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Article

Effects of Pyrolysis Temperature on Biochar Physicochemical and Microbial Properties for H2S Removal from Biogas

by
Rasa Vaiškūnaitė
,
Aušra Mažeikienė
* and
Kamyab Mohammadi
Department of Environmental Protection and Water Engineering, Faculty of Environmental Engineering, Vilnius Gediminas Technical University, 10223 Vilnius, Lithuania
*
Author to whom correspondence should be addressed.
Sustainability 2024, 16(13), 5424; https://doi.org/10.3390/su16135424
Submission received: 22 May 2024 / Revised: 20 June 2024 / Accepted: 22 June 2024 / Published: 26 June 2024
(This article belongs to the Section Waste and Recycling)
Browse Figures
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Abstract

:
Sludge is produced in sewage treatment plants and is still a problematic waste type after anaerobic digestion. A sustainable sludge management strategy would be to pyrolyze it and obtain biochar suitable for use in biofilters. This article examines the physical and chemical properties of biochar obtained by pyrolyzing sewage sludge at a temperature of 300–600 °C. The pyrolyzed sludge was used in the biofilter as a filler. The results demonstrated biochar packing materials after pyrolysis at 300 °C, 400 °C, 500 °C, and 600 °C, which exhibited porosities of 35%, 42%, 67%, and 75%, respectively. During the research study, it was established that the biofilter showed excellent efficiency (between 55 and 99 percent) when using carbon pyrolyzed at temperatures of 500 °C and 600 °C. In this study, the average growth rates of the number of sulfur-oxidizing microorganisms were 1.55 × 104 CFU/g at the first stage of the biofilter, 2.63 × 104 CFU/g at the second stage, 3.65 × 104 CFU/g at the third stage, 5.73 × 104 CFU/g at the fourth stage, and 2.62 × 104 CFU/g at the fifth stage. The number of sulfur-oxidizing microorganisms in the packing bed of biofilters during the 60-day period of the experiment constantly increased. The experimental results of H2S purification in biogas were compared with mathematical modeling results. These comparative results revealed a consistent trend: the model-estimated filter efficiency also reached 70–90 percent after 60 days of investigation.

1. Introduction

Sewage sludge is waste from wastewater treatment, and it is produced worldwide. Since the implementation of the Urban Wastewater Treatment Directive (1991) in Europe, an increasing amount of sludge has been produced in sewage treatment plants [1,2,3]. Municipal sludge, created during the treatment of wastewater, is a complex mixture of water, proteins, lipids, carbohydrates, nucleic acids, phenols, ash, various pathogens, heavy metals, polychlorinated biphenyls, dioxins, and other harmful substances. Current regulations limit the direct discharge of sludge, requiring the exploration of alternative disposal methods such as landfills, composting, and incineration. Landfill and composting can cause secondary pollution; thus, the thermochemical treatment of sludge, such as pyrolysis, gasification, or incineration, is a viable alternative. Pyrolysis, for example, is a cost-effective process that can simultaneously produce solid, liquid, and gaseous products. Additionally, it eliminates organic pollutants and pathogens, thereby decreasing waste.
H2S is a highly dangerous contaminant that can cause material spoilage and promote biogenic corrosion through the action of sulfur-oxidizing microorganisms depending on environmental conditions. When H2S oxidizes, it produces acidic sulfur dioxide (SO2), which can severely corrode metal surfaces. The corrosion process is hastened at high temperatures in a watery environment when metals interact with H2S. This leads to the creation of metal sulfide corrosion products that result in significant damage. When H2S dissolves in water, it creates an aqueous solution of H2S, which is a mild acid that dissociates in two stages. H2S attacks iron, forming an unstable FeS layer. In acidic conditions, FeS is removed from the surface, and H2S is reformed, thus intensifying corrosion.
H2S is removed from biogas in the biofilter during the transfer process from the gaseous to liquid phase, followed by diffusion into the biofilm on an inert packing material and biological degradation of the pollutants [4,5,6,7,8,9,10,11,12,13,14,15,16,17].
Biochar is a charcoal and carbon-rich material produced by partial oxidation (pyrolysis at ≤700 °C in the absence or limited supply of oxygen) of carbonaceous organic sources such as wood and plants, wastewater sludge, etc. The properties of biochar are influenced by various technological factors, particularly the pyrolysis temperature and the type of feedstock. These factors impact the pH, specific surface area, pore volume, cation exchange capacity, volatile matter, ash, and carbon content of the biochar. Higher pyrolysis temperatures result in the more efficient production of biochar, which has increased specific surface area, greater porosity, and higher pH and ash and carbon contents but decreased cation exchange capacity and volatile matter content. This is likely due to the high degree of decomposition of organic matter. The differences in the content of lignin and cellulose and the moisture content of the biomass are the reasons for these variations. The suitability of biochar as a sorbent for cleaning pollutants depends on its physical and chemical properties. For example, another advantage of pyrolysis is the production of biochar, a porous solid product that can be used for agricultural soil cleaning, heavy metal absorption, and pollutant removal [18,19,20,21,22]. For example, other researchers are currently studying the perspectives of converting biomass-derived polymers to functional biochar materials via pyrolysis. This is a mature and promising approach as these products can be widely utilized in many areas, such as carbon sequestration, power production, environmental remediation, and energy storage [23]. Biochar can be used for other purposes as well, for example, scientists have carried out very useful research that aimed to prepare biochar-derived carbocatalysts (BDCs) from three feedstocks (seaweed, microalga, and lignocellulosic biomass) and further evaluate the environmental impacts through life cycle assessment (LCA) methodology [24]. Other researchers have conducted studies involving the cultivation of Chlorella protothecoides for the adsorption of tetracycline and electrochemical studies on self-cultured Chlorella protothecoides. In this study, the use of biochar and Chlorella protothecoides to jointly adsorb tetracycline is of great significance for environmental protection and microalgae cultivation [25].
A biofiltration system using sulfur-oxidizing bacteria immobilized on granular biochar as packaging materials showed high efficiency in cleaning pollutants. During tests, even at high concentrations (200–4000 ppm), the system achieved over 98% H2S removal. The biofilter demonstrated excellent performance during operation. At initial pollutant concentrations of less than 350 mg/m3 and with a gas flow rate of 0.25–0.35 m3/h, the H2S removal efficiency reached almost 100%. The results suggest that a variety of microbial communities have the potential to efficiently eliminate H2S. Bacteria such as Rhodanobacter, Halothiobacillus, Mizugakiibacter, and Thiobacillus played an important role in the research on H2S removal [26,27,28,29,30,31]. When biogas contained 50–300 ppm hydrogen sulfide, a packed bed biofilter using wood pellets and supported by microbial cultures such as Thiobacillus thioparus, Thiobacillus thioxidans, and Thiobacillus novellus was utilized to remove H2S from the biogas. The biogas was passed through the biofilter column at various flow rates (0.003, 0.006, 0.075, 0.012, and 0.015 m3/h), with the spraying of liquid medium containing different microbial cultures (flow rate 12 l/h) at different temperatures (25, 30, 35, and 40 °C). It was observed that at low flow rates, between 0.003 and 0.006 m3/h, each of the mentioned microbial cultures achieved 100% pollutant degradation efficiency [29,30,31].
This study aims to assess the effectiveness of biochar derived from sewage sludge at various pyrolysis temperatures as adsorbents for removing hydrogen sulfide from biogas. It involves analyzing the key physical and chemical properties of selected packing materials. Specific experimental investigations focus on sulfur accumulation and its correlation with the outlet H2S concentration, as well as bacterial growth rates on these packing materials and their impact on the outlet hydrogen sulfide concentration in biogas at different biochar pyrolysis temperatures. Finally, the experimental results are compared to those obtained from the mathematical modeling of the outlet H2S concentration.

2. Materials and Methods

2.1. Completion of the Sewage Sludge Pyrolysis Process

In this research, samples sourced from the Vilnius Sewage Sludge Treatment Plant underwent a slow pyrolysis procedure at temperatures of 300 °C, 400 °C, 500 °C, and 600 °C for 7 h within a 300 L pyrolysis apparatus under an oxygen-free environment. Following the pyrolysis, the apparatus was naturally allowed to cool to room temperature for over 4 h. The heating rate to reach the designated pyrolysis temperatures was 400 °C per hour, with an 8 h holding time at the final temperature. Subsequently, the samples were cooled to 20 °C using nitrogen gas ventilation. The resulting biochar was mechanically ground and then sifted through a 1.0 cm mesh sieve.

2.2. The Determination of the Significant Physicochemical Properties of the Biofilter Packing Material

After the pyrolysis process, the sewage sludge was used as packing material in the biofilter. However, before that, the significant physical properties of packaging materials were determined. In this study, the most significant physical properties such as bulk density, porosity, specific surface area, ash content, moisture content, and volatile compound content, and chemical properties including chemical composition, pH, and electrical conductivity, will be assessed initially to identify and compare their differences.
In bulk density of the packing material (based on the ASTM D6683-19 standard procedure), a 100 mL glass cylinder is filled with crushed biochar and dried for eight hours at 80 °C in a drying cabinet to determine the density [32]. After that, the cylinder is shaken for one minute to compress the dried samples and fill all existing spaces in the cylinder.
The t-plot strategy is a significant method that was designed based on the New Source Performance Standards (NSPSs) and permits deciding the miniature and mesoporous volumes and the particular surface region of an example. The porosity of the biofilter packing material was determined using this method.
To evaluate pH (based on the ASTM D1293-18 standard test method), 5 g of each packing material at various conditions was taken and then blended with 100 mL of deionized water [33]. After leaving aside all mixtures for 5 min to obtain a good solution, all samples’ mixtures were placed in a shaker and left to integrate/shake at approximately 100 rpm (rotation per minute) for one hour. From that point onward, all samples were sifted utilizing < 0.55 μm paper channels.
In the case of electrical conductivity (based on the ASTM D1125-23 standard test method), samples that were sifted in the previous experiment can be analyzed for these physical properties [34]. The measurement was carried out at a room temperature of 24 °C.
In terms of nitrogen content existing in biochar samples, the Kjeldahl method (based on the ASTM-D 5231-92 standard procedure) is implemented [35]. The samples are digested via sulfuric acid, and nitrogen is converted into ammonium sulfate. Then, the nitrogen in ammonium sulfate is released as ammonia and converted to ammonium borate by boric acid and using sulfuric acid 0.1. Later, the amount of nitrogen is calculated by consumed acid. To evaluate the organic carbon content existing in the biochar samples, the Walkley–Black chromic acid wet oxidation method is utilized (based on the ASTM-D 5231-92 standard procedure).
The soil’s oxidizable organic carbon is oxidized using a 0.167 M solution of potassium dichromate (K2Cr2O7) in concentrated sulfuric acid. The intensity of the response raised the temperature, which is adequate to incite significant oxidation. At a wavelength of 600 nm, the remaining chromate is measured using spectrophotometry. Each experiment was performed using a furnace and repeated 3 times. The fixed carbon value, which is the remaining stable carbon content in biochar, was calculated from the difference between 100% of biochar and the sum of the percentage of moisture content, volatile compound content, and ash content. Eventually, the rest of the important chemical compositions (based on the ASTM-D 5231-92 standard procedure) that existed in biochar samples were evaluated using the SEM method (scanning electron microscopy) via an XRF (X-ray fluorescence) device, where the samples were exposed to radiation emitted from the X-ray tube. Specific X-radiation from the measured and collected count is compared to previous counts obtained from calibration samples in order to obtain a mass percentage of sulfur (S), magnesium (Mg), silicon (Si), calcium (Ca), aluminum (Al), phosphorus (P), iron (Fe), and potassium (K).
One crucial aspect to consider is the biochar yield, indicated by the organic carbon content (C) analyzed using the Walkley–Black method.

2.3. The Preparation for Biofiltration Processes

Biochar packing materials (separately measured for each pyrolysis temperature) were placed in a vertical plexiglass column (called a biofilter) that was 100 cm tall with an interior diameter of 14 cm (Figure 1).
Furthermore, moisture content, determined according to the ASTM-D 2867–99 standard procedure, ranged from 40% to 60% throughout the biofilter [36].
Subsequently, the operational conditions were controlled and evaluated, and they include the following: the size of the biofilter tube (cm), temperature of the biofilter (°C), inlet loading rate of biogas relative to the biofilter (ppm), inlet concentration of hydrogen sulfide relative to the biofilter (ppm), flow rate or velocity of biogas pumped into the biofilter (m/s), outlet concentration of hydrogen sulfide from the biofilter (ppm), empty bed retention time of the biofilter (s), moisture content or humidity of biochar (%), nutrient supply to biochar (g), volume of used biochar (cm3), biogas storage balloon (cm3), time to fill the balloon (min), and the duration of the experiment (days).
Regarding the nutrient supply solution used in this experiment to support the bacteria living in the biochar, K2HPO4 (0.02 g), (NH4)2SO4 (0.08 g), and Na2CO3 (0.39 g) are mixed with 1 L of deionized water, and the solution was constantly supplied to the biochar stages inside the biofilter. The height of the biochar bed is 10 cm. The inlet and outlet biogases were at the bottom and top of the biofilter column, respectively. The inlet concentration of H2S in the biofilter is controlled to be around 20–25 ppm, while the inlet loading rate (ILR) of the biogas was 0.16 m3 to 0.22 m3. The size of the biogas storage balloon is 0.16 cm3 (55Ø cm), and it takes 2 min 15 s, 2 min 55 s, and 4 min 30 s to completely fill the biofilter with biogas; it is then passed through a column at velocity/flow rates of 30 mL/min−1. The temperature of the biofilter, which was measured to be 27 °C to 30 °C, and velocity were analyzed using a testo 400 tool. Humidity/moisture content was controlled at around 40% to 60%, and the empty bed retention time (EBRT) was calculated using the equation below:
EBRT = (Height of biofilter)/(Velocity of inlet biogas)
Over 7 days, which is approximately 160 h, the outlet H2S concentration in the biogas from the biochar bed was continuously monitored using an infrared biogas composition analyzer, specifically the Shimadzu gas chromatography detector (model ITS04ATEX23415X).
The biochar volume implemented in this study was approximately 0.0055 m3 (divided into 5 stages, each containing almost 0.0011 m3 biochar particles with the same pyrolysis temperature). The calculated values for the empty bed retention time (EBRT) were 12.5 s, 16.6 s, and 25 s, corresponding to gas flow rates of 0.08 m/s, 0.06 m/s, and 0.04 m/s, respectively. Before starting the filter, the packing bed was prepared for the biological air purification process. Biogenic elements are added and activated biologically by passing H2S through it.
Biochar samples were continuously taken from the biofilter and soaked in 10 mL of sterile distilled water for microbial analysis. Sulfur-oxidizing bacteria were grown in a medium containing 5.0 g/L Na2S2O3, 2.0 g/L KNO3, 5.0 g/L KH2PO4, 0.5 g/L NaHCO3, 0.25 g/L MgCl2·6H2O, 0.25 g/L NH4Cl, 0.01 g/L FeSO4·7H2O, and 15.0 g/L agar powder at 30 °C for 2 days. Later, the number of microorganisms was expressed as CFU/g dry packing material.
The average experimental outcomes were derived from three replicates of each experimental treatment and expressed as the mean ± standard deviation. Data underwent analysis via analysis of variance, where values with a p-value below 0.05 were deemed acceptable. Analysis of variance was conducted to ascertain significant disparities in the adsorption capacity of sewage sludge biochar across various pyrolysis temperatures.

3. Results and Discussion

3.1. The Results of the Physicochemical Properties of Pyrolyzed Biochar

The results showed biochar packing materials after pyrolysis at 300 °C, 400 °C, 500 °C, and 600 °C, and specific bulk densities of 66 kg/m3, 67 kg/m3, 68 kg/m3, and 69 kg/m3, respectively, were observed, which are close to the results of [14] (71 kg/m3, 79 kg/m3, and 79 kg/m3 for 450 °C, 500 °C, and 550 °C, respectively). As pyrolysis temperatures increased, the bulk density increased as well. The higher temperature compacts biochar particles, reducing moisture and concentrating them in less space. The results demonstrated biochar packing materials after pyrolysis at 300 °C, 400 °C, 500 °C, and 600 °C, which resulted in porosities of 35%, 42%, 67%, and 75%, respectively. The obtained result confirms the author’s (Altikat) other observations in that a more porous structure was observed at higher pyrolysis temperatures when the pyrolysis temperature of the biochar was 400–600 °C [37]. Moreover, approximately similar porosities were attained (61% and 71% for 500 °C and 600 °C, respectively). The same porosities were observed with respect to magnetic biochar particles pyrolyzed at 100 °C, 200 °C, and 300 °C (30%, 31%, and 35%, respectively). As pyrolysis temperatures increased, the porosity of the samples subsequently enhanced. Regarding specific surface area measurements (based on the standard ASTM C1069-09 procedure), the data obtained from porosity assessment were used [38]. The analyzed bath temperature in both experiments was −186.898 °C. The outcome illustrated that biochar packing materials after pyrolysis at 300 °C, 400 °C, 500 °C, and 600 °C have specific surface areas of 10.97 m2/g, 22.28 m2/g, 29.76 m2/g, and 34.66 m2/g, respectively. Moreover, specific surface areas of 10 m2/g and 20 m2/g were reached for biochar after pyrolysis at 330 °C and 380 °C, respectively. Florent et al., while investigating the effect of pyrolysis temperature on biochar performance, observed that a temperature of 950 °C harmed the texture of the sample [39]. The obtained results demonstrated that as the pyrolysis temperature increases from 300 °C to 600 °C, values of 8.85, 9.12, 9.29, and 9.7 also demonstrated the value of alkaline biochar samples and their impact on amplifying removal efficiency [40,41,42]. When the pyrolysis temperature is enhanced, the overall pH of the sample will be impacted, which is why particles were more alkaline at higher temperatures. Additionally, as the pyrolysis temperature of biochar samples is enhanced, the total pH difference (gap) between the represented samples increases. Indeed, studies on biochar substrates prepared from anaerobically digested sewage sludge and fiber have highlighted the importance of surface alkalinity with respect to H2S removal, as the alkaline nature was suspected to facilitate H2S dissociation. The pH changes in the samples might be due to the composition (organic and inorganic materials) of the samples decomposed under different pyrolysis temperatures. Some compositions, such as hemicellulose and cellulose, decomposed below 400 °C. The results showed that as the pyrolysis temperature increases, the electrical conductivity of biochar samples at 300 °C to 600 °C achieved 310 μs/cm, 291.2 μs/cm, 287.1 μs/cm, and 285.9 μs/cm. We achieved electrical conductivity values of 271 μs/cm and 299 μs/cm for 600 °C and 700 °C biochar samples, respectively. As the temperature increases to obtain higher pyrolysis biochar particles, the electrical conductivity of the corresponding samples is significantly decreased. Metal oxides (heavy metals), such as Ca, Mg, Al, and Fe, and other elements, such as P, can act as H2S adsorption sites and undergo catalytic oxidation to convert H2S into elemental sulfur and sulfates. The results obtained using the Kjeldahl, Walkley–Black, and SEM methods are presented in Table 1 below.
Across pyrolysis temperatures of 300 °C, 400 °C, 500 °C, and 600 °C, the biochar yield decreased progressively, measuring approximately 28%, 22%, 17%, and 13%, respectively. Additionally, the nitrogen (N) content, assessed via the Kjeldahl method, decreased by 1% throughout the pyrolysis process. However, the sulfur (S) and potassium (K2O) levels remained consistent before and after pyrolysis. Interestingly, the amounts of calcium carbonate (CaO), aluminum oxide (Al2O3), phosphorus pentoxide (P2O5), iron oxide (Fe2O3), magnesium oxide (MgO), and silicon oxide (SiO2) experienced a notable increase in percentage after the initial pyrolysis at 400 °C. This trend continued for silicon oxide, suggesting that its proportion enlarges as nitrogen and carbon decrease, and this is likely due to chemical interactions. Notably, calcium (Ca), magnesium (Mg), and potassium (K) are recognized as active substances for hydrogen sulfide (H2S) absorption and chemical interactions, with their levels remaining consistent across all biochar samples post-pyrolysis. According to Yi et al., in addition to the catalytic decomposition of H2S to sulfur, iron and calcium compounds in the samples can also promote their conversion to sulfide and sulfate [43].

3.2. Biofiltration Performance and Kinetic Study of Hydrogen Sulfide Removal

Before conducting experimental studies of hydrogen sulfide removal, biofiltration performance and kinetic study were evaluated. That is, to develop a viable mathematical model of biofiltration performance and kinetic study, several simplifying assumptions are necessary. Here are the assumptions of the model [30,44]:
  • Negligible turbulence: Large turbulence is assumed to be negligible based on experimental data indicating a laminar flow pattern (with a Reynolds number between 0.2 and 0.5 and 0.5 for full-scale operations) in typical biofilters;
  • Homogeneous filter materials: The composition of the filter material, including porosity and water content, is assumed to be homogeneous;
  • Zero initial H2S concentration: It is assumed that the initial concentration of hydrogen sulfide (H2S) in the biofilter is zero;
  • Stoichiometric hydrogen sulfide production is assumed to follow a stoichiometric relationship, as described by the following equation: S + NaOH + Al + HCl → 3H2S + Al2O3;
  • Biomass distribution and density within the biofilter are assumed to be homogeneous.
The absorption and concentration reduction stages proceed as follows: Hydrogen sulfide is transferred from the gas phase to the biofilm surface surrounding the bed particles through a gas film. The equations provided below are accordingly used for calculations.
C_out = C_in∙(−k∙EBRT),
K = A m · X · μ · f x · D K · Y · t a n β ,
β = δ · √ X · μ K · Y · f x   · D ,
μ = μ m a x · S K S + S ,
In the provided equations, k represents the first-order reaction rate constant (s−1). K represents the Monod kinetic constant (kg·m−3). C_in stands for the hydrogen sulfide concentration in the biogas (g/cm). C_out denotes the hydrogen sulfide concentration at the gas film–biofilm interface (g/cm). fx is the rate of the diffusion of H2S in the biofilm relative to that in the water. D is the rate of H2S diffusion into the water (m2/s). m is Henry’s constant concerning a pollutant. µ is the specific growth rate of the biofilm (s−1). X is the biofilm density (kg/m). A represents the biofilm surface area. δ is the biofilm depth (m). Y is the yield coefficient, which equals the amount of biomass consumed. H is the height of the biofilter columns (m). V is the velocity of pumped air (m/s). [S] is the concentration of limiting substrate S for microbial bacteria growth (the value of [S] when µ/(µmax) is 0.5). µmax stands for the maximum specific growth rate (s−1).
R = μ m a x · C i n C i n + K S
In this context, R represents the bacterial growth rate (s−1). The next phase involves the conversion of hydrogen sulfide into harmless sulfur compounds, which are stored within bacterial cells. Bacteria flourish by consuming hydrogen sulfide compounds; thus, the bacterial growth rate is intimately connected with the rate of hydrogen sulfide consumption. As a mathematical modeling result, the thickness of the biofilm increases progressively over time. Such a conclusion was reported by other scientists [45,46] who conducted experimental microbiological studies. For example, when the concentration of hydrogen sulfide in the inlet biogas reaches 10–200 ppm, this indicates that the limiting substrate (sulfur) for sulfur-oxidizing bacteria has been reached. This leads to sulfur accumulation in the packing bed of the biofilter, negatively impacting the overall removal efficiency of the biofilter. According to this assumption, the number of sulfur-oxidizing microorganisms depends on the H2S concentration, and preventing sulfur accumulation in the packing bed is essential in order to reach high efficiency of air cleaning. In this study, steps were taken to prevent sulfur accumulation in the packing bed of the biofilter by maintaining a low [S] value and implementing biochar samples with higher pyrolysis temperatures [47,48,49].

3.3. Microbiological Studies of the Packing Material and Removal Efficiency of Hydrogen Sulfide

After the biofiltration efficiency and kinetics studies, microbiological studies of the packing material and hydrogen sulfide removal efficiency were conducted. By following these procedures, as the pyrolysis temperature of the biochar increases, the time taken to reach sulfur accumulation will be longer, ultimately resulting in a higher number of microorganisms in the biofilter packing material. However, in this study, after 60 days of the experiment, the average growth rates of several sulfur-oxidizing microorganisms were 1.55 × 104 CFU/g at the first stage of the biofilter, 2.63 × 104 CFU/g at the second stage, 3.65 × 104 CFU/g at the third stage, 5.73 × 104 CFU/g at the fourth stage, and 2.62 × 104 CFU/g at the fifth stage. The following graph provides clear evidence supporting these claims (Figure 2). Several sulfur-oxidizing microorganisms in the packing bed of the biofilter during the 60 days of the experiment constantly increased. This incremental growth is anticipated to continue until the environmental conditions are no longer conducive to their survival. The most important species involved in hydrogen purification is Thiobacillus sp. The following other bacterial cultures are involved in the sulfur-oxidizing process: Pseudomonas sp., Paenibacillus sp., Bacillus sp., Aeromonas sp., Previbacillus sp., Sphingobacterium sp., and Acinetobacter sp., etc.
Figure 3 shows the variation in hydrogen sulfide removal efficiency using biochar after different pyrolysis temperatures. In summary, different biochar samples were used and pyrolyzed at 300 °C, 400 °C, 500 °C, and 600 °C depending on the ability of each biochar to create an optimal environment for bacterial growth. During the conducted research, it was established that the biofilter showed excellent efficiency when using carbon pyrolyzed at temperatures of 500 °C and 600 °C.
As shown in Figure 3, the number of sulfur-oxidizing bacteria present in the biofilter packing material increased continuously over time when pyrolyzed carbon at 600 °C was used in the biofilter. However, it is very important to consider the sulfur accumulation according to the specific pyrolysis temperature of each biochar sample. Therefore, as shown in the graphs above, increasing the pyrolysis temperature of biochar samples increases the number of sulfur-oxidizing bacteria. This in turn has a positive effect on the removal efficiency of the biofilter in terms of removing hydrogen sulfide from the biogas. This is confirmed in research conducted by other scientists [50,51,52,53,54,55,56,57].
Biofilter studies have shown that the amount of hydrogen sulfide in biogas was effectively reduced during the 60 days of research (Figure 4). For example, the removal of hydrogen sulfide continuously increased by an average of about 20 percent in all stages of this filter. Such results were achieved using 600 °C pyrolyzed biochar, which was inhabited by a large number of sulfur-oxidizing microorganisms, as shown by the microbiological investigation. Finally, the experimental results of H2S purification in biogas were compared with the results of the mathematical modeling process using Equations (4)–(6). The mathematical modeling results only confirmed that the filter is effective: That is, the model’s calculated filter efficiency also reached between 70 and 90 percent after 60 days of testing. This comparison allows for an assessment of the effectiveness and accuracy of the mathematical model in predicting the purification process in the future. Similar results were reported in calculations by other authors [16,26,45,56].
During the research, it was found that the concentration of CH4 increased slightly during the desulfurization process. According to other researchers, the slight decrease in CO2 concentration in biogas was proportional to the recorded increase in CH4 concentration, from 60.0% to 61.7–63.5%, respectively. These observations will aid in the development of biochar as engineered sorbents for the elimination of hydrogen sulfide. It has also been observed that sewage sludge pyrolyzed at a temperature of 600 °C (after anaerobic digestion) is suitable for the production of effectively activated carbon adsorbents. The authors Choleva et al. note that simple chemical and physical activation processes of the produced BCs result in very effective H2S adsorbents that can be used in series to desulfurize and upgrade biogas [58,59,60,61,62,63,64].

4. Conclusions

The results from the BET method indicated that the specific surface area of biochar samples increased with respect to pyrolysis temperature, reaching up to 2.5 times the initial value at 600 °C compared to 300 °C.
Considering pH analyses, since aerobic sulfur-oxidizing bacteria prefer a more alkaline environment, biochar pyrolyzed at 600 °C exhibited the highest pH value (8.7), while biochar pyrolyzed at 300 °C exhibited the lowest pH value (7.85), with the biochar particles’ pH increasing with respect to pyrolysis temperature.
This analysis considered the influence of the outlet H2S concentration and its relation to sulfur accumulation in the packing bed, as well as the bacterial growth rate across all four samples. Finally, biofiltration performance and kinetic studies were employed to review the experimental results and propose more accurate data interpretation.
Several sulfur-oxidizing microorganisms in the packing bed of the biofilter during the 60 days of the experiment constantly increased. After 60 days of the experiment, the average growth rates of several sulfur-oxidizing microorganisms were 1.55 × 104 CFU/g at the first stage of the biofilter, 2.63 × 104 CFU/g at the second stage, 3.65 × 104 CFU/g at the third stage, 5.73 × 104 CFU/g at the fourth stage, and 2.62 × 104 CFU/g at the fifth stage. This incremental growth is anticipated to continue until the environmental conditions are no longer conducive to their survival.
Different biochar samples were used and pyrolyzed at 300 °C, 400 °C, 500 °C, and 600 °C depending on the ability of each biochar to create an optimal environment for bacterial growth. During the conducted research, it was established that the biofilter showed excellent efficiency (between 55 and 99%) when using carbon pyrolyzed at temperatures of 500 °C and 600 °C.

Author Contributions

Conceptualization: K.M. and A.M.; data curation: K.M.; formal analysis: K.M. and A.M.; investigation: K.M.; methodology: K.M.; project administration: R.V., K.M. and A.M.; resources: K.M. and A.M.; software: K.M.; supervision: A.M.; validation: K.M., R.V. and A.M.; visualization: K.M.; writing—original draft: K.M.; writing—review and editing: K.M., R.V. and A.M. All authors have read and agreed to the published version of the manuscript.

Funding

This research received no external funding.

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

The raw data supporting the conclusions of this article will be made available by the authors upon request.

Acknowledgments

This study was extracted from a doctoral dissertation conducted at the Vilnius Gediminas Technical University. I would like to express my sincere gratitude to the Department of Environmental Protection and Water Engineering for their invaluable contribution and support throughout the development of this article.

Conflicts of Interest

The funders had no role in the design of the study; in the collection, analyses, or interpretation of data; in the writing of the manuscript; or in the decision to publish the results. The author declares that he has no competing interests.

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Figure 1. Experimental studies on the biofiltration process for the removal of hydrogen sulfide from biogas: 1—samples of sewage sludge before pyrolysis; 2—samples of sewage sludge after pyrolysis; 3—biochar bed inside the biofilter; 4—biofilter.
Figure 1. Experimental studies on the biofiltration process for the removal of hydrogen sulfide from biogas: 1—samples of sewage sludge before pyrolysis; 2—samples of sewage sludge after pyrolysis; 3—biochar bed inside the biofilter; 4—biofilter.
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Figure 2. Number of sulfur-oxidizing microorganisms in the packing material of the biofilter during 60 days of the experiment.
Figure 2. Number of sulfur-oxidizing microorganisms in the packing material of the biofilter during 60 days of the experiment.
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Figure 3. Hydrogen sulfide removal efficiency using biochar after different pyrolysis temperatures and during the 60-day experiment.
Figure 3. Hydrogen sulfide removal efficiency using biochar after different pyrolysis temperatures and during the 60-day experiment.
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Figure 4. Hydrogen sulfide removal efficiency using biofilters during the 60-day experiment.
Figure 4. Hydrogen sulfide removal efficiency using biofilters during the 60-day experiment.
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Table 1. X-ray fluorescence (SEM method), Kjeldahl, and Walkley–Black examination results for identifying the chemical composition of biochar samples.
Table 1. X-ray fluorescence (SEM method), Kjeldahl, and Walkley–Black examination results for identifying the chemical composition of biochar samples.
Type of MethodX-ray FluorescenceKjeldahlWalkley–Black
Type of elementsSiO2CaOAl2O3P2O5Fe2O3K2OMgOSNC
Sewage sludge18.93%11.83%3.51%9.59%4.38%1.39%2.34%0.96%4.2%30.83%
Biochar after 400 °C25.67%17.45%5.23%13.43%6.38%1.78%3.39%0.78%2.1%24.17%
Biochar after 500 °C29.82%15.47%5.14%13.78%6.67%1.70%3.39%0.85%2.8%19.6%
Biochar after 600 °C30.27%15.80%5.39%14.38%6.6%1.70%3.50%0.81%2.1%15.41%
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Vaiškūnaitė, R.; Mažeikienė, A.; Mohammadi, K. Effects of Pyrolysis Temperature on Biochar Physicochemical and Microbial Properties for H2S Removal from Biogas. Sustainability 2024, 16, 5424. https://doi.org/10.3390/su16135424

AMA Style

Vaiškūnaitė R, Mažeikienė A, Mohammadi K. Effects of Pyrolysis Temperature on Biochar Physicochemical and Microbial Properties for H2S Removal from Biogas. Sustainability. 2024; 16(13):5424. https://doi.org/10.3390/su16135424

Chicago/Turabian Style

Vaiškūnaitė, Rasa, Aušra Mažeikienė, and Kamyab Mohammadi. 2024. "Effects of Pyrolysis Temperature on Biochar Physicochemical and Microbial Properties for H2S Removal from Biogas" Sustainability 16, no. 13: 5424. https://doi.org/10.3390/su16135424

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