Maximizing sustainability by energy recovery in anaerobic treatment systems

© 2024 IWAP. This is an Open Access book chapter distributed under the terms of the Creative Commons Attribution License (CC BY-NC-ND 4.0) which permits copying and redistribution for non-commercial purposes with no derivatives, provided the work is properly cited (https://creativecommons.org/licenses/by-nc-nd/4.0/). The chapter is from the book Anaerobic Treatment of Domestic Wastewater: Present Status and Potentialities, M.C. Tomei and J.M. Garrido (Editors). doi: 10.2166/9781789063479_0167 A. V. Haandel1* , S. L. Santos2, T. Allegue3 and J. M. Garrido4 1Department of Civil Engineering, Universidad Federal of Campina Grande, Campina Grande, PB, Brazil 2Department of Civil Engineering, Universidad Federal of Rio Grande do Norte, Natal, RN, Brazil 3Department of Chemical Engineering, Khalifa University, Abu Dhabi, United Arab Emirates 4Department of Chemical Engineering, Universidade de Santiago de Compostela, CRETUS, Santiago de Compostela, Galicia, Spain *Correspondence: [email protected] ABSTRACT Traditionally, wastewater treatment systems require considerable energy input, but this has changed since modern anaerobic treatment systems were developed in the second half of the last century. The production of biogas opens the possibility to produce electric power, which may make these systems independent of external energy and may even become net energy producers. A second source of energy production in wastewater treatment systems is sludge production. After drying the sludge its combustion can generate energy, though less than the energy potential from biogas, not only due to the reduced sludge mass, but also because solid fuel has a lower conversion efficiency than gaseous fuel. It has been demonstrated that in an efficient anaerobic treatment system operating under favourable conditions (warm climate, low-sulphate concentration in the influent) ∼1/6 of the influent organic material is converted into sludge whereas 4/6 is converted into biogas and 1/6 leaves the anaerobic treatment system with the treated effluent. Normally, the production of electric power from the biogas is more than sufficient to run the treatment system and an external source energy is not required. However, this does not mean that energy production is always a sound economic solution. Economic feasibility depends on the price of electric power and generally requires the operation of a large treatment system. In this chapter, equations are derived to estimate the composition of biogas and sludge in anaerobic wastewater treatment systems and specifically the influence of sulphate in the influent is discussed. Moreover, the processes for collection and treatment of biogas and sludge are presented. The potential for useful energy generation from anaerobic treatment systems is analysed and several examples are presented in order to highlight the feasibility of the energy recovery under different operating conditions. Keywords: anaerobic wastewater treatment, biogas production and composition, electric power from biogas and sludge, influence of sulphate in the influent, volatile solids production. 6.1 INTRODUCTION All biological wastewater treatment systems seek to reduce the concentration of organic material since its presence is highly undesirable. When raw wastewater is discharged on surface water, it leads to consumption of dissolved oxygen, and, depending on the ratio between surface water flow and organic Chapter 6 Maximizing sustainability by energy recovery in anaerobic treatment systems Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 168 Anaerobic Treatment of Domestic Wastewater: Present Status and Potentialities load, it can cause death of microorganisms. Conventional aerobic wastewater treatment systems are energy consumers, but anaerobic treatment systems offer the possibility of recovering the energy content of the influent organic matter, which can be profitably employed in wastewater treatment plants or sold as electric energy. As shown in Figure 6.1 the organic material of wastewater after the anaerobic treatment is distributed in three fractions: (1) a gaseous fraction composed of biogas, (2) a solid fraction composed of biological sludge and (3) a fraction dissolved in the effluent. In this distribution, the largest fraction is biogas, in which methane is the main component and can be used for electric energy generation. Sludge combustion can provide useful energy production after separation from the liquid phase and drying. The residual organic material in the effluent is composed of a biodegradable and a non-biodegradable fraction. The biodegradable components can be removed in a post-treatment unit as activated sludge or polishing ponds (see Chapter 5) whereas the non-biodegradable fraction will only be removed if the production of high-quality water is an objective of the treatment system. In that case advanced processes such as reverse osmosis can be used to produce water without impurities. In this chapter, an estimate is made of the composition of biogas and sludge as well as the potential for useful energy generation from these in anaerobic treatment systems. Moreover, the processes for collection and treatment of biogas and sludge are presented. 6.2 ENERGY RECOVERY IN ANAEROBIC TREATMENT SYSTEMS The production of biogas is the most characteristic feature of the anaerobic treatment. Biogas is composed mainly of methane and carbon dioxide, which are decomposable products of organic material and have a limited solubility in water. There are other components, whose magnitude depends on the operating conditions in digesters and the characteristics of wastewater. In this chapter, it will be shown that for sewage treatment contributions of nitrogen and water vapour in biogas may exceed that of carbon dioxide. If sulphate is present in wastewater hydrogen sulphide will also be formed, which generally has a low concentration, but is nonetheless important because it can cause several problems in anaerobic digesters as it will be reported in detail in Section 6.4.7. The destination of biogas depends on the produced quantity and the quality. If high amounts are available, it is possible to use biogas in a productive way, generating electricity and heat in combined heat and power units or refining and pressurizing it for subsequent use, for example for automotive purposes or injection in the gas grid. The technical feasibility of using biogas for power generation is not in question: there are many manufacturers who produce generators specifically designed for the use of biogas as fuel, with power generation ranging from 30 kW to over 5 MW. The transformation efficiency of these units is in the range of 35–40%. For very large biogas flows (production >5 MWel) there is the possibility of using gas turbines, which have a high initial cost, but with an efficiency of ≥50%. Figure 6.1 Distribution of organic material of influent wastewater in anaerobic treatment systems. Numbers indicate the estimated fractions under normal operational conditions. Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 169Maximizing sustainability by energy recovery in anaerobic treatment systems A more complicated application is the upgrading of biogas for producing pure methane. In this case, the purified biogas must be compressed to a pressure high enough for grid injection (16 bar) or to a high pressure of 200–250 atm to be used in cars which are adapted for use of compressed natural gas. The upgrading of biogas into methane has also been demonstrated, but the feasibility of this alternative depends on the price of conventional fuel, which is too low for widespread application. Raw biogas treatment processes are well established and the economic feasibility depends mainly on the benefit of the productive use, that is, the cost of fuel that otherwise would have to be purchased. In small treatment systems, it is usually not economically convenient using the biogas for energy production and it is preferable to burn it with a flare. It is worth noting that the combustion is important, and that omission of this step can result in the generation of greenhouse gases (GHG) by the anaerobic treatment higher than the one derived by the aerobic. The methane is a potent GHG with a carbon equivalence of 25:1, that is 1 mol of methane gas is 25 times more harmful than carbon dioxide. Example 6.1 Estimate the per capita production of gases contributing to the greenhouse effect in aerobic and anaerobic systems for a per capita contribution of 100 g chemical oxygen demand (COD)/day and making the following assumptions: (1) Aerobic: oxygen consumption: 0.72 g O2/g COD; aeration efficiency: 1 kWh/kg O2; fuel consumption for power generation: 0.2 kg CH4/day/kW of electric power. (2) Anaerobic: anaerobic digestion efficiency: 70% (3) Average elemental composition of sewage is C2H2O, close to the value of C2H19/9 ON1/9 determined by Henze and Harremoes (2001). Solution Assuming that carbon dioxide is produced in equimolar ratio with the consumption of oxygen in the aerobic system, it follows that the oxygen consumption of 72 g O2/inh/day or 72/32 = 2.25 mol/inh/ day generates 2.25 mol/inh/day of CO2. On the contrary, the energy consumed in oxygen transfer (1 kWh/kg O2) will be ∼72 Wh/day or 3 W. For power generation of 1 W, consumption of 5 g CH4/day is required and this generates 5/16 = 0.31 mol/day CO2 (other fuels have roughly the same contribution). For a power generation of 3 W, the production of CO2 is therefore 0.93 mol/inh/day leading to a total of 2.25 + 0.93 = 3.2 mol CO2/inh/day in aerobic systems. In anaerobic systems the per capita production is 0.7 × 100/4 = 17.5 g CH4/day or 17.5/16 = 1.1 mol CH4/day. If this amount is not burned, fatally it escapes into the atmosphere as a contributor to the GHG effect. Not considering the production of CO2 in anaerobic digestion (its major fraction does not pass into the atmosphere but remains dissolved), methane production is equivalent to 21 × 1.1 = 23 mol CO2/inh/day, a factor 23/3.2 = 7 times greater than the calculated value for the aerobic generation system! By burning the methane production of biogas, the CO2 production will be 1.1 mol/inh/day, ∼1/3 of the production in aerobic systems. If methane in the biogas is burnt, there is still the dissolved CH4, with amounts to ∼20 mg CH4/L generating from the digestion of 80 mg COD/L. If the dissolved CH4 is not metabolized, for example by methanotrophic bacteria it will eventually be released into the atmosphere. If a per capita contribution of 120 L/day is assumed, the dissolved methane would be 20/16 × 120 = 150 mmol CH4/ inh equivalent to 3.1 mol CO2/inh, equal to the production in aerobic systems. It can be concluded that even if all the biogas produced is burnt, the CO2 production of the burnt biogas and the release of dissolved methane leads to at least the same contribution of GHGs by anaerobic and aerobic treatment systems. Therefore, the sustainability of anaerobic systems for sewage treatment is actually questionable, due to the unavoidable release of GHGs into the atmosphere. In the above analysis, anaerobic digestion of aerobic sludge is not considered, but in reality, ∼30% of the organic Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 170 Anaerobic Treatment of Domestic Wastewater: Present Status and Potentialities influent load to aerobic systems is converted into methane and will increase the output of GHGs, so that in the end the two systems produce practically equivalent amounts of GHGs. If methane can be productively used, anaerobic systems become more attractive in terms of sustainability, because they reduce the use of fossil fuels. 6.3 SOLUBILITY OF GASES Biogas is a mixture of various gases and vapours that desorb from the liquid phase in an anaerobic digester because of the limited solubility of these products in water. To estimate the flow of biogas, it is necessary to know the production and solubility of its components. Calculation requires a first assumption of equilibrium conditions between the gaseous phase (biogas) and liquid phase in a bioreactor whose composition is assumed to be equal to one of the effluents, and that there is a uniform composition in both phases. In good approximation, this will normally be the case, even though the solid phase in the reactor can be highly stratified (Cozma etal., 2013). The solubility of gases in the liquid phase can be expressed with Henry’s law, stating that if a liquid and a gaseous phase are in equilibrium, there is proportionality between the saturation (maximum) concentration of a gas component in the liquid phase and its concentration in the gas phase: CkC sdg = (6.1) where Cs = saturation concentration of the dissolved gas (mg/L or mmol/L); Cg = gas concentration in the gas phase (mg/L or mmol/L) and kd = distribution coefficient (dimensionless). The distribution coefficient indicates the solubility of a gas: the higher the coefficient, the higher the solubility. The value of this coefficient has been determined and tabulated for the most important volatile compounds. Table 6.1 shows the values of the distribution coefficients for different temperatures of gases which can be biogas components. Concentrations may be expressed in any unit (g/L, mol/L, etc.). Equation (6.1) can also be written in a different form: CkP shp = (6.2) where kh = Henry’s constant; Pp = partial pressure of the component in the biogas (g/(m3 Pa) or g/J). The relationship between the distribution coefficient and the Henry constant is: kkWRT hd mg/J= () / (6.3) where Wm = molar weight of the gas; R = universal gas constant = 8.3 J/kmol; T = temperature (K). Table 6.1 Distribution coefficients of several biogas components. Gas Formula Molecular Weight (g/mol) Temperature (°C) kd20 kd25 kd30 Methane CH416 0.034 0.032 0.031 Carbon dioxide CO244 0.94 0.83 0.74 Nitrogen N228 0.017 0.016 0.015 Oxygen O232 0.034 0.032 0.030 Ammonium NH317 0.76 0.71 0.67 Hydrogen sulphide H2S34 2.87 2.55 2.27 Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 171Maximizing sustainability by energy recovery in anaerobic treatment systems 6.4 BIOGAS FLOW To estimate the flow of biogas generated in anaerobic treatment, firstly it is necessary to evaluate the volumetric flow of its components. In Figure 6.2 the gases normally present in biogas are presented. Along with methane, which is the main component, significant contributions of carbon dioxide, water and nitrogen are expected. In addition, there are other compounds which are less bulky, but whose presence is important because of their nature: O2, NH3 and H2S (Van Haandel & van der Lubbe, 2019). Regarding hydrogen sulphide, the contribution of this component depends on the concentration of sulphate in the water supply generating the sewage. Usually, this concentration will be small but there are regions where water resources with a high-sulphate concentration are the only source for public water supply. In these cases, sulphide production may still not be high, but due to sulphate reduction, there is a consumption of organic matter and methane production may be strongly reduced. Nitrogen, oxygen and ammonium are dissolved in influent whereas methane, carbon dioxide and hydrogen sulphide gas are generated in the anaerobic treatment unit. Being a vapour and not a gas at ambient temperature, the partial pressure of H2O is constant for a given temperature and is independent of the pressure of the gas phase. 6.4.1 Production potential of methane Methane production depends directly on the applied organic load and digestion efficiency. Stoichiometric production is calculated as 0.25 kg of CH4/kg digested COD. When it is assumed that ∼10% of influent COD is transformed into volatile solids, the production of methane can be calculated from the applied organic load if the removal efficiency of the organic material is known. The removal efficiency depends on a variety of factors: the nature of the organic material, environmental conditions, operating conditions and reactor characteristics in terms of design and construction. The daily mass of methane produced may be expressed as MCH ta te xv MS MS MS 44= () –– / (6.4) As a first estimate of the methane production, it can be assumed that all produced methane desorbs from the liquid phase. Knowing that the molar volume of a gas at atmospheric pressure is given by Figure 6.2 Schematic representation of the different components of biogas and the equilibrium between the gas and liquid phases. Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 172 Anaerobic Treatment of Domestic Wastewater: Present Status and Potentialities 22.4 × (273 + T)/273 (T in °C) and 1 mol of methane has a mass of 16 g, the produced methane volume is calculated as VTM T CH CH ti te x MS MS MS 44 22 4273 273 16 22 4273 273 =×+× =+ −− () () .// ./ vv () /64 (6.5) where V CH4 = first estimate of methane volume in the biogas (L/day); M CH4 = mass of the produced methane (g/day); MSti = mass of applied COD (g/day); MSte = mass of COD in the effluent (g/day); MSxv = COD mass converted into volatile sludge (g/day). Example 6.2 Determine the potential for per capita methane production from sewage by anaerobic digestion, making the following assumptions: (1) per capita organic matter production: 100 g COD/inh/day, (2) the organic material removal efficiency: 80%, (3) conversion into volatile solids: 10%. Solution From Equation (6.4): M CH ti te xv MS MS MS gCHinh/day/ 44 100 20 10 4175 4 =−− =−−= () () / /. . From Equation (6.5) at a temperature of 25°C: VT M CH CH 44 22 4273 273 16 22 4273 25 273 17 516 26 =×+× =× +× = () () .// ././ .. .7 4 Linh/day/ CH However, the estimated volume of the produced methane is not precise since part of the methane produced will remain dissolved in the liquid phase. The solubility of methane is low (as shown in Table 6.1) being the distribution coefficient kd = 0.03. To make a first estimate of dissolved methane concentration can be assumed that biogas is composed by only methane. As 1 mol of methane (or any other gas) has a volume of 24.4 L at a temperature of 25°C and atmospheric pressure, its concentration is 1/24.4 mmol/L = 41 mmol/L. Now, applying Equation (6.1) for the case of having only methane in the biogas is: Cs = 41 × 0.03 = 1.2 mmol/L = 20 mg/L. Thus, it is estimated that the dissolved methane is 20 mg CH4/L. As is known that for producing 1 mg CH4, about a mass of 4 mg of COD must be digested, it is estimated that a maximum of 80 mg COD/L will be found as dissolved methane after anaerobic digestion. This concentration is not negligible in the case of municipal sewage, which normally has a concentration of digestible COD in the range of 300– 500 mg/L. In Example 6.2, for a per capita contribution of 120 L/day, methane loss in the effluent would be 0.020 g CH4/L × 120 L/inh/day = 2.4 g CH4/inh/day or 2.4/17.5 = 14%. So the best estimate of CH4 production in the biogas would be MCH4 = 17.5 − 2.4 = 15.1 g/inh/day, whereas the per capita volume is VCH4 = 22.4 × (273 + 27)/273 × 15 1/16 = 23.1 CH4/L/day. It is important to notice that as the temperature becomes lower, the digested COD becomes smaller and the dissolved methane concentration becomes higher. As a result, the dissolved fraction increases and may be as high as 50% in case with a low influent COD and a low temperature. Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 173Maximizing sustainability by energy recovery in anaerobic treatment systems Example 6.3 Estimate the fraction of the produced methane that remains dissolved in a upflow anaerobic sludge blanket (UASB) reactor operating at a temperature of 15°C and a sludge age of 75 days, treating sewage with a COD of 500 mg/L in the influent. Solution The fractions of the influent COD which are discharged in the effluent (fs) and transformed into sludge (fx) can be estimated (from Equations (5.1) and (5.2)) as fs = 0.22 and fx = 0.11 (see also dos Santos, 2013). Hence, the digested fraction is fd = 1 − fs − fx = 0.67. Therefore, the digested COD is 0.67 × 500 = 330 mg/L, equivalent to 330/4 = 82 mg CH4/L. On the contrary, at 15°C the solubility of methane in water at 15°C is ∼23 mg/L (kd = 0.036), so that a fraction of 23/82 = 28% is retained in the liquid phase and 72% is released into the biogas. Figure 6.3 shows the methane fraction in the gas phase for different temperatures and influent COD concentration as a function of the sludge age. To make a more accurate estimate of the biogas volume and the partial pressures of methane and other components in the biogas, it is necessary to apply an iterative calculation, taking into account also the contributions of other gases. Firstly, the contribution of each component will be individually evaluated and then the iterative calculation procedure will be applied. 6.4.2 Carbon dioxide The presence of CO2 in the biogas is due to its production in the anaerobic digestion process. It can be assumed that the CO2 production in molar terms is equivalent to the methane production, considering an average carbon oxidation number close to zero for organic matter, accordingly with the elemental composition of sewage used for Example 6.1, but its presence in the gas phase is much smaller because of its greater solubility. The partial CO2 pressure can be estimated from the equilibrium of the carbonic system: CO HO HHCO 22 3 +↔+ +− (6.6a) Figure 6.3 Fraction of the methane production that desorbs from the liquid phase as a function of the sludge age for different temperatures and influent COD concentrations. Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 174 Anaerobic Treatment of Domestic Wastewater: Present Status and Potentialities or HCOH CO 32 1 −+          [] =/k (6.6b) or HCOH CO dg 32 1 −+          [] () =/ kk (6.6c) By applying Equation (6.3) this can also be written as pkK CO pp pH HCO H 2 1 310=     −+− () (6.7a) where pX = −log[X]; k1 = first dissociation constant of the carbonic system = 4.6 × 10−7; kH = Henry constant for CO2 = 0.034 mol/L/atm at 25°C; that is, pKH = 1.47 (Loewenthal & Marais 1976). Since at pH ≈ 7 the alkalinity is practically equal to the concentration of bicarbonate, Equation (6.7a) can also be written as pkK CO pp pH Alc H 2 1 10=× +− . () (6.7b) It is concluded that in the case of sewage treatment the CO2 partial pressure will be low: for instance for an alkalinity of 350 ppm or 7 meq/L (normal for domestic wastewater) and a pH = 7.0 a partial pressure of only 0.04 at 25°C is calculated. The CO2 pressure can be calculated provided there are data of alkalinity and pH, as well as for temperature. The presence of CO2 in the biogas normally is not problematic. In contrast dissolved CO2 in the liquid phase results in aggressive water, which can lead to serious concrete corrosion problems. 6.4.3 Water vapour The pressure of water vapour is directly related to the temperature. At 25°C, the water saturation pressure is 23.7 mm of mercury (3.1 Pa), or at a pressure of 760 mm or 1.013 kPa partial pressure is 0.031. Table 6.2 shows some values of water vapour pressure for different temperatures. The presence of water vapour in the biogas can cause problems, particularly when the ambient temperature greatly varies. During periods of low ambient temperature, water may condensate in biogas piping, because the biogas is generated in the relatively warm liquid phase. The biogas will have water vapour at the pressure of the liquid phase in the reactor, but out of the treatment system the biogas temperature decreases and water vapour will tend to become supersaturated and condensate. Apart from causing blockages, the condensed water tends to cause corrosion problems, especially if hydrogen sulphide is also present in the biogas. For those reasons, there will usually be a condensate retention device in the biogas transport system, which may be the hydraulic seal of the UASB reactor, if located at the lowest point of biogas piping. Table 6.2 Values of the saturation pressure of water vapour at different temperatures. Temperature (°C) Saturation Pressure Temperature (°C) Saturation Pressure mm Hg Partial Pressure mm Hg Partial Pressure 04.6 0.006 20 17.5 0.023 56.5 0.009 25 23.7 0.031 10 9.2 0.012 30 31.7 0.042 15 12.8 0.017 35 41 0.054 Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 175Maximizing sustainability by energy recovery in anaerobic treatment systems 6.4.4 Nitrogen In many cases, it may be assumed that wastewaters are saturated with atmospheric nitrogen (which has a partial pressure of 0.8 atm). In that case, the saturation concentration is 14.8 mg/L at 25°C (Table 6.1). If it is assumed that nitrogen is not produced or lost during anaerobic treatment, the influent nitrogen will leave the system in the biogas and in the effluent, thus from the mass balance: QC QC QC lal sg g =+ (6.8a) or CC QQC as glg =+ () / (6.8b) Hence CCkQQ ga dgl =+ () () // (6.8c) and Ck CQ Q sd ag l =+ () () // 1 (6.8d) where Ca = nitrogen concentration in the influent; Cg = nitrogen concentration in the gas phase; Qg/Ql = ratio between the biogas and wastewater flow. In order to calculate nitrogen in the effluent and in the biogas, it is necessary to know the concentration in the influent and the ratio of the flows of biogas and wastewater. Example 6.4 For the conditions of Example 6.2 and having a per capita sewage contribution of 120 L/day, estimate the partial pressure of nitrogen on the biogas. Solution The concentration of nitrogen in the influent is estimated from the distribution constant (kd = 0.016, Table 6.1). At a partial pressure of 0.8 bar, the concentration of nitrogen in the air is 41 × 0.8 = 34 mmol/L. Now the saturation concentration of nitrogen is calculated as CC sa mmol/Land mgN/L=×==0016 34 054148.. . Since Qg/Ql = 26.7/120 = 0.22 (Example 6.2), the nitrogen concentration in the biogas is CCkQQ ga dga // /mg/Lor/ mmol =+ =+ == () () () 14 80016 02263632822.. .. //L. Hence CkC sdg mg/L== ×=0016 63 10... The calculations indicate that nitrogen is efficiently removed by stripping in the anaerobic digester and almost all transferred to the gas phase, due to its very low solubility. At a temperature of 25°C and a pressure of 1 bar the concentration of a gas is 41 mmol, so that when the nitrogen has a concentration of 2.2 mmol/L in the gas phase, its partial pressure is 2.2/41 = 0.05 atm. Comparing the calculation results of the partial pressure of carbon dioxide, water and nitrogen (Sections 6.4.2 and 6.4.3), surprisingly it is observed that for anaerobic digestion of normal sewage in the produced biogas, the partial pressures for carbon dioxide, nitrogen and water vapour are all of the same order of magnitude. Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 182 Anaerobic Treatment of Domestic Wastewater: Present Status and Potentialities methane gas concentration in the biogas is now 0.87 × 41 = 36 mmol/L, whereas the saturation concentration in the liquid phase will be Cs = 0.03 × 36 = 1.08 mmol/L or 17 mg/L. Therefore, the COD concentration of digested organic material, which is found as dissolved methane in the liquid phase, is 4 × 17 = 68 mg COD/L. For a per capita contribution of 120 L/day, a mass of 68 × 120 = 8160 mg COD/day, equivalent to MCH4 = 8.16/4 = 2 g CH4/day, remains in the effluent. Moreover, for the assumed sulphate concentration of 20 mg S/L, the oxidation of 2 × 20 = 40 mg COD/L or 0.04 × 120 = 4.8 g COD/day is used for sulphate reduction, which reduces the production of methane by 4.8/4 mg CH4/L or 1.2 g CH4/day. Therefore, the mass of methane in the biogas can now be expressed as: MCH ti te xv CH doxi MS MS MS MS MS 444 100 20 10 82 48 4 =−−− =−−− − () () –/ ../===57 414254 /.gCHinh/day/ (6.14a) The new volume can now be calculated, duly not accounting the dissolved methane and the methane, which is not produced due to oxidation of organic material by sulphate: ′=×+× =+×= () () VtM t CH CH 44 22 4273 273 16 22 4273 273 14 25 16 23 .// ././ .. 2L/inh/day (6.14b) (c) New estimate of the biogas/effluent flow ratio Once the new estimates of the methane flow and partial pressure are obtained, the biogas/ effluent flow ratio, Qg/Qa, can be calculated as QV p gCHCHL/inh/day=′== 44 23 2087 26 7/./. . (6.14c) and QQ glL/L/././. /. == =26 7120 23 8087 120 022 (6.14d) (d) New estimate of the partial pressure of nitrogen Having a new estimate for the Qg/Ql ratio, the new values for the partial pressure of nitrogen is obtained, using the same procedure as in Example 6.3: pCkQQ Na dgl // 228 41 005=+ = () () // . (e) Compare the recalculated value of the composition and flow of the biogas with the values obtained in the previous iteration. Knowing that the values of the partial pressures of CO2 (=0.04) and H2O (=0.03) remain the same because they depend only on the pH + alkalinity and the temperature respectively, the new values of the partial pressure of CH4 (=1 − 0.04 − 0.03 − 0.05 = 0.87) and the biogas volume (=26.7 L/inh/day) are now compared with the values of the previous iteration (1 and 26.7 L/inh/day respectively). It can be observed that these values are different and, for this reason, a new calculation cycle is repeated from steps (a) to (d) and this procedure iterated until the calculated values of two successive iterations are very close, that is, different only for an established acceptable error. (f) Continue iteration calculations until a constant value is obtained for the variables Normally, two or three iterations are sufficient. Table 6.3 shows a series of iterative calculations relative to the data of Examples 6.2 and 6.4. It can be noted that the values become constant after the third iteration (bold), when four decimal numbers are used. The values of the partial pressures of CH4 and N2 show the slowest convergence. Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 183Maximizing sustainability by energy recovery in anaerobic treatment systems 6.5.2 High-sulphate concentration When the sulphate concentration is high, the methane production is affected because the organic material will be preferentially oxidized. In that case, the above iterative calculation procedure is extended in two aspects: (a) Methane production is now estimated by decreasing the availability of organic material for anaerobic digestion of 2 mg COD/mg S. For example, at a concentration of 100 mg/L of SO4–S or 120/1000 × 100 = 12 g S/inh/day: ′′ =+ −− − =× () () VT CH ti te xv SO MS MS MS MS 4 4 22 4273 273 26 4 24 6100 ./ / .( −−−−× =20 10 212177). .L/inh/day (b) Part of the formed hydrogen sulphide will pass to the gas phase, but the magnitude of this part depends on the ratio between the established biogas flow and the effluent flow. Figure 6.4 clearly shows that the fraction of hydrogen sulphide in the biogas depends on Qg/Ql ratio and pH. Once the sulphide fraction which migrates to the gaseous phase is established, the volume and the partial pressure of hydrogen sulphide can be calculated as well as the concentration of this component in the effluent. Example 6.7 Estimate for the data in previous examples the production of methane and biogas and the residual sulphide concentration in the liquid phase as a function of the sulphate concentration over the range from 0 to 200 mg S/L. Solution For each sulphate concentration, the problem requires an iterative calculation procedure that consists of the following steps: (1) Initially, it is assumed that the sulphate is converted to sulphide, the organic material demand to effect this redox process is estimated and the corresponding reduction in methane production is calculated (column 2 of Table 6.4): MS MSO MS MS MS MS MS oxid CH totti te xv oxid = =−−− () 2 4 4 4/ In reality, this estimate is imperfect because SRBs will grow and in this anabolic process SRBs, being heterotrophic, use organic material. At high-SO4 concentrations, this growth will not be insignificant because oxidative catabolism occurs, which generates much energy and thus anabolism is considerable. Table 6.3 Interactive calculations to establish the values of the volume and composition of biogas if there is little sulphate reduction. Iteration p CH4 M CH 4 QgQg/Qa p N 2 0 1 17.5 26.6875 0.222396 0.052616 10.8732 14.2500 21.7313 0.2074 0.0562 20.8697 14.2500 21.7313 0.2082 0.0559 30.8699 14.2500 21.7313 0.2082 0.0560 40.8699 14.2500 21.7313 0.2082 0.0560 Aq = 120 L/inh/day, temp = 25°C, MSinh = 100 g COD/day (70% digestion), methane production = 17.5 g CH4/inh/day, p CO 2 = 0.04, p HO 2 = 0.03. Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 184 Anaerobic Treatment of Domestic Wastewater: Present Status and Potentialities (2) Estimate the partial pressure of methane in the biogas using the values of the partial pressures of CO2, H2O, N2 and H2S. For N2 and H2S this will require an iterative calculation procedure. (3) Estimate the methane mass in the effluent as the difference between the total methane production and the mass that is transferred to the gaseous phase (column 4): MS MS MCH MS CH efl CH totCH eflCHefl 444 4 0020 120 4 4 , ,, . / =−×× = p (4) Estimate the methane volume corresponding to the daily mass of the produced methane in the gas phase (column 5): Vt CH CH totCHefl MS MS 44 4 22 4 273 273 64    ./ / , (5) With the aid of the estimates of the pressure and volume of methane, calculate the ratio between the gas and liquid flows, Qg/Ql (column 6): QQ Vp Q gl CH CH l // /  44 (6) Obtaining the Qg/Ql ratio, calculate the partial pressure of nitrogen and hydrogen sulphide (columns 7 and 8). (7) Knowing that the partial pressures of CO2 and water vapour do not depend on the flow of biogas, calculate the partial pressure of these two components. (8) Now, having the estimates for the partial pressures of CO2, H2O, N2 and H2S calculate a new estimate for the partial pressure of CH4. (9) With the aid of the new p CH 4 , for each sulphate concentration repeat steps 2–8 until the same result is obtained in two consecutive iterations. At the end of each iterative calculation the flow and composition of the biogas are obtained. Table 6.4 shows the results of the iterative calculations for different sulphate to sulphide conversions. Figures 6.6 and 6.7 are graphical presentations of Table 6.4. When analysing the values in Table 6.4 and Figures 6.6 and 6.7, the dramatic influence of sulphate reduction on the production of methane and biogas composition can be noted: the mass of the produced methane is reduced by a factor of 4 when the concentration of sulphate is increased from 0 to 200 mg S/L. For the admitted values of 100 g COD/inh/day and a per capita volume of 120 L/day the calculated COD of the influent is 100/120 = 0.833 g/L and a biodegradable concentration of 0.7 × 833 = 583 (70% efficiency). The data indicate that for a ratio S-SO4/COD in the influent greater than 200/583 ≈ 1/3 mg S/mg COD methane production is not feasible, which has also confirmed by the practical experience. Under these conditions, all the organic material is used for sulphate reduction. As for the estimation of concentration of hydrogen sulphide in the liquid phase, firstly the mass of sulphur in the biogas is calculated: for fixed influent concentration of 200 mg S/L, the partial pressure of H2S is 0.0275 and the concentration in the gas is 41 × 0.0275 = 1.15 mmol/L or 37 mg/L biogas. The ratio biogas/influent flow rate is Qg/Ql = 0.048 (column 6 of Table 6.4), thus the daily production per capita is 0.048 × 120 = 5.8 L/inh/day with a sulphur mass of 37 × 5.8 = 213 mg S/day. This is <1% of the influent sulphur mass (calculated by 213/(120 × 200)). For this reason, the mass of sulphur in the effluent, in good approximation, will be equal to the mass of sulphide generated in the anaerobic reactor or present in influent: 200 mg S/L. The concentration of 200 mg/L in the effluent will produce bisulphide and hydrogen sulphide gas. For a pH = 7.0 this ratio is equal, that is, C HSe 2 = CHS− = 100 mg/L. The concentration of hydrogen sulphide in the effluent and the reactor is very high: many researchers (Buisman & Lettinga, 1990; Rinzema, 1988) have shown that a Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 185Maximizing sustainability by energy recovery in anaerobic treatment systems concentration of 100 mg/L H2S severely inhibits methane production in anaerobic digesters. Thus, the production of methane could be even less than the calculations indicate since it was evaluated without considering the toxic effect of hydrogen sulphide gas. Table 6.4 Estimates of the production of biogas and the partial pressures of methane, nitrogen and hydrogen sulphide as functions of the reduced sulphate concentration for the same suppositions as in Table 6.3. 1 2 3 4 5 6 7 8 SO4 = mg S/L MSCH tot 4 g/inh/day pCH4 (dimensionless) MS CH g 4 g/inh/day VCH4 L/inh/day Qg/Qa (dimensionless) pN2 (dimensionless) pHS 2 (dimensionless) 017.5 0.873 15.40 23.54 0.1962 0.0569 0.0000 20 16.3 0.866 14.22 21.73 0.1811 0.0613 0.0026 40 15.1 0.858 13.04 19.93 0.1661 0.0663 0.0053 60 13.9 0.849 11.86 18.12 0.1510 0.0723 0.0079 80 12.7 0.840 10.68 16.33 0.1361 0.0794 0.0106 100 11.5 0.828 9.51 14.53 0.1211 0.0881 0.0134 120 10.3 0.815 8.34 12.75 0.1063 0.0988 0.0161 140 9.1 0.798 7.18 10.98 0.0915 0.1123 0.0189 160 7.9 0.778 6.03 9.22 0.0768 0.1301 0.0217 180 6.7 0.751 4.90 7.48 0.0624 0.1541 0.0246 200 5.5 0.714 3.79 5.79 0.0482 0.1881 0.0275 Figure 6.6 Influence of the sulphate concentration on the biogas production (see also Table 6.4). Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 186 Anaerobic Treatment of Domestic Wastewater: Present Status and Potentialities Example 6.8 A system composed of two UASB reactors, with a total volume of 7500 m3 treats municipal sewage. The water supply system is precarious and for that reason, the population tends to use also ground water, which is very rich in sulphate (1 g/L). As a result the average concentration of sulphate in the sewage is 200 mg/L (as SO4 =) or 67 mg S/L. Table 6.5 shows operational data. Evaluate the influence of sulphate on the methane production. Solution To estimate the effective production of biogas, it must be considered that the organic material of the influent is distributed into four fractions in the UASB reactor: (1) not removed and released into the effluent (2) transformed into volatile solids and released as such in excess sludge (3) oxidized in the redox process with sulphate reduction (4) digested forming methane. Therefore, one can calculate the methane production potential as: CH dtite xv ox4 44 ==−− − () SSSS S // (6.15) where CH4 = mass of the produced methane per L of influent (mg CH4/L); Sd = concentration of digested COD (mg COD/L); Sti = influent COD concentration (mg COD/L); Ste = effluent COD Figure 6.7 Influence of the sulphate concentration (reduced to sulphide) on production and composition of biogas (see also Table 6.4). Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 187Maximizing sustainability by energy recovery in anaerobic treatment systems concentration (mg COD/L); Sxv = COD transformed into volatile sludge in the UASB reactor (mg COD/L); Sox = oxidized COD by sulphate (mg COD/L). If there is 200 mg SO42−/L or 200/3 = 67 mg S-SO42−/L in the influent, with an oxidation capacity of 2 × 67 = 133 mg COD/L, it can prevent the production of 133/4 = 33 mg CH4/L. The Sxv concentration is estimated knowing that 1 g of volatile sludge has a COD of ∼1.5 g. According to Table 6.5, there is a production of 1200 kg Total Suspended Solids (TSS) sludge per day for an organic load of 6300 kg COD/day. Table 6.5 also shows that the anaerobic sludge has a volatile fraction of 2/3 so that the sludge production, in terms of Volatile Suspended Solids (VSS), is 2/3 × 1200 = 800 kg VSS/day, representing 1.5 × 800 = 1200 kg COD/day. As the applied load is 6300 kg COD/day the COD incorporated into the sludge is a fraction of 1200/6300 = 0.19 of the COD input, so it accounts for 0.19 × 650 = 124 mg/L. In the absence of sulphate in the influent, the expected anaerobic sludge production would be lower, since the sludge yield coefficient generated from the sulphate reduction is much greater (due to the oxidative catabolism). Hence, the methane production can now be calculated as: CH mg dta te xv ox4 44 650 307 124 133 4864 22 ==−− − =−−− == () () SSSS S // // CCH L 4/ However, not all of the methane produced will desorb: methane solubility in water is ∼20 mg/L when the partial pressure of methane in the gas phase is 1 atm. Therefore, in that case the concentration of methane that could desorb is only: CH CH mg/L g44 20 22 20 2=−=−= Knowing that the density of methane is ∼2/3 kg CH4/m3, the produced methane is calculated: V CH gCH CH mL Linfluent 44 15 3 4 =× =../ It can be concluded that there is practically no biogas production, not only due to the direct cause, the oxidation of organic material by sulphate, but also due to indirect reasons. The first one is that organic material is also lost to the higher sludge production. Another indirect effect of the presence of sulphate and the consequent reduction of biogas production is that the mixing of the bioreactor is reduced because the contribution of rising biogas bubbles is absent. The generation of biogas is actually the main driver of the reactor mixing and facilitates uniform distribution of influent organic material and contact between the biomass and the substrate. Since there is no contribution of biogas to mix Table 6.5 Operational data of an anaerobic treatment system for sewage with 200 mg/L of sulphate in the influent. Variable Unit Value Population Inh 90,000 Flow m3/day 9000 Organic load kg COD/day 6300 Influent COD concentration mg/L 650 Effluent COD concentration mg/L 307 Sludge production kg TSS/day 1200 Volatile fraction in the sludge mg SO4/L 0.67 Sulphate in the effluent ≈0 Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 188 Anaerobic Treatment of Domestic Wastewater: Present Status and Potentialities the reactor and thus to improve the contact between sludge mass and incoming organic material, it is conceivable that preferential channels will be formed, resulting in hydraulic short circuits. Under these conditions the fraction of the sludge present in stagnant zones receives little substrate and operates below its full capacity, whereas in other zones, where the flow of the liquid phase is accelerated, overload can occur. Both features tend to give a reduction of the organic material removal efficiency and particularly of methanogenesis. On the contrary, the concentration of 67 mg/L of sulphide itself can be toxic for methanogens, especially if the pH is slightly reduced, which again would lead to a decrease of methanogenic activity. It is concluded that the presence of 200 mg/L of sulphate has a disastrous effect on the efficiency of anaerobic sewage digesters. The problems extend beyond the performance of UASB reactors: post-treatment in either polishing ponds or aerobic reactors is very complicated. The high-sulphide concentration makes algae growth in polishing ponds almost impossible and exerts a large oxygen demand in aerobic reactors, thus increasing energy consumption. In fact, if aerobic post-treatment is applied, it would be preferable not to use the first anaerobic step and apply the aerobic treatment directly to the influent. Table 6.6 and Figure 6.8 show the dramatic influence of a high-sulphate concentration on anaerobic sewage treatment. Although without sulphate at least 65% of the influent COD is converted into methane with 11% remaining as dissolved methane in the liquid phase, with the high-sulphate concentration there is virtually no biogas production. 6.6 TREATMENT OF SULPHIDE PRODUCED IN ANAEROBIC DIGESTERS It is apparent from Figure 6.8 that the division of reduced sulphur mass depends primarily on the ratio Qg/Ql and also on the pH in the digester. As the ratio Qg/Ql depends on the concentration of the influent organic matter, this will be the main factor that determines whether the generated sulphide is found predominantly in the liquid phase or in the gas phase. Depending on the organic load, it can be necessary to remove the sulphide from the biogas or the effluent. These two alternatives will be discussed below. The removal of sulphide from contaminated gas above the liquid surface in the digester is also discussed. Figure 6.9 shows schematically the auxiliary units of the anaerobic digester to treat reduced sulphur that is generated in the anaerobic reactor. For the treatment of biogas three methods are distinguished: biological (Lin et al., 2018), physical and chemical (Zhang & Zhang, 2002). For removal of sulphide in the liquid phase the same methods can be applied. Table 6.6 Comparison between the performance of UASB reactors treating sewage with and without the presence of 67 mg S-SO42−/L in the influent, assuming 100% SO4 conversion. Mass of COD Sulphate: 67 mg S/L No Sulphate Assumptions kg COD/day Fraction kg COD/day Fraction In the influent 6300 16300 1 In the effluent 2980 0.47 1575 0.25 Efficiency = 75% In the sludge 1200 0.19 630 0.10 Sludge production = 10% Dissolved CH4720 0.11 720 0.11 20 mg/L of CH4 Oxidized 1200 0.19 00.00 2 kg COD/kg S In the biogas 200 0.03 3375 0.54 Methane mass (kg/day) 50 844 0.25 kg CH4/kg COD Vol CH4g (m3/day) 75 1265 1.5 m3/kg CH4 Electric energy (kW) 10 169 0.2 kW/(kg CH4/day) Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 189Maximizing sustainability by energy recovery in anaerobic treatment systems 6.6.1 Sulphide in biogas The most widely applied unit for hydrogen sulphide removal from biogas is probably the Thiopaques process in which the hydrogen sulphide is transformed into elemental sulphur. The process consists of three steps (see Figure 6.9): (1) In a counter current scrubber the biogas is placed in contact with an alkaline solution and the H2S in the biogas is absorbed and dissociates: HS OH HS HO 22 ++ → −− (6.16a) Figure 6.8 Distribution of the influent COD in different fractions with (left) and without (right) sulphate. Figure 6.9 Schematic representation of an anaerobic digestion system with possible auxiliary units for sulphide removal from biogas. Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 190 Anaerobic Treatment of Domestic Wastewater: Present Status and Potentialities (2) In a biological unit and a controlled aerobic environment HS− is oxidized to sulphur: HS OSOH −− +↓ +→05 2 . (6.16b) (3) The solid sulphur is separated from the liquid phase and becomes available as a nutrient or raw material. An alternative to remove smaller quantities of hydrogen sulphide in biogas is to add a small amount of oxygen to the mixture of biogas and then pass it through a unit operated with attached sulphideoxidizing bacteria able to convert aerobically the H2S in the biogas into sulphur. Physically, this biological unit can be located in the biogas accumulation zone above the liquid phase in the digester. This option is sometimes used for cleaning biogas in manure digesters, where biogas may be used directly to power generators for electricity production. Support media with a rough surface and high superficial area for better adhesion of biomass are employed for biogas treatment. The sulphide absorbed onto the solid surface accommodating the biomass will be oxidized by oxygen. The formed sulphur accumulates onto the solid surface and occasionally chunks of sulphur fall off and can be withdrawn for use of recovered sulphur or mixed with the effluent for agricultural reuse. Care must be taken not to add too much oxygen since it can make a potentially explosive mixture with the biogas when the volumetric ratio O2/CH4 is above 1:15–1:20. It is worth noting that the oxygen demand for the sulphide removal from the biogas is much smaller than this critical ratio. To inoculate biomass for the process, the solid surface can be covered with liquid manure, which always contains bacteria that oxidize hydrogen sulphide. The size of the hydrogen sulphide oxidation unit is small relative to the size of the digester. For a 1000 m3 digester, just a few square metres are enough to remove the hydrogen sulphide in the biogas. Physical removal of sulphides is possible because the hydrogen sulphide solubility in water increases with decreasing temperature. When the water vapour in the biogas has to be removed, the biogas may be cooled to ∼0°C: condensation of water will occur and hydrogen sulphide from the biogas will tend to dissolve in the formed liquid phase. To facilitate this dissolution, alkalinity is added to the condensed water. This alternative is attractive in regions with a cold climate, where ambient temperature is low for most of the year. In the chemical removal process (iron sponges process) the principle of operation is based on the reaction between iron oxide and hydrogen sulphide producing ferrous sulphide. In a reactor with partially oxidized iron filings the following reaction occurs: FeOHSFeS HO+→+ 22 (6.17a) When the removal capacity is exhausted, due to the conversion of ferrous oxide to ferrous sulphide, it can be regenerated by oxidation with air as indicated in Figure 6.10: FeSOFeOSO ++ →↑ 32 22 / (6.17b) Other oxidants may also be used for chemical treatment of the biogas, for example hypochlorite solutions, but in these cases, regeneration is usually not possible. 6.6.2 Sulphide removal from the effluent The methods employed to remove hydrogen sulphide from the effluent, as already mentioned, are biological, physical or chemical. A biological method consists of the oxidation of sulphide to sulphur with oxygen. A physical method is based on the application of a desorption unit to transfer the hydrogen sulphide dissolved in the effluent to air in an open tank or in a closed unit with air in counter current. For sulphide oxidation in an anaerobic reactor, it is possible to add oxygen directly to the reactor bulk phase through micro-aeration. In this case, a population of sulphide-oxidizing bacteria (Thiobacillus Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 191Maximizing sustainability by energy recovery in anaerobic treatment systems sp.) is established, and the sulphide in the liquid phase is oxidized to sulphur. The sulphur suspension is mixed with the sludge mass in the reactor and then discharged with the excess sludge. In this way, accumulation of a considerable concentration of sulphide in the liquid phase is avoided and thus indirectly the desorption of hydrogen sulphide to biogas will be limited. This method is particularly useful if much of the sulphur remains in the liquid phase, that is, when ratio Qg/Ql is small and when the sulphate concentration is high. An alternative is to operate a separate reactor where a low concentration of dissolved oxygen is maintained. Under these conditions a specific bacterial population (Thiobacillus sp.), which generates a suspend sulphur from sulphide will develop. Control of the dissolved oxygen concentration must be accurate to avoid oxidation to sulphate, which requires four times more oxygen (and therefore energy) and does not remove the sulphur from the liquid phase. Janssen etal. (1999) demonstrated the advantage of using the redox potential of the reactor to control the concentration of dissolved oxygen. If the oxygen transfer is carried out in such a way that there is not high turbulence, sulphur aggregates in large flocs that are removable by simple settling. Marked turbulence tends to generate a colloidal suspension and separation of the sulphur from the liquid phase may become problematic. The removal of the liquid phase can also be effected by desorption, a physical process taking place in a separate unit, where the digester effluent is transferred and placed in a counter current flux in the presence of air. This method is particularly effective when the influent has a low pH. In that case, mixing effluent (neutral pH) and influent (acid conditions) produces a mixture with low pH, in which hydrogen sulphide ion is converted into H2S. Like carbon dioxide, hydrogen sulphide gas is supersaturated in the mixture and it is transferred to the air phase. An industrial unit (tray aerator or scrubber) or simply an accumulation tank can be used to mix influent and effluent. Mechanical stirring or bubble aeration may be applied to accelerate desorption of volatile supersaturated compounds (CO2 and H2S). In case of a high-sulphide concentration in the effluent, the desorption unit may be closed to avoid spreading of bad odours. This implies that a solution must be given for the hydrogen sulphide transferred to the air. In the desorption unit there is also air oxygen transfer to the liquid phase of the mixture. This micro-aeration can facilitate sulphide oxidation in the digester itself when the influent and recirculation mixture is introduced. Sulphur in colloidal form, which may be formed, will be extracted with the effluent or with excess sludge. Chemical removal of sulphide from the liquid phase has been applied mostly for wastewaters with high-sulphate concentration (Song etal., 2004). Ferric chloride may be added to the influent and used Scrubber Biological reactorS -separat or Figure 6.10 Removal of H2S from biogas by absorption and partial oxidation. Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 198 Anaerobic Treatment of Domestic Wastewater: Present Status and Potentialities Screw press systems can directly dewater sludge without requiring any previous thickening stage, which could result in substantial reductions in capital and operating expenses. However, this is not an advantage for dewatering anaerobic sludge, as it does not require any thickening step, unlike conventional waste activated sludge. In contrast to other mechanical dewatering technologies, this technology exhibits lower dry solid content in the cake (17–25%). Nevertheless, it provides several benefits, including reduced energy consumption, lower operational and maintenance costs, compactness and decreased noise levels. 6.9.3.3 Belt filters Belt filters operate continuously and consist of three sections: a gravity drainage section, in which water is removed by gravity, a second section where moderate pressure is exerted, as the sludge is confined between the band cloths and finally, a high-pressure area where the sludge is subjected to higher pressures as the force exerted by the cloths and the rollers increases, thus generating a cake of dehydrated sludge. The solid content in these systems can achieve between 20 and 28% TSS. The energy requirement of this technology is between 10 and 25 kWh/t TSS, being lower than that required for centrifuges or filter presses. 6.9.3.4 Centrifuges Centrifugation process is widely used in STPs to separate excess water and generate a thickened cake with a higher solid content using bowl centrifuges. A polymer-conditioned sludge stream is constantly fed into a rotating bowl, where a solid cake and a diluted liquid stream are produced. The separated liquid, called centrate, must be returned to the STP water line, while the thickened sludge cake has a solids’ concentration often ranging between 20 and 30%. The energy requirement of this technology is between 30 and 60 kWh/t TSS. 6.9.3.5 Sludge-drying beds A sludge-drying bed is one of the oldest and simplest units for sludge dewatering and it has been applied to dewater the sludge purged from anaerobic sewage bioreactors (Figure 6.13). It gradually dewaters the sludge over the periods of several weeks by evaporation from the sludge surface in contact with the air and by drainage, using a system composed by a sand and gravel layer with drainage tubing located at the bottom of the bed. The collected water is usually recycled to the water line of the STP. The sludge to be dried is placed on the system in a 200–300 mm depth layer. Once dried, the solids are transported out of the STP to be used in agriculture or are landfilled. Conventional drying beds typically have rectangular shapes, with outer walls normally constructed with concrete blocks. Although the use of Figure 6.13 Sludge-drying bed in a municipal STP containing fresh sludge (left) and with dehydrated sludge (right). Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 199Maximizing sustainability by energy recovery in anaerobic treatment systems drying beds is associated with small plants in areas where cheap land is available and with hot or dry climates, it is possible to use them in temperate climates, always considering the rainfall in the area. Odour generation is usually very low when anaerobic sludge from UASB reactors is processed in drying beds since the sludge is stabilized as a result of the high sludge age in anaerobic bioreactors. Drying periods of 30–45 days can be applied, and the solids’ concentration generally lies between 30 and 45% and may achieve more than 50–60% TSS under dry climate conditions, attaining a partial sludge sanitization in terms of microbial indicators (Lopes etal., 2020). This technology requires more time and is less compact than mechanical devices. However, operational and maintenance costs are very low, its operation is simple and the percentage of solids in the dehydrated sludge is very high. 6.9.3.6 Sludge-drying lagoons Similar to drying beds, lagoons can be viewed as a cost-effective sludge dewatering technology in regions characterized by hot and dry climates, low population density and with cheap land availability. While it shares similarities to drying beds in terms of requiring vast land areas and labour-intensive mechanical removal of dewatered solids, its construction is simpler due to the absence of the need for filtrate drainage. However, drying lagoons are significantly less utilized than drying beds to treat digested sludge. The depth of the lagoon commonly ranges between 0.75 and 1.25 m. The main difference between this process and drying beds lies in the fact that evaporation is the primary mechanism driving the dewatering process, since no drainage can occur. This results in the need for longer drying times, ranging from several months to more than a year. Typically, the cake is removed with a concentration in the range of 25–30% of TSS (Table 6.8). Table 6.8 Comparison among the most common technologies used for dewatering waste sludge. Filter Press Screw Press Belt Filter Centrifuge Drying Beds Lagoons Consumption (kWh/t TSS) 20–40 5–15 10–25 30–60 – – Land occupied Low Low Low Low High High Cake solids (%) Up to 35 17–25 20–28 20–30 50–60 25–30 Operation mode Discontinuous Continuous Continuous Continuous Discontinuous Discontinuous Operational complexity High Intermediate Intermediate Intermediate Low Low Capital costs Intermediate Low Intermediate High Low Low Dewatering time Intermediate Short Short Short Long Very Long Maintenance High Low High Intermediate Low Low Chemical requirements High High High High Low or none – Odour problems – – – – Potential Potential Weather impact – – – – High High Groundwater pollution Low Low Low Low Intermediate High Noise Intermediate Low Intermediate High – – Source: Adapted from Von Sperling and Chernicharo (2005), and Metcalf and Eddy (2014). Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 200 Anaerobic Treatment of Domestic Wastewater: Present Status and Potentialities 6.10 ENERGY GENERATION FROM SLUDGE COMBUSTION Sludge generation in anaerobic treatment systems has a production of ∼0.15 kg TSS/kg COD applied, with an organic fraction of 55–70% and a concentration of 50–80 g TSS/L. The produced sludge may be separated from the water by mechanical processes (filtration, centrifugation), whereupon it becomes a semi-solid cake with a solid fraction of ∼20% and thus a residual humidity of ∼80%. If further drying is desired, thermal processes such as sludge-drying beds or external heat application must be applied. Depending on the use of the dried sludge and the cost of providing a final destination of the dried sludge it may be interesting to increase the solid fraction beyond the value that can be achieved by mechanical methods (15–25%). If the area for drying beds is not available, thermal drying may be applied, which is a process that produces a final sludge with no or very little water. It requires an external energy source, which can supply the heat required to evaporate the water from a sludge cake. Thermal drying can be completed with combustion of the dried sludge with air in an incinerator; in that case the energy source of the process is the combustion heat of the volatile sludge itself (see Figure 6.1). In that case, an important issue to be considered for sewage sludge incineration is the emission of toxic pollutants, due to the potential emission, among others, of dioxins, furans and nitrogen oxides, making the in-situ sludge incineration unaffordable. For this reason, centralized sludge incineration or co-incineration plants are used in many European countries, USA, and Japan. The possibility of producing electric energy from the dried sludge is limited based on the following analysis: in anaerobic treatment systems roughly 2/3 of the influent organic material are transformed into biogas and the remaining fraction (1/3) is equally divided between organic material residual in the effluent and organic material transformed into sludge, both accounting for a fraction of 1/6 of the influent organic load. Thus, the fraction of the organic material transformed into biogas (methane) is of the order of four times greater than the fraction transformed into volatile sludge ((2/3)/1/6) = 4). When it is desired to use energy to generate a usable product e.g. electric energy the difference becomes even greater for two reasons: (1) The biogas transformation efficiency is greater than that of sludge, because biogas is a gaseous fuel that can be used in an Otto-type explosion generator, characterized by an efficiency of 35–40%. In the case of sludge combustion, it is necessary to use a combination of a boiler to generate steam and a turbine to generate electricity. The efficiency of the boiler–turbine combination is at most ∼15%. Knowing that the chemical energy per unit mass of COD is ∼13.7 MJ, it is estimated that per unit mass of COD of the influent the potential of electricity generation is: 23 035040 13 734/. ... ××≈ () to MJ/kg COD whereas the energy generated from the sludge is no more than 16015137 034/. .. .××=MJ/kgCOD Therefore, it is estimated that the potential of electric energy generation from the biogas is ∼10 times higher than the one generated from the sludge combustion. (2) The sludge combustion energy is even lower because the sludge after drying still has residual humidity, which must be removed and it requires ∼2.5 MJ/kg H2O. Thus, if the sludge after drying has a humidity Ue and a volatile fraction fv, and knowing that fcv = 1.5 g VSS/g COD, it is calculated that the generation of the combustion heat is given by QUff cevcv MJkgCOD= () 1137 6–.// (6.20) Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 201Maximizing sustainability by energy recovery in anaerobic treatment systems On the contrary, the value to evaporate water from the dried sludge with a residual humidity Ue is given by QU f ee vMJ kgCOD=− () 25 61./ // (6.21) In Figure 6.14, the values of heat production by combustion (Qc) and the heat demand by water evaporation (Qe) are plotted as functions of the residual humidity Ue for organic sludge fractions of 60 and 70%. In the figure, an 80% heat utilization efficiency as well as the difference between the effective heat generated and the heat demand for evaporation are also reported. It can be seen that the difference is positive for a residual humidity of ∼60%, that is, the heat generated is sufficient to sustain sludge combustion when the residual moisture is <60%. This value can be easily obtained in drying beds, but not with the aid of mechanical methods of sludge dehydration. Example 6.10 Evaluate if the biogas production in a UASB reactor is sufficient to dry the produced sludge after mechanical dehydration to 80% humidity. Assume a sludge production of 0.1 g VSS/g COD, a VSS/ TSS ratio of 0.67 for the sludge and a digestion efficiency of 70%. Evaporation heat of water is 2.5 MJ/L and combustion heat is 13.7 MJ/kg COD Solution Since 1 g VSS has 1.5 g TSS and the humidity is 80%, the sludge production is 0.1 g VSS or 0.15 g TSS per g COD with 0.15/0.8 = 0.19 kg of wet sludge per kg COD. The required heat for evaporation of (0.19 − 0.15) = 0.04 kg H2O/kg COD is 0.04 × 2.5 = 0.1 MJ/kg COD. On the contrary, the digested COD is 0.7 g/g COD with a combustion heat for the formed methane of 0.7 × 13.7 ≈ 10 MJ/kg COD. However, only ∼50% of this biogas is actually captured and used, so that a more realistic energy production of 5 MJ/kg COD is estimated. It is concluded that the combustion heat of the biogas is ∼50 times greater than the energy demand for evaporation of the water in the sludge cake, so that even with low heat transmission efficiency, it should be possible to dry the sludge using only the combustion heat of the biogas to dry sludge, even though this does not seem to be applied in practice. Example 6.11 Evaluate if the heat production of sludge combustion from a UASB reactor is enough to have autosufficiency in an incinerator. Adopt the same assumptions as in the previous example. Solution The required heat for evaporation is the same as in the previous example. The combustion heat depends on the volatile fraction of the sludge. In Figure 6.13, the evaporation heat and the combustion heat are -1 -0.5 0 0.5 1 1.5 2 2.5 3 0.8 taeh/JM(OQDgk) Volale fracon f v = 60 % -1 -0.5 0 0.5 1 1.5 2 2.5 3 00.2 0.40.6 0. 8 Heat (MJ/kgDQO) Residual humidity Volale fracon f v = 70 % difference 0 0.2 0.4 0.6 Residual humidity difference 80 % Combustion heat Figure 6.14 Evaporation and combustion energy upon combustion of sludge cake as a function of humidity. The energy balance for 80% efficiency is also indicated, showing the need for external energy for humidity of  < 60%. Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 202 Anaerobic Treatment of Domestic Wastewater: Present Status and Potentialities shown as functions of sludge humidity for volatile sludge fractions of 0.6 and 0.7. In Figure 6.13 also shown is the energy balance for an assumed energy transmission efficiency of 80%. The figure shows that under these conditions the maximum humidity of the sludge is ∼60%, depending on the VSS/ TSS ratio, which is much less than the humidity obtained from mechanical dehydration, so that some external energy must be supplied, if it is desired that completely dry sludge is produced for maximum heat production by combustion. This external energy can be biogas in which case the energy balance is positive even for a very high initial humidity of the sludge. Example 6.12 Dry sludge is produced, and combustion is applied to generate steam to feed a turbine to produce electric energy. Estimate the production potential per capita using a top-quality boiler–turbine combination, which has an efficiency of 16%. The combustion heat of volatile sludge is 13.7 MJ/kg COD = 20 MJ/kg VSS. Volatile sludge production is 0.1 g VSS/g COD. Per capita contribution: 100 g COD/day. Solution The volatile sludge production per capita is 100 × 0.1/1000 = 0.010 kg VSS/inh/day with a per capita combustion heat of 20 × 0.01 = 0.2 MJ/day = 2.3 W. If the conversion efficiency is 16%, the per capita production potential is 0.16 × 2.3 = 0.37 W. This production potential is much smaller than the potential from gas, not only because the organic material is much smaller (16% of the influent COD for volatile sludge against 68% for methane), but also because the conversion efficiency of the solid fuel (16% for dry sludge) is much smaller than that for the gaseous fuel (40% for methane). 6.11 NUTRIENT RECOVERY FROM SLUDGE AND WASTEWATER STREAMS Phosphorus and nitrogen compounds are essential nutrients required by modern agriculture to improve yields. The world’s population increased from 2.5 billion inhabitants in 1950 to 8 billion at the end of 2022, and it is expected that it will reach 9.8 billion by around 2050. A large fraction of the phosphorus used in agriculture worldwide is obtained from a non-renewable resource, phosphate rock, mostly mined in Morocco (including Western Sahara), China, the USA, and Russia. In contrast, the production of nitrogen fertilizers is not restricted to a small number of countries and depends on the availability of energy to produce them through the Haber–Bosch and Ostwald processes. A large part of the developing countries is currently facing a great challenge: the increase in the prices of phosphorus fertilizers is limiting their use in agriculture and impacting food production in these regions (Brownlie etal., 2023). Thus, the recovery of nutrients from STPs and other waste resources would be important to guarantee the necessary food for the entire world population. Example 6.13 Estimate the phosphorus and nitrogen daily loads in a 90,000 inhabitants anaerobic STP and determine the fractions associated with the treated effluent and the anaerobic sludge generated. Consider a generation of 100 g COD/day, 10 g total nitrogen (TN)/day and 1.2 g total phosphorus (TP)/day and 100 L/day of sewage produced. Sludge yields 0.1 g TSS/g COD removed, COD removal percentage is 80. Nutrient content in the sludge is 0.05 kg N/kg TSS and 0.015 kg P/kg TSS. Solution Considering the above indicated figures, the amount of the incoming mass of TN and TP present in the raw sewage is (daily basis) TN inhgTN inh/daygTN/dayorkgTN/day in =× =90 000 10 900 000 900,/, TP inhgTP/inh/daygTP/dayorkgTP/day in =× =90 000 12 108 000 108,. , Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 203Maximizing sustainability by energy recovery in anaerobic treatment systems For anaerobic sewage treatment, the only mechanism contributing to N and P removal from wastewater is associated with the sludge formation. Thus, N and P associated with sludge generation should be: MS gCOD removedg COD fedgTSS/gCOD removed X= () × () × () ×08 01 90 0..,000 720 000 720 inh gTSS/daykgTSS/day=≈, TN kgN/kg TSSkgTSS/daykgTN/day X =×=005720 36. TP kgP/kg VSSkgTSS/daykgTP/day X =×=0012 720 10 8.. where MSX = anaerobic sludge generation (kg TSS/day); TNSV = total nitrogen in the sludge (kg N/ day); TPXV = total phosphorus in the sludge (kg P/day). The percentage of TN and TP associated with the sludge formation represents 4 and 10% of the daily nutrient load, respectively. In those STPs using anaerobic technologies, the recovery of phosphorus or nitrogen associated with the sludge is low, due to the low biomass yield of the process. In any case, the sludge generated in anaerobic STPs contains all the nutrients removed from the treated sewage. In fact, the use of sludge from STPs as agricultural fertilizer is a common practice in many countries around the world. Anaerobic sewage treatment is not the adequate solution to remove nutrients from the water phase. In recent years, many studies have tried to improve the removal of nutrients in these systems. One approach is the promotion of struvite precipitation, a phosphate salt containing ammonia and magnesium ions (MgNH4PO4·6H2O). This strategy is only effective when the treated wastewater contains high concentrations of orthophosphate ions, as is often the case with black water. The precipitation of struvite only compensates if the phosphorus concentration greatly exceeds 100 mg P/L, a level far from that of 5–15 mg P/L observed in sewage. In Table 6.9, the calculated concentrations of pollutants of raw and treated sewage for the analysed scenario are shown. As indicated, the calculated nutrient concentrations in the effluent and raw sewage are similar. In fact, there are many studies in the literature indicating that the observed removal of nitrogen and phosphorus in anaerobic bioreactors is almost negligible. Most of the nutrients contained in anaerobically treated sewage, without further post-treatment, remain in the effluent. The eutrophication of the receiving water bodies is one of the main problems of the discharge of anaerobically treated sewage. Nutrients contained stimulate algae blooms, harming water life. An alternative to take advantage of the nutrients and the water itself could be to store the treated waters in accumulation ponds, and use them for agricultural irrigation, as long as the distance between the fields and the STP does not make this application unfeasible. Water reuse is treated in Chapter 8, and could be a way to recover nutrients, and diminishing the use of chemical fertilizers. This practice, as well as the use of sewage sludge as fertilizer, is regulated by law in several countries. The World Health Organization publishes guidelines for the safe use of wastewater in agriculture (WHO 2006), to reduce the risks associated with the presence of pathogens and certain chemical compounds associated with wastewater and sewage sludge. For these reasons, in most cases, anaerobically treated sewage should undergo post-treatments to ensure adequate water quality, including a partial reduction of nutrients, especially nitrogen compounds, to guarantee a good food crop. Table 6.9 COD, TN and TP concentrations and nutrients loads of the analysed scenario. mg COD/L mg TN/L mg TP/L kg TN/day kg TP/day Raw 833.3 83.3 10.0 900 108 Effluent 166.7 80.0 9.0 864 97 Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 204 Anaerobic Treatment of Domestic Wastewater: Present Status and Potentialities One of the questions that could be raised is the comparison of the amounts of nutrients present in sewage, with the amount of fertilizer consumed. According to the International Fertilizer Association (IFA, 2023) during 2020, 21.8 and 112.3 Mt of phosphorous- and nitrogen-based fertilizers were consumed worldwide, which implies an average annual consumption per world inhabitant of 2.8 kg P and 14.4 kg N. It can be quickly calculated for our case study that an inhabitant generates 0.438 kg P and 3.65 kg N annually. Assuming that all nutrients from wastewater could be effectively reused, ideally 15% of P and 25% of N could be recovered, thereby reducing the consumption of chemical fertilizers. This is obviously a very simple calculation; surely the real figures that can be achieved are more modest, highlighting the great potential associated with the nutrients contained in urban wastewater. 6.12 CONCLUDING REMARKS It has been shown that anaerobic wastewater systems can be operated without the need of external energy, even when wastewater is diluted such as the case of sewage. The required energy for anaerobic treatment is obtained by using its organic material as an energy source. In the case of sewage ∼2/3 of the influent organic material is converted into methane and may be used for energy generation. About 1/6 of the influent organic material is converted into organic sludge and 1/6 is discarded in the effluent. Organic sludge may be incinerated to produce electric energy and heat, but the alternative of using it as an organic fertilizer (after drying) may be more attractive from the economic point of view. Nitrogen and phosphorus recovery from anaerobically treated wastewater can be carried out by using the reclaimed water and sludge generated in agriculture, according to local regulations. In most cases anaerobically treated sewage should undergo further treatment to ensure adequate water quality to guarantee a good crop. REFERENCES Braga A. F. M., Zaiat M., Silva G. H. R. and Fermoso F. G. (2017). Metal fractionation in sludge from sewage UASB treatment. Journal of Environmental Management, 193, 98–107, https://doi.org/10.1016/j. jenvman.2017.01.070 Brasil (2006). Conselho Nacional do Meio Ambiente. Resoluçao n 375. Brownlie W. J., Sutton M. A., Cordell D., Reay D. S., Heal K. V., Withers P. J. A., Vanderbeck I. and Spears B. M. (2023). Phosphorus price spikes: a wake-up call for phosphorus resilience. Frontiers in Sustainable Food Systems, 7, 1–8, https://doi.org/10.3389/fsufs.2023.1088776 Buisman C. J. N. and Lettinga G. (1990). Sulphide removal from anaerobic waste treatment effluent of a papermill. Water Research, 24(3), 313–319, https://doi.org/10.1016/0043-1354(90)90006-R Cooper C. D. and Alley F. C. (2011). Air Pollution Control: A Design Approach, 4th edn. Waveland Press, Long Grove, IL. Council Directive 86/278/EEC of 12 June 1986. European Union limits for metals in STP sludge. Cozma P., Wukovits W. and Mamaliga J. (2013). Analysis and modelling of the solubility of biogas components in water for physical absorption processes. Environmental Engineering and Management Journal, 12(1), 147–162, https://doi.org/10.30638/eemj.2013.017 De la Varga D., Díaz M. A., Ruiz I. and Soto M. (2013). Heavy metal removal in an UASB-CW system treating municipal wastewater. Chemosphere, 93, 1317–1323, https://doi.org/10.1016/j.chemosphere.2013.07.043 dos Santos S. L. (2013). Pós-tratamento de esgoto digerido visando reuso agrícola ou proteção de corpos receptores. PhD thesis, Federal University of Campina Grande (in Portuguese). Geng Y., Zhang C., Zhang Y., Huang D., Yan S., Sun T., Wang J. and Mao Y. (2021). Heavy metal(loid)s in sewage sludge in China: concentrations and spatial-temporal variations. Environmental Science and Pollution Research, 28, 29146–29156, https://doi.org/10.1007/s11356-021-12762-8 Grigoratos T. and Martini G. (2014). Non-exhaust Traffic Related Emissions – Brake and Tyre Wear PM. EUR 26648. Publications Office of the European Union, Luxembourg (Luxembourg). JRC89231. Henze M. and Harremoes P. (2001). Wastewater Treatment: Biological and Chemical Processes. Springer, Berlin. Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 205Maximizing sustainability by energy recovery in anaerobic treatment systems IFA, International Fertilizer Association (2023). https://www.ifastat.org/ (accessed on 23 May 2023) Janssen A. J. H., Lettinga G. and de Keizer A. (1999). Removal of hydrogen sulphide from wastewater and waste gases by biological conversion to elemental sulphur: colloidal and interfacial aspects of biologically produced sulphur particles. Colloids and Surfaces A: Physicochemical and Engineering Aspects, 151(1–2), 389–397, https://doi.org/10.1016/S0927-7757(98)00507-X Lens P. N. L. and Hulshoff Pol L. W. (2000). Environmental Technologies to Treat Sulphur Pollution. IWA Publishing, London (UK). Lin S., Hamish H. R., Tianwei H., Guo G., van Loosdrecht M. and Guanghao Chen C. (2018). Biological sulfur oxidation in wastewater treatment: a review of emerging opportunities. Water Research, 143(15), 399–415, https://doi.org/10.1016/j.watres.2018.06.051 Loewenthal R. E. and Marais G. V. R. (1976). Carbonate Chemistry of Aquatic Systems: Theory and Application. Ann Arbor Science Publishers, Michigan (USA). Lombardi A. T. and Garcia O. (2002). Biological leaching of Mn, Al, Zn, Cu and Ti in an anaerobic sewage sludge effectuated by Thiobacillus ferrooxidans and its effect on metal partitioning. Water Research, 36, 3193–3202, https://doi.org/10.1016/S0043-1354(02)00008-8 Lopes B. C., Rodrigues H. F., Costa T. E. R., Batista A. M. M., Filho C. R. M., Araújo J. C. and de Matos A. T. (2020). Air-drying bed as an alternative treatment for UASB sludge under tropical conditions. Journal of Water, Sanitation and Hygiene for Development, 10, 458–470, https://doi.org/10.2166/washdev.2020.041 Metcalf & Eddy (2014). Wastewater Engineering: Treatment and Resource Recovery, 5th edn. McGraw Hill, New York (USA). Rinzema A. (1988). Anaerobic Treatment of Wastewater with High Concentrations of Lipids or Sulphate. Doctoral thesis, Wageningen Agricultural University, Wageningen, The Netherlands. Song Z., Williams C. J. and Edyvean R. G. J. (2004). Treatment of tannery wastewater by chemical coagulation. Desalination, 164(3), 249–259, https://doi.org/10.1016/S0011-9164(04)00193-6 Souza L. C. F., Canteras F. B. and Moreira S. (2014). Analyses of heavy metals in sewage and sludge from treatment plants in the cities of Campinas and Jaguariúna, using synchrotron radiation total reflection X-ray fluorescence. Radiation Physics and Chemistry, 95, 342–345, https://doi.org/10.1016/j.radphyschem.2013.01.025 Van Haandel A. C. and van der Lubbe J. (2019). Anaerobic Sewage Digestion: Theory and Applications. International Water Association, London, UK. Von Sperling M. and Chernicharo C. A. D. L. (2005). Biological Wastewater Treatment in Warm Climate Regions. IWA Publishing, London (UK). WHO Guidelines for the safe use of wastewater, excreta and greywater (2006). World Health Organization. ISBN: 92 4 154686 7. Zhang S. and Zhang C. Z. (2002). Novel properties of ionic liquids in selective sulphur removal from fuels at room temperature. Green Chemistry, 4, 376–379, https://doi.org/10.1039/b205170m Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024 Downloaded from http://iwaponline.com/ebooks/book/chapter-pdf/1418711/9781789063479_0167.pdf by guest on 30 May 2024