Production of easily biodegradable carbon source for nutrient removal from wastewater through primary sludge hydrolysis - Impact on methane potential

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1 Water and Environmental Engineering Department of Chemical Engineering Production of easily biodegradable carbon source for nutrient removal from wastewater through primary sludge hydrolysis - Impact on methane potential Master s Thesis by Sílvia Juncà June 2010

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3 Vattenförsörjnings- och Avloppsteknik Institutionen för Kemiteknik Lunds Universitet Water and Environmental Engineering Department of Chemical Engineering Lund University, Sweden Production of easily biodegradable carbon source for nutrient removal from wastewater through primary sludge hydrolysis - Impact on methane potential Master Thesis number: by Sílvia Juncà Water and Environmental Engineering Department of Chemical Engineering June 2010 Supervisor: Associate professor Karin Jönsson Examiner: Professor Jes la Cour Jansen Picture on front page: Biological hydrolysis reactors used in the laboratory experiments. Postal address: Visiting address: Telephone: P.O. Box 124 Getingevägen SE Lund Sweden Telefax: Web address:

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5 Preface and Acknowledgements This report is the result of my Master Thesis, done during the spring term 2010, and represents the end of my degree in Chemical Engineering, carried out in the Technical University of Catalonia (Barcelona, Spain) and finished as exchange student in Lund University (Lund, Sweden). This Master Thesis has been hold in the Department of Chemical Engineering Water and Environmental Engineering, and its main topic refers to the area of wastewater treatment. First of all, I would like to thank to my supervisor Karin Jönsson for the support and guidance given during this months, and to my examiner Jes la Cour Jansen for the suggestions and interesting discussions. During this time, I also received support from more others, and without their help I would not have been able to present this work as it is now. Is the case of Gertrud and Ylva Persson, who helped me with the laboratory work and analysis whenever I needed; Åsa Davidsson, who gave me assistance with the methane production tests; Tobias Hey, who always showed interest in my work, provided me a useful collection of literature, and helped me with the statistics treatment of the data; Per Falås, who always showed interest in my work and with who I often discussed the interpretation of the results; and I could not leave without mention Aude Baillon-Dhumez and Laura Guerra, who have been exceptional support and have given me a hand in the moments when I have needed. Finally, I would like to express my gratitude to VASYD for providing the sludge samples needed to carry out the experiments. Lund, 11 th June. Sílvia Juncà V

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7 Summary Good quality carbon source is a prerequisite for efficient operation of biological nitrogen and phosphorus removal from wastewater. In order to be easily degradable by nutrient-removal microorganisms, the carbon source must be composed of soluble and/or short-chained organic compounds. If good quality carbon is not available in the influent of the wastewater treatment plant, it is possible to add it, either externally (purchasing external carbon source, such as ethanol or methanol) or using the influent slowly biodegradable organic matter present converted into easily biodegradable carbon via hydrolysis of sludge. This study is focused on biological hydrolysis of primary sludge, which was kept during 4 or 10 days in laboratory-scale reactors under anaerobic conditions. The sludge hydrolysis reached a maximum COD solubilisation of 13-22% in units of dissolved COD/initial COD total, and a maximum VFA production of 57-67% in units of dissolved COD VFA /dissolved COD. After 4 days of hydrolysis, when around the 85-89% of the maximum COD VFA /dissolved COD had already been reached, the hydrolysate was withdrawn and the remaining hydrolysed sludge was used for methane production. The methane potential of the hydrolysed sludge was compared to the methane potential of the fresh sludge (non-hydrolysed). The methane potential tests were performed following the method developed by Hansen et al. (2003) and modified by Davidsson (2007). The inoculum used for the methane production was from the mesophilic digesters at Sjölunda WWTP. The methane production of the sludge was calculated for every bottle based on its VS (or VSS when VS values were not available) before hydrolysis (Nml CH 4 /gvs(s) initial ), excluding the methane produced by the inoculum itself and expressed as accumulated production. The differences in methane production between fresh and hydrolysed sludge were analysed by comparing the measurements of the triplicate bottles whose average gave the maximum accumulated methane production. The obtained productions for fresh and hydrolysed sludge from every WWTP have been analysed statistically by means of the 95% confidence interval and with the unpaired t-test. The results of the methane production test showed that the 4-days-hydrolysed sludge from Klagshamn achieved a lower average accumulated methane production (308 Nml/gVSS initial ) than the fresh sludge (473 Nml/gVSS initial ; based in a guess of the VSS initial value). For the sludge from Källby, the hydrolysed sludge also achieved lower average accumulated methane production (246 Nml/gVSS initial ) than the fresh sludge (285 Nml/gVSS initial ), even though the overlapping of the confidence intervals indicates that this difference could not be significant. The situation for the sludge from Sjölunda was a bit different: the maximum average accumulated methane production was slightly higher for the hydrolysed sludge (303 Nml/gVSS initial ) than for the fresh sludge (282 Nml/gVSS initial ), but the 95% confidence intervals were also overlapped. From the statistical analysis of the results it has been concluded that for this experiment the hydrolysis lead to a statistically significant decrease in methane potential for the sludge from Klagshamn, but it has not been confirmed statistically that the hydrolysis lead to a negative impact in methane production for the sludge from Sjölunda and Källby. VII

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9 Table of contents Preface and Acknowledgements V Summary VII Table of contents IX 1 Background Nutrient removal from wastewater Sludge treatment 3 2 Aim 7 3 Literature review Importance of the substrate biodegradability for biological nutrient removal from wastewater Importance of substrate biodegradability in nitrogen removal Importance of substrate biodegradability in biological phosphorous removal Hydrolysis of sludge Primary Sludge Hydrolysis Return and Mixed Sludge Hydrolysis Influence of hydrolysis conditions in VFA production from sludge Temperature and ph Hydrolysis duration, sludge retention time and hydraulic retention time Sludge Concentration Methane potential of sludge and economical balance at WWTP 18 4 Materials and methods Experiment description Description of methods used Methane potential tests Method description Set-up volumes Calculation of the methane production End of digestion 30 5 Results and discussion Biological hydrolysis Control measures: ph and temperature Solids concentration Organic matter concentration Nutrients Methane Potential Methane production during experiment A Methane production during experiment B 58 Conclusions 65 Future Work 67 References 69 Table of abbreviations and units 73 Appendix 1: Hydrolysis 75 Appendix 2: Methane Potential Tests 85 Appendix 3: Scientific article 97 IX

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11 1 Background The use of water has become indispensable for human activity. The consequent formation of wastewater creates an important impact on the environment which has a repercussion in human life. The concern about wastewater treatment has lead to the development of processes in order to separate pollutants from water. Due to the complex wastewater composition, its treatment requires different steps, which are specifically focused on the removal of the different components. The current study is basically focused on two of these processes: nutrient removal and sludge treatment. Usually both processes are relatively independent, since they have different targets: the nutrient removal objective is to remove nitrogen and phosphorous from wastewater; the sludge treatment objective is to minimize the environmental impact of the subsequent waste product formed through the whole wastewater treatment process. This study will refer to an issue that both processes have in common: the use of primary sludge. Primary sludge is rich in organic matter, and can therefore be hydrolysed in order to produce carbon source which can be used in the nutrient removal processes, reducing the costs of adding an external carbon source. On the other hand, the primary sludge is rich in potentially degradable organic matter, which makes it highly convenient for methane production. The withdrawal of the sludge hydrolysate to be used in the nutrient removal process could have a negative effect on its methane potential, hence an economical conflict is created. 1.1 Nutrient removal from wastewater Nitrogen and phosphorous are two important elements sometimes present in large amounts in wastewater as nutrient salts. An excessive concentration in the discharged waters to the environment can lead to environmental problems such as eutrophication. The allowed discharge quantities of those compounds are regulated by law, and vary according to the country and region. In order to meet the permitted concentration values in the effluent, nitrogen and phosphorous compounds often have to be removed from wastewater. Removal of these nutrients can be carried out biologically or chemically (mainly used for phosphorous). In the biological nutrient removal processes (BNR), certain types of bacteria take up and transform these nutrients, making possible its separation from wastewater. Nitrogen Nitrogen can be present in wastewater in different forms: ammonium (NH 4 + ), nitrate (NO 3 - ), nitrite (NO 2 - ) or bounded in organic compounds. The majority of nitrogen present in wastewater is in the form of ammonium (Kemira Kemwater, 2003), or organically bounded nitrogen, which is rapidly converted in ammonia in the sewer net due to the molecular breakdown. Its removal process is divided in two steps: first converting ammonia into nitrate (nitrification) and in its turn converting nitrate into nitrogen gas (denitrification). During the nitrification, ammonium is oxidised into nitrate by a group of autotrophic bacteria also known as nitrifiers. The nitrification rate is related to temperature, achieving higher nitrification rates at higher temperature. During this process the nitrifying bacteria compete 1

12 against the BOD-reducing bacteria, which are heterotrophic and more effective. For this reason, low BOD/N ratios represent an advantage for the nitrifying bacteria, whose growth will be favoured in the situation of lack of substrate for the BOD-reducing bacteria, and consequently a more effective nitrification process can be performed. According to Kemira Kemwater (2003) BOD-reducing bacteria are favoured at ratios higher than 4, at which the population of nitrifying organisms is much smaller than at lower ratios. Once ammonium has been converted into nitrate, this can be removed from wastewater by denitrification, through reduction into nitrogen gas (N 2 ) while organic matter is oxidised. Most denitrifying microorganisms are facultative, which means that they can use either molecular oxygen (O 2 ) or nitrate/nitrite (NO 3 - /NO 2 - ) as an oxidising agent, but will always prefer to use oxygen if present in the water. For this reason the process requires anoxic conditions, which means absence of dissolved oxygen, but present as bound-up into nitrate or nitrite. Denitrifying bacteria are heterotrophic and need organic carbon as a substrate. In order to have high denitrification rates, it is important to ensure that the carbon source exists in an easily accessible form, which means both dissolved and short-chained. The better quality the carbon source has, the faster the denitrification process takes place. If there is a lack of easily biodegradable organic matter in the influent wastewater, a higher hydraulic retention time (HRT) in the denitrification basin will be needed in order to remove the required amount of nitrogen. These two processes are always performed in the same order due to its nature: first nitrification, second denitrification. Despite this fact, nitrification and denitrification basins do not need to be necessarily arranged in the natural order along the wastewater flow in WWTPs. Depending on which of the two processes is arranged first in the WWTP, several alternatives for nitrogen removal can be classified: pre-denitrification, post-denitrification (see Figure 1) or a combination of them. In the pre-denitrification system, heterotrophic bacteria can use the easily degradable organic matter present in the influent waste water as carbon source, as long as the COD/N ratio is sufficient. The needed COD/N proportion for denitrification is around 4-5 kg COD/kg N denitrified (Henze et al., 1997). On the contrary, in the post-denitrification system, the denitrification basin is placed after the nitrification basin (aerobic). In consequence the main organic fraction has already been oxidised in the previous aerobic stage, leading to a low COD/N ratio in the anoxic zone, which does not favour the denitrifiers. In order to ensure the growth and activity of the denitrifying bacteria, the addition of carbon source (such as methanol or ethanol) is hence necessary in this process configuration. NO 3 - C-source NH 4 + NO 3 - N 2 Anoxic NH 4+ NO 3 - Aerobic NH 4 + NH 4+ NO 3 - Aerobic NO 3 - N 2 Anoxic Pre-denitrification Post-denitrification Figure 1. Arrangement possibilities for nitrogen removal in WWTPs. 2

13 Phosphorous Phosphorous can be removed from wastewater by chemical precipitation. However, the purchasing of chemicals represents a substantial cost, and therefore an alternative biological method has gained interest. Biological phosphorous removal can be performed by a specific kind of bacteria able to take up unusually large amounts of phosphorous in the form of polyphosphates. This group of microorganisms is known as PAOs (Phosphate Accumulating Organisms), sometimes also called bio-p bacteria. The process is achieved by stressing bacteria exposing them alternately to aerobic and anaerobic conditions. Under anaerobic conditions, where there is no dissolved or nitrate-bonded oxygen available in the wastewater, bio-p bacteria show the ability to use their stored energy in form of polyphosphate chains, while other organisms cannot grow under lack of oxygen, and therefore the PAOs are favoured. When the energy accumulated in the polyphosphate chain bonds is used for the uptake of VFAs (i.e. to metabolise the organic matter), the phosphates are released to the surrounding water. Under aerobic conditions, dissolved oxygen is used for energy production, but then there is also a high competition between PAOs and other organisms, therefore bio-p bacteria have to use their previously stored substrate to build up new biomass, and to store energy in form of polyphosphate chains. By having the last process step as aerobic, stored phosphorous can be removed when removing the biomass. Usually phosphorous release during anaerobic conditions is not as high as the amount they store under aerobic conditions (Moser- Engeler et al., 1998), and hence there is a net uptake of phosphorous. Despite that this process is not always able to meet the standard effluent values, it might reduce the amount of precipitation chemicals needed, which makes this alternative highly suitable for phosphorous removal. In order to perform this process, also known as enhanced biological phosphorous removal (EBPR), the presence of simple organic compounds such as volatile fatty acids (VFAs) in the wastewater is needed, since they constitute an easily degradable carbon source and are an essential substrate under anaerobic conditions. If these compounds are not originally present in the influent wastewater, they can be added. The carbon source used for nutrient removal processes can be classified in two main types depending on its origin: External, which implies that the feed used comes from another site, and therefore needs to be purchased. For nitrogen removal, it usually consists of easily biodegradable organic compounds, most commonly methanol, ethanol, acetic acid or starch, often side-products from other industrial processes. The drawbacks are the purchasing and transport costs, and the extra sludge production. Internal, which consists on the organic matter present in the wastewater itself. Usually particulate organic matter is partially inaccessible for the denitrifying microorganisms, and therefore needs a high retention time in order to be metabolised. In order to achieve a higher reduction rate, the organic fraction of the sludge (primary or return sludge) can be added, if previously hydrolysed and consequently converted into a more easily degradable form. 1.2 Sludge treatment The concept of sludge is quite wide, and its origin needs to be specified in order to avoid confusion. Sludge can be produced either from potable water production or from sewage 3

14 treatment. During water and wastewater treatment processes, a complex mixture of undesired contaminants (suspended solids, organic matter, nutrients, bacteria, viruses, parasites, heavy metals, toxic substances, etc.) is separated from water. This separation is achieved by transforming dissolved substances and small particles into larger particles which can be more easily removed. This mixture of clustered larger particles and water, together with the organic matter already in form of particles in the influent is what constitutes the sludge. In the current study only sewage sludge will be referred to. Depending on the different processes applied in wastewater treatment, the sludge obtained and withdrawn from the process can be differentiated in several types. The following classification is commonly used: Primary sludge: formed during mechanical treatment. Secondary sludge: also known as biological sludge, formed during biological treatment. Tertiary sludge: also known as chemical sludge, formed during post-precipitation process, performed by addition of chemicals. All of them have to receive further treatment, and they are often mixed. Sludge treatment consists of different steps, usually performed in the following order: thickening, stabilisation and dewatering. Sludge processing represents an important part of the total costs at a WWTP. Thickening is carried out in order to reduce the volume and consequently the costs of the sludge treatment process. Stabilisation consists of decomposing the biodegradable compounds that remain biologically in the sludge, and can be performed by several different methods such as anaerobic digestion, aerobic stabilisation, composting, thermal or chemical oxidation, incineration, pyrolysis or liming. Since sludge has a high percentage of water (more than 90%), dewatering is needed in order to reduce the volume of the final product, and can be done by several methods, e.g. sludge bed drying or centrifugation. Anaerobic digestion The current study is focused on the anaerobic digestion. The process transforms raw sludge into a more hygienic material and the smell is highly reduced. The solid material can be further dewatered and handled as solid waste, or composted and further used as fertilizer for agriculture or forestry. The anaerobic digestion process also produces biogas, composed of carbon dioxide and methane, which represents a profitable product since it can be used for energy production. Anaerobic processes occur without presence of oxygen or nitrate, and are carried out by a varied group of microorganisms which live in a symbiotic relationship. According to several authors (Henze et al., 1997; Kemira Kemwater, 2003; Brinch et al., 1994) the anaerobic degradation process can be divided in four steps, described in Table 1. In the hydrolysis phase, the organic matter present in the form of particles and large dissolved molecules come to a breakdown into small dissolved molecules by the action of extra-cellular enzymes produced by bacteria. The volatile suspended solids (protein, carbohydrates and lipids) are hydrolysed into amino acids and monosaccharides (Brinch et al., 1994). In the acidogenesis phase, complex organic compounds are converted to simpler compounds (among them, volatile fatty acids) by the acid-forming bacteria. The alkalinity is reduced, and the ph decreases due to the formation of fatty acids. As described by Kemira Kemwater (2003), an excess of fatty acids production could lead to a sharp ph drop, which in its turn would lead to the process interruption. 4

15 During the acetogenesis phase, high molecular fatty acids and volatile fatty acids (except for acetate) are converted into acetate (Brinch et al., 1994). In the following methanogenesis phase, the products from the previous phases are consumed, and the alkalinity is slightly increased. The gas formed contains mainly methane and carbon dioxide, and its composition depends on the amount of carbon dioxide dissolved in the liquid phase, but usually around 90-95% of the energy produced from the resulting gas is due to its methane content. Step Type Bacteria Group Substrate(s) Product(s) Hydrolysis Enzymatic Extra-cellular Acidogenesis Acetogenesis Methanogenesis Biological Biological Biological Acid-forming bacteria Acetogenic bacteria Acetoclastic bacteria Particles, large dissolved molecules, carbohydrates, proteins and lipids. Carbohydrates, amino acids, long chain fatty acids. High molecular and volatile fatty acids Amino acids, fatty acids, monosaccharides (sugars), alcohols Butyric acid, propionic acid, acetic acid (VFAs), H 2 Acetate VFA (such as acetate) Methane, CO 2 Methane bacteria H 2, CO 2 Methane Table 1. Anaerobic process steps. Henze et al., (1997) described that the overall anaerobic process occurs at a maximum ph range of 6-8. For a ph below 6, the methane-forming bacteria activity drops rapidly, until reaching ph 5.5, point at which bacteria activity is practically stopped. 5

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17 2 Aim Within this study it will be investigated whether the biological hydrolysis of sludge with consequent withdrawal of the formed hydrolysate has an effect on the methane production from the remaining sludge. Sludge from three different Swedish WWTPs will be used; primary sludge from Klagshamn and Sjölunda WWTPs, and mixed sludge from Källby WWTP. The biological hydrolysis performance will be evaluated by means of quantifying the degree of solubilisation of the organic matter and the production of volatile fatty acids. Further, the suitability of the hydrolysate as a carbon source for biological nitrogen and phosphorous removal will be confirmed through evaluation of the ratio between nutrient content (ammonium and phosphate) and soluble COD in the hydrolysate. 7

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19 3 Literature review Human waste management is a highly focused topic in both environmental and engineering research. There exist many different approaches when it comes to treat wastewater and the consequently produced sludge. The current study will be focused into the following aspects: Importance of the substrate biodegradability for biological nutrient removal from wastewater. Hydrolysis of sludge (primary, return and mixed sludge). Influence of the hydrolysis conditions in VFA production from sludge. Methane potential of sludge and economical balance at WWTPs. 3.1 Importance of the substrate biodegradability for biological nutrient removal from wastewater Organic matter present in wastewater has a complex composition, and consists of biological and non-biological matter, which can be dissolved or suspended, and depending on its degradation rate in biological treatment plants can be classified as biodegradable or inert (so slowly degraded that it can be considered as non-biodegradable). A measure of the quantity of organic matter in wastewater is the COD, which according to Henze et al. (1997) can be divided into the fractions described in Figure 2: C COD = S S + S I + X S + X I Being C COD : Organic matter expressed in COD units. S S : dissolved organic matter (easily biodegradable) S I : dissolved biological inert organic matter X S : suspended organic matter (slowly biodegradable) X I : suspended biological inert organic matter Organic Matter (COD) S S Easily degradable S HA S I Inert X S Slowly degradable X I Inert Dissolved Suspended Figure 2. Fractions of the organic matter in wastewater. Fractions sizes do not correspond to the real proportions present in wastewater, since they can vary significantly. From the two biodegradable fractions, the dissolved part is considered as easily degradable and the suspended part is considered as slowly degradable. In its turn, the dissolved fraction can be subdivided into two other fractions: very easily degradable organic matter (S HAc ), which can be directly taken up by the microorganisms, and easily degradable organic matter (S S ), which must be hydrolysed before it can be absorbed. The fractions can correspond to different proportions according to the influent load, and vary along the wastewater treatment process as different 9

20 biological conversions take place. Henze et al., (1997) described the following model for organic matter conversion in conventional activated sludge plants: 1. The slowly degradable substrate is transformed into easily degradable material by hydrolysis, and is further converted to very easily degradable organic matter. 2. The very easily degradable material is used as substrate for biological growth, and is build up into biomass. 3. After growth, biomass cells death release slowly degradable matter and also some inert material, which is considered non-degradable for the common retention times in biological treatment plants. Several studies have demonstrated the influence by the carbon source type on nutrient removal rates. The biological removal processes that are most dependent on the form in which the carbon source is present in wastewater are denitrification and biological phosphorous removal Importance of substrate biodegradability in nitrogen removal Due to the denitrifying bacteria s heterotrophic constitution, the more easily biodegradable the organic matter, the higher denitrification rate can be achieved (g NO 3 -N/(kg VSS h)). Inert compounds present in wastewater are difficult to dissolve and cannot be taken up easily by bacteria. Another physical magnitude that explains the compounds biodegradability is the ratio surface/volume explained by Vavilin et al. (1996): the lower the ratio, the less space the particles have where its matter can be attacked by enzymes, and therefore its hydrolysis rate is slower than for small particles with higher surface/volume ratio. Some studies demonstrated high suitability of short-chained non-aromatic compounds as carbon source for denitrification. For example, Her et al. (1995) found that using this kind of compounds a total removal of nitrogen could be achieved at lower C/N ratio, meaning less quantity of carbon source used. Improved denitrification rates achieved in several studies by addition of different carbon sources are summarized in Table 2. Two different authors (Xu, 1996; Peng et al., 2006) showed that methanol achieved a lower denitrification rate when compared to ethanol or to acetate. Xu (1996) studied the addition of several different volatile fatty acids, and found that acetate showed a higher denitrification rate than propionate, butyrate and valerate, and an even higher rate was achieved when dosing a mix of VFAs. This phenomenon might be explained due to the fact that denitrifying organisms are composed by a diverse group of bacteria, therefore not all of them prefer the same compound as favourite substrate, as proved by Lee and Welander (1996) in their denitrification experiments using different substrates as carbon source. Elefsiniotis et al. (2004) found a pattern for preferential consumption order of specific VFA compounds during denitrification, being acetic acid the most preferred, followed by both butyric and propionic acid, and finally by valeric acid. Constantin and Fick (1997) tried with both ethanol and acetic acid as C-source for denitrification of high-nitrate concentrated industrial wastewater. In their denitrification experiments they concluded that a higher overall denitrification rate was achieved when using ethanol, but that acetic acid showed a higher denitrification rate when this is referred to the biomass. These results were justified with the fact that acetate can be directly used by the bacteria achieving therefore a higher denitrification rate, meanwhile ethanol favours the growth for microorganisms. 10

21 Norlander (2008) tested primary and return sludge hydrolysates as C-source, and achieved maximum denitrification rates similar to the maximum rate achieved using acetate. Carbon source used Methanol Denitrification rate achieved (mg NO 3- N/(gVSSh)) Maximum: a Apparent: 29 a Acetic Acid Maximum: 76 a Apparent: 48 a Methanol 3.2 Ethanol 9.6 Acetate 12 Methanol 12 Acetate 25.1 Propionate 15.1 Butyrate 21.6 Valerate 20.3 Mixed VFA 31.4 Acetate (21ºC) Biological hydrolysate Thermal/ Chemical hydrolysate Primary Sludge Biological Hydrolysate Acetate Propionate Fermentation products 5 6 (at 20ºC) (at 21ºC) (at 10ºC) 3.8 c 1.7 c 6.0 c Carbon source Nitrogen removal efficiency (N eff) Improvement in nitrate removal Authors 3.7 g COD consumed / g NO 3 --N removed Lee and 4.0 g COD consumed / g NO 3 --N removed 4.8 g COD consumed /g N removed 7.6 g COD consumed /g removed N 33% b 34% b Welander, 1996 Peng et al., 2006 Xu, 1996 Kristensen et al., (1992) Kristensen et al., (1992) Brinch et al., 1994 Moser- Engeler et al., 1998 a Based on g TSS (total suspended solids) instead of g VSS. Substrate enriched with yeast extract. b (Neff with C-source addition Neff Reference)/Neff Reference. Neff: Nitrogen removal efficiency. c Based in g COD total instead of g VSS. Table 2. Denitrification rates and efficiencies achieved by different types of carbon source used Importance of substrate biodegradability in biological phosphorous removal The polyphosphate accumulating organisms (PAOs) use the most easily degradable organic matter (VFAs) under anaerobic conditions, while the other species of bacteria (aerobic or facultative) are under a lag phase due to the lack of oxygen. The presence of VFA in the media is then essential for the growth of PAOs. A good parameter to evaluate the suitability of VFA for biological phosphorous removal is the phosphorous release, often also compared to substrate uptake, P/ S. Moser-Engeler et al. (1998) stated that if the ratio P released /S consumed is high, the PAOs accumulate a large amount of polyphosphate in the aerobic phase, and therefore a large amount of phosphorous can be 11

22 removed from wastewater. According to their experiences with different fatty acids as carbon source for EBPR, they found that acetate and propionate showed higher uptake rates, and it was concluded that PAOs consumed preferentially straight-chained rather than branched-chained fatty acids when exposed to an equimolar VFA mixture. 3.2 Hydrolysis of sludge The term hydrolysis can refer to similar but different concepts. The one most used in wastewater treatment refers to the breakdown of slowly degradable organic matter into more easily degradable compounds, similarly as Gujer et al. (1999) defined. This will be the meaning used within the current work. Sludge hydrolysis has lately gained interest because of a perpetually increasing price of external carbon sources. Several methods have been developed to hydrolyse the sludge: biological (anaerobic fermentation), enzymatic, thermal (breakdown of organic matter at ºC) or chemical (adding alkali, i.e. sodium hydroxide or lime; or adding acid, i.e. sulphuric acid or nitric acid). The main advantages offered by biological hydrolysis are the high fraction of readily biodegradable COD achieved and the simplicity of the process design. Æsøy and Ødegaard (1994) compared biological hydrolysis with thermal treatment and concluded that thermal treatment gives a higher degree of solubilisation of organic matter, but the fraction of readily biodegradable COD in the hydrolysate is smaller. Kristensen and Jørgensen (1992) studied the differences achieved in the hydrolysis of sludge by different hydrolysis methods: biological and thermal/chemical. They found that the dissolved COD in the biological hydrolysate consisted mainly of easily degradable COD; while the COD in the thermal/chemical hydrolysate was composed for a small fraction (around 20%) of easily degradable COD. These figures mean that with biological hydrolysate high denitrification rates similar to the rates obtained with acetate could be achieved; meanwhile the rates achieved with thermal/chemical hydrolysate were about half the values reached by acetate. The biological sludge hydrolysate mainly consists of acetic and propionic acids and it can represent up to the 60-80% of the soluble COD after hydrolysis (Henze et al., 1997). During hydrolysis, the degree of solubilisation of the COD increases gradually, due to the conversion of the long chained organic compounds into short chained ones. As the hydrolysis time increases, the dissolved COD tends to reach a stable concentration, and even starts decreasing. This decline in soluble COD production indicates the beginning of the methanogenesis phase (Hatziconstantinou et al., 1996). The sludge also experiences changes in its solids concentration. In a study carried out by Ferreiro et al. (2003), the VSS degradation during the hydrolysis was studied at different temperatures and initial sludge concentrations. The percentage of VSS degraded was around 40% (at T=20ºC), for the different initial VSS concentrations tested and with hydrolysis durations between days, meaning that the organic matter was partly solubilised during the hydrolysis. A lot of research has been previously done with different types of sludge, showing different results in COD-yield and nutrient removal efficiencies by the produced hydrolysate. Some studies (Ji et al., 2010) showed that both primary and mixed (primary and activated) sludge can be efficiently used for biological nutrient removal. 12

23 3.2.1 Primary Sludge Hydrolysis Primary sedimentation is a common process in wastewater treatment plants. The settling and further removal of a significant fraction of the organic matter and solids present in the incoming wastewater leads to a reduction of the subsequent biological treatment requirements, in this way improving the overall optimization of the organic carbon removal of the treatment plant (Hatziconstantinou et al., 1996). Sludge pre-treatment plays also an important role in the COD-yield obtained. Jönsson et al. (2008) found that the COD-yield achieved by sludge without pre-treatment was 19%, while chemically precipitated sludge only gave a COD-yield of 7% (at 10ºC) and 16% (at 20ºC). Even though the COD-yield appeared to be lower for the chemically precipitated sludge, the fraction of the soluble COD corresponding to VFA was substantially higher; reaching a value of 60% at 20ºC, meanwhile a value of only 43% was achieved for the sludge without chemical precipitation. These results were consistent with the values achieved by Kristensen et al. (1992), where the ratio mg VFA/mg COD was also around 60% in their hydrolysis experiments with chemically precipitated primary sludge; and with the results achieved by Daton and Wallergård (2003), who found a COD-yield of 20-24% at 20ºC (cited by Jönsson and Jansen, 2006). There are basically two ways to arrange primary sludge hydrolysis in WWTPs (see Figure 3): In-line alternative: primary sludge is kept in the primary clarifiers for a longer time than usual. Anaerobic conditions will occur in the sludge layer in the bottom of the basin, and the hydrolysis will take place. Sludge is lifted from the bottom of the basin to just below the surface by a pump, in order to wash the formed soluble COD out of the sludge. In order to avoid re-suspended solids in the outlet of the clarifier, the sludge lifting is done in the zone close to the inlet. Off-line alternative: primary sludge is pumped to a separate hydrolysis tank. The formed hydrolysate (containing soluble COD) is separated from the sludge, and can be dosed back in the biological process when and where it is needed. Primary Clarifier Sludge Anaerobic conditions Primary Clarifier Hydrolysis unit Hydrolysate In-line alternative Off-line alternative Figure 3. Arrangement alternatives for primary sludge hydrolysis. Separation unit A further example of how the hydrolysis of primary sludge can be arranged in WWTPs is the socalled HYPRO process. The HYPRO-concept uses the biologically hydrolysed sludge as carbon source in the denitrification process. The process uses a compact design for nutrient removal from wastewater, consisting on pre-precipitation, primary sludge hydrolysis and biological nitrogen removal. The idea was developed as a common project between research institutions and companies from Denmark, Sweden and Norway and it has been described by several authors, 13

24 among them Henze and Harremoës (1990) and Æsøy and Ødegaard (1994). During biological hydrolysis, a release of ammonium is commonly noticed. Rustrian et al. (1998) related the ammonium release observed during anaerobic digestion to the lysis of acidogenic cells. An amount of phosphorous is also released due to the hydrolysis process itself. It is favourable to have a low presence of both nutrients in the obtained hydrolysate, since an extra addition of both nutrients would decrease the positive effect of the hydolysate addition in nutrient removal processes. In the results from the primary sludge experiments carried out by Jönsson et al. (2008), the solubilised nitrogen in the form of ammonium followed a linear relationship of mg NH 4 -N / mg COD solubilised, meaning a COD/N ratio equal to 29. For phosphorous, there was a difference between sludge chemically precipitated or not. The soluble phosphate concentration in the hydrolysate from pre-precipitated sludge was negligible, while the solubilised phosphorous in the hydrolysate from sludge without chemical precipitation was up to mg PO 4 -P f / mg COD solubilised. Primary sludge has particular interest for its suitability for anaerobic digestion, including both hydrolysis and anaerobic digestion for methane production. Primary sludge is generally considered to have higher anaerobic degradability than waste activated sludge (Parkin & Owen, 1986; cited by Davidsson, 2007). Henze et al. (1997) stated that the maximum COD yield (soluble COD per total COD) can be up to 10-20% for primary sludge, while for activated sludge can be down to 2-6%. It has also been demonstrated that denitrification rates achieved (in the N-removal process) using carbon from primary sludge hydrolysate are higher (about 30%) compared to rates achieved using carbon from settled sewage (Hatziconstantinou et al., 1996). Primary sludge is one of the best types of sludge for methane production, since its organic matter content has not still been degraded (unlike return sludge). There exists the common conception that hydrolyzing sludge for production of internal carbon source must lead to a decrease in the gas production in the WWTP. This fact generates a conflict for economical balance of WWTPs with gas production if there is the possibility to utilize primary sludge for carbon source production via hydrolysis, since methane gas is one of the few profitable products obtained from the process. Recently, a study indicated that primary sludge hydrolysis with withdrawal of the hydrolysed fraction containing the solubilised carbon source not necessarily must lead to a significantly lower methane production (Jönsson et al., 2009) Return and Mixed Sludge Hydrolysis Waste activated sludge represents a fraction of the waste generated in WWTPs, for what its further treatment is needed, and hence its use for hydrolysis. Return sludge is richer in active biomass than the primary sludge, thus higher solubilisation rates could be achieved. For this reason, mixing both primary sludge (higher biodegradability) and return sludge (higher fraction of bacteria) could lead to a better hydrolysis efficiency, which has motivated several studies, e.g. the one carried out by Moser-Engeler et al. (1998), in which hydrolysis was performed with mixed sludge with a 60:40 ratio between primary and waste activated sludge, respectively. The process itself consists in adding a hydrolysis unit in the return sludge stream. As illustrated in Figure 4, either the total return sludge flow or a just a part of it can be hydrolysed. When only a 14

25 part of the return sludge flow is hydrolysed and re-circulated into the process the arrangement is known as Side-Stream Hydrolysis (SSH). Activated sludge process Activated sludge process Hydrolysis Unit Hydrolysis Unit Figure 4. Hydrolysis arrangement possibilities for return sludge. When hydrolysing the activated sludge to produce internal carbon source there is no need to separate the produced hydrolysate from the remaining sludge after the hydrolysis reaction. This is the main advantage in opposition to primary sludge hydrolysis, in which a substantial part of the hydrolysate remains in the sludge after phase separation (about 25%) (Jönsson and Jansen, 2006). Activated sludge has shown low tendency to biodegradation and low VFA yields (Ucisik and Henze, 2008). Andreasen et al. (1997) found a COD yield of 2.5% (in terms of soluble COD related to total COD), for activated sludge fermentation at ambient temperature 8-17ºC. Comparing this yield to the 9-16% yield achieved in primary sludge fermentation at 20ºC, it can be considered that activated sludge has less potential for easily organic degradable matter production, even though the yield for the activated sludge could be increased by increasing the fermentation temperature. The sludge amount used in order to achieve a certain VFA production is also important to consider: if larger quantities of return sludge are used, the total produced VFA can be the same as the amount produced from primary sludge, but consequently equipment with larger volumes is needed. Jönsson and Jansen (2006) found maximum COD yields of %, and VFA yields expressed as fraction of soluble COD of 25% (for long sludge age) and 50% (for short sludge age), for hydrolysis of return sludge without pre-sedimentation. The longer the sludge age, the higher the degree of mineralisation, and therefore lower the production of readily degradable organic matter. 3.3 Influence of hydrolysis conditions in VFA production from sludge Several studies have been carried out to find the optimal external factors in order to achieve the highest VFA production. A determining aspect is the hydrolysis method. Kristensen et al. (1992) analysed the difference between biological and thermal/chemical hydrolysate from activated sludge, and found that fraction of soluble VFA for the biological hydrolysate was 67% of its COD, while this fraction was only 20% for thermal/chemical hydrolysate. Brinch et al. (1994) stated that the optimal conditions for soluble carbon production in sludge hydrolysis are a residence time of 2-3 days and a process temperature of 25ºC. Under these conditions, the ph ranges between 5.3 and 6.0, and a dissolved COD yield of 10-15% in relation to the total COD was obtained. Around 80-95% of the dissolved COD consisted of VFA, being around 50-60% of this acetic acid. 15

26 3.3.1 Temperature and ph The anaerobic processes are temperature-dependent, since different biomass specific substrate removal rates have been observed at different temperatures. Three ranges can be defined: Cryophilic, under 25ºC; mesophilic, between 25-45ºC; and thermophilic, between 45-65ºC (Henze et al., 1997). The higher reaction rates (in terms of substrate removal per amount of biomass) occur at the thermophilic range. For this reason, in many cases performing the process at high temperatures can speed up the reaction rates, meaning that large tank volumes can be saved, so therefore hydrolysis might become economically profitable. Several studies have demonstrated that the yield of soluble COD is lower at lower temperature than at higher temperature. Jönsson and Jansen (2006) found a higher soluble COD yield at 20ºC than at 10ºC, in hydrolysis experiments with return sludge. In their studies with primary sludge hydrolysis, Karlsson and Smith (1990) achieved a soluble COD production of 20% for a HRT=1 d and T=37ºC (cited by Hatziconstantinou et al., 1996), whereas the soluble COD production found by Hatziconstantinou et al. (1996) was only of 10-15% for lower temperature, even with higher hydraulic retention times (HRT=2-3 days and a T=25ºC). Kristensen et al. (1992) reported also a higher soluble COD yield for a higher operation temperature (within the same HRT). Some results for studies involving temperature and hydraulic retention time are summarized in Table 3. In the study with primary sludge carried out by Ferreiro et al. (2003), the VFA yield found at 20ºC resulted to be mainly the same as the yield obtained at 35ºC, while at 10ºC the VFA produced was about 20% less. Additionally, they found that the quality (in terms of composition) of the readily biodegradable matter generated was not influenced by temperature. Moser-Engeler et al. (1998) obtained a first-order fermentation process with respect to particular fermentable components, and quantified the hydrolysis constant which turned out to be three times higher at 20ºC than at 10ºC. T= 10-15ºC T=20-25ºC T=30ºC T=37ºC HRT= 1 day 2% (Kristensen et al., 1992) 10% (Kristensen et al., 1992) 20% (Karlsson and Smith, 1990) HRT=2-3 days 8% (Kristensen et al., 1992) 5 6% (Hatziconstantinou et al., 1996) 10 15% (Brinch et al., 1994) 10-13% (Kristensen et al., 1992) Table 3. COD yield achieved by several authors, depending on temperature and HRT. The ph influence has found to be negligible within a range of 4.3 and 7.0 for VFAs production (Elefsiniotis and Oldham, 1991, cited by Ucisik and Henze, 2008). In the laboratory-scale experiments carried out by Moser-Engeler et al. (1998), it was found that a controlled ph=7 increased the hydrolysis (in terms of COD yield) during the first 10 days Hydrolysis duration, sludge retention time and hydraulic retention 16

27 time Sludge retention time and hydraulic retention time are two important parameters when talking about hydrolysis in semi-continuous or continuous processes. In batch hydrolysis experiments, where SRT and HRT do not have a realistic meaning, the hydrolysis duration (or fermentation period length) is a commonly used concept. Examples of COD yields in batch experiments achieved by several authors are summarized in Table 4. Temperature (ºC) 20 < 20 Hydrolysis Percentage of the duration (days) maximum COD yield (%) 5 60% (17% of COD total) 10 78% (22% of COD total) % (28% of COD total) Maximum soluble COD yield 28 % of COD total % mg % solubilised COD/ g VSS initial Authors Moser- Engeler (1998) Hatziconstant i-nou (1996) % 270 mg COD f/ g VSS Pottier (2006) Table 4. COD yields depending on the hydrolysis time in several batch experiments with primary sludge. Skalsky and Daigger (1995) experimented with sludge hydrolysis with SRT values from 2 to 6 days and at a temperature between 17-23ºC. They obtained the highest VFA production in units of mg VFA/mg VS feed for a SRT equal to 5 days. In addition, they stated that almost the maximum VFA production could be obtained at a SRT as low as 2 days. These results are consistent with the results achieved by Moser-Engeler et al. (1998), who carried out batch fermentation experiments during 16 days under different conditions: for the experiments at 20ºC, they found out that 17% of the total COD was produced as soluble carbon during the 5 first days; while this fraction just increased another 5% during the next 5 more days, being the maximum solubilised COD fraction around 28% of the total COD. Furthermore, they stated that the fermentation duration did not influence the VFA composition in the hydrolysate, and concluded that longer fermentation periods were not efficient. Hatziconstantinou et al. (1996) observed in their batch hydrolysis experiments that soluble COD production reached its maximum value within a period of 5 to 9 days, and during the first days 50-60% of the maximum quantity had already been achieved. Jönsson and Jansen (2006) found the higher release during the 6 th hydrolysis day in their experiments with return sludge, and in the primary sludge hydrolysis experiments done by Pottier (2006) the release reached the maximum between the 4 th and 5 th day Sludge Concentration Sludge concentration influence in anaerobic digestion has been studied by several authors. For example, Bouzas et al. (2002) (cited by Ucisik and Henze, 2008) found that a higher VFA production and more soluble COD was achieved with a higher solids concentration. Even though, if VFA production is analysed by means of efficiency, Skalsky and Daigger (1995) found a higher VFA production per VS in the feed at lower solids concentration. In their fermentation experiments of primary sludge at different initial solids concentration, the values of 0.22, 0.20, 0.16 and 0.12 mg VFA/mg VS feed were achieved for the experiments with an initial solids concentration of 0.43%, 0.87%, 1.30% and 2.60%, respectively (see Table 5). They stated that the 17

28 differences observed could be explained by several possible factors: a better biomass mixing could be achieved easily at more diluted solids concentration, which may enhance hydrolysis of particulate VS into soluble substrate as well as enhance substrate uptake. Another possibility would be that the inhibitory effect of fermentation products might be reduced at diluted concentrations. Solids concentration (%) VFA production (mg VFA/mg VS feed) Table 5. VFA production depending on solids concentration found by Skalsky and Daigger (1995). 3.4 Methane potential of sludge and economical balance at WWTPs Economical impact must always be evaluated when modifying a full-scale process, especially when it is necessary to incorporate new equipment. When discussing about primary sludge hydrolysis for production of easily degradable carbon source, there are several ways in which the process affects the whole WWTP economy: Internal production of carbon source suitable for nutrient removal leads to a decrease in the amount of external carbon source needed, and therefore the operating costs due to external product purchasing are substantially lowered. Moreover, the addition of external carbon source leads to an increase of sludge production and oxygen demand in the biological step, which might also represent an increased cost. Installation of new equipment in order to perform the sludge hydrolysis represents an investment cost which always has to be evaluated. Sludge hydrolysis and the subsequent withdrawal of its hydrolysate for its use in nutrient removal processes might lead to a decrease in its methane potential. In that case, sludge hydrolysis would lead indirectly to a loss in the economical profit generated by methane production. When considering the investment and operational costs related to the acquisition of a hydrolysis unit, the incoming wastewater characteristics play an important role. Canziani et al. (1995) stated that when the influent nutrient concentrations show important fluctuations, the operational cost of using a hydrolysis unit and its corresponding separation device could be higher than the cost of purchasing an external carbon source. When comparing the costs, a variable factor is the price of the external carbon source (such as methanol or ethanol); therefore the investment has to be specifically evaluated in every case. Methane production from organic waste has lately been a highly focused topic, since it might lead to a substantial economical profit. Both primary sludge and waste activated sludge are suitable for methane production via anaerobic digestion due to their high VS content (Davidsson, 2007). Even though, activated sludge has lower anaerobic degradability (Parkin and Owen, 1986, cited by Davidsson, 2007). A long retention time in the activated sludge process increases the mineralization degree of the sludge (increase of the non-biodegradable fraction and refractory 18

29 compounds) and as stated by Bolzonella et al. (2005) (cited by Davidsson, 2007) consequently leads to a decrease in the amount of gas produced. This differences in biodegradability can be noticed in the COD and VS reduction: for waste activated sludge, the COD and VS reduction usually do not exceed the 50% under mesophilic conditions, even if long retention times are used; meanwhile for primary sludge the typical degradation values are 40-60% of reduction for COD, and 40-70% of reduction for VS (Parkin and Owen, 1986, cited by Davidsson, 2007). 19

30 20

31 4 Materials and methods The experimental research consists of two different parts: first the biological hydrolysis of the sludge (under anaerobic conditions), and the succeeding production of methane through anaerobic digestion. Every experiment was carried out with sludge from three Swedish WWTPs: Klagshamn, Sjölunda and Källby. Sludge samples were taken from the bottom of the primary clarifiers, since the study is focused on primary sludge hydrolysis. Klagshamn and Sjölunda have a similar process configuration, and the sludge samples taken from the primary clarifiers are considered to be pure primary sludge. Sludge samples from Källby cannot be considered pure primary sludge since biological sludge and chemical sludge are re-circulated and added into the process before the primary clarifier, therefore it will be mentioned as mixed sludge (see Figure 5). Klagshamn and Sjölunda WWTPs FeCl3 Activated Sludge Process C-source (Methanol or Ethanol) Incoming wastewater Screening Solids > 3 mm Sand Trap Separated Sand Primary Clarifier Primary sludge Activated Sludge Basin Return Biologic Sludge Sedimentation Basin Waste activated Sludge Trickling filters (Only Sjölunda) Denitrification process Methane Production Källby WWTP Activated Sludge Process FeCl3 Incoming wastewater Screening Sand Trap Primary Clarifier Activated Sludge Basin Sedimentation Basin Chemical Treatment Solids > 3 mm Separated Sand Return Biologic Sludge Waste activated Sludge Chemical sludge Primary + Biological + Chemical sludge Methane Production Figure 5. Simplified principal process flow sheet for Klagshamn, Sjölunda and Källby (VASYD, 2009) Experiment description Primary sludge samples were kept in 2 L beakers arranged in duplicates (except for sludge from Källby during experiment A), which will be mentioned as hydrolysis reactors. For both experiments, in one of the two duplicate reactors the hydrolysis lasted for 4 days, while in the other duplicate the hydrolysis was prolonged until 10 days. The sludge hydrolysed during 4 days was further on used for methane potential test. The aim of the duplicates is to evaluate the hydrolysis evolution after stopping the hydrolysis process at the fourth day. An overall diagram of the experiment design is illustrated in Figure 6. 21

32 Original sludge from one Diluted sludge (if needed) Fresh sludge Reactor (4 days) Methane Potential test Hydrolysed sludge Remaining Hydrolysed Sludge Hydrolysis Reactor (10 days) Hydrolysed sludge Hydrolysate Figure 6. Diagram of the experimental process, which was carried out for the sludge from every WWTP. The experimental process was carried out in two times: experiment A (hydrolysis carried out between 1 st -10 th February 2010, methane potential tests carried out between 1 st February 23 rd March 2010); and experiment B (hydrolysis carried out between 19 th -29 th April 2010, methane potential tests carried out between 19 th March 31 st May The following denomination will be used to refer to the sludge used in the different parts of the experiment: Original sludge: the sludge as directly brought from the WWTP, without any dilution applied. Original diluted sludge: the original sludge once the desired dilution is made. Fresh sludge: sludge without hydrolysis treatment, used for the methane potential test. Hydrolysed sludge: sludge treated with biological hydrolysis. When describing the methane potential tests, it will be mainly referred to the sludge hydrolysed during 4 days and after separation and withdrawal of its hydrolysate. The sludge distribution in the reactors for both experiments is indicated in Table 6. Reactor Reference Experiment Sludge origin Type of sludge 22 Hydrolysis duration (days) KLH4 A A Klagshamn Primary sludge 4 SJÖ4 A A Sjölunda Primary sludge 4 KÄLL4 A A Källby Primary sludge + biological and chemical sludge 4 KLH10 A A Klagshamn Primary sludge 10 SJÖ10 A A Sjölunda Primary sludge 10 KLH4 B B Klagshamn Primary sludge 4 SJÖ4 B B Sjölunda Primary sludge 4 KÄLL4 B B Källby Primary sludge + biological and chemical sludge 4 KLH10 B B Klagshamn Primary sludge 10 SJÖ10 B B Sjölunda Primary sludge 10 KÄLL10 B B Källby Primary sludge + biological and chemical sludge 10 Table 6. Sludge distribution in the reactors. Solids concentration was analysed before starting up the experiments. In order to keep similar conditions in all the different reactors, the sludges that showed higher solids concentration were diluted with tap water. Taking into account that the suspended solids content (in grams) must be

33 the same before and after the dilution, the calculation described in Equation 1 was done: Dilution Factor = SS initial g l = V desired l SS desired g l V initial l [Equation 1] For experiment A, only the sludge from Sjölunda was diluted. The dilution factor was calculated in order to have a SS value around 22 g/l, an intermediate value between the concentrations for the original sludges from Klagshamn and Källby. For experiment B, and after experiencing practical problems in experiment A due to the high solids concentration, the sludge from both Sjölunda and Källby were diluted in order to have a SS value around 10 g/l, similar to the concentration for the original sludge from Klagshamn. The dilutions made in both experiments are showed in Table 7. Experiment Sludge origin Used in reactors SS initial (g/l) SS desired (g/l) Dilution Factor V initial (l) A Klagshamn KLH4 A, KLH10 A 24.1 No dilution A Sjölunda SJÖ4 A, SJÖ10 A A Källby KÄLL4 A 20.6 No dilution B Klagshamn KLH4 B, KLH10 B 8.6 No dilution B Sjölunda SJÖ4 B, SJÖ10 B B Källby KÄLL4 B, KÄLL10 B V desired (l) Table 7. Suspended solids concentration in the original sludge from different WWTPs and the dilution applied in every reactor. Hydrolysis conditions In experiment A, hydrolysis was performed at room temperature (temperature in the reactors: average 19.3ºC, minimum 15.0ºC, maximum 21.7ºC). In experiment B, temperature was controlled through water bath, with the water temperature set at 20ºC (temperature in the reactors: average 19.7ºC, minimum 18.6ºC maximum 20.4ºC). Due to the equipment dimensions, three of the reactors (KLH4 B, SJÖ4 B and KÄLL4 B) were installed in one water bath, and the other three (KLH10 B, SJÖ10 B and KÄLL10 B) in another water bath. The experiment B set-up can be seen in Figure 7 and Figure 8. In both experiments hydrolysis was performed under anaerobic conditions, which were achieved by covering the beakers with a lid and introducing nitrogen gas over the sludge surface. Stirring was kept at the lowest possible rate, around rpm. For experiment B, some arrangements were made in order to avoid evaporation in the reactors, which turned out to have an important effect during experiment A: Nitrogen flow was kept as low as possible, and as similar as possible in all the reactors. Sludge level was controlled by marking the sludge level every day after taking out sample. The level was checked next day. If it had decreased due to evaporation, tap water was added until reaching the level marked the previous day. This measure was applied from the second hydrolysis day, since no level decreasing could be seen in the first day. To avoid practical problems observed during experiment A due to the high sludge concentration (difficult phase separation, hydrolysate turbidity, difficult vacuum filtration for SS measurements), and also in order to keep similar conditions in the reactors, in experiment B the thicker sludges 23

34 (from Sjölunda and Källby) were diluted down to an approximate value of 10 g/l. Stirring Water bath Nitrogen gas Level mark Figure 7. Set-up of reactors KLH4 B, SJÖ4 B and KÄLL4 B. Figure 8. Set-up of reactors KLH10 B, SJÖ10 B, and KÄLL10 B. Sampling procedure Samples were taken out daily for dissolved substances analysis, and every 1-3 days for SS and VSS measurements. TS and VSS were measured only at the beginning and at the end of the hydrolysis in order to avoid disturbing the reactor by extraction of a substantial amount of sample needed for the measurement (around 90 ml). The sample volume was about ml when suspended solids measurements were done, otherwise just around ml were necessary. Control measurements for ph and temperature were done every day. The measurement planning is shown in Appendix 1.3, Table 27. Daily analysis procedure for each reactor 1. Check nitrogen gas flow into the beakers. 2. Take measurements of ph and temperature. 3. Take out ml of sample (depending on which measurements needed to be done). 4. Use sample directly for solids concentration: 30 g for each TS measurement and 5 ml for each SS measurement. From every sludge sample triplicate measurements were taken. 5. For COD total: Dilute about 50 times unfiltered sample for COD measurement with Dr Lange. From every sludge sample a single measurement was done. 6. For filtered substances (COD f, NH 4 -N f, PO 4 -P f and VFA f ). - Fill in eppendorfs with approximately 15 ml of sludge. - Centrifuge during 5-10 minutes at around 8000 rpm, until two phases are visibly separated. - Filter the supernatant with Munktell filters. - Dilute the times needed in order to fit the measurement in the measuring range for every type of Dr Lange test. From every sample a single measurement was done. 24

35 4.1.2 Description of methods used The description of methods used is done after Jönsson and Jansen (2006), with author s permission. Temperature and ph: integrated ph and Temperature sensor, WTW ph 320. Filtration: Separation of phases by centrifugation at 8500 rpm, followed by filtration with Munktell cellulose filters No. 1002, filtrating rate of 250 ml/min. Measurement of dissolved substances concentration: COD (filtered and total) NH 4 -N f, and PO 4 -P f ) by Dr Lange tests LCK 114, LCK 303 and LCK 349, respectively. Analysis measures were performed using a DR 2800 (Lange ). Dilution of the supernatant was needed in different proportions depending on the test and the current concentration. Solids Concentration: Following the Swedish standard SS-EN 872 (for SS and VSS) and SS (for TS and VS). Suspended solids: Whatman, filter of 55 mm diameter and 0.45 µm pore size. Dry matter at least one hour at 105ºC. Volatile suspended solids: After suspended solids measurement, ignition during 1 hour at 550ºC. Total solids: Aluminium plates. Dry matter at least 6 hours at 105ºC. Volatile solids: After total solids measurement, ignition during 1 hour at 550ºC. Volatile Fatty Acids: keep 0.9 ml of filtered sample with 0.1 ml of 10% phosphoric acid at cool temperature in 1.5 ml vials. Acetate and propionate analysis were done with a gas chromatograph (Agilent, 6850 Series GC System, equipped with a FID and column (30 m, 530 µm, 0.25 µm). The VFA concentrations are calculated as average of two measurements, and are expressed in units of COD (mg/l), for that reason they will be referred as COD VFA (mg/l). Calibration curves are shown in Appendix 1.2. Figure 9. VFA analysis with gas chromatograph (left). Dr Lange cuvettes LCK 349 (upper right). Vials for VFA analysis with sample (lower right). 25

36 4.2 Methane potential tests In order to evaluate the suitability of a certain organic waste for methane production, several methods have been developed. The principle usually consists in incubating a certain amount of organic waste together with anaerobic inoculum, and measuring the composition and volume of the gas generated. To evaluate the impact of primary-sludge hydrolysis in the further methane production, two sets of methane production tests were carried out in both experiments A and B. In the first sets fresh sludge was used, while in the second sets the sludge used had been hydrolysed during 4 days and separated from the supernatantt (which would be used as carbon source in BNR processes). Fresh sludge was diluted in the same proportions than in the hydrolysis experiment, in order to keep similar conditions in both fresh and hydrolysed sludge sets. In WWTPs, the hydrolysate is separated from the hydrolysed sludge, and transported to the BNR process to be entirely used as feed of internal carbon source. In order to keep the laboratory hydrolysis process realistic and as similar as it would be in a full-scale process, the whole sludge hydrolysate (supernatant) was discarded, and subsequently replaced by the same volume of tap water in order to maintain the sludge concentration. Centrifugation (during approximately 20 min at 3000 rpm) was the technique used for phase separation. The dark color of the separated hydrolysate indicated that it contained a substantial amount of organic matter (see Figure 10) Figure 10. Replacement of hydrolysate with water, after phase separation by centrifugation. Remaining sludge (left), separated hydrolysate (center) and tap water (right) Method description The methane potential test was performed according to the batch method used by Hansen et al. (2003), adjusted for rather non-homogeneous material (as is the case for primary sludge), according to the procedure described by Davidsson, The method basically consists in performing anaerobic digestion of the sludge using inoculum of methanogenic bacteria in sealed bottles, and measuring the methane produced periodically. 26

37 The inoculum The inoculum used in both sets of bottles was coming from the mesophilic phase of the anaerobic digesters in Sjölunda WWTP. Inoculum sample was brought from the WWTP 3-4 days before the start of every methane potential test, and kept in an incubator at 34.5ºC until the bottles set-up, with the purpose of degasification. Every set of bottles was started up at different times: the set using fresh sludge was stared up in parallel with the hydrolysis experiment, and the set containing hydrolysed sludge was stared up after 4 days of hydrolysis. For this reason, a different sample of inoculum was used for every set, and it cannot be considered to have exactly the same properties in all the sets. The bottles (Figure 11) The sludge was digested in 2 L glass bottles, which were kept in the incubator at 34.5ºC (mesophilic temperature). The bottles were sealed with a rubber septum and a screw cap. Anaerobic conditions were kept from the beginning of the digestion by introducing nitrogen gas right before sealing the bottles, once they had been filled up with the material to be digested. Since the methane production can vary significantly due to the sludge non-homogeneity, the bottles were arranged in triplicates. Every set of bottles consisted in: - 3 bottles containing primary sludge from Klagshamn (fresh or hydrolysed) and inoculum. - 3 bottles containing primary sludge from Sjölunda (fresh or hydrolysed) and inoculum. - 3 bottles containing mixed sludge from Källby (fresh or hydrolysed) and inoculum. - 3 bottles containing only inoculum, with the function of blank, in order to be able to calculate the real methane potential of the sludge only. - 3 bottles containing cellulose and inoculum. They are used as a reference, in order to verify the ability of the inoculum to digest an added substrate. Cellulose is chosen because of its slow digestion and as it has a methane potential similar to sludge (Davidsson, 2007). The triplicate bottles used for the methane production were filled up with the same amounts of material, meaning that in each group of triplicates containing hydrolysed sludge, the sludge used was taken from the same hydrolysis reactor. Figure 11. Bottles kept in the incubator at 34.5 ºC during a methane potential test. Figure 12. Injection of methane samples in the gas chromatograph with a gas syringe. 27

38 Measurements and equipment The equipment used for methane measurements was a gas chromatograph (Agilent 6850) equipped with a flame ionisation detector (FID) and a 30m, 0.32 mm, 0.25 µm column. The samples were extracted through the septum with a 0.5 ml gas syringe, Series A-2, 250 ul RN(VICI) Pressure-Lok (see Figure 12). The elastic properties of the septum makes repeated measurements of the methane content possible, since the small hole created during every sampling closes again due to the pressure inside the bottle. Every set of bottles is measured 2 times per week during the first 3 weeks, and only once a week in the following weeks, until no more methane is produced from the sludge. The digestion period required in order to determine the total methane potential can vary from 30 to 50 days (Davidsson, 2007). The lengths of the digestion period for every set of bottles for experiment A and B are shown in Table 8. Experiment A Experiment B Fresh sludge set Hydrolysed sludge set Table 8. Total digestion time (days) in every set of bottles of the methane potential tests. During a measurement session, every bottle is measured 3 times (each measurement corresponds to a different extracted sample), and the injected volume of every sample is 0.2 ml. The methane content value used for the methane production calculations is the average of the 3 measurements, and the calculations are related to the measurement values obtained from 100% methane obtained at the beginning of the measuring session. In every measuring session the following values are also noted: sampling time, room temperature, incubator temperature and atmospheric pressure. The methane produced increases the pressure inside the bottles, which can be noticed by the tension in the septum. When the pressure inside the bottles is around 1.5 times the atmospheric pressure, the methane must be released. A second measurement needs to be done after releasing the gas. This difference in methane content is used to estimate the amount of gas released, and taken into account in the methane production calculations. As it is not possible to know the exact quantity of methane released, in order to introduce less uncertainty in the calculated methane production the release does not need to be done if the septum does not look considerably tight. It is considered that a bottle has 1.5 times the atmospheric value when the methane concentration in the bottle reaches 15%, or when the measurement value is 1.4 times the last releasing value Set-up volumes In order to adjust the method developed by Hansen et al. (2003) for sludge samples, the following restrictions described by Davidsson (2007) are taken into account: - The total material volume is the same for all the bottles, and equal to 600 ml. - A proportion of 60:40 has to be kept between the VS from the inoculum, and the VS from the sludge, respectively. - In order to ensure the proportions, the volume of inoculum used would vary for every 28

39 type of sludge, since the VS content in them is different. As it is preferred to use the same volume of inoculums in all bottles in a set, the volumes are calculated as follows: the inoculum volume is calculated according to the sludge with least VS concentration, and this inoculum volume will be used for the rest of bottles (independently of their VS concentration), therefore the volume for the other sludges will be calculated according to the inoculum volume. - When needed, water is used in order to fill the bottles up to 600 ml and keep the 60:40 proportion between VS from inoculum and sludge, respectively. In order to calculate the volumes to use, TS and SS was measured just before starting the experiment. Since VS and VSS values are more or less equivalent for primary sludge, for practical reasons and due to the fact that the VS procedure takes several hours, in some occasions VSS values instead of VS were used for the calculations of the set-up volumes of inoculum and sludge used in the bottles. For experiment A, the VSS values from the different sludges could not be obtained in time before setting up the first set of bottles (fresh sludge), so the VSS value used was an approximation corresponding to 75% of the SS value. For the set of bottles containing hydrolysed sludge, the same volumes as for the first set were used, in order to keep the same conditions in both sets. For experiment B, all VSS values were taken before setting up the experiment, and the volumes were calculated according to them. At this occasion, the volumes to use were calculated for each set (the one containing fresh and the one containing hydrolysed sludge) according to the current VSS values of sludge (fresh or hydrolysed) and inoculum. The expressions used for the calculations and the calculated volumes for both A and B experiments can be found in Appendix Calculation of the methane production In order to be comparable with previous studies, methane production is expressed per gram VS at standard temperature and pressure conditions (STP) according to Equation 2 (Hansen et al., 2004): Where: X STP = X m T Standard P m T m P Standard [Equation 2] X STP is the methane measurement value at standard temperature and pressure (area). X m is the obtained methane measurement, at actual temperature and pressure (area). T m is the actual temperature of the incubator (K), since it is considered that the temperature of the bottle content is at the same as the incubator temperature during the measurements. P m is the actual atmospheric pressure (hpa). T standard is equal to K. P standard is equal to hpa. The methane measurement is the area of the methane signal given by the gas chromatograph. To 29

40 convert the measured area to volume of methane, the areas of the measurements given by 100% methane and the volume available in the upper part of the bottle (headspace) are used according to Equation 3. The produced volume of methane is calculated using the average of three measurements. Where: V produced = X STP V headspace X 100% V produced is the volume of methane produced in standard conditions (ml). [Equation 3] V headspace is the actual volume of the produced methane in the bottles, being the free space, subtracting to the total bottle volume the volume of the material being digested (ml) X 100% is the average value of three measurements of 100% methane (area). X STP is the average of three measurements at standard temperature and pressure (area) As bottles must be released when the interior pressure is too high, the methane production is calculated in reference to the last methane measurement (regardless whether the bottle has been emptied or not). By adding the methane produced between two measurements, the accumulated methane production is obtained. To know the real methane potential of the sludge only, the amount of methane produced in the bottles containing only inoculum (average of the three bottles) is subtracted from the amount of methane produced by the bottles also containing material (either sludge or cellulose). Once the total accumulated methane production from the material for the triplicate bottles containing sludge from the different WWTPs is obtained, the results are analysed statistically. First, the confidence interval for the three bottle measurements is calculated according to the expressions detailed in Appendix 1.1. A second statistical test is used in order to verify whether the values achieved for fresh and hydrolysed sludge can be considered equivalent, and thereafter conclude if the hydrolysis of the sludge affected its methane potential. This statistic test is known as unpaired t-test, and it is calculated with the software GraphPad Prism (version 5.03). The explanation of the test procedure is shown in Appendix 2.3, as well as the obtained parameters for the sludge from the three different WWTPs End of digestion At the end of the methane potential test, final measures for ph, COD f, COD tot, NH 4 -N f, TS and VS are taken from the bottle content, in order to evaluate the performance of the anaerobic digestion. The bottles are also weighted at the beginning and at the end, in order to verify the decrease in their content mass. Since the real initial values for these parameters were not taken from the bottles when starting up the test, some of the initial values in the bottles need to be calculated in order to be able to compare them with the final measured value. Initial ph and NH 4 -N f in the bottles are calculated with Equation 4, using the values from the original inoculum and sludge just before starting up the experiment. 30

41 X bottle = X inoculumv inoculum + X sludge V sludge +X water V water V total [Equation 4] (X: ph or NH 4 -N f value to be estimated.) The initial and final values of COD f, CODtot, TS and VS of the sludge used in the bottles will be compared and discussed. As the final measured values represent the mix of sludge and inoculum, the values for the bottles containing only inoculum (blank) are used for calculating the final concentration of the sludge, using Equation 5: x sludge =x bottle - x inoculum [Equation 5] Where: x sludge is the final content of COD f, COD tot, TS or VS from the sludge (g/l). x bottle is the final measured value COD f, COD tot, TS or VS content in a bottle (g/l). x inoculum is the average value of the final measurements in the bottles containing only inoculum (g/l). 31

42 32

43 5 Results and discussion The experiment results will be evaluated with two different objectives: quantity and suitability of the produced hydrolysate of different sludge origins as easily biodegradable carbon source; and the impact of the hydrolysate withdrawal on the methane potential of the remaining sludge. Therefore, this chapter will be separated between the biological hydrolysis and the methane potential. 5.1 Biological hydrolysis During the anaerobic hydrolysis of sludge there is a release of dissolved organic matter, ammonium and phosphate, which is monitored through measurements of COD f (and VFA f ), NH 4 -N f and PO 4 -P f, respectively, in the sludge hydrolysate. The concentration of these dissolved substances increases significantly during the first hydrolysis days, until the solubilisation rate decreases and the concentration tends to reach a maximum. A decrease in the COD f concentration is a sign of the beginning of the methanogenesis phase. During the first days of hydrolysis a decrease in ph is observed due to the formation of volatile fatty acids, typical of the acidogenic phase. Measurements of the dissolved substances, ph measurements and solids concentration measurements are used to evaluate the hydrolysis progression. To compare the hydrolysis performance in the current study with results from previous studies, dissolved substances measurements (COD f, VFA f, NH 4 -N f and PO 4 -P f ) are expressed in relation to volatile suspended solids, VSS (g/l) and in addition several ratios found in the literature are calculated. In experiment A, a non-typical trend was observed in the measurements: COD f, VFA f and NH 4 -N f concentration values showed a constantly increasing tendency, without reaching a stable level. At the last days of the hydrolysis experiment (after the 10 th day), the sludge in the reactors dried out, which was a clear sign that evaporation had an important effect during the experiment. Therefore, the constant increasase in the concentrations could be attributed to the evaporation of water in the reactors, reason for why the results cannot be considered reliable. Another phenomenon observed during experiment A was the increasing turbidity of the hydrolysate throughout the hydrolysis. From the sixth day and onwards, the hydrolysate started to become blackish (see Figure 13) and for every day it was more difficult to separate the sludge from the supernatant (needing more centrifugation time), which could illustrate the presence of small suspended particles in the hydrolysate. Figure 13. Hydrolysed sludge samples after centrifugation, in which can be noticed the high turbidity of the supernatant. Sample from the 9 th hydrolysis day during experiment A. 33