heavy metals in contaminated soils. Sorption is a widely used remediation strategy. Biochar is a sorbent produced from a renewable biomass containing abundant organic functional groups, pore structures, polytropic carbon structures, and synaptic plasticity [7,8]. These excellent characteristics provide biochar with a better sorption effect for heavy metals.
Nonetheless, the sorption of heavy metals onto biochar is complex. The primary mechanisms of heavy metal sorption onto biochar include electrostatic attraction, ion exchange, surface complexation, and precipitation. The sorption mechanisms depend on various factors, including the properties of the biochar (surface area, pore size distribution, and functional groups), the properties of the heavy metal ions (e.g., charge, size, and valence), and the environmental conditions (pH, temperature, and concentration) [9,10,11,12]. Among biochar’s properties, its organic functional groups are widely considered the main sorption sites for heavy metals [13,14]. For traditional pyrolytic biochar, the organic functional group content is limited. Therefore, many studies have used chemically modified biochar to increase the functional group content [15,16,17]. Unlike physical and biological modifications, the chemical modification of a biochar is more specific through the application of functional groups, such as hydroxyl, carboxyl, and amine groups, to the surface of the biochar to increase its sorption sites and capacity. The oxygen-containing functional groups contain many components and are highly focused on the sorption of heavy metals.
The oxidation treatment is one of the most universal chemical modification methods for enriching oxygen-containing functional groups [18,19,20,21]. Several common oxidants, such as H2O2, HNO3, and KMnO4, can be used to improve oxygen-containing biochar groups for heavy metal sorption [22,23]. Bleaching is a chemical oxidation modification method that allows for the oxidation of the biochar surface, although the bleaching treatment can remove a portion of the aromatic carbon fraction from the biochar. It was reported that bleached biochars demonstrated good adsorption performance for organic contaminants [24]. However, studies on the sorption of heavy metals are limited. Activation treatment is also a frequently used chemical modification method to enhance the capacity of contaminant sorption on biochars. Unlike bleached biochar, activated biochar generally possess a large specific surface area (SSA) to effectively remove environmental contaminants. As a biochar modification method, activation treatment produces biochar that are different from traditional activated carbon. Kołodyńska et al. reported that the heavy metal sorption capacity of biochar is higher than that of traditional activated carbon [25]. However, as a means of activation treatment, the sorption effect of heavy metals on activated biochar is unclear. Additionally, bleaching and activation treatments can provide more information on the carbon structure with respect to the oxygen-containing functional groups and aromatic carbons after modification treatment to benefit the understanding of heavy metal sorption on biochar.
Cu(II) and Pb(II) were selected as heavy metal representatives for the investigation of the sorption performance of bleached and activated biochars. The objectives of this study were (i) to compare the structures and properties of bleached and activated biochars and (ii) to evaluate the sorption capacities and underlying sorption mechanisms of Cu(II) and Pb(II) by bleached and activated biochars.

2. Materials and Methods

2.1. Preparation and Chemical Modified of Biochars

Rice straw was collected from an agricultural setting, dried, milled, and screened using 60 mesh sieves. To produce biochar, the rice straw, after pretreatment, was pyrolyzed with N2 at 300, 500, or 700 °C for 4 h in a chamber muffle furnace equipped with a gas cylinder (KSW, Yongguangming, Beijing, China). The original biochar samples were designated O300, O500, and O700. NaClO2, NaNO3, NaOH, and CH3COOH were purchased from Aladdin Bio-Chem Technology Co, Shanghai, China. For bleaching treatment, a mixture of 100 g of NaClO2, 100 mL of CH3COOH, and 1000 mL of deionized water was used to treat 10 g of each biochar sample for 5 h. Two additional sessions were conducted for treatment. The bleached biochar samples were identified with the prefix BL (bleached), specifically using the digits BL300, BL500, and BL700. For the activated treatment, solid NaOH (1 g biochar with 5 g NaOH) and original biochar were combined and heated for 10 min at 150 °C. This process was repeated three times. The activated biochars were identified with biochars activated via NaOH (AN), specifically labeled as AN300, AN500, and AN700. All samples were repeatedly washed with deionized water to 7.0 ± 0.5 and then freeze-dried for subsequent use.

2.2. Characterization of Original and Bleached Biochars

An elemental analyzer (Vario MicroCube, Elementar, Langenselbold, Germany) was used to measure the elemental content of each sample in triplicate. The atomic ratio was calculated using the specific value of the elemental content and molar mass of the element. A larger (N + O)/C value represents a higher polarity, and a larger H/C value represents a lower aromaticity for biochar particles. The original, bleached, and activated biochars were evaluated using an N2 physical adsorption instrument (Autosorb-1C, Quantachrome, Boynton Beach, FL, USA) with volumetric gas sorption to determine the SSA. X-ray photoelectron spectroscopy (XPS; PHI 5000 Versaprobe-II, Ulvac-Phi, Chigasaki, Japan) was used to measure the surface elemental compositions of all samples. We acquired the surface organic carbon bond component by fitting the peaks of the C 1s and O 1s spectra using the MultiPak software. Using a Fourier transform infrared (FTIR) spectrometer (Varian 640-IR, Agilent, Santa Clara, CA, USA), the functional groups of the biochar were examined before and after bleaching and the activation of chemical modifications. The FTIR spectra of all samples were measured in a dry environment between 400 and 4000 cm−1 with potassium bromide-based techniques. To better understand the structural characteristics of the original, bleached, and activated biochars, solid-state 13C nuclear magnetic resonance (NMR) spectroscopy (JNM-ECZ600R, JEOL, Tokyo, Japan) was used to obtain information about the organic carbon structures of the biochars.

2.3. Batch Sorption Experiment

Cu(NO3)2 and Pb(NO3)2 (Aladdin Bio-Chem Technology Co., Shanghai, China) solutions at a concentration of 50 mg/L were produced separately as Cu(II) and Pb(II) stock solutions in the background solution (0.01 mol/L NaNO3). Cu(II) and Pb(II) stock solutions were diluted to obtain nine concentrations ranging from 1 to 50 mg/L. The pH value is a crucial factor for the sorption process of heavy metals. The pH of the solution can influence the deprotonation of functional groups of biochars and the activity of heavy metal ions. Thus, we adjusted the background solution with NaOH and HNO3 to maintain a pH value of 6.0 ± 0.2 during the sorption of Cu(II) and Pb(II). Glass vials were agitated in a shaker at 25 °C for 5 days as part of the sorption experiment. Based on the preliminary experimental results, a 5-day period was considered sufficient for both compounds to reach sorption equilibrium. After 5 days, the mixtures were centrifuged at 1000× g for 10 min. The supernatant solution was filtered using a 0.45 μm membrane filter, and the concentrations of filtered solution in each vial were determined using flame atomic absorption spectroscopy (FAAS, Hitachi, Tokyo, Japan). All samples were analyzed in duplicate.

2.4. Data Analysis

Using SigmaPlot 10.0, we applied the Freundlich, Langmuir, Toth, and Sips models to fit the sorption isotherms of Cu(II) and Pb(II) on the biochars as follows:
Langmuir model: Se = KLQCe/(1 + KLCe)
Freundlich model: Se = KF Cen
Sips model: Se = (KSBCen)/(1 + BCen)
where Se (mg/g) is the equilibrium solid-phase concentration and Ce (mg/L) is the equilibrium aqueous-phase concentration. The Freundlich model coefficient is represented as KF [(mg/g)/(mg/L)n], where n is a nonlinear factor. In the Langmuir model, Q (mg/kg) is the sorption capacity and KL (mg/L) is the Langmuir model coefficient. To explain heterogeneous sorption, the Sips model was derived from the Langmuir and Freundlich models.

2.5. Density Functional Theory Calculation

The Cu(II)/Pb(II) sorption process at the molecular level and the energy shift between atomic contacts was simplified with Density Functional Theory (DFT) calculations. The DMol3 package in the Materials Studio 2019 software was used for theoretical calculations and geometric optimization. The electronic basis was set as DND 3.5 with the function of B3LYP. We established an aromatic model of biochars and a -COOH model of biochars to calculate the energies of complexation and cation–π interactions. The interaction energy (ΔE) was described as
ΔEcomplexation = ECOOH-ion − Eion − ECOOH
ΔEcation–π = EAr-ion − Eion − EAr

3. Results and Discussion

3.1. Elemental Composition

The total and surface elemental compositions of the original, bleached, and activated biochars prepared at 300, 500, and 700 °C are shown in Figure 1. The primary elements of these biochars consisted of C, O, and H, as well as a slight amount of N. For original biochars, the total C was increased and the total O was decreased with an increase in the pyrolysis temperature. The results are ascribed to the volatilization of hydrocarbon and oxygenated moieties during carbonization [26]. The alteration in the elemental compositions of bleached and activated biochars was similar to the original compositions with pyrolysis temperature, except for BL300. We observed that BL300 had the highest O and lowest C contents, suggesting that bleaching treatment effectively removed aromatic carbon and grafted more oxygen-containing groups onto the particles of O300. For activated biochars, a higher degree of carbonization was observed, which was supported by more C content. Compared with bleached biochars, activated biochars had more C and less O content.
The atomic ratios of (N + O)/C and H/C represent the polarity and aromaticity of the carbonaceous structure [27,28]. Figure 1 shows the changes in the polarity and aromaticity of original, bleached, and activated biochars with the pyrolysis temperature. The polarity was gradually decreased, and aromaticity was progressively enhanced, consistent with previous studies [29,30,31]. The bleached biochars possessed high polarity, likely facilitating their complexation with some heavy metals in the aqueous phase. In contrast, high aromaticity was observed for activated biochars [32,33]. It is worth investigating whether high aromaticity of the sorbent is beneficial to the immobilization of heavy metals [34,35]. The distributions of surface element atomic ratios were in line with the total elemental composition under the pyrolysis temperature. Interestingly, the surface O content of biochars was lower than that of the total O content because the surface elements were eliminated and rearranged more strongly through heat conduction. Furthermore, it was noted that bleaching and activation treatment changed the SSA of biochars due to the modification of porous structure. We observed that the SSA followed the following order at the same temperature (Table S1): activated biochar > original biochar > bleached biochar, which may have influenced the migration of heavy metals.
Deconvolution on the XPS spectra revealed the multiple structural patterns of the C 1s and O 1s peaks in biochars. The five peaks for C=C, C–C, C–O, C=O, and π–π are presented in Figure 2. The C=C and C–C peaks were observed at 284.1 and 284.8 eV, respectively. The C–O peaks, including the C–OH and C–O–C species, were observed at 286.5 eV, and the C=O peaks, such as O=C–OH, were found at 288.6 eV. At 291.6 eV, the deconvoluted peak was attributed to the π–π region [36]. The ratio of the C=C peaks of bleached biochar was higher than that of original and activated biochars, whereas the ratio of the C–C peaks of bleached biochar was lower. These results were caused by non-condensed aromatic removal via the bleaching treatment.
As shown in Figure 3, the deconvolution of the O 1s peak exhibited the presence of four peaks including –OH (528.4 eV), –C=O (530.4 eV), –C–O (531.7 eV), and O=C–O (532.9 eV). Among these species, C–O was the main oxygen-containing structure. With an increase in temperature, the C–O species evolved into C=O species. The C–O peak of bleached biochars was higher than that of the original and activated biochars. According to the area ratios of peak deconvolution (Table 1) and surface O content (Figure 1), the bleached biochars exhibited more oxygen-containing groups, such as carboxyl moieties, due to chemical oxidation. In contrast, activation of the original biochar gave rise to a reduction in the oxygen-containing groups and an increase in C–C species.

3.2. Functional Groups and Carbon Species

FTIR (Figure 4) and 13C NMR (Figure 5) spectra show the functional groups and carbon structure of the biochar after bleaching and activation. A broad absorption band at 3467 cm−1 represented the vibration of –OH. With an increase in temperature, a reduction in –OH stretching was observed. The vibration of –CH2 and C–H belonged to the alkyl structure of aliphatic carbon at 2956 and 680 cm−1, respectively. The activated biochars clearly had more –CH2 moieties compared with the original biochars. This result indicates that the aromaticity of the original biochars improved after alkaline activation. The peaks at 1720 and 1660 cm−1 referred to carboxylic and other C=O stretching, respectively. The C–O structure was observed distinctly at 1116 cm−1. These oxygen-containing groups were gradually diminished with pyrolysis temperature, which occurred in accordance with the polarity of biochars. Notably, BL300 and AN300 contained more oxygen-containing groups, especially the C–O structure, which was consistent with the results of the XPS spectra. However, the oxygen-containing groups of activated biochars were obviously diminished from 300 to 700 °C. In this temperature range, the superiority of the bleaching treatment was exhibited by the greater preservation of oxygen-containing groups compared with that under alkaline activation.
Quantitative solid-state 13C NMR spectroscopy was used to determine the molecular carbon structure of the biochar (Figure 5). Notably, the aliphatic and aromatic carbon components were quantified for rice straw, original, bleached, and activated biochars. The abundant aliphatic carbon was the representative structure of rice straw. The bleaching and activation treatments had little influence on the carbon structure of raw feedstock. The aliphatic and non-condensed aromatic moieties predominantly consisted of carbonaceous structure in low-temperature biochar, leading to the formation of micro-scale porous structures. With an increase in temperature, peaks at 108–165 ppm began to appear, indicating the presence of aromatic carbon after cyclization and condensation. Aliphatic carbon was gradually reduced and aromatic carbon significantly increased with increasing pyrolysis temperature. Notably, the aromatic carbon of the bleached biochars was much less than that of the original and activated biochars at the same temperature, especially for BL300. This result suggests that the non-condensed aromatic carbons were the primary carbon structure of O300 and were removed in great numbers via the bleaching treatment. Although the total aromatic carbon decreased with the bleaching treatment, this process also enhanced the mass ratios of condensed aromatic carbon to total carbon. In addition, the amount of aromatic carbon of the activated biochars was greater than that of the original biochars, which resulted from the strengthening of carbonization though the activation treatment.
The total and surface aromatic C and COOH/C=O content are illustrated via the 13C NMR and XPS calculations in Figure S1. Relative to the original biochars, the total aromatic C of the bleached biochars decreased, but the surface aromatic C increased. The accumulation of surface aromatic C in the bleached biochars may have contributed to the formation of cation–π interactions. The COOH/C=O content of bleached biochars was greater than that of the activated and original biochars, especially BL500 and BL700. This result is ascribed to structural oxidation though bleaching. The high COOH/C=O content of bleached biochars can provide more sorption sites for complexation with heavy metals. For activated biochars, the total and surface aromatic C were higher than those of the original biochars. Moreover, the total aromatic C of the activated biochars was higher than its surface aromatic C, indicating a higher degree of carbonization for the inner structure of biochar after the activation treatment. In addition, the surface COOH/C=O content of the activated biochars was less than that of the original and bleached biochars.

3.3. Cu(II) and Pb(II) Sorption

The sorption isotherms of Cu(II) and Pb(II) on the original, bleached, and activated biochars are presented in Figure 6. All sorption isotherms exhibited nonlinear patterns after taking the logarithm. The fitting results of the sorption isotherms with the Freundlich, Langmuir, and Sips models are listed in Table 2. The Langmuir model offered a better fit for the isotherms than the Freundlich model. The fundamental premise of the Langmuir model is that sorption occurs homogeneously on the adsorbent surface. This premise suggests that the homogeneous sites of biochars are conducive to surface sorption, especially for low-temperature biochars. The isotherms of low-temperature biochars were fitted well with the Langmuir model, as indicated by the high radj2 value (Table 2), which could be ascribed to the control of heavy metals by rich oxygen-containing groups. As shown by the FTIR spectra, low-temperature biochar contained many groups of C–O and C=O that were important sites for the sorption of Cu(II) and Pb(II). The Sips model provided the best description of the sorption process among the various models (Table 2).
The nonlinear coefficient (n) values indicated the heterogeneity of the effective sites for Cu(II) and Pb(II) sorption [37]. The smaller n values of Cu(II) and Pb(II) sorption caused by bleached biochars indicated more sites with heterogeneity among bleached biochars compared with activated and original biochars, due to a supplement from the oxygen-containing groups. Oxygen-containing groups are conducive to complexation with heavy metal, especially carboxyl groups. Lee et al. reported that acidic functional groups bind with Cd(II) through an inner- and outer-sphere complexation process [38]. In this research, the deprotonation carboxyl groups easily bound with metal ions through electrostatic attraction to form outer complexation. The electrostatic attraction between deprotonation carboxyl groups and Cu(II) and Pb(II) could improve the probability of forming stable inner-sphere complexes. The bleached biochars had a higher sorption capacity for Cu(II) and Pb(II) than that of the activated biochars. Especially for BL500 and BL700, the greater number of oxygen-containing groups provided many sorption sites for complexation with Cu(II) and Pb(II). According to the isotherms, Pb(II)’s sorption capacity was greater than that of Cu(II) sorption on the same biochar, which was associated with the metal ion structure. Pb(II) possesses a larger configuration of electron clouds than Cu(II) [39]. These large electron cloud increased the probability of interaction with the sorbent. Thus, the large electron clouds of the Pb(II) ion facilitated Pb(II)’s sorption on biochars. Although Cu(II) has a smaller electron cloud, the outer electron configuration of Cu(II) can provide an empty d orbital that is able to obtain external electrons from the sorbent to facilitate sorption [40]. This configuration decreased the difference in sorption capacity between Cu(II) and Pb(II).

3.4. Sorption Mechanism of Cu(II) or Pb(II)

The sorption capacity of Cu(II) and Pb(II) decreased slightly on BL300 and AN300, respectively, compared with that of O. This phenomenon was caused by the elimination of non-condensed aromatic carbon and fractional oxygen-containing groups for bleached and activated biochars, suggesting the importance of varied structural sites for Cu(II) and Pb(II) sorption. Figure 5b shows that bleaching caused an ~83% decline in the aromatic C concentration of BL300 as per the calculations of 13C NMR. An approximately 13% reduction in COOH/C=O components was also observed for AN300. Thus, Cu(II) and Pb(II) sorption was inhibited (Figure 6b,f). This result is attributed to a reduction in cation–π interactions and complexation for bleached and activated biochars, respectively. With an increase in temperature, both bleached and activated biochars exhibited higher sorption capacity to Cu(II) and Pb(II) than the original biochar. This result indicates that bleaching and activation successfully enhanced the sorption capacity of high-temperature biochar.
In addition, the SSA of the biochar was significantly enhanced as a function of temperature, which may affect sorption behavior. Peng et al. reported that SSA played a vital role in the sorption of heavy metal on modified biochar produced at high temperatures [41]. In our study, the SSAs of bleached and activated biochars were enlarged with the pyrolysis temperature, and the activated biochars exhibited the largest SSA. Thus, we investigated the relationship between the SSA value and single-point sorption coefficient (Kd). Figure S2a shows the lack of correlation between the SSAand Kd of Cu(II) and Pb(II) sorption, indicating that pore-filling could not be the main mechanism for the sorption of heavy metals. However, the significant and positive correlation of COOH/C=O content with Kd value (Figure 7a) suggests that the oxygen-containing groups were responsible for complexation with Cu(II) and Pb(II). Moreover, we observed a negative relationship between aromatic C content and the Kd value of high-temperature biochars (Figure 7b). This negative correlation was contributed by the large Kd values of BL500 and BL700. Bleaching treatment concentrated the condensed aromatic C of high-temperature biochars, which contributed to the sorption of Cu(II) and Pb(II) with BL500 and BL700. This correlation suggests that condensed aromatic C is an important structure for the formation of cation–π interactions between Cu(II) and Pb(II). Cao et al. reported that high aromaticity contributed to Pb(II) sorption due to the formation of cation–π bonds [42]. Yang et al. also reported that cation–π interactions could be formed between metal ions and aromatic carbon [43].
Figure 8 shows the molecular interpretation of the energies of complexation and cation–π interactions based on DFT calculations. The high binding energy represents the strong strength of the interaction with heavy metals. The results clearly showed that the Cu(II) combination provides stronger energies through complexation and cation–π interactions than the Pb(II) combination. The relatively high complexation energies of Cu(II) and Pb(II) were −12.4 and −9.3 eV, respectively. These results indicate that complexation is a crucial mechanism for Cu(II) and Pb(II) sorption. For Cu(II) sorption, the energy of cation–π interactions (−14.8 eV) was larger than that of complexation (−12.4 eV). This result suggests that Cu(II) was more conducive to absorption onto aromatic carbon due to cation–π interactions. Conversely, the complexation process plays a vital role in Pb(II) sorption, as the complexation energy was much higher than the cation–π interaction energy for Pb(II) sorption. Compared with the sorption energy of Pb(II), the stronger energy produced via combination with Cu(II) indicates that the stable immobilization of Cu(II) was due to complexation and cation–π interactions. This result suggests that Cu(II) ions may have difficulty achieving desorption from these biochar particles.
According to the sorption isotherm, however, the contribution of cation–π interactions for Cu(II) sorption did not exhibit a significant advantage over Pb(II) sorption. This result was likely due to the large electron cloud and smaller hydration radius of Pb(II), which enabled Pb(II) ions to interact with the biochar surface and to enter the nano-scale pores. Notably, BL500 and AN500 exhibited excellent sorption capacity for Cu(II) and Pb(II) because BL500 and AN500 possessed, respectively, a relatively high content of oxygen-containing groups and aromatic carbon. Overall, the sorption of Cu(II) and Pb(II) was simultaneously dominated by complexation and cation–π interactions.

4. Conclusions

In this study, Cu(II) and Pb(II) sorption was studied using biochars before and after bleaching and activation. The oxygen-containing groups of rice-straw-derived biochar were enhanced by bleaching, and the aromatic C content was increased by activation. Both bleaching and activation modifications helped improve the biochar’s sorption capacity on heavy metals, especially for high-temperature biochars. The high-temperature bleached biochars exhibited the greatest sorption capacity for Cu(II) and Pb(II). A significant and positive correlation was established between the sorption coefficient and oxygen-containing groups, suggesting the importance of oxygen-containing groups in regulating sorption. The configuration of Pb(II) with a larger electron cloud contributed a sorption capacity greater than that of Cu(II). Based on the DFT calculations, the energy of cation–π interactions was higher than that of complexation for Cu(II) sorption, but the opposite case was observed for Pb(II) sorption. Overall, the sorption mechanisms of Cu(II) and Pb(II) using bleached and activated biochars mainly involved complexation and cation–π interactions. This study provides technical guidance for the application of biochar modification to enhance the sorption of Cu(II) and Pb(II).

Supplementary Materials

The following supporting information can be downloaded at https://www.mdpi.com/article/10.3390/agronomy13051282/s1, Figure S1: Ratio of total aromatic carbon (a) and COOH/C=O (b) was acquired from the 13C NMR and the surface aromatic carbon (c) and COOH/C=O (d) of original, bleached, and activated biochars at 300, 500, and 700 °C from XPS C 1s spectra; Figure S2: Relationship between Kd of Cu(II)/Pb(II) sorption and SSA (a)/aromaticity (b) of original, bleached, and activated biochars at 300, 500, and 700 °C; Table S1: Specific surface area (SSA) of original, bleached, and activated biochars at 300, 500, and 700 °C with N2.

Author Contributions

Writing—Original Draft Preparation, Formal Analysis, Calculation, and Data Curation, J.Z.; Method, Calculation supporting, and Editing, L.W.; Supervision, Investigation, and Editing, G.C. All authors have read and agreed to the published version of the manuscript.

Funding

This research was supported by the National Natural Science Foundation of China (42107258 and 42107406), Yunnan Major Scientific and Technological Projects (202202AG050019), and China Postdoctoral Science Foundation (2020M673591XB).

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

Not applicable.

Conflicts of Interest

The authors declare no conflict of interest.

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Agronomy 13 01282 g001 550
Figure 1. Elemental compositions of original, bleached, and activated biochars at 300, 500, and 700 °C.
Figure 1. Elemental compositions of original, bleached, and activated biochars at 300, 500, and 700 °C.
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Figure 2. C 1s XPS spectra of original, bleached, and activated biochars at 300, 500, and 700 °C.
Figure 2. C 1s XPS spectra of original, bleached, and activated biochars at 300, 500, and 700 °C.
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Figure 3. O 1s XPS spectra of original, bleached, and activated biochars at 300, 500, and 700 °C.
Figure 3. O 1s XPS spectra of original, bleached, and activated biochars at 300, 500, and 700 °C.
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Figure 4. FTIR spectra of rice straw (a), original, bleached, and activated biochars at 300 (b), 500 (c), and 700 °C (d).
Figure 4. FTIR spectra of rice straw (a), original, bleached, and activated biochars at 300 (b), 500 (c), and 700 °C (d).
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Figure 5. The 13C NMR spectra of rice straw (a), original, bleached, and activated biochars at 300 (b), 500 (c), and 700 °C (d).
Figure 5. The 13C NMR spectra of rice straw (a), original, bleached, and activated biochars at 300 (b), 500 (c), and 700 °C (d).
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Figure 6. Sorption isotherms of Cu(II) (a–d) and Pb(II) (e–h) on original, bleached, and activated biochars at 300, 500, and 700 °C, respectively.
Figure 6. Sorption isotherms of Cu(II) (a–d) and Pb(II) (e–h) on original, bleached, and activated biochars at 300, 500, and 700 °C, respectively.
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Figure 7. Relationship between Kd and COOH/C=O (a) and aromaticity (b) of original, bleached, and activated biochars.
Figure 7. Relationship between Kd and COOH/C=O (a) and aromaticity (b) of original, bleached, and activated biochars.
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Figure 8. The energy of cation–π interaction and complexation between biochar and Cu(II) (a) and Pb(II) (b) based on DFT calculation.
Figure 8. The energy of cation–π interaction and complexation between biochar and Cu(II) (a) and Pb(II) (b) based on DFT calculation.
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Table
Table 1. The peak-differentiating results of C 1s and O 1s XPS peaks for original, bleached, and activated biochars.
Table 1. The peak-differentiating results of C 1s and O 1s XPS peaks for original, bleached, and activated biochars.
SamplesXPS C 1s (%)XPS O 1s (%)
C=CC–CC–O–COOπ–π Bond–OHC=OC–OO=C–O
O0.0071.814.811.42.022.4512.581.73.36
O3003.4164.422.87.841.5116.815.657.99.75
O50018.768.54.496.102.290.6826.454.918.0
O7002.8572.018.44.712.025.1641.440.812.7
BL38.025.323.513.20.000.0018.973.87.27
BL30034.334.29.4720.81.1715.54.3868.211.9
BL50022.736.720.519.01.1815.813.160.410.7
BL70025.739.616.016.42.3413.119.556.610.8
AN55.97.2028.08.840.001.3222.463.313.0
AN3002.1138.230.329.30.103.3149.339.87.60
AN50028.934.927.37.781.092.7446.837.812.7
AN7007.9283.61.845.950.731.8554.622.321.2
Table
Table 2. The fitting results for Cu(II) and Pb(II) sorption isotherms on original, bleached, and activated biochars using the Langmuir, Freundlich, and Sips models.
Table 2. The fitting results for Cu(II) and Pb(II) sorption isotherms on original, bleached, and activated biochars using the Langmuir, Freundlich, and Sips models.
Langmuir ModelKd a (L/g)Freundlich ModelKd (L/g)Sips ModelKd (L/g)
Q (mg/kg)KL (L/kg)radj2Ce = 5 mg/LKF bnradj2Ce = 5 mg/LKS (mg/g)Bnradj2Ce = 5 mg/L
CuO14.730.0860.9070.8881.4740.5970.8360.5830.0230.2210.3880.9741.141
O30027.490.1410.9622.2764.2230.5140.8941.3780.1112.4460.6460.9782.524
O5001.5550.8340.9180.2511.2970.8340.9180.8850.0120.2780.5380.9720.898
O7004.6580.5670.8080.6892.6430.5670.8080.9760.0180.3320.4100.9751.752
BL21.990.0350.9610.6550.9210.7500.9340.5180.0250.3040.6090.9810.634
BL30034.270.0680.9151.7452.7830.6410.8681.2180.0050.0910.2810.9672.311
BL50011.620.4870.9051.6475.9750.1970.9790.9400.1955.1930.7860.9853.207
BL7004.5660.6040.8220.6862.2770.0770.8980.2720.0110.2010.3490.9851.964
AN16.500.0820.9540.9601.8050.5500.8870.6400.0320.3820.5120.9881.014
AN3009.6530.1410.9270.7971.5380.4850.8240.4700.0900.6760.5390.9680.961
AN50024.790.0850.9791.4840.5261.3330.9791.1331.49×10 −120.5260.7500.9770.899
AN70027.150.0820.9091.5732.7110.5860.8291.3920.0180.3320.4100.9751.752
Langmuir modelKd (L/g)Freundlich modelKd (L/g)Sips modelKd (L/g)
Q (mg/kg)KL (L/kg)radj2Ce = 5 mg/LKFnradj2Ce = 5 mg/LKS (mg/g)Bnradj2Ce = 5 mg/L
PbO11.550.2490.9001.2802.3410.4750.8940.6990.1822.7321.2920.8961.164
O30070.660.3480.9138.97915.740.6580.8597.1531.13350.550.4250.9938.747
O50056.470.0060.9390.3130.3470.9290.9350.2950.0080.1300.7000.9410.242
O70047.990.0080.9060.3470.3860.9170.9000.3190.0020.0210.4500.9240.139
BL9.7730.2100.9411.0022.0010.4520.9640.5670.0872.1351.7700.9640.871
BL30055.640.2780.8676.46910.870.5850.7904.1860.37514.950.3800.9737.684
BL50033.130.4300.9364.52327.101.8600.975196.30.35547.790.5930.98822.67
BL70026.610.2480.9692.9495.9460.4390.9001.6340.9906.5880.2490.9671.328
AN30.490.0100.9500.2790.3450.8710.9430.2570.0100.1250.7060.9530.224
AN30030.780.1960.9263.0477.3661.5010.95523.350.38010.790.3050.9915.612
AN50016.390.0190.9730.2820.0851.6140.9600.3480.0000.0850.6200.9570.227
AN70029.710.0320.9620.8171.4810.6720.9810.6961.4891.4810.4380.9800.196
a Kd is the single-point sorption coefficient; b KF [(mg/g)/(mg/L)n].
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Zhao, J.; Wang, L.; Chu, G. Comparison of the Sorption of Cu(II) and Pb(II) by Bleached and Activated Biochars: Insight into Complexation and Cation–π Interaction. Agronomy 2023, 13, 1282. https://doi.org/10.3390/agronomy13051282

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Zhao J, Wang L, Chu G. Comparison of the Sorption of Cu(II) and Pb(II) by Bleached and Activated Biochars: Insight into Complexation and Cation–π Interaction. Agronomy. 2023; 13(5):1282. https://doi.org/10.3390/agronomy13051282

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Zhao, Jing, Lin Wang, and Gang Chu. 2023. "Comparison of the Sorption of Cu(II) and Pb(II) by Bleached and Activated Biochars: Insight into Complexation and Cation–π Interaction" Agronomy 13, no. 5: 1282. https://doi.org/10.3390/agronomy13051282

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