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Comparison of the lead and copper adsorption capacities of plant source materials and their biochars
Journal of Environmental Management 236 (2019) 118–124
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Research article
Comparison of the lead and copper adsorption capacities of plant source materials and their biochars
T
Myoung-Eun Leea,1, Jin Hee Parkb,1, Jae Woo Chunga,c,∗ a
Department of Urban System Engineering, Gyeongnam National University of Science and Technology (GNTECH), Dongjin-ro 33, Jinju, Gyeongnam 52725, South Korea Department of Environmental and Biological Chemistry, Chungbuk National University, Cheongju, 28644, South Korea c Department of Environmental Engineering, GNTECH, Dongjin-ro 33, Jinju, Gyeongnam, 52725, South Korea b
A R T I C LE I N FO
A B S T R A C T
Keywords: Biochar Gingko leaf Peanut shell Metasequoia leaf Lead Copper
Lead (Pb) and Cu are the most common pollutants found in industrial effluents which affect ecosystem and human health. To remove Pb and Cu from aquatic system, cost-effective and environmentally friendly adsorbents are required. Therefore, the study evaluated the adsorption of Pb and Cu by waste plant materials and their biochars. The adsorption kinetics and isotherms were applied to compare the Pb and Cu adsorption capacities using the gingko (Spiraea blumei) leaf (GL), peanut shell (PS), and Metasequoia leaf (ML), and their derived biochars (GB, PB, and MB, respectively). The GB showed a significantly higher Pb adsorption capacity than the other adsorbents. Maximum Pb adsorption by GB was 138.9 mg/g followed by GL (117.6 mg/g). The highest Cu adsorption (59.9 mg/g) was also achieved by GB followed by GL (57.8 mg/g). The carbonates and the phosphate functional groups in the GB and higher affinity of Pb to the functional groups contributed to higher Pb adsorption. The Pb adsorption kinetics on the plant source materials and their biochars followed a pseudo-second order model. The Pb and Cu adsorption capacities, with the exception of the GL, ML, and GB, are better explained by Langmuir-isotherm models. The carbonization did not always lead to better heavy metal adsorption. The Pb and Cu adsorption significantly reduced with carbonization of ML because of disappearance of oxygen containing functional groups. Therefore, appropriate method to prepare metal adsorbent should be selected depending on feedstocks and metal removal mechanisms. The GL is the most-abundant fallen leaf in the streets of the Republic of Korea; therefore, the use of the GL biochar for heavy-metal adsorption will also reduce the cost for waste disposal.
1. Introduction The presence of heavy metals in the aquatic environment is a major concern due to their extreme toxicity. Lead(II) and Cu(II), in particular, are among the most common pollutants found in industrial effluents (Ren et al., 2012). According to the US Environmental Pollution Agency, they are highly toxic and can cause a variety of negative effects on human health even at low dosages, for example, anemia, encephalopathy, hepatitis, and the nephritic syndrome (Wan et al., 2010). Therefore, the removal of Pb(II) and Cu(II) ions from aqueous solution is an important step in water purification process. Conventional technologies for the removal of heavy metals from wastewater include adsorption, chemical precipitation, redox process, ion exchange, solvent extraction, electrolyzation, filtration, electrochemical removal, membrane separation and reverse osmosis (Ren
et al., 2011; Guo et al., 2017; Bilal et al., 2018; Fakhre and Ibrahim, 2018). Among these methods, adsorption has attracted much attention as an effective purification and separation technique for treating wastewater. Removal of heavy metals from wastewater can be achieved by application of adsorption processes using a suitable adsorbent (Han et al., 2006; Mohan and Pittman Jr, 2006). Suitable adsorbent can be characterized by low cost, high renewability and high removal capacity (Chen et al., 2018). Recently, considerable attention has been paid to the investigation of different types of low-cost adsorbents especially using biochar. It has been reported that the biochar-contained functional groups and alkaline pH conditions are responsible for the removal of the heavy metal ions from water (Uchimiya et al., 2010). Removal of Pb and Cu using various biochars derived from different feedstocks was tested in other studies. Silver birch biochar and Scot pine biochar adsorbed
∗
Corresponding author. Department of Urban System Engineering, Gyeongnam National University of Science and Technology (GNTECH), Dongjin-ro 33, Jinju, Gyeongnam 52725, South Korea. E-mail address: [email protected] (J.W. Chung). 1 Myoung-Eun Lee and Jin Hee Park contributed equally to the manuscript. https://doi.org/10.1016/j.jenvman.2019.01.100 Received 23 August 2018; Received in revised form 4 January 2019; Accepted 26 January 2019 0301-4797/ © 2019 Elsevier Ltd. All rights reserved.
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pH, electrical conductivity (EC), surface morphology, and functional groups to identify their major physicochemical properties. For the measurement of the pH and the conductivity, 5 g of each sample and 25 mL of distilled water were mixed for 1 h in a 50 mL conical tube. The pH and EC were measured using the pH and EC electrodes after calibration. The surface morphology was analyzed using field emission scanning electron microscopy (FE-SEM, JEOL, Japan). The samples were degassed at 250 °C for 10 h and Brunauer-Emmett-Teller (BET) surface area was evaluated by N2 adsorption using a gas sorption analyzer (Autosorb-iQ2, USA). Fourier transform infrared spectroscopy (FT-IR, Perkin Elmer, USA) was used for an analysis of the functional groups on the biochar surfaces. FT-IR spectra of KBr pellets containing samples were recorded in the 4000–400 cm−1 region with a resolution of 4.0 cm−1. An elemental analysis for C, H, N, S and O was performed using the Flash 2000 Series elemental analyzer (Thermo Fisher Scientific, USA). The ash contents were measured from the combustion of the dry samples at 760 °C for 6 h. The phosphorus concentration of the samples was analyzed using the Optima 8300DV inductively coupled plasma-optical emission (ICP-OES, Perkin Elmer, USA) after a nitric acid digestion of the samples to determine the contribution of phosphorus to the removal of metals.
128.7 mg/kg and 107.0 mg/kg of Cu, respectively, although Pb adsorption by those biochars ranged from 1.29 to 4.49 mg/kg (Komkiene and Baltrenaite, 2016). Tobacco stem biochar was also effective in the adsorption of Pb, Cd and Cu (Zhou et al., 2018). Metal adsorption capacity of biochar can be different depending on the feedstock type and biochar production conditions such as pyrolysis temperature and time (Ahmad et al., 2014). Inyang et al. (2016) demonstrated that the feedstock is an important parameter that determines the adsorption capacity of biochars, and that heavy metal removal mechanisms vary depending on the biochar type. Feedstocks that have been used for biochar production include broiler litter, poultry manure, plant materials including straw and leaves, food wastes such as orange peel and peanut shell, industrial byproducts including paper sludge and sewage sludge, and agricultural wastes such as rice husk and wheat straw (Ahmad et al., 2014). Biochars are normally produced from waste materials because of the low cost. Waste-materialbased biochar production can not only reduce the waste amount, but also replace fossil fuels in the production of energy (McHenry, 2009). Therefore, the study focused on the evaluation of waste materials to select feedstocks for biochar production as a cost effective metal adsorbent and at the same time waste disposal and reuse were considered. Although biochar is known to be effective in heavy metal adsorption, the adsorption capacity may not be always higher than its feedstocks. Mechanisms of metal adsorption are different depending on feedstocks and carbonization of feedstocks may differently affect metal adsorption capacity. Since biochar production requires energy and time, feedstock as a raw material would be better to be used as a metal adsorbent if the feedstock has higher metal adsorption capacity than its biochar. However, few studies have been conducted on the comparison of metal adsorption capacity between feedstock and its biochar. Therefore, heavy metal adsorption capacity of feedstock and its biochar was compared in this study and interpreted according to adsorption mechanisms. Plant source materials that can be easily obtained in natural environments in Korea were selected as the raw materials for this study. Gingko and Metasequoia leaves were first used as feedstocks for biochar in this study. Although peanut shell biochar was previously tested for Pb and Zn immobilization in soil by Chao et al. (2018), it was included in this study for comparison with other feedstocks. Therefore, the objective of the present study is the comparison of the Pb and Cu adsorption characteristics using different feedstocks and their biochars for the possible application of plant source materials in the treatment of metal-contaminated water.
2.3. Lead and Cu adsorption of the plant source materials and their biochars The produced biochars were finely ground with a mortar and pestle and passed through a sieve (< 0.5 mm) to prepare them for the adsorption experiments. To investigate the Pb and Cu adsorption characteristics of the plant source materials and their biochars, a heavy metal solution (50 mg/L) was reacted with the adsorbent for 48 h on the basis of 1 g/L in a rotary shaker at 200 rpm at 25 °C. The Pb and Cu stock solutions were prepared with Pb(NO3)2 and Cu(NO3)2, respectively. To prevent the effect of the pH on the adsorption, the heavy metal solutions were prepared using 0.07 M sodium acetate and 0.03 M acetic acid in pH 5 buffer solution (Chen and Wu, 2004). To investigate the kinetic heavy metal adsorption characteristics of the plant source materials and biochars, 1 g/L of each of the plant source materials and their biochars were reacted with 50 mg/L of the pH 5 Pb and Cu solutions at 25 °C. The Pb and Cu concentrations were analyzed at 0, 5, 10, 20, 30, 40, 60, 90, 120, 180, 360, 540, 1440, and 2880 min during the adsorption experiment. The pseudo-first order model and the pseudo-second order model were used to characterize the adsorption kinetics (Ho and Ofomaja, 2006). To observe the isothermal adsorption characteristics of the heavy metal ions, 1 g/L of each of the plant source materials and their biochars were reacted with the heavy metal ion solution of 30, 50, 100, 150, 200, 300, 400, and 500 mg/L for 48 h and the equilibrium was established. After the adsorption, the samples were filtered through a 0.2 μm filter (Advantec, Japan) and diluted with a 1% nitric acid solution. The metal concentrations in the solution were analyzed using ICP-OES. To evaluate and compare the Pb and Cu adsorption capacities of the plant source materials and their biochars, the Langmuir model and the Freundlich model were applied to fit the experiment data (Liu and Zhang, 2009; Pellera et al., 2012).
2. Materials and methods 2.1. Production of biochars from various plant source materials Gingko (Spiraea blumei) leaf (GL), peanut (Arachis hypogaea Linn.) shell (PS), and Metasequoia (Metasequoia glyptostroboides) leaf (ML) were selected as the plant source materials for the Pb and Cu adsorption. The plant source materials can be easily obtained during the autumn of the Republic of Korea. The collected plant source materials were washed with deionized water and dried in a drying oven at a temperature of 105 °C for 24 h. To produce the biochar, 5–10 g of the dried feedstock was placed in a crucible and purged with nitrogen (N2) for 5 min, followed by a 90 min electric-furnace carbonization at 800 °C. In the authors' previous study, the optimal carbonization temperature and time for the heavy metal adsorption of the peat moss and Ginkgo leaf-derived biochar were 800 °C and 90 min, respectively (Lee et al., 2015; Lee et al., 2017); therefore, the temperature and time for producing biochar was set to 800 °C and 90 min.
3. Results and discussion 3.1. Characterization of the plant source materials and their biochars The carbonization of the plant source materials developed a porous structure, as evidenced by the scanning electron microscopy (SEM) data of Supplementary Fig. 1. A porous structure was formed from the tubular plant cell structure (Puga et al., 2015). The hydrolysis reactions of the plant source materials and the release of the degradation products led to morphological changes (Liu and Zhang, 2009). The SEM morphological textures of the feedstock and the biochar product on the surface showed a great difference. As observed in feedstock samples,
2.2. Characterization of the plant source materials and their biochars The plant source materials and their biochars were tested for the 119
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Table 1 Brunauer-Emmet-Teller (BET) surface area and elemental compositions of the plant source materials and their biochars (AC: activated carbon, GL: gingko leaf, PS: peanut shell, ML: Metasequoia leaf, GB: GL-derived biochar, PB: PS-derived biochar, MB: ML-derived biochar). BET surface area (m2/g)
Samples
Activated carbon Plant source materials
Biochar
AC GL PS ML GB PB MB
1069 2 16 17 310 790 496
Elemental composition (%)
pH
C
H
N
S
O
Ash
P
76.07 45.22 49.50 50.33 42.48 74.06 40.61
2.04 5.56 5.82 5.92 0.47 1.29 0.58
0.69 1.21 0.60 1.14 1.65 1.38 0.71
3.09 0.73 0.14 0.24 1.85 2.73 0.26
12.72 33.99 42.32 37.63 17.96 13.80 38.04
5.40 13.40 1.63 4.75 35.60 6.74 19.81
0.006 0.338 0.028 0.058 0.934 0.117 0.173
7.69 4.60 5.27 4.19 12.43 12.88 10.85
aromatic C]C, which is commonly found in biochars and is the result of cellulose-derived transformations (Wu et al., 2012). The peaks of the GB at 876 and 1450 cm−1 have been assigned to the bands of the outof-plane bending for the carbonates (Yuan et al., 2011; Xu et al., 2013). In the region of 600–1500 cm−1 regarding the plant source materials, various peaks are evident including C]C, C–H, C–O, C–C, and OH. The broad peak within the 1000–1250-cm−1 region can be attributed to the inorganic components such as the sulfates and the silicates. The bands at 1115 and 1059 cm−1 can be attributed to the phosphate P–O bond (Cantrell et al., 2012). Lee et al. (2017) reported that the functional groups such as the H-bonded OH groups, aliphatic CeH, CeH2, CeH3 stretching, and aromatic C]C was produced as the carbonization temperature was increased. Functional groups of GL and ML were similar except that NH-bending vibrations and CeN stretching at 1516 cm−1 was higher in ML. C]C stretching of the aromatic rings in lignin was found only in PS, which can be attributed to the ligneous nature of the feedstock in comparison to GL and ML. Specific characteristics of GB functional groups were the presence of carbonates and PeO band, which might contribute to higher Pb removal capacity of GB.
they exhibited a smooth and compact surface while biochar samples displayed a porous or pitted surface textures. However, significant differences in morphologies of biochars produced from various plant source materials were not observed. The pyrolysis conditions at the higher temperature created more crevices or pores on the surface of the biochar products, thus resulting in their greater pore properties as shown in Supplementary Fig. 1(GB, PB, MB). BET surface area of plant source materials also increased after carbonization, however, it was much smaller than the surface area of activated carbon (AC) (Table 1). Average pore width of plant source materials was 2.11 mm, which reduced to 0.85 mm after carbonization. Reduction in pore size along with increasing porosity and internal pores was caused by the volatilization during carbonization (Ahmad et al., 2012). The highest BET surface area was found for PB. Table 1 shows the elemental compositions of the plant source materials and their biochars in comparison with that of AC. The C content of PB of 74.1% is the highest among the tested samples, which is similar to the C content of the AC. The postcarbonization C content of the PS was increased, whereas the C contents of the GL and the ML were decreased by the carbonization. Generally, the carbonization of feedstock increases biochar C content as a result of the progressive dehydration of the hydroxyl groups during the transformation of the organic components into polycyclic aromatic C (Beach et al., 2015). However, when the ash content of feedstock is mostly nonvolatile in nature and the loss of the noncarbon volatile components (hydrogen [H], oxygen [O], and N) are compensated for by the nonvolatile ash components in the biochars, the C contents can be decreased after the carbonization (Beach et al., 2015). The decreases of the H and O contents after carbonization can be attributed to the bond cleavage and the polymerization. Carbonization leads to a loss of the easily decomposable components, resulting in a decrease of elements such as O, H, N, and sulfur (S) (Ippolito et al., 2015); however, the N and S contents of the PS and the GL increased after carbonization. The loss of the other elements such as the O and H might cause the concentration of N and S. Al-Wabel et al. (2013) reported an increase of the N contents with the increasing of the pyrolysis temperature, which can be explained by the incorporation of N in thermally stable complex structures. The FTIR-analysis data of the functional groups in the plant source materials and their biochars are shown in Fig. 1. The peak of the plant source materials at 3400 cm−1 has been assigned to the hydroxide (OH) stretching because of the rich cellulose amounts with the hydroxide groups, and water may remain in the sample. The peak of the plant source materials at 2934 cm−1 is due to the aliphatic C–H stretching, indicating the existence of cellulose. The peaks at 2920 and 2850 cm−1 of GL and the ML are attributed to the C–H (Chen et al., 2010). The peak of the PS at 1660 cm−1 occurs because of the C]C stretching of the aromatic rings in lignin (Chia et al., 2012). The bands of the GL and the ML at 1318 and 1620 cm−1 have been assigned to the N–H stretching (Chen et al., 2010). The peak of the biochars at 1440 cm−1 has been assigned to the
3.2. Lead and Cu removal in relation to the properties of the plant source materials and their biochars The aqueous-solution Pb and Cu removal capacities of the GL, ML, and GB were significantly higher than those of the Pb and Cu removal capacities of the AC, as shown in Fig. 2. The GL-derived biochar showed the highest Pb and Cu removal capacities although the BET surface area of GB was lower than the surface area of PB, MB and AC (Table 1). Therefore, Pb and Cu removal was not only governed by surface adsorption, but by other factors such as precipitation and complexation. The Pb removal capacities of the adsorbents were higher than that of the Cu. The Pb electronegativity constant of 2.33 is higher than the 1.90 of Cu, resulting in a higher specific adsorption (Shi et al., 2009). The solution pH after the adsorption experiment remained constant, ranging from 4.94 to 5.14 in all of the experimental solutions, and this indicated that the solution pH did not affect the adsorption capacity (data are not shown). The phosphorus (P) concentration of the GB that was analyzed according to the acid digestion was 0.93%, which was significantly higher than the P concentration of the GL (0.34%) and the other plant source materials (PS: 0.028%, ML: 0.058%) and the biochars (PB: 0.12%, MB: 0.17%). The higher P concentration of the GB might lead to the formation of a Pb–phosphate precipitate, as was evidenced by the phosphate peaks in the GB in Fig. 1. Cao et al. (2009) also showed a poorly crystalline β-Pb9(PO4)6 precipitate after the reaction of the aqueous Pb with a P-rich biochar. The GB carbonate might also contribute to the higher adsorption of the Pb through the formation of a hydrocerussite Pb3(CO3)2(OH)2 precipitation. The intensity of the carbonate peak in the FTIR spectrum of the GB was significantly stronger than those of the other biochars. Xu et al. (2013) concluded that a biochar feedstock 120
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Fig. 1. FTIR spectra of the activated carbon (AC) and the plant source materials (GL: gingko leaf, PS: peanut shell, ML: Metasequoia leaf), and their biochars (GB: GLderived biochar, PB: PS-derived biochar, MB: ML-derived biochar).
containing CO32− and PO43− was effective in the removal of heavy metals such as Pb, Cu, zinc (Zn), and Cd from aqueous solutions. The binding of Pb and Cu to plant source materials can be achieved through the surface complexation of OH– and amidogen (NH2). The intensity of the functional groups of OH– and NH2 in the GL and the ML was higher than that of the PS, as can be seen in Fig. 1. Although the FTIR spectra of GL and ML were similar, NH-bending vibrations and CeN stretching intensity was significantly higher in GL. Chen et al. (2010) also demonstrated that the Cu adsorption of Cinnamomum camphora-leaf powder was due to a surface complexation with the major contributions of the hydroxyl and amino functional groups. The lowest Pb and Cu adsorption rates of the AC, PS, and PB can be attributed to the absence of these functional groups.
as shown in Supplementary Fig. 2. The Cu adsorption kinetics of the PS were similar to the PB kinetics that reached the equilibrium after approximately 720 min. The Cu adsorption by the ML reached the equilibrium after approximately 480 min, and it slowed down after the carbonization. The Cu adsorption by the MB reached the equilibrium after approximately 1440 min, as shown in Supplementary Fig. 2. The Cu adsorption by the AC, GL, and GB reached the equilibrium faster than the other adsorbents, and the Cu concentration was almost constant after approximately 360 min; however, adsorbed Cu concentration at the equilibrium was the highest with GB, which is also shown in Supplementary Fig. 2. The pseudo-first and pseudo-second order models were applied to simulate the acquired adsorption kinetics data. The equations can be expressed as follows:
3.3. Lead and Cu adsorption kinetics of the plant source materials and their biochars
qt = qe (1 − e−kt ) Pseudo − first order model
The Pb adsorption kinetics of the PS and the ML were similar and reached the adsorption equilibrium after approximately 300 min (Supplementary Fig. 2). The Pb adsorption by the PB and the MB slightly increased after 300 min up to 1440 min. The Pb adsorption by the GB quickly reached the equilibrium after approximately 40 min, which was faster than the equilibrium time (180 min) of the AC and GL,
qt =
k 2qe2 t 1 + k 2qe t
Pseudo − second order model
(1)
(2)
where qt (mg/g) and qe (mg/g) are the amounts of the heavy metal that were adsorbed at the time t and at the equilibrium, respectively, and k1 (1/min) and k2 (g/mg·min) are the pseudo-first order and pseudosecond order adsorption rate constants, respectively. Fig. 2. Comparison of the Pb and Cu removal capacities of various plant source materials and their biochars (AC: activated carbon, GL: gingko leaf, PS: peanut shell, ML: Metasequoia leaf, GB: GL-derived biochar, PB: PS-derived biochar, MB: ML-derived biochar). The error bars represent the standard deviation of the mean values. The different letters above the metal histogram indicate significant differences at p < 0.05 (ANOVA and Tukey's tests using R3.3).
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Table 2 Pseudo-first order and pseudo-second order kinetics parameters for the Pb and Cu adsorption capacities of the plant source materials and their biochars (AC: activated carbon, GL: gingko leaf, PS: peanut shell, ML: Metasequoia leaf, GB: GL-derived biochar, PB: PS-derived biochar, MB: ML-derived biochar). Metal
Pb
Cu
Adsorbent
AC GL PS ML GB PB MB AC GL PS ML GB PB MB
qe (exp.) (mg/g)
First-order- rate constants
8.81 23.88 11.41 24.34 50.62 11.05 13.21 6.55 11.17 3.67 14.56 16.23 4.70 3.94
Second-order-rate constants
k1 (1/min)
qe (theor.) (mg/g)
R2
k2 (g/mg/min)
qe (theor.) (mg/g)
R2
0.0006 0.0036 0.0009 0.0023 0.0016 0.0020 0.0011 0.0041 0.0032 0.0013 0.0043 0.0013 0.0011 0.0013
1.89 7.80 2.73 7.15 1.35 2.59 7.41 2.07 4.85 2.91 5.02 2.43 3.23 3.20
0.655 0.915 0.602 0.762 0.545 0.766 0.894 0.955 0.913 0.894 0.932 0.617 0.788 0.952
0.0047 0.0023 0.0032 0.0002 0.0117 0.0054 0.0009 0.0082 0.0029 0.0046 0.0003 0.0056 0.2145 0.0011
8.70 24.04 11.26 24.39 50.51 11.06 13.14 6.60 11.29 3.80 14.62 16.18 4.66 4.10
0.997 1.000 0.997 0.999 1.000 0.999 0.993 0.999 0.999 0.985 0.999 0.999 0.988 0.985
groups were effective for Cu retention. Uchimiya et al. (2011) also showed that biochar produced at lower pyrolysis temperature resulted in the lowest equilibrium Cu concentration because of remaining oxygen containing functional groups compared to the biochar produced at higher pyrolysis temperature.
Table 2 shows the kinetic characteristics of the Pb and Cu adsorption capacities of the plant source materials and their biochars. The Pb and Cu adsorption capacities of the plant source materials and their biochars were better fitted to the pseudo-second order model than the pseudo-first order model, indicating that the rate limiting step for the Pb and Cu adsorption is the chemical adsorption between the heavy metals and the adsorbents (Pellera et al., 2012). When the pseudosecond order reaction model was applied, the R2 value was greater than 0.985 and the theoretical adsorption capacities of the Pb and the Cu were close to the experimental adsorption capacity (Table 2). The calculated amount of the Pb that was adsorbed at the equilibrium was significantly increased by the GL carbonization. However, Pb adsorption by PS did not increase due to the carbonization, and the Pb adsorption of the ML decreased by the carbonization. The Cu adsorption capacities of the biochars that are derived from the PS and the GL slightly increased compared with the adsorption capacities of their plant source materials. The Cu adsorption capacity of the ML significantly lowered after carbonization, which can be explained by the disappearance of the hydroxyl and amino functional groups, as can be seen in Fig. 1. Pyrolysis results in loss of oxygen functionality from dehydration, decarbonylation, and decarboxylation reactions and production of aromatic structures as a result of condensation (Uchimiya et al., 2011). Increasing pyrolysis temperature increased aromaticity (Lee et al., 2017). The adsorption of heavy metals is controlled by surface functional groups of biochars and oxygen containing functional
3.4. Lead and Cu adsorption isotherms of the plant source materials and their biochars The adsorption data of the Pb and the Cu regarding the plant source materials and their biochars were analyzed using the Langmuir- and Freundlich-isotherm models. The Langmuir and Freundlich isotherms can be written as follows:
qe =
bQ0 Ce Langmuir isotherm 1 + bCe
(3)
qe = Kf Cen Freundlich isotherm
(4)
where Q0 (mg/g) is the maximum amount of the adsorbed heavy metals, Ce (mg/L) is the heavy metal concentration in the solution at the equilibrium, and b (L/mg) is the Langmuir-adsorption constant that is related to the affinity of the binding sites. Kf ((mg/g)(L/mg)1/n) is the Freundlich adsorption constant and n is the Freundlich linearity constant that are related to the adsorption intensity. Table 3 and Supplementary Fig. 3 show the adsorption isotherm
Table 3 Isothermal parameters for the Pb and Cu adsorption onto plant source materials and their biochars (AC: activated carbon, GL: gingko leaf, PS: peanut shell, ML: Metasequoia leaf, GB: GL-derived biochar, PB: PS-derived biochar, MB: ML-derived biochar). Heavy metal
Adsorbent
Langmuir isotherm parameters Q
Pb
Cu
AC GL PS ML GB PB MB AC GL PS ML GB PB MB
0
33.11 117.64 33.33 108.69 138.89 30.67 34.01 17.51 57.80 9.36 43.85 59.88 10.43 9.31
Freundlich isotherm parameters 2
b
R
0.0058 0.0140 0.0154 0.0140 0.1607 0.0119 0.0176 0.0169 0.0070 0.0430 0.0131 0.0117 0.0466 0.0358
0.947 0.925 0.955 0.923 0.998 0.932 0.934 0.996 0.941 0.956 0.960 0.958 0.982 0.995
122
KF
n
R2
1.673 5.246 4.602 3.936 51.475 1.727 3.655 1.239 1.226 1.417 1.811 5.197 1.727 2.831
1.673 1.891 3.273 1.672 5.393 2.142 2.731 2.325 1.672 2.862 1.817 2.642 2.142 5.170
0.988 0.988 0.899 0.995 0.828 0.880 0.924 0.973 0.985 0.497 0.909 0.978 0.758 0.883
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doi.org/10.1016/j.jenvman.2019.01.100.
results of the heavy metals on the adsorbents. The Pb adsorption by the plant source materials and their biochars was better explained by the Langmuir model with the R2 values ranging from 0.922 to 0.997, with the exception of the ML and the GL. The Pb adsorption of the PS, GB, PB, and MB can be interpreted as the chemical adsorption on the homogeneous monolayer (Ruthven, 1984), indicating that the adsorption site on the adsorbents is not infinite (Liu and Zhang, 2009). The Pb adsorption of the GL and the ML and the Cu adsorption by the GL and the GB were better explained by the Freundlich isotherm model, indicating that the Pb and Cu adsorption capacities of the GL and the GB involved not only a surface adsorption but also CO32− and PO43− precipitations (Xu et al., 2013). The highest maximum Pb adsorption capacities in the Langmuir model were acquired using the GB (138.9 mg/g) followed by the GL (117.6 mg/g), and ML (108.7 mg/g), which was much higher than that of the AC (33.1 mg/g). The highest maximum Cu adsorption capacities were achieved using the GB (59.9 mg/g), followed by the GL (57.8 mg/ g) and the ML (43.8 mg/g). The affinity of the GB to Pb, as expressed by the parameter b, was the highest. The maximum Cu adsorption capacities did not increase after the carbonization of the plant source materials, and Cu adsorption of the ML significantly decreased after carbonization. The different adsorption capacities before and after carbonization depending on different plant source materials might be related to the metal removal mechanisms in aqueous solution. Lead removal involved adsorption with surface precipitation and Cu removal was achieved by adsorption on oxygen containing functional groups. Carbonization reduced oxygen containing functional groups and lead to the reduced Pb and Cu adsorption after carbonization of ML while concentrated P by carbonization of GL enhanced Pb adsorption by GB. Maximum Pb adsorption by GB, GL, and ML was much higher than Pb adsorption reported by sesame straw biochar (102 mg/g) (Park et al., 2016) and sludge-derived biochar (30.88 mg/g) (Lu et al., 2012). Therefore, Gingko and Metasequoia leaves are promising adsorbents for Pb. However, appropriate method to prepare metal adsorbent should be used because the effect of carbonization is different depending on different plant source materials.
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Simultaneous determination of trace Cd(II), Pb(II) and Cu(II) by differential pulse anodic stripping voltammetry using a reduced graphene oxide-chitosan/poly-l-lysine nanocomposite modified glassy carbon electrode. J. Colloid Interface Sci. 490, 11–22. Han, R., Lu, Z., Zou, W., Daotong, W., Shi, J., Jiujun, Y., 2006. Removal of copper(II) and lead(II) from aqueous solution by manganese oxide coated sand: II. Equilibrium study and competitive adsorption. J. Hazard. Mater. 137, 480–488. Ho, Y., Ofomaja, A.E., 2006. Pseudo-second-order model for lead ion sorption from aqueous solutions onto palm kernel fiber. J. Hazard. Mater. 129, 137–142. Inyang, M.I., Gao, B., Yao, Y., Xue, Y., Zimmerman, A., Mosa, A., Pullammanappallil, P., Ok, Y.S., Cao, X., 2016. A review of biochar as a low-cost adsorbent for aqueous heavy metal removal. Crit. Rev. Environ. Sci. Technol. 46, 406–433. Ippolito, J.A., Spokas, K.A., Novak, J.M., Lentz, R.D., Cantrell, K.B., 2015. Biochar elemental composition and factors influencing nutrient retention. In: Lehmann, J., Joseph, S. (Eds.), Biochar for Environmental Management: Science. Technology and Implementation, Routledge New York, pp. 139–163. Komkiene, J., Baltrenaite, E., 2016. Biochar as adsorbent for removal of heavy metal ions [Cadmium (II), Copper (II), Lead (II), Zinc (II)] from aqueous phase. Int. J. Environ. Sci. Technol. 13, 471–482. Lee, S., Park, J.H., Ahn, Y., Chung, J.W., 2015. Comparison of heavy metal adsorption by peat moss and peat moss-derived biochar produced under different carbonization conditions. Water Air Soil Pollut. 226, 9. Lee, M., Park, J.H., Chung, J.W., 2017. Adsorption of Pb(II) and Cu(II) by ginkgo-leafderived biochar produced under various carbonization temperatures and times. Int. J. Environ. Res. Publ. Health 14, 1528. Liu, Z., Zhang, F.S., 2009. Removal of lead from water using biochars prepared from hydrothermal liquefaction of biomass. J. Hazard. Mater. 167, 933–939. Lu, H., Zhang, W., Yang, Y., Huang, X., Wang, S., Qiu, R., 2012. Relative distribution of Pb2+ sorption mechanisms by sludge-derived biochar. Water Res. 46, 854–862. McHenry, M.P., 2009. Agricultural bio-char production, renewable energy generation and farm carbon sequestration in Western Australia: certainty, uncertainty and risk. Agric. Ecosyst. Environ. 129, 1–7. Mohan, D., Pittman Jr., C.U., 2006. Activated carbons and low cost adsorbents for remediation of tri- and hexavalent chromium from water. J. Hazard. Mater. 137, 762–811. Park, J.H., Ok, Y.S., Kim, S.H., Cho, J.S., Heo, J.S., Delaune, R.D., Seo, D.C., 2016. Competitive adsorption of heavy metals onto sesame straw biochar in aqueous solutions. Chemosphere 142, 77–83. Pellera, F.M., Giannis, A., Kalderis, D., Anastasiadou, K., Stegmann, R., Wang, J.Y., Gidarakos, E., 2012. Adsorption of Cu (II) ions from aqueous solutions on biochars prepared from agricultural by-products. J. Environ. Manag. 96, 35–42. Puga, A.P., Abreu, C.A., Melo, L.C.A., Beesley, L., 2015. Biochar application to a contaminated soil reduces the availability and plant uptake of zinc, lead and cadmium. J. Environ. Manag. 159, 86–93. Ren, Y., Yan, N., Wen, Q., Fan, Z., Wei, T., Zhang, M., Ma, J., 2011. Graphene/δ-MnO2
4. Conclusions The comparison of Pb and Cu adsorption showed that ginkgo leafderived biochar have significantly higher Pb removal capacity than plant source materials. The increased surface area and pH of adsorbent might enhance metal removal by biochars. Furthermore, concentrated phosphate of gingko leaf-derived biochar after carbonization increased the Pb removal capacity through a surface precipitation of the Pb. However, Cu removal was not significantly affected by carbonization because it might be governed by oxygen containing functional groups. In addition, Pb and Cu adsorption reduced after carbonization of Metasequoia leaf, which might be attributed to the disappearance of the active functional groups for the metal adsorption. The contradictory result of metal adsorption capacity after carbonization was attributed to different metal removal mechanisms depending on plant source materials. Since pyrolysis did not always lead to enhanced metal removal for all metals, both economic aspects and metal removal effect should be considered and tested for target metals when using biochar as an adsorbent for wastewater treatment. Nevertheless, the maximum adsorption capacities of the gingko leaf-derived biochar for Pb and Cu were approximately 4.2 and 3.4 times higher than those of AC, respectively, indicating that it is very effective metal adsorbent and its adsorbability is economical for the purification of metal contaminated water. Appendix A. Supplementary data Supplementary data to this article can be found online at https:// 123
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soil. J. Agric. Food Chem. 59, 2501–2510. Wan, M.W., Kan, C.C., Rogel, B.D., Dalida, M.L.P., 2010. Adsorption of copper (II) and lead (II) ions from aqueous solution on chitosan-coated sand. Carbohyd. Polym. 80, 891–899. Wu, W., Yang, M., Feng, Q., McGrouther, K., Wang, H., Lu, H., Chen, Y., 2012. Chemical characterization of rice straw-derived biochar for soil amendment. Biomass Bioenergy 47, 268–276. Xu, X., Cao, X., Zhao, L., 2013. Comparison of rice husk-and dairy manure-derived biochars for simultaneously removing heavy metals from aqueous solutions: role of mineral components in biochars. Chemosphere 92, 955–961. Yuan, J.H., Xu, R.K., Zhang, H., 2011. The forms of alkalis in the biochar produced from crop residues at different temperatures. Bioresour. Technol. 102, 3488–3497. Zhou, Z., Xu, Z., Feng, Q., Yao, D., Yu, J., Wang, D., Lv, S., Liu, Y., Zhou, N., Zhong, M.E., 2018. Effect of pyrolysis condition on the adsorption mechanism of lead, cadmium and copper on tobacco stem biochar. J. Clean. Prod. 187, 996–1005.
composite as adsorbent for the removal of nickel ions from wastewater. Chem. Eng. J. 175, 1–7. Ren, Y., Li, N., Feng, J., Luan, T., Wen, Q., Li, Z., Zhang, M., 2012. Adsorption of Pb(II) and Cu(II) from aqueous solution on magnetic porous ferrospinel MnFe2O4. J. Colloid Interface Sci. 367, 415–421. Ruthven, D.M., 1984. Principles of Adsorption and Adsorption Processes. Wiley Interscience, New York, NY. Shi, T., Jia, S., Chen, Y., Wen, Y., Du, C., Guo, H., Wang, Z., 2009. Adsorption of Pb (II), Cr (III), Cu (II), Cd (II) and Ni (II) onto a vanadium mine tailing from aqueous solution. J. Hazard. Mater. 169, 838–846. Uchimiya, M., Lima, I.M., Thomas Klasson, K., Chang, S., Wartelle, L.H., Rodgers, J.E., 2010. Immobilization of heavy metal ions (CuII, CdII, NiII, and PbII) by broiler litterderived biochars in water and soil. J. Agric. Food Chem. 58, 5538–5544. Uchimiya, M., Wartelle, L.H., Klasson, K.T., Fortier, C.A., Lima, I.M., 2011. Influence of pyrolysis temperature on biochar property and function as a heavy metal sorbent in
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Research article
Comparison of the lead and copper adsorption capacities of plant source materials and their biochars
T
Myoung-Eun Leea,1, Jin Hee Parkb,1, Jae Woo Chunga,c,∗ a
Department of Urban System Engineering, Gyeongnam National University of Science and Technology (GNTECH), Dongjin-ro 33, Jinju, Gyeongnam 52725, South Korea Department of Environmental and Biological Chemistry, Chungbuk National University, Cheongju, 28644, South Korea c Department of Environmental Engineering, GNTECH, Dongjin-ro 33, Jinju, Gyeongnam, 52725, South Korea b
A R T I C LE I N FO
A B S T R A C T
Keywords: Biochar Gingko leaf Peanut shell Metasequoia leaf Lead Copper
Lead (Pb) and Cu are the most common pollutants found in industrial effluents which affect ecosystem and human health. To remove Pb and Cu from aquatic system, cost-effective and environmentally friendly adsorbents are required. Therefore, the study evaluated the adsorption of Pb and Cu by waste plant materials and their biochars. The adsorption kinetics and isotherms were applied to compare the Pb and Cu adsorption capacities using the gingko (Spiraea blumei) leaf (GL), peanut shell (PS), and Metasequoia leaf (ML), and their derived biochars (GB, PB, and MB, respectively). The GB showed a significantly higher Pb adsorption capacity than the other adsorbents. Maximum Pb adsorption by GB was 138.9 mg/g followed by GL (117.6 mg/g). The highest Cu adsorption (59.9 mg/g) was also achieved by GB followed by GL (57.8 mg/g). The carbonates and the phosphate functional groups in the GB and higher affinity of Pb to the functional groups contributed to higher Pb adsorption. The Pb adsorption kinetics on the plant source materials and their biochars followed a pseudo-second order model. The Pb and Cu adsorption capacities, with the exception of the GL, ML, and GB, are better explained by Langmuir-isotherm models. The carbonization did not always lead to better heavy metal adsorption. The Pb and Cu adsorption significantly reduced with carbonization of ML because of disappearance of oxygen containing functional groups. Therefore, appropriate method to prepare metal adsorbent should be selected depending on feedstocks and metal removal mechanisms. The GL is the most-abundant fallen leaf in the streets of the Republic of Korea; therefore, the use of the GL biochar for heavy-metal adsorption will also reduce the cost for waste disposal.
1. Introduction The presence of heavy metals in the aquatic environment is a major concern due to their extreme toxicity. Lead(II) and Cu(II), in particular, are among the most common pollutants found in industrial effluents (Ren et al., 2012). According to the US Environmental Pollution Agency, they are highly toxic and can cause a variety of negative effects on human health even at low dosages, for example, anemia, encephalopathy, hepatitis, and the nephritic syndrome (Wan et al., 2010). Therefore, the removal of Pb(II) and Cu(II) ions from aqueous solution is an important step in water purification process. Conventional technologies for the removal of heavy metals from wastewater include adsorption, chemical precipitation, redox process, ion exchange, solvent extraction, electrolyzation, filtration, electrochemical removal, membrane separation and reverse osmosis (Ren
et al., 2011; Guo et al., 2017; Bilal et al., 2018; Fakhre and Ibrahim, 2018). Among these methods, adsorption has attracted much attention as an effective purification and separation technique for treating wastewater. Removal of heavy metals from wastewater can be achieved by application of adsorption processes using a suitable adsorbent (Han et al., 2006; Mohan and Pittman Jr, 2006). Suitable adsorbent can be characterized by low cost, high renewability and high removal capacity (Chen et al., 2018). Recently, considerable attention has been paid to the investigation of different types of low-cost adsorbents especially using biochar. It has been reported that the biochar-contained functional groups and alkaline pH conditions are responsible for the removal of the heavy metal ions from water (Uchimiya et al., 2010). Removal of Pb and Cu using various biochars derived from different feedstocks was tested in other studies. Silver birch biochar and Scot pine biochar adsorbed
∗
Corresponding author. Department of Urban System Engineering, Gyeongnam National University of Science and Technology (GNTECH), Dongjin-ro 33, Jinju, Gyeongnam 52725, South Korea. E-mail address: [email protected] (J.W. Chung). 1 Myoung-Eun Lee and Jin Hee Park contributed equally to the manuscript. https://doi.org/10.1016/j.jenvman.2019.01.100 Received 23 August 2018; Received in revised form 4 January 2019; Accepted 26 January 2019 0301-4797/ © 2019 Elsevier Ltd. All rights reserved.
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pH, electrical conductivity (EC), surface morphology, and functional groups to identify their major physicochemical properties. For the measurement of the pH and the conductivity, 5 g of each sample and 25 mL of distilled water were mixed for 1 h in a 50 mL conical tube. The pH and EC were measured using the pH and EC electrodes after calibration. The surface morphology was analyzed using field emission scanning electron microscopy (FE-SEM, JEOL, Japan). The samples were degassed at 250 °C for 10 h and Brunauer-Emmett-Teller (BET) surface area was evaluated by N2 adsorption using a gas sorption analyzer (Autosorb-iQ2, USA). Fourier transform infrared spectroscopy (FT-IR, Perkin Elmer, USA) was used for an analysis of the functional groups on the biochar surfaces. FT-IR spectra of KBr pellets containing samples were recorded in the 4000–400 cm−1 region with a resolution of 4.0 cm−1. An elemental analysis for C, H, N, S and O was performed using the Flash 2000 Series elemental analyzer (Thermo Fisher Scientific, USA). The ash contents were measured from the combustion of the dry samples at 760 °C for 6 h. The phosphorus concentration of the samples was analyzed using the Optima 8300DV inductively coupled plasma-optical emission (ICP-OES, Perkin Elmer, USA) after a nitric acid digestion of the samples to determine the contribution of phosphorus to the removal of metals.
128.7 mg/kg and 107.0 mg/kg of Cu, respectively, although Pb adsorption by those biochars ranged from 1.29 to 4.49 mg/kg (Komkiene and Baltrenaite, 2016). Tobacco stem biochar was also effective in the adsorption of Pb, Cd and Cu (Zhou et al., 2018). Metal adsorption capacity of biochar can be different depending on the feedstock type and biochar production conditions such as pyrolysis temperature and time (Ahmad et al., 2014). Inyang et al. (2016) demonstrated that the feedstock is an important parameter that determines the adsorption capacity of biochars, and that heavy metal removal mechanisms vary depending on the biochar type. Feedstocks that have been used for biochar production include broiler litter, poultry manure, plant materials including straw and leaves, food wastes such as orange peel and peanut shell, industrial byproducts including paper sludge and sewage sludge, and agricultural wastes such as rice husk and wheat straw (Ahmad et al., 2014). Biochars are normally produced from waste materials because of the low cost. Waste-materialbased biochar production can not only reduce the waste amount, but also replace fossil fuels in the production of energy (McHenry, 2009). Therefore, the study focused on the evaluation of waste materials to select feedstocks for biochar production as a cost effective metal adsorbent and at the same time waste disposal and reuse were considered. Although biochar is known to be effective in heavy metal adsorption, the adsorption capacity may not be always higher than its feedstocks. Mechanisms of metal adsorption are different depending on feedstocks and carbonization of feedstocks may differently affect metal adsorption capacity. Since biochar production requires energy and time, feedstock as a raw material would be better to be used as a metal adsorbent if the feedstock has higher metal adsorption capacity than its biochar. However, few studies have been conducted on the comparison of metal adsorption capacity between feedstock and its biochar. Therefore, heavy metal adsorption capacity of feedstock and its biochar was compared in this study and interpreted according to adsorption mechanisms. Plant source materials that can be easily obtained in natural environments in Korea were selected as the raw materials for this study. Gingko and Metasequoia leaves were first used as feedstocks for biochar in this study. Although peanut shell biochar was previously tested for Pb and Zn immobilization in soil by Chao et al. (2018), it was included in this study for comparison with other feedstocks. Therefore, the objective of the present study is the comparison of the Pb and Cu adsorption characteristics using different feedstocks and their biochars for the possible application of plant source materials in the treatment of metal-contaminated water.
2.3. Lead and Cu adsorption of the plant source materials and their biochars The produced biochars were finely ground with a mortar and pestle and passed through a sieve (< 0.5 mm) to prepare them for the adsorption experiments. To investigate the Pb and Cu adsorption characteristics of the plant source materials and their biochars, a heavy metal solution (50 mg/L) was reacted with the adsorbent for 48 h on the basis of 1 g/L in a rotary shaker at 200 rpm at 25 °C. The Pb and Cu stock solutions were prepared with Pb(NO3)2 and Cu(NO3)2, respectively. To prevent the effect of the pH on the adsorption, the heavy metal solutions were prepared using 0.07 M sodium acetate and 0.03 M acetic acid in pH 5 buffer solution (Chen and Wu, 2004). To investigate the kinetic heavy metal adsorption characteristics of the plant source materials and biochars, 1 g/L of each of the plant source materials and their biochars were reacted with 50 mg/L of the pH 5 Pb and Cu solutions at 25 °C. The Pb and Cu concentrations were analyzed at 0, 5, 10, 20, 30, 40, 60, 90, 120, 180, 360, 540, 1440, and 2880 min during the adsorption experiment. The pseudo-first order model and the pseudo-second order model were used to characterize the adsorption kinetics (Ho and Ofomaja, 2006). To observe the isothermal adsorption characteristics of the heavy metal ions, 1 g/L of each of the plant source materials and their biochars were reacted with the heavy metal ion solution of 30, 50, 100, 150, 200, 300, 400, and 500 mg/L for 48 h and the equilibrium was established. After the adsorption, the samples were filtered through a 0.2 μm filter (Advantec, Japan) and diluted with a 1% nitric acid solution. The metal concentrations in the solution were analyzed using ICP-OES. To evaluate and compare the Pb and Cu adsorption capacities of the plant source materials and their biochars, the Langmuir model and the Freundlich model were applied to fit the experiment data (Liu and Zhang, 2009; Pellera et al., 2012).
2. Materials and methods 2.1. Production of biochars from various plant source materials Gingko (Spiraea blumei) leaf (GL), peanut (Arachis hypogaea Linn.) shell (PS), and Metasequoia (Metasequoia glyptostroboides) leaf (ML) were selected as the plant source materials for the Pb and Cu adsorption. The plant source materials can be easily obtained during the autumn of the Republic of Korea. The collected plant source materials were washed with deionized water and dried in a drying oven at a temperature of 105 °C for 24 h. To produce the biochar, 5–10 g of the dried feedstock was placed in a crucible and purged with nitrogen (N2) for 5 min, followed by a 90 min electric-furnace carbonization at 800 °C. In the authors' previous study, the optimal carbonization temperature and time for the heavy metal adsorption of the peat moss and Ginkgo leaf-derived biochar were 800 °C and 90 min, respectively (Lee et al., 2015; Lee et al., 2017); therefore, the temperature and time for producing biochar was set to 800 °C and 90 min.
3. Results and discussion 3.1. Characterization of the plant source materials and their biochars The carbonization of the plant source materials developed a porous structure, as evidenced by the scanning electron microscopy (SEM) data of Supplementary Fig. 1. A porous structure was formed from the tubular plant cell structure (Puga et al., 2015). The hydrolysis reactions of the plant source materials and the release of the degradation products led to morphological changes (Liu and Zhang, 2009). The SEM morphological textures of the feedstock and the biochar product on the surface showed a great difference. As observed in feedstock samples,
2.2. Characterization of the plant source materials and their biochars The plant source materials and their biochars were tested for the 119
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Table 1 Brunauer-Emmet-Teller (BET) surface area and elemental compositions of the plant source materials and their biochars (AC: activated carbon, GL: gingko leaf, PS: peanut shell, ML: Metasequoia leaf, GB: GL-derived biochar, PB: PS-derived biochar, MB: ML-derived biochar). BET surface area (m2/g)
Samples
Activated carbon Plant source materials
Biochar
AC GL PS ML GB PB MB
1069 2 16 17 310 790 496
Elemental composition (%)
pH
C
H
N
S
O
Ash
P
76.07 45.22 49.50 50.33 42.48 74.06 40.61
2.04 5.56 5.82 5.92 0.47 1.29 0.58
0.69 1.21 0.60 1.14 1.65 1.38 0.71
3.09 0.73 0.14 0.24 1.85 2.73 0.26
12.72 33.99 42.32 37.63 17.96 13.80 38.04
5.40 13.40 1.63 4.75 35.60 6.74 19.81
0.006 0.338 0.028 0.058 0.934 0.117 0.173
7.69 4.60 5.27 4.19 12.43 12.88 10.85
aromatic C]C, which is commonly found in biochars and is the result of cellulose-derived transformations (Wu et al., 2012). The peaks of the GB at 876 and 1450 cm−1 have been assigned to the bands of the outof-plane bending for the carbonates (Yuan et al., 2011; Xu et al., 2013). In the region of 600–1500 cm−1 regarding the plant source materials, various peaks are evident including C]C, C–H, C–O, C–C, and OH. The broad peak within the 1000–1250-cm−1 region can be attributed to the inorganic components such as the sulfates and the silicates. The bands at 1115 and 1059 cm−1 can be attributed to the phosphate P–O bond (Cantrell et al., 2012). Lee et al. (2017) reported that the functional groups such as the H-bonded OH groups, aliphatic CeH, CeH2, CeH3 stretching, and aromatic C]C was produced as the carbonization temperature was increased. Functional groups of GL and ML were similar except that NH-bending vibrations and CeN stretching at 1516 cm−1 was higher in ML. C]C stretching of the aromatic rings in lignin was found only in PS, which can be attributed to the ligneous nature of the feedstock in comparison to GL and ML. Specific characteristics of GB functional groups were the presence of carbonates and PeO band, which might contribute to higher Pb removal capacity of GB.
they exhibited a smooth and compact surface while biochar samples displayed a porous or pitted surface textures. However, significant differences in morphologies of biochars produced from various plant source materials were not observed. The pyrolysis conditions at the higher temperature created more crevices or pores on the surface of the biochar products, thus resulting in their greater pore properties as shown in Supplementary Fig. 1(GB, PB, MB). BET surface area of plant source materials also increased after carbonization, however, it was much smaller than the surface area of activated carbon (AC) (Table 1). Average pore width of plant source materials was 2.11 mm, which reduced to 0.85 mm after carbonization. Reduction in pore size along with increasing porosity and internal pores was caused by the volatilization during carbonization (Ahmad et al., 2012). The highest BET surface area was found for PB. Table 1 shows the elemental compositions of the plant source materials and their biochars in comparison with that of AC. The C content of PB of 74.1% is the highest among the tested samples, which is similar to the C content of the AC. The postcarbonization C content of the PS was increased, whereas the C contents of the GL and the ML were decreased by the carbonization. Generally, the carbonization of feedstock increases biochar C content as a result of the progressive dehydration of the hydroxyl groups during the transformation of the organic components into polycyclic aromatic C (Beach et al., 2015). However, when the ash content of feedstock is mostly nonvolatile in nature and the loss of the noncarbon volatile components (hydrogen [H], oxygen [O], and N) are compensated for by the nonvolatile ash components in the biochars, the C contents can be decreased after the carbonization (Beach et al., 2015). The decreases of the H and O contents after carbonization can be attributed to the bond cleavage and the polymerization. Carbonization leads to a loss of the easily decomposable components, resulting in a decrease of elements such as O, H, N, and sulfur (S) (Ippolito et al., 2015); however, the N and S contents of the PS and the GL increased after carbonization. The loss of the other elements such as the O and H might cause the concentration of N and S. Al-Wabel et al. (2013) reported an increase of the N contents with the increasing of the pyrolysis temperature, which can be explained by the incorporation of N in thermally stable complex structures. The FTIR-analysis data of the functional groups in the plant source materials and their biochars are shown in Fig. 1. The peak of the plant source materials at 3400 cm−1 has been assigned to the hydroxide (OH) stretching because of the rich cellulose amounts with the hydroxide groups, and water may remain in the sample. The peak of the plant source materials at 2934 cm−1 is due to the aliphatic C–H stretching, indicating the existence of cellulose. The peaks at 2920 and 2850 cm−1 of GL and the ML are attributed to the C–H (Chen et al., 2010). The peak of the PS at 1660 cm−1 occurs because of the C]C stretching of the aromatic rings in lignin (Chia et al., 2012). The bands of the GL and the ML at 1318 and 1620 cm−1 have been assigned to the N–H stretching (Chen et al., 2010). The peak of the biochars at 1440 cm−1 has been assigned to the
3.2. Lead and Cu removal in relation to the properties of the plant source materials and their biochars The aqueous-solution Pb and Cu removal capacities of the GL, ML, and GB were significantly higher than those of the Pb and Cu removal capacities of the AC, as shown in Fig. 2. The GL-derived biochar showed the highest Pb and Cu removal capacities although the BET surface area of GB was lower than the surface area of PB, MB and AC (Table 1). Therefore, Pb and Cu removal was not only governed by surface adsorption, but by other factors such as precipitation and complexation. The Pb removal capacities of the adsorbents were higher than that of the Cu. The Pb electronegativity constant of 2.33 is higher than the 1.90 of Cu, resulting in a higher specific adsorption (Shi et al., 2009). The solution pH after the adsorption experiment remained constant, ranging from 4.94 to 5.14 in all of the experimental solutions, and this indicated that the solution pH did not affect the adsorption capacity (data are not shown). The phosphorus (P) concentration of the GB that was analyzed according to the acid digestion was 0.93%, which was significantly higher than the P concentration of the GL (0.34%) and the other plant source materials (PS: 0.028%, ML: 0.058%) and the biochars (PB: 0.12%, MB: 0.17%). The higher P concentration of the GB might lead to the formation of a Pb–phosphate precipitate, as was evidenced by the phosphate peaks in the GB in Fig. 1. Cao et al. (2009) also showed a poorly crystalline β-Pb9(PO4)6 precipitate after the reaction of the aqueous Pb with a P-rich biochar. The GB carbonate might also contribute to the higher adsorption of the Pb through the formation of a hydrocerussite Pb3(CO3)2(OH)2 precipitation. The intensity of the carbonate peak in the FTIR spectrum of the GB was significantly stronger than those of the other biochars. Xu et al. (2013) concluded that a biochar feedstock 120
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Fig. 1. FTIR spectra of the activated carbon (AC) and the plant source materials (GL: gingko leaf, PS: peanut shell, ML: Metasequoia leaf), and their biochars (GB: GLderived biochar, PB: PS-derived biochar, MB: ML-derived biochar).
containing CO32− and PO43− was effective in the removal of heavy metals such as Pb, Cu, zinc (Zn), and Cd from aqueous solutions. The binding of Pb and Cu to plant source materials can be achieved through the surface complexation of OH– and amidogen (NH2). The intensity of the functional groups of OH– and NH2 in the GL and the ML was higher than that of the PS, as can be seen in Fig. 1. Although the FTIR spectra of GL and ML were similar, NH-bending vibrations and CeN stretching intensity was significantly higher in GL. Chen et al. (2010) also demonstrated that the Cu adsorption of Cinnamomum camphora-leaf powder was due to a surface complexation with the major contributions of the hydroxyl and amino functional groups. The lowest Pb and Cu adsorption rates of the AC, PS, and PB can be attributed to the absence of these functional groups.
as shown in Supplementary Fig. 2. The Cu adsorption kinetics of the PS were similar to the PB kinetics that reached the equilibrium after approximately 720 min. The Cu adsorption by the ML reached the equilibrium after approximately 480 min, and it slowed down after the carbonization. The Cu adsorption by the MB reached the equilibrium after approximately 1440 min, as shown in Supplementary Fig. 2. The Cu adsorption by the AC, GL, and GB reached the equilibrium faster than the other adsorbents, and the Cu concentration was almost constant after approximately 360 min; however, adsorbed Cu concentration at the equilibrium was the highest with GB, which is also shown in Supplementary Fig. 2. The pseudo-first and pseudo-second order models were applied to simulate the acquired adsorption kinetics data. The equations can be expressed as follows:
3.3. Lead and Cu adsorption kinetics of the plant source materials and their biochars
qt = qe (1 − e−kt ) Pseudo − first order model
The Pb adsorption kinetics of the PS and the ML were similar and reached the adsorption equilibrium after approximately 300 min (Supplementary Fig. 2). The Pb adsorption by the PB and the MB slightly increased after 300 min up to 1440 min. The Pb adsorption by the GB quickly reached the equilibrium after approximately 40 min, which was faster than the equilibrium time (180 min) of the AC and GL,
qt =
k 2qe2 t 1 + k 2qe t
Pseudo − second order model
(1)
(2)
where qt (mg/g) and qe (mg/g) are the amounts of the heavy metal that were adsorbed at the time t and at the equilibrium, respectively, and k1 (1/min) and k2 (g/mg·min) are the pseudo-first order and pseudosecond order adsorption rate constants, respectively. Fig. 2. Comparison of the Pb and Cu removal capacities of various plant source materials and their biochars (AC: activated carbon, GL: gingko leaf, PS: peanut shell, ML: Metasequoia leaf, GB: GL-derived biochar, PB: PS-derived biochar, MB: ML-derived biochar). The error bars represent the standard deviation of the mean values. The different letters above the metal histogram indicate significant differences at p < 0.05 (ANOVA and Tukey's tests using R3.3).
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Table 2 Pseudo-first order and pseudo-second order kinetics parameters for the Pb and Cu adsorption capacities of the plant source materials and their biochars (AC: activated carbon, GL: gingko leaf, PS: peanut shell, ML: Metasequoia leaf, GB: GL-derived biochar, PB: PS-derived biochar, MB: ML-derived biochar). Metal
Pb
Cu
Adsorbent
AC GL PS ML GB PB MB AC GL PS ML GB PB MB
qe (exp.) (mg/g)
First-order- rate constants
8.81 23.88 11.41 24.34 50.62 11.05 13.21 6.55 11.17 3.67 14.56 16.23 4.70 3.94
Second-order-rate constants
k1 (1/min)
qe (theor.) (mg/g)
R2
k2 (g/mg/min)
qe (theor.) (mg/g)
R2
0.0006 0.0036 0.0009 0.0023 0.0016 0.0020 0.0011 0.0041 0.0032 0.0013 0.0043 0.0013 0.0011 0.0013
1.89 7.80 2.73 7.15 1.35 2.59 7.41 2.07 4.85 2.91 5.02 2.43 3.23 3.20
0.655 0.915 0.602 0.762 0.545 0.766 0.894 0.955 0.913 0.894 0.932 0.617 0.788 0.952
0.0047 0.0023 0.0032 0.0002 0.0117 0.0054 0.0009 0.0082 0.0029 0.0046 0.0003 0.0056 0.2145 0.0011
8.70 24.04 11.26 24.39 50.51 11.06 13.14 6.60 11.29 3.80 14.62 16.18 4.66 4.10
0.997 1.000 0.997 0.999 1.000 0.999 0.993 0.999 0.999 0.985 0.999 0.999 0.988 0.985
groups were effective for Cu retention. Uchimiya et al. (2011) also showed that biochar produced at lower pyrolysis temperature resulted in the lowest equilibrium Cu concentration because of remaining oxygen containing functional groups compared to the biochar produced at higher pyrolysis temperature.
Table 2 shows the kinetic characteristics of the Pb and Cu adsorption capacities of the plant source materials and their biochars. The Pb and Cu adsorption capacities of the plant source materials and their biochars were better fitted to the pseudo-second order model than the pseudo-first order model, indicating that the rate limiting step for the Pb and Cu adsorption is the chemical adsorption between the heavy metals and the adsorbents (Pellera et al., 2012). When the pseudosecond order reaction model was applied, the R2 value was greater than 0.985 and the theoretical adsorption capacities of the Pb and the Cu were close to the experimental adsorption capacity (Table 2). The calculated amount of the Pb that was adsorbed at the equilibrium was significantly increased by the GL carbonization. However, Pb adsorption by PS did not increase due to the carbonization, and the Pb adsorption of the ML decreased by the carbonization. The Cu adsorption capacities of the biochars that are derived from the PS and the GL slightly increased compared with the adsorption capacities of their plant source materials. The Cu adsorption capacity of the ML significantly lowered after carbonization, which can be explained by the disappearance of the hydroxyl and amino functional groups, as can be seen in Fig. 1. Pyrolysis results in loss of oxygen functionality from dehydration, decarbonylation, and decarboxylation reactions and production of aromatic structures as a result of condensation (Uchimiya et al., 2011). Increasing pyrolysis temperature increased aromaticity (Lee et al., 2017). The adsorption of heavy metals is controlled by surface functional groups of biochars and oxygen containing functional
3.4. Lead and Cu adsorption isotherms of the plant source materials and their biochars The adsorption data of the Pb and the Cu regarding the plant source materials and their biochars were analyzed using the Langmuir- and Freundlich-isotherm models. The Langmuir and Freundlich isotherms can be written as follows:
qe =
bQ0 Ce Langmuir isotherm 1 + bCe
(3)
qe = Kf Cen Freundlich isotherm
(4)
where Q0 (mg/g) is the maximum amount of the adsorbed heavy metals, Ce (mg/L) is the heavy metal concentration in the solution at the equilibrium, and b (L/mg) is the Langmuir-adsorption constant that is related to the affinity of the binding sites. Kf ((mg/g)(L/mg)1/n) is the Freundlich adsorption constant and n is the Freundlich linearity constant that are related to the adsorption intensity. Table 3 and Supplementary Fig. 3 show the adsorption isotherm
Table 3 Isothermal parameters for the Pb and Cu adsorption onto plant source materials and their biochars (AC: activated carbon, GL: gingko leaf, PS: peanut shell, ML: Metasequoia leaf, GB: GL-derived biochar, PB: PS-derived biochar, MB: ML-derived biochar). Heavy metal
Adsorbent
Langmuir isotherm parameters Q
Pb
Cu
AC GL PS ML GB PB MB AC GL PS ML GB PB MB
0
33.11 117.64 33.33 108.69 138.89 30.67 34.01 17.51 57.80 9.36 43.85 59.88 10.43 9.31
Freundlich isotherm parameters 2
b
R
0.0058 0.0140 0.0154 0.0140 0.1607 0.0119 0.0176 0.0169 0.0070 0.0430 0.0131 0.0117 0.0466 0.0358
0.947 0.925 0.955 0.923 0.998 0.932 0.934 0.996 0.941 0.956 0.960 0.958 0.982 0.995
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KF
n
R2
1.673 5.246 4.602 3.936 51.475 1.727 3.655 1.239 1.226 1.417 1.811 5.197 1.727 2.831
1.673 1.891 3.273 1.672 5.393 2.142 2.731 2.325 1.672 2.862 1.817 2.642 2.142 5.170
0.988 0.988 0.899 0.995 0.828 0.880 0.924 0.973 0.985 0.497 0.909 0.978 0.758 0.883
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doi.org/10.1016/j.jenvman.2019.01.100.
results of the heavy metals on the adsorbents. The Pb adsorption by the plant source materials and their biochars was better explained by the Langmuir model with the R2 values ranging from 0.922 to 0.997, with the exception of the ML and the GL. The Pb adsorption of the PS, GB, PB, and MB can be interpreted as the chemical adsorption on the homogeneous monolayer (Ruthven, 1984), indicating that the adsorption site on the adsorbents is not infinite (Liu and Zhang, 2009). The Pb adsorption of the GL and the ML and the Cu adsorption by the GL and the GB were better explained by the Freundlich isotherm model, indicating that the Pb and Cu adsorption capacities of the GL and the GB involved not only a surface adsorption but also CO32− and PO43− precipitations (Xu et al., 2013). The highest maximum Pb adsorption capacities in the Langmuir model were acquired using the GB (138.9 mg/g) followed by the GL (117.6 mg/g), and ML (108.7 mg/g), which was much higher than that of the AC (33.1 mg/g). The highest maximum Cu adsorption capacities were achieved using the GB (59.9 mg/g), followed by the GL (57.8 mg/ g) and the ML (43.8 mg/g). The affinity of the GB to Pb, as expressed by the parameter b, was the highest. The maximum Cu adsorption capacities did not increase after the carbonization of the plant source materials, and Cu adsorption of the ML significantly decreased after carbonization. The different adsorption capacities before and after carbonization depending on different plant source materials might be related to the metal removal mechanisms in aqueous solution. Lead removal involved adsorption with surface precipitation and Cu removal was achieved by adsorption on oxygen containing functional groups. Carbonization reduced oxygen containing functional groups and lead to the reduced Pb and Cu adsorption after carbonization of ML while concentrated P by carbonization of GL enhanced Pb adsorption by GB. Maximum Pb adsorption by GB, GL, and ML was much higher than Pb adsorption reported by sesame straw biochar (102 mg/g) (Park et al., 2016) and sludge-derived biochar (30.88 mg/g) (Lu et al., 2012). Therefore, Gingko and Metasequoia leaves are promising adsorbents for Pb. However, appropriate method to prepare metal adsorbent should be used because the effect of carbonization is different depending on different plant source materials.
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4. Conclusions The comparison of Pb and Cu adsorption showed that ginkgo leafderived biochar have significantly higher Pb removal capacity than plant source materials. The increased surface area and pH of adsorbent might enhance metal removal by biochars. Furthermore, concentrated phosphate of gingko leaf-derived biochar after carbonization increased the Pb removal capacity through a surface precipitation of the Pb. However, Cu removal was not significantly affected by carbonization because it might be governed by oxygen containing functional groups. In addition, Pb and Cu adsorption reduced after carbonization of Metasequoia leaf, which might be attributed to the disappearance of the active functional groups for the metal adsorption. The contradictory result of metal adsorption capacity after carbonization was attributed to different metal removal mechanisms depending on plant source materials. Since pyrolysis did not always lead to enhanced metal removal for all metals, both economic aspects and metal removal effect should be considered and tested for target metals when using biochar as an adsorbent for wastewater treatment. Nevertheless, the maximum adsorption capacities of the gingko leaf-derived biochar for Pb and Cu were approximately 4.2 and 3.4 times higher than those of AC, respectively, indicating that it is very effective metal adsorbent and its adsorbability is economical for the purification of metal contaminated water. Appendix A. Supplementary data Supplementary data to this article can be found online at https:// 123
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