Dynamic experiments of acid mine drainage with Rhodopseudomonas spheroides activated lignite immobilized sulfate-reducing bacteria particles treatment

Experimental materials

Preparation of immobilized particles

According to the group's previous study²³, SRB-LP and R-SRB-LP were prepared according to the optimal ratio (mass fraction of 3% lignite, 10% SRB) and (mass fraction of 3% lignite, 10% SRB, 10% R. spheroides) respectively. Firstly, 9% polyvinyl alcohol and 0.5% sodium alginate were dissolved in distilled water, kept under seal at room temperature for 24 h until sufficiently swollen and then stirred continuously in a constant temperature water bath (Hangzhou Hengyi instrument technology effective company, model HH-4) at 90 °C until no bubbles were produced. The lignite was then slowly added to the gel at a water bath temperature of 40 °C and fully stirred to uniform without bubbles. And then, the above materials were taken out and sealed in the room, and when cooled to 37 ± 1 °C, add the bacterial solution (SRB-LP with SRB concentrated solution only, R-SRB-LP with R. spheroides and SRB concentrated solution) and stirred well until no air bubbles. Then, the gel mixture was aspirated with a peristaltic pump (Shanghai Kachuaner Fluid Technology Company, model STP-F01A) and directly dropped into a 2% CaCl₂ saturated boric acid solution at pH 6. The crosslinking was carried out with a magnetic stirrer (Jintan District Xicheng Xinrui instruments, model JJ-1A) at a stirring rate of 100 rpm and the immobilized particles were taken out after 4 h, the particle size is about 0.5 cm. Rinsed with 0.90% physiological saline, and then the surface water was aspirated, and this operation was repeated three times to form SRB-LP and R-SRB-LP, respectively. The particles were activated with a modified Starkey-type medium without organic components for 12 h in an anaerobic environment before use.

Dynamic experiments

The dynamic test device is shown in Fig. 1 and consists of two sets of dynamic columns with a diameter of 54 mm and a height of 500 mm.

The upper and lower layers were placed with 50 mm quartz sand with a particle size of 3–5 mm, the middle layer of column 1# was 250 mm SRB-LP, and the middle layer of column 2# was 250 mm R-SRB-LP. Column 1# was set as the blank group and column 2# as the experimental group. The inlet valve was opened and the AMD was continuously supplied from bottom to top at a flow rate of 0.398 mL/min. Three groups of experimental devices were set up, and the temperature was room temperature (25 °C) for 30 days. The water samples of influent water and effluent water from two dynamic columns were taken at 8am every day, and the water samples were filtered through a 0.45 μm filter membrane, then measured the concentrations of SO₄²⁻, Zn²⁺, and Cu²⁺, pH, COD release and ORP value. The solution pH was measured by the glass electrode method (GB 6920-86), the concentration of SO₄^2– was measured by the barium chromate spectrophotometric method (HJ/T 342-2007), the concentrations of Cu²⁺, Zn²⁺ were measured by flame atomic spectrophotometer method (GB/T 9723-2007), the ORP value was determined by the glycol electrode method (GB/T 9354-1999), and the COD was measured by the rapid extinction spectrophotometry method (HJ/T 399-2007). The removal efficiency was calculated using w = (C₀ − C₁)/C₀ × 100%, where w represents the pollutant removal rate, %. C₀ represents the initial pollutant concentration, mg/L. C₁ represents the pollutant concentration after treatment, mg/L.

SO₄²⁻ reduction kinetics and Zn²⁺ adsorption kinetics

The SO₄²⁻ reduction kinetics and Zn²⁺ adsorption kinetics were studied.

SO₄²⁻ reduction kinetics experiment: 816 mg/L sulfate wastewater with pH of 4 was prepared. R-SRB-LP was added to the wastewater at the ratio of solid to liquid of 1:10, and continuous oscillation was carried out in a shaker at 30 °C and 150 r/min. The concentration of SO₄²⁻ was measured at 1, 3, 5, 7, 9, 12, 18, 21, 24, 30, 36, 48, 72, 96, 120 h.

Zn²⁺ adsorption kinetics experiment: 20 mg/L Zn²⁺ with pH of 4 was prepared. R-SRB-LP was added to the wastewater at the ratio of solid to liquid of 1:10, and continuous oscillation was carried out in a shaker at 30 °C and 150 r/min. The concentration of Zn²⁺ was measured at 1, 3, 5, 7, 9, 12, 18, 21, 24, 30, 36, 48, 72, 96, 120 h.

Reaction product characteristics

To explore the mechanism of AMD repaired by immobilized particles, the morphology of particles before and after reaction was characterized by JSM 7200F scanning electron microscope (SEM, produced in Japan), the surface functional groups of the particles before and after the reaction were analyzed by Niligao IN10 Fourier transform infrared spectrometer (FTIR, produced in USA), and the changes of the specific surface area of the particles before and after the reaction were analyzed by a Mac ASAP2460 analyzer (produced in USA) using N₂ adsorption and desorption.

ORP value analysis

Figure 2 provides the changes of Oxidation Reduction Potential (ORP) values during the reaction. ORP values could initially reflect the strength of SRB activity. It was reported that SRB grows vigorously when the ORP value is from – 50 to − 300 mV²⁴. The average ORP in column 1# was 246 mV. In the first stage, the ORP in column 1# was in the range of 200.78–174.71 mV, manifesting that SRB activity was low. SRB is a heterotrophic organism, and its metabolism needs the support of carbon source. At this stage, pH was suitable for SRB growth, the concentration of metal ions did not pose a threat to SRB activity, but the SRB activity was low, indicating that the organic matter in column 1# could not provide a large number of SRB growth, and the limited organic matter might come from the bacterial solution added in the preparation of particles. In the second and third stages, ORP increased from 288.33 mV to 299.73 mV, showing that the organic matter in the bacterial solution was fully utilized, and SRB was almost completely inactivated.

Figure 3 shows that the curve of column 2# was below 1# and the ORP was in the range of − 50 mV to − 300 mV from day 5 to 19, which indicates that SRB growth in column 2# was better than that in column 1#. Under the condition of not adding any carbon source, column 1# maintained the activity of SRB for only 5 days with poor activity due to the lack of continuous carbon source supply. However, SRB in column 2# remained active and had relatively high activity. This proves that R. spheroids could degrade lignite to form small molecular organic matter and provide carbon source for SRB. In the first stage, ORP decreased from 187.16 mV to − 98.83 mV. It has been reported that SRB activity is related to pH, which is most suitable for SRB growth at 6–7.50²⁵. At this stage, the pH of the solution gradually increased to this range, and the carbon source was sufficient, so the growth of SRB was getting better and better. In the second stage, ORP increased from − 149.14  to − 15.18 mV. At this stage, the carbon source was sufficient and the pH was suitable for SRB growth, but the activity of SRB decreased because of the inhibition of Cu²⁺ and Zn²⁺ on SRB. In the third stage, ORP increased from 49.03 mV to 278.60 mV. In this stage, the rate of SRB deactivation was accelerated due to low pH, toxicity of high concentrations of Cu²⁺ and Zn²⁺and shortage of carbon source.

COD release analysis

Figure 3 shows the change of COD content in the effluent of columns 1# and 2#. SRB consumes COD in the process of SO₄²⁻ reduction, Liu and Vossoughi et al.^26,27 showed that the reduction of 1 g SO₄²⁻ by SRB required 0.67 g COD. The average effluent COD from column 1# was 428.20 mg/L. The effluent COD increased from 423.62 to 646.72 mg/L in the first stage. Most of these COD came from the release of lignite²⁸. From the 1# column ORP, it can be seen that the activity of SRB is poor due to the lack of carbon source at this stage, indicating that SRB cannot effectively utilize the organic matter released by lignite. In the second and the third stage, the effluent COD decreased from 472.99 to 54.01 mg/L. The capacity of lignite to release organic matter was limited, the leaching of COD was less and less, so the effluent COD declined.

The average effluent COD of column 2# was 428.20 mg/L. Kunlanit et al.²⁹ indicated that COD was released when biomass materials were decomposed by anaerobic fermentation microorganisms, so the release of COD in column 2# was mainly attributed to the metabolic degradation of lignite by R. spheroides. In the first stage, the effluent COD of column 1# increased from 160.23 to 383.58 mg/L. R. spheroides was able to decompose lignite to provide carbon source for SRB growth. According to ORP, SRB growth gradually reached a vigorous state at this stage, and consumed more and more carbon source. While the effluent COD was growing at this time, which could indicate that the activity of R. spheroides was getting better and better. In the second stage, the effluent COD increased from 516.42 to 669.71 mg/L. From the 15th day, the effluent COD tended to be stable and gradually stabilized at about 680 mg/L. At this stage SRB was in a state of inhibition by metal ions, the carbon source consumed would be reduced. The COD remained flat indicating the ability to produce carbon source decreased in the system from day 18. This may be owing to diminished activity of R. spheroids. Quan³⁰ et al. found that metal ions could inhibit the activity of R. spheroides. Due to the reduced ability of lignite to adsorb metal ions at this stage, metal ions accumulated in the system and caused toxic effects on R. spheroids. Although the ability of the system to produce carbon decreased from day 18, it was clear from curve 2# that sufficient carbon was produced to support the growth of SRB. In the third stage, the COD decreased from 358.03 to 18.25 mg/L. At this stage, the adsorption of heavy metals by lignite was saturated, and the accumulation of metal ions in the system increased, resulting in the gradual inactivation of R. spheroides and the reduction of carbon source production. Since the growth of SRB requires carbon source, the lack of a carbon source in the third stage was one of the factors contributing to SRB death.

Analysis of pH enhancement effect

Figure 4 shows the effect of columns 1# and 2# on pH enhancement. The average pH of the effluent from column 1# was 5.88. In the first stage, the pH was in the range of 6.39–6.63. It can be seen from ORP and COD that SRB in column 1 # had almost no activity due to the support of no carbon source. Therefore, the slow increase of pH mainly depended on lignite, indicating that lignite has the ability to regulate pH, which is consistent with the results of Dong³¹. Lignite can release alkalinity to combine with H⁺ in solution to raise pH. In the second and third stage, the effluent pH decreased from 6.39 to 4.14. The decrease in pH was mainly due to the fact that the negative charge released by the lignite tended to saturate with the combination of H⁺, and the ability to regulate pH became weaker and weaker.

The average effluent pH of column 2# was 6.4 and the highest pH was 7.60. The effluent pH in the first stage was in the range of 7.24–7.60. The effluent pH of column 2# was higher than that of column 1 # because H⁺ was not only neutralized by alkaline substances released by lignite, due to the degradation of the lignite by R. spheroides continuously provided carbon source for the SRB and kept its activity, more H⁺ was consumed by the SRB as electron. The existence of a large number of active oxygen-containing functional groups on the lignite surface makes lignite negatively charged. The combination of H⁺ and negatively charged can improve the pH of the solution. From the 1 # column, the pH could be raised to about 6.50 only by lignite adsorption in the first stage, this range of pH was suitable for SRB survival, so lignite could remove the inhibition of acid on SRB and enhance SRB activity, which was consistent with ORP. At this stage, the adsorption capacity of lignite was strong, the SRB activity was enhanced and the consumption of H⁺ was increased, so the effect of pH improvement was good.

In the second stage, the effluent pH decreased from 7.11 to 6.20. At this stage, due to the limited negative charge contained in lignite, the removal effect of H⁺ decreased. However, from the 1# column, it can be seen that the pH could be increased to about 6 by lignite adsorption. It was in the suitable growth range of SRB, the existence of lignite could also maintain the stability of pH system. According to ORP and SO₄²⁻, the activity of SRB showed a downward trend, which was caused by the inhibition of metal ions on SRB. As SRB activity decreased, less and less H⁺ could be consumed as electrons by SRB, and the alkaline released by lignite was gradually neutralized by H⁺, the effect of pH improvement was getting worse and worse from the second stage.

In the third stage, the effluent pH decreased from 5.61 to 4.35. As the neutralisation of H⁺ by lignite at this stage tended to saturate, it gradually lost its ability to raise pH, making the pH system and its instability. Under acidic conditions (pH < 5), high concentrations of protons would lead to high diffusion pressure in the cell membrane of microorganisms³², more energy is used to pull protons out of the membrane through proton motive force to maintain the internal pH value, less energy is used for cell growth³³, the pH at this stage affected the SRB activity. Due to the reduced of H⁺ neutralization capacity of lignite and the reduced of electron consumption caused by SRB inactivation, the pH in the system decreased.

Analysis of SO₄²⁻ removal effect

Figure 5 shows the effect of SO₄²⁻ removal by 1# and 2# dynamic columns. Our previous experiment²⁸ showed that lignite had no ability to remove SO₄²⁻, so the removal of SO₄²⁻ by the system mainly depended on the reduction of SRB. SO₄²⁻ can also react to the activity of SRB. The average SO₄²⁻ removal rate of column 1# was 3.61%. In the first stage, the removal rate of SO₄²⁻ decreased from 21.39% to 0. Due to the absence of the carbon source in column 1#, the SRB only remained active for a few days and could last only a few days, so the SO₄²⁻ reduced by SRB through anaerobic reduction was less. The removal of SO₄²⁻ in the second and third stages was 0, which would also indicate that the SRB was deactivated at this stage, in line with the findings of the ORP value stage.

The average removal rate of SO₄²⁻ in column 2# was 59.08% and the highest removal rate was 93.97%. In the first stage, the SO₄²⁻ removal rate increased from 68.79 to 93.97%. At this stage, as the environment of the reaction system gradually reached the SRB comfort conditions, the SRB activity gradually raised, so the SO₄²⁻ removal rate gradually increased. In the second stage, the removal rate of SO₄²⁻ decreased from 87.23 to 54.61%. Although the activity of the SRB remained good at this stage, due to the inhibition of metal ions, the activity of SRB was affected, so the removal rate of SO₄²⁻ decreased. Moreover, metal sulfides generated by metal ions and S²⁻ were adsorbed on the surface by the lignite, which would encrust microorganism, restricting mass transfer between reactants and microbial enzymes, and ultimately making the removal of sulfate less effective³⁴. In the third stage, the removal rate of SO₄²⁻ decreased from 9.93 to 4.26%. Owing to the lack of carbon source in the system, pH decreased and metal ions accumulated had toxic effects on SRB, SRB gradually lost its activity, so the removal rate of SO₄²⁻ was low.

Analysis of Cu^{2 +}, Zn^{2 +} removal effects

Figure 6 shows the effect of columns 1# and 2# on the removal of Cu²⁺ and Zn²⁺. The average removal rates of Cu²⁺ and Zn²⁺ by column 1# were 58.87% and 47.05% respectively. In the first stage, the removal rate of Cu²⁺ by column 1# was 94.80%–89.59% and the removal rate of Zn²⁺ was 75.76%–72.49%. The removal of metal ions at this stage mainly depended on the adsorption of lignite, which has a large number of aerobic functional groups on its surface, and these functional groups can chelate with metal ions to adsorb them on the surface³⁵. Moreover, according to K_spCu(OH)₂ = 2.2 × 10^–20 and K_spZn(OH)₂ = 1.2 × 10^–17, some of the metal ions were removed in the form of hydroxide as the pH increased³⁶. In the third stage, the removal efficiency of Cu²⁺ declined from 83.64 to 10.78% and that of Zn²⁺ from 56.13 to 13.75%. In the middle and late stages of the reaction, the adsorption sites on the lignite surface gradually reduced until the adsorption saturation, and the removal effect of Cu²⁺ and Zn²⁺ continued to decrease.

The average removal rates of Cu²⁺ and Zn²⁺ by column 2# were 69.76% and 67.91%, respectively, and the highest removal rates were 98.52% and 97.03%. In the first stage, the removal efficiency of Cu²⁺ was 96.28–98.52%, and the removal efficiency of Zn²⁺ was 89.22–94.42%. The removal of Cu²⁺ and Zn²⁺ was stronger in column 2# than in column 1# because the biological removal of metal ions was present in column 2# in addition to the adsorption of lignite. R. spheroides can destroyed the structure of lignite and increased the adsorption sites of lignite to adsorb more metal ions³⁷. Li et al.³⁸ found that the effluent S²⁻ was less than 1 mg/L at pH above 6.5 during the effect of pH on S²⁻ in the AMD treatment test, indicating that the precipitation of sulfate ions with metal ions was easy at higher pH conditions. According to the pH, in the first stage, the pH could be increased to about 7.3 by the synergistic effect of lignite and SRB. Therefore, Zn²⁺ and Cu²⁺ in the system could be combined with reduced S²⁻ and removed in the form of CuS and ZnS³⁹. At the same time, SRB could adhere to the surface of lignite to form biofilm, and a large number of extracellular polymers (EPS) were generated. Beech et al.⁴⁰ found that bacteria themselves promote the binding of metals to S²⁻ when there was less metal sulfide in solution, cell walls and extracellular polymer (EPS) were able to bind metals thus accelerating solid formation. Therefore, the removal of Cu²⁺ and Zn²⁺ by column 2# at this stage was effective.

In the second stage, the removal efficiency of column 2# decreased from 92.57% to 55.39% for Cu²⁺ and from 83.64% to 55.39% for Zn²⁺. Kieu et al.⁴¹ indicated that 6.0–7.5 mg/L heavy metal ion concentration could inhibit the growth of SRB; above 7.5 mg/L heavy metal ion concentration was the toxic concentration that could lead to SRB death. In this stage, due to the limited adsorption sites on the surface of lignite, the adsorption properties of lignite for Cu²⁺ and Zn²⁺ gradually decreased, and the metal ions gradually accumulated. Although the carbon source was sufficient at this stage, the pH could be increased to about 6.50 under the combined action of lignite and SRB, which would not have a great impact on the activity of SRB and the formation of metal sulfate. The activity of SRB was reduced by the inhibition and poison of metal ions. The adsorption of metal ions by lignite decreased, and at the same time, the weakened activity of SRB resulted the reduced S^2- reduced, the generated metal sulfides declined, so the removal efficiency of Cu²⁺ and Zn²⁺ decreased.

In the third stage, the removal efficiency of Cu²⁺ decreased from 23.42% to 10.78% and the removal efficiency of Zn²⁺ decreased from 42.75% to 28.62%. At this stage, the adsorption performance of lignite tended to be saturated, and the adsorbed metal ions decreased. The accumulation of metal ions had a more serious toxic effect on SRB. So, this stage of SRB due to pH and metal toxicity, lack of carbon source, SRB activity decreased, S^2- reduced. At the same time, it can be seen from pH that lignite cannot maintain the stability of pH system at this stage, which affected the formation of metal sulfides. So, the removal efficiency of Cu²⁺ and Zn²⁺ by 2# column was low.

Reduction kinetics of SO₄²⁻ and adsorption kinetics of Zn²⁺

Kinetic studies are crucial for determining the dynamic reduction (adsorption) rate and controlling the reduction (adsorption) mechanism. Zero-order kinetics, first-order kinetics, pseudo-first-order kinetics and pseudo-second-order kinetics were used to simulate the relationship between reduction (adsorption) rate and time of SO₄^2-, and Zn²⁺, respectively.

In the formula, C_t represents the concentration of SO₄²⁻ at time t (h), mg/L. C₀ represents initial SO₄²⁻ concentration, mg/L. k₀ represents zero-order reaction rate constant, mg·L⁻¹·h⁻¹. k₁ represents the first-order-reaction rate constant, h⁻¹. q_e is the adsorption amount at adsorption equilibrium, mg g⁻¹. q_t is the adsorption amount when the adsorption time is time t (min), mg g⁻¹. k₃ is the pseudo-first-order kinetic reaction rate constant, min⁻¹. k₄ is the pseudo-second-order kinetic reaction rate constant, g (mg⁻¹ min⁻¹).

The kinetic fitting diagram is shown in Fig. 7. (a) As the zero-order kinetic fitting of SO₄^2- reduction. (b) As the first-order kinetic fitting of SO₄^2-. (c) As the pseudo-first-order kinetic fitting of Zn²⁺, and (d) as the pseudo-second-order kinetic fitting of Zn²⁺. According to the linear fitting results, the dynamic parameters and the correlation coefficient of the evaluation are shown in Table 1.

Obviously, in the reduction kinetics of SO₄²⁻, R² (0.991) of the first-order kinetics was greater than R² (0.941) of the zero-order kinetics, indicating that the first-order reaction kinetics model was more suitable for the reduction process of SO₄²⁻ by SRB in R-SRB-PL. The reduction of SO₄²⁻ is mainly affected by electron acceptors, and the main removal process of SO₄²⁻ in wastewater is SRB alienation reduction. In the adsorption kinetics of Zn²⁺, the pseudo-second-order kinetic R² is greater than the pseudo-first-order kinetic R². It is indicated that pseudo-second-order kinetics is more suitable for the adsorption process of metal ions by R-SRB-PL, and electron sharing or electron transfer may exist in the adsorption process of metal ions by R-SRB-PL.

SEM characterization analysis

The results of the SEM are shown in Fig. 8. At the end of the dynamic experiment, SRB-LP and R-SRB-LP were dried, and the changes of particle morphology before and after the reaction were explored by SEM scanning. There was no significant change in the surface state of lignite SRB-LP, indicating that SRB alone could not degrade lignite. After R-SRB-LP reaction, a large number of cracks and pores appeared on the surface of lignite, and the organic matter that could not be degraded by R. spheroides was precipitated on the surface of lignite. These results illustrated that R. spheroides could destroy the structure of lignite, and enhanced the adsorption capacity.

FTIR analysis

In order to further study the changes of surface functional groups of lignite in SRB-LP and R-SRB-LP, the SRB-LP particles and R-SRB-LP particles before and after the reaction were detected by FTIR, and the results are shown in Fig. 9. The peak at 3414.87 cm⁻¹ was attributed to the stretching vibration of enol-OH or N–H^42,43. The position of 2918.62 cm⁻¹ belonged to the antisymmetric stretching vibration of alkanes C–H. The peak at 1615.16 cm⁻¹ was benzene C=C stretching vibration⁴⁴. The position peaks of 1433 cm⁻¹ and 1340 cm⁻¹ were –CH₃ asymmetry peak, CH₂ planar bending peak or benzene skeleton vibration peak⁴⁵. 1112.66 cm⁻¹ position is assigned to the saturated C–O stretching vibration peak in tertiary alcohols. Peaks at 829.28 cm⁻¹, 663.21 cm⁻¹ represented the stretching vibrations of C–H substituents in the benzene ring structure.

It can be seen from Fig. 9a that in SRB-LP, the peak position only changed slightly, indicating that the structure of lignite in SRB-LP was not destroyed in a wide range. The reduction of peak positions around 3414.87 cm⁻¹, 2918.62 cm⁻¹, 1615.16 cm⁻¹, 1112.66 cm⁻¹, 829.28 cm⁻¹, 663.21 cm⁻¹ may be due to the replacement of some functional groups in the lignite by metal ions with H⁺ throughout the process, resulting in a reduction of some functional groups and leading to a reduction in peak positions.

It can be seen from Fig. 9b that in R-SRB-LP, the peak position changed greatly, not only the peak decreased but also the peak disappeared, which indicated that the lignite structure in R-SRB-LP was destroyed. The peak intensity at 3414.87 cm⁻¹ and 1615.16 cm⁻¹ decreased, indicating that during the degradation of lignite by R. spheroides, the structure of-OH and benzene ring in lignite structure was destroyed. The decrease in the vicinity of 2918.62 cm⁻¹ indicated that R. spheroides could break the C–C bond in the molecular structure of lignite, thus breaking the side chains of aromatic rings, some alkanes and olefins in the lignite structure. The disappearance of peaks at 829.28 cm⁻¹, 663.21 cm⁻¹ was attributed to the substitution effect of metal cations.

Due to the absence of R. spheroides in SRB-LP, there was little difference in the peak position before and after the reaction, while there was R. spheroides in R-SRB-LP, the difference of peak position before and after reaction was large and the peak disappeared. This shows that R. spheroides can decompose lignite and destroy lignite structure.

Specific surface area analysis

Specific surface areas of SRB-LP and R-SRB-LP are compared as shown in Table 2. As can be seen from Table 1, the BET specific surface area and Langmuir specific surface area did not change much after the SRB-LP reaction. However, the BET specific surface area of R-SRB-LP after reaction was 1.58 times that before the reaction and the Langmuir specific surface was 1.53 times that before the reaction. This indicated that R. spheroides changed the surface structure of lignite. Some functional groups were dissolved, thereby increasing the number of voids on the surface of lignite. The specific surface area increased, and the adsorption performance of lignite was improved.

Figure 10 shows the adsorption amount of R-SRB-LP versus relative pressure obtained from the test results. This adsorption equilibrium isotherm was consistent with the type II isotherm, which indicated that the adsorption process of lignite on metal ions was a free multilayer reversible adsorption process. The steep points of this isotherm are between 0.05 and 0.10, indicating the saturated adsorption amount of single molecule, which was equivalent to the completion of single molecule layer adsorption. The adsorption curve of R-SRB-LP particles after reaction was always above before reaction, which also proved that the adsorption performance of R-SRB-LP particles after reaction was better than that before reaction.