Biological Nutrients Eradication Employing Aerobic Granulation

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2018-07-13 Biological nutrients eradication employing Aerobic Granulation

Not available, Anrish

Anrish. (2018). Biological nutrients eradication employing Aerobic Granulation (Unpublished master's thesis). University of Calgary, Calgary, AB. doi:10.11575/PRISM/32661 http://hdl.handle.net/1880/107479 master thesis

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Biological nutrients eradication employing Aerobic Granulation

Anrish

A THESIS

SUBMITTED TO THE FACULTY OF GRADUATE STUDIES

IN PARTIAL FULFILMENT OF THE REQUIREMENTS FOR THE

DEGREE OF MASTER OF SCIENCE

GRADUATE PROGRAM IN CIVIL ENGINEERING

CALGARY, ALBERTA

JULY, 2018

ABSTRACT Nutrients (nitrogen and phosphorus), since mid-20th century, have been a rising matter of concern- impacting the quality of natural water bodies and stressing the wastewater treatment facilities. The increased availability of nitrogen and phosphorus are known to cause the excessive aging of lakes, eutrophication, thus depriving these bodies of oxygen, leading to various ecological, human health and socio-economic impacts.

The phosphorus discharged from untreated or under treated municipal wastewater considerably adds to eutrophication. Although the conventional processes for wastewater treatment are able to remove the organics yet contains residual nutrients in the effluent. These residual amounts exceed the desired effluent limitations, which are getting stringent. The following research provides an insight of an emerging biotechnology: Aerobic Granulation, for its suitability to remove phosphorus (growth limiting nutrient) and nitrogen existing in domestic wastewater.

The research focuses on investigation of different mechanisms supporting domestic level phosphorus removal in aerobic granular system. Granulation is a novel biotechnology based on self-immobilization of diverse microbes into a packed and dense structure giving it a unique ability to retain higher biomass, exceptional settleability, and resistance to toxicity and fluctuating organic loadings.

In comparison to the conventional treatment processes, granulation technology also offers smaller footprint and low operational cost. A characteristic property of the granules is that these can be tailored as per the treatment requirements. The stratified structure of the granules due to mass transfer diffusion provides favorable conditions for phosphorus removal. The co-existence of aerobic zone and anaerobic zone in the granules, as documented in the literature, supports the simultaneous removal of nitrogen and phosphorus.

The lab-scale bioreactor was setup in Hydraulics lab at University of Calgary from February 2016-

August 2016. Synthetic wastewater was used for the experiments with variation in COD and phosphorus concentration in the reactor for evaluating the significance of the operational factors.

The major phenomenon responsible for phosphorus removal at low C:P was biological assimilation.

At a high C:P ratio, biological accumulation and biologically induced phosphorus precipitation were identified as the phosphorus removal phenomena. Microbiological analysis through

Polymerase chain reaction of extracted 16r RNA genes indicated the presence of Rhodocyclaceae; phosphorus removing bacteria with a low temperature based biological phosphorus removal rate.

Data analysis illustrated a significance of C:P on aerobic granular phosphorus removal process.

The results reflect the suitability of aerobic granulation for municipal level phosphorus removal provided optimum operational conditions are maintained. A scale up version of this technology could be used for deeper insight into real domestic wastewater for advanced wastewater treatment processes.

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TABLE OF CONTENT ABSTRACT…………………………………………………………………………………………….....iii TABLE OF CONTENT…………………………………………………………………………………....iv LIST OF FIGURES…………………………………………………………………...…………...….....vi LIST OF TABLES………………………………………………………………………………………..viii LIST OF ABBREVIATIONS ...... ix CHAPTER 1- INTRODUCTION 1.1 Problem statement ...... 2 1.2 Nitrogen and Phosphorus……………………………………………..…………………….…5 1.3 Nutrients pollution: Status around world.………………………………………………….….7 1.4 Research objectives………..…………………………………………………………..…...... 8 1.5 Research questions…………………………………………………………….…...……..…...9 1.6 Scope and limitations…………….……………………………………..………………..…..10 1.7 Summary……………………………………………………………………………………..10 CHAPTER 2- LITERATURE REVIEW 2.1 Conventional phosphorus treatment technologies for phosphorus removal…………………13 2.1.1 Chemical removal………………………………………………...………………..…….14 2.1.2 Physical removal…………………………………………………………………………18 2.1.3 Biological treatment………………………………………………………………..…….19 2.1.3.1 Conventional biological phosphorus removal technologies…………………...……..21 2.1.3.2 Combined processes for biological phosphorus and nitrogen removal…………...…28 2.1.3.3 Disadvantages of conventional biological processes………………………………...30 2.2 Granular Processes……………………………………………………………………...……31 2.2.1 Anaerobic Granulation…………………………………………………………..……..…32 2.2.2 Aerobic Granulation………………………………………………………………………34 2.2.2.1 Formation of an aerobic granule…………………………………………..………….36 2.2.2.2 Mechanism of granulation………………………………………………….………...39 2.2.2.3 Factors affecting the formation of granules……………………………….……….…42 2.2.2.4 Application of aerobic granules…………………………………………....…………52 2.2.2.5 Nutrients removal……………………………………………………….…..………...54

2.2.2.6 Phosphorus removal via granulation……………………………………..…………...58 CHAPTER 3- MATERIALS AND METHODS 3.1 Reactor setup and operation……..…………………………………………………………...67 3.2 Media………………………………………………………………………………..….……72 3.3 Analytical methods…………………………………………………………………………..75 3.4 Experimental procedure……………………………………………………………….…..…77 CHAPTER 4- GRANULE DEVELOPMENT 4.1 Start-up phase……………………………………………………………………………...…80 4.2 Granule development…………………………………………………………………….…..83 4.3 Granular characteristics……………………………………………………………….….….86 4.3.1 Mixed liquor solids and volatile suspended solids (MLSS & MLVSS)……….….…..…86 4.3.2 Particle size……………………………………………………………………………....91 4.3.3 Sludge volume index…………………………………………………………………….93 4.3.4 Morphology……………………………………………………………………..……….98 CHAPTER 5- COD AND AMMONIA REMOVAL

+ 5.1 COD and NH4 removal………………..…………………………………………...……103 5.1.1 COD removal ………………..…………..……………………………………………104 5.1.2 Ammonia removal ….……………………………………...………………………….107 5.2 Discussion……………………………………...…………………..………………………110 CHAPTER 6- PHOSPHORUS REMOVAL 6.1 Phosphorus removal profiles……………………...……………………………………….118 6.2 Discussion…………………………………………………………………………..……...122 6.2.1 Biological assimilation………………………………………………………………….123 6.2.2 Biological accumulation……………………………………………………………..…129 6.2.3 Biologically induced phosphorus precipitation………………………………………...131 CHAPTER 7- CONCLUSION AND FUTURE SCOPE 7.1 Conclusion………………………………………………………………………………..………..144 7.2 Future work……………………………………………………………….……..……..….146 CHAPTER 8 - REFERENCES………………………………………………………………..148 8.1 Publications……………………………………………………………………………....162

LIST OF FIGURES Figure Page 1. Schematic representation of Phostrip process 23 2. Schematic representation of Modified Bardenpho process 24 3. Schematic representation of Pho-redox process 25 4. Schematic representation of oxidation ditch with anaerobic zone 26 5. Schematic representation of anaerobic/oxic process 27 6. Schematic representation of UCT process 28 7. Schematic representation of Westbank process 29 8. Aerobic granule consortium (10 µm) 37 9. Aerobic granules (2 µm) 37 10. Gradual process of granules formation 39 11. Acetate fed granules dominated by rod shaped bacteria 43 12. Glucose fed granules dominated by filamentous bacteria 44 13. Anaerobic and aerobic zone distribution in a granule 55 14. Photo of bioreactor and experimental facility 70 15.Schematic of reactor setup 71 16. Start-up of reactor seeded with activated sludge 78 17. Bioreactor in settling condition 78 18. Bioreactor in aerating mode 78 19. Particle size variation from day 1- day 60 82 20. Activated sludge particles 85 21. Young granules 85 22. Semi-mature granules 86 23. Rod-like mature granules 86 24. Fully mature granules 86 25. Variation of and MLVSS and MLSS throughout the experiment 91 26. Variation of MLVSS/MLSS throughout the experiment 92 27. Particle size analysis from day 1- day 183 94

28. Sludge settling through 1L flask for SVI test 97 29. SVI-30 variation from day 1- day 183 98 30. Variation in SVI (SVI5 and SVI10) throughout the experiment 98 31. Variation in SVI-30/SVI-5 throughout the experiment 99 32. Spherical and rod-like mature granules 102 33. COD analysis 108 34. Ammonia concentration analysis during the experiment 111 35. Effect of OLR on ammonia removal efficiency 113 36. Microbial composition of granules 115 37. Nitrite and Nitrate concentration in effluent 117 38. Phosphate influent vs phosphate effluent concentration 122 39. Variation of phosphate removal efficiency with time 122 40. Effect of COD concentration at a constant C:P ratio 128 41. Effect of P concentration at different C:P ratios 128 42. Variation of influent VFA concentration with time 131 43. Image of a granule sample taken through SEM 132 44. Image of granular interiors dominated by rod-shaped network of bacteria 133 45. Magnified image of granular interior 133 46. Image showing central interior granular space 134 47. Image depicting microorganisms structure in deep core of granule 134 48. Ionization energy vs counts of granular EDX spectrum 135 49. Changes in MLVSS/MLSS ratio 138 50. Microbial composition depicting major percentage of Rhodocyclaceae 139

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LIST OF TABLES

Table Page

1. SBR cycle of bioreactor 69

2. Operating conditions 71

3. Seed sludge characteristics 73

4. Composition of micro-nutrients 74

5. Feed composition at different OLR’s 75

6. Concentration of COD and COD removal efficiency 126

7. Concentration of P and P removal efficiency 126

8. Elemental composition of granular core through EDX analysis 135

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LIST OF ABBREVIATIONS AGS- aerobic granular sludge p value-probability value

APHA-American Public Health Association H/D-height/diameter

ATP- adenosine triphosphate MLSS- mixed liquor suspended solids

A2/O- anaerobic anoxic process MLVSS- mixed liquor volatile suspended solids

BOD- biochemical oxygen demand OLR-organic loading rate

CCD-central composite design PAO- phosphorus accumulating organisms

COD-chemical oxygen demand PHB- polyhydroxy butyrate

C:N:P-COD:NH3-N:Phosphorus SBR-sequencing batch reactor

DNA- deoxyribonucleic acid SNDPR-simultaneous nitrification denitrification and

phosphorus removal

DO- dissolved oxygen SOUR-specific oxygen uptake rate

DPAO- denitrifying phosphorus accumulating SEM-EDX- scanning electron microscopy-energy organisms dispersive X-ray

EBPR-enhanced biological phosphorus removal SRT- solids retention time

EPS-extracellular polymeric substances SVI- sludge volume index

ESOCs-emerging substances of concern TOC- total organic carbon

F/M- food/microorganisms TN- total nitrogen

FISH- fluorescence in situ hybridization UCT- university of Capetown

GAO- glycogen accumulating organisms UASB-upflow anaerobic sludge blanket

HAB- harmful algal blooms VFA- volatile fatty acids

HRT-hydraulic retention time XRD- X-ray diffraction

IC- internal circulation

CHAPTER 1- INTRODUCTION

1.1 Problem Statement Back in mid- 20th century several lakes and reservoirs saw an unusual form of pollution which led to questionable levels of discharged concentrations. With effective research, it was recognized that the water pollution problem which resulted in various ecological impacts was a process known as

Eutrophication. The process occurs because of increased availability of the nutrients, nitrogen and phosphorus, as a result of which an excessive growth of plants and algae is induced. This further leads to a depletion in the oxygen levels of lakes making them hypoxic.

Although eutrophication is a natural process which occurs on the geological time scale, the accelerated anthropogenic activities causes an exaggerated aging of lakes due to elevated concentration of the nutrients in the natural water bodies thus taking another form; cultural eutrophication. Alarming reports of minimal to no treatment of nitrogen and phosphorus from various industrial and especially municipal wastewater treatment facilities have revealed the fate of these contaminants in surface waters (Environment and Climate change Canada, 2017).

Due to excessive nutrient enrichment the lakes develop a bloom of aquatic plants, phytoplankton induced primarily by the discharge of phosphorus. When the dissolved oxygen concentration in the water bodies drops, it contributes to ecological problems including fish kill, toxicity, reduced aesthetic value of water body, and poses stress on wastewater treatment units.

As a response to the elevated nutrient levels, the overgrowth of blue-green algae cyanobacteria leads to the production of toxins. Cyanotoxins and mycrocystin, are very dangerous chemicals produced by blue-green algae like cyanobacteria and can destroy marine life, kill mammals, and cause extreme symptoms like diarrhea and liver damage in humans. These blooms have been seen to spread in drinking water supplies posing a threat to human health.

With rapid industrialization and urbanisation, discharge of nitrates and phosphates into environment further disrupt the natural biogeochemical cycles of the nutrients contributing further to nutrient pollution. Therefore, excessive nutrient inputs are a rising matter of concern putting high human health risk, stressing the wastewater treatment facilities and depleting the natural water bodies.

The conventional treatment technologies for wastewater treatment are based on floccular sludge systems. Some of the commonly used biological processes for nitrogen removal are- post-anoxic denitrification (endogenous driven), pre-anoxic denitrification (substrate driven), four stage

Bardenpho process, step feed process, oxidation ditch, two stage segregated process. These are all based on providing aerobic, anoxic and anaerobic conditions to favor nitrogen removal mechanism.

Similarly, for phosphorus removal, biological processes like Phostrip process, modified

Bardenpho process, A/O (anaerobic/oxic) process are used. These processes involve simultaneous nitrification denitrification and phosphorus removal (SNDPR) combining carbon, nitrogen and phosphorus removal and enhanced biological phosphorus removal (EBPR) (Ahn et al., 2009).

These processes are able to remove phosphorus to the desired concentration, however, require aerobic, anoxic and anaerobic stages.

Back in early twentieth century, when activated sludge was developed, and up to 1980s, the goal of conventional secondary treatment systems was to remove organics and TSS. It has been observed that the activated sludge process yields a total nitrogen level of 3-8 mg/L and total phosphorus concentration of 1-2 mg/L in the effluent, which is not desirable as the discharge limits continue to become stringent (Moore, 2010).

New technologies have been developed to overcome the demerits of conventional biological treatment systems. Granular biomass has evolved as a superior alternative to floccular sludge.

Mishima and Nakamura (1991) worked on the application of Aerobic upflow sludge blanket reactor (AUSB) process on treatment of municipal sewage. Granules were developed after 3 weeks of startup. An improvement in settling ability of the sludge was observed as granulation progressed. This study led to the evolution of Aerobic Granulation phenomenon.

Aerobic granules are aggregates of selective microorganisms from activated sludge having the ability to retain high biomass, augmented settling properties, require smaller prints and can remove organics as well as nutrients (Tay et al., 2002c). It is hypothesized that a granule contains cores of aerobic, anoxic and anaerobic layers supporting different microbial populations which can cause several biochemical reactions (Tay et al., 2002b). These reactions aid in the removal of nutrients- nitrogen and phosphorus, organics, and in some cases toxic compounds like phenol, heavy metals and ESOCs.

One of the main objectives of this study is to treat municipal level phosphorus concentrations and identify the operational parameters that affect the phosphorus removal mechanism in aerobic granulation. In contrast to conventional treatment processes which are based on providing separate- anaerobic and aerobic environments, the aerobic granular sludge has the capacity to remove phosphorus in single aerobic environment.

This can significantly reduce the operational cost and land space requirement in comparison to

EBPR processes. In depth investigation of P removal phenomenon will be accessed in the following chapters which could conclude the significance of aerobic granulation for municipal level phosphorus removal.

1.2 Nitrogen and Phosphorus Nitrogen is a critical element for living beings and terrestrial ecosystems. However, its excess presence in the aquatic ecosystem due to discharges from treated wastewater, fertilizers, direct storm water discharge or septic tanks may cause it to enter the surface waters. Nitrate is the most bioavailable form of nitrogen present in the water bodies and its increased availability is known to cause blue baby syndrome and stomach cancer.

Nitrate is primarily considered to be a limiting factor in the control of marine plant growth. Its presence also intervenes with the activity of aquatic animals and exaggerates plant and algal growth. Thus, a strict regulation has been put on its discharge. With regulations becoming stringent the total nitrogen concentration in the treated wastewater needs to be less than 0.1 mg/L (Moore,

2010) The traditional wastewater treatment facilities fail to meet these standards and require extensive tailoring.

Phosphorus, an essential nutrient for all living organisms has a vital role in biological metabolism of all biota and is the first nutrient to limit biological productivity (Wetzel 2001). Typically, a water body with restricted or low availability of phosphorus has a well balanced and diverse aquatic life. This reflects a limiting nature of the element which keeps the population of aquatic plant, algae and other vegetation in control.

However, the increasing anthropogenic activities break the limits of phosphorus discharge and hence a high concentration of phosphorus is introduced into these water bodies disrupting the balance of a healthy ecosystem (USEPA, 2017). It may enter lakes, estuaries and the ocean through primary sources such as industrial processes, agricultural run-off, municipal treated wastewater and detergents.

Naturally occurring phosphorus in the form of rocks, soil, decaying organic matter also contributes to nutrients pollution. These excessive amounts stimulate the growth of algae and aquatic plants forming a dense surface growth in the form of algal blooms (USEPA, 2017). The three available forms of phosphorus in an aquatic system are: inorganic phosphorus, particulate organic phosphorus and dissolved organic phosphorus.

The soluble inorganic form of phosphorus typically orthophosphate is directly utilized by aquatic biota for nutrition. It is then converted to organic form available for consumers and decomposers.

Generally, in fresh water bodies the organic phosphate concentration is most abundant. Usually when phosphorus concentrations exceed the desired limits, the first response of an aquatic ecosystem is increased productivity of plants, algae and biomass.

The few additional feedback of such an ecosystem may include anoxic conditions, production of toxins with potential to harm livestock, surge in biomass, decrease in biodiversity and increase in turbidity. The loss of phosphorus balance in the lake ecosystem involves complex mechanisms.

Input control of phosphorus is just one of the aspects that can help maintain the balance.

The amount of nutrients available in lakes is strongly influenced by the food web structure and the life persisting within it. It can also affect the animals, plants and human habitat associated with the food web. The changing food web due to natural causes and human intervention has led to the existence of non-native invasive species introduced in the Great Lakes (USEPA, 2017). These organisms compete with the native organisms for nutrients which has changed the quantity and forms of phosphorus available to the existing organisms.

1.3 Nutrients pollution: Status around the world Great Lake in the 1960s became a public concern due to the deteriorating water quality resulting from excessive phosphorus loading. Lake Erie, in particular, developed abnormal algal growth because of cultural eutrophication. Thus, Canada and United States proposed a joint approach for reducing phosphorus inputs to the Great Lakes. Implemented in 1972, the binational action was called the Great Lakes Water Quality Agreement.

The phosphorus levels in the Great lakes thus declined for two decades (USEPA, 2017). But 1990s saw a resurgence of severe algal growth due to excessive inputs of phosphorus resulting from urbanization and industrialization. Although it has been few decades since nutrients pollution affected the water quality, the problem continues to persist in the 21st century. Record breaking amounts of most deadly algal bloom in Western Lake Erie indicates the depth of problem.

In majority of Lake Erie areas, harmful algal growth is witnessed regularly thus priority attention is required to cope with the environmental, human health and economical effects caused as a repercussion (USEPA, 2017). The enormous algal spread-out has threatened the safety of the surrounding Northeast’s water supply, marine life, citizens and scientists believe it to be the biggest and most baneful algal bloom ever recorded in the fourth largest Great Lake. Paralytic shellfish poisoning, amnesic shellfish poisoning and fish kill are few of the events related to HABs spread throughout the world concentrated near the coastal regions (USEPA, 2017).

Recently Environment and Climate Change Canada along with its partners at provincial, domestic and binational level are working to address the phosphorus problem and HABS. Their top priority is setting objectives on phosphorus loading to Great Lakes. Focus of new joint phosphorus targets announced by Canada and United States in February 2016 is to reduce the phosphorus inputs to the lakes by 40% (USEPA, 2017).

This would further achieve minimization of hypoxic zones in the central basin of Lake Erie, regulating cyanobacteria biomass level produced in western basin to prevent production of severe concentrations of toxins unsafe for human health and lake ecosystem and maintaining the algal species in western and central basins contributing to a healthy aquatic ecosystem. The federal and provincial agreement, Canada-Ontario Agreement on Great Lakes water quality and ecosystem health revised in 2014 has time-bounded commitment for this issue and to protect the Great Lakes

(USEPA, 2017).

A new approach for nutrient management is required to target the continuously changing ecosystems of the Great Lakes. Modular technologies and choosing an adaptive model that can take care of the ecosystem upsets are required to keep up with the present ecosystems.

1.4 Research Objectives The stringent regulations laid on the disposal of wastewater and further the ecological and health issues associated with nutrient pollution, has emanated careful handling of the wastewater. The aim of this work includes the utilization of single phase for aerobic granular phosphorus treatment in contrast to the complex conventional processes requiring multiple aerobic, anaerobic and anoxic stages.

This biotechnology also targets at an effective removal of COD and ammonia. The research focuses on achieving the following objectives:

1. To determine a protocol for aerobic granule formation with high nutrient removal

capability

2. To investigate the removal efficiencies of COD, ammonia, and phosphorus in aerobic

granular bioreactor

3. To demonstrate the operational parameters affecting phosphorus removal capacity of

aerobic granular bioreactor

1.5 Research Questions The aerobic granulation technology is under detailed review to determine its effectiveness on different areas of wastewater treatment. Its application on nutrients removal has been distinctly presented in the literature (Dangcong et al., 2001, Tay et al., 2002a, Lin et al., 2003, Yang et al.,

2003, Tsuneda et al., 2004, Lin et al., 2005, de Kreuk et al., 2005, Qin et al., 2006, Ferrer et al.,

2008, Zhang et al., 2011). However, it still lacks the understanding of micro and macro level analysis to investigate the mechanisms occurring inside a granule for P-removal.

The conventional biological phosphorus removal processes in wastewater treatment are based on providing sequential anaerobic and aerobic stages for PAOs to grow and consume phosphorus.

This research involves utilization of 1-stage aerobic phosphorus removal in a SBR based granular bioreactor in contrast to multiple aerobic and anaerobic phases. The operational parameters affecting P-removal can then be optimized for future applications on a full scale treatment capacity.

Therefore, based on the thorough literature review and past data, the research aims at finding solutions to the following questions:

1. What is possibility of aerobic granulation to treat phosphorus in the range detected in

municipal wastewater in one aerobic phase?

2. What are the removal efficiencies of COD, ammonia and phosphate in the aerobic granular

bioreactor?

3. What are the possible mechanisms occurring in granules causing biological removal of

phosphorus?

4. What are the operational parameters optimum for the phosphorus removal process?

1.6 Scope and Limitations: From the above discussion, it is apparent that phosphorus is the dominant element in causing the nutrient pollution problem. Therefore, analyzing phosphorus removal is the deeper scope and various operational factors affecting its removal are investigated. Nitrogen removal efficiency is also evaluated alongside chemical oxygen demand since it is present in municipal wastewater.

However, there are certain limitations to the research and they are as follows:

1. Limiting the use of actual wastewater and use of synthetic wastewater for carrying out the

experiments.

2. The analysis would be restricted to influent and effluent analysis of phosphate, ammonia, COD

and VFA.

3. A lab scale reactor would be used and not a pilot scale.

4. The microbiological analysis of granules will be abstract and thus detailed study of

microbiological groups and their metabolic pathways will not be performed.

1.7 Summary Chapter 1 of this thesis includes the introduction which is further divided into problem statement and objectives of the research. Chapter 2 covers the literature review including various conventional treatment processes for nutrients removal and in-depth discussion about aerobic granulation. Chapter 3 gives details of the bioreactor design and the methods used for chemical analysis. The details of aerobic granule development and their characteristics were discussed in

Chapter 4.

The organics and ammonia removal profiles and the significance of aerobic granulation on their removal is discussed in Chapter 5. Chapter 6 provides a deeper insight into phosphorus removal and the causes of phosphorus removal phenomenon in aerobic granulation. This section covers microbiological and SEM-EDX analysis revealing major findings in granules. Chapter 7 includes conclusions and future scope of the research. And finally chapter 8 includes the references used and a list of publications.

CHAPTER 2- LITERATURE REVIEW

2.1 Conventional treatment technologies for Phosphorus removal The average phosphorus concentration found in municipal wastewater can vary between 4 mg/L and 12 mg/L as elemental phosphorus (Moore,2010). According to most of the wastewater effluent phosphorus standards set, the effluent concentration of phosphorus should range between 0.1 and

2 mg/L as P (Moore, 2010). In terms of percentage of phosphorus reduction 80-95% is desirable to meet most of the standards.

The secondary treatment normally removes only a fraction of influent phosphorus 1-2 mg/L and the excess is discharged as effluent (Moore, 2010). To treat these nutrients the treatment facilities in 1950s were amended to deal the evolving issue of eutrophication. As discussed earlier, the Great

Lakes Water quality agreement in 1978 proposed a strict limit on treatment plants discharging phosphorus in the lakes. A concentration of 1 mg/L phosphorus was stipulated per the agreement.

Similarly, the Anglian Water Authority in United Kingdom enforced a practice to reduce phosphorus concentration in the wastewater effluents entering lakes (USEPA, 2017). Chemical precipitation was adopted to remove phosphorus initially and remains the leading technology due to its flexible application and simplicity. Physical removal of phosphorus usually involves filtration.

The conventional biological technologies firmly established for treating the municipal wastewater aim at organics removal but were insufficient for the removal of nutrients as per the limits set.

Conventional activated sludge process typically removes 30-40% nutrients. Through a balanced combination of microbial action, adsorption and chemical precipitation occurring simultaneously in the process, phosphorus removal mechanism takes place (Yeuman et al., 1987).

Trickling filters are reported to remove 30% of the phosphorus. However, concerns of the capability of the slimes to adsorb phosphates by enhanced biological phosphorus removal process

13 is not significant even with amendments in design and operational conditions (Zanoni 1976).

Trickling filters and chemical treatment when used together were reported to achieve a 90% phosphorus removal efficiency (Ademoroti, 1985).

It was observed that activated sludge process renders a P-removal efficiency of around 80% without the addition of any chemicals elucidating the concept of luxury uptake of phosphorus under certain conditions (Lan et. Al 1983). Biological nutrients removal (BNR) processes gained much importance due to their efficiency in utilizing incoming organics as a substrate source for nutrients removal. These processes proved to be energy efficient and resource producers.

2.1.1 Chemical removal The chemical removal of phosphorus involves precipitating phosphorus as a solid followed by removing the solid. Ionic forms of calcium, aluminium and iron are commonly used for precipitation. When the divalent or trivalent metal salt is added to the wastewater, the insoluble metal phosphate gets precipitated which is then settled out by sedimentation. For settling the solids formed, anionic polymers are normally used.

The versatility of this process is reflected by how chemical precipitation can be applied at several stages during wastewater treatment. The chemicals can be added in primary stage before the primary settler, secondary process in activated sludge plant and tertiary system in the secondary effluent.

Pre-Precipitation

Modification of primary treatment process to include chemical precipitation is called the pre- precipitation. Addition of chemicals usually aluminium or iron salts in the primary stage of treatment in primary sedimentation tank can remove significant concentration of phosphorus. This happens as a result of adequate reaction time, thorough mixing, flocculation of suspended solids and organics in primary settler.

The total phosphorus removal varies from 60- 90% in primary stage with precipitation-flocculation in contrast to only 10-20% in conventional primary clarifier with no chemical addition. As a side benefit, BOD and suspended solids removal percentage also increase. The only disadvantage associated with addition of salts in this stage is that the phosphorus might not be in the form of orthophosphate thus precipitating with difficulty (Moore, 2010).

Usually salts like ferrous sulphate, ferric sulphate, aluminium sulphate, sodium aluminate or ferric chloride are added. This step is followed by flocculation using an anionic polymer which separates the solids. This process of chemical addition especially ferrous salts is prone to pH fluctuations thus a strong base is added between salt addition and polymer.

Equations (1), (2) and (3) represent the pre-precipitation process with different salts. To determine the dosage requirements, bench scale lab tests must be performed.

3+ 3−푛 + 퐴푙 + 퐻푛푃푂4 → 퐴푙푃푂4 + 푛퐻 (1)

3+ 3−푛 + 퐹푒 + 퐻푛푃푂4 → 퐹푒푃푂4 + 푛퐻 (2)

2+ 3− − 10퐶푎 + 6푃푂4 + 2푂퐻 →Ca10(푃푂4)6(푂퐻)2 (3)

Co-Precipitation

The process of dosing chemical directly in the aeration tank of an activated sludge process or in trickling filters is called the secondary or co-precipitation. In the secondary system where most of the biological processes take place, the phosphorus is removed by a combination of precipitation, adsorption, exchange and agglomeration. The dosing is done directly to aeration tank. Co- precipitation is favored mainly because of the ease of operation and flexibility to adapting discharge requirements (Moore, 2010).

Chemicals can be added freely at several points in activated sludge process assuming adequate dispersal. This process can yield an efficiency of as high as 95%. The phosphate is generally removed in the secondary sludge. A stronger chemical action takes place in this step since the organic phosphates have been transformed to orthophosphate resulting from primary addition of salts.

In this stage since the sludge in continuously being recirculated along with coagulation- flocculation and adsorption, the chemical consumption reduces (Moore, 2010). The only liability of this process is the addition of dissolved solids which behave as pollutants. Alum or aluminate is considered an effective salt which can be added in the aeration basin near the outlet.

Post Precipitation

Finally, the tertiary stage addition of chemical takes place following secondary treatment. The secondary effluent is treated commonly by lime treatment causing precipitation of phosphorus as hydroxyapatite. Single-stage and two-stage lime treatment systems are used in wastewater. The choice of single stage or two stage treatment depends on the degree of phosphorus removal required. In tertiary chemical precipitation, alongside phosphorus removal, water softening takes place by reaction of lime with alkalinity (Moore, 2010).

The water softening process leads to calcium removal which has a considerable effect on efficiency of the process. Equations (4), (5) represent the increase in pH value beyond 10 and (6) depict the reaction of calcium ions with phosphate to form hydroxyapatite.

퐶푎(퐻퐶푂3)2 + 퐶푎(푂퐻)2 → 2 퐶푎퐶푂3 + 2퐻2푂 (4)

푁푎퐻퐶푂3 + 퐶푎(푂퐻)2 → 퐶푎퐶푂3 + 푁푎푂퐻 + 2퐻2푂 (5)

2+ − − 5 퐶푎 + 4푂퐻 + 3퐻푃푂4 → 퐶푎5푂퐻(푃푂4)3 + 3퐻2푂 (6)

In this case, the formation of CaCO3 aids the process in two ways: first the lime dosage required for operating the process, secondly the CaCO3 formed serves as a settling agent can be determined by the lime consumption. The excess calcium formed reacts with phosphorus to precipitate hydroxylapatite [Ca5(OH)(PO4)3].

Lime treatment in this process represents a higher capital and operating cost in comparison to chemical addition to treatment plant. However, this process can be easily reliable and adds flexibility to operation since there are individual sedimentation and flocculation units so minor upsets in conventional plant would not affect the tertiary treatment. Lime treatment is a preferred process because of its ability to produce effluent with lesser phosphorus than all other methods thus, giving a high degree of removal (Moore, 2010).

Since lime treatment also removes hardness and alkalinity, the total dissolved solids content of the water decreases. Alum and iron salts are the other metal salts used in tertiary treatment of phosphorus. The only reason why post-precipitation is not a favourable approach is its high chemical cost and formation of mixed biological and chemical sludge making them difficult to separate in next stages and requiring large sedimentation tanks.

Separation and recovery of phosphorus from sludge also requires complex mechanisms.

Technologies such as clarification and filtration are normally used to obtain solids separation at most of the wastewater treatment plants. In primary or secondary clarifiers, gravity separation is adopted and use of flocculation zones is recommended (Moore, 2010).

Filtration using membrane filters, media filters is also practiced as a final step to reduce solids and achieve a very low concentration of phosphorus in the effluent. Excessive sludge production happens as a result of chemical precipitates formed and handling such large quantities of sludge poses a problem. Chemical precipitation has been widely used and proven technology because of its simplicity. But factors like high chemical costs, increased chemical slurry production and chemical handling costs limit its use (Moore, 2010).

2.1.2 Physical Removal This removal process is considered a tertiary treatment at wastewater facilities in order to meet the stringent discharge limits. Tertiary treatment aiding phosphorus removal generally includes sand filtration, membrane filters and cloth media filters. In the case of sand filters, the wastewater is allowed to pass through the sand where it seeps through the sand thus removing particles larger than the pore size. These filters are usually designed with layers of different materials like geotextile fabric, gravel, wood.

Dissolved phosphorus can be removed via these filters where they are enhanced with materials that have the ability to cause the adsorption or precipitation of phosphorus from the wastewater.

Deep sand filters have large space and time requirements which is a restriction in their use.

Membrane filters have been in use recently utilizing less space and more effective in removing wastewater pollutants (Reardson, 2006).

The basic idea behind using membrane filters is to use different media as membranes which have different pore size and filtration capacity. The pore size of the membranes decreases downwards in the filter which creates a pressure gradient to force wastewater through the different layers of filter leaving contaminants behind. In order to clean these membranes, they are simply backwashed.

Interestingly, the dissolved phosphorus concentrations drop to less than 0.1 mg/L in the treated effluent. For membrane bioreactors and tertiary membranes, an achievable limit of 0.04 mg/L has been set on TP (total phosphorus) concentrations in treated wastewater. Inspite of all the advantages served by this modular process, high cost and energy requirements and complicated equipment required puts a restriction to its use.

2.1.3 Biological Treatment In the late 1950s a research conducted led to the development of biological phosphorus removal process. It was observed that under certain conditions activated sludge could remove exceedingly high concentrations of phosphorus than what was required for its normal microbial and biomass growth (Greenburg et al., 1955) Another research concluded the effect of changing aeration in

ASP on soluble phosphorus concentration decrement to 1 mg/L (Srinath et al., 1959).

This phenomenon was later discovered to be ‘luxury uptake’ responsible for excess biological phosphorus removal (Levin and Shapiro 1965). A wide range of processes have been developed based on the phenomenon of enhanced biological phosphorus removal (EBPR) and it has been widely recognized as an accepted mechanism for P removal.

EBPR in activated sludge process occurs in response to provision of an anaerobic/anoxic phase ahead of the existing aerobic zone which favors the formation of phosphorus accumulating

19 organisms (PAO). PAOs have been identified as clusters of coccobacillus shaped microorganisms containing polyphosphate. Their uniqueness lies in the fact that PAOs can uptake and store food in anaerobic conditions in contrast to other microorganisms which can only uptake food until oxygen is made available.

They are naturally occurring heterotrophic bacteria favoring anaerobic and aerobic conditions for optimum growth. Acinetobacter spp. was identified as first group of PAO aiding the biological phosphorus removal process (Fuchs and Chen, 1975). In the anaerobic stage, sufficient substrate in the form of volatile fatty acids (VFAs) must be available for the PAOs to assimilate the acids and release phosphorus in the solution.

Energy in the form of stored polyphosphates is used by the bacteria to uptake the acids and produce intracellular PHB (Polyhydroxybutyrate) which takes place concomitantly with phosphorus release. In the next aerobic phase the PHB is oxidized and energy is released. PAOs use this released energy to assimilate phosphorus back into biomass of bacteria. Thus, the phosphorus gets trapped in the biomass and the sludge can be wasted and stored phosphorus can be removed from the system (Moore, 2010).

It is the later uptake of phosphorus by PAOs which causes the net reduction in phosphorus. The availability of simple organic carbon like VFA in the anaerobic phase predominantly affects the biological phosphorus removal. The VFAs are usually formed when municipal wastewater gets fermented. The fermentation process can either occur naturally in anaerobic zone or VFAs can be added artificially using external sources like acetate, propionate, butyrate.

EBPR process is thus a combination of biochemical reactions occurring in anaerobic and aerobic phases (Moore, 2010). A maximum phosphorus removal efficiency of 90% is reported through

20 this mechanism. However, the inconsistency of this mechanism requires an additional tertiary treatment process to be added in the wastewater treatment plant. The tertiary treatment could be chemical precipitation or filtration depending on the concentration of effluent discharging from the secondary process.

Since the phosphorus obtained is biologically bounded, careful sludge management techniques are required. More complex plant arrangements might be also be required if the treatment efficiency is low and does not meet desired discharge limits. Generalizing, all the biological phosphorus removal technologies are based on the principle of directing treatment process from anaerobic to aerobic conditions to favor the augmented growth of PAOs to luxury uptake phosphorus.

The biological processes provide sustainability being renewable energy source and offer economic advantages. This approach has revolutionized the role of wastewater treatment plants from being waste generators to resource producers. With regulatory requirements getting stringent the sustainable biological technologies are focusing on resource recovery and reuse, decrease energy cost and reduce the emission of greenhouse gases. More stress is laid on recycle, reuse and recovery as elements of the wastewater treatment plants.

2.1.3.1 Conventional Biological Phosphorus removal technologies The following process are all based on providing anaerobic and aerobic conditions to the PAOs for EBPR mechanism to occur.

Phostrip Process

Phostrip process was developed in 1965 by Levin and Shapiro is a combination of biological and chemical removal phenomenon. In this process, phosphorus is stripped in the first step and subsequently precipitated by chemical precipitation in the following step to remove the phosphorus. A percentage of returned activated sludge is directed to the anaerobic tank under different detention times, where the phosphorus is released. Thereafter, the effluent or supernatant from the first step is mixed with lime to cause the precipitation (Liu et al., 2010).

An effluent phosphorus concentration of less than 1 mg/L can be obtained through this process

(Liu et al., 2010). Since the phosphorus removed is in the form of lime sludge, the handling of such sludge becomes easier than phosphorus rich biological sludge. The chemical dose required for this process is less than direct chemical addition in aeration basin because lime dosage is dependent on alkalinity rather than the amount phosphorus being removed.

Thus, lower chemical costs and practicability of handling lime sludge favors this process. The system’s inability to utilize influent BOD due to lack of internal recycling and requirement of additional tank for phosphorus precipitation brought this process in controversy. A schematic representation of Phostrip process can be found below in Figure 1.

Influent Aeration Clarifier Effluent

Phosphorus Direct Sludge Recycle Enriched Sludge Waste sludge

Elutriation: Phosphorus Stripped Sludge Recycle Anaerobic Stripped Sludge Recycles Phosphorus Primary Effluenc Stripper Supernatant Return

Lime Feed System

Supernatant Return Reactor- Clarifer

Waste Chemical Sludge

Fig. 1 Schematic representation of Phostrip process

Modified Bardenpho process

In the early 1973 James Barnard developed an internal recycle setup utilizing influent BOD for denitrification which subsequently got recognized as standard process for nitrogen removal. He also laid stress on directing influent BOD in anaerobic zone before aerobic zone could achieve better biological phosphorus removal. Based on his demonstration, a four-stage process was designed intended for nitrogen removal.

The anoxic-aerobic-anoxic-aerobic for nitrification and denitrification was recognized as

Bardenpho Process (Moore, 2010). This process later was modified to add an anaerobic zone ahead of the four-stage process for phosphorus removal known as Modified Bardenpho Process. This process is based on activated sludge process with the addition of four zones to accomplish biological phosphorus and nitrogen removal (Liu et al., 2010).

With the presence on anaerobic zone before the anoxic-aerobic-anoxic-aerobic zones, an anaerobic-aerobic sequencing environment is created which supports the mechanism of biological phosphorus removal. A part of the recycled activated sludge from secondary clarifier is mixed with influent wastewater and enters anaerobic tank. This initiates the process of phosphorus release in the anaerobic tank.

The sludge from the anaerobic zone then enters the anoxic zone where denitrification takes place due to presence of organics in the form of influent BOD mixed with internally recycled sludge from 2nd aerobic zone. About 70% of nitrate can be removed in this anoxic stage (Liu et al., 2010).

This stage is followed by an aerobic phase where several reactions occur simultaneously.

Luxury uptake of phosphorus, nitrification and removal of organics (BOD) occur in the aerobic nitrification phase. A second anoxic stage is successive to the aerobic phase where additional removal of nitrates takes place in order to minimize nitrate fed back to anaerobic stage. The presence of nitrate in anaerobic zone can interfere with the VFA availability to PAOs thus affecting their growth.

A final aerobic stage provided ensures the aeration of mixed liquor to avoid anaerobic conditions which can lead to phosphorus release in the secondary clarifier (Liu et al., 2010). A schematic representation of Modified Bardenpho process can be found in Figure 2 below.

Effluent Internal Recycle

Influent Anoxic Aeration Anoxic Aeration Anaerobic (Denitrification) (Nitrification) Clarifier

Sludge Recycle

Waste Sludge (Phosphorus-Rich) Fig. 2 Schematic representation Modified Bardenpho process

Pho-redox process (A/O)

This process involves addition of an anaerobic zone at the head of aeration tank in conventional activated sludge setup. The returned activated sludge is recycled directly in the anaerobic zone. To avoid nitrification, a low solids retention time is maintained in this process to make sure no nitrates are present in returned activated sludge. If nitrates are present in the recycled sludge, the anaerobic zone can be divided into anoxic zones and anaerobic zones where denitrification and biological phosphorus removal can take place respectively (USEPA,2010).

A schematic representation of Pho-redox process can be found below in Figure 3.

Fig. 3 Schematic representation of Pho-redox process

Oxidation ditch with anaerobic zone

This process is a modification of a standard oxidation ditch by adding an anaerobic zone ahead of the ditch or within the ditch. A continuous flow of wastewater and biomass occurs in the different channels of a ditch. The aerators installed in the channels aid in moving the water and maintain

25 oxygen. Oxidation ditches typically have large footprint and less energy requirements since no additional pipes are needed for its operation.

Low alkalinity is one of the issues witnessed in a ditch with only phosphorus removal as nitrification depletes alkalinity (Moore, 2010). Schematic representation of oxidation ditch with anaerobic zone can be found below in Figure 4.

Fig. 4 Schematic representation of oxidation ditch with anaerobic zone

Anaerobic/Oxic Process

This process is a single system which involves combination of anaerobic, anoxic and aerobic zones. The anaerobic condition in the zone causes release of phosphorus by PAOs and subsequent uptake of phosphorus in the aerobic zone with consumption of BOD. The anaerobic and oxic zones

26 are sub divided into a number of equally sized compartments ensuring no back mixing due to plug flow in the several sub compartments (Liu et al., 2010).

The phosphorus is removed in the form of waste sludge. The effluent phosphorus concentration is usually controlled by the amount of sludge wasting which further depends on the solids retention time. Optimum conditions for A2/O process involves maintain high organic loading rates and short

SRTs (Sedlak RI 1991). Schematic representation of anaerobic/oxic process can be found below in Figure 5.

Fig. 5 Schematic representation of anaerobic/oxic process

UCT Process (University of Capetown)

The process consists of three phases: anaerobic, anoxic and aerobic stage. The returned activated sludge here is recycled to the anoxic compartment where denitrification takes place in contrast to recycling in anaerobic stage in other processes. This was modified to account for elimination of nitrate in the anaerobic stage which can act as an electron acceptor and interfere with the functioning of PAOs.

An internal recycle of nitrate rich sludge from aerobic stage to anoxic zone influent for denitrification. The denitrified supernatant is then internally recycled back to the anaerobic tank for effective phosphorus removal (Liu et al., 2010). Schematic representation of UCT process can be found in Figure 6.

Fig. 6 Schematic representation of UCT process

2.1.3.2 Combined process for biological phosphorus and nitrogen removal The sludge is allowed to strip by adding a stripper in the combined biological system to provide anaerobic contact to the returned activated sludge followed by clarification to separate the stripped sludge. A very high phosphorus removal efficiency can be obtained via this process in comparison to anaerobic-aerobic process without stripper. A low organic loading rate favors the process.

Westbank process

In this process, a pre-anoxic zone is followed by anaerobic zone, second anoxic zone and finally an aerobic zone. The purpose of pre-anoxic zone is to reduce the nitrates entering the anaerobic zone. The influent is discharged into the pre-anoxic, anaerobic and second anoxic zones where denitrification, phosphorus removal and denitrification takes place respectively. The effluent from aerobic zone is fed back to second anoxic zone to enhance denitrification.

This design has an average size footprint due to smaller anoxic zone in comparison to Bardenpho process (Moore, 2010). A schematic representation of Westbank process can be found below in

Figure 7.

Pre-anoxic

Fig.7 Schematic representation of Westbank process

Sequencing batch reactor (SBR)

SBRs can be used for biological phosphorus removal by adding an anaerobic period within the

SBR cycle. SBR is a fill-and-draw system where the tank is allowed to fill with the influent followed by aeration where the reactions occur. The contents in the tank are then made to settle for a certain period of time and at the end of withdrawal it is subsequently decanted. An idle phase may/may not follow the decant phase.

For the purpose of phosphorus removal, an anaerobic period is added following the feeding stage where synthesis of organic carbon and release of phosphorus takes place (Liu et al., 2010). A gentle mixing is provided in the anaerobic stage. Since the aeration is the next step, uptake of phosphorus takes place since aerobic conditions persist. SBRs can also be used to accomplish biological

29 nitrogen removal. A high degree of organics and nutrients removal can be achieved using SBR

(Liu et al., 2010).

2.1.3.3 Disadvantages of conventional biological processes EBPR based processes are known to behave erratically because of many reasons. One probable reason is the competition between PAOs and another microbiological group GAOs (Glycogen accumulating organisms). GAOs behave similarly to PAOs in terms of their metabolic activities.

The GAOs compete with the PAOs for anaerobic consumption of substrate with no contribution to P removal.

GAOs accumulate glycogen instead of uptaking phosphorus under aerobic conditions (Zheng et al., 2014). Thus, the presence of GAOs is regarded as detrimental in the EBPR processes. Organic loading rate and its increased presence is also known to be unfavorable. Freitas et al observed the negative role of organic loading on P removal when acetate was spiked in the system. The surge in acetate lead to an increase in anaerobic P release which was consequently not uptaken by PAOs in the aerobic stage.

The reason for the degraded P uptake was assumably due to outgrowth of heterotrophic bacteria in the aerobic stage due to excessive exposure to organic substrate outcompeting PAOs due to their triggered growth rate. The presence of nitrate in wastewater treatment plants treating carbon, phosphorus and nitrogen is known to cause EBPR failures. The presence of nitrate in the anaerobic conditions can suppress the activity of PAOs.

Further the competition between PAOs and denitrifiers is propelled for organic substrate (Guerrero et al., 2012). In addition, environmental factors like temperature, HRT (Hydraulic retention time),

SRT (solids retention time) and internal recycling time are known to affect the EBPR process.

Temperature directly affects the competition between PAOs and GAOs metabolism. At a temperature higher than 20oc, the GAOs have an advantage in terms of carbon uptake over the

PAOs.

GAOs are considered to be mesophilic while PAOs are assumed to be psychrophilic. This drives the growth of GAOs leading to EBPR instabilities (Zheng et al., 2014). A long SRT leads to decrease in efficiency of P removal as GAOs dominate over PAOs thus upsetting the EBPR system. On the other hand, short SRT could lead to a washout of PAOs causing a dominance of heterotrophic bacteria in the system (Zheng et al., 2014).

A low hydraulic retention time proves beneficial for EBPR process. Song et al., (2008) observed a high removal efficiency of nitrogen and phosphorus by reducing the HRT. Because of the HRT reduction, the F/M ratio increased thereby boosting the P and N treatment capacity of the system.

Zheng et al., (2014) also suggested the use of granules due to their tolerance to shock loadings in concentrated wastewater and toxicity.

The granulation process has been discussed in detail as follows.

2.2 Granular Processes Granulation also known as Bio-granulation, thoroughly studied since 1980’s, can potentially replace the conventional activated sludge process. Microbial granulation is a phenomenon of cell- to-cell adhesion through physical, biological and chemical reactions forming dense microbial aggregates.

The microbial aggregates formed are called granules and they are a niche for different bacterial species containing millions of organisms per gram biomass. The microorganisms inhabiting the granule can perform various roles with respect to decomposition of organics, nutrients-nitrogen

31 and phosphorus and other inorganic compounds found in industrial wastewater. The granules can be cultured under anaerobic and aerobic conditions and hence widely known as Anaerobic

Granulation and Aerobic Granulation processes respectively.

2.2.1 Anaerobic Granulation Anaerobic granulation was discovered in upflow anaerobic sludge blanket reactor (UASB) where granular sludge was a prominent characteristic of the technology. The anaerobic microorganisms play a major role in self-immobilization clustering into dense consortia; granules, due to the anaerobic conditions. The different anaerobic bacterial species residing the granules perform complex interactions to completely degrade the wastes.

Upflow anaerobic sludge blanket technology (UASB) has been developed since 3 decades with approximately 900 units operating all over the world (Alves et al., 2000). Anaerobic granular sludge, known to contain millions of microorganisms, is the most important aspect of anaerobic granulation. It has an exceptional capacity to retain high biomass concentrations, can remove high potential organic pollutants and is an energy alternative giving biogas production (Show et al.,

2004).

In comparison to the traditional anaerobic treatment, UASB offers rich biomass concentrations retention despite of system’s upflow velocity due to wastewater flow and biogas production in the reactor. Since the granules contain selectively cultured microorganisms with good settling properties, the requirement of long hydraulic retention time is not necessary as the sludge retention quality is exquisite.

UASB is a single tank configuration taking a small amount of space and thus being desirable because a high number of organisms could be maintained in the reactor which can treat highly

32 concentrated organic pollutants by rapid transformations (Liu et al., 2002). A typical UASB reactor consists of a tank containing granular sludge bed at the bottom, fluidized zone, gas and liquid separator and finally a settling compartment.

The wastewater enters the bottom of the reactors and passes upwards through the sludge blanket where the organics are degraded and biogas is produced. As a result of biogas production, the fluidized zone forms where further decomposition of contaminants take place. The biogas passes upwards towards the gas liquid separator and gets separated from the liquid. As the sludge approaches maturity, the microorganisms capable of conglomeration proliferate and form granules.

The granules with dense structure and good settling sink back into the sludge bed, while the ones which are dispersed wash out of the reactor. Different mechanisms have been put forward to explain the formation process of anaerobic granules. The mechanisms are broadly physical, chemical and biological forces responsible for holding the microbial aggregates together. A study was performed on brewery wastewater treatment using modified two stage UASB reactor called internal circulation (IC) reactor.

It revealed its potential to treat low strength brewery wastewater and high strength potato processing wastewater (Pereboom &Vereijken 1994). The removal efficiency of COD for brewery wastewater with average influent COD of 1.7 kg COD/m3 was rendered about 80%. And that for potato processing wastewater with influent COD of 6-8 kg COD/m3 a removal efficiency of 85% was achieved.

Another study conducted by Show et al., (2004) on accelerated UASB startup reflected the ability of UASB to remove 80% of influent COD concentration of 2500 mg/L. Several other studies on

33 slaughter house wastewater (Sayed 1987), soybean wastewater (Yu et al., 1998), paper mill wastewater (Pereboom &Vereijken 1994) indicated the reliability of UASB for treatment of high organic pollutants.

Full scale UASB industrial applications include wastewater from food industry, distilleries, breweries, pulp and paper industries, sewage and chemical industries (Schellinkhout & Collazos

1992; Driessenet al. 1994; Zoutberg & de Been 1997). Inspite of the benefits offered by anaerobic granulation problems like long start-up periods, strict requirement of optimum conditions for bacterial growth like high temperature, unsuitability for nutrients removal and deficiency in handling low strength wastewater limit its usage on full scale (Tay et al., 2000).

2.2.2 Aerobic Granulation It is apparent from the above discussion that granular processes are recommended for biological treatment of wastewater since a high biomass can be retained in the reactors which further require small space for setup. With high biomass, a large number of reactive bacterial species are available to degrade the organics and transform the contaminants which could be highly concentrated.

Aerobic granulation phenomenon was discovered in a pilot aerobic upflow sludge blanket reactor study (Mishima and Nakamura, 1991). After identifying the self-immobilization process through

UASB, much attention was given to aerobiosis of activated sludge to achieve immobilization process to achieve similar and better advantages as UASB. The pilot AUSB lead to the formation of aggregates of size 2-8 mm and revealed exceptional settling properties.

A high BOD removal efficiency was observed alongside good suspended solids removal (Mishima and Nakamura, 1991). The aggregates showed the property of self-coagulating with the presence

34 of mild shear force and dissolved oxygen. The filamentous microorganisms produced in the reactor started binding to the aggregate leading to the formation of aerobic granules.

The aerobic granules have also been developed in sequencing batch reactors with potential to remove COD, nitrogen and phosphorus (de Kreuk et al., 2005) (de Kreuk et al., 2006). The disadvantages associated with anaerobic granulation can be overcome using aerobic granulation as the aerobic granulation incorporates the advantages of anaerobic granulation and surmount the drawbacks of anaerobic granules.

Aerobic granulation is thus a novel biotechnology and can be defined as self-immobilization of diverse microbes into a packed and dense structure giving it a unique ability to retain higher biomass, exceptional settleability, and defiance to toxicity and fluctuating organic loadings (Tay et al., 2001b). Aerobic granulation process aims at triggering the growth of healthy microorganisms which can cause self-coagulation and sustain the dynamics of the granular system to effectively remove organics and nutrients found in wastewater.

A characteristic property of these granules is that these can be tailored or customized as per the treatment requirement. In other words, the granules can be trained to fit different operational and environmental conditions. In comparison to conventional activated sludge process which is predominated by fluffy filamentous bacteria, the granular sludge grows as compact aggregate having the capacity to retain more biomass concentrations and volumetric loading rates (de Kreuk

2006).

These systems require economical smaller footprints offering various economic advantages without the requirement of large clarifiers as is the case in activated sludge.

2.2.2.1 Formation of an aerobic granule Aerobic granule formation can take place via two pathways: gradual process and direct process.

The gradual process involves a slow mechanism in which the reactor is implanted with activated sludge which develops into compact sludge aggregates. As the operation continues, the sludge in the reactor gets compact, dense and granular. This is finally followed by formation of mature aerobic granules.

This process was tracked through advanced imaging by Tay et al., (2009). Scanning electron microscope (SEM) was used to produce images of aerobic granule which revealed the composition and topography of the granule sample. The image analysis of activated sludge sample at start-up phase revealed its loose and irregular structure with presence of filamentous bacteria. After 1 week of reactor operation, the imaging reflected clustered shape of aggregates formed.

The granular sludge was detected after 2 weeks of operation where relatively clear boundaries of sludge sample were found. And finally, at the end of week 3, aerobic granules with round and definitive shape dominated the reactor (Tay et al., 2009). Images of glucose fed aerobic granule taken using SEM depicted a congested structure of microbes closely connected. The third week results of SVI indicated enhanced settling ability of granules in comparison to start-up phase sludge (Tay et al., 2009).

Fig.8 Aerobic granule consortium (10 µm) Source-Liu et al., (2002)

Fig.9 Aerobic granules (200 µm) Source- Liu et al., (2008)

The direct process is a fast start-up strategy for the formation of aerobic granules. The observation of the fact that UASB start-up could be augmented by seeding fully mature anaerobic granules in the reactor, lead to the realization of similar strategy for fast cultivation aerobic granules. This reduces the granulation time significantly and achieve faster start-up (Liu et al., 2005).

A study conducted by Zhu et al., 2003 on structural and activity behavior of aerobic granules under

7 weeks anaerobic idle time, revealed the activity recovering capacity of granules when subjected to storage. It has been reported that granules stored in a refrigerator at 4oC for a period of 4 months

37 can maintain their granular shape and physical strength (Tay et al., 2002). Thus, rapid cultivation of aerobic granular sludge has been successfully implemented so far depending on the storage stability of the granules.

2L of freeze-stored aerobic granules were added as seed sludge to the activated sludge to obtain pilot scale start-up of the aerobic granulation process (Liu et al., 2005). The results revealed new granules development in the reactor after 5 days of operation with full microbial activity recovery

-1 of granules within 2 days of operation indicated by increased SOUR value of 94.5 mg O2 g VSS h-1 (Liu et al., 2005).

Another study targeting on quick start up by adding 50% crushed aerobic granules to the floccular sludge demonstrated a short granulation time of 18 days (Yuan et al., 2011). It was hypothesized that granules and flocs when mixed, formed matrices which aided the granulation process while maintaining COD and nitrogen removal capacity (Yuan et al., 2011). This mechanism was further proved by Bond et al., (2011) with the elucidation that the granules proved to act like a carrier media to the flocs for attachment and the granulation process was monitored by confocal laser scanning microscopy.

A recent study was performed on pilot scale SBR start-up by inoculating 25% fully developed aerobic granules after the activated sludge attained good settling properties (Yang et al., 2014).

The granulation time shortened to 18 days with formation of AGS having mean size 1.58 mm and sludge volume index (SVI) of 67.4 ml/g. The possible mechanism proposed for the granulation process here is crystal nucleus theory, where inoculated granules act as a media for microbial growth.

Thus, direct process can be served as fast granulation method overshadowing the possible long start-up times and disintegration problem.

Fig. 10 Gradual process of granules formation Source Tay et al., (2001a)

2.2.2.2 Mechanism of granulation The formation of aerobic granules is a cell aggregation phenomenon occurring with a combination of physical, chemical and biological forces (Khan et al., 2013). The bacterial aggregation can be between genetically identical or distinct species. The granulation mechanism is a five-stage process: microorganism multiplication, appearance of flocs, flocs cohesion, maturation of flocs and granulation (Hailei et al., 2005).

The notion that bacteria prefer a dispersed state over accumulated state, made the granulation mechanism unclear. The fact that bacterial conglomeration was taking place led to the elucidation of initiating force making the bacterial adhesion (Khan et al., 2013). A theory was put forward by

Beun et al., (1999) which explained the possible phenomenon of bacterial lysis under an oxygen deficit environment leading to the formation of granules.

Later it was discovered that the primary force for granulation was the cell surface hydrophobicity

(Liu et al., 2003). Cell hydrophobicity was recognized to be a prominent adhesion force causing triggered interactions of the cells. Results reflecting twofold increase in hydrophobicity of granular sludge in comparison to conventional flocs were obtained (Liu et al., 2003). Hydrophobicity being a physicochemical property of the cell surface plays major role in self-immobilization and adhesion of cells such as biofilms and aerobic granules (Kos et al., 2003).

The hydrophobicity of bacteria leads to accelerated interactions between the cells contributing in their aggregation. Recently, extracellular polymeric substances (EPS) has been identified as a substantial compound supporting the agglomeration of bacterial cells under specific environmental conditions. EPS has been recognized to be natural polymer secreted by microorganisms containing substances such as exopolysaccharides, proteins, humic acid, DNA and lipids.

EPS causes adhesion of the cells due to its property of matrix, responsible for physicochemical properties like hydrophobicity, enhances polymeric interaction and maintains structural integrity of the granules (Wang et al., 2006). During the starvation phase of SBR cycle, EPS was reported to act as a carbon and energy source and its presence shielded the granule from extreme environmental conditions.

EPS also favors more biomass retention due to increased entanglement of microorganisms in the gel networks formed through chemical bonding on EPS molecules (Liu et al., 2004). The high polysaccharides content promotes cell-to-cell interactions strengthening microbial structure through the polymeric matrix formed and resulting in aerobic granules (Liu et al, 2004). However, later a study conducted by Yu et al., (2009) reported the retarding effect of EPS attached to flocs in comparison to the EPS secreted by aggregated cells.

The EPS attached to the flocs inhibited cell-to-cell interactions thereby delaying the granulation process. The possible explanation of this role of EPS could be its hydrophobic or hydrophilic nature depending on its composition. If EPS is composed of uronic acid which is a hydrophobic protein, it becomes hydrophobic and cause aggregation or attachment to surfaces (Brian and

Malcolmm 2002; Khan et al., 2013).

However, under certain conditions, some cyanobacteria transition the hydrophobic nature to hydrophilic, thus causing detachment from surfaces. The role of EPS should be further elucidated to understand the intermolecular polymeric interactions (Sarma et al., 2017).

Proton translocation is another possible granulation mechanism that has been recently reviewed.

The proton pump activation by fermentation of substrate creates a proton gradient across the cell of the membrane. The proton pump recirculates into the cell membrane and leads to the production of adenosine triphosphate (ATP). When the cell becomes abundant of ATP, the proton pool is discharged through the external membrane to the surrounding environment.

This transforms the naturally negative charge of the cell to positive charge turning the hydrophilic characteristics of cell to hydrophobic. Hydrophobic cells repel water and become dehydrated and attaches to the neighbor hydrophobic cells as a result of a repulsive force originated between water

41 molecules and hydrophobic cells (Tay et al., 2000). The gradient formed thus has the capability to neutralize the charges on the cell surface, dehydrate it and further make it slightly hydrophobic

(Sarma et al., 2017).

Therefore, this process can be summarized into four stages: Bacterial cell surface dehydration, nascent granule formation, development of mature granules and post maturation stage (Tay et al.,

2000). As discussed above, hydrophobicity plays a major role in cells adhesion, thus proton translocation could be a predecessor force to hydrophobicity responsible for granule formation.

This accounts for the role of more than one mechanism for cell adhesion.

Quorum Sensing, a phenomenon of production and detection of autoinducer molecules, plays a vital role in formation of biofilms and bacterial adhesion (Ren et al., 2010). The autoinducer molecules or signal chemicals accumulate in the form of cells. These molecules on reaching a threshold, induce gene expression to assist in the growth and function of biofilms. Similarly, granules, a special form of biofilm, have the potential to produce quorum sensing chemicals that affects the cell attachment capability and further formation of granules.

Experiments performed by Ren et al., (2010) support the fact that the formation of quorum sensing signal molecules from the immature and mature granules induces the gene expression of bacterial cells in suspension which further mediates in attached growth and formation of stable granule.

2.2.2.3 Factors affecting the formation of granules A variety of factors have been known to govern the granulation process. With advanced research, it has been ascertained that a combination of operational, environmental, engineering and microbiological factors influences the aerobic granulation phenomenon. The reactors with airlift

42 or bubble column configurations have been popularly used in sequential batch mode for granules development (Liu et al., 2011).

Composition of substrate

The composition of substrate and specially the type of carbon source typically affects the microstructure and the diversity of microbial species (Khan et al., 2013). The various substrates used to grow granules are glucose, ethanol, acetate, propionate, phenol and synthetic wastewater.

Since different sources have been used to grow granules, it can be concluded that the nature of substrate is not sensitive in aerobic granulation process.

A study conducted by Tay et al., (2009) illustrated structural difference between glucose fed granules and acetate fed granules. The glucose sourced granules exhibited a filamentous, loose, fluffy and string like structure whereas the acetate fed granules illustrated non-filamentous, compact structure with predominant rod like species. Nitrifying bacteria and inorganic carbon source have also been used as substrate to develop aerobic granules.

The granules possessed accelerated ability for nitrification (Tay et al., 2002).

Fig. 11 Acetate fed granules dominated by rod-shaped bacteria Source-Tay et al., (2010)

Fig. 12 Glucose fed granules dominated by filamentous bacteria Source-Tay et al., (2010)

Organic Loading Rate

The role of organ ic loading rate in activated sludge process has been considered of utmost importance. As the OLR increases, the size of activated sludge flocs increases which further affects physical properties like settling velocity, density, specific surface and porosity. Therefore, OLR alteration can affect the performance of the reactor in activated sludge process (Tay et al., 2004).

Similarly, for biofilms the significance of OLR has been well developed by the fact that thickness of biofilm is directly proportional to the loading rate of substrate. A higher growth rate of biofilm has been characterized with increase in OLR (Tay et al., 2004). Thus, organic loading rate can be established as a vital operational parameter affecting the design of biological wastewater treatment processes including aerobic granulation.

In case of anaerobic granules formation, a high OLR favors their growth in upflow anaerobic sludge blanket reactor systems (Campos and Anderson, 1992). In comparison to anaerobic granules, the aerobic granules have been known to form within a wide range of organic loading rate, normally varying between 2.5 kg COD/m3 day and 15 kg COD/m3day (Tay et al., 2004).

It seems that aerobic granulation is less sensitive to substrate loading rate applied which further depends on the nature of aerobic bacteria (Tay et al., 2004). However, physical properties of aerobic granules are in direct relation with OLR applied. A study conducted by Tay et al., (2003) indicated granular size increment from 1.6mm to 1.9 mm when OLR was changed from 3 kg

COD/m3 day to 9 kg COD/m3 day explaining the possibility of accelerated growth of aerobic bacteria when more substrate is provided.

The morphology of aerobic granules was independent of OLR since the granules developed at different loading rates appeared to have similar roundness. Other physical properties like dry mass density, SVI and specific gravity were found to be alike even when different OLRs were applied.

The notable characteristic which was most affected by OLR was structural strength of the granule.

As the OLR increased, the physical strength of the aerobic granule deteriorated.

This could be explained by the understanding that an increased substrate loading triggers the biomass production which in turn augments the microbial growth. This high microbial growth intervenes in the microbial community structure thereby reducing the strength of the granule.

Another study conducted by Li et al., (2008) reported the variation in OLR affected the granular morphology, bacterial species and structural integrity.

Striking differences were observed between granules developed at high OLR and low OLR. In the case of high OLR, large and loose granules were formed with lowest species diversity and short granulation time. Whereas at low OLR, the granulation time was longer but small and dense granules formed with active species diversity (Gao et al., 2011). At very high OLRs upto 21 kg/COD m3 day, the granules tend to disintegrate due to change in aggregation tendency (Adav et al., 2009).

Hydrodynamic shear force

Shear force plays an influential role on the formation and structural stability of the granules. It also affects other properties of granules including aspect ratio, EPS secretion and specific oxygen uptake rate (SOUR). The microbial structure and metabolism of the microorganisms are shear force dependent as well. Shear fore is usually measured in terms of superficial upflow air velocity.

In a column type SBR a threshold shear force of more than 2.1 cm/sec is a requisite for formation of aerobic granules (Tay et al., 2001a).

A high shear force is favorable for formation of granules with high granule stability. Regular, round and compact aerobic granules are developed as a result (Tay et al., 2001 a). This could be explained by aggregating tendency of microorganisms. Higher the shear force, greater will be the tendency of the microbes to agglomerate leading to quicker formation of granules (Beun et al.,

1999).

A very high superficial air velocity selects for the formation of smaller and compact granules which are better for granule formation and operation of the reactor since limitations in diffusion are less probable. The aspect ratio representing the roundness of the granules also increased with high superficial sir velocity. The production of EPS was discovered to be a function of shear force.

As the shear force increased, the extracellular polysaccharide content increased and in turn improved the hydrophobicity of granule. Thus, the structure of the aerobic granule improved significantly (Tay et al., 2001a). Tay et al. (2001a) reported the effect of shear force on catabolic activity i.e. SOUR. It was suggested that shear force enhances the respiration activity of the microorganisms and therefore the SOUR increases. Hydrodynamic shear force thusly holds absolute importance in maintaining the structural strength of the aerobic granule.

Settling time

The settling time acts as hydraulic selection pressure which assists in selecting the large, compact flocs or granules and rejects the loose, fluffy flocs out of the system. Short settling times in the range of 2-10 minutes are desirable for laboratory scaled reactors (Tay et al., 2001b). Short settling times are proven to retain the cells with high hydrophobicity and surge the EPS content in sludge

(Gao et al., 2011).

It has been reported that reducing the settling time to 5 minutes achieved faster granulation (Adav et al., 2009). Another study demonstrated the formation of large aerobic granules by application of short settling time (Qin et al., 2004). On the contrary, a settling time of 15 minutes led to difficulties in formation of aerobic granules (Lei et al., 2004).

A different approach was adopted by Sheng et al. (2010), where a percentage of the biomass was discharged from the reactor and established this as an effective parameter controlling granulation.

A daily discharge of 10% was reported which selected for removal of slow settling biomass and hence more substrate was available for large, compact biomass leading to selective bio granulation.

Dissolved Oxygen (D.O)

The role of dissolved oxygen concentration on aerobic granulation is debatable. Evidence shows that aerobic granules can be formed in SBR at D.O concentration of as low as 0.7-1 mg/L

(Dangcong et al., 1999). Granules can also be nominally developed at dissolved oxygen concentrations of upto 5 mg/L (Tay et al., 2009). On the contrary, Hailei et al., (2006) reported a high D.O level of at least 2.5 mg/L is imperative for granulation. Therefore, further investigation is required for investigating the effect of D.O concentrations on aerobic granules.

Aerobic starvation

The microorganisms present in SBR system are subject to periodic changes due to sequential cycles of feeding, aeration, settling and discharge. The aeration period is divided into two consecutive phases: degradation phase and starvation phase. The depletion of substrate takes place in the degradation phase and the availability of external substrate to microorganisms tends to cease in starvation phase (Tay et al., 2001b).

This starvation phase holds significance in the formation of aerobic granules as it accelerates the granulation process. It was established that starvation conditions make the bacterial cells more hydrophobic, causing them to aggregate and bond (Kjelleberg and Hermansson 1984).

Aggregation can be considered as strategic action of bacterial cells to withstand starvation.

Starvation induces surface characteristic changes in microorganisms and triggers the aggregation ability.

It was hypothesized that a long starvation phase had an affirmative effect on granulation in SBR

(Tay et al., 2001 b). Another similar study was performed by Liu and Tay (2007) to examine the exposure of aerobic granules to different starvation times. This study concluded that aerobic granules formed faster with short starvation time than longer starvation times, thus indicating starvation time as a minor factor affecting the formation process of aerobic granules.

However, in terms of stability values, the longer starvation time seemed more suitable than shorter starvation. Furthermore, too short starvation period was ineffective for long term stability of granules. Li et al., (2006) proposed that starvation initiates aerobic granulation and further aids anaerobic metabolism of facultative microorganisms facilitating granulation process.

Presence of divalent cations

Divalent cations such as Ca2+ and Mg2+ are known to stimulate granulation process via two pathways. One, due to strong Vander Waals forces, these cations neutralize the negative charge on the cell surface of bacteria and second by acting as a cationic bridge between bacteria. Divalent ions like Ca2+ have also been related with EPS and their complex interactions.

Ca2+ binds itself to the negative charges on EPS molecules and stimulates microbial aggregation.

Alginate molecule is an anionic polysaccharide present in EPS which plays an important role in development of granules. Calcium ions link with these anionic polysaccharides and supports the generation and stability of granules (Sarma et al., 2017). A high quantity of polysaccharides production was also observed with the presence of calcium ion, further helping in the formation and maintenance of aerobic granules (Jian et al., 2003).

Polysaccharides forms a strong matrix as the OH- functional group present in it combines with

Mg2+ to form a stable polymeric framework that assists in maintenance and stability of granules

(Li et al., 2009). Therefore, the role of divalent ions in promoting granule formation, early aggregate formation, faster granulation and reduced start up time is vital through enhancement of physicochemical forces.

Hydraulic Retention time

Hydraulic retention time (HRT) is another significant factor affecting the granule formation process. A short HRT was suggested by Beun et al., (1999) for aerobic granulation in sequencing batch reactor as it aided in removing the suspended biomass from the reactor. To evaluate the effect of HRT on the stability of aerobic granules, a study was conducted by Pan et al., (2003).

Four different values of HRT i.e. 2, 6, 12, 24 hours were adopted to ascertain parameters like diameter, strength, integrity coefficient, hydrophobicity and EPS production. The granules formed with smaller HRTs possessed better characteristics; mean diameter 1.2-3.5 mm, granule strength

1.054-1.065, integrity coefficient 98%, hydrophobicity of 65%-69% and amount of EPS 6.1-6.5 mg PS per mg PN. Further the HRT of 6 hour was chosen as optimum for microbial granulation.

Other factors affecting granulation

Other factors like temperature, pH, volumetric exchange ratio, H/D ratio and F/M ratio seem to have a minor effect on aerobic granulation. Temperature has some effect on the performance of a granular system. As reported by Hailei et al., (2006), too high (41o C) and too low (26o C) temperature caused a decrement in biomass in the reactor system. Between the temperature range

29-38o C, the effect was found to be insignificant.

Lower temperature ranges have also been experimented by de Kreuk et al., (2005). Unstable granules were formed at 8o C while a system with startup temperature of 20o C that got eventually lowered down to 15o C and 8o C, the granules formed were comparatively stable. Temperatures of

25o C, 30o C and 35o C were used as comparison by Song et al., (2009). 30o C was reported as optimum temperature as the granules possessed a compact structure, better settling rate and high microbial activity. Hence, an optimum temperature suitable for granulation should range between

28o C and 35o C (Khan et al., 2013).

The pH controls aerobic granulation to some extent since the selection of pH affects the type of species growing in the granular system. A low acidic pH medium favors fungus dominating granules while a high alkaline pH medium favors bacterial dominating granules (Yang et al.,

2008). A pH around 7.5 is necessary for aerobic granulation and a pH above 8.5 tends to inhibit granulation process (Hailei et al., 2006).

Food/microorganisms ratio (F/M) is assumed to control aerobic granulation to some degree. A high F/M ratio promotes faster granulation, improved granule stability and larger size. And a lower

F/M ratio lead to slow formation of granules (Lobos et al., 2008). Since large granules suffer diffusion limitation, a certain strategy should be applied at different stages of granulation. Li et al.,

(2011) suggested applying a high F/M at the early stage and low F/M at the advancing stages of granulation.

Volumetric exchange ratio defined as the volume of mixed liquor ejected through each cycle of

SBR, is known to influence granulation. 40%, 50% and 80% exchange ratio could be typically maintained (Khan et al., 2013). A high volumetric exchange ratio in SBR column has been proven effective for rapid granulation yielding better characteristics of granules in terms of results on EPS, mean size, SVI and granule fraction (Wang et al., 2006).

Volumetric exchange ratio needs to be controlled depending on different stages of granulation. A high volumetric exchange ratio could result in excessive biomass washout if the granular sludge does not tend to settle. A high H/D ratio of the reactor has been reported to produce high protein to polysaccharide ratio and stronger granular structure (Zhu et al., 2008). However, Kong et al.,

(2010) adopted different ratios of H/D at different settling velocities, and they discovered similar physical properties, microbial populations and granulation speed. This implies that H/D ratio did not have an influential effect on aerobic granulation.

2.2.2.4 Applications of aerobic granulation The diverse microorganisms present in the aerobic granules are multi-functional. The granular system provides flexibility over the treatment process as microbes adapt to the environment and function as per the treatment requirement. Aerobic granulation has been applied to treat organics in municipal and industrial wastewater, nutrients removal, heavy metals, toxic pollutants removal.

Now it is being modified to be applied on radioactive waste, emerging substances of concern, hazardous waste treatment by application of different operational and optimized parameters.

High Strength wastewater treatment

Aerobic granulation has been applied to treat wastewater from slaughter house, brewery, municipal and landfill leachate with optimum results. Liu et al., (2008) stated COD concentrations of upto

3000 mg/L could be utilized easily to form and maintain granules. Industrial wastewater from dairy products was also used to develop aerobic granules with organic loading rate of 7 kg/m3 day. An overall COD removal efficiency between 85% and 95% was achieved (Arrojo et al., 2004).

In order to treat high strength industrial wastewater containing phenol, aerobic granular sludge was used. The reactor reached a steady state after 2 months of operation with stable phenol removal efficiencies (Jiang et al., 2004). The application of aerobic granulation for high strength wastewater treatment was examined in a study, by step increasing the organic loading rate from 6 to 9, 12 and 15 kg COD/m3 day (Moy et al., 2002).

The organic loading was raised after the COD removal efficiency attained a stable value of minimum 89% for atleast 14 days. The aerobic granules sustained the high loading rates upto 15 kg COD/m3 day with glucose as substrate and 9 kg COD/ m3 day with acetate as substrate. The irregular structure obtained for glucose fed granules allowed for better diffusion and penetration

52 of nutrients into the granular cores. A stable COD removal efficiency of 92% was acquired which ascertained the applicability of aerobic granules to treat high strength wastewater.

Degradation of toxic substances

Toxic organic compounds such as phenol is a major pollutant found in industrial wastewater. Due to its toxic nature, the microbial growth and biodegradation of phenolic substrate are retarded.

However, the microbe’s agglomeration phenomenon could be utilized to degrade the phenolic compounds from wastewater. Jiang et al., (2004) first demonstrated the ability of aerobic granules to degrade phenol at a phenol loading rate of 1.5 g/L/d.

The bio-granules exhibited excellent phenol degradation capability by degrading phenol concentration of 500 mg/L to less than 0.2 mg/L in the effluent. Even at phenol concentrations of as high as 2000 mg/L, the rate of phenol degradation was significant. Another startling study was performed by Moussavi et al., (2010), where phenol removal efficiency of 99% was obtained with influent phenol concentrations of upto 1000 mg/L using a total cycle time of 17 hours.

A concentration of 5000 mg/L of phenol was effectively degraded by the aerobic granules without any inhibitory effects (Ho et al., 2010). Aerobic granulation has been applied on the treatment of wastewater containing extremely toxic chlorinated phenolic compounds. A high concentration of

4-chlorophenol was completely removed via acclimated aerobic granules with high removal rates

(Carucci et al., 2009).

Wang et al., (2007) reported degradation of another toxic compound, 2-4 dichlorophenol (2,4-

DCP) by aerobic granular sludge with a removal efficiency of 94%.

Adsorption of heavy metals

Heavy metals pose a serious threat to the environment because of their toxic nature. Aerobic granules act as a biomaterial for their removal and this attribute is given by the physical properties of the granules. A strong microbial structure coupled with large surface area and high porosity makes it suitable to use as a biosorbent for heavy metals and dyestuff. A cationic dye, malachite green was removed by aerobic granules by adsorption phenomenon.

Each gram of aerobic granules could adsorb 56.8mg of malachite green at a temperature of 30oC and alkaline pH (Sun et al., 2008). Adsorption capacity of Cd2+ and Zn2+ via aerobic granules was investigated by Liu et al., (2002, 2003). It was found that the maximum adsorption capacity of

Cd2+ was 566 mg/g and that of Zn2+ was reported as 270 mg/g. The adsorption capacity of the granule was estimated to be directly related to the initial concentration of the metal and the concentration of granules.

Later Xu et al., (2006, 2008) conducted a study on analyzing mechanism of Cd2+, Cu2+, Ni2+ and

Cr3+ biosorption on aerobic granules. The primarily phenomenon responsible for adsorption were traced to be chemical precipitation, bonding to EPS and ion exchange.

2.2.2.5 Nutrients removal Aerobic granular sludge has been widely used to remove nitrogen and phosphorus present in municipal and industrial wastewater. The co-existence of aerobic, anoxic and anaerobic zones due to the mass transfer gradient inside the granule support the growth of different trophic bacteria such as nitrifiers, denitrifiers and anaerobes (Tay et al., 2002). These micro environments support different bacterial forms cause nitrification, denitrification and phosphorus accumulation.

This process can occur either in addition to or independent of organic matter removal depending on wastewater composition and operational conditions.

Fig. 13 Anaerobic and Aerobic zone distribution Source-Toh et al., (2002)

Simultaneous removal of nitrogen and organic carbon was performed by Yang et al., (2003) where different N/COD (5/100 – 30/100) ratios were adopted in SBR. Results indicated the presence of heterotrophic, nitrifying and denitrifying populations in the granule. Granules developed at high

N/COD ratio yielded active nitrifying and denitrifying populations, whereas, heterotrophic populations were inhibited at high ratios.

Although a high organics removal efficiency was obtained, denitrification was prominently dependent on concentration of dissolved oxygen. A D.O concentration of less than 0.5 mg/L along with liquid-granule mixing was suggested to achieve high denitrification efficiency. Nitrifying granulation mechanism was later proposed by Tsuneda et al., (2004) in aerobic upflow fluidized bed reactor in the presence Fe as a trace metal.

The 2-step mechanism compromises of (1) Formation of granule core due to aggregation of nitrifying bacteria and Fe precipitation. (2) production of mature nitrifying granules as a result of

55 nitrifying bacteria multiplication and moderate shear force. This mechanism was supported by

FISH (Fluorescence in situ hybridization) analysis of change in nitrifying bacteria distribution inside the granules.

To study the effect of low oxygen concentration on N removal in AGS reactor, a step reduction in oxygen from 100% to 50%, 40%, 20%, 10% was carried out. Results demonstrated an increase in

N removal from 8% to 45% with oxygen concentration reduced from 100% to 40%. However, the decrement in oxygen concentration lead to instability of granules along with reduction in biomass as a result of washout (Mosquera et al., 2005).

De Kreuk et al., (2005) monitored the long and short term effect of oxygen saturation on N removal efficiency under anaerobic-aerobic SBR operation. When the saturation level was decreased to

20%, a high nitrogen removal efficiency of 94% was obtained. Long term analysis of the reactor indicated the dependency of N-removal efficiency on the diameter of granules. A smaller granule diameter resulted in low N-removal efficiency.

It was further observed that a granule size more than 1.3 mm led to an optimum N-removal efficiency. The overall purpose of size limitation is to have an increased anoxic zone in the granular micro zones. Decrease in oxygen saturation or concentration leads to the formation of a bigger anoxic zone containing DPAO (Denitrifying phosphorus accumulating organisms) which causes increased denitrification.

For SND efficiency, the ratio between volume of aerobic layer to the volume of anoxic core is important for long term effects. The effect of minimum settling velocity on granulation and nitrogen removal was reported by Kim et al., (2006) under alternating aerobic and anoxic

56 conditions in a SBR. Two settling velocities of 0.6 m/h and 0.7 m/h were chosen as operational parameters.

With increase in the settling velocity from 0.6 m/h to 0.7 m/h, the response in terms of COD removal, denitrification and granulation time, improved significantly. The COD removal efficiency reached a value of 92% from 82%. The denitrification efficiency increased from 78% to 97% and the SVI decreased from 85 mL/g to 50 mL/g and the granulation time shortened effectively.

The effect of alternating aerobic-anaerobic conditions on cultivation of microbial granules and nitrogen removal capacity was examined at different ammonium-N loadings (0.15-0.45 kg/m3 day). A COD removal efficiency of greater than 95% was obtained, however, the denitrification capacity was primarily depending on the availability of an external carbon source.

In the presence of ethanol, a complete denitrification (N removal efficiency of 99.9%) was achieved, while in the case of no external source, partial denitrification occurred. This study further indicated that the microbial aggregation is a protection strategy of sensitive nitrifying populations which could survive the shortcomings of conventional nutrient removal systems. Aerobic granular sludge membrane bioreactor was used for organics removal and simultaneous nitrification and denitrification (SND).

Influent TOC concentration in the range 56.8-132.6 mg/L yielded TOC removal of 84.7-91.9%.

+ For influent NH4 in the range 28.1-38.4 mg/L, the ammonia removal obtained was 85.4-99.7% and total nitrogen removal was 41.7-78.4% (Wang et al., 2008). A recent study on high strength ammonium wastewater treatment illustrated the potential of aerobic granules to remove high ammonium concentrations at different COD/total nitrogen (C/TN) ratios.

During a 12-hour cycle, aerobic granules treated high concentrations of COD (9860 mg/l) and

+ NH4 (2000 mg/L) with an efficiency of 94.5% and 59.6% TN. It was further reported that the possible mechanisms for ammonium removal were: granular adsorption, air stripping, conversion by nitrification/denitrification (Yu et al., 2014). In depth investigation of microbiological aspects of nitrogen removal mechanism is further required to obtain and optimize aerobic granular systems on pilot as well as full scale.

2.2.2.6 Phosphorus removal via granulation All the current Enhanced biological phosphorus removal processes (EBPR) have been utilized on full scale. They often require large reactor volume due to requirement of anaerobic and aerobic conditions and are based on suspended biomass cultures. Unstable and unreliable operation in full scale treatment has been recognized (Lin et al., 2005). On the contrary, granular based P removal systems can offer reduced land space requirement as aerobic conditions are found adequate as per the research findings.

Thus, granular based system also offer an impressive alternative over the conventional suspended growth P removal systems due to granular strong physical structure, high biomass retention capacity and ability to bear shock loadings (Beun et al., 1999, Tay et al., 2001b).

Phosphorus accumulating granules were developed in SBR at different P/COD ratios under alternating anaerobic aerobic conditions to mimic EBPR system (Lin et al., 2003).

As the substrate ratio (P/COD) increased, the size of the granules decreased with their structure

3- getting denser. The cyclic analysis through COD and PO4 profiles revealed the consumption of organic carbon and release of phosphate in the anaerobic stage and accelerated phosphate assimilation in the aerobic stage. The SVI decreased with an increase in substrate P/COD ratio.

SVI of 12 mL/g was reported at the highest P/COD ratio of 10/100 demonstrating the importance of substrate ratio on structure and settleability of granules.

As a comparison, non P-accumulating granules have a SVI in the range of 50-100 mL/g, therefore the P accumulating granules clearly have a better settleability and compactness. Further in terms of microbial activity, a low SOUR (specific oxygen uptake rate) value was obtained at high P accumulation indicating restricted respirometric activity favors phosphorus accumulation.

It was reported that the uptake of P by the granules was in the range of 1.9%- 9.3% by weight, comparable to conventional EBPR processes.

Dulekgurgen et al., (2003) also conducted a similar study on enhanced biological phosphate removal using granular sludge in SBR with anaerobic-aerobic cyclic operation. Sufficient time past the start-up phase, the granular sludge developed, rendered 95% of carbon removal, 71% nitrogen removal and 99.6% phosphate removal. Rod shaped bacteria assumed to be PAO was detected through microbiological analysis alongside coccoid shaped clusters resembling GAO.

Later, determination of elemental composition and distribution of P-accumulating granules (Liu et al., in 2003) was re-examined by Liu et al., (2005). The phenomenon responsible for phosphorus accumulation in the granules was biological storage and chemical precipitation of phosphorus in the granules was less than 10% of total phosphorus accumulation.

The granules primarily assimilated phosphorus from the liquid solution. Polyvalent cations Ca2+ and Mg2+ accumulation was also found in the granules, and their content was proportionally related to P content in the granules. At a low saturation level of 20%, de Kreuk et al., (2005) found out the phosphate removal efficiency of 94% could be obtained. Slow growing

59 microorganisms such as PAO were selected at low oxygen concentration which proved to contribute towards granule stability.

It was established that selection of bacterial populations with low growth rate could lead to granule stability. So, feeding the substrate under anaerobic conditions was performed for substrate storage. This in turn caused enrichment of PAOs in the granules due to alternating anaerobic feeding and aeration phase leading to a positive effect on phosphate removal. The P content of the granules was estimated to be 0.20 g P/g SS which is higher than normal biofilm values of 0.02-0.14 reported in literature (de Kreuk et al., 2005).

Therefore, precipitation in the granule is one of the reasons for phosphate removal. Interestingly, the phosphorus uptake was the maximum at a 100% saturation and decreased under anoxic conditions. At lower oxygen conditions, the oxygen penetration decreased and the anoxic volume in the granule increased. This in turn lead to a decrease in aerobic volume and thus the phosphate uptake rate declined.

Murnleitner et al., (1997) also reported a lower phosphorus uptake under anoxic conditions as compared to aerobic conditions. This area requires further research to clarify on preferences of aerobic conditions for effective phosphorus removal. Nutrient rich industrial wastewater was treated using granular sludge via simultaneous nitrification denitrification and phosphorus removal (SNDPR) (Yilmaz et al., 2007).

SNDPR requires alternating anaerobic and aerobic conditions, where P uptake and conversion of ammonia to nitrogen gas takes place concomitantly. With COD, nitrogen and phosphorus loading rate of 2.7 g COD/L day, 0.43 g N/L day and 0.06 g P/L day respectively, the removal

60 efficiencies of total COD, total nitrogen and total phosphorus obtained were 68%, 86% and 74% respectively.

The dip in total removal efficiency was due to high suspended solids content in the effluent. The nutrient removal efficiency could be credited to the presence of large anoxic micro zone located at the center of the granule which facilitated the SNDPR process. Although, chemical precipitation was evaluated during anaerobic phase of the SBR cycle, its significance in nutrients removal was found to be minor due to the subsequent dissolution in the aerobic period.

Observations of change in microbial communities under varying pH in a granulation EBPR SBR system were determined using FISH (Ahn et al., 2009). The granulation mechanism at pH of 7.5 and 7 was proposed as: enrichment of existing flocs with Accumulibacter microcolonies and formation of granules by co-aggregation of the microcolonies. As the pH was lowered to 6.5, the granules increased in size.

It was noted that glycogen accumulating organisms (GAO) were predominant at low pH value of

6.5. This was observed by an increased consumption of glycogen as an energy source in anaerobic period resulting in replacement of poly-phosphorus thus reduced P release. Switching the pH back to 7.5, the granules became unstable and disintegrated. This was supported by the presence of filamentous bacteria Chloroflexi in the loose biomass formed.

A microbiological analysis on granulation mechanism was performed by Barr et al., (2010) on a phosphorus removing aerobic granular system. Three lab scale SBR reactors were used to treat two different types of wastewater.

The initial stage of granulation revelated the formation of white granules and yellow granules.

The white granules had a compact and smooth structure with 97.5% of Candidatus

Accumulibacter phosphatis ppkI gene phylotype and yellow granules had a loose and irregular structure with mixed microbial community of 12.3% Candidatus Accumulibacter phosphatis and

57.9% Candidatus Competibacter phosphatis ppkI gene (Barr et al., 2010).

Thus, the formation mechanisms were found to be different for the two types of granules. The white granules were assumed to form by multiplication of individual microcolony to a granule consisting of one type of bacterial population. While the yellow granules were hypothesized to form from aggregation of multiple microcolonies to a microcolony segregated granules containing mixed microbial communities (Barr et al., 2010).

Microbial community evolution was investigated during granulation in an anaerobic-aerobic phosphorus removal system. Growth of polyphosphate accumulating granules was observed due to anaerobic-aerobic conditions (Bin et al., 2010). At the initial stage of granules formation, filamentous rod-shaped bacteria were predominant whereas the mature granules were dominated by coccoid bacteria. With increase in the growth of granules, the microbial community got highly diversified.

The primary microorganisms responsible for nutrient removal were identified to be

Accumulibacter, Nitrosospira and Thauera. The microorganisms responsible for maintenance of granule structure were mainly Rhodocyclus and Hyphomicrobaceae. The phosphorus uptake rate by day 90 decreased in comparison to the high P uptake rate in activated sludge due to the high

MLSS and decreased aerobic zone because of oxygen restriction in large granules (Bin et al.,

2010).

Biological nutrients removal was carried out via two different granulation strategies to treat domestic wastewater. The 100% floccular SBR achieved removal efficiencies of 85% and 94%

62 at loading rate of 78 g N/m3 day and 13.5 g P/m3 day respectively. However, the nutrient removal rates in 90% floc SBR + 10 % crushed granules, was more stable even under high loading rates.

When nitrogen loading of 122 g N m-3 day-1 and phosphorus loading of 20.8 g P m-3 day-1 were applied in the 90% floc SBR, removal efficiencies of 85% and 94% respectively were obtained

(Verawaty et al., 2012). Angela et al., (2011) reported the phenomenon of biologically induced phosphorus precipitation in aerobic granular sludge. Direct spectral and optical analysis performed on stable 500 days granules revealed the presence of hydroxyl-apatite, detected in the core of the granules.

The precipitates were found in the form of mineral clusters saturated with calcium and phosphorus predominantly. Among the overall P removal, 45% was accounted for biologically induced precipitation under anoxic-aerobic operating conditions. Towards the end of the study, the drop in MLVSS/MLSS ratio from 80% to 67% indicated mineral accumulation. A removal efficiency of 82% for ortho-phosphate was achieved at the end of 540 days.

Further in the same study, Raman analysis revealed hydroxyapatite as the major mineral precipitated in the aerobic granule core (Angela et al., 2011). Hydroxyapatite was detected due to its best fitting spectrum in intensity and wave number. Further, cut mature granules were analyzed through SEM-EDX (Scanning electron microscopy, Energy dispersive X-ray detector).

Inorganic precipitates were found throughout the granules in different zones but were concentrated in the center.

These precipitates were found to be composed of Ca and P supporting the presence of calcium phosphate in the core of granules. Finally, XRD analysis confirmed the presence of

63 hydroxyapatite through diffractogram interpretation of peaks with spectrums (Angela et al.,

2011).

Although wide application of aerobic granulation has been focused on phosphorus removal using alternating anaerobic-aerobic phases to achieve EBPR bio-granular system, the possibility of opting only ‘aerobic phase’ for P removal through aerobic granulation has not been considered.

In addition, the lack of research on phosphorus removal in municipal wastewater needs contemplation.

Pilot scale granular SBR was used to treat real wastewater (40% domestic and 60% industrial)

+ and the results indicated only COD and NH4 analyses missing phosphorus evaluation (Liu et al.,

2011). Wagner et al., (2013) studied aerobic granulation to treat domestic wastewater on a full

+ scale SBR reactor. COD and NH4 removal profiles were observed but this study lacked reports on phosphorus removal.

Li et al., (2014) applied aerobic granulation on a full scale SBR treating municipal wastewater and reported only 50% removal of TP claiming absence of anaerobic phase. As discussed in the previous sections, the diffusion limitation offers a layered microstructure in a granule. The layered granule provides aerobic-anaerobic environment required for traditional phosphorus removal.

Instead of requiring two different phases, if the presence of one aerobic phase could achieve phosphorus removal, it will render an evolution in phosphorus removal mechanism. Phosphorus precipitation could also be considered as a contributory factor mediating aerobic phosphorus removal. The present research focuses on treatment of phosphorus in municipal wastewater

64 through single aerobic phase allowing ease of operation and ruling out chances of failure as observed in some conventional phosphorus treatment methods.

Various engineering parameters would be recognized to optimize the phosphorus removal process in SBR while maintaining the stability of granules. Lastly, the possible mechanisms shall be considered to deduce the various factors contributing towards P removal under novel conditions.

CHAPTER 3- MATERIAL S AND METHODS

3.1 Reactor Set-up and Operation The experimental setup constituted a column type laboratory scaled reactor with a working volume of 4.6 L. A SBR mode was operated with a cycle time of 3 hours. The height and diameter of the reactor was 80 cm and 8 cm respectively, with H/D ratio of 10. The SBR cycle comprised of feeding, aeration, settling and decant. The settling time was decreased as the settling ability of the sludge improved while approaching granulation.

Upon reaching a steady state, the settling time was set at 5 minutes. An upflow shear force of 1.7 cm/sec was adopted throughout the experiment. It provided a gentle bubble pattern inside the reactor, which assisted in providing the optimum shear to the granules. The air flow pattern was uniform and smooth as a result. This shear force was high enough to keep the granules suspended in the reactor and at the same time moderate to avoid breakage of granules promoting stability. An air pressure control valve was installed with the air pump to regulate the air pressure in the reactor.

The operational conditions were controlled and fixed values were chosen with regards to the experiments performed in the literature. Tay et al., (2001) performed an extensive study on the role of shear force on metabolism and structure of granules. It was discovered that a superficial air velocity greater than 1.2 cm/sec must be provided to develop aerobic granules. Adopting a shear force greater than 1.2 cm/sec also affects the structure and compactness of granules.

Further a settling time of 5 minutes was chosen based on the study conducted by Qin et al.,

(2003) which indicated the importance of settling time on formation of aerobic granules. It was based on the hypothesis that a short settling time favors the selection of particles with good settling properties whereas the poor settling particles being lighter will be washed out. At a settling time of 5 minutes, the granules formed had a defined shape and compact structure.

The reactor configuration in terms of H/D ratio is known to directly affect the flow pattern of liquid and sludge aggregates in the column. A higher H/D ratio of the reactor was adopted to maintain a prolonged trajectory of the circular flow pattern which further favors a better hydraulic abrasion to microbial aggregates (Beun et al., 2002).

The synthetic wastewater was introduced from the bottom of the reactor with a volume of 2.5

L/cycle. An albor pump with a maximum flow rate of 0.33 L/min was used for feeding the reactor. A volumetric exchange ratio of 40% was maintained thus approximately 2 L of the effluent was discharged directly into the drain. A solenoid valve was used for discharging the effluent and the effluent port was located at a height of 40 cm from the bottom of the reactor.

The SBR mode was chosen to provide a better solid-liquid separation of the microbial aggregates. The cyclic operation involving filling, aeration, settling and decanting provides hydraulic selection of the microbial aggregates which further has an influential role on granule formation. The details of SBR cycle and each phase time can be found in Table (1).

The HRT (Hydraulic retention time) was maintained at 7.5 hours according to cycle time and volumetric exchange ratio. The experiments were performed at a room temperature (20-25o C).

To maintain the pH within a range of 7-7.5, buffer solution of KH2PO4 and K2HPO4 was added in the feed. The reactor was aerated using a fine cubical bubble aerator at a superficial air velocity of 1.7 cm/sec.

The operating conditions maintained in the reactor can be found in Table 2 below. Pan et al.,

(2003) investigated the role of hydraulic retention time on the growth and stability of granules. It was established that a HRT between 2 to 12 hours favored a selection pressure that promotes the

68 formation of stable aerobic granules in terms of settling ability. Thus, a hydraulic retention time of 7.5 hours was chosen for the experiment.

The pH was maintained neutral and temperature was set at room temperature as per Tay et al.,

(2002 c) and Beun et al., (2001). Initially after inoculation, the SBR was operated with a cycle consisting of 8 minutes feeding, 120 minutes aeration, 20 minutes settling and 7 minutes of decant. After 60 days of operation, the settling time was gradually reduced to 5 mins with cycle split into 8 minutes feeding, 123 minutes of aeration, 5 minutes settling and 7 minutes decant.

The synthetic wastewater was stored in a water tank with a capacity of 200 L. The configuration of the experimental setup can be found in Figure (15).

Table 1. SBR Cycle of Bioreactor

Operational Day 1-5 Day 6-16 Day 17-59 Day 60 -184

cycle

Fill 8 min 8 min 8 min 8 min

Aeration 2.42 hr 2.55 hr 2.62 hr 2.67 hr

Settle 20 min 12 min 8 min 5 min

Decant 7 min 7 min 7 min 7 min

Total cycle 3 hr 3 hr 3 hr 3 hr

Fig. 14 Photo of Bioreactor and experimental facility

Fig. 15 Schematic of reactor setup

Table 2. Operating conditions

Parameters Aerobic reactor

pH 7-7.5

Temperature 20-25oC

COD (mg/L) 500-4000

TP (mg/L) 3-16.5

HRT (hr) 7.5

3.2 Media The reactor was inoculated with 2 L of returned activated sludge [50% (v/v)] collected from

Pinecreek Wastewater Treatment plant, Calgary, Canada. The Pinecreek Treatment Plant treats upto 100 mega-litre of wastewater per day and includes two bioreactors to carry out Biological

Nutrient Removal (BNR) process. The characteristics of the seed sludge are summarized below in Table 3.

The objective of the study is accessing nutrients and specifically phosphorus removal from municipal wastewater, so a synthetic substrate with similar composition was used. The synthetic wastewater was composed of the following compounds: sodium acetate (CH3COONa), ammonium chloride (NH4Cl), dipotassium phosphate (K2HPO4), monopotassium phosphate

(KH2PO4), calcium chloride dihydrate CaCl2.2H2O, magnesium sulfate heptahydrate

MgSO4.7H2O, ferrous sulfate heptahydrate FeSO4.7H2O and 1 mL/L trace elements.

The feed was prepared in reference to Tay et al., (2002 c). To maintain pH in the range 7-7.5 in the system, a buffer capacity was provided by adding a combination dipotassium phosphate and monopotassium phosphate in the feed which also acted as a source of phosphorus. The trace elements or micronutrients solution contained ZnCl2, CuCl2, MnSO4.H2O,

(NH4)6.Mo7O24.4H2O, AlCl3, CoCl2.6H2O, and NiCl2 (50 mg/L individually).

The feed composition (Tay et al., 2002 c) was changed according to OLR requirements at different points of the experiment. The substrate load in the range of 1.2- 12.8 kg COD/m3 day in terms of COD was applied throughout the experiment. The details of different substrate composition can be found below in the Table 4 and 5. A C:N:P (COD: Ammonia: Phosphate) ratio of 100:4.5:0.5 was maintained to sustain an active microbial culture in the system for growing ‘base’ granules.

Organic loading rate of 6.4 kg COD/m3 day was utilized to develop base granules. A study conducted by Tay et al., (2004) highlighted the effect of different organic loading rate on formation and characteristics of granules. An OLR between 4 kg COD/m3 day and 8 kg COD/m3 day was found most suitable for developing stable aerobic granules with good characteristics.

Tay et al., (2002 c) used an OLR of 6 kg COD/m3 day to study aerobic granulation in a sequencing batch reactor mode. Based on these studies, an OLR of 6.4 kg COD/m3 day was adopted for the present experiments.

Table 3. Seed Sludge Characteristics

Sludge Characteristics Values

MLSS (mg/L) 5902

MLVSS (mg/L) 4945

MLVSS/MLSS 0.81

SVI30 (mL/g) 166

Mean Particle Size (µm) 114

pH 6.43

COD (mg/L) 8410

+ NH4 (mg/L) 8.71

3- PO4 (mg/L) 278

VFA (mg/L) 941

Table 4. Composition of micro-nutrients

Element Quantity (g/L)

H3BO3 0.05

ZnCl2 0.05

CuCl2 0.03

MnSO4.H2O 0.05

(NH4)6.Mo7O24.4H2O 0.05

AlCl3 0.05

CoCl2.6H2O 0.05

NiCl2 0.05

Table 5. Feed Composition at different OLR’s

COD COD COD COD COD COD

500mg/L 1000mg/L 1500mg/L 2000mg/L 3000mg/L 4000mg/L

OLR 1.6 3.2 4.8 6.4 9.6 12.8

(kg COD/m3

day)

Elements

CH3COONa 0.73 1.46 2.19 2.93 4.39 5.86

NH4Cl 0.08 0.17 0.26 0.35 0.52 0.7

K2HPO4 0.0075 0.015 0.0225 0.03 0.045 0.06

KH2PO4 0.00625 0.0125 0.01875 0.025 0.0375 0.05

CaCl2.2H2O 0.03 0.03 0.03 0.03 0.03 0.03

MgSO4.7H2O 0.025 0.025 0.025 0.025 0.025 0.025

FeSO4.7H2O 0.02 0.02 0.02 0.02 0.02 0.02

Microelements 1 mL/L 1 mL/L 1 mL/L 1 mL/L 1 mL/L 1 mL/L

3.3 Analytical Methods The influent and effluent samples were monitored for pH, temperature, soluble COD, ammonia nitrogen, phosphate and VFA. The pH and temperature were analysed using YSI MultiLab IDS

4010-3. Soluble COD (sCOD), ammonia nitrogen (NH3-N), phosphate (P-PO4) concentrations were analyzed based on Standard Methods using relevant Hach kits (APHA, 2012). Volatile fatty acids (VFAs) concentration as acetic acid was monitored by Hach kit (TNT872) considering sodium acetate as sole carbon source.

Phosphate (P-PO4) concentration was also analyzed by ion chromatography (Metrohm 930

Compact IC Flex). Samples for sCOD, NH3-N, P-PO4, VFA analyses were filtered through

0.45µm Whatman filter paper. Mixed liquid suspended solids (MLSS), mixed liquor volatile suspended solids (MLVSS), sludge volume index (SVI) were analyzed in accordance with

Standard Methods (APHA, 2012).

The elemental analysis of granules, surface analysis and their high-resolution images were taken using SEM-EDX technique. The microbial composition of the granule sample was analysed through Polymerase chain reaction of extracted 16r RNA genes.

The particle size of sludge/granules was determined by laser particle size analysis system;

Malvern Mastersizer 2000 (detection limit  2000 µm). As a laser beam moves through a sample, the angular variation in the intensity of the scattered light is measured by the laser diffraction system. The data obtained related to angular scattering intensity measurements is analysed to calculate the particle size based on Mie theory of light scattering.

The size of the particle is expressed as volume equivalent sphere diameter (Malvern Panalytical,

2018). 40 mL of well mixed sludge sample was collected during the middle of the aeration cycle from the SBR. The sample was added to the wet sample dispersion unit Hydro 2000. The sample was transferred to the Mastersizer through tubes for size measurements. A result analysis report was obtained for each sample with a particle size distribution curve corresponding to sample volume percentage. The value of particle diameter at 50% d(0.5) in the cumulative distribution was used as median diameter.

Particles of size greater than 2000µm were measured manually with the aid of a measuring scale

(ruler). Air compressor (Airmax KA20) was used to diffuse air into the bioreactor. A cubical

76 fine-pore diffuser (Pentair Aquatic Eco-Systems® AS40) was used to diffuse air into the reactor.

For controlling the aeration rate and air pressure, a rotameter (Cole Parmer M489770) and air pressure gauges were used (Wika 60 Psi).

3.4 Experimental Procedure To evaluate the performance of aerobic P removal SBR and characteristics of the granules developed, a laboratory scaled reactor built with plexiglass was operated. The synthetic wastewater with a volume of 100 L was prepared biweekly and was stored in the feed tank. An albor pump was used to transfer the feed from the tank to the reactor at a flow rate of 0.33 L/min with an upflow regime.

The feeding was followed by aeration for 2.42-2.67 hours depending on SBR cycle. A cubical fine pore diffuser was used at the bottom of the reactor for a distributed air flow regime. An air flow rate of 1.7 cm/sec was adopted and minor adjustments in the air flow meter was made through rotameter and air pressure gauges. After aeration, the sludge/granules were allowed to settle and the settling time was adjusted as the experiment progressed.

The settling time was chosen as one of the selection pressures and it was decreased from 20 min to 5 min as a short settling time is known to favor aerobic granulation (Liu et al., 2004). As the granules settled, the supernatant was collected in the reactor which was finally discharged. The effluent was decanted directly into the drain via a solenoid valve connected at a height of 40 cm from the bottom of the reactor such that the volumetric exchange ratio value of 40% was obtained.

Approximately 2.5 mL of effluent was wasted per cycle and the effluent samples were regularly taken during decanting. The cycle lasted 3 hours and each operational phase was controlled by

automatic timers plugged to each of the pump, air compressor and valves. The images of the

bioreactor in settling mode and aeration mode can be found below in Figures 17 and 18

respectively.

Fig. 16 Start-up of reactor seeded with activated sludge

Fig. 17 Bioreactor in settling condition Fig. 18 Bioreactor in aeration mode

CHAPTER 4- GRANULE DEVELOPMENT

4.1 Start-up phase The bioreactor was commenced by inoculating with 2.5 L of activated sludge and fed with another 2 L of feed. To acclimatize the sludge to the feed, the reactor was allowed to continuously aerate for approximately 16 hours. A substrate loading of 6.4 kg COD/m3 day was adopted as a base OLR to grow granules. The SBR cycles were started after the acclimatization and the settling time was initially set at 20 minutes to avoid sludge washout during the decant phase.

As the experiment progressed, the sludge settling improved and initial granules were observed on day 9 after the start-up. These granules formed were small, discrete and fluffy in the beginning with average size of 200 µm. Experiments conducted in the literature conclude the age of granule formation as 1 week-3 weeks. The first study conducted by Mishima and Nakamura on aerobic granulation revealed a granule formation time of 3 weeks along with improved sludge settling ability as the granulation progressed.

Beun et al., (2001) reported the formation of granules in a SBR within 1 week of start by inoculation with activated sludge from a wastewater treatment plant. It was also initially observed that a mixture of granules, filaments and flocs were prevalent as was observed in the present research. Similarly, in this study, sludge primarily contained a mixture of tiny granules with predominated flocs and filaments.

A particle size distribution analysis (Figure 19) revealed the sludge particle size of 200 µm which qualified to the definition of aerobic granules. According to de Kreuk et al., (2007) the minimum size of the sludge qualifying for being considered a granule should be 0.2 mm. Liu et al., (2010) used SVP-SB200 (sludge volume percentage with size below 200µm) value below

50% as an indicator of granule dominant sludge.

Meanwhile on day 10, the settling time was reduced to 12 minutes to select for healthy rapidly settling sludge and washout fluffy biomass from the system. After 21 days of operation, the flocs disappeared and the reactor was predominated by small string like granules. The particle size of sludge gradually increased as observed by particle size analysis results (Figure 19). Semi-mature granules with a median particle size of 500µm were observed on day 35.

To ensure the presence of a healthy biomass in the granular sludge and achievement of fully mature stable granules, the settling time was further reduced to 8 minutes. Biomass selection for granules development was made by controlling settling time as one of the influential selection pressures. By changing the settling time, the particles with better settling properties were retained in the reactor while the floccular particles were washed out.

Day 60 marked the formation of mature, dense, rod-like granules with a particle size of 800 µm and SVI of 61 mL/g (Figure 19). Tay et al., (2001 b) obtained mature, homogenous and defined acetate and glucose fed granules at 3 weeks of operation. In the present study, the gradual transition of floccular sludge to granular sludge proves the importance of settling time as a selection parameter for granulation.

Beun et al., (2001) chose a settling time such that bigger particles with settling velocity greater than 10 m/hour were reserved in the reactor. As a result, the selection of biomass granules led to rapid development of granules in the reactor. Li et al., (2006) fixed a settling time of 5 minutes during the first 14 days of the experiment which resulted in washout of seed sludge followed by decreased MLSS concentration.

Garrido et al., (2004) chose a different operational strategy which resulted in a severe washout of biomass from the reactor during the start-up. A significantly short settling time and hence a fast

81 discharge time was selected for the reactor, which proved to be unsuitable specially during initial

1-2 weeks when the system is slowly stabilizing. Hence, a steady decrease in settling was chosen to preserve the biomass in the reactor.

A steady state in terms of granule formation was achieved after day 60 and thus the settling time was fixed at 5 minutes for the entire experiment as the granules showed rapid settling. At this stage, the biomass concentration in the reactor was stable at 62 mg/L. The granules grew into a compact and dense shape as the experiment progressed with time.

Rod-like granules were observed with a strong structural stability. During the initial 15-20 days of the experiment, the reactor inside walls developed an attached biomass growth which slowly diminished with time as the sludge approached granulation. The color of the sludge in the reactor transitioned from brown to white and finally to yellow when mature granules were formed. A similar color change was observed by Wang et al., (2006) and Li et al., (2006) when synthetic wastewater was used as feed.

Fig. 19 Particle size variation from day 1- day 60

4.2 Granule development Before the reactor start-up, the SBR system suffered through various system washouts. It was estimated that several factors affecting granule formation must be weighed higher in comparison to minor factors presented in the literature. These factors included evaluation of appropriate air flow velocity according to the column configuration, control of air pressure, providing an aerator which gives a uniform bubble flow pattern and control of settling time to avoid sludge washout and retention of healthy biomass.

Even though detailed research has been conducted on aerobic granulation and its applications, however, the most crucial factors contributing in the mechanism of granulation must be deeply investigated.

The reason behind sudden biomass washouts and system failures could be a result of improper engineering parameters provision, poorly understood microbiology of transitioning granules and more recently role of EPS. One of the objectives of this study was to investigate the primary engineering parameters leading to granulation. The SBR was initially (Sept-Dec 2016) setup in 3 separate phases and each phase suffered through a strong biomass washout.

The several reasons observed were sludge bulking, mechanical problems in solenoid discharging valves, excessive biomass growth, blockage of diffuser pores resulting in irregular air pattern and high superficial air velocity. Going through the past observations, one parameter was commonly involved that was predicted to cause the system failure; Superficial air velocity and air bubble pattern in the reactor.

Upflow air velocity, was deemed to be the main source of hydrodynamic shear force in bubble column reactor. Shear force was recognized as an effective parameter inducing aerobic granulation which further affects the granular structure and metabolism (Tay et al. 2001).

An increase in upflow air velocity led to a decrease in granular size due to high attrition between the particles resulting in more detachment. Increased production of cell polysaccharides was observed which contributes to the compactness and strength of the granules. A high biomass was retained in the reactor with rounder and strong granules as a result of high upflow velocity (Tay et al., 2003).

It is supposed that an increase in aeration rate leads to an increase in particle-gas-liquid interactions, and thus increases attrition on the surface of microbial particles (Pishgar et al.,

2017). In the present study, the start-up of SBR system failed initially and it was later figured out that air bubble pattern was being ignored. A pilot scale study was performed on the importance of air bubble pattern for growth of aerobic granules.

It was assumed that granulation did not occur in the pilot reactor with the lower number of diffusers due to a defective aeration pattern even if an identical specific upflow air velocity was maintained under both conditions (Pishgar et al., 2017). The modified air diffuser configuration led to a well-distributed air bubble pattern that could completely cover the cross-sectional area of the pilot reactor at different heights.

This explains the start-up failures initially in the lab-scale reactor (R1) when the aeration system was defective and the air bubble distribution was not uniform. A cubical 30µm pore diffuser was used which produced a uniform fine bubble pattern and was resistant to clogging. The upflow air velocity was lowered down from 3 cm/sec to 1.7 cm/sec which resulted in a smooth aeration pattern. Lowering the upflow air velocity was necessary to achieve a uniform aeration pattern considering the small cross-sectional area of the reactor.

The control of air pressure was also a viable parameter which led to distinctive aeration pattern, and thus affected aerobic granulation . An air pressure of 3 psi was adopted (controlled through a pressure valve) which further contributed to granule formation. The results of another study

(unpublished data) showed that substantially lower air pressure (below 5 Psi) could lead to the most homogenous aeration pattern, and thus faster and steadier granulation (Pishgar et al., 2017).

Settling time, as mentioned earlier, was chosen as selection parameter to retain healthy sludge in the reactor. It was initially set at 20 minutes and changed to 12 minutes, 8 minutes and finally 5 minutes. This strategy has been practiced by few other researchers as the rate of sludge discharged in SBR is dependent on the settling time adopted (Beun et al., 2001; Li et al., 2006).

Several other minor factors like cleaning the diffuser weekly, cleaning the solenoid valves, regularly changing feed and effluent pipes to avoid losses through cracks, controlling timer accuracy, frequent feed preparation were practiced in order to maintain proper cycles of SBR and hence functioning of the reactor. The sustainment of these factors led to development of tiny granules within 1 week of operation, confirmed by particle size analysis of the sludge sample which revealed a size of 200 µm. A smooth transitioning of flocs to aggregates and finally to granules was observed.

Fig. 20 Activated sludge Fig. 21 Young granules particles

Fig. 22 Semi-mature granules Fig. 23 Rod-like mature granules

Fig. 24 Fully mature granules

4.3 Granular characteristics

4.3.1 Mixed Liquor suspended solids and volatile suspended solids (MLSS & MLVSS) The biomass and non-biodegradable suspended matter concentration can be expressed in terms of MLSS and MLVSS values. An increasing trend in these values over time usually indicate adequate availability of active biomass to treat the applied quantity of organic waste. In granulation, a rising trend of MLSS concentration has been strongly observed.

Infact, granulation pitches up the concentration of biomass as it ensures stable biomass being preserved in the reactor resulting from improved settling of granular sludge (Tay et al., 2002).

The rate at which the MLSS changes as the sludge approaches granulation stage can be dependent on various factors such as the type of wastewater (synthetic vs. real), initial MLSS concentration, OLR, F/M ratio (Pishgar et a., 2017).

Li et al., (2011) repoted an influencial effect of biomass loading rate or F/M ratio on the rate of granulation. A higher F/M ratio causes accelerated granulation while a lower F/M ratio results in slower granulation rate. It was further suggested to adopt a specific F/M ratio at different stages of granulation to obtain dense and stable granules (Li et al., 2011). As a result, a high F/M ratio in the early stages of granulation and a lowering it in the later stages could be a viable strategy for rapid granule formation and sustainence.

The reactor was fed with synthetic feed so the substrate concentration was controlled in comparison to the real wastewater with more chances of concentration variation. The OLR was changed step-wise from higher to lower concentration as the experiment progressed. Thus, the

MLSS concentration might also be affected by the change in the substrate loading.

This is relatable to the concept of adopting higher F/M ratio in the beginning to a lower F/M ratio in the later stages which led to development of small and stable granules (Li et al., 2011).

Likewise dense and stable granules were formed in the present research. The activated seed sludge fed to the reactor possessed MLSS and MLVSS values of 5902 mg/L and 4945 mg/L respectively.

The activated sludge was initially acclimatized for 16 hours and then resumed with regular SBR cycles. The immature granules appeared in approximately 9 days and at this stage the MLSS and

MLVSS surged. The slight increase in the values could be observed due to onset of granulation process. However, on day 14th, the reactor suffered from biomass washout due to valve blockage and hence the MLSS and MLVSS values dropped.

Although the reactor lost a certain percentage of biomass, the granulation process was ongoing.

Particle size analysis revealed the granules size of 400 µm on day 20 which confirmed the progressing granulation. A similar phenomenon was observed by Corral et al., (2004) when a major percentage of sludge washed out of the system during the first 7 days due to maintenance of very short settling time. 3 weeks later small granules appeared in the reactor with an average size of 1.05 mm and SVI of 60 mL/g.

Li et al., (2006) reported a sudden decrease in MLSS concentration due to washout following an adoption of short settling time. Despite of the washout, small granules of size 100 µm were observed in the reactor with increase in biomass concentration as the operation proceeded. The partial washout in the SBR system caused a sludge color change in the reactor from brown to white as observed by Wang et al., (2006) and Li et al., (2006).

It could be hypothesized that biomass selection took place with rejection of unhealthy sludge from the system. Thus, the reactor retained the good settling granular sludge and rejected the less settling fluffy sludge. The MLSS and MLVSS values from day 21 to day 35 remained low but stable. An average value of 2182 mg/L was observed for MLSS which suggests that the biomass or sludge was properly retained in the system alongside slight increase in the size of granules.

An increase in MLSS and MLVSS was observed on day 50th and a size of 600 µm was obtained at this point. Tay et al., (2002) reported the direct relationship between granulation and concentration of biomass. It was observed that biomass concentration in terms of MLVSS

88 increased from the stage the of granule formation to the stage of stable granules retention, signifying granulation caused an increase in biomass concentration.

Day 54th saw a considerable increase in MLSS and MLVSS. The biomass concentration almost doubled and the granules grew in size with a clear outline. Notably, the increase in biomass concentration occurred in a span of 4 days which indicates each cycle in a SBR contribute towards granulation.

The MLSS and MLVSS gradually increased as the experiment was advancing. A MLSS as high as 16,193 mg/L was acquired on day 97 indicating the growth and stability of granules in a biomass sufficient environment. This in turn leads to improvement in reactor performance because of abundant biomass being preserved in the reactor (Tay et al., 2002). The MLSS and

MLVSS values decreased slightly by day 145 but resumed to pitch up to 20,640 mL/g MLSS by the end of experiment on day 184.

The MLVSS/MLSS ratio signifies the concentration of active microorganisms present in the reactor. It is further related to the concentration of OLR provided to the system. A general decreasing trend of MLVSS/MLSS ratio was observed throughout the experiment with a few hikes. As the ratio decreased, an increase in granular size was observed. The peaks in the trend were corresponding to the days 111 and 122 when MLVSS and MLSS value were similar with small difference between them.

An increase in the size of granules was consistent. Muda et al., (2008) established the significance of the MLVSS/MLSS trend in the reactor. A decrease in MLVSS/MLSS ratio was adjacent to an increase of solid inert particles inside the granules. This was as a result of a triggered accumulation of inert particles within the granules.

The size of the granules hiked to 0.15 cm towards the end of the experiment. Stable, clearly outlined, small and dense granules were maintained in the reactor for a span of 184 days with small system upsets yet stable performance in terms of organics and nutrients removal. The

MLVSS, MLSS and MLVSS/MLSS profiles throughout the experiment can be found in Figure

25 and Figure 26 below.

MLVSS MLSS Formation of 25000 stable, dense Improvement in granules reactor performance 20000

Increase in biomass 15000 concertation

10000 Biomass washout

Biomass Biomass concentratuin(mg/L) 5000

0 0 20 40 60 80 100 120 140 160 180 200 Days

Fig. 25 Variation of MLSS and MLVSS throughout the experiment

1 0.9 0.8 0.7 0.6

0.5 Ratio 0.4 Increase in inorganic 0.3 components of granules 0.2 0.1 0 0 20 40 60 80 100 120 140 160 180 200 Days

Fig. 26 Variation of MLVSS/MLSS ratio throughout the experiment

4.3.2 Particle size The particle size distribution of the activated and granular sludge was measured to evaluate the distribution of granule size. Median diameter d(0.5) was used a representative value for particle diameter at 50% in the cumulative distribution. Particle size is an important indicator of granulation process. The transition of an unstable, fluffy and flocculent sludge particle to a dense, compact and stable granule involves a combination of various phenomenon including particle size increment.

The particles increment is a result of flocs cohesion or agglomeration due to binding forces like surface hydrophobicity and EPS. As reported by de Kreuk et al., (2007) and Li et al., (2010) a sludge particle is considered a granule if its size is greater than or equal to 200 µm. Thus, an aerobic granule is a particle with size equal to or beyond 200 µm.

Generally, an increasing trend in particle size distribution denotes granulation process. On the other hand, small and dense granules are favorable over large sized granules in context of stability. Tay et al., (2001 b) obtained glucose and acetate fed granules of size 7 mm in about three weeks. Wang et al., (2006) reported a granule size of 5 mm with brewery wastewater in about 9 weeks. These results were similar to the particle size reported by Tay et al., (2001a) and

Zheng et al., (2005).

Various technologies are available for size measurement depending on the particle size to be measured. Laser diffraction, image analysis, Raman spectroscopy etc. Laser particle size analysis system; Malvern Mastersizer 2000 was utilized to detect various size range of the particles. The returned activated sludge seeded in the aerobic SBR composed of flocculent, fluffy particles and exhibited a size of 114 µm.

As the granulation of activated sludge commenced in the SBR, the particle size of the sludge rapidly changed to 200 µm within 9 days of operation and small immature granules were obtained. Since the sludge particles attained a size of 200 µm, granulation occurred on day 9. A particle size distribution curve at different points throughout the experiment, shown in Figure 27, reflects the change in size of the granular sludge.

The transition is granule size was clear and rapid with every passing stage. A particle size of

346.54 µm was measured on day 21 which escalated to 500 µm on day 35 when semi-mature granules were formed. On day 54, a particle size of 700µm was observed and this phase marked the beginning of fully mature granules formation alongside reduction of sludge flocs. The particle size augmented rapidly after this phase as granules became stable and well-defined.

A granular size of 940 µm with fully mature discrete granules was examined on day 73 which pitched up to 1246.09 µm on day 145 when small, rock-like, dense granules were observed. This size for the fully developed granules thus continued to grow till the end of experiment on day

183 and a size of 1500 µm was observed. The following graph represents the number of particles that fall into each of the size ranges d (0.1), d (0.5) and d (0.9). d (0.1) represents 10% of particles in the sample are smaller than particle size obtained. d(0.5) represents 50% of the particles in the sample are larger than particle size obtained. Similarly, d (0.9) indicates 90% of the particles in the sample are smaller than particle size value obtained.

d (0.5) d (0.1) d (0.9)

2500

2000

1500

200 µm 1000 1500 µm

1246 µm Particle Particle size (um) 700 µm 500

0 0 20 40 60 80 100 120 140 160 180 200 Day

Fig. 27 Particle size analysis from day 1- day 183

4.3.3 SVI (Sludge volume index) Sludge volume index (SVI) is an indicator of settling and compacting characteristics of sludge and aerobic granules. Granules settle significantly faster in comparison to activated sludge thus their settling characteristics vary from standard activated sludge. In the case of granules, SVI 5

SVI 10 and SVI 30 (or SVI) terminologies have been commonly reported in the literature because of a faster settling rate (Beun et al., 2002; Schwarzenbeck et al., 2004; de Kreuk et al.,

2005).

In the present research, changing trend in SVI 5, SVI 10 and SVI 30 have been reported. SVI 5 indicates the sludge volume index at 5 minutes and SVI 10 indicates sludge volume index at 10 minutes. SVI values in the range 100-200 mL/g are representative of the sludge in most of the activated sludge plants producing a clear, good-quality effluent whereas for granular sludge an

SVI5 range of 30-60 mL/g has been reported (Schwarzenbeck et al., 2004).

According to de Kreuk et al., (2005) a combination of SVI at 10 and 30 minutes should be adopted to ascertain settleability characteristic of the granular sludge. The difference in the values of SVI10 and SVI30 indicates formation of granules and sludge compacting during and after settling. A trend in SVI of sludge gives an excellent indication about granule formation and growth.

A decreasing trend of SVI with time reflects the stage of sludge approaching granulation with extremely low values of SVI indicating complete granulation. The aerobic granules possess a higher settling velocity in comparison to activated sludge flocs, thus when the flocs turn into aggregates, their settling velocity becomes higher leading to faster settling.

A settling velocity higher than 10 m/hour has been reported for aerobic granules (Beun et al.,

2002) and therefore a maximum of 5 to 10 minutes settling is required for mature granular sludge to reach a terminal SVI due to faster sludge bed consolidation (Schwarzenbeck et al., 2004). The dynamics of SVI30/SVI5 ratio as hypothesized by Erşan and Erguder (2014) could also ascertain evolvement of granules.

SVI30/SVI5 ratio in the range 0.68-1 has been reported in granular reactors where mature granules emerged after 15 – 20 weeks of operation (Schwarzenbeck et al., 2004; Schwarzenbeck et al., 2005). In the present study, the sludge volume index value decreased progressively as the sludge approached granulation through 183 days run.

The SVI value of seeded activated sludge dropped from 166 mL/g to 138 mL/g during the first two weeks of operation during which initial granules were observed. During the third week, the system suffered from a minor washout due to which the settling property of the sludge degraded and a rise in SVI was observed until day 35. Sludge bulking occurred as a consequence of washout and hence the surge in SVI was noticed.

The increase in SVI could be attributed to a number of other related factors including poor hydrodynamic shear force, improper air flow pattern in the reactor and blockage in diffuser.

Despite the sludge bulking, granular growth was observed which was reflected from particle size analysis results during the third and fourth week.

Following day 54, the SVI rapidly decreased as fully mature aerobic granules were formed at this stage. This was relatable to a corresponding increase in MLSS concentration and particle size distribution. The decreasing trend in SVI was observed throughout the end of experiment as stable aerobic granules were maintained in the system for 184 days. SVI value of 22 mL/g was obtained on day 82, 23 mL/g on day 112 and as low as 9 mL/g was observed on day 184 of the experiment.

The SVI30/SVI5 ratio, as conceptualized by Schwarzenbeck et al., (2005) was found to be an average value of 0.78 in the present research which complies to the reported range of 0.68-1

(Schwarzenbeck et al., 2004; Schwarzenbeck et al., 2005). However, a high value of 0.86 for

SVI30/SVI5 was observed during and after 8th week of operation since mature, stable granules were progressing.

A high value indicates the granular sludge is dense and compact with better settling granules predominating in the reactor. The trend of SVI (SVI30, SVI10 and SVI5) and SVI30/SVI5 is demonstrated graphically in Figure 29, 30, 31 and it can be observed that SVI dropped gradually over time as the settling property of the sludge improved.

Fig. 28 Sludge settling through 1L flask for SVI test

180 160 140 120 100

80 SVI (mL/g) SVI 60 40 20 0 0 20 40 60 80 100 120 140 160 180 200 Day

Fig. 29 SVI (SVI30) variation from day 1- day 183

SVI5 SVI10

350

300

250

200

150

SVI (mL/g) SVI SVI5- 100 SVI5- 22mL/g 9 mL/g

0 0 20 40 60 80 100 120 140 160 180 200 Day

Fig. 30 SVI (SVI5 and SVI10) variation from day 1- day 183

1 0.9 0.8 0.7

5 0.6

/SVI 0.5 30 SVI30/SVI5= 0.86 SVI 0.4 Mature granule formation 0.3 0.2 0.1 0 0 20 40 60 80 100 120 140 160 180 200 Time (d)

Fig. 31 Variation of SVI30/SVI5 in the experiment

4.3.4 Morphology The initial returned activated sludge fed to the reactor exhibited a typical loose, fluffy and irregular morphology with predominant filamentous organisms (images not shown). The sludge was brown in color (Figure 20) which slowly changed to white and later yellow during the operation. The dispersed structure of the sludge gradually turned to small aggregates type morphology during the 1st week of operation. The small brown colored aggregates were spherical and defined with improved settling property.

During the 3rd week, the brown spherical aggregates turned off-white with a slight increase in the size. These spherical immature granules were present along with flocs and were dense as well as stable. Although the biomass concentration dropped at this stage, the development of granules was an ongoing phenomenon. As the size of the granules kept on increasing, they became rounder and rigid.

On day 60, the granules grew to a size of 800µm and possessed spherical, dense and compact shape with good settling property. The granules were white in color and were uniformly distributed in the bioreactor due to the superficial air velocity provided. Tay et al., (2009) established a relationship between structure of granule and hydrodynamic shear force.

It was found that the aspect ratio or roundness of acetate-fed granules increased with a surge in superficial air velocity. On the other hand, the density and strength of granule were also in direct relation with shear force. Thus, it was established that the granular structure is regulated by hydrodynamic shear force maintained in the reactor.

For the present study, a superficial air velocity of 1.7 cm/sec was fixed throughout the experiment. Since the biomass concentration pitched up significantly on day 54, the granules grew bigger in size and acquired flaky shape with dense structure. 60% of the granules were spherical but the remaining 30% which were presumably flocs gained rice-like morphology. At this stage, the color of the sludge in the bioreactor started turning to yellow.

The sludge was properly retained in the system and thus the granules grew to a size of 1200 µm until day 97. Clearly-defined, dense, spherical and rod-like granules were observed (Figure 32).

Since the hydrodynamic shear force was kept constant throughout the experiment but the OLR was changed hence the roundness of the granules was affected. However, the change in morphology did not affect the structural strength of the aerobic granule and stable granules were retained in the bio system.

A cross section along the diameter of the granule was cut and it was found that the granule was substrate loaded and not hollow inside. The granules retained the same shape with a combination

99 of spherical and rod-like granules for a few days until day 123. At this stage, the substrate load was reduced and hence the increment in the size of granules was small.

The granules persisted to grow, but the growth rate declined due to reduction in organic loading which favored in transformation of granules to a spherical and denser morphology. The change in shape could be attributed to that fact that the shear force per granule surged as the organic loading deceased. And as the hydrodynamic shear force rose, the granules suffered more attrition and collision which further lead to detachment of loose fragments from granules.

As observed by Tay et al., (2003) the aspect ratio or roundness of granules was directly dependent of increase in hydrodynamic shear force. Since the shear force per granule pitched, the morphology improved creating spherical, dense and stable granules. The granules continued to become circular and similar shape was observed on day 157 with more density of the granule shown by excellent settling property.

The granule was uniformly loaded with substrate which was proved by dissecting a granule along its diameter to observe the interior cores. The substrate was well diffused in the granule and there were no hollow cores. The granules attained a rock-like structure and were slightly brown, dense and possessed clearly defined shape. The morphology was maintained throughout the end of experiment till day 184 and the granules were confined and rock-like. And finally, small, stable, well-defined, spherical and dense aerobic granules were obtained through a span of

7 months.

It can be thus concluded that the development of granules was an evolution process from flocs to aggregates and finally to granules with each phase possessing its particular shape. The mature granules further change shapes in response to changes in substrate loading, hydrodynamic shear

100 force and other working conditions. However, once the granules are acclimatized to the system, the morphology remains almost constant unless major changes are made in the system.

Fig. 32 Spherical and rod-like mature granules

101

CHAPTER 5- ORGANICS AND AMMONIA REMOVAL

102

+ 5.1 COD and NH4 removal The layered structure of aerobic granules due to oxygen transfer limitation allows organics removal, nitrification and denitrification to take place simultaneously. It has been hypothesized that the outer aerobic layer supports COD oxidation and nitrification whereas the inner anoxic/anaerobic layer favors denitrification. Beun et al., (2001) showed the potential of aerobic granular sludge to remove COD, cause nitrification and denitrification with optimum DO value of 40%.

A high biomass retention in the aerobic granules selects for slow growing organisms (Winkler et al., 2012). The selection of slow growing organisms was maintained by selecting high substrate

N/COD ratio which resulted in a lower specific growth rate and a higher nitrification activity

(Liu et al., 2003). This in turn improves the stability of granules by selecting for small sized granules having more compact structure.

Yang et al., (2003) investigated the capability of hybrid aerobic granules to simultaneously remove organics and nitrogen from wastewater. A complete removal of organic carbon and nitrogen was observed in the SBR based system with a stable operation. Different N/COD ratios

(5/100-30/100) were adopted to develop microbial granules and it was found that nitrifying, denitrifying and heterotrophic populations co-existed in the granules.

The influent COD of 500 mg/L was removed by a value of around 95% and complete conversion of ammonium nitrogen to nitrate was observed under aerobic conditions (Yang et al., 2003).

Further for denitrification, DO concentration and mixing were the two factors affecting the efficiency of N removal. A DO concentration of 0.5 mg/L and mixing to ensure contact between granules and soluble nitrate caused a complete denitrification (Yang et al., 2003).

103

Similarly, Kreuk et al., (2005) worked on simultaneous organics and nutrients removal by aerobic granular sludge. Low oxygen concentration yielded high COD and N removal of 100% and 94% respectively. The nitrogen removal was additionally found to be dependent on diameter of the microbial granules.

Coma et al., (2012) obtained granules treating domestic wastewater concurrently removing nutrients in two SBRs. One was 100% floccular and the other was seeded with 10% crushed granules (95% floccular). The study concluded the capability of a granular system to cause nutrients removal (nitrification, denitrification and phosphorus removal) in VFA limited domestic wastewater.

In the present study, the influent and effluent samples were analyzed for soluble COD, VFA,

+ NH4 to determine the removal efficiencies obtained through the granular sludge developed in the SBR. Different organic loading rates were applied at various stages of the experiment to access the capability of aerobic granules to treat organics and nutrients at each loading rate. The organic loading rate in terms of COD was analyzed in the SBR. The nitrogen loading rate was

+ measured as NH4 for influent and effluent samples. The DO content was not controlled in the reactor and the saturation was close to 100%

5.1.1 COD removal The COD removal and VFA profiles were analyzed through the end of the experiment. The samples were filtered through 0.45 µm filter and soluble COD and VFA were measured. The influent COD, effluent COD and removal efficiencies were analyzed and are reported below

(Figure 33). A hydrodynamic shear force of 1.7 cm/sec was adopted with a ceramic diffuser pumping out uniformly distributed air pattern in the reactor.

104

The operational conditions were kept constant as mentioned above in Chapter 3. To monitor the effect of changing organic loading rate on COD removal efficiency, different OLR values were adopted. The COD removal efficiency was then monitored corresponding to each OLR. Figure

33 shows the variation of influent COD with time. The OLR in terms of substrate was varied at different stages of the experiment to access the capacity of granular sludge in removing COD, nitrogen and phosphorus. A gradual variation was made at every stage without shocking the granular system.

A COD of 2000 mg/L was opted initially for a span of 35 days to foster the development of granules and obtain stable average sized granules. The feed was prepared in reference to Tay et al., (2002c). After the granules obtained, stabilized and acclimatized the system, the OLR was increased and a COD of 3000 mg/L was provided to the system. The granular sludge rapidly responded to the increased COD and effective removal efficiency was obtained.

This COD was maintained from day 36 until day 60 and during the 26 days, the granular sludge showed high degree of organic C removal rendering an average removal efficiency greater than

96%. Day 61-Day 91, the substrate load was raised with an influent COD of 4000 mg/L in order to determine the high organics removal tendency of the granules. The stabilized granules in the system adapted to the high COD, responded well to the increased substrate load and could remove the organics in the synthetic wastewater.

During the high substrate period, an average COD removal efficiency of 99% was observed. This explains the superior COD removal efficiency of aerobic granular sludge even with elevated loads. At this point, in order to achieve municipal wastewater substrate levels, the COD was gradually decreased from 4000 mg/L to 2000 mg/L to 1500 mg/L to 1000 mg/L finally to reach a

105 value of 500 mg/L. The gradual decrease was adopted to avoid shock loadings to the system so that the granules could acclimatize to the changing COD.

Day 92-99 an influent COD of 2000 mg/L was maintained which was further changed to a COD of 1500 mg/L from day 100 to 120. The removal efficiencies obtained during this phase, were an average of 98.69% which again proved the granular sludge ability to effectively remove organics present in wastewater. A COD of 1000 mg/L was adopted from day 121 to 134.

Finally, from day 135 to 183, municipal level COD of 500 mg/L was chosen to evaluate different concentrations of phosphorus that could be removed by the aerobic granules. Figure 33 represents the effluent concentrations of COD at various points of the experiment corresponding to different influent concentrations. All the values reported are quite low in comparison to the influent COD values rendering a high average removal efficiency. Figure 33 denotes COD removal efficiency throughout the experiment.

The removal efficiencies, independent of the OLR, were excellent and above 90% on average.

The acetate used as an external C source in the substrate was completely utilized for assimilation for biomass growth in the granular reactor. This explains the almost complete degradation of influent COD by the granules.

106

COD Influent COD Effluent COD removal efficiency 5000 100

OLR= 1.6 kg COD/m3 day 4000 Granulation: 80 OLR=6.4 kg COD/m3 d 3000 60 OLR=12.8 kg COD/m3 d

COD (mg/L)COD 2000 40

C:N:P 100:4.5:0.5 (%) efficiency Removal 1000 20 OLR=9.6 kg COD/m3 d

0 0 0 20 40 60 80 100 120 140 160 180 200 Time (d) Fig. 33 COD analysis

+ 5.1.2 Ammonia (NH4 ) Removal profiles Ammonium levels in influent and effluent were monitored to analyze the nitrogen removing ability of the aerobic granules. Beun et al., (2001) reported N-removal by monitoring the feast

+ and famine periods in SBR. It was observed that NH4 concentration gradually declined during the feast period and the NOx underwent denitrification due to the presence of acetate. After the

+ depletion of acetate, the NH4 concentration decreased rapidly in comparison to feast period and this was as a result of nitrification.

During the famine period, denitrification was achieved due to decrease in NOx concentration. It was further concluded that a creating a balance between nitrification and denitrification by controlling DO in the reactor would result in maximum N removal. Dangcong et al., (2001)

107 evaluated that the ammonium present in the feed was totally nitrified with nitrate being the only end product. A total N removal efficiency of 75% was obtained in the experiment.

In the present experiment, the ammonium levels were changed corresponding to COD levels to maintain the right C:N:P (COD: NH3-N: P) ratio. In order to maintain the C:N:P ratio of

100:4.5:0.5, the influent ammonium levels were raised alongside COD and P. Figure 34 shows the influent ammonium levels at different days throughout the experiment. The graph follows a

+ similar trend as influent COD due to proportionate increase in NH4 .

Day 1 to day 35 an influent ammonium concentration of 92 mg/L was maintained as a COD of

2000 mg/L was adopted. This ammonium concentration was changed to 138 mg/L between day

36 and day 60, when the COD was raised to 3000 mg/L. Further, during day 61-91 the ammonium concentration was raised to a value of 184 mg/L in order to analyze the granular sludge capacity to remove elevated concentrations of organics and nutrients.

Following day 92, the substrate load was gradually lowered to reach a concentration close to

+ municipal wastewater levels. NH4 concentration of 92 mg/L was adopted until day 99 which was subsequently changed to 69 mg/L from day 100-day 120. This was followed by an adoption of 46 mg/L corresponding to a COD of 1000 mg/L until day 134. And finally, the ammonium level was set at a constant value of 23 mg/L until the end of experiment, to mimic the concentrations present in municipal wastewater.

The effluent ammonium concentrations were satisfactorily low with an average ammonium concentration of 7.87 mg/L (Figure 34). Figure 34 demonstrates ammonia profiles in the influent, effluent and the removal efficiencies. The removal efficiencies were in the range 75%-99% with

+ high removal efficiency at lower influent COD and NH4 values.

108

The results indicate the potential of aerobic granular sludge to treat different ammonium concentrations with an average removal efficiency of 90.76%. This also shows the potential of aerobic granulation to treat ammonia with the presence of single aerobic phase in comparison to consecutive anoxic and aerobic phases or regulating DO levels provided in the literature (de

Kreuk et al., 2005, Coma et al., 2011, Beun et al., 2001).

+ In the present experiment, during the stage of granule development, an average NH4 removal

+ efficiency of 89.28% was obtained. On increasing the substrate load with influent NH4 of 138 mg/L, the removal efficiency pitched up to 92.6 %. During the next phase, when influent ammonium was maximum with a value of 184 mg/L a removal efficiency of 91.88% was obtained.

The substrate load was reduced at this point and lower ammonium concentrations were adopted in the influent. The removal efficiency dropped to 67.42% upon lowering the COD to 1500 mg/L and it increased to 85.87% at a COD of 1000 mg/L. And finally, when concentrations were lowered down to municipal levels maintaining a COD of 500 mg/L, the removal efficiency improved to 95.85%.

109

NH3-N Influent NH3-N Effluent NH3-N removal efficiency 250 100

200 80 Highest removal efficiency at 150 lowest COD and 60 NH3-N loading

rate

N (mg/L) N -

100 40 NH3 + Effluent NH4 < 7.87 mg/L (%) efficiency Removal 50 20

0 0 0 20 40 60 80 100 120 140 160 180 200 Time (d) Fig. 34 Ammonia concentration analysis during the experiment

5.2 Discussion The hypothesis of the presence of anaerobic zone in the interior core of aerobic granular sludge results in integrated single phase aerobic nitrogen removal. The simultaneous nitrification and denitrification occurs due to the presence of anoxic and aerobic zones in the granule thus leading to an effective nitrogen removal in one phase.

Winkler et al., (2012) suggested the presence of various bacterial populations in different layers of granule. Nitrifiers were hypothesized to be present in the outer most oxygen rich layer whereas denitrifiers and PAOs were assumed to be located in the inside anoxic zone (Winkler et al., 2012). The majority of COD and ammonia removal using aerobic granular sludge in this study is hypothesized to be due to assimilation as a low COD: NH3-N :P was maintained.

110

Monitoring the ammonia levels in the influent and effluent, it can be clearly stated that aerobic granules showed a good potential in ammonia removal. An average removal efficiency of

90.76% was obtained through the 183 days of experimental run. The influent ammonia concentration was varied at several points along the experiment and the effect on effluent ammonia concentration was observed.

The parameters contributing to the effective ammonia removal were identified throughout the experiment. Some of the engineering parameters including hydrodynamic shear force, pH and temperature were maintained constant. They were chosen based on the findings in literature and thus proved effective.

A hydrodynamic shear force of 1.7 cm/sec was maintained through the entire experiment which further helped in controlling the granular size and density. These two factors are known to affect the nitrification and denitrification rate thus affecting nitrogen removal rate (Ning et al., 2011). A steady granular size and density maintains the volume of aerobic and anoxic zones inside the granule which majorly controls the nitrogen removal process in an aerobic granule.

The hydrodynamic shear force did not directly affect the nitrogen removal efficiency. But its dependency on sub factors (granular size and density) that maintain the micro-zones inside the granule influence the ammonia removal process. The pH was maintained neutral value of 7 and it did not have a significant impact on nitrogen removal efficiency. The pH value was maintained constantly neutral as a very high (alkaline) or a very low (acidic) pH could negatively affect the nutrient removal efficiency (Lashkarizadeh et al., 2016).

The reactor was setup at a constant room temperature of 20oC. The temperature although was found independent of ammonia removal efficiency in this study but affect the settling capacity

111 and morphology of the granules. Zhiwei et al., (2008) studied the influence of temperature on ammonia removal and found out the increase in temperature positively affects the speed of nitrification and denitrification ability of granular sludge. The ammonia removal efficiency in the reactor thus increased from 68% to 82% with a temperature increment from 25oC to 30oC.

The C:NH3-N ratio of 100:4.5 was kept constant and the OLR was varied between 1.6-12.8 kg

COD/m3day. COD values of 500 mg/L, 1000 mg/L, 1500 mg/L, 2000 mg/L, 3000 mg/L and

4000 mg/L were adopted corresponding to each value of OLR. Figure 35 illustrates the effect of

OLR on ammonia removal efficiency in aerobic granular sludge. It can be observed that the highest removal efficiency of ammonia was obtained at the lowest OLR of 1.6 kg COD/m3 day.

OLR vs NH3-N removal efficiency 120

100

N removal Nremoval efficiency -

20 NH3

0 0 1.6 3.2 4.8 6.4 8 9.6 11.2 12.8 OLR (kg COD/m3 day)

Fig. 35 Effect of OLR on ammonia removal efficiency

The ammonia removal efficiency gradually dropped until an OLR of 4.8 kg COD/m3 day was adopted and from this point steadily rose with increasing OLR. At an OLR of 6.4 kg COD/m3 day, a removal efficiency of 89% was achieved. Further at OLR of 9.6 and 12.8 kg COD/m3 day, removal efficiencies of 92% and 91% were obtained respectively. A minimal significance of

112

OLR on ammonia removal can be observed as a break point (OLR 4.8 kg COD/m3 day) can be seen with lowest removal efficiency.

Analyzing the general trend, it can be concluded that a high OLR (ranging from 6.4-12.8 kg

COD/m3 day) favors the simultaneous organic and nitrogen removal. It can be hypothesized that the nitrification occurs in the hypothesized aerobic layer where nitrifiers are assumed to exist.

While the denitrification occurs in the inner anoxic layer of a granule (Winkler et al., 2012)

The C:N ratio was kept constant at 100:4.5 throughout the experiment. This nutrient ratio was provided to ensure a maximum functionality of microorganisms for achieving optimum efficiency. It can thus be assumed that majority of the ammonia removal was as a result of biological assimilation. Since the N/COD ratio maintained was low, it can be estimated that nitrogen was being consumed for cellular growth of granular sludge bacteria.

However, the effective removal efficiency of ammonia via granular sludge indicates the role of additional phenomenon. The microbiological sequencing of granular samples indicated the dominant presence of denitrifying bacteria Rhodobacteraceae (Fig 36). Its presence in the granule could suggest the nitrogen removal phenomenon. Rhodobacteracea besides being an anaerobic denitrifying bacterium promotes attachment due to the EPS secretion and thus add to the aerobic granules formation process (Lv et al., 2014).

113

Flavobacteriaceae, Unclassified, 10% 6%

Rhodanobacteraceae, 7%

Rhodobacteraceae, 23%

Rhodocyclaceae, 30%

Fig. 36 Microbial composition of granules examined through Polymerase chain reaction of extracted 16r RNA genes

Zhao et al., (2013) found the strains of heterotrophic denitrifiers in aerobic granules developed at high COD and N loadings. In the present study, the presence of Rhodobacteracea led to the internal denitrification in the granules yielding low effluent nitrogen levels as well as promoted attachment for granules formation.

The oxygen gradient inside the granules and the formation of aerobic-anoxic-anaerobic layers could also presumptively support the removal of ammonia. A control in D.O concentration has been proved effective in N removal process. A distinguishable effect of low D.O concentration has been reported in simultaneous organics and N removal (Liu et al, 2003, de Kreuk et al., 2005,

Dancong et al., 2001).

The low D.O concentration proves beneficial for denitrification as a higher denitrifying bacteria activity is observed at a low D.O concentration (Liu et al., 2003). Beun et al., (2001) reported the control of D.O led to the small penetration depth of oxygen and simultaneously the concentration

114 of oxygen was just appropriate to cause nitrification in the outer aerobic layer of granule.

Ultimately an anoxic zone was formed in the centre of granule supporting denitrification process.

A similar phenomenon could have taken place in the present research since denitrifying bacteria was dominantly located inside the granule. The effluent samples were analyzed for nitrate and considerably low concentrations of nitrate were found (Figure 37). This implies that the denitrification phenomenon occurred inside the granule which led to the conversion of nitrate to nitrogen gas.

Although, nitrogen gas levels were not monitored from the bioreactor, but the drop in ammonia and nitrate levels in the effluent signify the probable nitrification-denitrification in granules. The outermost aerobic layer in the granule might have supported the nitrification reaction causing conversion of ammonia to nitrite and further to nitrate. The anoxic layer provided an environment for denitrification reaction leading to reduction of nitrate to nitrogen.

This fact could be supported by the presence of denitrifying bacteria found in microbiological analysis. Additionally, the nearly negligible levels of nitrate in the effluent could signify the probable conversion of nitrate to nitrogen gas. Figure 37 represent the nitrite and nitrate levels found in the effluent. Influent ammonia concentration between 23 mg/L-184 mg/L were adopted.

Even at higher levels of influent ammonia, an average removal efficiency of 91% was obtained proving that granular sludge has the capacity to treat nitrogen in a single aerobic phase.

A removal efficiency of 95% was achieved at an influent ammonia concentration of 23 mg/L.

Thus, the treatability of ammonia present in municipal wastewater using aerobic granulation is efficient. In depth studies on granular microbiology will reveal the characteristics of microorganisms causing nitrification and denitrification within different layers of granule.

115

Nitrate Effluent Nitrite Effluent 10

8 No nitrite or nitrate 7 accumulation 6

3 Effluuent concentration concentration (mg/L)Effluuent 2

0 0 20 40 60 80 100 120 140 160 180 200 Time (d)

Fig. 37 Nitrite and Nitrate concentration in the effluent

116

CHAPTER 6- PHOSPHORUS REMOVAL

117

6.1 Phosphorus removal profiles Oxygen transfer restrictions in the granule leads to the formation of microbial stratification which makes favorable conditions inside the granule for satisfactory nutrients removal. Since it is known that simultaneous nitrification and denitrification can occur within the same granule, a part of phosphorus accumulating organisms have the capability to accumulate polyphosphate under anoxic conditions.

Therefore, denitrifying and phosphorus accumulating organisms (DPAOs) can be flourished in the granule for the purpose of phosphorus removal in the presence of less oxygen (Coma et al.,

2011). Biologically induced precipitation was discovered to be an accountable process causing

45% of the total P removal under selected operating conditions (Angela et al., 2011). In the

3- present study, phosphorus in terms of PO4 was analyzed for influent and effluent values.

Phosphorus removal is one of the main focus of this study and hence different concentrations of phosphorus were adopted at various points throughout the experiment to establish the capability of aerobic granules to treat phosphorus using a simple aerobic phase. Eliminating addition of anaerobic or anoxic phases, VFA addition, DO control or idle period as adopted in literature

(Angela et al., 2011, de Kreuk et al., 2005, Coma et al., 2011), this study focuses to simplify the phosphorus removal process in aerobic granulation.

The phosphorus levels were initially set so as to maintain C:N:P ratio and later on varied, fixing

C:N ratio at municipal levels. The influent phosphorus concentration was varied in 10 phases

(Figure 38). During phase 1, day 1-35, the influent phosphorus concentration was set at 11 mg/L with a phosphate concentration around 30.38 mg/L. This phosphorus concentration was set corresponding to a COD of 2000 mg/L.

118

The phosphorus concentration was raised in phase 2 with an average phosphate concentration of

45.58 mg/L (P= 14.8 mg/L) from day 36-day 60. The phosphorus concentration during the phase

3 (day 61-91) was further increased to a phosphate concentration of 60 mg/L (P= 19 mg/L) at a

COD level of 4000 mg/L. Day 92-99 the substrate load was lowered at a COD of 2000 mg/L, during when the phosphate concentration of 30.38 mg/L (P=11 mg/L) was fixed again.

Further, during phase 5, the phosphate was lowered to an average of 25 mg/L (P= 8.15 mg/L).

During phase 6 (day 121-day 134), the COD was altered to 1000 mg/L and the phosphate was set at 16.87 mg/L (P= 5.5 mg/L). Finally, to achieve municipal levels of wastewater concentrations, a COD of 500 mg/L was fixed and various phosphorus concentrations (2.75 mg/L – 16.5 mg/L) were adopted to check the potential of aerobic granular sludge in removing the different concentrations.

The phosphorus concentrations were gradually increased from 2.75 mg/L to 16.5 mg/L at a constant COD of 500 mg/L. During phase 7 (day 135-144), the influent phosphate concentration of 8.43 mg/L (P= 2.75 mg/L) was selected. Day 145-157 phase 8, the phosphate concentration was raised to 16.87 mg/L to access removal of 5.5 mg/L of phosphorus. The phosphate concentration was surged furthermore to 25 mg/L (P= 8.15 mg/L) in phase 9 and lastly a phosphate concentration of 45.58 mg/L was chosen corresponding to a phosphorus concentration of 16.5 mg/L during phase 10 (Figure 38).

Figure 38 represents the effluent phosphate concentrations corresponding to the different influent concentrations adopted. Given that the phosphate concentration in the range 8.43 mg/L- 60 mg/L were adopted, all the effluent concentrations were found to be below 9 mg/L. The granular sludge exceedingly showed the potential of removing phosphorus with the presence of single aerobic phase reducing the complexity of traditional phosphorus removal processes.

119

Figure 39 shows the removal efficiency of phosphate throughout the experiment. Evaluating the efficiencies at different phases of the experiment as per the influent phosphate concentration adopted, it can be found that during phase 1 the average removal efficiency obtained was

96.16%. This phase consisted of start-up and granule development with moderate influent phosphorus levels, leading to a high phosphate removal efficiency.

During phase 2, when the organic carbon load as well as influent phosphate was increased, almost complete removal of phosphate was observed rendering a high average removal efficiency of 99.59%. A similar trend was observed during phase 3, where a high COD of 4000 mg/L and high influent phosphate were adopted. An average removal efficiency of 98.72% was achieved in this phase. After this stage, the substrate load was reduced to gradually achieve municipal level loads.

During phase 4 a high removal efficiency of 99.62% was observed and similarly in phase 5,

99.55% of influent phosphate was effectively removed. As the substrate load was reduced, the influent phosphate concentrations were lowered and thus during phase 6 and phase 7, a removal efficiency of 99.62% and 99.39% respectively was observed. At phase 7, the organic carbon load was fixed and phosphorus was varied in different phases to determine how granular sludge could handle the increasing phosphorus loads at a municipal level COD and N.

During phase 8, 98.17% average removal efficiency was observed. However, when the phosphorus load was surged to an influent phosphate concentration of 25 mg/L, the removal efficiency dropped and an average value of 93.32% was rendered. Similarly, when phosphorus load was further raised to an influent phosphate of 45 mg/L during phase 10, the removal efficiency of 79.27% was obtained.

120

Phosphate influent Phosphate effluent

100 90 80 Fixed C:P Variable C:P 70 60 Granule 50 maturation 40

Phosphate (mg/L)Phosphate 30 20 10 0 0 20 40 60 80 100 120 140 160 180 200 Time (d)

Fig. 38 Phosphate influent vs phosphate effluent concentration analysis with time

100

90 High removal efficiency at high 80 influent COD and P concentrations 70 60 50 40 30 Removal Removal efficiency (%) 20 10 0 0 20 40 60 80 100 120 140 160 180 200 Time (d)

Fig. 39 Variation of phosphate removal efficiency with time

121

6.2 Discussion The important aim of the study was to determine the potential of aerobic granular sludge to remove phosphorus levels present in municipal wastewater. As mentioned in the previous

+ 3- sections, high removal efficiencies for COD, NH3 and PO4 were obtained throughout the experiment. It was discovered that phosphorus removal was rather a simple phenomenon observed in the aerobic granular sludge.

A single aerobic phase was adopted in the SBR with approximately 2 hours of continuous aeration. The inclusion of anaerobic or anoxic phase was eliminated to identify the sole potential of aerobic granules to treat phosphorus considering their stratified aerobic-anoxic-anaerobic structure. The identification of various operational parameters that caused the effective phosphorus removal would be discussed in the following section.

The successful application of phosphorus removal mechanism under this study and its further research could simplify phosphorus removal in wastewater treatment at municipal level. Aerobic granules are known to contain certain functional group of bacteria such as nitrifiers, denitrifiers and phosphorus accumulating organisms concentrated in specific zones.

As a result of substrate gradients, shear stress phenomenon and protozoa presence on the outer granule layers, ecological niches are developed supporting species diversity. It has additionally ben reported that different sections of sludge bed in the column selects for different microbial population. The substrate gradient is further expected to form more ecological niches leading to an increase in biodiversity (Winkler et al., 2012).

The various conversions of C, N and P occur within different layers of granular biomass due to segregation of bacterial populations in the granules (Winkler et al., 2012). Simultaneous COD, N

122 and P removal is achieved by different biotransformation mediated by numerous microbial groups existing in aerobic granules (de Kreuk et al., 2005).

6.2.1 Biological assimilation An influent C:N:P ratio of 100:4.5:0.5 was maintained during major part of the experiment.

Evaluating the effluent results, it was found that similar C:N:P ratio occurred for the effluent carbon, nitrogen and phosphorus values. At each stage when the OLR was changed, the C:N:P ratio was kept constant and the removed concentrations of COD, nitrogen and phosphorus were analyzed which reflected that the ratio of consumption rate was equivalent to influent substrate rate.

This suggests that the organics and nutrients were dominantly taken up by the microorganism for their cellular needs and growth. Granular sludge has the capacity to retain high biomass and therefore the assimilation capacity per granule peaks up absorbing more substrate. The metabolism was primarily responsible for phosphorus uptake when OLR between 12.48 kg

COD/m3 day – 3.12 kg COD/m3 day was adopted.

Phosphorus removal efficiency above 95% was attained elucidating complete assimilation of P by granular sludge biomass. Influent phosphate values as high as 68 mg/L were removed by the metabolism process in aerobic granules at high OLR. However, when municipal level concentrations were adopted in the influent at an OLR of 1.6 kg COD/m3 day and higher C:P, phosphorus removal mechanism in addition to biological assimilation was observed.

The COD was fixed at 500 mg/L and phosphorus concentration was gradually increased from

2.75 mg/L to 16.5 mg/L. A C:P ratio of 100:0.5 implies a COD and phosphorus concentration of

500 mg/L and 2.75 mg/L respectively was required for metabolic removal of P via granular

123 sludge. When the influent phosphorus was raised from 2.75 mg/L to 5.5 mg/L, at C:P ratio of

100: 1.1, a removal efficiency of 98.17% was obtained.

Further the influent phosphorus was fixed at 8.25 mg/L at C:P ratio of 100: 1.65 which rendered a removal efficiency of 82.77%. The influent phosphorus in the next phase was increased to 16.5 mg/L at C:P ratio of 100:3.3 and a removal percentage of 79.27% was obtained. When a C:P ratio of 100:4.5 was adopted with an influent phosphorus concentration of 22 mg/L, no phosphorus removal was observed.

The granules developed at high substrate P/COD ratio were found to be saturated with phosphorus accumulating bacteria. Infact a high P/COD ratio in activated sludge process is known to enhance the growth of PAOs and improve their phosphorus storage capacity. For the granules developed at P/COD ratio of 1/100, the elemental composition obtained was similar to the basic prokaryotes as the phosphorus provided was consumed for the metabolism process only.

Biological accumulation (90%) was dominantly responsible for the phosphorus accumulation as minimal precipitation was found in the granules. Comparable results were observed in the present research in terms of output. However, the operational mode involved single aerobic phase instead of two: anaerobic and aerobic phases in order to relate an EBPR process.

Liu et al., (2005) investigated the effect of P/COD ratio on the composition of aerobic granules accumulating phosphorus. An anaerobic aerobic sequential operation was maintained in the reactors and P/COD of 1/100, 2.5/100, 5/100, 7.5/100 and 10/100 was kept. The P content in the microbial granules was estimated and it was found that a high P/COD ratio led to a higher accumulation of P in the aerobic granules.

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It can be clearly interpreted that substrate ratio of 100:0.5 and 100:1 involved P utilization for growth of microorganisms as seen in study conducted by Liu et al, (2005). This explains the almost complete removal of P and high removal efficiency achieved from phase 1 through phase

Besides a high COD and P values (maintaining C:N:P of 100:4.5:0.5) were adopted in phase 2 and phase 3 to determine the adequacy of aerobic granules to remove higher substrate values and to determine their acclimatization capability to increased substrate loads. As apparent from the results, the system seemed to respond effectively to higher loads as good COD as well as P removal efficiency was obtained in phase 2 and phase 3.

Since municipal level phosphorus treatment with aerobic granulation was the primary goal, the

COD was reduced gradually to reach a value of 500 mg/L and variations in P value were made to determine the maximum value of P which could be treated by aerobic granules cultivated over a span of 5 months.

The phosphorus concentrations of 2.75 mg/L and 5.5 mg/L at a COD of 500 mg/L were easily removed by aerobic granule, given that the concentrations were low and they were assimilated for biomass growth. When concentrations of 8.25 mg/L and 16.5 mg/L were adopted, the removal efficiency dropped slightly yet removing more than 80% of the influent phosphorus.

To evaluate the effect on COD concentration on efficiency of phosphorus removal two procedures were performed: changing the COD at different levels and monitoring the efficiency of P removal, second altering the influent phosphorus concentrations at a fixed COD of 500 mg/L and monitoring % P removal. For COD, a range of 1000 mg/L- 4000 mg/L was adopted,

125 while the phosphorus concentration was varied in the range 2.75 mg/L – 16.5 mg/L at a fixed

COD of 500 mg/L.

The impact of the COD was evaluated at concentrations of: 1000 mg/L, 1500 mg/L, 2000 mg/L,

3000 mg/L and 4000 mg/L. A constant C:N:P was maintained at all these concentrations and the effluent P concentration was monitored for its efficiency of removal. In case of phosphorus concentration impact at constant COD of 500 mg/L, a total number of 4 concentrations were adopted and evaluated: 2.75 mg/L, 5.5 mg/L, 8.25 mg/L and 16.5 mg/L.

Table 6. Concentrations of COD and P removal efficiency

COD(mg/L) % P removal 1000 99.65 1500 99.55 2000 98.01 3000 99.59 4000 98.72

Table 7. Concentrations of P and P removal efficiency

P (mg/L) % P removal 2.75 99.39 5.5 98.17 8.25 82.77 16.5 79.27

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100

80 C:N:P 100:4.5:0.5 60

40 Insignificant effect

P removal efficiency (%) efficiency removal P 0 0 1000 2000 3000 4000 5000 COD concentration (mg/L)

Fig 40. Effect of COD concentration at a constant C:P ratio

120 100:0.5 100:1.1 100 100:1.6 100:3.3 80

60 Significant effect 40

P removal efficiency (%) efficiency removal P 0 0 5 10 15 20 P concentration (mg/L)

Fig. 41 Effect of P concentration at different C:P ratios

Analyzing the Figure 40, no trend can be observed between the COD concentration adopted and

P removal efficiency. The P removal rate is almost same at all the COD concentrations thus it can be concluded that varying COD concentration did not affect the P removal values. This is

127 because a constant C:N:P was maintained at every COD concentration. As mentioned in the previous section, the ratio between carbon and phosphorus was as low as 100:0.5.

This suggests that the phosphorus was used for assimilation and thus being removed efficiently at all COD concentrations chosen. As long as the ratio C:P is maintained at the same value, the phosphorus removal rate remains constant. However, if even higher levels of COD were chosen, the impact might have been different on the response. It is thus suggested to adopt for higher levels and determine the effect on phosphorus removal efficiency.

Evaluating the effect of COD concentrations above 4000 mg/L was beyond the scope of this research, hence higher level COD concentrations were not examined against phosphorus removal. Evaluating the effect of P concentration at different C:P ratios, it can be stated that the trend is significant between different phosphorus concentrations adopted and the phosphorus removal rate (Figure 41).

With the gradual increase in influent phosphorus concentration, the removal efficiency in the aerobic granular system dropped. This could be explained by two factors. One is the C:P ratio and the other is granular capability. As the phosphorus concentration was increased, the C:P ratio altered from 100:0.5. Thus, the assimilation capacity of granular sludge degraded as the phosphorus concentration increased.

The organics (COD) availability for P degradation was not enough thus the efficiency of phosphorus dropped. At each COD, a specific maximum concentration (threshold) of P can be removed. If the P concentration exceeds that threshold, the granular capability to remove phosphorus drops drastically. In this research, a P concentration of 16.5 mg/L was removed by

128 granules with an average efficiency of 80%. It can thus be established that the granular capability to remove maximum phosphorus (16.5 mg/L) at a COD of 500 mg/L is 80%.

Thus, a substantial relationship existed when phosphorus concentration was varied at fixed COD and effluent phosphorus concentration was monitored. As the influent phosphorus concentration was increased from 8.25 mg/L to 16.5 mg/L, the phosphorus removing capability of aerobic granular sludge degraded which is evident from the P-removal efficiency. When influent P concentration higher than 16.25 mg/L was maintained, P removal rate dropped to zero and therefore the experiment was terminated.

6.2.2 Biological accumulation The phosphorus removal at higher substrate ratios with C:P ratio of 100:1.65 and 100:3.3 suggests that phenomenon additional to metabolism were reasonably accountable. Biological accumulation as reported by Liu et al., (2005) might be a contributing phenomenon of P removal when higher C:P ratios were adopted. The aerobic granules cultivated over a span of 5 months might have developed capability to uptake phosphorus from the feed solution as anaerobic as well as aerobic conditions prevail inside the granule supporting the phosphorus uptake.

During the feeding period and the presence of anaerobic conditions inside the granule, there is a possibility of ortho-phosphate release while uptake of phosphorus during aeration period coupled with aerobic zone inside the granule.

The presence of VFA is important for EBPR process. Phosphorus accumulating organisms utilize VFA as food source under anaerobic conditions which results in the release of orthophosphate. Thus, in the present study the consumption of VFA due to the presence of anaerobic

129 core inside the granule could have possibly caused luxury phosphorus uptake. In order to analyze the VFA consumption, measurements were taken for influent and effluent samples.

Similar profiles as COD were observed in case of volatile fatty acids analysis. Figure 42 represents the influent and effluent VFA analysis. The influent VFA values were in the range

75%-90% of influent COD values. As the influent COD values were varied according to OLR, the detected VFA values were affected as well.

The decrease in effluent VFA suggest that VFA was utilized in the anaerobic core of the granule for release of orthophosphate. And subsequent removal of phosphorus might have occurred in the aerobic environment (aeration phase) provided to the granule. All the detected values of effluent VFA (Figure 42) were low (below 35 mg/L) signifying an apparent consumption of

VFA by anaerobic core in aerobic granule.

The effluent VFA trend was similar to effluent COD and on an average low VFA values were consistently observed in the effluent.

VFA Influent VFA Effluent 5000

4000 No VFA in effluent indicates VFA consumption during

3000

2000 VFA (mg/L) VFA 1000

0 0 50 100 150 200 Time (d) Fig. 42 Variation of influent VFA concentration with time 130

6.2.3 Biologically induced phosphorus precipitation Biologically induced phosphorus precipitation has recently been demonstrated as an important P removal process causing 45% of the total removal discovered through spectral and optical analysis (Angela et al., 2011).

The chemical components of the precipitates in the granules were determined by SEM-EDX

(Scanning electron microscopy, Energy dispersive X-ray detector). The high-resolution images taken for the granular sample through SEM can be found in Fig 43, 44, 45, 46 and 47. EDX reports revealed that P was found in the precipitates close to the center of the granules supporting the probable precipitation of phosphorus in the granular core (Figure 48, Table 8).

Fig. 43 Image of a granule sample taken through SEM

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Fig. 44 Image of granular interiors dominated by rod shaped bacteria

Fig. 45 Magnified image of granular interior

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Fig. 46 Image showing central interior granular space

Fig. 47 Image depicting microorganisms structure in deep core of granule

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Fig. 48 Ionisation energy vs counts of granular EDX spectrum

Further EDX spectral analysis on several locations of central mineral showed phosphorus, calcium and oxygen as major elements (peaks shown in Figure 48). The presence of calcium and phosphorus in central core indicate the possibility of calcium phosphate precipitation in the core of microbial granules. The results of elemental composition in granular core can be found below in Table 8.

It was found that carbon, oxygen and phosphorus was dominant near the central core with mass percentage of 44.81%, 38.18% and 7.22% respectively (Table 8).

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Table 8. Elemental composition of granular core through EDX analysis

Mass Mass Norm. Atom Element At. No. [%] [%] [%] Carbon 6 24.96 44.81 54.6 Oxygen 8 21.27 38.18 34.92 Sodium 11 3.51 6.3 4.01 Phosphorus 15 4.02 7.22 3.41 Calcium 20 0.48 0.86 0.32 Nitrogen 7 1.46 2.62 2.74 55.71 100 100

The presence of phosphorus suggests some form of phosphorus precipitation in the granular cores. Although a small percentage of phosphorus was found, however its presence indicates biologically induced precipitation. De Kreuk et al., (2005) established the high phosphorus removal by hypothesizing chemical precipitation inside the granules as one of the phosphorus removal processes (De Kreuk et al., 2005). Increased ash content in the granules was representative of the possibility of precipitation occurring in the granules.

30-41% of ash content was found in PAO enriched granules grown under anaerobic-aerobic phases in comparison to 6% when anaerobic feeding was absent from the cycle. Further the presence of calcium in influent tap water and high amounts of soluble phosphate in the anaerobic feeding possibly led to formation of calcium phosphate precipitates (De Kreuk et al., 2005).

Precipitated phosphate was extracted and a phosphate concentration of 2.6% was detected out of which 1.3% was calcium or magnesium bound. It was finally concluded that the presence of phosphorus accumulating organisms in the granules have the capability to improve the biologically induced precipitation.

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The EDX analysis on the granules obtained in present research revealed calcium mass percentage of 0.86% (Table 8). Although the quantitative composition of calcium is low, the possibility of calcium phosphate precipitation is existent as tap water also contains a calcium concentration high enough to cause precipitation.

The presence of anaerobic zone in the deep core of granules, can cause the release of phosphate inside the granule which on combining with calcium leads to the formation of precipitates. Since the anaerobic period is absent, the phosphorus uptake process might have been slow. However there exists a positive possibility as elemental composition clearly indicates the presence of phosphorus and calcium.

Manas et al., (2012) also conducted a study to determine the distribution and chemical proportion of phosphorus precipitates in granular sludge. SEM-EDX analysis was performed on the mineral clusters for the granule samples and calcium was the major constituent found in the granule with average 17-38% of mineral fraction. The second major component was phosphorus with a fraction of 7-15%. Further the zonal EDX analysis of the mineral core of granules revealed different chemical composition in the inner and external layers.

A low Ca:P ratio was found in the external layers and a higher ratio was found in the internal layers near the core indicating the presence of calcium hydroxyapatite. Therefore, a strong possibility of internal phosphorus precipitation occurs as one of the contributing phosphorus removal mechanisms in the present research.

Coma et al., (2012) reported improved phosphorus removal when a surge in inorganic constituent of the granules was observed. This was reflected from the rising difference between

MLSS and MLVSS values and hence low MLVSS/MLSS ratio (Figure 49).

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The dominance of phosphorus accumulating organisms leads to a low MLVSS/MLSS ratio as a result of higher polyphosphate storage in comparison to biomass containing heterotrophic and nitrifying bacterial populations (Taya et al., 2011). In the present research, as the system approached stability, the MLVSS/MLSS gradually decreased (after week) indicating increase of inorganic component in the granules.

Particularly, when high C:P ratios were adopted, the MLVSS/MLSS was considerably low signifying the probable presence of PAOs(Taya et al., 2011). Figure 49 shows the drop in

MLVSS/MLSS ratio. This could also be attributed to biological precipitation of phosphorus inside the granules leading to an increase in inorganic component of the granule.

MLVSS/MLSS 1.4 Ratio: 0.6 on average; Biological precipitation 1.2 Ratio: 0.8 on average; Possible presence of PAOs and thus increase in inorganic components 1 of granules

0.8

0.6 MLVSS/MLSS MLVSS/MLSS ratio 0.4 Onset of accumulation of inorganic constituents in granules 0.2

0 0 20 40 60 80 100 120 140 160 180 200 Time (d)

Fig. 49 Changes in MLVSS/MLSS ratio

The microbiological analysis of granules through DNA extraction revealed the presence of denitrifying bacteria Rhodobacteraceae (23%). Rhodocyclaceae (30%) was majorly found in the

137 granules responsible for the production of Exopolysaccharide which promotes granulation

(Figure 50). A major percentage of unclassified bacteria were found which could not be identified. However, it seems that this group of bacteria has an influential role in the granular system and further research could reveal its nature.

Flavobacteriaceae, Unclassified, 10% 6%

Rhodanobacteraceae, 7%

Rhodobacteraceae, 23%

Rhodocyclaceae, 30%

Fig. 50 Microbial composition of granules depicting major percentage of Rhodocyclaceae

Lv et al., (2014) revealed the presence of anaerobic Rhodocyclaceae was found to be mainly concentrated in the core of mature granules. Rhodobacteraceae was found abundant during granulation stage in young granules. Rhodocyclaceae was found abundant on central section of granule suggesting its dominating presence in the core of granules. The presence of family

Rhodocyclaceae produces EPS and proteins contributing to the core stability in granules (Lv et al., 2014).

The presence of Rhodocylaceae, a denitrifying bacteria, has been strongly related to the existence of excellent ability to accumulate phosphorus (Yeping et al., 2016). In the recent study

138 by Li et al., (2013) Pseudomonas sp. and Rhodocyclus sp. were predominantly found in MBR for enrichment of denitrifying phosphorus bacteria. The Rhodocyclaceae has been indicated as a phosphorus removing bacteria with a low temperature based biological phosphorus removal rate.

The microbiological analysis of Rhodocyclaceae is another contributing factor explaining the phosphorus removal under simplified aerobic conditions. The apparent presence of

Rhodocyclaceae with versatile properties of denitrifying, EPS secretion and phosphorus removal adds to the phenomenon of phosphorus removal obtained in the 184 days operated granular reactor.

Kong et al., (2004) studied the ecophysiological study of Rhydocyclus related PAO in EBPR based activated sludge plants. The organims were prevalent in all the plants with a different percentage and their responses were based on biochemical models suggested for PAOs through their lab scale analysis. The Rhodocyclus related PAO dominantly accumulated substrates like acetate, pyruvate and propionate over compounds like ethanol and butyrate.

It was further shown that Rhodocyclus related PAO could uptake orthophosphate and cause polyphosphate accumulation in the presence of oxygen, nitrate or nitrite. Hence the capability of

Rhodocyclus to cause denitrification in activated sludge EBPR plants was also signified.

Experiments adopting higher phosphorus concentrations and accessing the performance for a longer period is suggested for clear results indicating the extremities to which phosphorus can be treated at municipal level. The future prospects of the work include evaluating the P concentrations over a prolonged time.

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Engineering Factors affecting phosphorus removal

A variety of operational factors affecting the granulation process were taken into consideration for the P-removal in aerobic granular bioreactor. These engineering parameters included hydrodynamic shear force, settling time, HRT (hydraulic retention time), pH and temperature.

Hydrodynamic shear force, HRT, pH and temperature were fixed variables. Whereas, settling time was a changed variable.

Hydrodynamic shear force is a crucial factor which affects other operational factors in granulation. Superficial air velocity greater than 1.2 cm/sec and upto 3.6 cm/sec have been popularly adopted in the past for nutrients removal process in granulation. High shear force has been identified to produce more stable and dense granules (Tay et al., 2001a).

Formation of extracellular polysaccharides (EPS), increase in SOUR (Specific oxygen uptake rate), higher cell surface hydrophobicity around 90% was also reported as a result of high shear force value (Zhang et al., 2007). However, for maintenance of P removing granules, it was observed that a higher value was infact unsuitable for the sustainability of granules.

A fixed hydrodynamic shear force of 1.7 cm/sec was adopted throughout the experiment period.

The aeration pattern obtained at this shear force was smooth and uniformly distributed in the reactor. A gentle shear force allowed the granules to sustain their layered structure and avoided distortion. The aerobic granules were also uniformly suspended in the reactor therefore the chemical reactions occurring within and outside of granules were consistent throughout the reactor.

Tay et al., (2001a) examined the role of shear force on characteristics of aerobic granules. It was established that shear force had a substantial effect on formation of polysaccharide,

140 hydrophobicity, specific gravity and SOUR. A suitable shear could promote the formation of polysaccharide and polysaccharide is also produced in the presence of Rhodocylaceae (Lv et al.,

2014).

Thus, triggered growth of Rhodocyclaceae could be beneficial for P removal as the microorganisms has been associated with phosphorus removal and accumulation as discussed in the previous section. Low shear force values have generally been adopted in nutrients removal studies in the literature (Yilmaz et al., 2007). Although, the role of shear force on granule formation and stability is substantial. However, its direct effect on P removal was not majorly accountable. Minor changes in shear force did not dominantly affect the P removal capacity of granules.

Settling time was another operational parameter that was altered initially during the development of granules and fixed later when stable granules were developed. The variations in settling time had a profound effect on aerobic granule development but not on P-removal efficiency. It was set at a fixed value of 5 minutes after day 54 when mature aerobic granules were formed. A 5- minute settling time seemed suitable for P removing aerobic granulation system for the retention of healthy granular biomass. After the granules were fully grown and mature, the settling time was fixed to keep the granular system stable and to induce selection pressure for retaining good bioparticles.

Short settling time seems optimum for P removal as the granular system remains stable. A detailed study about settling time induced microbiological selection of certain bacteria during P removal is required for getting a deeper insight of the affect.

141

The pH was not controlled in the system and was simply maintained at neutral value of 7 in the feed. Similarly, the reactor was setup at room temperature between 20oC and 25oC and hence the temperature was monitored and not controlled.

142

CHAPTER 7- CONCLUSION AND FUTURE SCOPE

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7.1 Conclusion The application of the most recent technology aerobic granulation and its practice for domestic level phosphorus cessation has been discussed so far. The principle conclusions obtained from this research based on phosphorus removal process under single aerobic phase via aerobic granulation have been presented as follows.

1. Granule development was observed as a gradual process. The initial young aerobic

granules were observed on day 9 qualifying the defined 200µm size alongside

improvement in settling ability. The base granules were developed under a substrate load

of 6.4 kg COD/m3 day. The activated sludge turned into young granules which further

developed into semi-mature granules. After day 60, fully mature granules with a defined

shape were developed. An increasing trend in particle size of granular sludge, increase in

biomass concentration (MLSS and MLVSS) and decrement in SVI was positively observed

during granulation.

Settling time played an influential role in inducing a selection pressure on sludge for

granules formation. It was gradually decreased from 20 minutes to 5 minutes, until stable

aerobic granules were formed in the reactor. The settling time variation lead to the retention

of healthy biomass in the reactor whereas the fluffy flocs were washed out of the system.

The better settling sludge aggregated and developed into stable aerobic granules.

Hydrodynamic shear force of 1.7 cm/sec at an air pressure of 3 psi was deemed to be an

effective fixed parameter for development and maintenance of stable granules. This shear

force was nominal to keep the granules suspended in the reactor and further to avoid

distortion of the granules contributing to granular stability. For an engineering application,

the hydrodynamic shear force of 1.7 cm/sec corresponds to an air flow rate of 4.76

144

litre/min. An average aeration period of 2.5 hours was adopted. For a full scale application,

the air flow rate and time should be calculated depending on reactor volume and

configuration.

Hydraulic retention time of 7.5 hours was fixed throughout the various stages of

experiment and hence was identified to be appropriate for granular development. This

value was roughly adopted to mimic HRT of activated sludge systems (4-8 hours) treating

domestic wastewater. Since the HRT value was adopted purely based on literature

recommendations, variations in HRT were not practiced. A 7.5 hour HRT ensured an

average frequency of change in wastewater thus hydraulic selection pressure on biomass

was not dominant.

2. A high removal efficiency of COD, ammonia and phosphate were observed during the

experiment. Effluent with low COD, ammonia and phosphorus was produced indicating

removal through biological assimilation predominantly. The C:N:P ratio was fixed at a

value of 100:4.5:0.5 for the first 8 weeks of the experiment which led to stable granules

maintenance. This proves that providing a right ratio of organics to nutrients can lead to

stable granular growth and maintenance. A high biomass retained in the aerobic granular

sludge assisted in removing high concentrations of COD, ammonia and phosphorus.

3. The C:P ratio significantly affected the P removal efficiency of the aerobic granular system.

Different C:P ratios of 100:0.5, 100:1.1, 100:1.65 and 100:3.3 were adopted. To evaluate

the effect of P concentration at different C:P ratios, influent phosphorus concentrations

were changed in the range of 5.5 mg/L- 16.5 mg/L at a fixed COD of 500 mg/L. The

phosphorus removal efficiency was then evaluated at each C:P ratio. A significant effect

of C:P ratio was observed on phosphorus removal efficiency. At a low C:P ratio of 100:0.5

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and 100:1.1, a higher P removal efficiency was observed. While at a high C:P ratio of

100:1.65 and 100:3.3, a low phosphorus removal efficiency was observed.

4. At low C:P ratios, the primary factor causing phosphorus removal was hypothesized to be

biological assimilation. An increased assimilation capacity of the aerobic granules was

observed due to a high biomass retention property of the granules.

5. At high C:P ratio, P removal occurred through two different pathways: (1) biological

accumulation; (2) biologically induced chemical precipitation.

Biological accumulation was assumed responsible at high C:P with probable presence of

phosphorus accumulating bacteria. Microbiological analysis revealed the presence of

Rhodocyclaceae; identified as phosphorus removing bacteria. At low C:P, biological

assimilation contributed in enhanced P removal

Chemical precipitation was verified through SEM-EDX spectrum. EDX analysis revealed

that the major component of granule core was phosphorus, with a mass percentage of

7.22%. Mass transfer diffusion caused chemical gradients inside the granule, leading to

favourable environment for phosphorus precipitation.

7.2 Future Work Prospects of the research include- 1. An in-depth and combined investigation of microbial and performance analyses should be

performed to identify all the species in the granules and relate the function of each to the

system performance. A detailed microbiological analysis of granular interiors to detect the

microorganisms contributing to phosphorus accumulation inside the granular layers. The

role of unclassified bacteria detected in the analysis seems to have an influential role in P

removal process which needs further investigation. Rhodocyclaceae and Rhodobacteraceae

146

have been identified in the aerobic granules via gene sequencing. Further investigation on

their role in phosphorus removal will help in optimization of granular P-removal process.

2. The experiment should be replicated to further investigate and thus confirm the effect of

C:P ratio on P removal efficiency over a prolonged time period. This also includes

evaluating the maximum phosphorus concentrations that could be treatable by aerobic

granular sludge at a municipal level wastewater concentration.

3. The effect of other influential factors, such as HRT, hydrodynamic shear force, and oxygen

saturation level should be examined and accounted for.

4. Contribution of chemical precipitation in P removal should be quantified and the

underlying mechanism should be verified

Future investigation would necessitate expansion of lab scale unit to a bigger facility with a better process instrumentation and control. The granulation technology is currently being developed for its applications on high strength industrial wastewater treatment, removing of emerging substances of concern (ESOCs), treatment of oil sands process, prevent fouling of engineered biofilms and excellent outcomes are expected.

147

CHAPTER 8- REFERENCES

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PUBLICATIONS ▪ Pishgar R., Kanda A., Gress G.R., Gong H., Tay J.H. (2017) Startup of Aerobic Granulation Technology: Troubleshooting Scale-up Issue. In: Mannina G. (eds) Frontiers in Wastewater Treatment and Modelling. FICWTM 2017. Lecture Notes in Civil Engineering, vol 4. Springer, Cham

▪ Pishgar R., Kanda A., Gress G.R., Gong H., Tay J.H. (2018) Effect of aeration pattern and gas distribution during scale-up of bubble column reactor for aerobic granulation. Journal of Environmental Chemical Engineering (under review, Manuscript Ref. #JECE- D-18-01280)

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