An acid modified approach to enhance the adsorption capacity of low-rank coal (lignite) is proposed to reduce the risk of heavy metal ions within the wastewater. Adsorption kinetics, adsorption thermodynamics, adsorption coefficient and density functional theory DFT calculations were studied in this paper, respectively. The results indicate that the adsorption capacity of lignite was enlarged after HNO3 modification, and Pseudo-second order kinetics model and Langmuir isothermal adsorption model can be used to describe the adsorption process. The surface chemical properties of lignite play a dominant role rather than the specific surface area and total pore volume in the Pb(II) cation adsorption process, and it is suggested that the adsorption of Pb(II) cation by Raw lignite (RL) and Modified lignite (ML) is mainly completed by chemical adsorption. The Fourier transform infrared spectroscopy (FTIR) characterization showed that the surface oxygen functional groups of lignite increased after modification. The results of interaction energies between the model molecule and Pb(II) cation show that Pb(II) cation and -C-O-C are most easily combined, followed by -COOH, and -C = O is the weakest.

  • The adsorption mechanisms of raw and modified lignite and Pb(II) cation were summarized.

  • The original hemidirected structure of Pb(II) cation is broken, and a polar covalent bond is formed.

  • A large number of flocculent precipitates were generated rapidly with the chelating reaction between humic acid molecules and Pb(II) cation.

  • The adsorption of Pb(II) cation by lignite was a chemical adsorption process.

Graphical Abstract

Graphical Abstract
Graphical Abstract

With the development of China's economy and the expansion of industrial scale, a large number of industrial wastewaters containing heavy metals have been produced, especially in smelting, painting, electrolysis, alloy, electroplating and other industries (Yang & Li 2017). Heavy metal ions have the characteristics of strong toxicity, carcinogenicity, easy enrichment and difficult degradation, which have serious harm to the surrounding ecological environment and human health (Fu & Wang 2011). Therefore, industrial wastewater containing heavy metal ions has become the focus and difficulty of wastewater treatment in China (Wang & Chen 2015).

At present, the commonly used heavy metal wastewater treatment methods mainly include chemical precipitation methods (hydroxide precipitation method, sulfide precipitation method, ferrite method, etc.) (Chen et al. 2018), electrochemical methods (electrocoagulation method, micro-electrolysis, electroreduction method, etc.) (Tran et al. 2017), membrane separation methods (electrodialysis method, nanofiltration method, reverse osmosis method, etc.) (Abdullah et al. 2019), adsorption methods (physical adsorption method, chemical adsorption method, resin adsorption and biological adsorption method, etc.) and ion exchange methods (Dabrowski et al. 2004; Stafiej & Pyrzynska 2007; Nascimento et al. 2009). Compared with other methods, adsorption techniques are widely used for the removal of certain classes of pollutants from wastewater, especially those that are not easily biodegradable (Pehlivan & Arslan 2007; Li & Zhou 2018), and searching for cheap, efficient and environmentally friendly adsorbent materials is the focus of the adsorption method. As a conventional adsorbent, commercial activated carbon has a good performance, but the high cost restricts its widespread use. As a natural porous media, the adsorption performance of coal is of interest to scholars, and experimental results showed that coal as a kind of adsorbent material has achieved a good treatment effect on wastewater.

Lignite is a kind of coal with low coalification degree, loose composition structure and developed internal pore structure, which is similar to the traditional adsorbents (Polat et al. 2006). Based on the high specific surface area, high porosity and rich oxygen functional groups of lignite, it can be used as adsorbent for organic and inorganic wastewater treatment (Uçurum 2009; Doskočil & Pekař 2012). But the ability to treat wastewater is limited, and the properties are unstable. Therefore, in order to improve its adsorption capacity, surface modification was used (Mohan & Chander 2006; Huang et al. 2019). However, the adsorption mechanism of modified lignite in the adsorption process is not clear, so it is necessary to explain it through experiments.

In this paper, based on the changes of surface physical and chemical properties of lignite before and after modification, the adsorption capacity and adsorption type of lignite were studied by adsorption process kinetics and thermodynamics. Sodium humate was used to simulate lignite molecules to verify the role of chemical adsorption reactions such as coordination chelating in the adsorption of Pb(II) cation by lignite. The adsorption morphology of pollutants on the coal surface, the adsorption activation sites on the coal surface and the spatial structure of the adsorbent were explained.

Materials

In this experiment, lignite (Xilinguole, Inner Mongolia) (45–75 μm) was selected as adsorbent, and HNO3 (aladdin, ≥ACS, 70%) was selected as the surface modification agent. Pb(NO3)2 was used to represent wastewater which contains Pb(II) cation.

Methods

Preparation of modified lignite

Firstly, 5.0 mol·L−1 HNO3 solution was prepared, and 100.0 g lignite was weighed and added into a 1,000 mL conical flask. Then 500 mL HNO3 solution was added and stirred in a magnetic stirrer at 298 K and 300 rpm for 4.0 h. After modification, the coal sample was washed with deionized water to neutral and dried in a vacuum drying oven at 348 K for 24 h. Finally, HNO3 modified lignite was obtained.

FTIR test

1.0 g raw lignite and modified lignite samples were weighed by an electronic analytical balance, and the particle size was in the range 45–74 μm. After the sample preparation was completed, FTIR was performed using Nicolet IS 10 Fourier transform infrared spectrometer.

Adsorption kinetics

100 mL Pb(II) cation solution with a concentration of 300 mg·L−1 was accurately measured by a volumetric cylinder, and added to a 250 mL conical flask. The pH of Pb(II) cation solution was adjusted to 4.0. 0.5 g of RL and ML were weighed by an electronic analytical balance, and added to the conical flask, shaking at 298 K for 15, 30, 60, 120, 140, 360, 540 and 720 min, respectively. The mixed solution was filtered by 0.45 μm filter membrane, and the residual Pb(II) cation ion concentration in the filtrate was determined by UV spectrophotometer.

The adsorption process was fitted by pseudo-first-order kinetic equation and pseudo-second-order kinetic equation (Sun et al. 2018). The linear expressions are shown in Equations (1) and (2), respectively.
formula
(1)
formula
(2)
where qe (mg·g−1) is the equilibrium adsorption amount, qt (mg·g−1) is the adsorption amount at time t (min), K1 (min−1) is the pseudo-first-order kinetic adsorption rate constant (min−1), and K2 is the pseudo-second-order kinetic adsorption rate constant, g·mg−1·min−1.

Adsorption thermodynamics

100 mL Pb(II) cation solutions with a concentration of 50, 100, 200, 300, 400, 500, 600 mg·L−1 were accurately measured by a volumetric cylinder, and added to a 250 mL conical flask. The pH of Pb(II) cation solution was adjusted to 4.0. 0.5 g of RL and ML were weighed by an electronic analytical balance, and added to the conical flask, shaking at 298 K for 120 min. The mixed solution was filtered by 0.45 μm filter membrane, and the residual Pb(II) cation ion concentration in the filtrate was determined by UV spectrophotometer.

The adsorption process was fitted by Langmuir and Freundlich isothermal adsorption models. The linear expressions are shown in Equations (3) and (4), respectively (Shrestha et al. 2013).
formula
(3)
formula
(4)
where KL is a constant related to the adsorption activation energy, L·mg−1, qm is the monolayer adsorption amount, mg·g−1, Ceq is the concentration of adsorbate in the solution at equilibrium, mg·L−1, KFr is the parameter characterizing the adsorption capacity in Freundlich, is a parameter for evaluating adsorption superiority.
The adsorption free energy ΔG° can be calculated by the following equations.
formula
(5)
formula
(6)
formula
(7)
where R is the ideal gas constant (8.314 J·mol−1·K−1), T is the absolute adsorption temperature, K. K0 was obtained by the KFr in the Freundlich model, and the ΔG° value was calculated by the KFr value of the Freundlich model (Gürses et al. 2014).

Simulation details

Density functional theory (DFT) calculations of Pb(II) cation and different oxygen-containing functional groups of lignite were implemented in the Forcite and DMol3 program (Material Studio version 2018 software, Accerly Corporation). First, Ben-COOH, Ben-C = O, Ben-C-O-C and Pb(II) cation models are built on MS visualizer; the molecular models are shown in Figure 1 (Meuser et al. 2015). Then, the calculation between Ben-COOH, Ben-C = O, Ben-C-O-C and Pb(II) cation adsorption locator was performed, respectively. Choose universal as the forcefield during all the simulation process, and choose surface region defined by atom set as the location. Finally, Forcite anneal was performed for each system, and DFT calculation was performed after annealing. The initial temperature was 298.0 K, the mid-cycle temperature was 1,098.0 K, heating ramps per cycle was 16, and dynamics steps per ramp was 625 during Forcite anneal. GGA-PBE was selected as functional in the DMol3 calculation (Moreira da Costa et al. 2012; Gao et al. 2017; Ribeiro et al. 2019).

Figure 1

Structure of the hemidirected [Pb(H2O)4]2+ (a), Ben-COOH (b), Ben-C = O (c) and Ben-C-O-C (d). Where H is white, C is gray, O is red and Pb is black.

Figure 1

Structure of the hemidirected [Pb(H2O)4]2+ (a), Ben-COOH (b), Ben-C = O (c) and Ben-C-O-C (d). Where H is white, C is gray, O is red and Pb is black.

Adsorbent characterization

The industrial analysis and elemental analysis of lignite are shown in Table 1.

Table 1

Proximate analysis and ultimate analysis of experiment lignite

Proximate analysis (Wt/%)
Ultimate analysis (Wt/%)
MadAadVadFCadCadHadOadNadSad
8.72 10.98 35.65 44.65 68.78 5.09 24.40 1.02 0.71 
Proximate analysis (Wt/%)
Ultimate analysis (Wt/%)
MadAadVadFCadCadHadOadNadSad
8.72 10.98 35.65 44.65 68.78 5.09 24.40 1.02 0.71 

ad, air drying basis; A, ash; FC, fixed carbon; M, moisture; V, volatile matter.

It can be seen from Table 1 that the content of Oad reaches 24.40%, indicating that there are a large number of oxygen-containing functional groups in the coal. The type and quantity of functional groups on coal surface have great influence on chemical adsorption of coal, and the infrared spectra of lignite before and after modification are shown in Figure 2.

Figure 2

FTIR spectra of raw lignite and modified lignite.

Figure 2

FTIR spectra of raw lignite and modified lignite.

In the range of 3,200–3,600 cm−1, this section is mainly the absorption band of phenolic hydroxyl (Ar-OH) in coal; after HNO3 modification, the width and intensity of the peak are significantly enhanced, indicating that the content of phenolic hydroxyl in modified coal samples is increased. In the range of 1,400–1,800 cm−1, there is a strong absorption peak at 1,620 cm−1 and a weak absorption peak at 1,710 cm−1. The peak at 1,620 cm−1 is caused by the stretching vibration of aromatic ring C = C and the conjugated stretching of benzene ring. The peak intensity at 1,710 cm−1 has a good correlation with-COOH in coal, which is the characteristic peak of carboxyl group in coal. Compared with the infrared spectra of lignite under different HNO3 modification conditions, it was found that the peak intensity at 1,710 cm−1 was significantly enhanced. With the deepening of HNO3 modification, the peak gradually changed from a weak absorption peak to a strong absorption peak, indicating that the content of carboxyl group in the lignite structure increased during HNO3 modification. In the range of 900–1,200 cm−1, the characteristic absorption peaks in this region are relatively complex, which are usually composite superposition absorption peaks. From the infrared spectra of the tested coal samples, the waveform in this region is disordered, and there is no particularly prominent strong absorption peak, but there are two weak absorption peaks at 1,090 cm−1 and 1,030 cm−1, which can be attributed to the stretching vibration of C-O in coal and the stretching vibration of C-O-C in ethers in coal (Georgakopoulos 2003). Compared with the raw coal, there is no new spectral peak in this region, but the strength is slightly enhanced, indicating that the content of oxygen-containing functional groups on the surface of lignite after HNO3 modification is increased (Qi et al. 2011; Feng et al. 2017).

Adsorption kinetics

The kinetic curves of Pb(II) cation adsorption by RL and ML are shown in Figure 3.

Figure 3

Dynamic curves of raw lignite and modified lignite.

Figure 3

Dynamic curves of raw lignite and modified lignite.

It can be seen from Figure 3 that at the beginning of adsorption, the adsorption rate of Pb(II) cation by lignite was fast; this was because there were a large number of adsorption active sites on the surface of lignite and on the inner surface of macropores at the beginning of adsorption. Pb(II) cation quickly occupied the outer surface of the adsorbent and diffused to the inner pore. After 120 min of adsorption, the adsorption rate became smooth and finally reached the adsorption equilibrium. With the adsorption time going on, the adsorption active sites on the adsorbent surface reduced, and the adsorption rate of Pb(II) cation with the surface functional groups decreased. Considering the Pb(II) cation adsorption capacity and time efficiency, the Pb(II) cation adsorption time was determined to be 120 min in subsequent experiments.

The adsorption process was fitted by the pseudo-first-order kinetic equation and the pseudo-second-order kinetic equation, respectively. The results are shown in Table 2.

Table 2

Dynamic fitting parameters of raw lignite and modified lignite

SamplesPseudo-first-order kinetic equation
Pseudo-second-order kinetic equation
/mg·g−1K1/min−1R2/mg·g−1K2/g·mg−1·min−1h/mg·g−1·min−1R2
RL 38.44 0.0048 0.6501 36.76 0.0028 4.53 0.9999 
ML 57.52 0.0044 0.6612 56.49 0.0033 10.62 0.9999 
SamplesPseudo-first-order kinetic equation
Pseudo-second-order kinetic equation
/mg·g−1K1/min−1R2/mg·g−1K2/g·mg−1·min−1h/mg·g−1·min−1R2
RL 38.44 0.0048 0.6501 36.76 0.0028 4.53 0.9999 
ML 57.52 0.0044 0.6612 56.49 0.0033 10.62 0.9999 

It can be seen from Table 2 that the correlation coefficient R2 of the pseudo-second-order kinetic equation for the adsorption of Pb(II) cation by lignite is higher than pseudo-first-order kinetic equation, and reaching 0.9999, it indicates that the adsorption process can be described by pseudo-second-order kinetic equation. The value of initial adsorption rate h (mg·g−1·min−1) in the pseudo-second-order kinetic rate model was ML (10.62) > RL (4.53), which indicates that the initial adsorption rate of Pb(II) cation adsorption on ML was higher than that on RL.

Adsorption thermodynamics

The variation curves of Pb(II) cation adsorption by RL and ML at different temperatures are shown in Figure 4.

Figure 4

Thermodynamics curves of raw lignite and modified lignite. Raw lignite (a), Modified lignite (b).

Figure 4

Thermodynamics curves of raw lignite and modified lignite. Raw lignite (a), Modified lignite (b).

Figure 4 presents that with the increase of adsorption temperature, the adsorption capacity of RL and ML on Pb(II) cation increased, but the increase was not significant, which indicates that the adsorption capacity of lignite can improve with the increase of temperature, but the effect is little. The adsorption process was fitted by Langmuir and Freundlich adsorption isotherm equations, respectively, and the results are shown in Table 3.

Table 3

Adsorption isotherm equations fitting parameters of raw lignite and modified lignite (298 K)

SamplesLangmuir
Freundlich
qm /mg·g−1KL/L·mg−1RLR2KF/mg·g−11/nR2
RL 60.98 0.0256 0.1288 0.9982 1.32 0.83 0.8177 
ML 88.50 0.0655 0.0366 0.9965 9.56 0.53 0.8840 
SamplesLangmuir
Freundlich
qm /mg·g−1KL/L·mg−1RLR2KF/mg·g−11/nR2
RL 60.98 0.0256 0.1288 0.9982 1.32 0.83 0.8177 
ML 88.50 0.0655 0.0366 0.9965 9.56 0.53 0.8840 

It can be seen from Table 3 that the R2 fitted by the Langmuir model was greater than 0.99, which was higher than that fitted by the Freundlich model. It indicates that the adsorption process can be described by the Langmuir isothermal adsorption model better, and Pb(II) cation was adsorbed on the surface of modified lignite in the form of a monolayer. The adsorption sites on the surface of modified lignite were uniform and the adsorption capacities were the same. At the same time, comparing the maximum theoretical monolayer adsorption capacities, the maximum adsorption capacities were 60.98 and 88.50 mg·g−1, respectively. That is to say, the adsorption capacity of lignite for Pb(II) cation was enhanced after modification by HNO3. The ΔG°, ΔH° and ΔS° of the adsorption process were calculated to explore the reaction trend and characteristics of the adsorption process. The results are shown in Table 4.

Table 4

Thermodynamics fitting parameters of raw lignite and modified lignite

SampleT(K)Kd(ml·g−1)lnKdΔG°(kJ/mol)ΔH°(kJ/mol)ΔS°(J/mol. K)
RL 298 2.15 0.76 −1.89 11.04 43.42 
308 2.51 0.92 −2.36 
318 2.84 1.04 −2.57 
ML 298 12.19 2.50 −6.19 47.11 178.65 
308 20.66 3.03 −7.75 
318 40.38 3.70 −9.78 
SampleT(K)Kd(ml·g−1)lnKdΔG°(kJ/mol)ΔH°(kJ/mol)ΔS°(J/mol. K)
RL 298 2.15 0.76 −1.89 11.04 43.42 
308 2.51 0.92 −2.36 
318 2.84 1.04 −2.57 
ML 298 12.19 2.50 −6.19 47.11 178.65 
308 20.66 3.03 −7.75 
318 40.38 3.70 −9.78 

Where Kd is calculated when C0 = 300 mg·L−1.

The relationship of lnKd and 1,000/T is shown in Figure 5.

Figure 5

Relationship between lnkd and 1,000/T of Pb(II) cation adsorbed by RL and ML.

Figure 5

Relationship between lnkd and 1,000/T of Pb(II) cation adsorbed by RL and ML.

The ΔG° of Pb(II) cation adsorption by lignite is negative, which indicates the adsorption process is spontaneous. The absolute value of ΔG° increases with the increase of temperature, which indicates that the adsorption reaction trend increases with the increase of temperature. However, the increase in adsorption capacity was not obvious, so the adsorption temperature was not a significant factor affecting the adsorption effect of lignite on Pb(II) cation. Adsorption process ΔH° >0, which suggests that the adsorption process is endothermic, so the adsorption reaction can be prompted when the temperature increases. Generally, the ΔH° of the chemical adsorption process is in the range of 20–400 kJ·mol−1. It can be seen from Table 4 that the adsorption enthalpy changes of RL and ML for Pb(II) cation are 11.04 kJ·mol−1 and 47.11 kJ·mol−1, respectively. The enthalpy changes are in this range or close to this range, indicating that the adsorption of Pb(II) cation by lignite is mainly chemical adsorption. The entropy changes of the reaction system during the adsorption process ΔS° >0 indicates that the disorder of the solid–liquid interface of the adsorption system increases during the adsorption process.

Adsorption coefficient

Based on the physical and chemical properties of lignite, the adsorption mechanism of Pb(II) cation on modified lignite was explored. In this paper, the relationship between specific surface area, total pore volume, Zeta potential and adsorption coefficient lnKd was mainly discussed, and the results are shown in Figure 6.

Figure 6

Relationship between adsorption coefficient and specific surface area (a), Total pore volume (b), |Zeta potential| (c) of lignite.

Figure 6

Relationship between adsorption coefficient and specific surface area (a), Total pore volume (b), |Zeta potential| (c) of lignite.

As shown in Figure 6(a) and 6(b), the specific surface area of ML is lower than that of RL. However, with the deepening of modification, the adsorption capacity of modified lignite for Pb(II) cation is increasing. It can be judged that the specific surface area and total pore volume of lignite are not the main influencing factors to improve the adsorption efficiency of Pb(II) cation. There is no strong correlation between the adsorption coefficient and the specific surface area and total pore volume of lignite, which fully shows that the pore filling effect has little contribution to the adsorption of Pb(II) cation in the process of modified lignite adsorption.

As shown in Figure 6(c), there was a positive correlation between Zeta potential and adsorption coefficient. The adsorption capacity of lignite increased with the increase of Zeta potential on the surface of modified lignite, which indicates that the chemical properties of lignite surface contributed greatly to the adsorption of Pb(II) cation. Futhermore, it can be seen from the figure that there is no good linear correlation between the Zeta potential of the adsorbent surface and the adsorption coefficient lnKd, which shows that the surface chemical properties of lignite are not the only factor affecting the adsorption amount of Pb(II) cation.

In order to further verify the role of coordination chelating and other chemical adsorption reactions in the adsorption of Pb(II) cation by lignite, sodium humate was used to simulate lignite molecules (Martyniuk & Więckowska 2003; Zhang et al. 2012). 1.0 mL Pb(II) cation solution with a concentration of 300 mg·L−1 was taken and added into 0.1 mg·L−1 sodium humate solution. The change of sodium humate solution is shown in Figure 7.

Figure 7

Comparison of before and after the addition of Pb(II) cation to sodium humate solution, initial (a) and final (b).

Figure 7

Comparison of before and after the addition of Pb(II) cation to sodium humate solution, initial (a) and final (b).

It can be clearly seen that after the addition of Pb(II) cation, a large number of flocculent precipitates are generated rapidly with the chelating reaction between humic acid molecules and Pb(II) cation. Through the above discussion, it can be considered that Pb(II) cation adsorption by modified lignite is mainly completed by chemical adsorption. Chemical adsorption mainly involves ion exchange, coordination and chelation between metal ions in aqueous solutions and surface functional groups of adsorbents.

(1) Coordination and chelation reactions

In the coordination and chelating reaction mechanism, humic acid reacts with heavy metal ions to form chelates (heavy metal ions compete with H+ at the coordination point of lignite surface), and the reaction equation is as follows.

This kind of reaction promoted the dissociation of H+. After the reaction, the concentration of H+ in the solution increased and the pH decreased. Under the condition of low pH, the dissociation of functional groups on the surface of lignite was inhibited due to the high concentration of H+, which made the chelating reaction difficult. When the pH of the solution increased, the coordination and chelating reactions were promoted, and the adsorption capacity of the adsorbent for Pb2+ increased.

(2) Ion-exchange mechanism

In the formation process environment, coal comes into contact with inorganic minerals such as Ca2+ and Na+, such that the coal surface active functional groups exist in the form of calcium and sodium salts. The main binding force is ionic bonding, so when this kind of salt and heavy metal ions meet, it is easy for the sodium and calcium to be replaced and go into solution. At the same time, when the concentration of metal ions in the solution is high, the metal ions also exchange with H+ on the surface of lignite to replace H+. The reaction is as follows.
formula

In addition, the dissociation of functional groups on the surface of lignite makes the surface of lignite adsorbent negatively charged, and there is a large electrostatic attraction between lignite adsorbent and positively charged metal ions, which provides a driving force for lignite to adsorb metal ions.

Interaction energies of the model molecule and Pb(II) cation

Through the above analysis, we can understand that Pb(II) cation is mainly chemisorbed on the surface of lignite. Through DMol3 calculation, the interaction energy ΔE between surface functional groups of lignite and Pb(II) cation can be calculated by the following equation.
formula
(8)
where E(A…B) refers to the energy of the entire system containing the model molecule and Pb(II) cation. E(A) and E(B) represent the energy of the model molecule and Pb(II) cation, respectively (Zhu et al. 2020). The result is shown in Table 5.
Table 5

Interaction energies of model molecule and Pb(II) cation

ComplexE(A…B) (Ha)E(A) (Ha)E(B) (Ha)ΔE (Kcal/mol)
Ben-COOH….Pb(II) cation −20,253.38 −420.49 −19,832.83 −37.54 
Ben-C = O….Pb(II) cation −20,178.16 −345.27 −19,832.83 −36.25 
Ben-C-O-C….Pb(II) cation −20,218.61 −385.72 −19,832.83 −41.49 
ComplexE(A…B) (Ha)E(A) (Ha)E(B) (Ha)ΔE (Kcal/mol)
Ben-COOH….Pb(II) cation −20,253.38 −420.49 −19,832.83 −37.54 
Ben-C = O….Pb(II) cation −20,178.16 −345.27 −19,832.83 −36.25 
Ben-C-O-C….Pb(II) cation −20,218.61 −385.72 −19,832.83 −41.49 

It can be seen from Table 5 that the ΔE of Ben-C-O-C….Pb(II) cation was the minimum, which suggests that Pb(II) cation is more easily adsorbed on Ben-C-O-C. Combined with the FTIR of raw and modified lignite, the surface modification increased the oxygen-containing functional groups on the surface of lignite, thus promoting the adsorption of Pb(II) cation, which further proved that the surface modification method can improve the adsorption capacity of lignite (Li et al. 2018). The equilibrium distance between model molecule and Pb(II) cation is shown in Figure 8.

Figure 8

Equilibrium distance between [Pb(H2O)4]2+ and Ben-COOH (a), Ben-C = O (b) and Ben-C-O-C (c), respectively. Some atoms are hidden.

Figure 8

Equilibrium distance between [Pb(H2O)4]2+ and Ben-COOH (a), Ben-C = O (b) and Ben-C-O-C (c), respectively. Some atoms are hidden.

It can be seen from Figure 8(a) and 8(b) that Ben-COOH and Pb(II) cation, Ben-C = O and Pb(II) cation are mainly connected by hydrogen bonds, and the adsorption equilibrium distances between H and O atoms are 1.462 Å and 1.395 Å, respectively. From Figure 8(c), the original hemidirected structure of Pb(II) cation is broken, and a polar covalent bond is formed between Ben-C-O-C and Pb(II) cation, and the equilibrium distance is 2.544 A. The bond energy of the polar covalent bond is much larger than that of the hydrogen bond energy, so the adsorption configuration between Ben-C-O-C and Pb(II) cation is more firm. It is further proved that the adsorption of Pb(II) cation on the surface of lignite is mainly completed by chemical adsorption, and functional groups -C-O-C and Pb(II) cation are most easily combined, followed by-COOH, while -C = O is the weakest.

In this paper, lignite was used as adsorbent to remove the heavy metal ion Pb(II) cation in wastewater, and the adsorption process of Pb(II) cation on the surface of lignite was studied through adsorption kinetics, adsorption thermodynamics, adsorption coefficient variation and DFT calculation. The main conclusions are as follows.

  1. The adsorption capacity of lignite was improved by HNO3 surface modification, and the surface functional groups such as -COOH, -C = O and -C-O-C were increased because of surface oxidation of lignite. The adsorption of Pb(II) cation on lignite surface conforms to pseudo-second-order kinetic equation and Langmuir isothermal adsorption model. When the adsorption time was 120 min, the adsorption process basically reached equilibrium. The adsorption process of Pb(II) cation on the surface of lignite is endothermic and spontaneous, and the increase of temperature is conducive to improving the adsorption capacity of lignite.

  2. Sodium humate was used to simulate lignite molecules, and it can be clearly seen that after the addition of Pb(II) cation, the chelating reaction between humic acid molecules and Pb(II) cation rapidly generates a large number of flocculent precipitates. The result can be considered that Pb(II) cation adsorption by modified lignite is mainly completed by chemical adsorption and mainly involves ion exchange, coordination and chelation between metal ions and adsorbent.

  3. Interaction energies between Ben-COOH, Ben-C = O, Ben-C-O-C and Pb(II) cation are −37.54, −36.25, −41.49 kcal/mol, there are mostly hydrogen bonds between Ben-COOH, Ben-C = O and Pb(II) cation, and between Ben-C-O-C and Pb(II) cation, a polar covalent bond is formed. Pb(II) cation on the surface of lignite is mainly completed by chemical adsorption; there are lots of functional groups in the surface of raw lignite, and surface modification with HNO3 can increase the oxygen-containing functional groups of lignite and improve its adsorption capacity.

This work was supported by the National Natural Science Foundation of China (No. 51974324), the Fundamental Research Funds for the Central Universities (No.2020QN08) and the Fundamental Research Funds for the Central Universities (No. 2020YQHH02).

Data curation, P. Wang; Formal analysis, P. Wang; Investigation, P. Wang, W. Liu, and H. Xu; Methodology, W. Liu and H. Xu; Project administration, W. Liu and H. Xu; Resources, W. Liu and H. Xu; Software, P. Wang; Writing – original draft, P. Wang; Writing – review & editing, P. Wang.

The authors declare no conflict of interest.

All relevant data are included in the paper or its Supplementary Information.

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