“A person who never made a mistake never tried anything new.”
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TABLE OF CONTENTS
ABSTRACT ... 3
CHAPTER 1 : THE ACTIVATED SLUDGE TREATMENT PROCESS ... 6
1.1 INTRODUCTION ... 6
1.2 ACTIVATED SLUDGE PROCESS ... 7
1.3 BIOLOGICAL TREATMENTS OF COD AND NUTRIENT REMOVAL ... 14
1.4 THE PROBLEM OF THE EXCESS SLUDGE PRODUCTION ... 17
References ... 22
CHAPTER 2 : OBJECTIVES AND THESIS STRUCTURE ... 24
2.1 MOTIVATION AND OBJECTIVES ... 24
2.2 THESIS STRUCTURE ... 25
CHAPTER 3 : MATERIALS AND METHODS ... 26
3.1 EXPERIMENTAL SET-UP ... 26
3.2 ANALYTICAL METHODS ... 29
3.3 ACTIVATED SLUDGE MODELLING ... 29
References ... 30
CHAPTER 4 : ANAEROBIC MEMBRANE BIOREACTOR SYSTEM (ANMBR) ... 31
4.1 INTRODUCTION ... 31
4.2 OBJECTIVES ... 36
4.3 MATERIALS AND METHODS ... 37
4.4 EXPERIMENTAL RESULTS ... 44
4.5 CONCLUSIONS ... 52
References... 53
CHAPTER 5 : OXIC-SETTLING-ANAEROBIC PROCESS (OSA)... 56
5.1 INTRODUCTION TO THE OSA PROCESS ... 56
5.2 OBJECTIVES ... 65
5.3 MATERIALS AND METHODS ... 65
5.4 ACTIVATED SLUDGE MODELLING ... 81
5.5 EXPERIMENTAL RESULTS AND DISCUSSION ... 93
5.6 MODELLING RESULTS ... 113
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References... 133
CHAPTER 6 : CONCLUSIONS AND RECOMMENDATIONS ... 140
APPENDIX 1: PROJECT OF OSA PILOT PLANT ... 143
APPENDIX 2: CHARACTERIZATION OF TERENZANO WWTP ... 152
APPENDIX 3: VALIDATION OF THE MODEL WITH BIOWIN SOFTWARE ... 154
LIST OF ABBREVIATIONS ... 158
LIST OF FIGURES... 160
LIST OF TABLES ... 164
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ABSTRACT
The main costs in wastewater treatment are the sludge disposal and the supply of O2 in
aerobic processes. Various approaches are being proposed to reduce the excess sludge production in bioreactors of wastewater treatment plants (WWTP): among others OSA (Oxic-Settling Anaerobic) technology showed to reduce the excess sludge production till 60%. Anaerobic processes, achieving high organic matter removal efficiencies and biogas generation, without O2 need, realise also a low biomass
production. One example of them is the Anaerobic Membrane Bioreactor (AnMBR) technology, which combines an anaerobic biological treatment with a membrane separation process. This PhD project focused mainly the OSA technology, developed at Trieste University with the collaboration and financial support of a wastewater treatment company, but also consider the AnMBR technology that was developed at the Barcelona University. The latter was realised from October 2014 to March 2015, working with a laboratory scale AnMBR, fed with winery synthetic wastewater at low temperatures as novelty (from the start-up at 35°C, later at 25°C and finally at 15°C), and evaluating its removal efficiency and methanogenic activity. The COD removal was of 80% and 71% at 25°and 15°C, respectively and without suspended solids in the effluent. The quality of the effluent exceeded very often the legal requirement due to the VFA accumulation and the methane retained in the liquid phase. The methanogenic activity decreased at low temperatures as expected although SMA obtained at 25°C was similar to that at 35°C. On the other hand, at 15°C the activity decreased significantly. Reducing the temperature the microbial population changed from Methanosaeta to Methanosarcina, because of the higher amount of VFA in the AnMBR at lower temperatures: this contributed the development of an acetoptrophic methanogen with a higher growth rate under high acetate concentration. The operation at low temperatures needs further research, especially in terms of microbiology, because the slow growing anaerobic microorganisms would be subjected to seasonal changes of temperatures and organic load. Therefore, the most adequate inoculum should be identified to reduce the start-up periods and the procedure to cope these temperatures variations also needs to be determined. At low temperatures this technology leads to a loss of methane dissolved in the liquid phase. A post-treatment could be considered to recover it for economic and environmental reasons.
4 research is the use of a real wastewater fed in an OSA pilot plant. Indeed, in recent years almost all the studies report the feasibility of the OSA process in laboratory-scale plant fed with synthetic wastewater, whereas very few experimental works are related to OSA pilot plants using real wastewater.
The research concerned a first period (1st and 2nd year) with the project and construction of an automated OSA pilot plant, located close to the CAFC wastewater treatment plant (12,000 PE) of Terenzano, a village about 70 km from Trieste, and fed with the same wastewater of the real plant. In the last part of the study (3rd year), the OSA pilot plant was studied (and modelled), by applying respirometric techniques, evaluating the quality of the effluent, the sludge reduction feasibility and the effect on biomass activity. During the start-up phase the pilot plant operated as Conventional Activated Sludge (CAS) system, with sludge inoculum coming from the Terenzano plant. The CAS scheme consisted of a reservoir tank of 0.3 m3, an aerobic reactor of 0.5 m3 and a settler of 0.8 m3. The influent flowrate was of 1.44 m3d-1 with a hydraulic retention time of 8 h. Samples of influent, effluent and from aerobic reactor were analysed two times a week for TCOD, sCOD, NH4+-N, TSS, VSS. After 3 months the pilot plant
configuration was adapted to the OSA system, connecting an anaerobic reactor of 2 m3 into the sludge return line and monitoring the plant by calculations of sCOD and NH4+-N
removal efficiencies. Respirometric tests were done to estimate the required kinetic parameters for plant modelling: the used biomass samples (activated sludge) were taken from the pilot plant aerobic reactor (working as CAS and as OSA) and from the anaerobic reactor of the OSA. A model based on ASM3, was calibrated using the respirometry results and AQUASIM software. The model, with the different kinetic parameters for CAS pilot plant and OSA pilot plant, was implemented and validated into the BioWin software, simulating both periods of the pilot plant.
The results showed that the COD and ammonia removals remained high during the CAS and OSA conditions, even if for the OSA system both removal efficiencies decreased slightly compared to those of the CAS pilot plant. Better removal efficiency of the pilot plant (as CAS and OSA system) was obtained working with high organic loading. Moreover, the insertion of the sludge holding tank did not worse the settling properties of the whole system.
5 mechanism of the excess sludge reduction in this study was due to the enhanced endogenous decay.
In this research a simplified ASM3-based model, with 8 wastewater components and 12 biological processes was used. The calibration of the model was conducted by using the respirometric assays of the pilot plant that consisted of OUR, AUR and NUR test runs (calibration data-set). Parameter optimization and the numerical solution of model equations were performed using the Aquasim program (Reichert, 1998).
11% yield reduction efficiency was achieved in the OSA pilot plant process compared to the CAS pilot plant system. However, the CAS system operated with a continuous influent flowrate only for one month (start-up phase) and, probably, it was not in steady state conditions when the respirometric assays were conducted.
The calibrated mathematical model was implemented into BioWin software and then validated using a data set of 1 month in case of the CAS system and 3 months for OSA process. The predicted profiles versus time reflected enough the operating performance of the pilot plant operating as CAS and OSA process.
The last batch experiment of the OSA pilot plant confirmed that the insertion of the sludge holding tank promoted the slowly growth of the PAOs. On the other hand, this event was evidenced in the final part of the research and, therefore, further investigations should be taken into account in the future. If this phenomenon will persist, it should be better to perform supplementary respirometric tests in order to calibrate a different model more accurate than the mathematical one proposed in this study, as could be the ASM2d.
In order to provide more conclusive data on the efficiency and feasibility of the OSA process, the pilot plant should be continuously operated for as long as possible to find the best operational conditions for the sludge reduction.
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CHAPTER 1 : THE ACTIVATED SLUDGE TREATMENT PROCESS
1.1 INTRODUCTION
Municipal wastewater contains organic compounds, such as carbohydrates, proteins and fats, nitrogen in the form of ammonia and phosphorus in the form of phosphate from human waste and some detergents (now forbidden in Italy). Municipal wastewater also contains other types of particulate and dissolved matter and all these components have to be dealt with a wastewater treatment plant. Wastewater treatment plant is a plant where the wastewater is collected with the objective to reduce the contaminant concentration under the regulatory legislation limits. The Italian legislation corresponds to “Testo Unico Ambientale, Decreto Legislativo n.152/2006” (Table 1.1 and Table 1.2) and its aims are the following:
Organic matter removal (BOD, COD);
Nutrients removal as nitrogen and phosphorus; Colloidal substances removal;
Pathogens microorganisms decrease.
In general a wastewater treatment plant is divided in two sections:
A water line, where the contaminants concentration can be removed;
A sludge line, where the by products from wastewater treatment plants, such as screenings, grit and sewage sludge can be also treated.
There are three main levels of wastewater treatment:
Primary (mechanical) treatment is designed to remove gross, suspended and floating solids from raw sewage. This level is sometimes referred to as “mechanical treatment”, although chemicals are often used to accelerate the sedimentation process. Primary treatment can reduce the BOD of the incoming wastewater by 20-30% and the total suspended solids by some 50-60%. Primary treatment is usually the first stage of wastewater treatment;
7 pond and constructed wetland systems, trickling filters and other forms of treatment which use biological activity to break down organic matter;
Tertiary treatment can remove more than 99% of all the impurities from sewage, producing an effluent of almost drinking-water quality. An example of a typical tertiary treatment process is the modification of a conventional secondary treatment plant to remove additional phosphorus and nitrogen.
Table 1.1 Emission limits for municipal wastewater treatment plant- D.Lgs. 152/ 2006.
P.E. 2000-10000 ˃10000
Contaminants Concentration
[mgL-1] Removal efficiency [%] Concentration [mgL-1] Removal efficiency [%]
BOD5 without
nitrification ≤25 70-90 ≤25 80
COD ≤125 75 ≤125 75
Suspended solids ≤35 90 ≤35 90
Table 1.2 Emission limits for municipal wastewater treatment plant delivered in sensible area- D-Lgs.152/2006.
P.E. 10000-100000 ˃100000
Contaminants Concentration [mgL-1] Removal efficiency [%] Concentration [mgL-1] Removal efficiency [%]
Total phosphorus ≤2 80 ≤1 80
Total nitrogen ≤15 70-80 ≤10 70-80
1.2 ACTIVATED SLUDGE PROCESS
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1.2.1 Process description
Wastewater is introduced into a reactor where an aerobic bacterial culture is maintained in suspension. The aerobic reactor provides suitable environment for a mixture of bacteria and other microorganisms to aerobically metabolize the biodegradable contaminants in the incoming wastewater. Oxygen is supplied to the activated sludge by diffused air. After a specified period of time, the mixed liquor (MLSSV) flows to a settling tank. The settling tank is an integral part of the activated sludge system. It has two main functions: it separates the biomass from the water in order to produce a good quality of the effluent and it thickens the biomass. Part of the thickened biomass is returned (Figure 1.1) to the biological aerobic reactor to maintain an appropriate biomass concentration (MLSSV). Another portion of the thickened solids is removed daily or periodically, because the process produces excess biomass that would accumulate with the non-biodegradable solids contained in the influent wastewater. The operation of the settling tank is crucial for the whole treatment plant (Gerardi and Wiley, 2002).
Figure 1.1 Schematic diagram of the conventional activated sludge process. 1.2.2 Operation of Activated Sludge Process
9 appropriate limit to maintain the microorganisms active, and ensure that the reactor contents are appropriately mixed to keep the solids in suspension. The return activated sludge (RAS) rate is a critical control variable as it redistributes the sludge between the settling tank and the aerobic reactor, such that the healthy population of biomass is maintained in the aerobic reactor. If the RAS rate is too low, solids can remain in the settling tank, resulting in solids loss. If the rate is too high, the aerobic reactor can become hydraulically overloaded, causing reduced aeration time and poor performance.
1.2.3 Microorganism types in biological wastewater treatment
In municipal wastewater treatment incorporating biological nutrient removal, two basic categories of organisms are of specific interest: the heterotrophic organisms and the autotrophic nitrifying organisms, including the ammonia and nitrite oxidizers. The former group utilizes the organic compounds of the wastewater as electron donor and either oxygen or nitrate as terminal electron acceptor, depending on whether or not the species is obligate aerobic or facultative: the heterotrophic organisms obtain their energy and material requirements from the same organic compounds. In contrast, the latter group, which are autotrophs, obtains its energy and material requirements from different inorganic compounds: the energy from oxidizing ammonia in the wastewater to nitrite and nitrate and the material from dissolved carbon dioxide in the water. Being obligate aerobic organisms, only oxygen can be used as an electron acceptor and therefore the nitrifying organisms require aerobic conditions. In activated sludge plants, the heterotrophic organisms dominate more than 98% of the active organism mass in the system. Therefore, in terms of sludge production and oxygen or nitrate utilization, the heterotrophs have a dominating influence on the activated sludge system. In contrast to the autotrophs, the heterotrophs obtain the energy requirements and material from the same organic compounds, irrespective of the type of external terminal electron acceptor. This difference in the metabolism of the autotrophic and heterotrophic organisms is the principal reason why the cell yield, the organism mass formed per electron donor mass utilized (see section 1.2.3.2), is low for autotrophs (0.10 mg VSS mg-1 NH4+-N nitrified)
10 Bacteria have attracted the greatest attention in microbiology because they are easy to cultivate and to study. Fungi are also important in the treatment of some industrial wastewaters and in composting of solid wastes. Protozoans are found in almost every aquatic environment and are widely distributed: they play an important role in the treatment of wastewater. Also rotifers, crustaceans and worms live on bacteria, algae and small protozoans (Vilaseca et al. 2001).
1.2.3.1 Microbial metabolism
Metabolism is the sum of biochemical transformations included the relationship between anabolic and catabolic reactions and the behavior of the microbial culture determined by anabolism and catabolism. Catabolism is a reaction series that reduces the complexity of organic compounds with the release of energy. Anabolism involves the use of energy produced during catabolism to synthesize cellular materials. Energy transfer between these two reactions takes place in the form of adenosine triphosphate (ATP). Metabolism refers to the totality of organized biochemical activities carried out by an organism.
The bacterial kinetics of microorganism growth regulate the substrate utilization and the production of new biomass that increases the total suspended solids inside the reactor. The organic substrate concentration of the wastewater can be determinate through two parameters, as COD (Chemical Oxygen Demand) and BOD (Biochemical Oxygen Demand), that measure the amount of oxygen needed to oxidize, chemically (the first one) or biologically (the second one) the organic carbon. Often, the activated sludge amount inside the reactor is measured as Volatile Suspended Solids (VSS). The Total Suspended Solids (TSS) include both the inert inorganic compounds and the non-biodegradable fraction, which can be determined by subtracting the VSS, which represent the organic fraction, from the TSS.
The net microorganism growth rate “𝑟𝑥” is often represented as a combination of the biomass growth𝑟𝑥𝑔(see equation 1-2) and the biomass decay 𝑟𝑥𝑑 (see equation 1-6):
𝑟𝑥= 𝑟𝑥𝑔− 𝑟𝑥𝑑 1-1
11 𝑟𝑥𝑔 = 𝜇 =
𝜇𝑚𝑆
𝐾𝑠+𝑆 X 1-2
where
𝜇 = specific biomass growth rate; 𝜇𝑚 = maximum specific growth rate;
𝑋 = biomass concentration;
𝑆 = growth limiting substrate concentration in solution; 𝐾𝑠 = half saturation coefficient.
In Monod’s model, the growth rate is related to the concentration of a single growth-limiting substrate through the parameters 𝜇𝑚 and 𝐾𝑠. In addition to this, Monod also related the yield coefficient (𝑌𝑥 𝑠⁄ ) (equation 1-3) to the specific rate of biomass growth
(𝜇) and the specific rate of substrate utilization (q) (equation 1-4).
𝑌𝑥 𝑠⁄ = 𝑑𝑋𝑑𝑆 1-3
where X is the biomass and S the substrate. 𝜇 =𝑌𝑥 𝑠⁄
𝑋 𝑑𝑆
𝑑𝑡 ≅ 𝑌𝑥 𝑠⁄ 𝑞 1-4
The heterologous substrate concept assumes that the growth rate can be affected simultaneously by more than one substrate. A “Double Monod” model (equation 1-5) originally proposed by Mc Gee et al. (1972) was used to describe this phenomenon. 𝜇 = 𝜇𝑚𝑎𝑥 𝑆1
𝐾1+𝑆1 𝑆2
𝐾2+𝑆2 1-5
where 𝑆1and 𝑆2 represent the substrates.
The decay is described by a first order rate equation:
𝑟𝑥𝑑 = 𝐾𝑑𝑋 1-6
where 𝐾𝑑 = the decay coefficient and 𝑋 = biomass concentration.
1.2.3.2 Bacterial growth and biomass yield: Observed versus synthesis yield
In biological treatment processes, cell growth occurs concurrent with the oxidation of organic or inorganic compounds. The ratio of the amount of biomass produced to the amount of substrate consumed (g biomass g-1 substrate) is defined as the biomass yield (Y) and is typically defined relative to the electron donor used.
12 biomass yield is based on the actual measurements of biomass production and substrate consumption and is actually less than the synthesis yield because of cell loss concurrent with cell growth. The term is different from the synthesis biomass yield values, because it contains other organic solids from the wastewater that are measured as VSS but are not biological. The synthesis yield is the amount of biomass produces immediately upon consumption of the growth substrate or oxidation of the electron donor in the case of autotrophic bacteria. The synthesis yield is seldom measured directly and is often interpreted from evaluation biomass production data for reactors operation under different conditions. Synthesis yield values for typical bacteria in wastewater treatment are 0.40 g VSS g-1COD. The observed yield accounts for the actual solids production that would be measured for the system.
1.2.4 Process analysis
The concentration of the microorganism in the effluent is assumed insignificant. For the mass balance, and for the development of the kinetic model for the activated sludge process, some other assumptions are taking into account:
The waste stabilization is carried out by the microorganisms occurs in the aerator unit;
The calculation of the sludge retention time, for the conventional activated sludge process, is made considering only to the aerobic reactor volume;
The hydraulic retention time, HRT, for the aerobic reactor is defined as: 𝐻𝑅𝑇 =𝑉𝑅
𝑄0 1-7
where: 𝑉𝑅 = Volume of the aerobic reactor; 𝑄0 = Influent flow rate.
For the CAS system with excess sludge waste, the sludge retention time (SRT or 𝜃𝑐) is defined as the amount of microorganisms in the reactor over the quantity of microorganisms removed plus the microorganisms wasted from the system per day: 𝜃𝑐 =
𝑉𝑅𝑋
𝑄𝑤𝑋𝑟+𝑄𝑒𝑋𝑒 1-8
where: 𝑄𝑒 = Flow rate of the effluent;
𝑄𝑤 = Flow rate of the liquid containing the biological solids to be wasted from the CAS system;
13 𝑋𝑒 = Effluent microorganism concentration;
𝑋𝑟 = Microorganism concentration in the sludge return line.
The mass balance (Figure 1.2) for the biomass in the CAS system can be written as follows:
𝑑𝑋
𝑑𝑡𝑉𝑅 = 𝑄0𝑋0− [(𝑄0− 𝑄𝑤)𝑋𝑒− 𝑄𝑤− 𝑋𝑟] + 𝑟𝑥𝑉𝑅 1-9
where:
𝑟𝑥 is the net growth rate of microorganisms within the system reported in equation 1-1;
𝑑𝑋
𝑑𝑡 is the rate of change of biomass concentration in reactor;
𝑉𝑅 is the reactor volume; 𝑄0 is the influent flowrate;
𝑋0 is the concentration of biomass in influent; 𝑄𝑤 is the waste sludge flowrate;
𝑋𝑒 is the concentration of biomass in effluent;
𝑋𝑟 is the concentration of biomass in return line from settling.
Assuming zero microorganisms concentration in the influent prevalence of the steady state condition (𝑑𝑋
𝑑𝑡 = 0) and rearranging equations 1-1 and 1-9, a final expression of
sludge retention time is obtained:
1
𝜃𝑐= −𝑌𝑞 − 𝑘𝑑
1-10
where: 𝑞 =𝑟𝑆𝑈
14 The mass balance (Figure 1.2) around the activated sludge process scheme regarding substrate can be written as follows:
𝑑𝑆
𝑑𝑡𝑉𝑅 = 𝑄0𝑆0− 𝑄0𝑆 + 𝑉𝑅(𝑟𝑆𝑈) 1-11
where: 𝑆0 is the influent soluble substrate concentration, 𝑄0 is the flowrate, S is the
substrate and VR is the volume of the reactor.
Figure 1.2 Activated Sludge Flow Diagram and parameters
1.3 BIOLOGICAL TREATMENTS OF COD AND NUTRIENT REMOVAL 1.3.1 Organic matter removal process
The organic matter removal is performed inside an aerobic reactor (as it was mentioned previously) by the activated sludge. This biological aerobic oxidation process is achieved by two steps, oxidation reactions with synthesis of cells and endogenous respiration, in accordance with the stoichiometry shown in the following equations.
Oxidation and synthesis: 𝐶𝑂𝐻𝑁𝑆 + 𝑂2+ 𝑛𝑢𝑡𝑟𝑖𝑒𝑛𝑡𝑠 𝑏𝑎𝑐𝑡𝑒𝑟𝑖𝑎 → 𝐶𝑂2+ 𝑁𝐻3+ 𝐶5𝐻7𝑁𝑂2+ 𝑜𝑡ℎ𝑒𝑟 𝑒𝑛𝑑 𝑝𝑟𝑜𝑑 1-12 Endogenous respiration: 𝐶5𝐻7𝑁𝑂2+ 5𝑂2 𝑏𝑎𝑐𝑡𝑒𝑟𝑖𝑎 → 5𝐶𝑂2+ 2𝐻2𝑂 + 𝑁𝐻3+ 𝑒𝑛𝑒𝑟𝑔𝑦 1-13
in which COHNS symbolises the waste to be treated and 𝐶5𝐻7𝑁𝑂2 the produced new
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1.3.2 Nitrogen removal process
Nitrification
In wastewater the nitrogen present is primarily combined in proteinaceous matter and urea as organic nitrogen. Decomposition by heterotrophic bacteria, known as ammonification, readily converts organic nitrogen to ammonia nitrogen (NH4+-N).
The main organisms involved in nitrification processes are the Nitrosomonas and Nitrobacter. These organisms are autotrophs since they derive energy for growth and synthesis from the oxidation of inorganic nitrogen and carbon (CO2) compounds, rather
than from organic compounds. Both of these groups have rather specific environmental requirements in terms of pH, temperature, and dissolved oxygen and reproduce at much slower rates than heterotrophic bacteria. Various heavy metals and organic compounds have been found to inhibit the growth of nitrifiers. Nitrosomonas can only oxidize ammonia nitrogen to nitrite nitrogen (N-NO2), while Nitrobacter is related to the oxidation
of nitrite nitrogen to nitrate nitrogen (NO3--N).
The oxidation of NH4+-N to NO3--N occurs in two steps, as represented by the following
equations: 2𝑁𝐻4++ 3𝑂2 𝑁𝑖𝑡𝑟𝑜𝑠𝑜𝑚𝑜𝑛𝑎𝑠 → 4𝐻++ 2𝐻2𝑂 + 2𝑁𝑂2− 1-14 2𝑁𝑂2−+ 𝑂 2 𝑁𝑖𝑡𝑟𝑜𝑏𝑎𝑐𝑡𝑒𝑟 → 2𝑁𝑂3 1-15
The overall reaction may be represented by combining equations 1-14 and 1-15:
𝑁𝐻4++ 2𝑂2→ 𝑁𝑂3−+ 2𝐻++ 𝐻2𝑂 1-16
The biomass synthesis reaction can be represented according to equation 1-17:
𝑁𝐻4++ 4𝐶𝑂2+ 𝐻𝐶𝑂3−+ 𝐻2𝑂 → 𝐶5𝐻7𝑁𝑂2+ 5𝑂2 1-17
in which C5H7NO2 is the empirical formula of a bacterial cell. By combining equations
1-16 and 1-17, the overall oxidation and assimilation reaction is equation 1-18:
𝑁𝐻4++ 1.863𝑂2+ 0.098𝐶𝑂2 → 0.0196𝐶5𝐻7𝑁𝑂2+ 0.98𝑁𝑂3−+ 0.0941𝐻2𝑂 + 1.98𝐻+ 1-18 From the equation 1-16 it is noticed that for each gram of ammonia nitrogen (as N) converted, 4.25 g of O2 are utilized, 0.16 g of new cells are formed, 7.07 g of alkalinity
as CaCO3 are removed and 0.08 g of inorganic carbon are utilized in the formation of
16 Denitrification
Biological denitrification involves the oxidation of many organic substrates in wastewater treatment using nitrate or nitrite as the electro acceptor instead of oxygen. In the absence of dissolved oxygen or under limited DO concentrations, the nitrate reductase enzyme in the electro transport respiratory chain is induced, and helps to transfer hydrogen and electrons to nitrate as the terminal electron acceptor. The nitrate reduction reactions involve the following steps from nitrate to nitrite, to nitric oxide, to nitrous oxide and to nitrogen gas:
𝑁𝑂3− → 𝑁𝑂2− → 𝑁𝑂 → 𝑁2𝑂 → 𝑁2 1-19
In biological nitrogen removal processes, the electron donor is typically one of three sources:
The biodegradable COD in the influent wastewater;
The biodegradable COD produced during the endogenous decay; An exogenous source such as methanol or acetate.
The reaction stoichiometry for different electron donors is shown as follows:
Wastewater COD (the term 𝐶10𝐻19𝑂3𝑁 often used to represent the biodegradable organic matter in wastewater).
𝐶10𝐻19𝑂3𝑁 + 10𝑁𝑂3− → 5𝑁2+ 10𝐶𝑂2+ 3𝐻2𝑂 + 𝑁𝐻3+ 10𝑂𝐻− 1-20
Methanol
5𝐶𝐻3𝑂𝐻 + 6𝑁𝑂3− → 3𝑁2+ 5𝐶𝑂2+ 7𝐻2𝑂 + 6𝑂𝐻− 1-21
Acetate
5𝐶𝐻3𝐶𝑂𝑂𝐻 + 8𝑁𝑂3− → 4𝑁2+ 10𝐶𝑂2+ 6𝐻2𝑂 + 8𝑂𝐻− 1-22
In all the above heterotrophic denitrification reactions, one equivalent of alkalinity is produced per equivalent of NO3--N reduced, which equates to 3.57g of alkalinity (as
CaCO3) production per g of nitrate nitrogen reduced. The reduction of nitrate to nitrogen
17
1.3.3 Phosphorus removal process
Phosphorus removal from wastewater is important to prevent eutrophication because phosphorus is a limiting nutrient in most freshwater. Therefore, this process is an integral part of modern wastewater treatment plants for nutrient removal from both municipal and industrial wastewater.
The enhanced biological phosphorus removal (EBPR) process is characterized by cycling the activated sludge through alternating anaerobic and aerobic conditions. The groups of microorganisms that are largely responsible for P removal are known as phosphorus-accumulating organisms (PAOs). These organisms are able to store phosphate as intracellular polyphosphate, leading to P removal from the bulk liquid phase via PAO Cell removal in the waste activated sludge. Phosphorus–accumulating organisms are encouraged to grow and consume phosphorus in systems that use a reactor configuration that provided PAOs with a competitive advantage over other bacteria.
The phosphorus removal in biological systems is based on the following points (Metcalf and Eddy, 2003):
Numerous bacteria are capable of storing excess amounts of phosphorus as polyphosphates in their cells;
Under anaerobic conditions, PAOs will assimilate fermentation products into storage products within the cells with the concomitant release of phosphorus from stored polyphosphates;
Under aerobic conditions, energy is produced by the oxidation of storage products and polyphosphate storage within the cell increases.
1.4 THE PROBLEM OF THE EXCESS SLUDGE PRODUCTION
The conventional activated sludge process has been applied to deal with an enormous variety of wastewater but the effect is an awesome excess sludge production, because the continuous stabilization of organic matter in wastewater results in the production of more microorganisms that are needed to maintain the activated sludge at its desired concentration in the aeration reactor. The extra microbial production is called excess activated sludge.
18 98/15/EEC) (Diaz-Cruz et al. 2009). The EU-28 sewage sludge production during the year 2010 was about 10 million tons of dry solids (http://epp.eurostat.ec.europa.eu) (Figure 1.3) and it is expected that, up to 2020, the previous value will increase exceeding 13 million tons of dry solids (Kelessidis and Stasinakis, 2012). The treatment and disposal of this sewage sludge are challenging waste management problems common to many countries and could account for the 25%-60% of the total plant operation costs (Ettiene, 2012).
The main options for sludge treatment and disposal as landfill, incineration and agricultural reuse have disadvantages as the production of gaseous emissions and hazardous compounds (Foladori et al. 2010). In addition, sludge may contain heavy metals (Tchobanoglus et al. 2003) and trace organic chemicals that can be toxic. Controlling parameters such as increasing the sludge retention time (SRT) and dissolved oxygen (DO) concentration can only yield marginal improvement but may increase plant operation costs (Wei et al. 2003).
Figure 1.3 Sewage sludge production and disposal, (http://epp.eurostat.ec.europa.eu).
19
1.4.1 Strategies for excess sludge reduction
The technologies developed to achieve a reduction in sewage sludge production can be classified in three categories, depending on the treatment line where they are located in the conventional activated sludge plant. Strategies belonging to the first category are applied in the wastewater handling units. Strategies included in the second category are employed in the sludge treatment. The third category focuses on the management of the sludge generated.
Different strategies are also currently developed for sludge reduction in an engineering way based on these mechanisms: lysis-cryptic growth, uncoupling metabolism, maintenance metabolism, and predation on bacteria. However, in order to reach a significant sludge reduction, some techniques have been proven to be not energy saving technologies, while others can negatively affect the effluent quality of the process due to the formation of by-products (Ferrentino et al. 2016).
When certain external forces are applied, microbial cells undergo lysis or death during which cell contents (substrates and nutrients) are released into the medium, providing an autochthonous substrate that is used in microbial metabolism. The term cryptic growth is because the biomass can growth due to this substrate. The net biomass growth in activated sludge system can decreased under cryptic growth conditions. Lysis cryptic growth has two stages: lysis and biodegradation. Lysis is the rate limiting step of the lysis cryptic growth and the increasing of lysis efficiency causes reduction of sludge production. Sludge lysis and cryptic growth can be developed with physical, chemical and the combination of physical and chemical methods such as ozonation, chlorination, combination of thermal/ultrasonic treatment and membrane, heat treatment in the increasing of oxygen concentration. Some of these treatments may result in the production of toxic by-products (Mahmood and Elliott, 2006)
1.4.1.1 Uncoupling Metabolism
20 the production of sludge can be reduced. Uncoupling metabolism can be obtained by adding chemical uncoupling.
In literature there are several studies (Liu et al. 2003, Wei et al. 2003, Ye et al. 2003) about the use of biochemical uncouplers, as the 3,3’,4’,5-tetrachlorosalicylanilide (TCS), in the activated sludge process, also if the toxicity of phenolic compounds is well known (Clarke et al. 2011). TCS is identified as being bioaccumulative, persistent and toxic to aquatic organisms (Liu et al. 2003).
Figure 1.4 Simplified relationship between catabolism and anabolism. 1.4.1.2 Endogenous Metabolism
Bacteria use energy from the substrate biodegradation for the maintenance requirements when external substrate is available. Nevertheless, only a part of cellular constituents can be oxidized to carbon dioxide and water to produce the energy needed for cell maintenance. When the external substrate is completely depleted, storage compounds are used for maintenance purposes: so, bacteria induces endogenous metabolism (Foladori et al. 2010). In the maintenance metabolism, microorganisms satisfy their maintenance energy requirements in preference to producing additional biomass, and this recognition has revealed possible methods for sludge reduction during biological wastewater treatment. In other words, sludge production is inversely proportional to metabolic activity. The main advantage of endogenous metabolism is less biomass production due to the transforming of substrate to CO2 and H2O (Wei et al.
2003).
1.4.1.3 Microbial predation
22
References
Clarke B.O., Smith S.R., 2011. Review of “emerging” organic contaminants in biosolids and assessment of international research priorities for the agricultural use of biosolids. Environ. Int., 37, 226-247.
Diaz-Cruz M.S., Garcıa-Galan M.J., Guerra P., Jelic A., Postigo C., Eljarrat E., Farre M., Lopez de Alda M.J., Petrovic M., Barcelo D., 2009. Analysis of selected emerging contaminants in sewage sludge. Trends in Analytical Chemistry, 28(11), 1263– 1275.
Etienne P., Yu L., 2012. Biological Sludge Minimization and Biomaterials/Bioenergy Recovery Technologies. Wiley, Hoboken, New Jersey.
Ferrentino R., Langone M., Merzari F., Tramonte L., Andreottola G., 2016. Review of anaerobic side stream reactor for excess sludge reduction: Configurations, mechanisms and efficiency. Critical Reviews in Environmental Science and Technol., 46(4), 382-405.
Foladori P., Andreottola, G. & Giuliano, Z., 2010. Sludge Reduction Technologies in wastewater. Treatment Plants. IWA Publishing, London, UK.
Gerardi M. and Wiley A., 2002. Settleability problems and loss of solids in the activated sludge process. Wiley& sons Publication, USA.
Kelessidis A., Stasinakis A.S., 2012. Comparative study of the methods used for treatment and final disposal of sewage sludge in European countries, Waste Manage., 32, 1186-1195
Lee N.M., Welander T., 1996. Reducing sludge production in aerobic wastewater treatment through manipulation of the ecosystem. Water Res., 30, 1781-1790. Liu Y., 2003. Chemically reduced excess sludge production in the activated sludge
process. J. Chemosphere, 50, 1-7.
Low EW, Chase HA., 1999. Reducing production of excess biomass during wastewater treatment. Water Res., 33(5), 1119–32.
Mahmood T., Elliot A., 2006. A review of secondary sludge reduction technologies for the pulp and paper industry. Water Res., 40 (11), 2093-2112.
McGee RD, Drake JF, Fredrickson, AG, Tsuchiya HM., 1972. Studies in intermicrobial symbiosis, Saccharomyces cerevisiae and Lactobacillus casei. Can. J. Microbiol., 18, 1733-1742.
23 Monod J., 1942. Recherches sur la croissance des cultures bactériennes. Hermann et
Cie, Paris, France.
Perez-Elvira S.I., Nieto Diez P., Fdz-Planco F., 2006. Sludge minimisation technologies. Rev. Environ. Sci. Biotechnol., 5(4), 375-398.
Rensink J.H., Rulkens W.H., 1997. Using metazoa to reduce sludge production. Water Sci. Technol., 36, 171-179.
Semblante G. U., Hai F.I., Bustamante H., Price W.E., Nghiem L.D., 2016. Effects of sludge retention time on oxic-settling-anoxic process performance: Biosolids reduction and dewatering properties. Bioresource Techonol., 218, 1187-1194. Spellman, F. R., 2003. Hand book of water and wastewater treatment plant operation.
Lewise publishers, USA.
Tchobanoglous G., Burton F.L., 1991. Wastewater Engineering: Treatment, Disposal and Reuse. McGraw-Hill Series in Water Resources and Environmental
Engineering.
Ye F.X., Shen D.S., Li Y., 2003. Reduction in excess sludge production by addition of chemical uncouplers inactivated sludge batch cultures. J. Appl. Microbiol., 95, 781-786.
Vilaseca. 2001. Observacion microscopica de fangos activados en los tratamientos de depuracion biologica. Boletin Intexter (U.P.C) no. 119.
24
CHAPTER 2 :
OBJECTIVES AND THESIS STRUCTURE
2.1 MOTIVATION AND OBJECTIVES
As stated in Chapter 1, organic matter and nutrients, present in urban and industrial wastewater, should be removed or valorised to reduce its impact on the environment. Conventional wastewater treatments are focused on the removal of these pollution sources, obtaining a huge production of excess sludge, which must be removed from the activated sludge systems to maintain the desired microbial population. This sludge disposal, followed by the supply of oxygen in aerobic processes, are the largest operating costs in wastewater treatment. Therefore, any technology or strategy to reduce energy consumption must involve the reduction of sludge output and oxygen input.
These considerations were the motivation of the present work, which deals with the application of two different technologies:
1. The anaerobic membrane bioreactor technology (AnMBR, reported in Chapter 4), in which the costs for aeration are avoided and a low sludge production is obtained;
2. The oxic-settling-anaerobic process (OSA, reported in Chapter 5), in which the main goal is the reduction of the excess sludge production.
In order to reach the general objective, the following purposes are introduced through Chapter 4 and Chapter 5.
Chapter 4: Anaerobic Membrane Bioreactor system (AnMBR)
To examine the operation of an AnMBR treating synthetic winery wastewater at 25ºC and at 15ºC, (psychrophilic temperature) simulating the winter season. After 45 days of start up at 35°C, the temperature was decreased to 25°C for 45 days and to 15°C for 60 days.
To carry out several anaerobic digestion batch tests to assess the activity of biomass at different temperatures (35°C, 25ºC and 15ºC).
To study the quality of the effluent in order to assess its potential to be reused. Measure the production of biogas at each temperature.
25 Chapter 5: Oxic-Settling-Anaerobic process (OSA)
To project and construct an automated OSA pilot plant located close to a real wastewater treatment plant.
To make the start-up phase of the pilot plant, operating as Conventional Activated Sludge (CAS) system, with sludge inoculum and wastewater coming from the real wastewater treatment plant.
To change the CAS pilot plant configuration to the OSA process inserting an anaerobic sludge holding tank in the sludge return line.
To monitor both configurations (CAS and OSA) by sCOD and NH4+-N removal
efficiencies.
To perform respirometric tests on biomass samples from CAS and OSA configurations.
To calibrate a mathematical model using the respirometry results, estimating the required kinetic and stoichiometric parameters for the pilot plant (working as CAS and OSA).
To evaluate the reduction of excess sludge production on the basis of the observed sludge yield (Yobs) and of the heterotrophic biomass aerobic yield.
To simulate both periods of the pilot plant with BioWin software.
2.2 THESIS STRUCTURE
This dissertation is divided into six chapters:
Chapter 1 introduces the conventional activated sludge treatment process and the problem on the excess sludge production.
Chapter 2 shows the motivation of the research and the general objectives of the thesis.
Chapter 3 introduces the materials and methods used during these three years of research.
Chapter 4 reports the activity on AnMBR technology that was developed at the Barcelona University.
Chapter 5 describes the main activity on the OSA pilot plant that was performed at the Trieste University.
26
CHAPTER 3 : MATERIALS AND METHODS
In order to reach the objectives proposed in Chapter 2, the following materials and methods were used through Chapter 4 and Chapter 5.
3.1 EXPERIMENTAL SET-UP
Different experimental set-up were used in this study:
a. an anaerobic membrane bioreactor (AnMBR) (Chapter 4) treating synthetic winery wastewater;
b. an oxic-settling-anaerobic (OSA) pilot plant (Chapter 5) treating municipal wastewater;
c. a respirometer set up (Chapter 5) to obtain the kinetic and stoichiometric parameters of the pilot plant process.
3.1.1 Anaerobic membrane bioreactor
The AnMBR was set-up as a conventional stirred anaerobic digester of 5L, coupled with an external membrane unit (Orelis, Rayflow Module) of 100 cm2 of membrane area (Figure 3.1). The digester was a jacketed vessel mechanically stirred at 100 rpm and heated at the desired temperature by recirculating water from a heated water bath (HUBER 118A-E). Influent wastewater was fed from a 10 L tank with a cooling system to avoid early degradation. Digester feeding was performed keeping the digester in contact with a 500 mL cylinder at a constant volume of wastewater: thus, the working volume inside the digester was kept at 5-6 L. Since the membrane unit was placed outside the digester, biogas was quantified with an on-line measuring device (Ritter MGC-1) connected to the headspace of the digester.
27
3.1.2 Oxic-settling-anaerobic pilot plant
Chapter 5 is involved in the project, construction and start-up of the automated OSA pilot plant, which was assembled with the following stainless steel units (see figure 3.2): A Reservoir tank (A) of cylindrical shape, with a volume of 300 L, and fed with
municipal wastewater (WW) by an intermittent pump.
An Aerobic reactor (B) fed with the wastewater (WW) of the reservoir tank and the recirculated activated sludge from the anoxic reactor (D). It has a cylindrical shape with internal diameter of 1,150 mm and height of 1,100 mm: it is also provided of four spillways located at different levels, to allow work volumes of 500, 700, 880 and 1,070 L. The mixing in the reactor is provided by four air diffusers. Dissolved oxygen, ORP and temperature bayonet probes are located in the vertical wall of the reactor.
A Settling tank (C) fed with the treated WW of the aerobic reactor. It has a conical shape with a total height of 1,590 mm. A low speed (0.1 rpm) paddle agitator along the vertical axis of the settling tank is provided to facilitate the sludge concentration and to avoid sludge arching.
An Anoxic reactor (D) fed with the sludge coming from the settling tank. It has a cylindrical shape, with internal diameter of 1,400 mm and height of 1,325 mm. The reactor gas tight cover is equipped of a mechanical stirrer (SRA1) to maintain the sludge suspended with homogeneous composition. The produced biogas is collected in a gasometer. The reactor is provided of temperature, ORP and pH sensors.
An reservoir tank (E) fed with the settler effluent water. It has a cylindrical shape with conic bottom, internal diameter of 700 mm and height of 700 mm.
28
3.1.3 Respirometric set-up
The respirometric batch experiments (see Figure 3.3) were performed using a Plexiglas reactor with a volume of 800-1000L, provided of a magnetic stirrer (1) and thermally controlled with a thermostatic bath (2). The reactor was filled with the activated sludge (from the pilot plant aerobic reactor) and different substrates. Dissolved oxygen (DO) concentration was measured by electro-chemical Clark-type probes (5). Aeration was provided by membrane pumps (10). The respirometer was arranged also with pH, temperature (6) and ORP probes, which were connected to a data logger acquisition unit (7).
Figure 3.3 Experimental respirometer described in section 5.3.6 (Vitanza et al. 2016b): 1. magnetic stirrer; 2 thermostatic water-bath; 3 oxygen porous diffuser; 4 mixed liquor; 5 DO probe;
29
3.2 ANALYTICAL METHODS
Chapter 4: Anaerobic Membrane Bioreactor system (AnMBR)
The analytical methods used in this investigation were performed for measurements of Chemical Oxygen Demand (COD), Alkalinity, pH and Total solids, according to the Standard Methods for the examination of Water and Wastewater (APHA, 2005).
Volatile fatty acids (VFAs) and biogas compositions were determined by gas chromatographic technique.
Biomethane potential tests (BMP) were performed following the guidelines of Angelidaki et al. (2009).
Fluorescence in situ hybridization (FISH) technique was carried out according to López-Palau (2012).
Chapter 5: Oxic-Settling-Anaerobic process (OSA)
In this research, wastewaters, effluents and activated sludges were analysed for Nitrogen (ammonium and nitrate), Phosphate and COD by means of Hach-Lange test cuvettes.
Mixed liquor total and volatile suspended solids (MLSS and MLSSV) were measured according to Standard Method for Examination of Water and Wastewater (APHA, 2005). Gas samples were analysed by a landfill gas analyser.
3.3 ACTIVATED SLUDGE MODELLING
Chapter 5: Oxic-Settling-Anaerobic process (OSA)
A mathematical model for the activated sludge pilot plant, based on ASM No.3, was calibrated by using the respirometric assays of the same pilot plant (calibration data-set). Parameter optimization and the numerical solution of model equations were performed using the Aquasim program (Reichert, 1998).
30
References
Angelidaki I., Alves M., Bolzonella D., Borzacconi L., Campos J.L., Guwy aJ., Kalyuzhnyi S., Jenicek P., van Lier J.B., 2009. Defining the biomethane potential (BMP) of solid organic wastes and energy crops: a proposed protocol for batch assays. Water Sci. Technol., 59, 927-34.
APHA, 2005. Standard Methods for the Examination of Water & Wastewater. American Public Health Association, Washington DC. ISBN: 0875530478.
EnviroSim Associates Ltd., Canada. BioWin 5.0 Software.
Lopez-Palau S., 2012. Biological granulation technology for wastewater treatment. University of Barcelona.
31
CHAPTER 4 : ANAEROBIC MEMBRANE BIOREACTOR SYSTEM
(AnMBR)
4.1 INTRODUCTION
As mentioned in Chapter 1, the prediction of excess sludge production is an important issue for the design and modeling of wastewater treatment plants but it is difficult to achieve because of the numerous processes involved and the variability of the factors that influence it. Energy considerations have spurred a reevaluation of wastewater treatment processes.
Not only treatment efficiency and effluent quality are important but also the energy required for the treatment of the excess sludge. The largest operating cost in wastewater treatment is that associated with sludge disposal followed by the supply of oxygen in aerobic processes. For this, any technical or strategy to reduce energy consumption must involve reducing sludge output and oxygen input (Etienne et al. 2012).
Traditional anaerobic processes are an excellent example to achieve high organic matter removal without oxygen requirement and therefore, with low energy demand because (Metcalf and Eddy, 2003):
no aeration is required;
there is a low biomass production due to the low biomass yield of anaerobic organisms;
there is a reduction of the cost of handling, stabilization and final disposal of the sludge (biological sludge minimization);
there is a recovery of energy from the methane gas produced in the process. Anaerobic wastewater treatment processes are typically conducted within mesophilic (28-45°C) or thermophilic (45-60°C) temperature ranges because most of the biochemical reactions involved in organic matter biodegradation proceed slower under psychrophilic (≤ 25°C) conditions.
4.1.1 Fundamentals of anaerobic wastewater treatment process
Anaerobic wastewater treatment process consists of several interdependent, complex sequential and parallel biological reactions, during which the products from one group of microorganisms are taken as the substrates for the next, resulting in transformation of organic matter mainly into a mixture of methane and carbon dioxide.
32 1. Hydrolysis: Complex particulate organic matter is decomposed into simple soluble molecules which can pass through the cell walls and the membranes of the fermentative bacteria;
2. Fermentation or Acidogenenesis: The dissolved compounds are converted into a number of simple compounds (volatile fatty acids, alcohols, lactic acid, CO2, H2,
NH3 and H2S) which are then excreted;
3. Acetogenesis: The fermentation products are converted into acetate, hydrogen and carbon dioxide by acetogenic bacteria;
4. Methanogenesis: Acetate and hydrogen /carbon dioxide are converted into methane and CO2 by methanogenic bacteria.
Figure 4.1 Path of anaerobic digestion.
The acetogenic bacteria grow in close association with the methanogenic bacteria during the fourth stage of the process. The reason for this is that the conversion of the fermentation products by the acetogens is thermodynamically possible only if the hydrogen concentration is kept sufficiently low. This requires a close relationship between both classes of bacteria.
To ensure a balanced digestion process it is important that the various biological conversion processes remain sufficiently coupled during the process to avoid the accumulation of any intermediates (as volatile fatty acids, VFA) in the system.
33 The combination of membrane separation technology and an anaerobic bioreactor (AnMBR) may allow for a sustainable municipal wastewater treatment, without the extra costs for aeration associated with the conventional activated sludge process, and with the added benefits of lower sludge production and solid-free effluent. Moreover, the quality of the effluent should be better and the permeate can be reused after.
4.1.2 AnMBR
AnMBR combines an anaerobic biological treatment system with a membrane separation process. Organic components are digested in the anaerobic reactor and energy rich biogas is produced. Incorporating membranes to anaerobic membrane bioreactor wastewater treatment can results in a superior effluent quality in terms of COD, SS and pathogen in comparison with the conventional anaerobic processes (Ozgun et al. 2013).
AnMBRs were first introduced in the 1980s in South Africa. At the present time there is an increasing interest in the field of AnMBR (Visvanathan et al. 2012).
There are some advantages as the separation of the solids retention time (SRT) from the hydraulic retention time (HRT) (optimizing biological performance), no suspended solids in the final effluents (improving quality and reliability) and the suitability of the effluent for post treatment and reuse applications (Ozgun et al. 2013).
However, it is important to note that AnMBR process does not remove nutrients and therefore additional treatment may be required in watersheds where nutrient effluent limits are in place. On the other hand, as macronutrients are not removed, the permeates of AnMBRs are positively of interest for agricultural use (Martinez-Sosa et al. 2011). Some problems related to the AnMBR technology are: the strong influence of temperature on the process, low operational fluxes, rapid membrane fouling, high capital and operational costs. (Stuckey et al. 2012).
4.1.2.1 Operational conditions 4.1.2.1.1 Temperature
34 AnMBRs can tolerate slight temperature fluctuations. Martinez-Sosa et al. (2011) reported that bioreactor temperature influenced methane recovery in an anaerobic externally submerged MBR treating municipal wastewater.
4.1.2.1.2 pH
Most AnMBR systems operate at near neutral pH since anaerobic digestion takes place within the pH range of 6.5-8.5 with an optimum interval between 7 and 8. Such a desired pH was usually achieved through neutralization, which could require an excessive use of chemicals because hydrolysis and acidogenesis phases normally decrease the pH value (Lin et al. 2013).
4.1.2.1.3 HRT (Hydraulic retention time)
HRT is an important parameter from an economic perspective as it has a strong influence on capital costs, considering that shorter HRTs allow smaller reactors. Liao et al. (2006) found that the lowest HRTs for a wide range of wastewaters was 8-12 h.
4.1.2.1.4 SRT (Sludge retention time)
SRT is one of the main operational parameters determining both treatment performance and membrane fouling (Lin et al. 2013). SRTs in AnMBRs can be as low as 25 d as high as 335 d (Liao et al. 2006). High SRTs are more desirable since it corresponds to less sludge production and higher sludge concentrations in the reactor. However, long SRTs may also affect methanogenic sludge activity owing to a decrease in viable microorganism concentration. High SRTs may stimulate cell lysis and increase the release of inert decay products and soluble microbial products (SMP) leading to an increase of effluent COD concentration (Barker et al. 1999).
The effect of high SRTs especially on membrane filtration performance is still a research topic that needs to be further investigated.
4.1.2.1.5 Organic loading rate (OLR)
35
4.1.2.2 Membrane fouling and flux
Membrane fouling in AnMBR can be reversible and /or irreversible. It can take place on the membrane surface or into the pores. As soon as the membrane surface comes into contact with the biological suspension, the biosolid deposition onto membrane surface takes place.
The biosolid is a combination of several components in the reactor, such as soluble organics, colloidal particles from the feed and cell lysis and inorganic precipitates. These in turn are influenced by a range of parameters such as the composition of the biological system, membrane type, hydrodynamic conditions, reactor operation conditions, process performance targets and the chemical system (Stuckey et al. 2012).
The fouling is called reversible when it is possible to remove the biosolid from the membrane by an appropriate physical cleaning.
The irreversible fouling is normally caused by strong attachment of particles, which is difficult or impossible to be removed by physical cleaning methods, and it is generally removed by chemical cleaning.
Membrane fouling remains the critical obstacle limiting the more widespread application of AnMBR in wastewater treatment. It could decrease system productivity, cause frequent cleaning which might reduce the membrane lifespan, and result in higher replacement costs and increase the energy requirement for sludge recirculation or gas scouring (Lin et al. 2013).
36
4.1.3 Research at Barcelona University
In the past, the AnMBR technology was performed at the Barcelona University to remove the organic matter from winery wastewater at mesophilic temperature due to its advantages; removing the organic matter with the option to produce energy (biogas) and improving the quality of the effluent for the reuse applications (Basset et al. 2016). AnMBR technology has been introduced for industrial wastewater treatment with high organic content from distilleries, septic tanks, food and paper industries since 1990s (Skouteris et al. 2012).
Winery wastewater is an industrial wastewater characterized by a high content in biodegradable organic matter and by a strong seasonal variability. During summer season, the production of winery wastewater is high and contains elevated organic matter, whereas, during the wintertime, the production decreases with the consequence that organic loading matter is quiet low.
Only if winery wastewater has a high organic load (COD over 3 gL-1) can be treated at high temperature by AnMBR technology because the biogas obtained would cover the heating expenses (Basset et al. 2014). The problem arises during the winter season when winery wastewater contains less organic loading matter so the produced biogas would not be enough to maintain the bioreactor and the supply of external energy could be necessary. In these conditions, the winery wastewater would have similar characteristics as the urban wastewater.
4.2 OBJECTIVES
The main objective of this study is to examine the operation of an AnMBR treating synthetic winery wastewater at 25ºC and at 15ºC, (psychrophilic temperature) simulating the winter season.
After a start-up phase at 35°C, the temperature will be decreased to 25°C and then to 15°C.
37
4.3 MATERIALS AND METHODS 4.3.1 AnMBR configuration
The bench-scale AnMBR used in this study is shown in Figure 4.2 and described in Basset et al., 2016.
The digester was a jacketed vessel mechanically stirred at 100 rpm, heated/cooled at a desired temperature (35°C, 25°C and 15°C) by recirculating water from a heating/cooling water bath (HUBER 118A-E).
The reactor feeding was performed by pressure equilibrium keeping the digester in contact with a 500 mL cylinder at a constant volume of wastewater. Thus, the working volume inside the reactor was kept at an average of 5-6 L.
Figure 4.2 Scheme of experimental AnMBR.
38
(a) (b)
Figure 4.3 (a) Reactor and communicating vessel. (b) Membrane module.
The Figure 4.3 shows the AnMBR lab scale with the communicating vessel (a) and the membrane module working in the system (b). The synthetic wastewater containing around 1,500 mgCOD L-1, was prepared with diluted white wine (Artiga et al. 2005) and NH4Cl and K2HPO3 in accordance to the ratio COD/N/P of 800/5/1. Moreover, alkalinity
was added (500-1,000 mgCaCO3 L-1) to keep the pH at neutral values. Influent synthetic
wastewater was placed in a 10L reservoir tank located in a cool box in order to avoid early degradation. Nevertheless, sometimes significant oscillations in COD concentration were detected. To try to reduce this degradation, synthetic wastewater was provided every 1-2 days.
As mentioned previously, the bioreactor was operated at 35°C for 40 days, for the following 45 days at 25°C, and for 60 days at 15°C.
4.3.2 Analytical methods
The analytical methods used in this investigation were performed according to the Standard Methods for the examination of Water and Wastewater (APHA, 2005).
39
4.3.2.1 Volatile fatty acids (VFAs)
Concentrations of volatile fatty acids (VFAs) (formic acid, acetic acid, propionic acid,
butyric acid, isobutyric acid, valeric acid and isovaleric acid) were determined by a Shimadzu GC-2010+ gas chromatograph equipped with a capillary column Nukol (0.53 mm ID; 15 m length) and a flame ionization detector (FID). Carrier gas was helium at a rate of 36.9 mL min-1 and 17.6 kPa.
Biogas composition, as percentage of methane and carbon dioxide, was determined by a second Shimadzu GC-2010+ gas chromatograph equipped with a capillary column Carboxen 1010 Plot (0.53 mm ID; 30 m length) and a thermal conductivity detector (TCD). The analysis program was as follows: hold 6 min at 40°C; increase to 230°C at a rate of 25°C min-1 and hold 2 min at this temperature. Injector and detector temperature was set at 200°C and 230°C, respectively. Helium was the carrier gas at 47 mL min-1 and 20.4 kPa.
4.3.2.2 Suspended solids content (SS)
Total suspended solids (TSS) and volatile suspended solids (VSS) were determined following the reference methods 2540D and 2540E, respectively. A known volume of sample (V) was filtered through a 1.2 μm Millipore standard filter, previously weighted (W1). Later, the filter with the TSS was placed at 105°C during 4h, afterwards in a
desiccator for 10 minutes and then it was weighted (W2). TSS concentration was
calculated according to Equation 4-1. Finally, the filter with TSS was burned at 550°C for 15 minutes, after that put in a desiccator for 10 minutes and weighted (W3). Thus, the
VSS value could be obtained from Equation 4-2. 𝑇𝑆𝑆[𝑔𝐿−1] =𝑊2−𝑊1
𝑉 4-1
𝑉𝑆𝑆[𝑔𝐿−1] =𝑊2−𝑊3
𝑉 4-2
4.3.2.3 Chemical oxygen demand (COD)
The COD indicates the quantity of matter present in a wastewater sample that is susceptible to be oxidised. This parameter is expressed as mgO2L-1, so that the COD is
40 compounds. Silver sulphate was used to catalyse the reaction and mercuric sulphate to avoid the interference of chloride (Equation 4-4).
𝐶𝑛𝐻𝑎𝑂𝑏𝑁𝑐+ (2𝑛 3 + 𝑎 6− 𝑏 3− 𝑐 2) 𝐶𝑟2𝑂7 2−+ (8𝑑 + 𝑐)𝐻+ → 𝑛𝐶𝑂 2+ ( 𝑎+8𝑑−3𝑐 2 ) 𝐻2𝑂 + 𝑐𝑁𝐻4 ++ 2𝑑𝐶𝑟3+ 4-3 6𝐶𝑙−+ 𝐶𝑟 2𝑂72−+ 14𝐻+ → 3𝐶𝑙2+ 2𝐶𝑟3++ 7𝐻2𝑂 4-4
Five standards of potassium biphtalate with 0, 50, 250, 500 and 1,000 mg CODL-1, treated with the same procedure of the samples, were analysed to determine the calibration curve.
Each sample was prepared with 2.5 mL of the wastewater mixed with 1.5 mL of sodium dichromate 0.04 molL-1 (with 80 gL-1 of mercuric sulphate) and 3.5 mL of sulphuric acid (with 10 gL-1 of silver sulphate). It was maintained at 150°C for 2h in a digester to ensure the complete reaction and then cooled down to allow the solids formed to settle at room temperature.
Finally, all the samples and the standards were analysed by means of a Shimadzu UV-1203 spectrophotometer at λ of 620 nm.
4.3.2.4 Alkalinity
The alkalinity was measured using an automatic titration device (CRISON pH Burette 24) equipped with a pH meter (CRISON Basic 20) as shown in Figure 4.4. The method consists in a titration of 25 mL of sample with standard acid (HCl 0.1M) to desired end point. The alkalinity in this research is expressed as mg CaCO3 L-1 and calculated with
Equation 4-5.
Figure 4.4 Alkalinity apparatus measurement. 𝐴𝑙𝑘𝑎𝑙𝑖𝑛𝑖𝑡𝑦[𝑚𝑔𝐶𝑎𝐶𝑂3𝐿−1] =𝑚𝐿𝐻𝐶𝑙 ∙ 0.1𝑁 ∙ 50.000
41
4.3.3 Biomethane potential test
The biochemical methane potential (BMP) is a procedure developed to determine the methane production of a given organic substrate during its anaerobic decomposition (Raposo et al., 2011).
(BMP) assays were performed in order to evaluate biomass activity at each temperature. In each experiment, the inoculum used was taken from the AnMBR after operating at the selected temperature for 30 days. The tests were carried out in 100 mL serum bottles closed with a PTFE/Butyl septum fixed around the rim of the bottle by an aluminum crimp cap (Figure 4.5), following the guidelines of Angelidaki et al. (2009). The bottles, containing 50 mL of inoculum, were filled up to 80 mL with substrate (synthetic winery wastewater prepared with diluted white wine with 5.41gL-1COD).
The COD substrate/COD inoculum (F:M) ratio selected was close to 1.5 g COD g-1COD.
Alkalinity of 1,000 mg CaCO3 was also supplied.
Several physicochemical parameters were determined in order to characterize the initial and final conditions of the BMP test, which were COD, TSS, and SSV.
A blank test, adding only biomass and deionized water, was necessarily prepared to evaluate biomass endogenous activity. Before closing the bottles, nitrogen was added to avoid the presence of oxygen. The digesters were placed at three different temperatures (35°C, 25°C and 15°C) and manually mixed twice a day. The biogas production during the test was obtained by means of a vacumeter (Ebro – VAM 320) and adjusted to normal conditions (0°C and 1 atm) once subtracted the vapor pressure. Finally, the methane content was analyzed by gas chromatography (Shimadzu GC-2010+).
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4.3.4 Fluorescence in situ hybridization (FISH)
The different populations of microorganisms present in the sludge samples of the reactors were studied by means of fluorescent in situ hybridization (FISH). It is a cytogenetic technique used to detect the presence of DNA sequences in chromosomes. It consists of chemically preparing a short strand of a specific sequence of nucleic acids, an oligonucleotide, and appending a coloured fluorescent marker at its end. Cells are then made porous to the marked oligonucleotide, which binds to its complementary strand of RNA. After removing the unbound markers, bacteria containing the target genetic material emit light that can be observed under a fluorescent microscope.
The development of this technique was carried out according to López-Palau 2012, and fluorescent signals were recorded with a TCS-SP2 confocal laser scanning microscope (Leica, Germany), equipped with a DPSS 561 nm laser for the detection of Cy3 and one Argon ion laser for the detection of 6-fam. The specific oligonucleotide probes used were: EUB338 for Bacteria (6-fam); ARC915 for Archaea (Cy3); MX825 for Methanosaeta spp. (6-fam); MS821 for Methanosarcina (Cy3); MG1200b for Methanomicrobials spp. (6-fam); and MB311 for Methanobacterials (minus Methanothermus) (Cy3). Samples at each temperature were taken to determine the changes on the microbial population. The procedure included the fixation and permeabilization of the sample, hybridization of the targeted sequence to the probe, washing steps to remove unbound probe and finally the detection of labelled cells by microscopy (Figure 4.6).
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4.3.5 Membrane characterization: critical flux
