Abstract

Biochar, as a cost-efficient adsorbent, is of major interest in the removal of heavy metals from wastewater. Herein, batch experiments were conducted to investigate the performance of biochar derived from rice straw for the removal of Ni(II) as a function of various environmental conditions. The results showed that Ni(II) sorption was strongly dependent on pH but independent of ionic strength and the effects of electrolyte ions could be negligible over the whole pH range. Ionic exchange and inner-sphere surface complexation dominated the sorption of Ni(II). Humic/fulvic acids clearly enhanced the Ni(II) sorption at pH <7.2 but inhibited the sorption at pH >7.2. The sorption reached equilibrium within 10 hours, and the kinetics followed a pseudo-second-order rate model. Any of the Langmuir, Freundlich, or Dubinin-Radushkevich isotherm models could describe the sorption well, but the Langmuir model described it best. The maximum sorption capacity calculated from the Langmuir model was 0.257 m·mol/g. The thermodynamic parameters suggested that Ni(II) sorption was a spontaneous and endothermic process and was enhanced at high temperature. The results of this work indicate that biochar derived from rice straw may be a valuable bio-sorbent for Ni(II) in aqueous solutions, but it still requires further modification.

INTRODUCTION

Currently, increasing amounts of industrial and agricultural effluents are discharged into the environment, resulting in more inorganic and organic pollution of soil and water. Most organic pollutants are biodegradable, but inorganic pollutants, especial heavy metals, are non-biodegradable and can be accumulated through the human and animal food chains. This accumulation of heavy metals is a serious threat to human health, aquatic organisms and ecological safety over a long period (Dong et al. 2014). Nickel (Ni) mainly derives from steel factories, batteries manufacturing, and metal and alloy production (Gautam et al. 2015). It is well known that as a micronutrient, Ni is part of the synthesis of vitamin B12, and an activator in some enzyme systems. However, high dosages can cause serious diseases such as hepatitis, nephritic syndrome, anaemia, and diarrhoea (Kasprzak et al. 2003). It is therefore necessary to reduce ionic Ni to permissible limits before its discharge into the natural environment.

The United States Environmental Protection Agency (EPA) provides current maximum allowable contaminant levels for many heavy metals. To meet the EPA standards, various removal techniques such as ion exchange, precipitation, sorption, and electrocoagulation are used in sewage treatment (Inyang et al. 2015). Among these techniques, the sorption approach has been extensively adopted due to its ease of operation, low cost, and high efficiency. However, different adsorbents exhibit different sorption abilities. For instance, the sorption behaviours of Ni on a series of adsorbents, including activated carbon (Kadirvelu et al. 2001), clays (Gupta & Bhattacharyya 2006), Na-attapulgite (Fan et al. 2009), carbon nanotubes (Chen et al. 2009), mordenite (Yang et al. 2011), magnetic nanoparticles (Panneerselvam et al. 2011), graphene oxides (Chen et al. 2016a), and montmorillonite (Chen et al. 2016b) showed a large variation. Not only that, many feedstock materials in nature have the potential to be cost-efficient adsorbents. Thus, potential adsorbents deserve further exploration.

Biochar can be produced from agricultural residues, the majority of which are waste or by-products from the harvesting and processing of crops (Inyang et al. 2015). Thus, there are dual benefits for environmental protection to utilize agricultural wastes as adsorbents for the removal of heavy metals. Furthermore, many low-cost adsorbents have been used to remove contaminants from water, such as tea factory waste (Malkoc & Nuhoglu 2005), diatomite (Sheng et al. 2011), sepiolite (Liang et al. 2011), diatoms and algae (Bird et al. 2011), peat (Zhou et al. 2012), coke plant wastewaters (Zhou et al. 2017), and zeolite (Ciosek & Luk 2017), which are widely applied to adsorb metal ions and other pollutants from water. It is therefore important to study these kinds of low-cost adsorbents. In this paper, we selected the residual wastes of Oryza sativa, which are scarcely utilized in the field, as a feedstock to obtain biochar through pyrolysis techniques. The sorption properties of this biochar for heavy metals in water are rarely reported. For these reasons, the sorption of Ni(II) on biochar derived from rice straw was investigated under various experimental conditions including contact time, pH, ionic strength, foreign cations and anions, solid content, and temperature. The sorption kinetics was simulated by a pseudo-second-order model to illustrate the possible kinetic interaction mechanism. The equilibrium sorption behaviour of biochar was studied using sorption isotherms. Experimental data were fitted to the Langmuir, Freundlich, and Dubinin-Radushkevich (D-R) models to determine the interaction mechanism between Ni(II) and biochar. The thermodynamics of the sorption process was also investigated to demonstrate the characteristics of Ni(II) sorption on biochar. This exploration for the sorption capacity and underlying mechanisms of rice straw-derived biochar will improve its long-term performance in field applications.

EXPERIMENTAL

Materials

Rice straw, one of the most common agricultural residues, was selected as the feedstock for biochar in this research. Specifically, the aboveground biomass of Oryza sativa L. without seeds was acquired from farmlands in Shaoxing, Zhejiang province, China. Then, the individual plants were washed with tap water, air dried in a hot-air oven at 80°C for 12 hours, crushed, and ground to <1.0 mm particle size using a fodder grinder. Finally, the ground particles were slowly pyrolysed at 600°C in a pyrolyser without admission of air for 2 hours at a heating rate of 7°C/min. All biochar samples were cooled to room temperature and stored in desiccators before use.

A stock solution of Ni ions was prepared by dissolving Ni(NO3)2·6H2O in deionized water. All reagents used in this experiment were of analytical grade with a mass fraction purity of 0.99 and underwent no further purification.

We obtained humic acid (HA) and fulvic acid (FA) by extracting them from the soil in Hua-Jia County (Gansu province, China). It has been proved that both HA and FA are ubiquitous in aquatic environments and have all kinds of functional groups that allow HA and FA to complex with heavy metals and to interact with adsorbents, which finally influences the sorption and transportation of heavy metal by adsorbents (Tang et al. 2013).

Characterization of biochar

Both scanning electron microscopy (SEM) images and transmission electron microscopy (TEM) images of biochar derived from rice straw were obtained using a field emission scanning electron microscope (JSM-6360LV, Japan) and a transmission electron microscope (JEM-1011, Japan), respectively. Simultaneously, the biochar sample was characterized using a Fourier transform infrared (FTIR) spectrometer (NEXUS, USA) in the wavenumber range of 4,000–400 cm−1. A spectral resolution of 2 cm−1 and an energy ratio range of 25–40% were set up to determine the surface functional groups of the biochar.

Batch sorption experiments

The sorption behaviour of biochar from rice straw for Ni(II) was investigated by batch adsorption experiments at three temperatures, i.e., 293 K, 313 K, and 333 K. Briefly, the stock suspensions of Ni(II) ions solution, biochar, NaClO4, KClO4, LiClO4, NaNO3, NaCl, HA, FA, and Milli-Q water were added to polyethylene tubes in order to obtain the desired concentrations of the different constituents. The solution pH was adjusted with 0.01 or 0.1 M HClO4 or NaOH solutions. The batch experiments were carried out in 15 mL polyethylene tubes using the desired pH value, ionic strength, coexisting electrolyte ions, contact time, and temperature. The ratio of the amount of biochar to the solution volume was 0.45 g/L. After the equilibrium time, the solid and liquid phases were separated by centrifugation at 9,000 rpm for 30 min. The concentration of Ni(II) in the supernatant was measured with a Shimadzu 6,300 atomic adsorption spectrophotometer (Japan).

All experiments were run in duplicate or triplicate and the experimental data were averaged to ensure experimental repeatability and improve data accuracy. The relative errors of the experimental data were less than 5%. The Ni(II) removal efficiency by biochar [Ni(II) sorption (%) = (C0Ce)/C0 × 100], distribution coefficient [Kd=(C0Ce)/Ce × V/m], and sorption amount onto biochar [qe = (C0Ce) × V/m] were calculated from the initial Ni(II) concentration (C0, mol/L), the final or equilibrium Ni(II) concentration (Ce, mol/L), the biochar mass (m, g), and the suspension volume (V, L).

RESULTS AND DISCUSSION

Characterization of biochar

The SEM and TEM images for biochar are shown in Figure 1. Both images show the irregularly porous structure of biochar. The SEM image (Figure 1(a)) shows that the surface and the edge of biochar are rough and various blocky structures randomly pile up. The TEM image (Figure 1(b)) also shows the loose inner structure of biochar. All these characteristics may endow the rice straw-derived biochar with favourable sorption abilities. The FTIR spectrum of biochar is shown in Figure 1(c). The broad band at 3,439 cm−1 is attributed to the O-H stretching vibration of biochar (Tong et al. 2011; Kołodyńska et al. 2017). The peaks at 1,426 and 1,036 are assigned to carboxyl OO and alkoxy CO stretching vibrations, respectively (Hu et al. 2017a). The bands at 795, 520, and 468 may be assigned to the integrated crystalline structure of biochar (Zong et al. 2013), which needs to be further studied.

Figure 1

SEM image (a), TEM image (b) and FTIR spectrum (c) of biochar derived from rice straw.

Figure 1

SEM image (a), TEM image (b) and FTIR spectrum (c) of biochar derived from rice straw.

Effect of contact time

The effect of contact time on Ni(II) sorption by biochar and the ratio of t to qt is shown in Figure 2. It is evident that the amount of Ni(II) adsorbed on biochar rapidly increases during the initial 5 hours of contact time, and then the sorption rate slows down until equilibrium is reached. The occurrence of fast sorption during the initial contact time may arise from the availability of large numbers of sites on the biochar surface. With the sites gradually being occupied, the kinetic rate of sorption becomes slower. The mechanisms may involve the less reactive sites on the biochar surface, diffusion into internal sites, and the formation of complexes or precipitation (Chen et al. 2016b). These results imply that the sorption of Ni(II) on biochar is a chemical but not a physical process. In addition, considering the kinetic rates, 24 hours of stirring time was selected to assure that the sorption completely reached equilibrium in the batch experiments.

Figure 2

Time dependence of Ni(II) sorption on biochar and pseudo-second-order rate calculation of Ni(II) sorption on biochar. T = 293 ± 2 K, pH = 6.0 ± 0.2, I (NaClO4) = 0.01 mol/L, m/V = 0.45 g/L.

Figure 2

Time dependence of Ni(II) sorption on biochar and pseudo-second-order rate calculation of Ni(II) sorption on biochar. T = 293 ± 2 K, pH = 6.0 ± 0.2, I (NaClO4) = 0.01 mol/L, m/V = 0.45 g/L.

To study the specific rate constant of Ni(II) sorption on biochar, pseudo-second-order models were used to fit the data. The linear form of the pseudo-second order kinetic was calculated as follows (Ho & McKay 1999; Ho 2006):  
formula
(1)
where qe and qt represent the Ni(II) sorption ability (mol/g) at equilibrium time and time t (h), respectively; and k [g/(mol·h)] is the rate constant. As shown in Figure 2, there is an extremely significant linear relationship between t/qt and t (r2 = 0.995, P < 0.001), suggesting that the kinetic sorption process can be well described by a pseudo-second-order rate equation. This result indicates that the sorption of Ni(II) onto biochar is a chemical process, which is also the rate-limiting step.

Effect of the initial pH and ionic strength of solutions

The pH of a solution can influence functional groups on the adsorbent surface and the species of an adsorbate (Chen et al. 2016a), and thus pH plays a key role in the sorption process. The effect of the initial pH on Ni(II) sorption with 0.001, 0.01, and 0.1 NaClO4 solutions is shown in Figure 3. It is obvious that the sorption of Ni(II) on biochar derived from rice straw is strongly affected by the initial pH values of the solutions at pH <8.8. In addition, the sorption percentage increases with increasing pH, but the sorption curve over the whole pH range here can be divided into three regions: (1) the sorption slowly increases from approximately 8% to 16% over the pH range from 4 to 6; (2) the sorption rapidly increases from 16% to a maximum value of 90% over the pH range from 6 to 8.8; and (3) the sorption percentage maintains its highest level at pH >8.8. A similar sorption curve was found in Ni(II) sorption on montmorillonite (Chen et al. 2016b). This result suggests that the sorption of Ni(II) may be dominated by different sorption mechanisms. For example, at low pH values, cation exchange and inner-sphere surface complexation may be the main mechanisms. The oxygen-containing functional groups on biochar are protonated and Ni(II) ions are adsorbed through ion exchange between Ni2+ and Na+/H+. With an increase of pH, deprotonation occurs, resulting in the charge on the biochar surface changing from positive to negative. This process can benefit the sorption of the positively charged Ni(II) ions through electrostatic attraction. These mechanisms were also found in some other studies: the oxygen-containing groups (C-O, −OH, C = O) in oak bark biochars provided negatively charged surface sites (COO and OH) to attract Pb2+ (Mohan et al. 2007); pyrolysis of sugar beet tailings may yield electron-donor functional groups (C-OH, C-O, C-O-R), which can promote the sorption of chromium through reducing Cr(VI) to Cr(III) (Dong et al. 2011); some cationic nutrients such as Na, K, and Mg in oak and pine wood feedstock materials can increase the cation-exchange abilities of biochars and enhance Pb sorption through ion exchange at low pH values (Mohan et al. 2014).

Figure 3

Effect of pH and ionic strength on Ni(II) sorption on biochar surface. T = 293 ± 2 K, CNi(II)initial = 10 mg/L, m/V = 0.45 g/L.

Figure 3

Effect of pH and ionic strength on Ni(II) sorption on biochar surface. T = 293 ± 2 K, CNi(II)initial = 10 mg/L, m/V = 0.45 g/L.

According to the distribution of Ni(II) species calculated from the relative hydrolysis constants shown in Figure 4, it can be seen that Ni(II) exists in the species of Ni2+, Ni(OH), Ni(OH)2, Ni(OH)3, Ni(OH)42− at different pH values. At pH <8.8, Ni2+ is the predominant species. Therefore, the relatively low sorption of Ni2+ on biochar mainly arises from the competition between H+/Na+ and Ni2+ on the biochar surface. In other words, Ni2+ is mainly removed through ion exchange. Similar results were found in many studies, which attribute them to the ion-exchange process (Tan et al. 2008a; Chen et al. 2016b).

Figure 4

The distribution of Ni(II) species as a function of pH values. T = 293 ± 2 K, CNi(II)initial = 10 mg/L, m/V = 0.45 g/L.

Figure 4

The distribution of Ni(II) species as a function of pH values. T = 293 ± 2 K, CNi(II)initial = 10 mg/L, m/V = 0.45 g/L.

As seen in Figure 3, there is no drastic difference in the Ni(II) sorption on biochar in the three different solution concentrations of NaClO4 over the pH range of 4–11. The pH-dependent and ionic strength-independent sorption for Ni(II) ions suggests Ni(II) sorption on biochar is mainly due to inner-sphere surface complexation (Hayes & Leclie 1987). The precipitation curve of Ni(II) at the concentration of 10 mg/L is also shown in Figure 3. It is clear that Ni(II) starts to form a precipitate at pH ∼8.8 if it is not adsorbed on biochar. At pH <8.8, the percentage of Ni(II) removal reaches its maximum value. Thus, Ni(II) removal from solution to biochar is not attributed to precipitation. Nevertheless, in view of the strong sorption of Ni(II) on biochar, the local Ni(II) concentration on the biochar surface is very high, and the formation of (co)precipitates on the biochar surface is also possible (Tan et al. 2014). Currently, the EXAFS technique provides insight into the formation of surface (co)precipitates and can help to distinguish the precipitates from inner-sphere surface complexes (Tan et al. 2008b). Therefore, further investigation is necessary at the molecular level to get an insight into the sorption mechanisms.

Effect of coexistent electrolyte ions

In practice, biochars derived from different materials display different sorption capacities and pathways for heavy metals under various aqueous conditions. In other words, the sorption process is determined by the identity of the adsorbent and the adsorbate, and the physicochemical conditions of the aqueous environment. To illustrate the effect of coexistent electrolyte ions on the sorption of Ni(II), the sorption behaviour of Ni(II) was investigated in multiple solutions as a function of the pH value and is shown in Figure 5. Ni(II) sorption on biochar is not affected by the electrolyte cations in the suspension, as can be seen from Figure 5(a). However, Chen et al. found the opposite result when investigating the sorption of Ni(II) on montmorillonite (Chen et al. 2016b). This effect of cations in solution on the sorption of heavy metals may be determined by different adsorbents. Thus, the role of cations in the sorption of Ni(II) on biochar is negligible over the entire pH range. The effect of ionic strength on the sorption of Ni(II) also proves that the sorption is chemical sorption and inner-sphere surface complexation rather than physical adsorption. Figure 5(b) shows the sorption of Ni(II) on biochar as a function of pH in 0.01 mol/L NaCl, NaNO3, NaClO4, respectively. These results demonstrate that there are no significant differences among the three electrolyte anions. Generally, the negatively charged anions can form complexes with the oxygen-containing groups on the surface of a solid. However, the influences of Cl, NO3, and ClO4 on the sorption of Ni(II) are weak, suggesting that inner-sphere surface complexes are formed. This discovery is similar to Ni(II) sorption on montmorillonite (Chen et al. 2016b), U(VI) sorption on bentonite (Xiao et al. 2013), and Eu(III) sorption on graphene oxide (Hu et al. 2017a) but is different from Eu(III) sorption on titanate nanotubes (Sheng et al. 2012). This indicates that the sorption behaviour of a heavy metal is co-determined by the characters of the adsorbate, adsorbent, and environment. In addition, the mechanism of specific anions on biochar is difficult to distinguish using macroscopic investigation only. Thus, we need to further investigate related mechanisms at the molecular level to gain in-depth information.

Figure 5

Effect of coexistent electrolyte cations (a) and anions (b) on Ni(II) sorption on biochar surface at different pH values. T = 293 ± 2 K, CNi(II)initial = 10 mg/L, m/V = 0.45 g/L.

Figure 5

Effect of coexistent electrolyte cations (a) and anions (b) on Ni(II) sorption on biochar surface at different pH values. T = 293 ± 2 K, CNi(II)initial = 10 mg/L, m/V = 0.45 g/L.

Effect of solid content

Figure 6 illustrates the effects of the solid content on the Ni(II) sorption percent and sorption amount on biochar. One can see that the sorption percent increases from ∼6% to 34% as the solid content increases from 0.1 g/L to 1.2 g/L. In general, the greater the applied solid content, the more available binding sites there are. Therefore, more Ni(II) ions can be bound with an increase in the solid content. It should be noted that the removal percent of Ni(II) on biochar reaches its highest level at m/V >0.9 g/L. To reduce the purification cost of effluent, the appropriate content of biochar derived from rice straw to remove Ni(II) at pH 6.0 is 0.9 g/L.

Figure 6

Effect of solid content on the Ni(II) sorption percentage and Ni(II) sorption amount on biochar derived from rice straw. pH = 6.0 ± 0.2, T= 293 ± 2 K, CNi(II)initial = 10 mg/L, I (NaClO4) = 0.01 mol/L.

Figure 6

Effect of solid content on the Ni(II) sorption percentage and Ni(II) sorption amount on biochar derived from rice straw. pH = 6.0 ± 0.2, T= 293 ± 2 K, CNi(II)initial = 10 mg/L, I (NaClO4) = 0.01 mol/L.

As shown in Figure 6, the Ni(II) sorption amount decreases with rising biochar dosage. One possible reason is that the biochar particles disperse well in the solution at a lower solid dosage, which causes all the biochar surface sites to be exposed for more Ni(II) binding. Thereby, there is a higher sorption amount at a lower solid dosage. However, at a higher solid dosage, the collision among biochar particles can cause aggregation and decrease their dispersion ability in the solution. This process would reduce the total specific surface area and increase the path length of diffusion, finally resulting in the reduction of the availability of binding sites and Ni(II) sorption amount on biochar (Huang et al. 2008; Yang et al. 2010). In addition, some weak-linked Ni(II) on the biochar surfaces may be desorbed due to the collision between individual solid particles (Dong et al. 2014), which is also adverse to the sorption of Ni(II) on the biochar surface.

Effect of humic substances

Humic substances (HA/FA), ubiquitous in the natural environment, can mediate the sorption and transport of metal ions by forming complexes with metal ions (Sheng et al. 2016). Here, the pH-dependence of Ni(II) sorption on biochar in the absence and presence of HA/FA is shown in Figure 7. The presence of HA/FA enhances Ni(II) sorption on biochar at pH <7.2, it inhibits sorption at pH >7.2. There is no significant difference between the two sorption curves in the presence of HA and FA over the whole pH range. Similar effects of HA/FA have been found in many studies, including Ni(II) sorption on Na-attapulgite (Fan et al. 2009), Th(IV) sorption on NKF-6 zeolite (Wang et al. 2016), Eu(III) and UO22+ sorption on graphene oxide (Hu et al. 2017a, 2017b), and so on. These findings suggest the effects of HA/FA on sorption may be independent of the properties of adsorbents and the adsorbates but dependent on the pH of the solution. At low pH values, it is easy to attract the negatively charged HA/FA onto the positively charged biochar due to the stronger complexation ability between HA/FA and biochar compared with the ability between HA/FA and Ni(II). Subsequently, the electrostatic properties of the solid-aqueous interface are modified by the HA/FA adsorbed onto the biochar surface, which then provides a more favourable environment for Ni(II) sorption (Strathmann & Myneni 2005). Thereby, the addition of HA/FA enhances Ni(II) sorption at low pH values. However, at high pH values, the solubility of HA/FA increases and the charge on the biochar surface becomes negative, which causes the negatively charged HA/FA to be difficult to adsorb on the biochar surface. In addition, a part of the free stable HA/FA-Ni(II) complexes are formed in solution. Therefore, the sorption of Ni(II) on biochar significantly decreases at high pH values.

Figure 7

Effect of both HA and FA on Ni(II) sorption onto biochar derived from rice straw. pH = 6.0 ± 0.2, T= 293 ± 2 K, CNi(II)initial = 10 mg/L, I (NaClO4) = 0.01 mol/L, m/V = 0.45 g/L.

Figure 7

Effect of both HA and FA on Ni(II) sorption onto biochar derived from rice straw. pH = 6.0 ± 0.2, T= 293 ± 2 K, CNi(II)initial = 10 mg/L, I (NaClO4) = 0.01 mol/L, m/V = 0.45 g/L.

Adsorption isotherms and thermodynamic study

Ni(II) sorption isotherms on biochar at 293, 313, and 333 K are show in Figure 8. The sorption amount of Ni(II) is the lowest at 293 K and highest at 313 K, indicating that high temperature is beneficial for the sorption of Ni(II) on biochar. To illustrate the sorption characteristics of Ni(II) on biochar, the Langmuir, Freundlich, and D-R models (Chen et al. 2016b) were simulated and the results are shown in Figure 8(a)8(d). The linear forms of the three isotherm models can be expressed as follows:  
formula
(2)
 
formula
(3)
 
formula
(4)
where qmax (mol/g) is the Ni(II) maximum sorption capacity on a per-weight basis of biochar; KL (L/mol) is the Langmuir affinity parameter; KF (mol1−n·Ln/g) and n represent the Freundlich affinity-capacity parameter and exponent, respectively; β is the D-R activity constant and ɛ is the Polanyi potential, which is calculated by RT ln(1 + 1/Ce). From the D-R model, the bonding energy of the ion-exchange mechanism (E, kJ/g) is calculated by (2β)−1/2. Here, R is the universal gas constant [8.314 J/(K·mol)] and T (K) is the absolute temperature in kelvin.
Figure 8

Sorption isotherm (a), linearized Langmuir isotherm (b), linearized Freundlich isotherm (c), and linearized Dubinin-Radushkevich isotherm (d) for Ni(II) sorption onto biochar derived from rice straw at 293 K, 313 K, and 333 K. pH = 6.0 ± 0.2, CNi(II)initial = 10–35 mg/L, I (NaClO4) = 0.01 mol/L, m/V = 0.45 g/L.

Figure 8

Sorption isotherm (a), linearized Langmuir isotherm (b), linearized Freundlich isotherm (c), and linearized Dubinin-Radushkevich isotherm (d) for Ni(II) sorption onto biochar derived from rice straw at 293 K, 313 K, and 333 K. pH = 6.0 ± 0.2, CNi(II)initial = 10–35 mg/L, I (NaClO4) = 0.01 mol/L, m/V = 0.45 g/L.

The corresponding parameters of the three isotherms are also shown in Table 1. According to the curves of the simulated models and parameters listed in Table 1, it was found that the sorption isotherms fit very well to the three isotherm models. Especially for the Langmuir model, its regression coefficients of determination (R2) are higher than those in the other two models, demonstrating that the Langmuir model best fits the isotherm. The result implies that the binding energies on the biochar are uniform and Ni(II) is adsorbed on the biochar surface by forming an almost complete monolayer coverage (Langmuir 1918). In addition, the values of qmax calculated from the Langmuir model increase with an increase of temperature, which reveals that high temperature is beneficial to Ni(II) sorption on biochar. In the D-R model, the qmax values are different from the qmax values in Langmuir model. This discrepancy may be attributed to the different assumptions between the two models.

Table 1

The isotherm parameters derived from the Langmuir, Freundlich, and D-R models for Ni(II) sorption onto biochar at three temperatures. pH = 6.0 ± 0.2, CNi(II)initial = 10 mg/L, I (NaClO4) = 0.01 mol/L, m/V = 0.45 g/L

T (K) Langmuir model
 
qmax (mol/g) b (L/mol) R2 
293 1.889 × 10−4 3.811 × 104 0.9868 
313 2.389 × 10−4 8.408 × 104 0.9903 
333 2.567 × 10−4 3.736 × 104 0.9990 
 Freundlich model
 
T (K) KF (mol1−n·Ln/g) R2 
293 7.741 × 10−3 0.4806 0.9355 
313 7.042 × 10−3 0.4302 0.9313 
333 1.300 × 10−3 0.1994 0.9516 
 D-R model
 
T (K) qmax (mol/g) β R2 
293 8.988 × 10−4 0.0267 0.9481 
313 1.003 × 10−3 0.0236 0.9448 
333 5.088 × 10−4 0.0105 0.9678 
T (K) Langmuir model
 
qmax (mol/g) b (L/mol) R2 
293 1.889 × 10−4 3.811 × 104 0.9868 
313 2.389 × 10−4 8.408 × 104 0.9903 
333 2.567 × 10−4 3.736 × 104 0.9990 
 Freundlich model
 
T (K) KF (mol1−n·Ln/g) R2 
293 7.741 × 10−3 0.4806 0.9355 
313 7.042 × 10−3 0.4302 0.9313 
333 1.300 × 10−3 0.1994 0.9516 
 D-R model
 
T (K) qmax (mol/g) β R2 
293 8.988 × 10−4 0.0267 0.9481 
313 1.003 × 10−3 0.0236 0.9448 
333 5.088 × 10−4 0.0105 0.9678 

Here, Ni(II) maximum uptake capacity (i.e., qmax) of rice straw-derived biochar fitted by the Langmuir model was compared with the capacities of other sorbent materials. As displayed in Table 2, the qmax values of Ni(II) uptake on rice straw-derived biochar is greater than that of a series of materials including bagasse (Gupta et al. 2003), CMC-bentonite (Liu et al. 2012), Na-rectorite (Tan et al. 2008c), Alternanthera philoxeroides biomass (Wang & Qin 2006), Phragmites australis biochar (Liu et al. 2016), calcium-alginate (Huang et al. 1996), and magnetic nanoparticles (Sharma & Srivastava 2011), while it is lower than Chrysanthemum indicum biochar (Vilvanathan & Shanthakumar 2017) and Citrus limetta peel biochar (Vidhya et al. 2017). The result implies that rice straw-derived biochar is favourable but still needs a further-modified adsorbent to raise the sorption capacity. For example, another bio-sorbent (pine sawdust) exhibited an adsorption capacity more than three times higher when it was modified by citric acid compared with the untreated pine sawdust (Zhou et al. 2015).

Table 2

Adsorption comparison of biochar and several reported materials for Ni(II)

Materials qmax (mg/g) References 
Bagasse 1.70 Gupta et al. (2003)  
CMC-bentonite 2.89 Liu et al. (2012)  
Na-rectorite 7.18 Tan et al. (2008c)  
Alternanthera philoxeroides biomass 9.73 Wang & Qin (2006)  
Phragmites australis biochar 9.75 Liu et al. (2016)  
Calcium-alginate 10.50 Huang et al. (1996)  
Magnetic nanoparticles 11.53 Sharma & Srivastava (2011)  
Rice straw biochar 11.96 This work 
Chrysanthemum indicum biochar 29.44 Vilvanathan & Shanthakumar (2017)  
Citrus limetta peel biochar 86.99 Vidhya et al. (2017)  
Materials qmax (mg/g) References 
Bagasse 1.70 Gupta et al. (2003)  
CMC-bentonite 2.89 Liu et al. (2012)  
Na-rectorite 7.18 Tan et al. (2008c)  
Alternanthera philoxeroides biomass 9.73 Wang & Qin (2006)  
Phragmites australis biochar 9.75 Liu et al. (2016)  
Calcium-alginate 10.50 Huang et al. (1996)  
Magnetic nanoparticles 11.53 Sharma & Srivastava (2011)  
Rice straw biochar 11.96 This work 
Chrysanthemum indicum biochar 29.44 Vilvanathan & Shanthakumar (2017)  
Citrus limetta peel biochar 86.99 Vidhya et al. (2017)  

However, rice straw biochar has two prominent merits compared with the other two superior adsorbents: (1) the feedstock is easier to acquire; (2) it simultaneously solves the problem of environmental pollution from agriculture in China. In addition, the EPA provides the maximum contaminant levels (MCLs) of some heavy metals in ground water and drinking water, including lead (0.015 mg/L), mercury (0.002 mg/L), cadmium (0.005 mg/L), chromium (0.100 mg/L), zinc (5.000 mg/L), copper (1.300 mg/L), manganese (0.050 mg/L), and so on (Inyang et al. 2015). In this study, the rice-straw derived biochar can adsorb ∼6.787 mg/L Ni (the maximum value simulated by Langmuir) from water. The data suggest that the biochar has the potential to meet the standard. Finally, it is noted that natural bio-sorbents are generally considered to be a cost-effective adsorbent alternatives for the removal of divalent heavy metal ions (Zhou et al. 2015). Therefore, rice straw-derived biochar deserves to be further explored as a new bio-sorbent.

To further investigate the thermodynamic properties of Ni(II) sorption on biochar, the sorption isotherms were employed to calculate the thermodynamic parameters. Gibbs free energy (ΔG0) is calculated from the equation (Yang et al. 2011):  
formula
(5)
where K0 represents the equilibrium constant of the sorption reaction, and the values of lnK0 are obtained by plotting lnKd versus Ce. Standard entropy (ΔS0) and the average standard enthalpy (ΔH0) are calculated from the slope and intercept of the plot of lnK0 versus 1/T, respectively (Li et al. 2012).  
formula
(6)

The obtained thermodynamic parameters are listed in Table 3. The positive value of ΔH0 indicates that the adsorption of Ni(II) on biochar is endothermic. In fact, the process of ions attaching to a solid surface is exothermic. However, the hydration sheaths of Ni(II) ions are partially destroyed in the sorption process, which requires energy (Chen & Wang 2006). Because the endothermic nature of the desolvation process exceeds that of the enthalpy of adsorption to a considerable extent, the final sorption is endothermic. The value of ΔG0 is negative and becomes more negative as the temperature increases, implying that the sorption process is spontaneous and more efficient at higher temperatures. A positive value of ΔS0 reflects the partial structural changes of Ni(II) and the biochar during the sorption process, leading to an increase in randomness at the biochar-solution interface. All the above analyses indicate that the sorption process of Ni(II) on biochar is spontaneous and endothermic.

Table 3

The thermodynamic parameters for Ni(II) sorption on biochar at three temperatures. pH = 6.0 ± 0.2, CNi(II)initial = 10 mg/L, I (NaClO4) = 0.01 mol/L, m/V = 0.45 g/L

T (K) ΔGo (KJ/mg) ΔSo [J/(mg·K)] ΔHo (KJ/mg) 
293 −17.810 15.763 13.191 
313 −20.272  15.338 
333 −24.114  18.865 
T (K) ΔGo (KJ/mg) ΔSo [J/(mg·K)] ΔHo (KJ/mg) 
293 −17.810 15.763 13.191 
313 −20.272  15.338 
333 −24.114  18.865 

CONCLUSIONS

In this paper, biochar derived from rice straw was characterized by using SEM, TEM, and FTIR. A batch technique was adopted to investigate the performance of biochar in the removal of Ni(II) as a function of various environmental conditions including pH, ion strength, foreign cations and anions, contact time, solid content, humic substances, and temperature. The sorption of Ni(II) on biochar rapidly reached equilibrium and kinetic studies suggested that the sorption can be well described by a pseudo-second-order kinetic model. Ni(II) sorption was strongly dependent on pH but independent of ionic strength over the whole pH range. Ion exchange and inner-sphere surface complexation dominated the sorption process over the whole pH range. Solid content was also an important factor in controlling the sorption amount of Ni(II) on biochar and the appropriate content of biochar was 0.9 g/L. The effect of HA/FA on Ni(II) sorption was dependent on pH: the sorption was enhanced at low pH values but was reduced at high pH values. The thermodynamic data derived from temperature-dependent sorption isotherms suggested that the sorption of Ni(II) on biochar was an endothermic and spontaneous process. The above results are important for the application of biochar derived from rice straw to effluent management.

ACKNOWLEDGEMENTS

Financial support from the National Natural Science Foundation of China (31700476) is gratefully acknowledged.

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