Evaluation of the Denitrification Process in the Sewer Pipeline, Dalby-Lund, Sweden

Transcription

1 Water and Environmental Engineering Department of Chemical Engineering Evaluation of the Denitrification Process in the Sewer Pipeline, Dalby-Lund, Sweden Master s Thesis by Farnaz Edalat November 2008

2

3 Vattenförsörjnings- och Avloppsteknik Institutionen för Kemiteknik Lunds Universitet Water and Environmental Engineering Department of Chemical Engineering Lund University, Sweden Evaluation of the Denitrification Process in the Sewer Pipeline, Dalby-Lund, Sweden Master Thesis number: by Farnaz Edalat Water and Environmental Engineering Department of Chemical Engineering November 2008 Supervisor: Carl-Johan Legetth Examiner: Professor Jes la Cour Jansen Picture on front page: 1 Activated sludge process in Dalby WWTP (Photo: F.Edalat) Postal address: Visiting address: Telephone: P.O Box 124 Getingevägen SE Lund Sweden, Telefax: Web address:

4

5 Acknowledgement This master thesis is a part of the Master of Science program in Water Resources Engineering at Lund University. It has been conducted at Tyréns AB in collaboration with Water and Environmental Engineering, Department of Chemical Engineering, Lund University. I would like to thank Carl-Johan Legetth, my supervisor at Tyréns AB, who gave me this opportunity to work in a real project in this company. He has been really supportive during my work and providing me all the information I needed. I am especially grateful to Jes la Cour Jansen, my examiner at Water and Environmental Engineering, Department of Chemical Engineering, for his guidance in theoretical issues, excellent feedbacks and his patience in all steps of this thesis. I would like to thank Hans Carlsson and Stefan Aguayo for their time and advices and all the staff at Tyréns AB in Water Department, Ann-Christin, Nanna and Cornelia for creating comfortable and pleasant working environment for me. I appreciate all supports and encouragements from my family in all levels of my studies in Sweden although being far away. I would like to say thank you to Hassan, the wisest friend. Lund, November 2008 i

6 ii

7 Summary The Dalby-Lund sewer pipe project is planned to transport partially treated wastewater from Dalby WWTP to Källby WWTP in Lund. The underground sewer pipe has been hydraulically designed at Tyréns AB and will be constructed as soon as all issues with the property owners have been resolved. Due to the population growth, Lund municipality, which is in charge of this project, decided to transport wastewater from Genarp, Veberöd and Björnstorp communities to Dalby WWTP and eventually to Källby WWTP. The post-denitrification method will be conducted, in which the nitrification process will occur in Dalby WWTP while denitrification is supposed to occur in the sewer pipe under anoxic conditions. The aim of this study is to evaluate whether in-sewer microbiological transformations can efficiently reduce the nitrogen concentration under anoxic conditions in the sewer pipe. The denitrification process has been studied both in bulk water and biofilm phases of the sewer pipe. EFOR computer program has been used to simulate biological nitrogen removal processes in the bulk water phase while calculations of the nitrogen removal in the biofilm phase have been done manually. In the computer models, different scenarios have been defined to investigate the nitrate removal efficiency in the sewer pipe in different situations. Scenarios modeling the sewer pipe as anoxic chambers could decrease the initial nitrate concentration in the sewer pipe which is 17.6 g NO 3 /m 3 to 16 g NO 3 /m 3 in the outlet. Scenarios, in which bypassed raw wastewater was injected to the sewer pipe, nitrogen compound concentrations did not show significant removal efficiency because the removed nitrate was substituted by the high ammonium concentration in the raw wastewater. Despite the addition of the required external carbon source, methanol, to the sewer pipe, nitrate could not be efficiently removed in the sewer pipe and the denitrification rate was limited to 0.67 g NO 3 /m 3 h. Results indicated that the available organic matter in the sewer pipe was the limiting factor for redox reactions so that nitrate removal efficiency in the outlet was 22%. By injecting the required external carbon source to the sewer pipe, the biofilm was capable of removing 6.81 g NO 3 /m 3 in the outlet. The nitrate utilization rate was estimated to 0.13 g NO 3 /m 2 h in the biofilm phase of the sewer pipe. The results indicate a low denitrification rate and consequently low nitrate removal in the Dalby- Lund sewer pipe. The low denitrification rates in this project can be explained by three general reasons. The first one is that all the produced biomass in the sewer pipe is discharged to Lund s sewer network and there is not any possibility to return it to the sewer pipe to increase biomass concentrations in the sewer pipe system. In fact, the low biomass concentration in the sewer pipe has limited the denitrification rate along the sewer pipe. Another reason is that only rapidgrowing bacteria can take part in the organic matter degradation in the short residence time, 6 hours, in the sewer pipe. The third reason is that the molecular diffusion of the substrates into the biofilm layer is a slow process limiting the efficiency of the biomass in the sewer pipe. iii

8 iv

9 Table of contents 1. Introduction Objective and methodology Limitations 2 2. Background 3 3. Literature study; Biological nitrogen removal Biological treatment Nitrification Denitrification Different methods of denitrification C/N ratio Sewer as a reactor Kinetics of redox processes Anoxic condition Anaerobic condition Corrosion risk Odor nuisance and health problems Anoxic condition prevalence to prevent sulfide formation Sewer as a biofilm reactor Integrated operation of sewer system and WWTP Treatment processes in Dalby WWTP Present approach Future approach Biological transformations in the bulk water EFOR software Modeling methodology Model description Model parameters Model components description C/N ratio in the sewer pipe Model components characteristics 32 v

10 5.3. Different Alternatives for optimizing denitrification Results Nitrate concentration in the outlet COD concentration in the outlet Ammonium concentration in the outlet Biological transformations in the biofilm Denitrification kinetics in the biofilm Methodology Results Physical properties of the sewer pipe Kinetic constants in a denitrifying biofilm reactor Mass transfer limitation Nitrate removal efficiency in the biofilm Addition of an external carbon source Discussion Denitrification in the bulk water Denitrification in the biofilm Application of an external carbon source Sulfide formation risk Effects of the project on Källby WWTP Conclusion Recommendation 55 References 57 Appendixes 59 vi

11 1. Introduction Municipal wastewater commonly contains high amounts of nitrogen in different forms such as ammonium (NH 4 + ), organic compounds and in much lower extent nitrite (NO 2 - ) and nitrate (NO 3 - ). These mostly originate from metabolic activities in the human body, specifically from food processing, and can be found both in organic (amino acids, proteins and urea) and inorganic compounds (ammonium ions). Presence of nitrogen-containing waste in the effluent can cause several environmental problems. Nitrogen compounds, ammonium, nitrate and nitrite, are favorable for the growth of plants such as algae. Dissolved oxygen oxidizes the ammonium ions to nitrite and nitrate resulting in dissolved oxygen depletion. In addition, the nitrogen compounds especially ammonia (NH 3 ) are very toxic substances. At high ph, ammonia is the dominant product of the chemical equilibrium between ammonium and ammonia which may cause toxicity if it is discharged to the environment. Eutrophication is another environmental concern, nutrients such as phosphorous and nitrogen compounds cause a rapid growth of plants. When these plants die, the undecomposed parts of the plants accumulate in the freshwater and endanger the aquatic life (Gerardi, et al., 2003). Sewer networks and wastewater treatment plants (WWTP) as integrated parts of the urban wastewater system are important to enhance the sustainability of the wastewater management in the society. Although sewer networks are inherently designed to collect and convey wastewater to the WWTP, microbial transformations of wastewater along the sewer pipes are inevitable. The presence of different types of bacteria in the wastewater and their tendency to grow and reproduce explains the necessity of designing the sewer system based on not only hydraulics and physical points of view but also on chemical and microbiological points. These microbial transformations in sewage pipes may adversely affect the sewer, WWTP and also the human health. In gravity sewers, wastewater commonly contains dissolved oxygen which provides aerobic condition in the sewer pipes. Anaerobic respiration occurs when dissolved oxygen and nitrate ions are consumed in the sewage system, mostly pressurized mains, which can cause several corrosion and odor problems. Addition of nitrate to the sewer pipes, to provide artificially anoxic conditions, is studied as a solution to prevent anaerobic conditions and/or oxidize the generated sulfides in the network (Yang, et al., 2004; Yang, et al., 2005; Mathioudakis, et al., 2006). All of these studies show high denitrification efficiency in the sewer network resulting in nitrate consumption at a high rate while harmful effects of prevailing anaerobic respiration were not reported. In the Dalby-Lund sewer pipe project, the sewer pipe will not be used to transfer the raw wastewater to the WWTP but it is planned to convey the partly treated wastewater (nitrified wastewater) from a conventional WWTP to another WWTP with more capacity. Since nitrified wastewater contains a high concentration of nitrate ions, it is expected that denitrification takes place along the sewer pipe at a high rate. In this scenario, activated sludge method in the nitrification stage in the WWTP and denitrification kinetics in the sewer networks are applied as the two main techniques to evaluate the process efficiency. 1

12 1.2. Objective and methodology The aim of this study is to evaluate the denitrification process efficiency along the sewer pipeline in Dalby-Lund project. In this study, the nitrate removal efficiency in the outlet is estimated to assess the in-sewer microbiological transformation under anoxic conditions. Furthermore, investigating the effects of using bypassed raw wastewater and an external carbon source to compensate organic matter deficiency and enhance the denitrification process in the sewer pipe is a case of study in this report. In the first part of the report, a literature study based on the scientific results achieved in correlated studies with this project will be conducted. The nitrogen removal in the activated sludge method, microbial transformations in water and biofilm phases in the sewer networks in different aerobic, anoxic and anaerobic conditions and also their effects on the connected WWTP will be investigated. In order to evaluate the nitrate removal efficiency under anoxic condition in the sewer pipe, different solutions have been assessed to optimize the denitrification process in the bulk water and the biofilm phases of the sewer pipe. The denitrification process in the bulk water phase of the sewer pipe will be investigated by simulating the treatment process in EFOR, computer modeling software. The nitrate removal in the sewer wall biofilm will be manually calculated to assess whether the denitrification process occurs in a high extent in the biofilm Limitations As far as applying the denitrification method under sewer conditions is considered to be a relatively new technique, some limitations outlined below came up during assessing the process. - Although the microbiological transformations of the wastewater in the sewer system have been studied by many researchers, mostly covering the kinetics of aerobic and anaerobic conditions, there is still a lack of information for those of transformations under anoxic condition. - All the calculations in this report are based on the activated sludge method due to the limited knowledge regarding the in-sewer denitrification process under anoxic conditions. - As this project has not begun yet, the data from the nitrification process at Dalby WWTP, which is required to evaluate the denitrification process, is not available. Therefore, EFOR software has been applied to model the nitrification process and its results are used in the evaluation of the denitrification process in the sewer pipe. - EFOR only has the capability of simulating microbial transformation in bulk water phase of wastewater. The denitrification process in biofilm phase has been calculated manually. 2

13 2. Background Dalby is located in Lund municipality, 10 km south-east of Lund, and the population was reported to be 5517 in 2005 (Statistiska Centralbyrån, 2006). Figure 1 shows Lund municipality and Dalby community (Lund municipality, 2008). Dalby wastewater treatment plant situated in a cultivated area one kilometer south west of Dalby was designed in 1981 and constructed in The plant is capable of handling wastewater from 7500 connected users. The statistics show that until late 2007 ( ), the plant received extra wastewater from Alba margarine factory which generates wastewater equivalent to 130 persons (70 g BOD/(person).day) in addition to Dalby s 5506 residents which were already connected to the WWTP (Dalby Miljörapport, 2007). Figure 1: Lund municipality (Lund municipality, 2008). 3

14 Due to the anticipated future population growth, Lund municipality, which is in charge of the Dalby WWTP, has decided to increase the WWTP capacity and enhance the treatment quality. Because of high costs and operational limitations to expand Dalby WWTP, it has been decided to transport the partially treated wastewater through a pipeline from Dalby to Källby WWTP in Lund. A post-denitrification method is going to be used in the Dalby WWTP. After specific pretreatment and nitrification processes which will be conducted in the Dalby WWTP, nitrified wastewater will be pumped into the pipe to be denitrified under anoxic conditions. In other words, the pipe is applied not only for transporting wastewater to the Källby WWTP but also to enhance its quality by denitrification method under sewer condition. The pipeline has been designed in two different sections, the first section which transfers nitrified wastewater under pressure and the second part under the influence of gravity. It is designed to be joined to the sewer network in Lund. In the first phase, only Dalby s wastewater will be transported to the Källby WWTP. In the next phase, wastewater from Genarp, Veberöd and Björnstorp will be connected to the Dalby WWTP (See Figure 1). The pipeline which is going to connect the two WWTPs together has been hydraulically designed at Tyréns AB. 4

15 3. Literature study; Biological nitrogen removal 3.1. Biological treatment: Wastewater contains a relatively large diversity of bacterial types which in aerobic, anoxic and anaerobic conditions degrade organic wastes for their cellular growth and reproduction. The required oxygen demand which micro-organisms utilize to degrade organic matter is expressed as biochemical oxygen demand (BOD). Oxygen demand is measured in 5 or 7 days (BOD 5 or BOD 7 ) and indicates the extent of the pollution in the wastewater. Different bacteria decompose specific organic matters due to their properties (Folder, 2006). Among high concentrations of bacteria in the activated sludge process organotrophic bacteria degrade organic matters in the wastewater and produce energy and carbon dioxide. A group of bacteria which use inorganic substances in the wastewater as an energy source for reproduction are called Chemolithotrophs bacteria. Nitrifying bacteria, which are included in this category, oxidize ammonium to nitrite and nitrate (Gerardi, et al., 2003) Nitrification Nitrification is a biological process in which ammonium ions (NH 4 + ) are oxidized to nitrite ions (NO 2 - ) and eventually nitrate ions (NO 3 - ) in aerobic respiration. The nitrification process progresses by two general groups of bacteria, ammonium oxidizer and nitrite oxidizer bacteria, in two specific steps (Expression 1 and 2) (Gerardi, et al., 2003): NH O 2 NO H + + H 2 O + Energy (1) NO O 2 NO Energy (2) In the nitrification process, nitrite does not extensively accumulate so nitrification mostly progresses through ammonium oxidation to nitrate in a combination of two energy yielding stages (Expression 3): NH O 2 NO H + + H 2 O + Energy (3) Nitrogen can be mostly found in the form of ammonia (NH 3 ) and ammonium (NH 4 + ) in the wastewater but ammonia cannot participate in the nitrification process. There is a constant total quantity of ammonia and ammonium ions in the wastewater and at a specific ph, ph=9.4, their concentrations are equal. Ammonium ions are the dominant form in the influent at a lower ph and vice versa (Gerardi, et al., 2003). As can be seen in the Expression 1, the nitrification process produces acid which may decrease ph drastically if alkalinity is low. In order to raise the ph to the required level lime can be added to the nitrification process (Lindquist, 2003). 5

16 The temperature can affect the nitrification rate extensively. At low temperatures, down to 4 C, bacteria reproduce very slowly resulting in the need for a higher sludge age in the nitrification basin. In general, some factors such as low temperature, short retention time in the aeration basin, sludge discharge of soluble organic matters and low dissolved oxygen level can adversely affect nitrification rate (Lindquist, 2003; Henze, et al., 1996) Denitrification In the aerobic respiration, ammonium ions are oxidized to nitrate ions during the nitrification process. Therefore the nitrogen concentration in the wastewater does not decrease and the nitrogenous compounds are just transformed to other forms. The preferred oxidizing agent, free molecular oxygen, takes part in chemical reactions which produces much of the energy required for bacteria growth. When free oxygen molecules are depleted in the system (less than 3 g/m 3 ), bacteria begin to utilize nitrate ions and nitrite ions as electron acceptors, which have been produced in the nitrification stage, to degrade substrates. Denitrifying bacteria, which are almost 80 % of the total concentration of bacteria in the activated sludge process, use nitrite or nitrate to degrade organic substrates. This procedure, known as denitrification, removes nitrogen from the wastewater by transforming nitrate ions to nitrogen gas in two stages (Expressions 4 and 5) (Gerardi, et al., 2003). NO cbod NO CO 2 + H 2 O (4) NO cbod N 2 + CO 2 + H 2 O (5) Denitrifying bacteria may produce nitric oxide and nitrous oxide gas in an incomplete denitrification procedure. In the denitrification process nitrate ions and nitrogen gas are considered as initial and latter components respectively. Nitrite ions, nitric oxide and nitrous oxide can be produced as secondary products. The possible nitrogenous compounds in denitrification process are presented in Table 1 (Gerardi, et al., 2003). Table 1: Possible nitrogenous compounds in the denitrification process (Gerardi, et al., 2003). Nitrogenous Compound Formula Nitrate ion - NO 3 Nitrite ion - NO 2 Nitric oxide NO Nitrous oxide N 2 O Molecular nitrogen N 2 6

17 Different methods of denitrification Denitrification takes place in an anoxic condition in the absence of free molecular oxygen and presence of nitrate and/or nitrite ions. Therefore denitrification is commonly carried out in a combined system with nitrification. The methods are classified as pre-denitrification and postdenitrification processes, depending on which one of the nitrification and denitrification stages has been applied in advance. These two methods have advantages and drawbacks which are considered in different cases. In pre-denitrification, the organic matter in the influent entering the denitrification chamber is used as an internal carbon source so no external one is needed. Therefore in the absence of dissolved oxygen (anoxic condition) in the denitrification basin, nitrate is converted to nitrogen gas which leaves the wastewater as bubbles. Consequently the denitrified wastewater is conveyed to the nitrification basin. In the nitrification process, ammonium ions are oxidized to nitrite and nitrate ions in an aerobic respiration with dissolved oxygen, which is continuously provided by aerators in the aeration basin. Internal recycling is applied to transfer the high concentration of nitrate to the denitrification basin and enhance the denitrification process under anoxic conditions. Eventually, the returned sludge, which is rich in bacteria, goes back to the denitrification basin while influent (raw wastewater) provides organic matter as the required carbon source in the denitrification process. In post-denitrification, ammonium ions are converted to nitrite and nitrate ions in the presence of dissolved oxygen in the nitrification basin. Nitrified wastewater conveyed to the denitrification basin contains low concentrations of organic substrates since BOD-reducing bacteria degrade organic matters at a high rate in the nitrification basin. Therefore, easily biodegradable substances (as methanol, ethanol, etc) are commonly applied as the external carbon source in the post-denitrification method. Although methanol is relatively expensive to apply in this process, it can extensively enhance the nitrogen removal process. Consequently, nitrate ions transform to nitrogen gas in anoxic conditions. The returned sludge goes back to the nitrification stage to balance the nutrient/bacteria ratio in the system (Gerardi, et al., 2003; Henze, et al., 1996). The denitrification rate depends on some factors such as adequate substrate and denitrifying bacteria, absence of dissolved oxygen, ph, temperature and redox potential. The substrate removal rate of organic matters in the denitrification process can be estimated with Expression 6 (Lindquist, 2003; Henze, et al., 1996). (6) : Volumetric nitrate removal rate (kg COD/m 3.d) : Maximum specific growth rate (d -1 ) : Maximum yield constant (g COD Biomass /g COD Substrate ) : Saturation constant (g/m 3 ) : Biomass concentration (kg COD Biomass / m 3 ) : Dissolved biological easily-degradable organic matter (kg COD Substrate / m 3 ) 7

18 C/N ratio The organic substrates are used as the internal source of carbon and energy in the denitrification process. The carbon/nitrogen ratio (C/N ratio) is an important factor in the denitrification process since denitrifying bacteria need certain amounts of organic matter to transform nitrate to nitrogen gas. The C/N ratio as a rough estimation can be calculated as COD concentration available in the influent (COD/N ratio). This ratio evaluates the possibilities of the process and the concentration of the external carbon source if required. Transformation of nitrite and nitrate to nitrogen gas, production of sludge and respiration with oxygen are three main reactions in the denitrification process. All these reactions consume organic matter which indicates the required concentration of organic matter in the anoxic system. Expression 7 expresses that consumption of 1 mole nitrate needs 1.25 mole COD (substrate) which corresponds to 2.86 kg COD/kg NO 3 -N. Hence the C/N ratio can be calculated by dividing total COD-consumption of wastewater with total nitrogen removal in the denitrification process (Expression 8) (Henze, et al., 1996). NO COD 0.5 N 2 + CO 2 (7) C/N = (2.86. NO 3 -N + F SP + F O2 ) / (NO 3 -N + F SP. f B,N ) (8) NO 3 -N: Denitrified amount of nitrate (kg COD/d) F SP : Sludge production (kg COD/d) F O2 : Oxygen respiration of organic matters (kg O 2 /d) f B,N : assimilated nitrogen in sludge production (kg N/kg COD) The optimum C/N ratio for different types of organic matters in the denitrification process is presented in Appendix 1. The COD-reduction occurring by oxidation of the organic matter with oxygen in the nitrification process is considered as waste of organic matters. This reaction which results in a higher required C/N ratio in practice in the denitrification process takes into account the efficiency factor (f C/N ). This factor, which expresses the available portion of the internal carbon in the system, depends on the design of the plant and the control of the process. The efficiency factor for organic matter in different denitrification plants can be found in Appendix 2. The C/N ratio in practice can be calculated by Expression 9 (Henze, et al., 1996). (9) If the C/N ratio in the system is relatively lower than the C/N ratio in practice, the internal carbon source is not enough for complete nitrate reduction in the denitrification process. Therefore an external carbon source such as methanol should be applied to compensate the carbon source deficit. 8

19 3.2. Sewer as a reactor Municipal sewer networks collect wastewater and storm water from different sources such as household s wastewaters and transport them to the wastewater treatment plants. Whereas the sewer network has been applied to reduce the odor nuisance and hygienic problems, the microbiological transformations in the sewer system can cause extreme pipe corrosion, odor nuisance and health problems. Despite the fact that WWTPs are assumed to be the only section of the process dealing with microbiological treatment, the sewer network also takes part in the physiochemical treatment processes. This means that the treatment procedure is considered as starting at the kitchen sink (Vollertsen, et al., 2005). The biological transformation in sewers can proceed either in the suspended water phase, the biofilm, the sewer sediments or the sewer atmosphere and walls in contact with the air phase shown in Figure 2 (Hvitved-Jacobsen, 2002). Generally speaking, in the bulk water phase of the sewer, only fast growing bacteria can significantly take part in the biological processes due to short residence time of the wastewater in the sewer system. Both fast and slow growing bacteria can contribute in microbial transformations in other phases of the sewer (biofilm and/or sediment) (Hvitved-Jacobsen, 2002; Vollertsen, et al., 2005). Figure 2: The sewer as a reactor for biological processes (Hvitved-Jacobsen, 2002). Wastewater can be transported in sewers by both gravity and pressure. In gravity sewers, the length and slope of the pipe are critical factors in the design phase. If the slope of the pipe is not big enough, the risk of sediment depositions can come up along the pipe. Partially filled gravity sewers can be re-aerated to provide an aerobic condition along the sewers. Wastewater transferred under pressure is full flowing without the possibility of re-aeration which experiences anaerobic or anoxic (if nitrate has been added) respiration (Hvitved-Jacobsen, 2002). 9

20 Kinetics of redox processes Living microorganisms (biomass) in wastewater need energy, for maintaining life and reproduction; this is provided by degradation of organic matter. The degradation process is a catabolism by oxidation of substrates which depends on the availability of electron acceptors. The transformations of the electron donor (organic matters) and the electron acceptors are a case of study for in-sewer processes. In these reduction-oxidation (redox) processes, substrates acting as electron donors are oxidized by electron acceptors such as dissolved oxygen, nitrate or sulfate. Basically, the design properties and operational functions of the sewer network define which redox condition will prevail in the system (Hvitved-Jacobsen, 2002). The electron acceptors and the process condition in which they may occur can be found in Table 2 (Hvitved-Jacobsen, et al., 2002). Table 2: The electron acceptors and the possible process conditions in sewer networks (Hvitved- Jacobsen, et al., 2002). Process conditions Electron acceptor Typical sewer system characteristics Aerobic + Oxygen Partly filled gravity sewer Aerated pressure sewer Anoxic - Oxygen Pressure sewer with addition + Nitrate of nitrate Anaerobic - Oxygen Pressure sewer - Nitrate Full flowing gravity sewer + Sulfate Gravity sewer with low slope Generally, substances with high redox potential tend to take part in reduction-oxidation reactions. Dissolved oxygen has a high redox potential and produces high energy during the oxidation process. So as long as aerobic conditions prevail in the sewer networks, oxygen is utilized as an oxidizing agent and the biomass yield is relatively high. The aerobic respiration in sewers is fast and has a high biomass yield. As soon as oxygen is depleted, nitrate which also has a relatively high redox potential acts as an oxidizing agent in an anoxic respiration. Anoxic respiration is not common in sewers since the concentration of nitrate is generally low in the wastewater. If neither oxygen nor nitrate is present, as in anaerobic condition, sulfate will take part in the degradation of substrate. This reaction will produce hydrogen sulfide, volatile substances and alcohol. In the anaerobic condition, the oxidation rate is lower and the biomass yield is smaller compared to the aerobic and anoxic heterotrophic processes. Beside the sulfate reduction, organic matter can also take part in redox processes in the anaerobic respiration, as both an electron acceptor and donor. This reduction-oxidation process is known as fermentation, during 10

21 which certain organic compounds transform to other forms of organic compounds for instance volatile fatty acids (Vollertsen, et al., 2005). In the following parts, the kinetics of anoxic and anaerobic conditions in sewer networks is studied. (Since aerobic condition is not the case of study in this project, microbiological processes in aerobic respiration are not discussed here.) Anoxic condition Anoxic conditions occur in the sewer network when dissolved oxygen is depleted in the wastewater but nitrate is available, in sufficient concentrations. However, nitrate concentrations are not relatively high in the conventional wastewater. Therefore, aerobic and anaerobic situations are the most probable conditions which can prevail in the sewer system. In order to prevent anaerobic condition in the sewer network, anoxic conditions can be induced in sewers by the addition of nitrate to the wastewater. This procedure can control the risk of sulfide formation and also enhance the degradation process of microorganisms as electron acceptors. Since the nitrogen uptake rate (NUR) is in the same range as the oxygen uptake rate (OUR), measured in electrons transferred, the anoxic respiration is an alternative option to be used in sewers instead of providing aerobic condition by aeration methods. Depending on where transformations are happening, either in the water phase or in the biofilm of the pipe, the denitrification rate can be estimated. According to Hvitved-Jacobsen (2002), the denitrification rate in sewers, which has considered bulk water and biofilm phases, can be calculated by the following expressions. : Nitrate removal rate in the water phase and the biofilm (g NO 3 /m 3 h) : Nitrate removal rate in the water phase under substrate nonlimited conditions (g NO 3 /m 3 h) : Nitrate removal rate in the biofilm under substrate nonlimited conditions (g NO 3 /m 3 h) (10) : Total nitrate removal (g NO 3 /m 3 h) : Anoxic residence time in the pipe (h) (11) Despite aerobic and anaerobic conditions, very few studies have been done on the denitrification process in sewer networks. However, nitrate addition to prevent anaerobic respiration and its deteriorative effects have been reported in many studies. Some relevant case studies regarding to the denitrification in sewer networks are discussed in the following section. 11

22 Abdul-Talib, et al., (2002) has studied the denitrification process in the bulk water phase of municipal wastewater in sewer networks. Different concentrations of nitrate were introduced as the electron acceptor and results show that as long as nitrate concentration is above 5 g NO 3 - N/m 3 the denitrification process was not limited by the nitrate concentration. The insignificant accumulation of nitrous oxide and nitric oxide (1% and 0.01% of total nitrogen) in the system revealed that nitrite and nitrate ions react as electron acceptors in the denitrification process. The results of nitrogen compounds concentration tests in both stages are presented in Figure 3. In this study, denitrification proceeded in two distinct stages. Nitrate and nitrite removal rates can be estimated from stage 1 and 2 in Figure 3. In the first stage, nitrate was utilized as the electron acceptor and produced nitrite and nitrogen. The nitrate reduction rate, 0.22 mmol/h, was much higher than the nitrite reduction rate, 0.14 mmol/h. This was the reason why the nitrite accumulation occurred. Nitrate inhibited nitrite reduction in stage 1 and nitrite accumulation had its maximum concentration while nitrate was depleted in the system. When nitrate was depleted, stage 2, nitrite had been utilized at a much higher rate, 0.24 mmol/h, than the nitrate and nitrite utilizations in stage 1. Figure 3: Variation of nitrate, nitrite and nitrous oxide during denitrification of the bulk water phase (Abdul-Talib, et al., 2002). Denitrification kinetics, which is presented in this study, differs from the denitrification process occurring in the activated sludge method in WWTPs. In the activated sludge method nitrate is utilized and nitrogen gas is produced without any significant accumulation of nitrite or other nitrogen compounds. It could be because of the higher amount and variety of biomass in the activated sludge system, compared to sewer systems, and/or more acclimatization of biomass in the anoxic condition in the activated sludge method. 12

23 The denitrification rate in the water phase, when sufficient substrates and electron acceptors are available, were found in the range of g NO 3 -N/m 3 h. In order to provide anoxic respiration in the length of one meter of the pipe, the total denitrification rate in a sewer pipe can be calculated by Expression 12 (Abdul-Talib, et al., 2002): (12) : Total denitrification rate : Denitrification rate in the biofilm and water phase respectively : Wetted area to volume ratio of the sewer By assuming 2 g NO 3 -N/m 3 h and 0.15 g NO 3 -N/m 2 h for r Film and r Water in this equation, the required nitrate for prevailing anoxic condition in the sewer pipe can be estimated (Abdul-Talib, et al., 2002) Anaerobic condition Anaerobic conditions occur when oxygen and nitrate are depleted in the wastewater and bacteria use sulfate (SO 4 2- ) as the electron acceptor, which are reduced, and produce hydrogen sulfide (H 2 S). This reaction mostly progresses in the biofilm and sediment phases in the sewer, where the biomass residence time is longer, and diffuses to the bulk water (Expression 13). Anaerobic respiration mostly prevails in full-flowing gravity sewers and pressure mains. (Vollertsen, et al., 2005). SO Organic carbon HCO 3 - (carbon dioxide) + H 2 S (13) Beside sulfate reduction, fermentation and methanogenesis also occur, at a low rate, under anaerobic conditions. In fermentation, organic matter is used both as electron donors and acceptors. Although degradation of organic sulfur compounds such as certain proteins and amino acids results in hydrogen sulfide production, the most of the H 2 S concentration is produced by sulfate reduction. Studies indicate that high temperature and residence time in sewers are critical factors for increased sulfide formation in the system. Sulfide concentrations of 0.5, 3, 10 g S/m 3 in sewers are considered as low, moderate and high respectively in terms of problems which may occur (Hvitved-Jacobsen, 2002). Sulfate concentration and the quality and quantity of biodegradable organic matter can affect the progress of sulfate reduction. Typically sulfate can be found in concentrations greater than 5-15 g S/m 3 in all wastewaters (Hvitved-Jacobsen, 2002). Although sulfate-reducing bacteria adopt themselves to both high and low temperatures, their temperature dependency in this reduction process is rather high and is evaluated by temperature coefficient (α). Hydrogen sulfide formation is also dependent on ph since the dissociation constant for HS - and H 2 S is close to 7. This means that HS - and H 2 S concentrations are the same at ph 7. Area-to-volume ratio, flow velocity and anaerobic residence time are other factors controlling sulfide formation in biofilms 13

24 and sediments in the sewer network. (Hvitved-Jacobsen, 2002; Vollertsen, et al., 2005; Yang, et al., 2005) In pressure pipes, anaerobic condition occurs right after the initial aerobic phase which is caused by pumping. Since pressure sewers are fully-flowing pipes and there is no air phase in the sewer system, anaerobic conditions occur much sooner than in gravity sewers. The sulfide production can be estimated by Expression 14 (Hvitved-Jacobsen, 2002): (14) : Sulfide production transformed to the water phase (g S/m 3 ) : Sulfide concentration at the end of the sewer section (g S/m 3 ) : Sulfide concentration at the start of the sewer section (g S/m 3 ) : Areal sulfide production rate (g S/m 2 h) : Area to volume ratio (m -1 ) : Aerobic residence time in the pipe (h) In gravity sewers, the anaerobic condition can be controlled by aeration methods. The hydrogen sulfide is not produced in the gravity sewers when the dissolved oxygen concentration exceeds g O 2 /m 3 (Hvitved-Jacobsen, 2002). Presence of an anaerobic condition in sewers can cause adverse effects in both sewer networks and connected treatment plants. Health issues, odor problems, corrosion of concrete and metal pipes and anaerobic wastewater conveyed as influent to WWTPs are example of effects which may occur Corrosion risk The production and deterioration effects of hydrogen sulfide (H 2 S) have been studied for many years as one of the critical concerns in sewer networks (Boon, 1995; Hvitved-Jacobsen, 2002; Yongsiri, 2005; Vollertsen, et al., 2005; Yang, et al., 2004; Yang, et al., 2005; Mathioudakis,et al., 2006). Although dissolved sulfide in the wastewater does not threaten the sewer pipe walls, partial pressure of hydrogen sulfide which becomes higher in the water phase results in the emission of hydrogen sulfide in the air phase of the sewers. Consequently Chemoautotrophic bacteria oxidize hydrogen sulfide, with the available oxygen at the moist surface of the pipes, and produce sulfuric acid (Expression 15) (Vollertsen, et al., 2005). H 2 S + 2 O 2 H 2 SO 4 (15) The produced sulfuric acid in the air phase reacts with alkaline cement in the concrete pipes and causes substantial damages. The product in this reaction is gypsum which does not have a strong structural resistance and consequently results in reduction in sewer wall thickness and ultimately a reduction in structural integrity. High temperature and high concentrations of hydrogen sulfide 14

25 increase the corrosion rate in the sewer network. The serious concrete corrosions have been reported despite low concentration of the H 2 S concentration ( mg/l). It occurs because the ventilation is not appropriate to reduce the emitted hydrogen sulfide gas in the sewer and the turbulent wastewater flow enables H 2 S emission to the sewer atmosphere (Boon, 1995; Hvitved- Jacobsen, 2002). Yongsiri, et al., (2005) investigated the effects of wastewater constituents on the emission rate of hydrogen sulfide (H 2 S) in the domestic wastewater under sewer conditions and figured out the impurities interference in the reduction rate of H 2 S emission. Vollertsen, et al., (2008) has studied the hydrogen sulfide formation and also the corresponding corrosion in concrete surfaces. In a series of experiments a specific concentration of hydrogen sulfide gas was injected to the sewer pipes. Within a few months high concrete corrosion, which was much faster than the emission of H 2 S gas to the sewer atmosphere of a gravity sewer, had been observed. The research indicated that the corrosion rate, to a large extent, is controlled by the hydrogen sulfide emission rate to the sewer atmosphere and that corrosions continue even in low concentration of H 2 S in the gas phase Odor nuisance and health problems Odor-causing substances produce inorganic gases, such as ammonia (NH 3 ) and hydrogen sulfide (H 2 S), and volatile organic compounds (VOCs) in anaerobic conditions. A list of malodorous substances which may be found in different steps of the sewer processes are presented in Appendix 3. Hydrogen sulfide is a poisonous component with odor of rotten eggs. It can be emitted to the sewer atmosphere in manholes, wet-wells and mainly in gravity sewers. Beside odor nuisance, many different health problems may occur with high concentrations of H 2 S, which are generated in the sewer networks in an anaerobic respiration. Some odor and health problems by H 2 S are presented in Appendix 4 (Hvitved-Jacobsen, 2002) Anoxic condition prevalence in sewer to prevent sulfide formation In order to prevent anaerobic conditions, which consequently reduce the corrosion and odor nuisance risks in sewer networks, or oxidize the produced sulfide in an already existing anaerobic conditions (anoxic sulfide oxidation), nitrate can artificially be injected to the sewer system (Zhang, et al., 2008; Bentzen, et al., 1995; Yang, et al., 2004; Yang, et al., 2005; Mathioudakis, et al., 2006). Since the nitrate-reducing bacteria reproduce more quickly than the sulfate-reducing bacteria, nitrate, as electron acceptor, is the dominant electron acceptor in redox processes in this method. Yang, et al., (2005) has studied the anoxic sulfide oxidation under sewer conditions through some batch experiments. The experiments included both fresh wastewater and wastewater from the inlet of a WWTP which was transferred 30 km in a sewer network. The sulfide concentration in the fresh wastewater had been artificially provided by adding sodium sulfide (Na 2 S). Sodium 15

26 nitrate (NaNO 3 ), as the nitrate solution, had been added to all samples after 30 minutes, which provided the initial nitrate concentration of about 50 g NO 3 -N/m 3. The results showed sulfide oxidation in all experiments but at different rates. The average oxidation rates in fresh wastewater were 0.48 and 0.62 g S/m 3 h at ph 7 and 8.5 respectively. Nitrite accumulation was observed as an intermediate product during nitrate reduction. The sulfate reduction and nitrite accumulation rates in the sample, which was 3 days under anaerobic condition in sewer system (ph = 8.5), have been shown in Figure 4 (Yang, et al., 2005). This experiment indicated reduction of nitrite accumulation while sulfide was depleted in the system which implied that nitrite accumulation was a contributing factor in the anoxic sulfide oxidation. The highest oxidation rate had been seen in the fresh wastewater samples. Sulfide oxidation rates in the samples taken from the sewer network, which had spent 6 days under anaerobic conditions, were extremely lower than those of the fresh wastewater (only 10 % of the oxidation rate in the fresh wastewater). It can be inferred that the longer presence of the wastewater in the anaerobic condition of the sewer network resulted in a reduction of biomass of sulfide reduction bacteria (SOB) and eventually a lower sulfide oxidation rate. The required nitrate concentration for anoxic sulfide oxidation was only 10 % of the total nitrate concentration which was consumed for these experiments. This implied that heterotrophic denitrification processes occurred in these experiments. In this study nitrite accumulation has been observed within the anoxic sulfide oxidation, which was also reported in the anoxic respiration process (Abdul-Talib, et al., 2002). Figure 4: Anoxic sulfide oxidation and nitrite accumulation rate in two scenarios with different initial sulfide concentrations (Yang, et al., 2005). Mathioudakis, et al., (2006) has investigated the effects of nitrate addition in reducing hydrogen sulfide concentrations, both in laboratory and field experiments. In the laboratory experiments, two different situations are studied to evaluate inhibition of sulfide formation and/or the rate of anoxic oxidation of sulfide. In the first scenario, potassium nitrate was initially added to the batch reactor without sulfide. Results showed inhibition of sulfide formation in the system. In the second scenario, in which potassium nitrate is applied to the anaerobic wastewater, nitrate reduced the sulfide concentration in the reactor and prevented sulfide production. The results of the two batch reactor experiments at two different temperatures are illustrated in Figure 5 (Mathioudakis, et al., 2006). As can be seen in Figure 5, the dashed lines represent the excess 16

27 nitrate concentration after dissolved sulfide has been completely oxidized. The excess concentration has been consumed in a heterotrophic denitrification process. Figure 5: The nitrate and dissolved sulfide concentrations in lab experiments (Mathioudakis, et al., 2006). In the field experiments, the wastewater samples were taken from a 6.7 km combined sewage network (pressure and gravity sewers) in which different concentration of ammonium nitrate had been added at the first station (pumping station) for 4 to 8 hours. The results of measured dissolved sulfide are presented in Figure 6 (Mathioudakis, et al., 2006) and indicate a reduction of dissolved sulfide by applying different dosage of ammonium nitrate. The nitrate concentration at the inlet of the connected WWTP was quite low in all cases in the field experimet and did not affect the subsequent treatment processes in the WWTP. Figure 6: The dissolved sulfide concentration along the sewer network with and without addition of ammonium nitrate (Mathioudakis, et al., 2006). 17

28 3.3. Sewer as a biofilm reactor A biofilm is a dense layer of living microorganisms produced by adhering to a solid medium. The firmly attached cells of bacteria form a fixed film on the solid surface of the sewer pipe which protect the bacteria from washing out. Living microorganisms needing source of energy for growth and reproduction utilize the substances in the wastewater carried out through the sewer. This slow reaction is progressed by molecular diffusion of the oxidant and reductant compounds in the biofilm layer which provides the possibility of redox reactions and nutrient removal. In an active and thick biofilm, substrates can diffuse from the bulk water phase into the biofilm phase of the sewer pipe to a certain depth. Redox converions can only occur as deep as both oxidant and reductant compounds penetrate into the biofilm layer. The efficiency factor is a dimensionless value for evaluating geometry and diffusion kinetices in the biofilm layer. It is defined as a criterion to assess the degree of penetration for different substances in the biofilm. In 100 percent efficient biofilm, substrates can fully penetrate into the biofilm and the nutrient removal occurs according to the zero order reaction. On the other hand, partially efficient film, in which substances cannot penetrate to the biofilm s bottom results in a half order reaction in the biofilm layer. Kinetic constants such as diffusion coefficient, D, and area specific rate constant for zero and half order reactions, k 0A and k 1/2A, determine the penetration and removal rate for different substances. The stoichiometric coefficient, a kinetic constant in biofilms, represents the contribution of different species in the chemical reactions. For instance if the stoichiometric coeficient for species A is twice that of B in a chemical reaction, twice the concentration of species A will take part in the reaction. A list of kinetic constants for oxygen and organic substances in biofilms is presented in Appendix 5 (Henze, et al., 1996; Huisman, et al., 2003). A redox reaction will occur in the biofilm where both electron acceptors and donors are available. As various substances have different diffusion coefficients and rate constants, the distance which oxidant and reductant compounds can penetrate into the biofilm is different. As soon as one of the substrates is used up, no more redox reaction and eventually nutrient removal will occur. Depending on which substrate penetrates the shortest distance into the biofilm, the reaction is reductant or oxidant limiting. By using Expression 16 one can estimate whether oxidant or reductant substances are limiting in the biofilm phase of the sewer network (Henze, et al., 1996). (16) > : Reductant is potentially limiting for removal = : Neither reductant nor oxidant or both of them are limiting for the removal < : Oxidant is potentially limiting for removal Substances with high redox potential take part in redox reactions and yield energy. As deep as dissolved oxygen can penetrate into the biofilm layer, while enough electron donors (reductant) are available, aerobic condition is the dominant redox process. Once oxygen is depleted, nitrate 18

29 will take part to degrade substrate as the electron acceptor if both electron acceptors and donors penetrate deep enough into the biofilm. In a thick biofilm, where organic matters is still available, anaerobic condition can prevail in the sewer network where neither oxygen nor nitrate is present. The mass transfer in biofilm is presented in Figure 7 (Huisman, et al., 2002). Figure 7: The sketch of mass transfer in the ideal case (Huisman, et al., 2002). Huisman, et al., (2002 and 2003) has investigated the role of the biofilm on the sewer wall in microbiological transformations in the sewer network. The results indicate that despite the fact that the biofilm is active and fully penetrated, the observed redox reactions are at a lower rate compared to maximum redox reactions. The effectiveness factor (η f,i ) is expressed as the ratio of the observed and maximum redox reactions. Thereby, the reaction rate in the biofilm has been calculated by Expression 17. (17) : The consumption of substrate i in the biofilm (g.m -3.d -1 ) : The specific biofilm area (m 2.m -3 ) : Biofilm thickness (m) : Stoichiometric coefficient for compound i and process j (g.g -1 ) : Process rate for process j within the biofilm (g. m -3.d -1 ) The penetration depth of different substrates into the biofilm ( ) is estimated by Expression 18. It is calculated for all substances and the smallest one which indicates limiting substrate in the system is selected for estimating the redox reactions within the biofilm. A saturation term is used to control the penetration depth according to initial concentration of the substrates. (18) 19

30 3.4. Integrated operation of the sewer system and WWTP Sewer networks and WWTPs are conventionally evaluated as two separated phases of the urban wastewater management. Since the wastewater, that has been collected by the sewer system, is transported to the WWTP, WWTP can be affected by probable changes which may occur in the wastewater in the sewer system. Enhanced management of sewage networks, effective treatment processes in WWTPs and low concentrations of nutrients in the effluent are main goals in the wastewater management strategy. These goals bring forward the necessity of an integrated design of sewers and WWTPs (Huisman, et al., 2003). The combined sewer network of Copenhagen and the overload risk in the influent to the WWTP has been studied by Harremoës, et al., (2002). This scenario was modelled in the software MOUSE-SAMBA. In order to avoid overflows of the influent during wet seasons, different avilabilities of discharging wastewater to the local water bodies has been investigated. The first option is at the entrance from which stormwater can be discharged and needs restructuring. The next feasible discharge, before treatment stages, is applied in emergency cases only. There is a detention tank, which has four times the capacity of the biological basin, storing biologically treated wastewater to regulate discharging. Applying a detention tank to store wastewater and storm water temporarily, when the overflow occurs, can effectively reduce the overflow risk, by up to 50-80%. If the flow exceeds the capacity of the biological basin, the bypass from influent will discharge the excess wastewater to the marin waters in Öresund. The local rivers and lakes have a higher priority, compared to marine areas, to be protected environmentally because they are much more sensitive to the the nutrient accumulation and eutrophication. The integration of in-sewer microbial transformations and treatment processes in the WWTP will result in a sustainable model for designing wastewater management systems. Therefor the sewers should not only be designed to transport wastewater but also to achieve a better wastewater quality, by in-sewer transformations, to enhance the treatment processes in the WWTP and in the effluent (Vollertsen, et al., 2002). As mentioned before, sewer networks which have been applied to transport wastewater to WWTPs are subjected to microbiological transformations during different conditions (aerobic, anoxic and anaerobic). Huisman, et al., (2003) has investigated the transformations in sewer networks which influenced the connected WWTP. In this scenario, biofilm kinetics and its contribution to bacteria growth is modelled with ASM3 software. In an active and thick biofilm, degradation of substrates apparently depends on how deep the substrate can penetrate in the biofilm layer. During the dry season, in which the sewer network conveys a normal flow of wastewater, biofilm grows on the wall of the pipes which contains high concentrations of organic and inorganic materials. When the sewage pipes are exposed to high flows, during a rain event, all organic and inorganic particles which have accumulated on the sewer walls will be washed out and this may cause clogging in the system and raise the oxygen uptake rate in the WWTP. This study indicates that if nitrate is available in the wastewater, denitrification will proceed in a high rate in the biofilm layer, which can denitrify the daily nitrogen load by as much as 30%. After the rain event, the flow rate, the oxygen concentration in the sewer and the COD concentration increased which eventually resulted in reduction of efficiency in the nutrient removal process in the WWTP. The higher oxygen concentration in the wastewater was due to 20

31 the less active biomass, and more effective aeration, because of higher flow rate, and a higher concentration of oxygen in the wastewater. Ahnert, et al., (2005) has studied the effects of in-sewer transformations on the WWTP processes by considering its temperature dependancy in different seasons. This study indicated that a low concentration of dissolved oxygen in the sewers in the summer resulted in a denitrification of the low nitrate concentration in the system. COD concentration in the influent of the WWTP reduced during summer while the in-sewer degradation of biodegradable organic matters increased which resulted in less biodegradable substances in the biological process in the WWTP. The WWTP which has worked efficiently in winter periods with low temperature may face operational problems at high temperatures during summer due to in-sewer transformations in the sewer network. Higher production of excess sludge, reduction of substrate in the WWTP for biological purposes and an increasing demand of the required external carbon source for denitrification are some examples of inconveniences which may arise in the WWTP. 21

32 22

33 4. Treatment processes in Dalby WWTP Total influent to the wastewater treatment plant in Dalby is reported to be m 3 /year and the concentration of constituents in the influent wastewater is presented in Table 3 (Dalby Miljörapport, 2007). Table 3: Average and maximum concentration of constituents in the influent wastewater (Dalby Miljörapport, 2007). Parameter Average concentration (mg/l) Maximum concentration (mg/l) BOD Total nitrogen Total phosphorous Treatment processes in the Dalby WWTP are planned to be changed to enhance the management and quality of wastewater in Dalby community. Here, the treatment techniques, both present and future, are presented Present approach i) Pre-treatment In order to prevent the blockage of pipes and damage to mechanical equipments, stone, sand and large pieces of waste are removed in pre-treatment processes. Two pre-treatment steps are used in Dalby WWTP, sand trap and screen. In the sand trap, stones and sands settle by gravity and be removed from the system. It is V-shaped in cross-section and is deep enough to compensate for turbulence. Subsequently, wastewater passes through the screen in which heavier solids such as grit and large suspended solids are removed. ii) Biological treatment Biological treatment in Dalby WWTP comprises BOD removal and biological nitrogen removal which are carried out by a pre-denitrification process. The denitrification tank has a volume of 300 m 3 is used to reduce nitrate to nitrogen gas under anoxic condition. The organic matter in the influent wastewater is as internal carbon source in the denitrification process. Subsequently denitrified wastewater entering the nitrification tanks with a volume of 600 m 3 is aerated by bottom aerators. Figure 8 shows the nitrification tanks in the Dalby WWTP. In the presence of dissolved oxygen, ammonium is reduced to nitrite and nitrate ions. High concentration of nitrate and biomass in the nitrification basin is internally recycled to the denitrification basin required for denitrification process. 23

34 Figure 8: The nitrification tank which is used, in the present approach, in the Dalby WWTP. iii) Settling tank: In the next step, sludge settles in two settling tanks from where sludge is partly returned back to the denitrification tank to balance the biomass concentration in the biological process. The excess sludge is removed from the system. After sedimentation, sludge is centrifuged and lime is added to the sludge to increase the ph and destroy the pathogens. Eventually sludge is stored and transferred to Källby WWTP to be used in fields as a fertilizer. iv) Chemical treatment The procedure proceeds with chemical treatment in which iron chloride (FeCl 3 ) is added as a coagulant. The treated wastewater is passed through a filter and consequently discharged out to 24

35 the Dalby rivulet. Schematic plan of the treatment processes in Dalby WWTP is illustrated in Figure 9 (Dalby Miljörapport, 2007). 1: Sand trap 2: Screen 3: Biological nitrogen removal (pre-denitrification) 4: Sedimentation tank 1 5: Chemical treatment 6: Sedimentation 2 7: Filter 8: Return sludge 9: Excess sludge 10: Centrifuge Figure 9: Schematic plan of treatment processes in Dalby WWTP (present approach) (Dalby Miljörapport, 2007). Measurements from 2007 indicate high BOD and nitrogen removal efficiency in the Dalby WWTP. Average BOD and nitrogen in the outlet, for the year, was measured to be 2 and 10 mg/l respectively. Total sludge production in this WWTP was reported to 815 ton/year in 2007 (Dalby Miljörapport, 2007) Future approach The structure of treatment processes in Dalby WWTP will be technically changed to a partial treatment method. The biological nitrogen removal will be changed to the post-denitrification method. The nitrification will be carried out in Dalby WWTP while denitrification is planned to 25

36 occur in a pipe conveying the nitrified wastewater under anoxic condition to sewer network in Lund. Different sections of the treatment processes in Dalby WWTP in the future are presented below. i) Pre-treatment No significant changes will be implemented in the pre-treatment steps. Sand trap and screen are two initial treatment processes in this WWTP respectively. There are some problems with the screen which will be changed if it cannot work properly. ii) Biological treatment Only the nitrification stage of the post-denitrification process will be carried out in the WWTP. Total capacity of the two nitrification tanks and one former denitrification tank which were used in the previous treatment scenario, 900 m 3 in total (all three tanks), will be applied as three nitrification tanks aerated by bottom aerators to 2 mg O 2 /l. In high flow cases, two extra tanks which were formerly used as settling tanks can increase the nitrification process capacity to 1780 m 3. Chemical treatment will be completely removed from the treatment processes and the tank will be applied as a settling tank. The sludge from the settling tank will be returned back to the nitrification process and the excess sludge will be conveyed to the pipe to be used in the denitrification process. iv) Sewer pipe design: The underground sewer pipeline, which is supposed to transport nitrified wastewater in under anoxic condition, has been designed to be built from Dalby WWTP to Lund. It joins the sewer networks, right before Gastelyckan in the south-east of Lund, and eventually is transported to Källby WWTP. The design phase of the sewer pipeline project has been conducted at Tyréns AB and the pipe consists of two distinct sections with different dimensions and material. The first part of the pipeline is made of polyethylene, since wastewater is pumped into the pipe and plastic pipe is appropriate for flows under pressure. The other section is a gravity sewer in concrete which will be connected to the sewer network in Lund. These two pipes are 0.45 m and 0.6 m in diameter and are 7.36 km and 0.87 km long respectively which will be coupled together by an ordinary manhole located in a monitoring station. The minimum slope in the gravity pipe is 0.5% and it provides the appropriate velocity along the pipeline to prevent sedimentation in the pipe (Publikation P90, 2004). The map and schematic profile of the sewer pipeline in Dalby-Lund project are illustrated in Figures 10 and

37 Figure 10: Map of the sewer pipeline in Dalby-Lund project (Tyréns AB archive, 2008). Figure 11: The schematic profile of the sewer pipeline. 27

38 iii) Pump station and detention tank Although the sewage network in Dalby was separately designed for collecting wastewater and storm water, there is still a risk of overflow in Dalby WWTP in wet seasons. The overflow occurs because there are still some connections between these two sewer systems and also storm water can infiltrate into the wastewater sewers in some cases. Pumping wastewater to the sewer pipeline is planned with the pumping rates 50 and 120 l/s in normal and high flow cases respectively. The possible extreme high discharge of the wastewater to the Dalby WWTP is estimated to 200 l/s. This overflow resulted in applying a method where a detention tank is used to store the excess wastewater before it is conveyed to the pumping station. Based on calculations that have been done by Tyréns AB, a 6000 m 3 detention tank with 50 and 120 l/s pumping rate in normal and high flow situations, can handle the influent wastewater to the WWTP. 28

39 5. Biological transformations in the bulk water 5.1. EFOR software The Dalby-Lund project is not executed, so the biological treatment is still carried out based on the pre-denitrification method. In order to evaluate the treatment efficiency of the planned scheme in the WWTP and sewer pipe, computer modeling has been applied to simulate the case of study in the future and analyze the feasible situations. EFOR is the computer model which is used in this study. It is a computer tool, by the DHI Water and Environment group, with the possibility to model, analyze and predict the different sections of a conventional WWTP. It is based on reliable mathematical models such as ASM-1 and ASM- 3 for biological processes in the activated sludge method (EFOR, 2003). The possibility to model almost all features of a WWTP such as aerobic, anoxic and anaerobic basins, intermediate and final settling basins provides a realistic configuration for operational optimization and predicting feasible drawbacks. In general, EFOR is a user-friendly and easy to apply computer program which is very flexible with limited or missing data. The missing data can be estimated by mass balance methods or default data which are roughly fitted to the model (EFOR, 2003). EFOR as a software is only capable of simulating the biological treatment processes occurring in the bulk water phase of the sewer pipe. Hence, all the results achieved by simulations are evaluated to predict the denitrification rate in the bulk water Modeling methodology Model description The components of Dalby WWTP and the sewer pipe which have been modeled in EFOR are presented in Figure 12. All of the parameters have been evaluated in normal cases to make an overview of the whole process. Figure 12: Model of components in the Dalby WWTP and sewer pipe in EFOR. 29

40 The physical structure of the project has been modeled in the design section in EFOR. There one can make a schematic plan with combinations of different components such as activated sludge tanks, settling tanks, etc and link them together by pipes. In the next step, all the treatment processes have to be assigned by a control loop defining their specific properties. In each control loop sensors, controllers, control devices and setpoints should be defined and linked together (EFOR, 2003). All the simulations are evaluated at a steady state situation. The wastewater characteristics e.g. flow, BOD, COD and nitrogen and phosphorous concentrations presented in Table 3 are inserted in to the model. Temperature is fixed to the constant value of 20 C to evaluate the maximum potential for the denitrification process. The mechanical pre-treatments, sandtrap and screen, have not been considered in the simulations because of their low influence in the treatment processes. Inlet, three nitrification tanks, sedimentation tank, sewer as anoxic chambers, outlet and the dosing unit are defined as the treating components in this model Model parameters In the near future persons will be connected to the WWTP and the average inflow to Dalby WWTP is estimated to be 80 l/s. The average concentration of BOD, nitrogen and phosphorous are considered the same as those in the present approach. These concentrations, presented in Table 3, are inserted in this model as initial concentrations in inflow wastewater to Dalby WWTP Model components description Here, different components of this project and their assigned characteristics in the control loops are presented. All characteristics of the features in the treatment system which are used in the computer model, both in the WWTP and the sewer pipe, should be defined. Activated sludge tanks consist of 3 aerobic tanks (AS1, AS2, AS3) in the WWTP and two anoxic tanks (AS4 and AS5) representing the pressure and gravity sewer pipe. The settling tank is shown as SS1 in Figure 12. The capacities of all tanks and aeration rates in nitrification tanks are presented in Table 4. Table 4: Properties of activated sludge tanks in the model. Activated sludge tank Volume (m 3 ) Aeration rate (kg O 2 /h) AS AS AS AS AS

41 The valve is the controlling device to regulate the bypassed raw wastewater to the sewer pipe. It can be set to bypass different amounts of wastewater. For instance, 0% open means that no wastewater is bypassed to the sewer pipe and 100% of the wastewater goes directly to nitrification tanks. Three pumps are used in the model for controlling the return sludge, excess sludge and dosing unit named pump 1, 2 and 3 respectively. The capacity of the pumps used in this model is presented in Table 5. Table 5: Properties of pumps used in the model. Control device Capacity (m 3 /h) Connecting components Pump SS1 to AS1 Pump 2 20 SS1 to AS4 Pump 3 0.2/0.4 Dosing 1 to AS C/N ratio in the sewer pipe: In the post-denitrification process, the internal carbon source required for the denitrification process can be limiting because organic matters in the wastewater are utilized by bacteria in the aerobic basin. In order to evaluate whether the internal carbon source is sufficient for the denitrification process, C/N ratio is calculated. By using Expression 9, the (C/N) practice is calculated as the sufficient ratio in the denitrification process by considering 3 kg COD/kg N and 0.3 for (C/N) optimum and f C/N respectively (Appendixes 1 and 2). The initial concentrations of soluble COD (SCOD), of which soluble inert COD has been subtracted, and nitrate in the sewer pipe have been estimated to 21 g COD/m 3 and 17.6 g NO 3 /m 3 in EFOR. Therefore the required COD concentration, in the sewer pipe, to remove initial nitrate under anoxic conditions can be calculated. The result shows that the available internal carbon source (21 g COD/m 3 ) is not sufficient for the denitrification process and an external carbon source should be injected to the sewer pipe. The nitrate which can be removed by using the available internal carbon source is calculated. Methanol is used as the external carbon source to remove the remained. (C/N) optimum and f C/N in the post-denitrification process, with an external carbon source (methanol), are considered 2.5 kg MeOH/kg N and 0.8 respectively (Appendixes 1 and 2). 31

42 The results indicate that 48.5 g MeOH/m 3 (0.004 m 3 MeOH/h) is required to be added to the sewer pipe to efficiently remove nitrate under an anoxic condition Model components characteristics All the characteristics of the control loops assigned for the nitrification tanks, return sludge, excess sludge, bypassed raw wastewater and dosing unit are presented in Table 6. Table 6: Characteristics of control loops used in the model. Control loop Sensor Controller Control device Setpoint AS1 Oxygen in AS1 On/off (Inverse) Diffuser (AS1) 2 (g O 2 /m 3 ) AS2 Oxygen in AS2 On/off (Inverse) Diffuser (AS2) 2 (g O 2 /m 3 ) AS3 Oxygen in AS3 On/off (Inverse) Diffuser (AS3) 2 (g O 2 /m 3 ) Return sludge Flow in inlet Percentual (Direct) Pump 1 100% Excess sludge Suspended solid On/off (Direct) Pump (g SS/m 3 ) Bypassed wastewater Flow in inlet Percentual (Direct) Valve 1 10% or 20% External carbon source Timer Timer Pump m 3 /h The on/off controller set in inverse mode for the nitrification tanks regulates the oxygen concentration in the system. Inverse mode is used to decrease the aeration when the oxygen concentration is higher than the setpoint, set to 2 g O 2 /m 3 and vice versa. The dosing unit constantly injects methanol, as an external carbon source, to the sewer pipe. Pump 3 is set to inject methanol, and m 3 /h, to the sewer pipe. Valve 1 regulates the proportion of wastewater conveyed to nitrification basins and to the sewer pipe. It is set to bypass 10% and 20% raw wastewater to the sewer pipe. This means that only 90% and 80% of the total inflow will enter the nitrification tanks. Return sludge is pumped from the settling tank (SS1) to the first nitrification tank using pump 1. Pump 2 is in charge of transferring excess sludge from the settling tank (SS1) to the sewer pipe. The setpoint is set to 4500 g/m 3 suspended solids in the first nitrification tank. A direct on/off controller changes the pumping rate until the concentration of suspended solids, in the settling tank, reaches the setpoint. 32

43 The pressure and gravity sewer pipes are modeled as anoxic tanks without aeration and no control loops need to be assigned Different alternatives for optimizing the denitrification The goal of this project is to transfer wastewater from Dalby WWTP to Källby WWTP in Lund under sewer conditions and take advantage of in-sewer microbial transformations under anoxic conditions. In order to optimize the denitrification process in the sewer system, different approaches are modeled to achieve a realistic and economically beneficial scenario. Ideas for optimizing the denitrification process in the sewer pipe are presented in different scenarios to evaluate the nitrate reduction efficiency in the outlet. Scenario 1 The simplest approach is to model the sewer as two anoxic tanks without aeration, for pressure and gravity sewers respectively. In this case, the capacities of two anoxic tanks are considered as the same as the capacity of the pressure and gravity pipes. Since the gravity sewer is half full, half of the capacity of this section of the pipe is used. Therefore the volumes of the anoxic tanks are 927 and 123 m 3. Scenario 2 The anoxic tanks, representing the pressure and gravity sewers, are not modeled as a unified model. These two sections are equally divided to more chambers with less capacity. Using a model with multiple tanks instead of one tank, especially for tanks with high capacities, can model the sewer pipe closer to the plug flow reactor (See section 6.2) and result in a more realistic model for the denitrification process. In this case the pressure pipe, which has more capacity than the gravity pipe, is divided into four anoxic tanks and the gravity pipe is divided into two anoxic tanks. This means that there are six anoxic tanks interacting in the system. The only change in this scenario compared to scenario 1 is that the capacities of the anoxic tanks have been changed. The first four chambers are 231 m 3 and the two subsequent chambers are 62 m 3. Scenario 3 In this scenario two anoxic tanks are applied to model the sewer pipe. A specific proportion of the inflow, raw wastewater, is bypassed directly to the sewer system as an internal carbon source. This amount is controlled by a controller device, valve, and can easily be modified. In this scenario 10% and 20% of the raw wastewater is bypassed to the sewer pipe to evaluate the denitrification rate under anoxic conditions. Scenario 4 A dosing unit is defined to the system which provides the possibility of adding external carbon, methanol, to the sewer system modeled as two anoxic tanks. It has been linked by a pump to the anoxic tanks which can regulate the concentration of the external carbon. Since the methanol concentration setpoint cannot be set to less than 0.01 m 3 /h in the EFOR program, methanol dilution (to 1% of its original concentration) is used in the model. Setpoint is 0.4 m 3 /h, which is 33

44 the maximum concentration of methanol that can be applied in the sewer pipe. Different concentrations of methanol, and m 3 /h, are injected to the sewer pipe to investigate the in-sewer nitrate removal efficiency. Due to applying methanol dilution, the concentrations of injected methanol inserted to the model are 0.2 and 0.4 m 3 /h. Remember that the methanol concentration could not be set to less than 0.01 m 3 /h, therefore and m 3 /h of 100% methanol are considered 0.2 and 0.4 m 3 /h with 1% methanol in the model. Combination of these scenarios may bring forward new ideas that can effectively reduce nitrate concentrations in the outlet and enhance denitrification processes in the sewer pipe. Two main schemes of combination of these scenarios are: Scenario 5 Applying raw wastewater and external carbon, methanol, to the first scenario evaluates the denitrification rate under in-sewer anoxic conditions. In this scenario, 10% raw wastewater is bypassed into the sewer pipe, modeled as two anoxic tanks, while m 3 /h methanol is injected to the system. Scenario 6 In this case, raw wastewater and methanol is added to the second scenario, modeling the pressure and gravity pipes as 4 and 2 anoxic tanks respectively, to assess the nitrate removal efficiency in the outlet. The amount of bypassed wastewater and injected methanol is the same as in the fifth scenario, 10% wastewater and m 3 /h respectively Results Nitrate concentration in the outlet The nitrate removal efficiency is studied by comparing the initial concentration of nitrate in the sewer pipe with nitrate concentrations in the outlet in different scenarios. The initial nitrate concentration in the sewer pipe and nitrate concentrations in the outlet in different scenarios are presented in Figure 13. The model of the sewer pipe as two anoxic tanks with defined capacity, scenario 1, does not show significant nitrate removal in the sewer pipe. As can be seen in Figure 13, the concentration of nitrate after the nitrification process and the concentration in the outlet are estimated to 17.6 and 16 g NO 3 /m 3 respectively. The nitrate removal efficiency is only 9% in scenario 1. The idea of using multichamber anoxic tanks, scenario 2, for modeling the sewer pipe does not significantly enhance the denitrification process due to organic matter deficit in the sewer pipe (see Figure 13). The nitrate concentration in the outlet in scenario 2, 15.9 g NO 3 /m 3, shows a slight difference compared to that of scenario 1. This indicates that the idea of using more chambers with less capacity for modeling anoxic chambers in scenario 2 can model the sewer pipe closer to a plug flow model and reality. 34

45 Figure 13: Initial nitrate concentration in the sewer pipe and nitrate concentrations in the outlet in different scenarios. The idea of bypassing raw wastewater to the sewer pipe to provide enough organic matter for the denitification process is evaluated in scenario 3. The model has been simulated with 10% and 20% bypassed raw wastewater into the sewer pipe. As can be seen in Figure 13, the nitrate concentration in the outlet decreases to almost 12.8 and 10 g NO 3 /m 3 by bypassing 10% and 20% respectively. Since 10% and 20% of the total discharge to the sewer pipe comes from raw wastewater which contains a very low nitrate concentration, the initial nitrate concentration in the sewer pipe decreases. The results presented in Figure 13 illustrate the nitrate concentration in the outlet by applying methanol as the external carbon source. The nitrate concentration decreases slightly to 15 and 13.8 g NO 3 /m 3 by applying and m 3 /h methanol. Results indicate that injecting required methanol, m 3 /h, for removal of the initial nitrate concentration in the sewer pipe can only remove the nitrate with 22% efficiency. The nitrate removal efficiency in the sewer pipe is assessed by combining scenarios to evaluate the effects of applying raw wastewater and external carbon sources simultaneously. In scenario 5, 10% raw wastewater and m 3 /h methanol are added to the sewer pipe model in the first scenario. The result indicates that the nitrate concentration decreases to 11.8 g NO 3 /m 3 in the outlet. In the scenario 6, the same situation is conducted for the second scenario, 10% raw wastewater and m 3 /h methanol. The nitrate reduction shows a slight difference compared to scenario 5 and decreases to 10.5 g NO 3 /m 3 in the outlet. Scenarios in which bypassed raw wastewater is added to the sewer pipe (Scenarios 3, 5 and 6), the nitrate concentration in the outlet decreases more efficiently than that of in other scenarios. This can be explained by the lower initial nitrate concentration in the sewer pipe because only 80% or 90% of the raw wastewater is nitrified in the nitrification process in these scenarios. 35

46 COD concentration in the outlet The COD concentration in the outlet is investigated in different scenarios to evaluate the contribution of organic matters in the denitrification process in the sewer pipe. Because the soluble COD (SCOD), of which soluble inert COD has been subtracted, can take part in redox reactions, the SCOD is used to compare different scenarios. Initial SCOD concentration in the sewer pipe is modeled to 21 g SCOD/m 3. Scenario 1 and 2 show a slight decrease in COD concentrations in the outlet, 20.5 g COD/m 3, compared to the initial COD concentration in the sewer pipe. The addition of 10% and 20% raw wastewater in the scenario 3 increases the COD concentration to 25.5 and 30.5 g SCOD/m 3 in the outlet. Since raw wastewater contains high COD concentrations, the SCOD concentration in the outlet increases significantly. The SCOD concentration in the outlet in scenarios 1, 2 and 3 can be found in Figure 14. Figure 14: Initial SCOD concentration in the sewer pipe and SCOD concentrations in the outlet in different scenarios. The concentration of SCOD in the outlet increases by injecting methanol as an external carbon source in scenario 4. The additions of and m 3 /h methanol to the sewer pipe increase the SCOD concentration in the outlet to 21.5 and 22.2 g SCOD/m 3 (See Figure 14). In both scenarios 5 and 6, in which 10% raw wastewater and m 3 /h methanol has been added to the sewer pipe, the SCOD concentrations increase compared to scenario 4. The SCOD concentrations in scenarios 5 and 6 increase to 26.2 and 26.4 g COD/m 3 respectively. As can be seen in Figure 14, although bypassed wastewater and an external carbon source, methanol, has been added to the sewer pipe to provide enough organic matter for the 36

47 denitrification process, SCOD concentrations in the outlet are increasing in all scenarios without efficient nitrate removal Ammonium concentration in the outlet The concentration of ammonium in the outlet is assessed so that the removal efficiency of nitrogen compounds in the denitrification process can be evaluated. The initial ammonium concentration in the sewer pipe is relatively low, 1.9 g NH 4 -N/m 3. The ammonium concentration in the outlet in scenarios 1 and 2 are modeled to 1.9 and 1.88 g NH 4 -N/m 3 respectively. The results are illustrated in Figure 15. Figure 15: Initial ammonium concentration in the sewer pipe and ammonium concentrations in the outlet in different scenarios. Bypassing raw wastewater to the sewer pipe, scenario 3, significantly increases the ammonium concentrations in the outlet due to a high ammonium concentration in the raw wastewater. Accordingly, bypassing 10% and 20% raw wastewater results in 3.1 and 4.8 g NH 4 -N/m 3 in the outlet. Addition of methanol, which is investigated in the scenario 4, shows a slight decrease in ammonium concentrations compared to the initial ammonium concentration in the sewer pipe. In fact, the ammonium concentrations are reported to 1.5 and 1.1 g NH 4 -N/m 3 by injecting and m 3 /h methanol to the sewer pipe (See Figure 15). Scenarios 5 and 6 result in lower ammonium concentrations compared to scenario 3. In these scenarios, ammonium concentrations are modeled to 2.7 and 2.3 g NH 4 -N/m 3 in the outlet. 37

48 38

49 6. Biological transformations in the biofilm 6.1. Denitrification kinetics in the biofilm Denitrification occurs when dissolved oxygen is depleted and nitrate is the dominant electron acceptor in the sewer system. Easily-degradable organic matters can penetrate into the biofilm and reduce nitrates by denitrification reactions. The penetration rates of the nitrate and organic substances into the biofilm depend on their molecular diffusion coefficients and the stoichiometric coefficient between these two substances. In cases which nitrate and organic matter substrates diffuse into the bottom of the biofilm, the nitrate removal occurs according to the zero order reactions with 100 percent efficiency. In reality, the biofilm is partially penetrated and the nitrate removal is based on half order reactions unless the concentration of substrates is extremely high. In this study, the biofilm is assumed to be partially penetrated so that the nitrate removal rate is estimated according to the half order reactions. The list of kinetic constants in a denitrifying biofilm reactor is presented in Table 7 (Henze, et al., 1996). Table 7: Kinetic constants in a denitrifying biofilm reactor (20 C) (Henze, et al., 1996). K 1/2,NO3 D ν NO3 D NO3 ν NO3 /D COD K 0f,NO3 Substance (gno 1/2 3 m -1/2 d -1 ) (10-4 m 2 /d) (kgcod/kgno 3 ) (kgcod/kgno 3 ) (kgno 3 /m 3 d) Acetic acid Methanol Glucose Unspec. COD Methodology In an ideally mixed reactor, the concentrations of substances are considered constant along the reactor and in the effluent. In this project, as concentrations of substrates change through the sewer pipe due to denitrification reactions in the biofilm, a plug flow reactor (PFR) can model bio-chemical reactions in continues flowing systems much closer to reality. The bio-chemical reactor is represented as a series of infinitely thin coherent plugs transferred along the axial direction of the chemical reactor. Each plug consists of a certain substrate concentration less than the plugs before and more than those after. Each plug of the differential volume, considered as an infinitesimally small batch reactor, is assumed to be perfectly mixed in the radial direction but not in the axial direction in a PFR model (Michigan University, 2000). The plug flow reactor model is applied to describe bio-chemical transformations conveyed in the sewer pipe. In the ideal PFR model residence time, θ, is constant and measured as the average time that a discrete plug travels in the pipe. In reality, the turbulent flow and axial diffusion may cause mixing in the axial direction in the reactor and result in errors in the PFR model. Therefore, some assumptions which are listed below are required to simplify the problems. 39

50 Assumptions in the plug flow reactor: - Plug flow - Steady state - Constant density - Isothermal conditions (Constant temperature) The schematic diagram and the mass balance of a plug flow reactor are presented in Figure 16 (Henze, et al., 1996). Figure 16: Left: Schematic diagram of a plug flow reactor (PFR). Right: Mass balance for a biofilm reactor (Henze, et al., 1996). The mass balance of substrates in an infinitesimal plug of the fluid, axial length dx, can be found in Expression 19. In Removed = Out : Flow through the reactor : Concentration of substrate in the differential plug of the fluid (dv) : Cross section area of the reactor : Volumetric removal rate of the reactor As a matter of fact, the mass balance can be stated as Expressions 20 and 21 when dx approaches zero. Or dimensionless: (19) (20),,, 40

51 (21) : Residence time : Initial concentration of substrate : Volume of the reactor : Length of the reactor The reaction rate in the partially efficient biofilm following the half order reaction is calculated by Expression 22. : The specific biofilm area : Removal rate per unit area : Half order area specific rate constant Concentrations of substrates decrease along the biofilm reactor while redox reactions occur in the biofilm layer. By combining Expressions 20 and 22, one can calculate the concentrations of substrates in the outlet by using Expression 23. : Concentration of substrate in the outlet (22) (23) In the denitrification process in the biofilm phase of the sewer pipes, organic matter deficiency is commonly the limiting factor reducing nitrate removal efficiency under anoxic conditions. There are three prerequisites to ensure that the denitrification process occurs along the sewer pipe. First of all, the ratio of available concentration of easily-degradable organic matters to the nitrate concentration in the wastewater should be evaluated (Expression 9) to ensure whether enough organic matters are present in the sewer pipe. Moreover, organic matters have to penetrate into the biofilm as far as nitrate does to reduce nitrate concentrations in the sewer pipe. Otherwise the organic matters will be limiting even if the required C/N ratio is available in the sewer pipe. Even though these two prerequisites are achieved with the initial organic matter concentration, the denitrification process in the biofilm consumes organic matters along the sewer pipe and may result in organic matter deficiency in the sewer pipe. The concentration of substrates in each cross section of the sewer pipe depends on the residence time that the wastewater travels in the sewer pipe. Residence time in each cross section is a factor of the length of the sewer pipe in that specific section. Providing that organic matters can penetrate into the biofilm as far as nitrate does in the outlet, enough organic matters are available in the sewer pipe which can reduce the nitrate concentration without limiting the denitrification process in the entire sewer pipe. 41

52 6.3. Results Physical properties of the sewer pipe Physical properties of two sewer pipes with different diameters and lengths have been calculated to estimate the denitrification rate in the biofilm as a bio-reactor. The gravity sewer pipe is considered half full in these calculations. A list containing the physical properties of both sewer pipes is presented in Table 8. Table 8: Physical properties of plastic and concrete sewer pipes. Parameter Plastic pipe Concrete pipe Unit Diameter m Length m Discharge l/s Volume m 3 Biofilm area m 2 Cross section area m 2 Residence time h Kinetic constants in a denitrifying biofilm reactor In order to calculate the denitrification efficiency in the biofilm phase of the sewer pipe as a bioreactor, kinetic constants in the denitrifying biofilm should be calculated. Diffusion coefficients for nitrate and COD, D NO3 and D COD, are considered 0.5*10-4 m 2 /d (Appendix 5). The stoichiometric coefficient between nitrate and organic matter in the denitrification reactions, ν NO3,COD, represents the degree that two chemical compounds, electron acceptor and donor, participate in the reaction (Henze, et al., 1996). The stoichiometric coefficient for the denitrification process between COD and nitrate is calculated by assuming C 18 H 19 O 9 N as the average composition of the organic matter in the wastewater. The observed yield constant in a low-loaded nitrification-denitrification plant is 0.4 kg biomass/kg organicmatter. As the result, the reaction between the organic matter, C 18 H 19 O 9 N, and the biomass, C 5 H 7 NO 2, is presented in Expression 25. C 18 H 19 O 9 N O 2 18 CO H 2 O + H + + NO 3 - (24) 1 mole organic matter 19.5 mole O 2 C 18 H 19 O 9 N+8.4 NO H NH C 5 H 7 NO N CO H 2 O (25) COD of organic matter = 19.5 mol COD/ mole organic matter 1 mole organic matter*19.5 mol COD/ mole organic matter = 19.5 mole COD converted by means of 8.4 mole NO 3 -. Thereby, the stoichiometric coefficient, ν NO3,COD, is: 42

53 The half order area specific rate constant for NO 3, k 1/2,NO3, is calculated while the zero order rate constant for COD is estimated to 250*10 3 kg COD/m 3 d (Appendix 5). (25) Mass transfer limitation The rate of biological transformations in the biofilm depends on how far the electron acceptor and donor can penetrate into the biofilm layer. In a denitrifying biofilm reactor, nitrate and organic matter, as oxidant and reductant substrates, can take part in denitrification reactions as long as both substrates are present in the biofilm. Hence, the shortest distance that substrates penetrate into the biofilm is the criterion to determine which substrate is limiting. The initial concentrations of nitrate and easily-degradable COD in the sewer pipe have been estimated to 17.6 g NO 3 /m 3 and 21 g COD/m 3 in the computer model. By substituting available concentrations in Expression 16, one can estimate whether nitrate or COD is the limiting substance in the denitrification process. Therefore, COD is the limiting factor in the denitrification process occurring in the biofilm. The required organic matter, as the internal carbon source for the complete denitrification in the sewer pipe, can be estimated according to Expression 9. Values for (C/N) optimum and f C/N are considered 3 kg COD/kg N and 0.3 respectively (Appendixes 1 and 2). 43

54 In order to achieve complete denitrification, the organic matter should penetrate as far as nitrate does. The estimated COD concentration in the sewer pipe is not the limiting factor in the denitrification process anymore. The organic matter can penetrate deep enough into the biofilm as it is calculated below Nitrate removal efficiency in the biofilm The denitrification process can occur in the biofilm reactor while both the organic matter and nitrate are present. Since the COD is the limiting factor in the removal process, a certain amount of initial nitrate will take part in the denitrification reactions and will be removed. Here, the concentrations of COD and nitrate in the outlet are calculated according to Expression 23. The results show that g NO 3 /m 3 will remain in the outlet and the denitrification process in the biofilm can only remove 3.44 g NO 3 /m 3 which is 20% of the initial nitrate concentration Addition of an external carbon source Providing that enough organic matters are available in the sewer pipe, both nitrate and organic matters are sufficient for the denitrification process in the denitrifying biofilm reactor. Therefore, methanol as the external carbon source has been used to evaluate the nitrate removal efficiency in the biofilm. By calculating (C/N) practice in Expression 9, the required methanol for removing the remained nitrate concentration in the sewer pipe can be estimated. (C/N) optimum and f C/N are considered 2.5 kg MeOH/kg N and 0.9 respectively (Appendixes 1 and 2). 44

55 Therefore g MeOH/m 3 is required to efficiently remove nitrate in the sewer pipe. According to Table 7, diffusion coefficients for nitrate and methanol are considered 0.6*10-4 and 1*10-4 m 2 /d and the zero order rate constant of methanol in the biofilm, k 0vf,Me, is 80*10 3 g MeOH/m 3 d. Moreover, the stoichiometric coefficient between methanol and nitrate in the denitrification process, ν NO3,Me, is suggested 3.5 g MeOH/g NO 3 -N (Henze, et al., 1996). The half order rate constant of methanol is calculated to: The penetration rate of methanol and nitrate into the biofilm should be evaluated to ensure that the methanol concentration is not the limiting factor for the nitrate removal (Expression 16). Theoretically, complete denitrification is expected to occur in the biofilm by injecting g MeOH/m 3 to the sewer pipe. Here, the nitrate and methanol concentrations in the outlet are calculated. The results indicate that there are still considerable concentrations of nitrate and methanol in the outlet. By applying methanol as an external carbon source, 6.81 g NO 3 /m 3 which is almost 48% of the remained nitrate in the biofilm can be removed. Applying the required external carbon source for the denitrification process in the biofilm phase of the sewer pipe, g MeOH/m 3, indicates that the initial methanol concentration can penetrate into the biofilm without being limiting. Along the sewer pipe, methanol concentrations should be high enough to not limit the denitrification process in the biofilm. Therefore, if the 45

56 methanol concentration in the outlet is still high enough to penetrate into the biofilm as far as nitrate does (Expression 16), the methanol concentration is not limiting along the sewer pipe. The result shows that the methanol concentration is sufficient for the denitrification process in the biofilm phase along the sewer pipe even though nitrate reduction efficiency is only 48% in the outlet. 46

57 7. Discussion 7.1. Denitrification in the bulk water The bulk water phase of the sewer pipe as an anoxic unit is modeled in EFOR computer program to evaluate the in-sewer microbiological transformations and the denitrification efficiency in the sewer pipe. Moreover, different scenarios which provide higher concentrations of organic matter are simulated to assess whether using these scenarios can enhance the denitrification rate in the sewer pipe. As it is expected, the results do not show significant nitrate removal efficiency in the outlet in scenarios 1 and 2 because the available SCOD concentration in the sewer pipe is quite low. In fact, the nitrate utilization rate in the sewer pipe is estimated to 0.27 g NO 3 /m 3 h. Bypassing raw wastewater to provide the sufficient internal carbon source for the denitrification process is limited to 10 and 20 percent of the total discharge to the sewer pipe. Bypassing higher percentage of the raw wastewater may severely decrease the wastewater quality in the outlet. Despite the nitrate reduction in the outlet in scenario 3, the total concentration of nitrogen compounds does not show significant reduction because of a high ammonium concentration in the raw wastewater. As can be seen in the results (sections and 5.4.3), the total nitrate and ammonium concentrations in the outlet is 15.9 and 14.8 g N/m 3 in scenario 3. Therefore, the nitrogen removal efficiency is only 9% and 16% when 10% and 20 % raw wastewater are bypassed to the sewer pipe respectively. According to calculations estimating the required external carbon source for complete denitrification, and m 3 /h methanol is injected to the sewer pipe in scenario 4. Despite the fact that injected methanol, at m 3 /h, to the sewer pipe is expected to completely remove nitrate in the outlet, the results only show 22% nitrate reduction efficiency. The denitrification rate is estimated to 0.67 g NO 3 /m 3 h in the sewer pipe. In scenarios 5 and 6, the total nitrate and ammonium concentrations in the outlet, 14.5 and 12.8 g N/m 3, show how same situations model a closer scenario to the plug flow model and eventually reality in scenario 6. In the study that has been done by Abdul-Talib, et al., (2002) (See section 3.2.2), the denitrification rate in the bulk water phase of sewer networks was estimated to g NO 3 /m 3 h, when enough substrates were available. In Dalby-Lund sewer pipe, the results show a rather lower denitrification rate in the bulk water phase even if sufficient substrates are available. In the activated sludge process, living microorganisms degrade organic matter for their cellular growth. Therefore, the success of this process depends upon providing high concentrations of bacteria by increasing the sludge concentration in the process. In this way recycling sludge to the activated sludge tank is a solution to use both rapid- and slow-growing microorganisms in substrates degradation. In the Dalby-Lund project, sludge discharged to the sewer pipe leaves the system in the outlet without being recycled; thereby, the biomass concentration in the sewer pipe is very low. Moreover, since the residence time in the sewer pipe is short, 6 hours, only rapid- 47

58 growing microorganisms can participate in the denitrification process and remove nitrate. These are the reasons that even though enough external carbon source for the complete denitrification process has been added to the sewer pipe, the denitrification rate is relatively low and there is still a high nitrate concentration in the outlet Denitrification in the biofilm Since the internal organic matter concentration is low in the sewer pipe, there is a competition between bulk water and biofilm phases of the sewer pipe to degrade present organic matters for the denitrification process. In this case study, the total concentration of the initial organic matter in the sewer pipe is assumed to be available for the denitrification process in both bulk water and biofilm phases of the sewer pipe. Redox-zone, where redox reactions occur, is determined by the degree of penetration of the nitrate and organic matter into the biofilm. As the organic matter cannot penetrate into the biofilm as far as nitrate does, it will be the limiting substrate in the denitrification process. The results show that only 3.44 g NO 3 /m 3, that is 20% of the initial nitrate concentration, can take part in denitrification reactions in the biofilm. The high concentration of nitrate in the outlet, g NO 3 /m 3, indicates that the denitrification rate in the biofilm phase of the sewer pipe is relatively low, g NO 3 /m 2 h. Therefore, biofilm as the reactor cannot efficiently participate to reduce the nitrate concentration in the sewer pipe. The idea of injecting the external carbon source, methanol, into the wastewater has come up to compensate the organic matter deficiency and consequently increase organic matter penetration into the biofilm. The required methanol to remove remained nitrate in the sewer pipe, g NO 3 /m 3, is estimated to g MeOH/m 3 which can penetrate into the biofilm without limiting denitrification reactions. Results show that the nitrate concentration decreases to 7.35 g NO 3 /m 3 in the outlet. That is, only 48% of the nitrate will be removed even if the required organic matter is added to the sewer pipe. The result shows a low denitrification rate, 0.13 g NO 3 /m 2 h, in the biofilm phase of the sewer pipe. Diffusion is the predominant transport of substrates from bulk water into the biofilm. The diffusion-kinetic rate of different substrates is a slow process which may limit the penetration of substrates into the biofilm. For instance, the diffusion coefficients for nitrate and COD, which are considered 0.7*10-4 and 0.4*10-4 m 2 /d respectively, are relatively low. Thereby, even if a high biomass concentration is available in the bulk water phase of the sewer pipe, the slow molecular diffusion process limits the substrates penetration rate and eventually the nitrate removal (Henze, et al., 1996). In the Dalby-Lund sewer pipe, the low biomass concentration results in an even worse scenario for the diffusion process into the biofilm. In the case which the required methanol for the denitrification process has been injected to the sewer pipe, high nitrate and methanol concentrations are still present in the outlet. Generally, microorganisms are gradually growing on the sewer wall during the dry weather and make an active and thick biofilm layer. However, the biofilm layer can be eroded and washed out from the sewer pipes during wet periods. Although the biofilm formation on the sewer wall will be started right after the rain event, its formation is rather slow (Ahnert, et al., 2005). Due to long 48

59 wet periods in south Sweden, the denitrification rate in the biofilm layer of Dalby-Lund sewer pipe may be limited because of biofilm erosion during rain events Application of an external carbon source The initial easy-degradable organic matter concentration in the sewer pipe is rather low and is not sufficient for the denitrification process. According to section in the literature study, in the pre-denitrification process, internal organic matters are consumed in the nitrification process so the external carbon source is commonly added to the denitrification process. In this project, the effects of injecting the external carbon source, methanol, to the sewer pipe has been studied to evaluate the nitrate removal efficiency in the sewer pipe. Different scenarios with addition of methanol to the sewer pipe are modeled to evaluate the denitrification process in the bulk water phase. The results show only 22% nitrate removal efficiency even if required methanol is injected to the system. Moreover, the results of methanol addition to the sewer pipe do not significantly decrease the nitrate concentration in the sewer wall biofilm. Methanol is costly to use at a high extent for denitrification processes. Using methanol in this project will not effectively enhance the denitrification process in the sewer pipe. Therefore using methanol is not worth the increased financial expense Sulfide formation risk As stated before (section 3.2.3), once dissolved oxygen and nitrate are depleted in the sewer network, under anaerobic conditions, sulphate will take part as an electron acceptor in redox reactions. The produced hydrogen sulfide in these reactions can cause odor nuisance and sewer wall corrosion. Since the denitrification rate in the sewer pipe is rather low and there are still significant concentrations of nitrate in the outlet, the entire pipe is considered to be under an anoxic condition. As the anaerobic condition seems unlikely to occur in the sewer pipe, there is almost no risk of hydrogen sulfide formation and its deteriorative effects in the entire sewer pipe Effects of the Dalby-Lund project on Källby WWTP In the Dalby-Lund sewer pipe project, the partially treated wastewater is conveyed from Dalby WWTP, through the sewer pipe, to the sewer network in Lund. Consequently it will be mixed with a high volume of raw wastewater in Lund s sewer network before being discharged to Källby WWTP. This project will undoubtedly affect both technical and biological aspects of Källby WWTP. Some of these effects are outlined below. 49

60 Flow In this project a considerable volume of wastewater is transported to Källby WWTP, either in dry or wet periods. The wastewater pumping rate into the sewer pipe is set to 50 and 120 l/s in normal and high flow cases respectively. The average inflow to Källby WWTP is reported to m 3 /d in 2007 (Dalby Miljörapport, 2007) so the discharged inflow of wastewater to Källby WWTP will increase 13% and 32% in dry and wet periods respectively. Peak flows in WWTPs depend on the time that the wastewater spends in the sewer network. Daily peak flows commonly occur in late morning and in the evening (Pescode, 1992). In this project, the wastewater, transported from Dalby, is discharged to Källby WWTP with 6 hours delay. This delay may result in peak flow distribution which is more favorable in Källby WWTP. Biological Nitrogen removal Municipal raw wastewater commonly contains a high concentration of ammonium and a low concentration of nitrate. On the contrary, the wastewater transferred from Dalby WWTP contains considerable concentrations of nitrate, 17.6 g NO 3 /m 3, and insignificant ammonium concentrations, 1.9 g NH 4 -N/m 3, in normal cases. The existing nitrate in the outlet of the sewer pipe will be diluted in a higher volume of the raw wastewater in Lund (more than 7 times) and will considerably decrease the nitrate concentration. However, the biological nitrogen removal is conducted according to the pre-denitrification method in Källby WWTP so that nitrate can be removed in the denitrification process. Biological phosphorous removal The nitrate concentration in the wastewater can consume easily-degradable organic matters which are favorable for phosphorous-accumulating bacteria. Therefore, the biological phosphorous removal efficiency is reduced since high concentrations of easily-degradable organic matters are degraded in the denitrification process. Moreover, when nitrate is present during the anaerobic period, the metabolism of phosphorous-accumulating bacteria is adversely affected by nitrate (Henze, et al., 1996). Although the nitrate concentration decreases by dilution with a high volume of the wastewater in the sewer network in Lund, it can still influence the biological phosphorous removal process in Källby WWTP. Besides, increasing inflow which is discharged to Källby WWTP affects the chemical treatment process so that a higher iron chloride concentration is required to be added for the phosphorous removal. Sludge production Sludge handling is one of the important processes in WWTPs. The municipal raw wastewater commonly contains a high suspended solid concentration, which mostly originates from human activities. The sludge production rate in biological treatment processes directly relates to the number of connected persons to the WWTPs. The present approach in Dalby WWTP produced 187 ton sludge in 2007, originating from 5636 connected persons to Dalby WWTP (Dalby Miljörapport, 2007). In a future approach, Dalby WWTP will be capable of handling wastewater from persons. As a rough estimation, sludge production in the future approach will be approximately 2.5 times more, 470 ton TS/year. The produced sludge will be diluted in the sewer 50

61 network in Lund before being discharged to Källby WWTP and could affect primary sedimentation and biological removal processes in Källby WWTP. 51

62 52

63 8. Conclusion Dalby-Lund sewer pipe project is planned to transport nitrified wastewater from Dalby WWTP to Källby WWTP under anoxic conditions. In-sewer microbiological transformations are expected to occur along the sewer pipe to reduce the nitrogen concentration in the outlet at the junction with the sewer network in Lund. However, the denitrification rate in the sewer pipe is relatively low and not significantly capable of removing nitrate concentrations under anoxic conditions. The produced biomass in the sewer pipe is fully discharged to the sewer network in Lund and there is not any possibility to recycle it to the sewer pipe. The low biomass concentration in the sewer pipe severely affects the denitrification process and consequently nitrate removal in the outlet even if the required external carbon source is added to the sewer pipe. The short residence time which the wastewater spends in the sewer pipe does not allow slowgrowing microorganisms to take part in the denitrification process and degrade organic matters. Therefore, only limited concentrations of bacteria are able to use organic matters as a source of energy under anoxic conditions and remove nitrate in the sewer pipe. The diffusion of different substrates into the sewer wall biofilm is a relatively slow process which restricts the denitrification process in the sewer pipe. Although the required external carbon source has been added to the sewer pipe, nitrate and methanol are not able to penetrate deep enough into the biofilm for the denitrification process to occur due to the low efficiency of the biomass in the sewer pipe. Lund municipality which is in charge of this project does not intend to use an external carbon source in the post-denitrification process in the sewer pipe. Therefore, the nitrate utilization rate will be limited to estimated values in this study, 0.27 g NO 3 /m 3 h and g NO 3 /m 2 h in the bulk water and biofilm phases of the sewer pipe respectively. Because of all these limitations, it seems that the sewer pipe will only transfer nitrified wastewater to the sewer network in Lund without any significant nitrate removal in the outlet. By construction of the Dalby-Lund sewer pipe, the main purpose of this project, that is transporting wastewater from Dalby and other communities to Källby WWTP, will be fulfilled. However, the microbiological transformation in the sewer pipe cannot efficiently remove nutrients under anoxic conditions. 53

64 54

65 9. Recommendations Dalby-Lund sewer pipe project is hydraulically designed but it is not executed yet. Therefore, all the achievements in this study are based on theoretical knowledge of the denitrification process under sewer conditions since there is no possibility to experimentally investigate the denitrification rate in the sewer pipe. In addition, the pre-denitrification process is conducted in Dalby WWTP in the present approach so no experimental results are available to estimate the concentrations of substances after the nitrification process in the future approach. For this reasons, author strongly recommends that substrate concentrations in the nitrification process and nitrate removal efficiency in the sewer pipe should be experimentally evaluated when the Dalby- Lund sewer pipe is a reality. The treatment processes in Källby WWTP will be affected by this project as soon as wastewater from Dalby is transferred to Lund. In this regard, the capacities and capabilities of different treatment processes in Källby WWTP, primary sedimentation, biological nitrogen and phosphorous removal, chemical treatment and sludge digestion, should be evaluated. The available data from treatment processes conducted in Dalby WWTP and wastewater characteristics from Dalby and other communities joining this project in the second phase are limited. Dalby s wastewater will be discharged to the high-capacity WWTP in Lund, Källby WWTP, in the near future affecting technical and biological aspects of treatment processes. Thereby, providing more detailed data from Dalby WWTP and the sewer pipe is necessary for further investigations. EFOR computer software is only capable of modeling the denitrification process in the bulk water phase of the sewer pipe. Therefore, calculations of the denitrification process in the sewer wall biofilm were done manually. This was inconvenient because all the aspects for the denitrification process cannot be precisely considered with manual calculations. The new version of the MOUSE TRAP computer program is capable of simulating denitrification processes under sewer conditions but is not commercially released yet. It is a good idea to simulate this project in the MOUSE TRAP computer software in the future to assess the denitrification process under sewer condition. 55

66 56

67 Reference: - Abdul-Talib, S. Hvitved-Jacobsen, T. Vollertsen, J. & Ujang, Z., Anoxic transformations of wastewater organic matter in sewers - process kinetics, model concept and wastewater treatment potential. Water Science and Technology, [Online]. 45 (3), p Ahnert, M. Kuehn, V. & Krebs. P., Identification of Overall Degradation in Sewer Systems from Long-Term Measurements and Consequences for WWTP Simulations. Water Science and Technology, [Online]. 52 (3), p Bentzen, G. Smith, A.T. Bennett, D. Webster, N.J. Reinholt, F. Sletholt, E & Hobson, J., Controlled dosing of nitrate for prevention of H 2 S in a sewer network and the effects on the subsequent treatment processes. Water Science and Technology, [Online]. 31 (7), p Boon, A.G., Septicity in sewers: causes, consequences and containment (basic information in H2S generation). Water Science and Technology, [Online]. 31 (7), p DHI Water and Environment, EFOR User Guide. - Folder Collected for Urban Water Course at Department of Chemical Engineering, Environmental and Water Engineering, Lund University, Lund, Sweden. - Gerardi, M.H., Nitrification and Denitrification in the Activated Sludge Process. [ebook]. New York: Wiley and sons Inc. Available at Lund University s Online Library. - Harremoës, P. Andersen, H.S. Dupont, R. Jacobsen, P. & Rindel, K., Analysis of scenarios for sewerage, wastewater treatment and prioritised load on environment from the Greater City of Copenhagen. Water Science and Technology, [Online]. 45 (3), p Henze, M. Harremoës, P. la Cour Jansen, J. & Arvin, E.,1996. Wastewater Treatment. 2nd ed. New York: Springer Verlag. - Hvitved-Jacobsen, T. Vollertsen, J. & Matos, J.S., The sewer as a bioreactor - a dry weather approach. Water Science and Technology, [Online]. 45 (3), p Hvitved-Jacobsen, T., Sewer Processes: Microbial and Chemical Process Engineering of Sewer Networks. [e-book]. Florida: CRC Press. Available at Lund University s Online Library. - Huisman, J.L. Krebs, P. & Gujer, W., Integral and unified model for the sewer and wastewater treatment plant focusing on transformations. Water Science & Technology, [Online]. 47 (12), p Huisman, J.L. & Gujer, W., Modeling wastewater transformation in sewers based on ASM3. Water Science & Technology, [Online]. 45 (6), p Lindquist, A. ed., About Water Treatment. Helsingborg: Kemira Kemwater. 57

68 - Mathioudakis, V.L. Vaiopoulou, E. & Aivasidis, A., Addition of Nitrate for Odor Control in Sewer Networks: Laboratory and Field Experiments. Global Nest Journal, [Online]. 8 (1), p Michigan University, Asynchronous learning, Plug Flow Reactor (PFR).Available at: - Miljörapport, Dalby avloppsreningsverk, Lund Kommun. - Miljörapport, Källby avloppsreningsverk, Lund Kommun. - Svenskt Vatten, Mars Dimensionering av allmänna avloppsledningar, Publikation P90. - Pescode, M.B., Wastewater treatment and use in agriculture - FAO irrigation and drainage, Paper 47. [Online] Food and Agriculture Organization of the United nations.available at: - Statistiska Centralbyrån, Folkmängd Per Tätort och Småort Per Kommun [Online], Available at: - Vollertsen, J. Hvitved-Jacobsen, T. Ujang, Z.& Abdul-Talib, S., Integrated design of sewers and wastewater treatment plants. Water Science and Technology, [Online]. 46 (9), p Vollertsen, J. Nielsen, A.H. Yang, W. & Hvitved-Jacobsen, T., Effects of in-sewer processes: a stochastic model approach. Water Science and Technology, [Online]. 52 (3), p Yang, W. Vollertsen, J. & Hvitved-Jacobsen, T., Anoxic control of odour and corrosion from sewer networks. Water Science and Technology, [Online]. 50 (4), p Yang, W. Vollertsen, J. & Hvitved-Jacobsen, T., Anoxic sulfide oxidation in wastewater of sewer networks. Water Science and Technology, [Online]. 52 (3), p Yongsiri, C. Vollertsen, J. & Hvitved-Jacobsen, T., Influence of wastewater constituents on hydrogen sulfide emission in sewer networks. Journal of environmental engineering, [Online]. 131 (12), p Zhang, L. Schryver, P.D. Gusseme, B.D. Muynck, W.D. Boon, N. Verstraete, W., Chemical and biological technologies for hydrogen sulfide emission control in sewer systems: A review. Water research, [Online]. 42 (1-2), p

69 Appendix 1 Optimum (C/N) ratio for different types of organic matter to be used for denitrification (Henze, et al., 1996). Organic matter (C/N) optimum Unit Organic matter in wastewater kg BOD/kgN 4-5 kg COD/kgN Organic matter in sludge kg BOD/kgN kg COD/kgN Methanol kg MeOH/kgN kg COD/kgN mol MeOH/ mol N Acetic acid kg HAc/kgN kg COD/kgN mol HAc/ mol N Appendix 2 The efficiency factor (f C/N ) for organic matter for different plant designs for denitrification (Henze, et al., 1996). Activated sludge plant f C/N Separate Post-denitrification Post-denitrification (with extra carbon) Recycle Alternating Alternating, one-tank Simultaneous

70 Appendix 3 Sulfur and Nitrogen-Containing Odorous Compounds in the Influent Wastewater at a Treatment Plant (Hvitved-Jacobsen, 2002) Compound Average Concentration (μg/l) Range of Concentrations (μg/l) Hydrogen sulfide Carbon disulfide Methyl mercaptan Dimethyl sulfide Mar Dimethyl disulfide Dimethylamine Trimethylamine 78 - n-propylamine 33 - Indole Skatole Appendix 4 Odor and Human Health-Related Effects of Hydrogen Sulfide in the Atmosphere (Hvitved- Jacobsen, 2002) Odor or Human Effect Concentration in Atmophere (ppm) Treshold odor limit Unpleasent and strong smell Headache, nausea and eye, nose and throat irritation Eye and respiratory injury Life threatening Immediate death >700 60

71 Appendix 5 Kinetic constants for oxygen and organic matters in the biofilm reactor (Henze, et al., 1996). D ν NO3,COD k 0vf Substance (10-4 m 2 /d) (g COD/g O 2 ) (kg COD/m 3 d) Oxygen Acetic acid Methanol Glucose Unspec. COD Unspec. BOD

72 Appendix 6: Article Evaluation of in-sewer denitrificationin in a transmission pipe for nitrified wastewater. Case study: Dalby-Lund, Sweden F.Edalat Water and Environmental Engineering, Department of Chemical Engineering, Lund University, P.O.Box 124, SE , Lund, Sweden farnazedalat@yahoo.com Abstract Denitrification in the Dalby-Lund sewer pipe, transporting nitrified wastewater to the sewer network in Lund, has been evaluated. The nitrification process will be conducted in Dalby WWTP while the sewer pipe will be used as a denitrification unit. EFOR computer software was used to simulate the denitrification process in the bulk water phase of the sewer pipe. Model scenarios in which internal and external carbon sources were injected to the sewer pipe were simulated. The highest denitrification rate in the bulk water phase of the sewer pipe was 0.67 g NO 3 /m 3 h when the required external carbon source, m 3 /h methanol, was added to the sewer. The contribution of the sewer wall biofilm in the nitrate removal under anoxic conditions was evaluated manually. The organic matter was the limiting substrate in the biofilm; thereby, g/m 3 methanol was estimated to compensate the organic matter deficiency. The denitrification rate was estimated to 0.13 g NO 3 /m 2 h in the denitrifying biofilm reactor. Consequently, the organic matter deficit and low biomass concentration in the sewer pipe has limited the denitrification capacity of the sewer pipe. Keywords In-sewer processes; denitrification; nitrate; anoxic; methanol; nitrate removal; biofilm; EFOR; sewer; denitrification rate 1. Introduction Wastewater treatment plants had been considered as the only part of the urban wastewater management for the biological processes while sewer pipes only collected and transported wastewater to WWTPs. However, sewer pipes, as bio-chemical reactors, take part in microbiological transformations of organic matters in the wastewater, while being transported (Abdul-Talib et al., 2002; Hvitved-Jacobsen et al., 2002; Yang et al., 2004). Since microbiological transformations in sewer pipes can affect the quality of the wastewater, integrated wastewater treatment processes are considered as starting at the kitchen sink (Hvitved-Jacobsen, 2002). Living microorganisms, present in the wastewater, degrade organic matters as a source of energy for their cellular growth. In-sewer microbiological processes may occur under aerobic, anoxic or anaerobic conditions. Dominant electron acceptors, oxygen, nitrate or sulfate, tend to take part in redox reactions in the sewer network. As long as aerobic conditions prevail in the sewer network, the free molecular oxygen participates in the redox reaction with a high biomass yield. In the absence of oxygen, nitrates, having relatively high redox potential, degrade organic matters under anoxic conditions. While the oxygen and nitrate are depleted, sulfate is the dominant electron acceptor in the degradation of substrates, producing hydrogen sulfides and 62

73 volatile fatty acids. The hydrogen sulfide formation in the sewer network may cause intensive corrosion, odor and health problems (Hvitved-Jacobsen et al., 2002; Vollertsen et al, 2005; Yang et al., 2005). Anoxic conditions do not normally prevail in the sewer network due to a low amount of nitrate in the municipal wastewater. However, it may occur if nitrate is deliberately injected to the sewer system to prevent hydrogen sulfide formation (Yang et al., 2005; Mathioudakis et al., 2006; Yang et al., 2004). Cases in which high nitrate concentrations are available in the sewer system, the denitrification process may be limited due to deficiency of organic matters. In activated sludge, the denitrification rate is evaluated by a carbon/nitrogen ratio (C/N) in the denitrification process. If the internal carbon source is not sufficient for the denitrification process, mainly in a post-denitrification method, external carbon is applied to compensate the internal carbon source deficiency. External carbon sources, e.g. methanol, ethanol and acetate, are expensive to use in the denitrification process and their applications are limited to special cases. However, a few studies are conducted to investigate the in-sewer microbiological transformations in the sewer network under anoxic conditions. The aim of this study is to evaluate the denitrification rate in both bulk water and biofilm phases of the sewer pipes under anoxic conditions. Due to lack of information about the denitrification kinetics in the sewer network, under anoxic conditions, this study is conducted based on theoretical concepts of the activated sludge method. Material and method Present approach in Dalby WWTP Dalby WWTP serves 5636 connected residents and average inflow to this WWTP is reported 30 l/s in 2007 (Dalby Miljörapport, 2007). In Dalby WWTP, the wastewater is treated with pretreatment processes, pre-denitrification processes and subsequently a chemical treatment process. An external carbon source is not added to the denitrification process. The chemical treatment is conducted by adding iron chloride as a coagulant. Eventually the excess sludge, which is centrifuged and decreased in ph by lime, is used as fertilizer. Wastewater composition The wastewater entering Dalby WWTP is municipal wastewater. The annual average concentration of its constituents has been presented in Table 1. Table 1: Influent wastewater composition to Dalby WWTP. Average concentration Parameter (mg/l) BOD Total N 46 Total P 9.3 Project description The anticipated population growth in Dalby in the near future has brought forward the necessity of increasing the capacity of Dalby WWTP. However, due to high costs and operational 63

74 problems, Lund municipality, which is in charge of this WWTP, has decided to transport the wastewater through a sewer pipeline to Källby WWTP in Lund. In this project, biological nitrogen removal in Dalby WWTP will be changed from the pre-denitrification to the postdenitrification process. The nitrification process will be conducted in Dalby WWTP, however, the denitrification is supposed to occur in the sewer pipe under anoxic conditions while being transported to Källby WWTP. After executing this project, Dalby WWTP will be capable of handling an average incoming flow of 80 l/s wastewater from connected residents. This will increase to 200 l/s in high flow cases. The treatment scheme, conducted in Dalby WWTP in the future approach, consists of pre-treatment processes, a nitrification process, sedimentation, a detention tank and a pumping station. In order to control the overflow risk in wet periods, a detention tank with 6000 m 3 capacity will be used. The underground sewer pipe is designed in two distinct sections coupled together by an ordinary manhole situated in a monitoring station. The first part is made of polyethylene and is under pressure whereas the second part is a gravity sewer in concrete. They are 0.4 and 0.6 m in diameter and 7.36 and 0.87 km long respectively. The nitrified wastewater and the excess sludge will be pumped into the sewer pipe at a rate of 50 and 120 l/s in normal and high flow cases respectively. Further raw wastewater can be bypassed to the sewer. The sewer pipe, joining the sewer network in the south-east of Lund and blended with a high volume of municipal wastewater, will eventually be discharged to Källby WWTP. Computer modelling EFOR software has been applied to simulate the treatment processes in Dalby WWTP and the sewer pipe as anoxic units. EFOR is capable to model the denitrification process only in the bulk water phase of the wastewater. The layout of the model in EFOR is presented in Figure 1. AS1, AS2 and AS3 represent nitrification tanks in the Dalby WWTP while AS4 and AS5 represent the sewer pipe as two anoxic tanks. Pumps 1, 2 and 3 control the amount of return sludge, excess sludge and methanol respectively. Valve 1 is a control device to manage the proportion of the bypassed raw wastewater to the sewer pipe. Figure 1: Model of the treatment processes in Dalby WWTP and the sewer pipe in EFOR. Three different scenarios are separately modeled in EFOR to evaluate the nitrate removal efficiency and the denitrification rate in the sewer pipe. 64

75 Scenario1 The sewer pipe is modeled as two anoxic chambers representing pressure and gravity pipes respectively. The volumes of the chambers are set to the capacities of the pipes, 927 and 123 m 3. Scenario 2 Certain amount of the raw wastewater (10% of the total inflow) is bypassed to the sewer pipe as the internal carbon source. Scenario 3 Required amount of the external carbon source, m 3 /h methanol, is injected to the sewer pipe. Manual calculations The denitrification process in the biofilm phase of the sewer pipe has been evaluated manually. The denitrification process in the sewer wall biofilm occurs when sufficient organic matters are available for denitrification reactions in the sewer pipe. Moreover, organic matters should penetrate the biofilm as deep as nitrate does. The required organic matter in the sewer pipe (C/N Practice ) for the denitrification process can be calculated as (Henze et al, 1996), : Optimum C/N ratio for different types of organic matters in the denitrification process and : The efficiency factor for the organic matter in different plant designs for denitrification. The denitrification rate in the biofilm phase of the sewer pipe depends on the nitrate and organic matter diffusion rate into the biofilm layer. The shortest distance in which either nitrate or organic matter penetrate into the biofilm determines which substrate is limiting for nitrate removal. The expression that distinguishes the limiting substrate for nitrate removal, either nitrate or organic matter, is expressed as (Henze et al, 1996), : Diffusion coefficient, : Concentration of substrate and : Stoichiometric coefficient. Sign > means that the carbon source is potentially limiting for the nitrate removal while sign < implies that the nitrate is potentially limiting for the removal. In this study, the substrates are assumed to partially penetrate into the biofilm layer and the nitrate removal is based on the half order reaction. Along the sewer pipe, concentrations of organic matter and nitrate decrease because of insewer microbiological transformations under anoxic conditions. Thereby, the sewer pipe is considered as a plug flow reactor along its axial direction. According to the mass balance of an infinitesimal plug of the fluid, dx, the substrates concentrations in the outlet and the denitrification rate in the sewer pipe can be estimated as (Henze et al, 1996), 65

76 : The specific biofilm area, : Half order area specific rate constant, : Initial concentration of substrate, : Concentration of substrate in the outlet, : Residence time and : Wetted area to volume ratio of the sewer. Injecting an external carbon source to the denitrification process is a solution to increase the organic matter concentration in the system. In this study, the denitrification rate in the biofilm is evaluated when the required external carbon source, methanol, is added to the sewer pipe. 3. Results Denitrification rate in the bulk water The denitrification rates in the computer models are evaluated by the nitrate removal efficiency in the outlet of the sewer pipe. The scenario 1 does not show significant nitrate removal compared to the initial nitrate concentration in the sewer pipe which is estimated to 17.6 g NO 3 /m 3. The nitrate concentration decreases to 16 g NO 3 /m 3 in the outlet and the denitrification rate is estimated to 0.27 g NO 3 /m 3 h in the sewer pipe. In scenario 2, bypassing 10% of the raw wastewater to the sewer pipe does not decrease the total nitrogen compounds concentrations (nitrate and ammonium) in the outlet; however, only 90% of the total inflow is nitrified in the nitrification process so initial nitrate concentration in the sewer pipe is less than in scenario 1. In fact, the high ammonium concentration in the raw wastewater increases the nitrogen compounds concentrations in the outlet. Despite less nitrate concentrations in the outlet compared to the scenatio1, bypassing the raw wastewater does not effectively enhance the nitrogen removal efficiency in the sewer pipe. Injecting the external carbon source, methanol, to the sewer pipe in scenario 3 shows a low nitrate removal efficiency in the outlet. The result indicates that the nitrate concentration decreases to 13.8 g NO 3 /m 3 in the outlet and the denitrification rate is 0.63 g NO 3 /m 3 h in the sewer pipe. As a result, the nitrate removal efficiency is estimated to 22% even if the required methanol concentration for the complete denitrification process is added to the sewer pipe. The nitrogen compounds concentrations in the outlet, both ammonium and nitrate, in different scenarios are illustrated in Figure 2. In order to evaluate the contribution of soluble COD (SCOD), of which soluble inert COD has been subtracted, in the denitrification process, the SCOD concentrations in the outlet of the sewer pipe, in different scenarios, are compared. The initial concentration of the SCOD in the sewer pipe, 21 g COD/m 3, does not show a significant change in the outlet in the scenario 1. On the contrary, the SCOD concentration in the outlet increases in scenario 2 when raw wastewater with high SCOD concentration is bypassed to the sewer pipe. Scenario 3 also shows a higher SCOD concentration in the outlet compared to scenario 1, due to its low consumption in the sewer pipe. Figure 2 presents SCOD concentrations in the outlet in different scenarios. The low nitrate removal efficiency in different scenarios implies that the denitrification rate is quite low in the bulk water phase in the Dalby-Lund sewer pipe despite the fact that sufficient organic matters are available in the sewer pipe. 66

77 Figure 2: Left: Nitrate and ammonium concentrations in the outlet in different scenarios. Right: SCOD concentrations in the outlet in different scenarios. Denitrification rate in the sewer wall biofilm The initial concentrations of SCOD and nitrate in the sewer pipe are estimated to 21 g SCOD/m 3 and 17.6 g NO 3 /m 3 respectively, obtained in the computer model. By assuming (C/N) Optimum = 9 kg SCOD/kg N (Henze et al, 1996), results show that the initial SCOD concentration in the sewer pipe is not sufficient for removing the initial nitrate concentration. Therefore the internal carbon source is the limiting factor for the nitrate removal in the biofilm phase of the sewer pipe. The diffusion coefficients for nitrate and SCOD are considered 0.5*10-4 m 2 /d and the stoichiometric coefficient, ν NO3,COD, is estimated to 5.3 g COD/g NO 3 -N (Henze et al, 1996). As a result, the available internal organic matter in the wastewater can only remove 20% of the initial nitrate in the sewer pipe, 3.44 g NO 3 /m 3. Therefore, the denitrification rate in the denitrifying biofilm reactor is limited to g NO 3 /m 2 h due to deficiency of organic matters. The denitrification rate is also evaluated if the required external carbon source, methanol, is injected to the sewer pipe to compensate the internal organic matter deficiency. The required methanol for removing the remained nitrate concentration in the sewer pipe is estimated to g MeOH/m 3 by assuming (C/N) Optimum = 2.78 kg MeOH/kg N (Henze, et al., 1996). The diffusion coefficients for nitrate and methanol are considered 0.6*10-4 and 1* 10-4 m 2 /d respectively and the stoichiometric coefficient, ν NO3,Me, is estimated to 3.5 g MeOH/g NO 3 -N (Henze et al, 1996). This amount of the methanol can penetrate into the biofilm layer as far as the remained nitrate does. This means that the organic matter is not the limiting factor for the nitrate removal anymore. The concentrations of nitrate and methanol in the outlet are calculated to 7.35 g NO 3 /m 3 and g MeOH/m 3 which indicate that methanol is not the limiting factor in the outlet and consequently along the sewer pipe. Results show that even though the methanol concentration is sufficient for the denitrification process in the biofilm along the sewer pipe, nitrate cannot efficiently be removed. In fact, the denitrification rate in the sewer pipe increases to 0.13 g NO 3 /m 2 h while considerable amounts of nitrate and methanol are present in the outlet. 67