development and optimization of an a-stage – daf...

Faculteit Bio-ingenieurswetenschappen

Academiejaar 2015 – 2016

Development and optimization of an A-stage – DAF

system

Ine De Saedeleer

Promotor: Prof. dr. ir. Korneel Rabaey & Prof. dr.ir. Bart De Gusseme

Tutor: Cristina Cagnetta & Dr. ir. Jo De Vrieze

Masterproef voorgedragen tot het behalen van de graad van Master in de

bio-ingenieurswetenschappen: Milieutechnologie

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vermelden bij het aanhalen van resultaten uit deze scriptie.”

Gent, juni 2016

Promotor Promotor Author

Prof. Dr. ir. Korneel Rabaey Prof. Dr. ir. Bart De Gusseme Ine De Saedeleer

i

Acknowledgements

This past year, I have gained so much valuable knowledge that will guide me in the path to

come. I would have never gained this knowledge or been able to finish this thesis without the

help and support of many people, whom I would like to thank.

To start, I would like to thank Cristina Cagnetta, who was always there to keep me motivated,

to help me when the reactor was broken again, to give me helpful suggestions on writing this

thesis and just learning me everything that I needed to know to bring this thesis to a good end.

I will always keep in mind what she said after 5 months of hard labour without result: “It is not

a matter if it will float, it is just a matter when it will float”. Thank you for helping me and

always staying positive no matter what, without you this thesis would not have been completed.

I would also like to thank Dr. ir. Jo De Vrieze, for guiding me in writing this thesis as well as

his valuable help in the lab. Prof. dr. ir. Korneel Rabaey inspired me two years ago, in his

lectures on wastewater treatment, which led me to chose a thesis topic at CMET. Prof. dr. ir.

Bart De Gusseme provided scientific and practical guidance as well as helpful solutions when

problems arose and created with me the story of my thesis. Without prof. dr. Arne Verliefde

providing me the DAF unit, this thesis would not have existed in the first place, so thank you

for lending it to me.

I would also like to thank everyone at CMET, for creating such a pleasant environment to work

in and all the helpful hands whenever I needed guidance in the lab or more specifically when I

could not unscrew something. A warm thanks to all the thesis students, in particular Jan De

Kezel, with whom you could always have a chat, who where always there whenever you needed

any help and became good friends along the year. All of you made it bearable whenever I had

a hard day in the lab and made this thesis a nice experience.

At last, I would like to thank my parents, sister, boyfriend and friends for supporting me when

I needed it, always listening whenever I complained and just making me laugh and happy.

Although I’m excited to start a new challenge next year, finishing my thesis at CMET was the

perfect ending of my bio-engineering studies. Thank you all for the unforgettable time this last

year!

iii

Abstract The current activated sludge process does not exploit the maximal resource recovery potential

of the wastewater. The adsorption bio-oxidation (AB) process could be better suited, since the

highly loaded first A-stage operated at short hydraulic retention time (∼ 30 min) and sludge

retention time (< 1 day) results in large incorporation of influent organic matter into highly

biodegradable sludge with minimal energy input. The high food-to-microorganism ratio

however, employed in the A-stage determines the poor settling characteristics of the A-sludge,

resulting in suboptimal solid/liquid separation by the conventional settler. This suggests that

optimization of sludge separation by replacing the settler with a different technique could lead

to a higher resource recovery through obtaining a more concentrated sludge after separation. In

this thesis the innovative combination of dissolved air flotation (DAF) for solid/liquid

separation with the A-stage was explored. No research has been done so far on this combination,

necessitating the investigation of its feasibility, and optimizing the system in order to achieve

the most optimal solid/liquid separation.

This study demonstrated that using DAF for solid/liquid separation of A-sludge is feasible, and

that high removal efficiencies could be obtained. First, the optimal concentration of coagulant

to be used for proper floc formation, susceptible to flotation, was determined. An optimal

concentration of 75 mg L-1 FeCl3.6H2O and 50 mg L-1 AlCl3.6H2O, both in combination with 3

mg L-1 cationic polymer, was found to give the highest decrease in volatile suspended solids

(VSS) and turbidity. When an organic polyelectrolyte was used as coagulant (Zetag 7651), a

concentration of 10 mg L-1 was found optimal, and this concentration achieved a removal

efficiency of 84% in terms of total chemical oxygen demand (tCOD), a 93% removal of total

suspended solids (TSS), and 91% removal of VSS. These high removal efficiencies resulted in

a concentrated floated sludge layer with a tCOD content of 22 g L-1 and a TSS concentration of

16 g L-1.

This concentrated sludge layer was thereafter digested and fermented to assess the potential of

recovering energy and volatile fatty acids (VFA). The floated sludge proved to be efficient in

producing VFA, since a maximal production of 401 mg VFA COD g-1 CODfed was obtained,

corresponding with 40 % conversion of the COD to VFA, when Zetag 7651 was used for the

coagulation of the sludge. The CH4 production, however, was rather low, since only a digestion

efficiency of 35 % was achieved when sludge coagulated with FeCl3.6H2O was digested,

iv

presumably due to the use of an unsuited and too old inoculum for the digestion of this particular

sludge.

This study demonstrated that by using DAF for separation of A-sludge, a high removal

efficiency of COD, TSS and VSS can be achieved, resulting in a concentrated sludge layer that

shows considerable potential for resource recovery in the form of CH4 and VFA production.

v

Samenvatting Het conventioneel actief slib proces slaagt er niet in de maximale hulpbronnen uit het afvalwater

te recupereren. Het adsorptie-bio-oxidatie proces zou hiervoor beter geschikt kunnen zijn,

gezien de hoogbelaste A-trap, uitgevoerd bij korte hydraulische retentietijden (∼ 30 min) en

slibleeftijd (< 1 dag), resulteert in een grote incorporatie van influent organisch materiaal in

goed biodegradeerbaar slib met minimale energie input. De hoge slibbelasting toegepast in de

A-trap leidt echter tot slechte bezinkingskarakteristieken van het A-slib, resulterend in een

suboptimale scheiding van slib en water door de conventionele bezinkingstank. Dit suggereert

dat een optimalisatie van slib afscheiding, door het gebruik van een alternatieve techniek ter

vervanging van de bezinkingstank, zou kunnen leiden tot een hogere recuperatie van

hulpbronnen door het verkrijgen van een meer geconcentreerd slib na afscheiding. In deze thesis

werd de innovatieve combinatie van een A-trap met dissolved air flotation (DAF) voor

slib/water afscheiding onderzocht. Er werd tot nu toe nog geen onderzoek uitgevoerd naar deze

combinatie, wat het nagaan van de haalbaarheid en de optimalisatie van het systeem

noodzakelijk maakt, om op die manier de meest optimale slib/water afscheiding te verkrijgen.

Deze studie demonstreerde dat het gebruik van DAF om A-slib af te scheiden haalbaar is en dat

hoge verwijderingsrendementen konden verkregen worden. Eerst werd de optimale coagulant

concentratie voor voldoende vlokvorming, gevoelig voor flotatie, bepaald. Een optimale

concentratie van 75 mg L-1 FeCl3.6H2O en 50 mg L-1 AlCl3.6H2O, beiden in combinatie met 3

mg L-1 kationisch polymeer, bleek te leiden tot de grootste afname in vluchtig zwevende stoffen

(VSS) en turbiditeit. Wanneer een organische poly-elektrolyt werd gebruikt als coagulant

(Zetag 7651) bleek een concentratie van 10 mg L-1 optimaal te zijn. Deze concentratie bereikte

een verwijderingsrendement van 84 % in termen van chemische zuurstofvraag (COD), een 93

% verwijdering van totale zwevende stoffen (TSS) en een 91 % verwijdering van VSS. Deze

hoge verwijderingsrendementen resulteerden in een geconcentreerde drijvende sliblaag met een

totale COD concentratie van 22 g L-1 en een TSS concentratie van 16 g L-1.

Deze geconcentreerde sliblaag werd hierna vergist en gefermenteerd om het potentieel om

energie en volatiele vetzuren (VFA) te recupereren vast te stellen. Het drijvende slib bleek

efficiënt te zijn in het produceren van VFA, gezien een maximale productie van 401 mg VFA

COD g-1 CODgevoed werd bereikt, corresponderend met een 40 % conversie van COD in VFA,

wanneer Zetag 7651 werd gebruikt als coagulant. De methaan productie daarentegen was vrij

vi

laag, gezien slechts een digestie efficiëntie van 35 % werd bereikt wanneer FeCl3.6H2O werd

gebruikt als coagulant, vermoedelijk door het gebruik van een ongeschikt en te oud inoculum

voor de digestie van dit slib.

Deze studie demonstreerde dat door het gebruik van DAF voor de afscheiding van A-slib een

hoog verwijderingsrendement van COD, TSS en VSS kon bereikt worden, resulterend in een

geconcentreerde sliblaag die een aanzienlijk potentieel voor hulpbronnen recuperatie in de

vorm van methaan en VFA vertoonde.

vii

List of abbreviations

AB process

A/S ratio

AD

BMP

BOD

CODH

COD

CSTR

CAS

DAF

DO

F/M

HRT

MLVSS

NTU

FID

OLAND

SBR

Adsorption-bio-oxidation process

Air to solids ratio

Anaerobic digestion

Biochemical methane potential

Biological oxygen demand

Carbon monoxide dehydrogenase

Chemical oxygen demand

Continuous stirred tank reactor

Conventional activated sludge system

Dissolved air flotation

Dissolved oxygen

Food-to-microorganism

Hydraulic retention time

Mixed liquor volatile suspended solids

Nephelometric turbidity unit

Flame ionisation detector

Oxygen-limited autotrophic nitrification-denitrification

Sequencing batch reactor

viii

SRT

SVI

sCOD

SS

tCOD

TS

VFA

VS

VSS

WWTP

Sludge retention time

Sludge volume index

Soluble COD

Suspended solids

Total COD

Total solids

Volatile fatty acids

Volatile solids

Volatile suspended solids

Wastewater treatment plant

Index

Abstract iii

Samenvatting v

List of abbreviations vii

Part I Literature review 1

1. Abundant resource potential in wastewater 1

2. Activated sludge process 3

2.1. Conventional activated sludge process 32.2. Settling problems in secondary settlers 5

2.3. Shift towards a more sustainable WWTP 6

3. The AB process 8

3.1. Detailed description of the AB process 83.2. Benefits of AB process 9

3.3. Coagulation-flocculation 11

4. Air flotation 13

4.1. Dissolved air flotation 134.2. Applications of DAF in wastewater treatment 16

4.3. Benefits of DAF compared to conventional sedimentation 16

5. Anaerobic digestion and fermentation 18

5.1. Anaerobic digestion- a microbial process 20

5.2. Influence of iron and aluminium 21

6. Research questions 23

Part II: Materials and methods 25

1. CSTR with concentrated SYNTHES 27

2. Coagulation: jar testing 29

3. System set-up 31

3.1. Materials 323.2. SBR cycle 33

343.3. Influent 34

4. Development and optimization experiments 35

4.1. Development of A-stage DAF 35

4.1.1. Experiment 1 CSTR 354.1.2. Experiment 2 Real conditions 35

4.1.3. Experiment 3 Coagulant and flocculant 364.1.4. Experiment 4 pH control 36

4.1.5. Experiment 5 Increasing the VSS 364.1.6. Experiment 6 Air dissolution 36

4.2. Optimization of A-stage DAF 39Determining optimal concentration of AlC3.6H2O, FeCl3.6H2O and Zetag 7651 39

5. Batch tests 40

5.1. Fermentation test 40

Fermentation of the floated sludge 40

5.2. Biochemical methane potential (BMP) test 41

6. Chemical analyses 42

6.1. Chemical Oxygen Demand 42

6.2. Solids analysis 42

6.3. VFA analysis 436.4. Gas analysis 43

6.5. Turbidity 43

Part III Results 45

1. CSTR with concentrated SYNTHES 47

2. Coagulation jar testing 49

2.1. Concentration and type of coagulant 49

2.2. Flocculation time 50

3. Development and optimization experiments 51


3.1.1. Type of A-sludge 513.1.2. Type of coagulant/flocculant 52

3.1.3. pH of the MLVSS 523.1.4. COD influent 52

3.1.5. VSS of the MLVSS 523.1.6. Amount of bubbles 53

3.1.7. Air dissolution into the liquid 533.2. Optimization of A-stage DAF 54

3.3. DAF performance 563.4. Comparison DAF with settling experiments 57

3.5. Sludge characterization 58

4. Batch tests 60

4.1. Fermentation batch test 604.2. BMP test 64

Part IV Discussion 67

1. A-stage: state of the art 69

2. CSTR 71

2.1. CSTR with concentrated SYNTHES 71

2.2. Coagulation jar testing 71

3. Development and optimization of A-stage DAF 73


3.2. DAF: Benefits compared to conventional settling 763.3. Optimal conditions for flotation 78

4. Batch tests 80

4.1. Fermentation batch test 80

4.2. BMP test 82

5. General conclusions 85

6. Future perspectives 87

Part I Literature review

Abundant resource potential in wastewater 1

1. Abundant resource potential in wastewater

In our present world, 2.5 billion people lack access to improved sanitation, due to the water

crisis (WHO, 2015b). According to Wallace & Gregory (2002), by 2050 66% of the world’s

population will experience water shortage and water requirements for food production will be

a major issue. The growing world population, urbanization, climate change, limited fresh water

supplies, and cost of water treatment emphasize the need for reuse of wastewater in agriculture

and industry (WHO, 2015a). Instead of discharging un-treated wastewater, it is actually feasible

to produce high quality freshwater from wastewater, and implement the wastewater treatment

as a part of the production process, instead of seeing it as an end-of-pipe solution (Verstraete et

al., 2009).

Not only water scarcity is a global issue, attention is also shifted to the scarcity of nutrients and

energy. By 2020 the European Union aims to (1) reduce its greenhouse gas emissions by at

least 20%, (2) increase the share of renewable energy to at least 20% of total energy

consumption, and (3) achieve total energy savings of 20% or more. Fluctuating energy prices

and dependence on energy are driving forces for research towards more energy efficient

processes and the recovery of energy from waste streams (European Commission, 2010).

Since the water industry in developed economies uses 3 to 5% of the total electricity demand,

energy self-sufficiency of a wastewater treatment plant (WWTP) through biogas production by

anaerobic digestion can contribute significantly to the 2020 goals (Curtis, 2010). The collection

and treatment of wastewater requires 0.6 kWh per cubic meter wastewater treated (McCarty et

al., 2011). Thermodynamic calculations indicate that this demand is unnecessary, the chemical

energy in municipal wastewater exceeds this requirement by almost a factor 4 (1.96 kWh per

cubic meter) (Scherson & Criddle, 2014). The conversion of this chemical energy into useful

energy by an anaerobic digester is a perfect on-site energy recovery possibility (Wett et al.,

2007).

Since discharging phosphorus and nitrogen into rivers, lakes and seas can lead to severe

deterioration of environmental quality through eutrophication, increasingly stringent

regulations are a result of the growing awareness of P and N control in wastewater (Oehmen et

Abundant resource potential in wastewater 2

al., 2007). Enhanced biological phosphorus removal is potentially the most cost-effective

method to remove P from wastewater, because it avoids the use of chemicals and excess sludge

production (Morse et al., 1998). Its success highly depends on the abundant availability of

volatile fatty acids (VFA). Fermentation of waste sludge instead of anaerobic digestion can be

a valuable source for the production of VFA (Morgan-sagastume et al., 2011). To reduce costs

and limit contribution to the environmental footprint of the WWTP, the use of VFA originating

from the on-site fermentation of sludge can be considered as the most sustainable source (Q.

Yuan et al., 2009). The VFA are not only a carbon source for biological P removal, but because

it’s readily bioavailable, it’s also most suitable for denitrification and contributes in this way

also to the sustainable removal of N from the wastewater (Cokgor et al., 2009).

Typically, wastewater is treated by the conventional activated sludge system (CAS) to remove

organics and nutrients. In view of resource recovery and energy efficiency the AB-system

(translated as adsorption-bio-oxidation system) gains increasing interest. Especially the highly

loaded A-stage supports the high rate entrapment of organics, without excessive aerobic

stabilization (Wett et al., 2007). The A-sludge is traditionally separated by a settler but due to

the high food-to-microorganism (F/M) ratio of the sludge, filamentous organisms (for example

Sphaerotilus) become predominant, leading to poor settling (Ramalho, 2012). This poor settling

results in low activated sludge concentrations, determining the need for thickening before being

used in a fermenter or digester (Bolzonella et al., 2005; Wanner, 1994). Interests are shifted to

other separation techniques, one of them being flotation. Instead of gravitationally letting the

sludge settle, air is used to float the sludge (Wang et al., 2010). El-Gohary & El-Ela (1980)

compared biological- coagulation sedimentation and biological coagulation flotation

techniques and showed that higher removal values of chemical oxygen demand (COD) could

be achieved.

Activated sludge process 3

2. Activated sludge process

The activated sludge process was developed in England, more than 100 years ago, by

Harden and Lockett and is, at present, the most commonly used technology for

biological wastewater treatment (Martins et al., 2004). Depending on the design of the

process, an activated sludge WWTP can achieve biological nitrogen and phosphorus

removal, besides removal of organic carbon (Gernaey et al., 2004).

2.1. Conventional activated sludge process

In the activated sludge process, first the organic material is converted into sludge by

microorganisms (MO) in an aeration tank, and, second, the sludge is settled and

separated from the clear effluent (Çakici & Bayramoǧlu, 1995; Gernaey et al., 2004).

The activated sludge process is a key part of the whole WWTP, but several other

physical/biological and/or chemical unit steps are necessary to obtain effluent of good

quality in order to protect human health and the environment (Tchobanoglous et al.,

1991). The conventional activated sludge (CAS) process consists of three stages:

primary, secondary and tertiary treatment (Figure I.1).

Typically, the first stage of a WWTP is a pre-treatment of the water, which involves a

grit remover to remove solid materials, such as plastics and stones, which could

otherwise damage or clog the plant’s pumps and skimmers. After this a sand-trap

follows to settle out the sand to prevent abrasion. Then, a primary clarifier removes

settleable material and floating greases and fats can be skimmed off (Tchobanoglous et

al., 1991). The primary treatment can remove about 90% of the settleable solids and

10-30 % of organic carbon, and to a lesser degree, some organic nitrogen and

phosphorus (Wang et al., 2010). Primary treatment can decrease the COD:N and

COD:P ratio substantially leading to a decrease in the amount of P and N that can be

removed without the addition of an extra carbon dosage. Therefore, short retention

times in the primary clarifier are often chosen to reduce the COD removal and ensure

further on nutrient removal (Nowak, 2002; Randall & Barnard, 1998).


Secondary biological treatment involves micro-organisms that grow in an aeration tank,

converting approximately half of the COD into new biomass (sludge) and half into CO2,

due to the low sludge-specific loading rate. Typically, next to aerated zones also anoxic

zones are implemented in the overall system to ensure that nitrification, denitrification

and enhanced biological phosphorus removal can be obtained (Ge et al., 2012). All

these different biological processes result in the increasingly complex complete

treatment of wastewater (Verstraete & Vlaeminck, 2011).

In the secondary settler the clear effluent can be separated from the sludge by

gravitational sedimentation. Part of the sludge (excess sludge) is regularly removed to

maintain a desired sludge age, while another portion of the settled sludge is recycled

back to the aeration tank to maintain enough microorganisms in the system (Bratby &

Marais, 1976; Çakici & Bayramoǧlu, 1995).

In some cases, the secondary treatment is not sufficient to meet the regulatory effluent

requirements, and a tertiary treatment is necessary. This tertiary treatment can involve

ultrafiltration, reverse osmosis, carbon adsorption or several other possibilities. The

plant effluent is often disinfected with chlorine, ozone or UV light to destroy pathogenic

microorganisms before it’s discharged into the receiving water bodies (Wang et al.,

2010).

Figure I.1: Conventional activated sludge process (Katsoyiannis et al., 2006)


2.2. Settling problems in secondary settlers

Typical issues in the activated sludge process are the separation of sludge from clear

effluent in the secondary settler and insufficient sludge thickening (Bratby & Marais,

1976). These problems are related to the performance of the settler, which makes this

process a crucial part of the whole WWTP. Potential improvements of the secondary

settler in space requirements, retention times and especially sludge thickening are

limited, resulting in no significant improvements since the development of the settler

in the last 50 years (Wang et al., 2005).

Bulking sludge is one of the most common challenges in the CAS process. It is caused

by the excessive growth of filamentous bacteria, and leads to slow settling and poor

compaction (Martins et al., 2004; Novàk et al., 1993), resulting in a very low suspended

solids (SS) concentration at the bottom of the settler and in the return sludge. Due to

this low SS concentration, it is very difficult to maintain or achieve a sufficient amount

of microorganisms in the aeration tank, and the overall performance of the WWTP

could possibly decline. Bulking sludge can also give rise to solids carried over with

effluent from the secondary settler, which results in a decrease of effluent quality

(Bratby & Marais, 1976).

Currently, solutions to bulking sludge mainly include non-specific methods, such as

chlorination, and specific methods involving strategies to suppress the growth of

filamentous bacteria, for example keeping the dissolved oxygen concentration greater

than or equal to 2 mg L-1. Non-specific methods will give rise to additional cost, while

specific strategies are difficult to achieve in most cases (Guo et al., 2010). The most

applied strategy to prevent bulking sludge is the use of a selector, where a feast and

famine regime is applied to suppress the filamentous microorganisms (Chudoba et al.,

1973). Although successful in some cases, the selector does not guarantee control over

bulking sludge (Linne & Chiesa, 1987)

Rising sludge is a second phenomenon occurring in the secondary settler. It is due to

hydrolysis of particulate degradable COD and decay of biomass, leading to

denitrification in secondary settlers, which results in the formation of nitrogen gas

bubbles (Koch et al., 1999). These nitrogen gas bubbles lead to the flotation of SS


resulting in increased SS in the effluent and an overall reduction of the effluent quality

(Henze et al., 1993).

2.3. Shift towards a more sustainable WWTP

The focus of a WWTP used to be solely on achieving a good effluent quality that

protected human health and satisfied the regulatory demands. Since sustainability,

especially the more efficient use of energy, has gained increasing attention, the focus

has shifted also towards maximum recovery of all nutrients and energy present in

wastewater (Angenent et al., 2004). A new challenge WWTPs face is to minimize

energy consumption, while maintaining the desired quality of the effluent (Silvestre et

al., 2015).

The CAS process has several flaws in its design, which leaves room for improvement.

Besides the substantial environmental footprint, it does not utilize all the potential

opportunities where resources can be recovered. Given that also the energy efficiency

of most WWTP’s is low due to the high aeration demands, and therefore also the cost-

effectiveness, it’s clear that driving forces for rethinking the conventional system are

abundantly present. The CAS as it’s known now is not sustainable, and further research

into more efficient and sustainable alternatives is crucial (Verstraete & Vlaeminck,

2011).

A more sustainable approach is the ZeroWasteWater concept, where wastewater is

short cycled and considered as a source of fresh water, nutrients and energy (Figure

I.2). A big issue in the recovery of energy and nutrients is the high dilution of the waste

stream that enters the WWTP. A possible solution is to separate sewer systems for

sanitary and storm water, but this will mean high investment cost, as well as radical

infrastructural works. A more applicable solution is the advanced concentration of the

sewage arriving in the WWTP by means of dissolved air flotation, chemically enhanced

primary sedimentation or biological sorption. This biological sorption is exactly what

takes place in the A-stage of the Adsorption-Belebungsverfahren process (AB-process),

and this process shows great potential in recovering the high amount of energy and

nutrients present in wastewater (Verstraete & Vlaeminck, 2011).


Figure I.2: Major pathways left: wasting a lot of resources, right: resource recovery in

light of ZeroWasteWater (Verstraete & Vlaeminck, 2011)

The use of an AB-process instead of CAS supports the entrapment of the organics

without excessive aerobic mineralization. This in combination with aeration control,

based on on-line effluent ammonia concentrations, and the implementation of an

energetically attractive nitrogen treatment with oxygen-limited autotrophic

nitrification-denitrification (OLAND) may result in an overall positive energy balance

of the WWTP (Wett et al., 2007).

The AB process 8

3. The AB process

Developed during the energy crisis in the 70’s by Dr. Botho Boehnke, the AB-process

or adsorption/bio-oxidation process can play a significant role in the future energy self-

sufficient WWTP. It consists of a first more innovative stage, a high rate adsorption

stage, followed by a second stage, where bio-oxidation takes place, more similar to a

CAS system (Figure I.3). The principles of the process were intended at first to treat

industrial wastewaters with too extreme conditions to treat in a CAS, but the potential

to treat also domestic sewage was quickly realized (Böhnke et al., 1997).

Figure I.3: Simplified scheme of a typical AB process

3.1. Detailed description of the AB process

The highly loaded first A-stage is characterized by a very high food-to-microorganism

(F/M) ratio ranging from 2 to 10 kg biological oxygen demand (BOD) kg-1 volatile

suspended solids (VSS) day-1, extremely low sludge retention time (SRT) values less

than 1 day and a short hydraulic retention time (HRT) of 30 minutes or less. These

characteristics aim at a fast removal of organic material (BOD removal of 50-60%) and

at the same time a high production of sludge (Salomé, 1990).

Removal of the COD occurs mainly in the first stage due to adsorption, coagulation and

flocculation (Böhnke et al., 1997). Biological oxidation plays only a minor role (15 -

20% of the COD converted to CO2). Due to this low biological oxidation, the loss of

The AB process 9

organic material through mineralization is prevented, and higher sludge yields can be

observed (De Graaff & Roest, 2012).

Since oxygen is necessary only for the biological growth (no oxygen is required for the

adsorption) the aeration costs can be severely reduced, and A-stage reactors can operate

near a dissolved oxygen (DO)-level of zero (De Graaff & Roest, 2012). It has been

shown that to obtain the highest removal rates, no pre-treatment stage should precede

the adsorption stage to avoid removal of organic material that is advantageous in the

A-stage. (Böhnke et al., 1997).

The low loaded second B-stage has completely different characteristics than the A-

stage, and can best be compared to a CAS. A F/M ratio of less than 0.1 kg BOD kg-

1VSS day-1 and a high SRT are ideal for the growth of autotrophic nitrifiers to ensure

nitrification. If denitrification is desired, sufficient COD needs to be present in the A-

stage effluent to support the growth of the denitrifiers (Böhnke et al., 1997).

3.2. Benefits of AB process

The biggest advantage of the AB-system lies in its 10-20% lower energy demand

compared to CAS (Salomé, 1990). Considering that 60% of the energy usage of CAS

is due to aeration, the AB-process finds its economic advantage specifically in the lower

aeration demand of the A-stage. Since biological nitrogen removal represents a major

part of the overall electrical energy demand, even further energy reductions can be

accomplished by using more innovative techniques to remove nitrogen. By

implementing partial nitritation and anammox, 30-40% of the overall nitrogen removal

costs can be saved (De Clippeleir et al., 2011). This reduction of cost is not only due to

a significant reduction of aeration costs, but also due to the redundancy of external

carbon dosage for denitrification. The oxygen limited autotrophic

nitrification/denitrification (OLAND) is a configuration of this process, and it is

feasible to combine this process with a low HRT, which characterizes the A-stage (De

Clippeleir et al., 2011) Energy-positive sewage treatment becomes possible with this

configuration.

Due to the high F/M ratio and short SRT an overall smaller volume of the treatment

plant is required to obtain similar effluent qualities as in CAS (Böhnke et al., 1997).

The AB process 10

Salomé (1990) concluded that installing an AB process instead of a CAS process can

reduce the costs with about 25%. Moreover, A-stage reactors are fairly insensitive to

fluctuations in organic load, pH or toxic shocks, and give rise to a process stability of

the overall system as well as a fairly constant load that passes through to the B-stage

(De Graaff & Roest, 2012).

The AB-systems aim at high-rate incorporation of the organics into sludge without

extensive oxidation, which leads to energy rich A-stage sludge that can be easily

digested through anaerobic digestion (Verstraete et al., 2009; Wett et al., 2007). All

these benefits contribute to the overall reduction of the environmental footprint, and

play an important role in nutrients and energy recovery besides the purification of

wastewater.

Due to the high F/M ratio and low oxygen levels in the A-stage settling problems may

occur. Not a lot of studies have been carried out regarding the settleability of A-sludge

and some contradictory opinions can be found in literature. According to Salomé (1990)

the sludge volume index (SVI) of A-sludge is low (40-60 mL g-1) which should

theoretically result in quickly settling flocs and clear supernatant. Ramalho (2012),

however, showed that the high F/M ratio leads to the prevalence of filamentous

organisms that do not settle well, and remain in suspension almost indefinitely,

especially in combination with low oxygen concentrations, typically present in the A-

stage (Wilen & Balmer, 1999) (Figure I.4).

Figure I.4: Relationship between F/M ratio and sludge volume index

The AB process 11

3.3. Coagulation-flocculation

Coagulation and flocculation are well known processes in wastewater treatment to

combine small particles into larger aggregates to enhance the separation of sludge from

clear water (Li et al., 2006; Wang et al., 2005). Although the distinction between

coagulation and flocculation in practice has little meaning, because both stages occur

very rapidly, they are intended to describe different processes (Jiang & Graham, 1998).

Coagulants used for wastewater treatment are predominantly based on inorganic salts

of iron and aluminium (Jiang & Graham, 1998). These hydrolyse rapidly to form a

range of metal hydrolysis species, which are then adsorbed by negatively charged

particles. Next to inorganic salts also synthetic organic polymers, known as

polyelectrolytes can be used for the coagulation (Wang et al., 2005). These polymers

offer some significant advantages when compared to inorganic salts. Although the salts

are readily available and cheap, they are often required in high doses, and a large

amount of base is needed to change the pH of the solution to achieve its isoelectric point

(Sharma & Sanghi, 2012). Polymers can be dosed in much lower amounts than the salts,

and will increase the toughness of the flocs. Disadvantages on the other hand are a

greater sensitivity to incorrect dosing and higher costs (Bolto & Gregory, 2007).

The overall coagulation/flocculation consists of three sequential steps: coagulant

formation, particle destabilization and particle aggregation. Coagulation occurs

typically in a rapid-mixing stage that ensures the rapid hydrolysis of the coagulant and

the contact with suspended particles, leading to particle destabilization. After this

destabilization, a slow-mixing phase begins where particles are able to aggregate to

larger flocs, this is defined as flocculation (Jiang & Graham, 1998; Wang et al., 2005).

There are three mechanisms of coagulation: charge neutralization, sweeping, and

bridging. Many hydrolysis products of metal salts are cationic and can interact strongly

with negatively charged particles causing charge neutralization and, thereby,

destabilization of these particles. In contrast, when the dose of metal salts is sufficiently

high to cause precipitation of amorphous metal hydroxides, this leads to the possibility

of sweep coagulation in which particles can be enmeshed in the growing precipitate

and removed (Duan & Gregory, 2003). When long chain polyelectrolytes are used for

coagulation, they can strongly adsorb to particles, due to their opposite charge. These

The AB process 12

adsorbed polymers can have loops and tails extending into solution that can attach to

other particles bridging the particles together into sludge flocs (Figure I.5) (Bolto &

Gregory, 2007). These three mechanisms can operate simultaneously, depending on the

nature of the particle surface and the type of coagulant.

Figure I.5: Polymer bridging

Air flotation 13

4. Air flotation

Flotation technology has been employed successfully since the end of the 19th century

in ore processing, and also has been applied for several environmental applications,

including the treatment of wastewater (Kiuru, 2001; Rodrigues & Rubio, 2007).

The process of flotation consists of four basic steps (Wang et al., 2005):

1. Bubbles generation in the wastewater

2. Contact between the gas bubbles and the particles (or oil droplets) suspended in

the water

3. Attachment of the particles (or oil droplets) to the gas bubbles

4. Rise of the air/solids combination to the surface, where the floated material is

skimmed off

There are five different types of flotation systems, their classification based on the

method of bubble formation:

1. Dissolved air

2. Induced air

3. Froth

4. Electrolytic

5. Vacuum

Of these five, dissolved air flotation is the most commonly used flotation process in

industry and municipalities, and has undergone significant improvements in the last 50

years. (Wang et al., 2005).

4.1. Dissolved air flotation

Dissolved air flotation (DAF) is a unit operation that has been established as an efficient

removal technique of suspended solids from water in which air micro-bubbles attach to

the particles to make them float. Dissolved air flotation was first used for drinking water

clarification in the 1920’s, and played a significant role in water treatment in the 60’s

and 70’s in Scandinavian countries since DAF is highly effective in removing light

Air flotation 14

suspended solids if cold waters are treated. Finland for instance has more than thirty

plants where DAF is used as a clarification process in potable water treatment

(Heinänen et al., 1995; Kiuru, 2001). Three generations of configurations followed after

the first introduction, whereby modifications in tank design were the most significant

improvements.

A DAF system consists essentially of a flotation unit and a saturator (Figure I.6) (Bratby

& Marais, 1976). Typically, 15-50 % of the raw or treated water is saturated by

dissolving air under pressure in the saturator. Typical pressure differences of 4-5 bar

are selected, because of operational limits, as well as limits in efficiency since at higher

pressures the microbubbles are larger and the efficiency of separation then becomes

lower. This recycle flow is then introduced in the flotation tank through nozzles or

special valves (Edzwald, 2010; Wang et al., 2010). Because of the sudden pressure

drop, small microbubbles are generated with diameters between 10 to 100 µm. These

small air bubbles give the water a milky appearance, and is, therefore, called white

water.

Figure I.6: Overview of a conventional dissolved air flotation plant (Edzwald, 2010)

The flotation unit is meant to separate the solid phase from the liquid, and is divided

into two zones, separated by a baffle. The contact zone is at the front side of the tank,

and provides the opportunity for collision and subsequent attachment of the

microbubbles to the particles. The white water blanket will act as a micro-bubble filter,

and will retain particles, while water flows through it (Kiuru, 2001). In the following

separation zone free bubbles and bubble-particle aggregates have the opportunity to rise

to the surface of the tank (Figure I.7). This will give rise to a sludge layer being formed

Air flotation 15

at the surface that can be skimmed off, while the clear effluent (subnatant) can be

withdrawn from the bottom of the tank (Edzwald, 2010).

Figure I.7: Separation and contact zone in the flotation unit (Edzwald, 2010)

Coagulation and flocculation strongly influence the efficiency of the DAF-system.

Without addition of coagulants the particles and air bubbles carry negative zeta

potentials, causing repulsing forces. Important is that the coagulation will lead to little

or no particle charge, as to keep the electrostatic forces low. Stable attachment of a

particle to a bubble also requires hydrophobicity, which can be obtained by optimum

coagulation conditions (Edzwald, 2010). Stable bubble-particle agglomerates will not

form without a certain hydrophobicity of the particles (Gochin & Solari, 1983).

Another crucial parameter that influences the overall efficiency of DAF is the air to

solids (A/S) ratio . An increase in bubbles attached to the particles leads to increased

flotation velocity and subsequent increased degree of separation. To increase this ratio,

either the pressure in the saturator or the amount of water recycled should be increased,

leading to an additional turbulence and possible disturbance of the flow pattern in the

flotation unit. This turbulence can decrease the overall process efficiency, and,

Air flotation 16

therefore, the A/S ratio has to be optimal to obtain a high efficiency of separation (Wang

et al., 2010).

4.2. Applications of DAF in wastewater treatment

DAF has many different applications in wastewater treatment, ranging from enhanced

primary treatment to sludge thickening (Bratby & Marais, 1976; Ødegaard, 2001).

When used in primary treatment, the treatment results are equal to, or better than those

obtained by sedimentation (Ødegaard, 2001). The DAF can also replace the secondary

settler with better solid/liquid separation, since smaller particles can be removed.

Moreover, it is used to thicken or dewater waste activated sludge, obtaining a

concentration up to 7% solids with no chemical addition (Bratby & Marais, 1976).

Another application is tertiary treatment of secondary effluents in which the preferred

set-up is a combination of flotation and sand filtration in one single tank. The

advantages of this set-up are high removal efficiencies and cost savings, when

compared to other tertiary treatment processes (Wang et al., 2010). These examples

show the wide-range applicability and viable implementation of DAF in the WWTP, as

an alternative to conventional sedimentation.

4.3. Benefits of DAF compared to conventional

sedimentation

Many of the problems that can occur in a secondary settler are not inherent to the

flotation process (section 2.2). Bulking sludge should give similar results to normal

sludge in the flotation process, and any formation of gas (including nitrogen bubbles)

can only be advantageous for the process (Bratby & Marais, 1976).

El-Gohary & El-Ela (1980) compared the efficiency of a biological trickling filter

coupled with DAF or with a conventional settler (El-Gohary & El-Ela, 1980). This

work clearly showed the promising potential of combining a biological system with a

DAF for organics removal. Increasing organic loading rates did not have a significant

impact on BOD and COD removal efficiency of the combined system. This could mean

that DAF is more suitable for fluctuating organic loads, and offers a higher reliability

Air flotation 17

than settlers to treat sewage wastewater with daily fluctuations in organic load

(Munksgaard & Young, 1980).

Moreover, using the same amount of coagulant doses (alum and FeCl3) under the same

operating conditions, a higher removal efficiency of COD, biological oxygen demand

(BOD) and phosphate was achieved using DAF. Reduced amounts of coagulant are

necessary when DAF is used, compared to conventional settling (El-Gohary & El-Ela,

1980). This does not only lead to an economical benefit, but can also have an effect on

the anaerobic digestion of the sludge. The sludge separated by using DAF will also

have a higher solids content in comparison with a settler, due to the higher separation

efficiency (El-Gohary & El-Ela, 1980; Wang et al., 2005) The higher solids content

will lead to a higher biogas production, and can help in closing the energy-balance of

the WWTP.

Implementation of DAF instead of a settler can contribute to the improvement of the

environmental footprint of the WWTP, since a significant reduction of surface area and

volume requirement is realized. Also, the flotation retention time is short (10-12

minutes), compared to gravity settling (1-2 h) (El-Gohary & El-Ela, 1980). Although

the operational cost of DAF is slightly higher than a settler, due to the pressurization,

this is compensated by the considerably lower installation cost (Wang et al., 2005).

It is clear that DAF shows significant potential to be implemented in a WWTP, and has

the potential to overshadow the conventional settlers that are currently used

Anaerobic digestion and fermentation 18

5. Anaerobic digestion and fermentation

At WWTP level, recovery of bioenergy from wastewater is typically achieved through

anaerobic digestion (AD) to recover energy through biogas. Domestic sewage could

also be a source of products, such as nutrients or volatile fatty acids (VFA) that can be

used as building blocks for the production of valuable products (Angenent et al., 2004).

The AD is a process with a complex pathway where complex organic compounds are

converted into biogas under strict anaerobic conditions. Biogas is a mixture of 60-70

volume% CH4, 30-40 volume% CO2 and traces of H2S and H2. The produced biogas

can be used at WWTP level, to directly maintain the temperature of the anaerobic

digester, to fuel a combined heat and power engine or can be upgraded and injected in

the grid or used as transport fuel (Silvestre et al., 2015). Biogas upgrading involves the

removal of CO2 and other contaminants (H2S, H2, etc.), and is necessary to increase its

energy content, to make the transport over large distances economically and

energetically viable, as well as to avoid corrosion (Appels et al., 2008).

A typical implementation of AD in a WWTP can be seen in Figure I.8. The main

advantages of anaerobic digestion over other treatment techniques is the simultaneous

stabilization and minimization of the amount of sludge that has to be disposed of, while

producing renewable energy in the form of biogas (Appels et al., 2008). A major

drawback of anaerobic digestion is the high sensitivity of methanogens; sudden change

in pH, temperature or an increase in organic loading or toxic compounds could lead to

entire system failure (Wijekoon et al., 2011).


Figure I.8: Flowchart of the sludge processing line in a WWTP (Appels et al., 2008)

Due to the low market price of energy and the fact that biogas production cannot satisfy

the energy demands of our whole society, fermentation could be a more economically

attractive process (Angenent et al., 2004). The VFA produced through fermentation are

short-chain fatty acids consisting of six or fewer carbon atoms, and they have a wide

range of applications. They can be employed in the production of bioplastics, bioenergy

and the biological removal of nutrients from wastewater. Since the importance of a

transition from fossil fuel based economy to a bio-based economy is well recognized,

the current chemical commercial production of VFA should be gradually replaced by

biological production through fermentation (Lee et al., 2014).


5.1. Anaerobic digestion- a microbial process

The AD process consists of four sequential steps where a complex community of

bacteria will convert the organic compounds into a mixture of CH4, CO2 and new cells

(Figure I.9). The first step is the hydrolysis (i) in which complex macromolecules

(lipids, proteins and polysaccharides) are converted into simple soluble monomers

(fatty acids, aminoacids and sugar). Hydrolysis is generally the rate limiting step.

Several pre-treatment methods including thermal, mechanical and chemical methods,

have been extensively studied to increase the kinetics of this step (Feng et al., 2014).

These monomers are then fermented to a mixture of VFA and in less quantities to

alcohols during the acidogenesis/fermentation (ii). The VFA are subsequently

converted into acetic acid, CO2 and H2 during acetogenesis (iii). The H2 partial pressure

is a crucial parameter in this last reaction since the production of acetic acid is only

thermodynamically possible when the H2 partial pressure is sufficiently low, more

specifically lower than 10-4 atmosphere (Droste & Masse, 2000) (Gerardi, 2003). In the

final step the hydrogenotrophic methanogens (iv) convert H2 and CO2 to CH4 while the

acetoclastic methanogens (iv) convert the acetate to CH4 and CO2 (Angenent et al.,

2004). The hydrogenotrophic methanogens are crucial in keeping the H2 partial pressure

low to ensure stable and efficient digestion performance (Demirel & Scherer, 2008).

Figure I.9: Pathway of anaerobic digestion


The microbial consortium involved in the AD includes several different bacteria as well

as archaea. All methanogens are strictly anaerobic archaea while the other steps in the

AD pathway are carried out by bacteria, all of them having their own optimal

environmental conditions for growth (Liu & Whitman, 2008). The methanogenic

archaea, for example, are extremely sensitive to pH with an optimum between 6.5 and

7.5. Since the fermentative bacteria are less sensitive to pH and can operate in a wider

range of pH values, adjusting the pH can be a possible strategy of inhibiting the

methanogens and favoring fermentation to VFA (Appels et al., 2008).

5.2. Influence of iron and aluminium

As mentioned in Part I, section 4.3, a reduction of coagulant dose is possible when

implementing DAF instead of a conventional settler, and this can have its consequences

on the anaerobic digestion of the sludge.

Iron is needed by many micro-organisms for many reactions, and it is reported that

supplementation of iron enhances anaerobic digestion; in particular the acetogenic

methanogens have a high requirement of iron (Hoban & van den Berg, 1979; L. Zhang

& Jahng, 2012). The positive influence of iron can be attributed to multiple reasons

including the essential role of Fe in several metabolizing enzymes and the

detoxification effect of iron on sulfide inhibition ( Zhang & Jahng, 2012). In particular,

the carbon monoxide dehydrogenase (CODH) complex, key enzyme in the

methanogenesis from acetate, is composed of iron components (Zandvoort et al., 2006).

Thus the use of iron as coagulant enhances methanogenesis, increasing the CH4

production. However, in fermentation, iron does not seem to have any influence on the

VFA production. Replacing the iron by another coagulant, for example polyelectrolytes

or aluminium, could be a strategy to decrease the methanogenic activity and in its turn

reduce the hindering of the fermenting bacteria (Cagnetta et al., 2016).

Aluminium on the other hand shows the opposite effect of iron. Studies show that

aluminium can have an inhibitory effect on both methanogenic and acetogenic bacteria,

leading to a low conversion of the organic matter into methane. It is, however, also

reported that the bacteria may overcome the inhibition of Al, possibly showing the

ability of adaptation of the micro-organisms. The inhibitory effect can possibly lead to


an accumulation of VFA and a decrease in bicarbonate alkalinity, resulting in a decrease

of pH (Cabirol et al., 2003).

Research questions 23

6. Research questions

The AB-system is ideal as wastewater treatment technique to effectively treat the

wastewater, while also producing high amounts of biodegradable sludge that can then

be further digested or fermented in order to recover as much resources as possible.

Especially the A-stage is the innovative part, where a short SRT and high organic

loading rate lead to a minimal loss of biomass by endogenous respiration and a maximal

potential of resource recovery. Conventionally, the sludge from the A-stage is separated

by means of traditional gravitational settling. However, due the low sludge age, A-

sludge comes with some inherent settling problems, mainly due to the formation of

filamentous bacteria. In this thesis, dissolved air flotation is proposed as an innovative

technique to separate the A-sludge from the clean water. The separation of A-sludge by

flotation has not yet been investigated, so the primary goal of this thesis is to develop

and optimize a lab-scale DAF system combined with an A-stage reactor.

Coagulation and flocculation are essential processes to obtain robust sludge flocs

susceptible to flotation. The influence of three different coagulants, namely iron,

aluminum and a poly-electrolyte will be reviewed for their effect on flotation of A-

sludge. The harvested A-sludge will then be digested and fermented, whereby the

possible influence of the different coagulants on both processes can be evaluated.

In short, the main research questions are as followed. Is separation of A-sludge by DAF

feasible, and if so, can the organics recovery be higher than with a conventional settler?

Is there a difference in separation performance by using different coagulants? Are

digestion and fermentation effected by the use of DAF for separation of the sludge?

Does the use of different coagulants effect the digestion and fermentation behaviour of

the floated sludge?

25

Part II: Materials and methods

CSTR with concentrated SYNTHES 27

1. CSTR with concentrated SYNTHES

To obtain a mixed liquor volatile suspended solids (MLVSS) mimicking the

composition of a real A-stage an Erlenmeyer of 5 L was operated as a continuous stirred

tank reactor (CSTR) without retention of biomass (Figure II.1). As influent,

concentrated synthetic wastewater, SYNTHES, developed by Aiyuk and Verstraete

(2004) was used. This synthetic wastewater mimics favourably well actual raw

domestic sewage (Aiyuk & Verstraete, 2004) (Table II.1). Synthetic wastewater was

chosen to avoid big fluctuations in composition of the influent. Concentrated batches

of 10 - 20L SYNTHES were prepared, stored at 4 °C before usage and stored at room

temperature during usage.

The reactor was operated at a SRT equal to an HRT of 1 day. The effluent was collected

in vessels for further use and daily samples of the MLVSS and influent were taken and

analysed for chemical oxygen demand (COD), total suspended solids (TSS) and volatile

suspended solids (VSS). The Erlenmeyer was mixed by a magnetic stirrer, and aerated

with a cylindrical aeration stone (Scalare Oxy-tech., 70x15mm diffuser). A peristaltic

Watson Marlow pump (323 s/d, Germany) was used to pump in the influent as well as

pump out the effluent to obtain the desired SRT of 1 day. The pump was timed to

operate five minutes every hour, and to supply a daily amount of feed of 5L as well as

extract 5L of MLVSS.

Figure II.1 Scheme of the continuous stirred tank reactor


Table II.1: Composition of concentrated SYNTHES (Aiyuk & Verstraete, 2004)

Component Concentration (mg L-1)

Chemical compounds

Urea 1600

NH4Cl 200

Na-acetate.3H2O 2250

Peptone 300

MgHPO4.3H2O 500

K2HPO4.3H2O 400

FeSO4.7H2O 100

CaCl2 100

Food ingredients

Starch 2100

Milk powder 2000

Dried yeast 900

Soy oil 500

Trace metals

Cr(NO3)3.9H2O 15

CuCl2.2H2O 10

MnSO4.H2O 2

NiSO4.6H2O 5

PbCl2 2

ZnCl2 5

Overall parameters

COD total 8000

COD soluble 2500

COD particulate 5500

pH 7.1

Coagulation: jar testing 29

2. Coagulation: jar testing

A mechanical flocculator (Flocculator SW1, Stuart Scientific, FS685, United Kingdom) with 6

peddles was used for the coagulation experiments. The 6 beaker jar tests were conducted using

metal salts of FeCl3.6H2O (Sigma-Aldrich -236489) or AlC3.6H2O (Merck 1.01084.1000) as

coagulant, or a cationic polyelectrolyte Zetag 7651 (Ciba Specialty Chemicals, Groot-

Bijgaarden, Belgium) as flocculant.

FeCl3.6H2O and AlC3.6H2O give cationic hydrolysis products when dissolved in water, which

are adsorbed on negative particles and result in charge neutralization. At higher dosages and

optimal pH, FeCl3.6H2O and AlC3.6H2O hydroxide precipitates can enmesh impurities

resulting in removal of particles (Duan & Gregory, 2003). Zetag 7651 is a cationic flocculant

based on polyacrylamide with high molecular weight and high charge density. Due to the high

charge density, both charge neutralization and bridging could occur, resulting in removal of

particles (Bolto & Gregory, 2007). Zetag 7651 can thus also be used solely, as primary

coagulant, instead of being used as a flocculant aid and will be referred to as coagulant.

Stock solutions of the respective coagulants were made, and were used for a maximum of 7

days after preparation to avoid hydration/ hydrolysis in water of FeCl3.6H2O and AlC3.6H2O to

various hydroxide species and avoid disentangling of the polymer chains of Zetag (Bolto &

Gregory, 2007; Duan & Gregory, 2003).

The jar tests were intended to investigate the optimal dosage of coagulant that destabilized the

negatively charged particles, and allowed the sludge to form bigger flocs. The jar test was

conducted in four consecutive steps at room temperature. A volume of 200mL of MLVSS

originating from the CSTR was initially stirred in a beaker for 1 minute at 270 rpm to achieve

complete mixing. Then a certain amount of was added into the beakers to achieve a desired

concentration coagulant (FeCl3.6H2O / ALC3.6H2O concentration: 1, 2, 4, 6, 8 and 10 g L-1;

Zetag 7651 concentration: 20, 30 and 40 mg L-1). Next, rapid stirring at 270 rpm for 2 minutes

was carried out to ensure sufficient contact between the coagulant and the MLVSS. Then, gentle

mixing at 35 rpm for 15 minutes was included to allow floc growth (flocculation time). The

formed flocs were left to settle for 5 minutes to achieve separation between the subnatant and

the sludge in the decantation phase. Samples of the subnatant were taken for further analysis of

turbidity. The conditions for optimized coagulation were determined by measuring the turbidity

Coagulation: jar testing 30

of the subnatant, and for these optimal concentrations a lower flocculation time of 5 and 10

minutes at 35 rpm was also tested.

System set-up 31

3. System set-up

The set-up consisted of two main parts, the A-stage reactor and the DAF unit. The reactor served

as both biological and flotation unit, and was run in sequencing batch reactor (SBR)

configuration (unless stated otherwise) (Figure II.2). The DAF unit was used to pressurize

liquid and generate microbubbles with a size of 50-60 µm. All experiments are performed in a

temperature controlled room of 20°C. This was chosen to eliminate the possible influence of

temperature on the different processes.

Figure II.2: System set-up: 1. A-stage reactor, 2. Dissolved air flotation unit

1.

2.

System set-up 32

3.1. Materials

Pumping was performed using 5 peristaltic pumps for influent, effluent, waste, DAF and

coagulation (3 pumps Watson Marlow 323 s/d, Germany as P_in, P_effl and P_waste, Watson

Marlow 101 F, Germany as P_cg and Masterflex L/S economydrive model 7090-42 as P_DAF).

Aeration was achieved using an aquarium air pump with a cylindrical aeration stone (Scalare

Oxy-tech, 70x15mm diffuser). As A-stage reactor a 2L glass beaker was used, and mixing was

carried out using an overhead stirrer (Ika RW16). A pressure vessel (Prosep Zaventem-

Belgium) was used as DAF unit, which could be pressurized from the top using compressed air

(Figure II.3).

Four valves (Bürkert valves type 6013 for liquid and type 6027 for gas) were used to control

the flow of liquid or air. The two liquid valves controlled the liquid going to the DAF, as well

as the pressurized liquid coming from the DAF to the reactor. The first gas valve controlled the

compressed air going into the DAF to pressurize the liquid, the second acted as an escape valve

to release the pressure from the DAF. By using these valves, the system could be controlled by

a computer instead of manually regulated valves.

The pumps together with the stirrer, valves and the aeration pump were controlled by a

homemade digital timer driven by Labjack soft-and hardware (Labjack, 2015). The software

was programmed to turn the equipment on and off according to the timing of 1 cycle (see Part

II, section 4.1)

System set-up 33

Figure II.3: Scheme of the set-up

3.2. SBR cycle

The A-stage was operated as a SBR in which each cycle consisted out of 4 consecutive phases:

(i) fill, (ii) contact, (iii) float and (iv) draw (Figure II.4). This fill and contact phase lasted for

20 minutes with progressive filling of the SYNTHES together with mixing and aeration. This

favoured the conversion of the organics in wastewater into biomass. After the contact phase the

coagulant and/or flocculant was added to the reactor. A rapid mixing phase, followed by gentle

mixing occurred. After the gentle mixing phase, either 800 mL of MLVSS from the reactor or

clean water (depending on the experiment) was send to the DAF, and pressurized up to 5 ± 0.5

bars. After pressurization, the pressurized liquid was released in the reactor and the floating

phase followed. Clear effluent and floating sludge could be pumped out, and collected for

further analyses.

System set-up 34

3.3. Influent

As influent for the A-stage reactor, complex synthetic wastewater SYNTHES was used (Table

II.1). Concentrated batches of 10 – 20 L SYNTHES were prepared, and diluted ten times to

obtain a COD of on average 700 mg L-1.

Figure II.4 Sequencing batch reactor cycle

Development and optimization experiments 35

4. Development and optimization experiments

The main goal of this research was to develop and optimize an A-stage DAF. In the A-stage the

goal is to convert the soluble COD in the influent into biomass without excessive respiration.

During the coagulation phase, different coagulants and/or flocculants in different

concentrations can be used to provide good floc formation. After the coagulation phase, the

DAF is applied to separate the sludge to produce a clear effluent in the A-stage.

4.1. Development of A-stage DAF

4.1.1. Experiment 1 CSTR

The A-stage reactor was filled with 2 L effluent from the CSTR. Next, 4 g L-1 of AlC3.6H2O

was added as coagulant. The reactor content was then mixed fast (mimicking 270 rpm of the

flocculator, Part II, section 2) for 5 minutes to ensure sufficient contact between the MLVSS

and the coagulant, followed by a gentle mixing phase (mimicking 35 rpm of the flocculator,

Part II, section 2) of 15 minutes to provide floc formation (coagulation phase). After the

coagulation phase, 800 mL of MLVSS (40% of the reactor content) was sent to the DAF and

pressurized till 5 ± 0.5 bar for 5 minutes. Then, the pressurized liquid was released again in the

A-stage reactor.

4.1.2. Experiment 2 Real conditions

The reactor was filled with 0.5 L of effluent from the CSTR as inoculum or 0.5 L A-sludge

from the full-scale WWTP Nieuwveer (Breda, the Netherlands). A contact phase of 20 minutes

followed, in which the reactor was filled with concentrated (COD of 7000 mg L-1) or non-

concentrated SYNTHES (10-fold diluted) wastewater, while being stirred and aerated. After a

20 minutes contact phase, the aeration was stopped and a certain dosage of AlCl3.6 H2O

(concentrations in Table II.2) was added to the reactor. Then the reactor content was mixed fast

for 5 minutes, followed by a gentle mixing phase of 15 minutes. After the coagulation phase,

800 mL or 1200 mL (40 % or 60 % of the reactor content) of MLVSS was sent to the DAF and


pressurized till 5 ± 0.5 bar for 5 minutes. Then the pressurized liquid was released again in the

A-stage reactor.

4.1.3. Experiment 3 Coagulant and flocculant

The same procedure as for experiment 2 was used, but after the addition of a certain

concentration of AlC3.6H2O as coagulant, different concentrations of cationic (Kemira

Superfloc C-492) or anionic polymer (SNF Floerger, China) were added as a flocculant (Table

II.3). After the contact phase, the aeration was kept on during the coagulation phase, and the

pressurization in the DAF lasted for only 1 minute. After different trials, instead of sending 800

mL of MLVSS to the DAF and pressurize this, 800 mL of clean tap water was sent to the DAF,

pressurized, and then released back into the reactor, which in this experiment contained only

1.2 L of MLVSS instead of 2L.

4.1.4. Experiment 4 pH control

The same procedure as in experiment 2 was applied, but before the coagulation phase the pH

of the non-concentrated A-sludge was adjusted to a value of 8.40 using a 2 M NaOH solution.

4.1.5. Experiment 5 Increasing the VSS

A sample of 2 L of concentrated (± 12 g L-1 VSS) and non-concentrated (2-4 g L-1 VSS) A-

sludge from the WWTP Nieuwveer (Breda, the Netherlands) was brought into the reactor,

followed by a coagulation phase or without coagulation phase (concentrations in Table II.2). A

volume of 800 mL of this sludge was pumped to the DAF, and pressurized till 5 ± 0.5 bar for 5

minutes. After this, the pressurized liquid was brought back into the reactor.

A summary of experiment 1 to 5, with respective concentrations of coagulant, can be found in

Table II.2.

4.1.6. Experiment 6 Air dissolution

The DAF unit was filled with carrier material and the set-up was adjusted to put the liquid into

the DAF unit from the top instead of the bottom.


Table II.2: Summarizing table of the tested conditions; experiment 1-5

Conditions AlCl3.6H2O

(g L-1)

Al3+

(mmol L-1) VSS

(g L-1)

Experiment 1: 2L effluent CSTR 4 16.57 ± 2

Experiment 2:

0.5L effluent CSTR +

1.5L SYNTHES WW

4

3

2

1

0.75

0.5

0.2

16.57

12.43

8.28

4.14

3.11

2.07

0.83

1 - 4

Experiment 2:

0.5L A-sludge + 1.5L SYNTHES WW

4

0.25

16.57

1.04 1 - 4

Experiment 4:

pH adjustment A-sludge till pH of 8.40 1 4.14 2 - 4

Experiment 2:

0.5L A-sludge +

1.5L concentrated SYNTHES WW

4

1

16.57

4.14 ± 1.5

Experiment 5:

2L A-sludge (Nieuwveer)

0

1

0.25

0.00

4.14

1.04

2 - 4

Experiment 5: 2L concentrated A-sludge 0

1

0.00

4.14 ± 12

Experiment 2:

Pumping in DAF 60% MLVSS,

0.5L A-sludge + 1.5L SYNTHES WW

0.75 3.11 1 - 4


Table II.3: Tested concentrations of AlC3.6H2O in combination with cationic or anionic polymer in experiment 3

AlC3.6H2O ( g L-1) Cationic / anionic polymer (mg L-1)

0 1/ 2/ 3/ 4

0.1 1/ 2/ 3/ 4

0.05 1/ 2/ 3/ 4

0.075 2

0.03 2


4.2. Optimization of A-stage DAF

Determining optimal concentration of AlC3.6H2O, FeCl3.6H2O and Zetag 7651

The reactor was filled with 250 mL of A-sludge from the WWTP Utrecht (the Netherlands) as

inoculum. The DAF unit was filled with 100 mL of tap water and pressurized till 5 ± 0.5 bar.

During the filling and contact phase, 750 mL of SYNTHES wastewater was added to the

reactor, while the MLVSS in the reactor was aerated and stirred for 19 minutes. After 19

minutes, AlC3.6H2O or FeCl3.6H2O was added to the reactor as coagulant in combination with

a cationic polymer (Kemira Superfloc C-492) as flocculant or Zetag 7651 was added solely as

coagulant (Table II.4). A fast mixing phase of 1 minute, during which aeration and mixing were

kept on followed, and then a small amount of about 20 mL pressurized liquid was released in

the reactor. The mixer was then kept on, while the aeration was turned off, during which the

content of the DAF unit was released in a waste vessel, filled up again with 500 mL of tap water

and pressurized till 5 bar ± 0.5 bar. After 5 minutes mixing, the pressurized liquid was released

in the reactor and the flotation phase followed. After 5 minutes of flotation, clear effluent and

floated sludge could be pumped out and collected for further analysis in terms of total COD

(tCOD), soluble COD (sCOD), TSS and VSS.

Table II.4: Tested concentrations of AlC3.6H2O or FeCl3.6H2O in combination with Kemira

Superfloc 2000 and tested concentrations of Zetag 7651

AlC3.6H2O or FeCl3.6H2O (mg L-1) Kemira Superfloc C-492 (mg L-1)

30 3

50 1/2/3

75 3

Zetag 7651 (mg L-1)

5

7.5

10

Batch tests 40

5. Batch tests

5.1. Fermentation test

Fermentation tests were carried out to quantify the amount of VFA that can be produced from

the fermentation of the sludge. Fermentation batch tests were performed in triplicate in serum

flasks sealed with a rubber stopper and aluminium sealer. Total and working volumes were 120

mL or 80 mL and 80 mL or 60 mL, respectively. The tests were run at 35 ºC and pH 7 or pH

10. To control the pH, the flasks were opened on day 2, 4 and 7, and the pH was adjusted using

a 3 M NaOH solution. On day 0, and always after opening the flasks, the headspace was flushed

with N2, and then bottles were kept at 35 ºC, and shaken at 120 rpm for 9 days. Liquid samples

of 4 mL or 2 mL were taken at day 0, 2, 4, 7 and 9 for pH measurement and VFA analysis.

Volumetric biogas production and gas composition were measured on day 2, 4, 7 and 9.

Fermentation of the floated sludge

Mixed culture inoculum from a fermenter (CSTR) treating A-sludge (HRT 20 days, pH 10, T

35 ºC, obtained after a stable working period of 150 days) was used as inoculum for the

fermentation of the floated sludge. Controls with only inoculum or only substrate were also

included to normalize the final results. For the experiment, 10 mL of inoculum was mixed with

50 mL of floated sludge as substrate and 20 mL of tap water and added to 120 mL serum flasks.

Some fermentation batch tests were performed with 10 mL inoculum, mixed with 30 mL of

substrate and 20 mL of tap water and added to 80 mL serum flasks. For the controls, either the

inoculum or the substrate was replaced by tap water. All tests were carried out in triplicate.

Batch tests 41

5.2. Biochemical methane potential (BMP) test

The BMP tests were carried out to estimate the anaerobic biodegradability of the floated A-

sludge and to quantify the biogas production. The tests were performed in triplicate in 120 mL

serum flasks under mesophilic conditions (35 °C) with a working volume of 80 mL. Inoculum

or substrate were replaced by tap water in the control tests. The inoculum was obtained from

an in-house anaerobic sludge digester. The inoculum sludge was diluted with tap water until a

VSS concentration of 10 g VSS L-1 was obtained. The amount of substrate added to the flasks

was calculated based on a constant ratio of 0.5 g COD g-1 VSS.

Flasks were sealed with a rubber stopper and aluminium sealer, and then connected to glass

columns, in which biogas production was measured by means of water displacement. Biogas

composition was evaluated at the end of the experiment when the cumulative biogas production

stabilized (i.e. no observation of further gas production).

Chemical analyses 42

6. Chemical analyses

6.1. Chemical Oxygen Demand

The Chemical oxygen demand (COD) is the amount of oxygen required to oxidize organic

carbon completely to CO2 by chemical means. The COD analysis on sludge samples was

performed according to Standard Methods for the examination of water and wastewater (5220-

C, APHA, 1992). This describes the classical oxidation method using an excess of K2Cr2O7 in

an acid environment and high temperature in the presence of silver sulphate as catalyst. After

oxidation, the excess of unreduced K2Cr2O7 is titrated with an FeNH3SO4 solution with ferroin

as indicator. Samples were diluted till the range 100 - 900 ppm COD could be performed for

COD determination.

The COD analysis of non-coloured samples was measured using Nanocolor® kits (CODE;

Macherey-Nagel) of the type COD - 1500 with an analysis range of 100 - 1500 mg COD L-1.

For determination of sCOD, the samples were filtered with 0.45 µm pore diameter filters, and

then also analysed using Nanocolor® kits

6.2. Solids analysis

Total suspended solids (TSS) and volatile suspended solids (VSS) analysis was performed

according Standard Methods 2540D and E (APHA, 1997) in the range up to 20 g L-1. Total

solids (TS) and volatile solids (VS) were determined following Standard Method 2540G

(APHA, 1992). TS were determined by drying the sample in a crucible for 24 h at a temperature

of 105 °C. VS were measured by keeping the crucible at 550 °C for 1.5 h in a muffle furnace

(Nabertherm® LE6/11/B150, Germany).

TSS were determined by filtering a measured volume of sample over a 0.45 µm filter and drying

it at 105 °C for at least 1 h. VSS were determined by keeping the dried filter at 550 °C for 1.5

h in a muffle furnace.

The TS and TSS (g L-1) are quantified by subtracting the weight of the crucible / filter from the

weight after drying at 105 °C and dividing it by the volume of the original sample. The VS and


VSS (g L-1) are quantified by subtracting the weight of the crucible / filter after ashing at 550

°C from the weight after drying at 105 °C and dividing by the volume of the original sample.

6.3. VFA analysis

The VFA were analysed according to Andersen et al. (2014). The C2 - C8 fatty acids (including

isoforms C4 - C6) were measured by gas chromatography (GC-2014, Shimadzu®, The

Netherlands), equipped with a DB-FFAP 123-3232 column (30m x 0.32 mm x 0.25 µm;

Agilent, Belgium) and a flame ionization detector (FID). Detection limits are: acetate: 30 mgL-

1, propionate: 10 mg L-1, others: 2 mg L-1, upper detection limit: 1000 mg L-1. Liquid samples

were conditioned with sulfuric acid and sodium chloride and 2-methyl hexanoic acid as internal

standard for quantification of further extraction with diethyl ether. The prepared sample (1 µL)

was injected at 200 ºC with a split ratio of 60 and a purge flow of 3 mL min-1. The oven

temperature increased by 6ºC/min from 110 ºC to 165 ºC where it was kept for 2 min. The FID

had a temperature of 220 ºC. The carrier gas was nitrogen at a flow rate of 2.49 mL min-1.

6.4. Gas analysis

The gas phase composition was analysed with a Compact GC (Global Analyser Solutions,

Breda, The Netherlands), equipped with a Molsieve 5A pre-column and Porabond column

(CH4, O2, H2 and N2) and a Rt-Q-bond pre-column and column (CO2, N2O and H2S).

Concentrations of O2, N2, CH4, H2 and CO2 were determined by means of a thermal

conductivity detector with a lower detection limit of 1 ppm for each gas component.

6.5. Turbidity

A Hanna instruments portable microprocessor turbidity meter HI 93703 was used to measure

turbidity of the clean effluent collected after flotation. This instrument is used to measure the

light scattered by non-dissolved particles. It measures the light intensity under an angle of 90°

with respect to the incident light beam. This intensity is transformed by a microprocessor to a

FTU (Formazine Turbidity Unit) = NTU (Nephelometric Turbidity Unit) value. The range is

divided in two scales: 0 - 50 NTU and 50-1000 NTU. The resolution is respectively 0.01 and

1 NTU. The accuracy is 0.5 NTU for the range of 0 to 10 NTU, 5 NTU for the range of 10 to

50 NTU and 50 NTU for the range of 50 to 1000 NTU.

45

Part III Results


1. CSTR with concentrated SYNTHES

The goal of this experiment was to obtain MLVSS with similar composition as real A-stage

MLVSS. To this aim, a CSTR was operated at a SRT of 1 day for 41 days.

The profile of TSS and VSS of influent and reactor was fairly dynamic throughout the whole

experiment (Figure III.1 and Figure III.2). The TSS and VSS of the influent highly depended

on whether a new or older batch of SYNTHES wastewater was used. The lower TSS and VSS

of the community, in comparison with the influent, relates to the fact that there was no retention

of biomass in the CSTR, and all grown bacteria were washed out daily, resulting in a lack of

time to establish a stable community. At day 22, the reactor experienced a wash-out of biomass

due to pump failure, after which the VSS concentration dropped to 1 g L-1. The VSS

concentration restored slowly after this wash-out and attained values similar to those before the

wash-out (Figure III.1). The average TSS and VSS of the reactor was 2.27 g L-1 and 2.02 g L-1

respectively.

0.01.02.03.04.05.0

0 10 20 30 40

VSS

(g L

-1)

Time (d)

VSS REACTOR

VSS INFLUENT

Figure III.1: Reactor and influent volatile suspended solids (g L-1)


The COD of the reactor stabilized after approximately ten days and reached an average

concentration of 4.96 g O2 L-1 (Figure III.3). The COD of the influent fluctuated between 5 and

9 g O2 L-1, with an outlier of 12 g O2 L-1, throughout the whole experiment, again depending

whether a new or older batch of SYNTHES was used, and reached an average concentration of

7.23 g O2 L-1. COD removal obtained a value of 31.45 % during the experiment, which could

again be contributed to the lack of retention of any biomass in the reactor.

Figure III.3: Influent and reactor COD (mg O2 L-1)

0

2

4

6

8

10

12

14

0 10 20 30 40

g O

2L-1

Time (d)

INFLUENT

REACTOR

0.01.02.03.04.05.0

0 10 20 30 40

TSS

(g L

-1)

Time (d)

TSS REACTOR

TSS INFLUENT

Figure III.2: Reactor and influent total suspended solids (g L-1)

Coagulation jar testing 49

2. Coagulation jar testing

Based on turbidity measurements of the supernatant, the optimal concentration, for settling of

the sludge, of AlC3.6H2O, FeCl3.6H2O and Zetag 7651 was determined.

2.1. Concentration and type of coagulant

In these experiments, 2 operational parameters were varied, the concentration of coagulant and

the type of coagulant. The supernatant had the lowest residual turbidity when a concentration

of 1 g L-1 AlC3.6H2O, 2 g L-1 FeCl3.6H2O and 40 mg L-1 Zetag 7651 was present in the liquid

(159 ± 37, 130 ± 38 and 186 ± 27 NTU, respectively) (Figure III.4).

Figure III.4: Turbidity measurements (NTU) for different concentrations of AlC3.6H2O, FeCl3.6H2O and Zetag (n=3)

10 20 30 40

0

50

100

150

200

250

300

350

400

450

500

0 2 4 6 8 10 12

Concentration Zetag (mg L-1)

Turb

idity

(NTU

)

Concentration AlCl3.6H20, FeCl3.6H2O (g L-1)

AlCl₃FeCl₃PE (Zetag 7651)

Coagulation jar testing 50

2.2. Flocculation time

The influence of flocculation time was also investigated. Gentle mixing at 35 rpm was applied

for 5, 10 and 15 minutes and residual turbidity of the supernatant was measured to determine

the optimal flocculation time. This was respectively 10 minutes for AlC3.6H2O and 15 minutes

for Zetag 7651 (Figure III.5). When using FeCl3.6H2O as coagulant, almost no settling occurred

after 5 and 10 minutes, therefore, 15 minutes was concluded to be the optimal time for the

flocculation.

Figure III.5: Influence of different flocculation times (min) on turbidity (NTU) with AlC3.6H2O and Zetag 7651 (n=3). FeCl3.6H2O coagulated sludge did not settle after 5 and 10

minutes, thus turbidity is not shown.

0

50

100

150

200

250

300

0 5 10 15 20

Turb

idity

(NTU

)

Time (min)

AlCl₃

PE (Zetag 7651)


3. Development and optimization experiments


To develop the A-stage DAF, 7 operational parameters were investigated for their role in A-

sludge flotation. In each experiment, 1 of these parameters was varied, while the remaining

parameters remained constant. This allowed that unbiased conclusions could be made regarding

every parameter and their influence on the flotation.

3.1.1. Type of A-sludge

In experiment 1, sludge obtained from the CSTR was used as inoculum. This experiment did

not give rise to floating sludge. Instead of this sludge, real A-sludge coming from the WWTP

Nieuwveer (Breda, The Netherlands) was also used as inoculum. Again no flotation was

observed, instead sedimentation of the sludge occurred (Figure III.6).

Figure III.6: Sedimentation instead of flotation of the sludge when A-sludge

is used as inoculum after 5 minutes (left) and after 15 minutes (right)


3.1.2. Type of coagulant/flocculant

In experiment 3, AlC3.6H2O was used in combination with a cationic and anionic flocculant to

improve floc formation. It was also decided to try and pressurize 800mL of water instead of

MLVSS, since the flocs were subjected to severe shear stress during the pumping in the DAF,

and could be broken further down during the pressurization. To obtain stronger flocs, aeration

was kept on during the slow mixing phase, in order to increase the strength of the flocs and

make them more susceptible to flotation. None of these measures lead to successful flotation

but instead resulted in sedimentation.

3.1.3. pH of the MLVSS

Dosing of AlC3.6H2O lead to a significant pH decrease, depending on the concentration the pH

could drop to 3.7 ± 0.1. Addition of AlC3.6H2O to wastewater will always lead to a pH decrease,

but normally the buffering capacity of the water is high enough to avoid strong pH decreases.

The buffering capacity of the synthetic wastewater used in all experiments, however, was too

low for the high concentrations of AlC3.6H2O that were applied, so manual pH correction was

necessary.

It was chosen to conduct the experiment by controlling the pH at a neutral value. The pH of the

MLVSS was first adjusted with NaOH from a pH of 6.2 ± 0.1 to a pH of 7.7 ± 0.1. By doing

this, the pH decrease caused by the addition of AlC3.6H2O would result in a pH of 6.5 ± 0.3.

This pH control did not lead to successful flotation.

3.1.4. COD influent

In experiment 2, both concentrated (∼ 7 g L-1 COD) and non-concentrated SYNHTES (∼ 800

mg L-1 COD) were used as influent of the A-stage reactor, yet no flotation took place.

3.1.5. VSS of the MLVSS

Since employing concentrated SYNTHES as influent did not really influence the sludge growth,

gravity concentrated A-sludge (∼ 12 g L-1 VSS) was used. This ensured a higher sludge

concentration inside the reactor, resulting in abundant flocs for a dense sludge blanket. But

again, no flotation was observed.


3.1.6. Amount of bubbles

In experiment 2, instead of transferring 40% of the MLVSS to the DAF to be pressurized, 60%

of the MLVSS was sent to the DAF. No successful flotation followed.

3.1.7. Air dissolution into the liquid

To achieve proper dissolution of air into the liquid, the contact surface between air and liquid

had to be increased. To accomplish this, the DAF unit was filled with carrier materials. This led

to successful flotation of the A-sludge. Moreover, to create strong flocs resistant to shearing

due to vigorous contact between bubbles and particles, a small amount of pressurized liquid

was already released in the reactor during the coagulation phase.

A summary of all investigated parameters, and the result in terms of flotation can be found in

Table III.1.

In all further experiments executed, the DAF was filled with carrier material and 400mL of tap

water was send to the DAF instead of MLVSS, to avoid disruption of sludge flocs due to the

shear stress of the pump, when MLVSS would be send to the DAF. All experiments were

carried out in batch mode, but in continuous operation, instead of using tap water, part of the

cleared effluent after one cycle could be recycled back to the DAF, pressurized and used for

creation of microbubbles.

Table III.1: Summarizing table of all investigated parameters and the result on flotation

Operational parameter Result

Type of A-sludge Sedimentation

Type of coagulant/ flocculant Sedimentation

pH of the MLVSS Sedimentation

Influent COD Sedimentation

VSS of the MLVSS Sedimentation

Amount of bubbles Sedimentation

Air dissolution into the liquid Flotation


3.2. Optimization of A-stage DAF

To determine the optimal concentrations of all three coagulants that gave rise to the best

flotation, turbidity and VSS were measured of the MLVSS before DAF and of the effluent

obtained after DAF. From these measurements, removal efficiencies of turbidity and VSS could

be calculated, and could be used as parameters to choose the optimal concentrations.

For FeCl3.6H2O, first, 50 mg L-1 was tried with varying concentrations of polymer. Since the

combination with 3 mg L-1 gave the highest decrease in VSS, this concentration was kept

constant, and a varying concentration of FeCl3.6H2O was tried. The maximum decrease in VSS

and turbidity was 90 ± 1 %, obtained at a concentration of 75 mg L-1 FeCl3.6H2O in combination

with 3 mg L-1 polymer (Table III.2).

Table III.2: Decrease in VSS and turbidity (%) for FeCl3.6H2O (n=2)

FeCl3.6H2O

(mg L-1)

Polymer

(mg L-1) Decrease in VSS (%) Decrease in turbidity (%)

50 1 76 ± 7 85 ± 5

50 2 65 ± 0 75 ± 4

30 3 79 ± 1 66 ± 7

50 3 82 ± 8 77 ± 11

75 3 90 ± 1 90 ± 1

The same rationale as for FeCl3.6H2O was followed for AlC3.6H2O. The maximum decrease in

VSS and turbidity was 94 ± 1 and 93 ± 3 %, respectively, obtained at a concentration of 50 mg

L-1 AlC3.6H2O in combination with 3 mg L-1 polymer (Table III.3).


Table III.3: Decrease in VSS and turbidity (%) for AlC3.6H2O (n=2)

AlC3.6H2O

(mg L-1) Polymer (mg L-1)

Decrease in VSS (%) Decrease in turbidity (%)

50

30

2

3

89 ± 1

90 ± 7

86 ± 1

83 ± 5

50 3 94 ± 4 93 ± 3

75 3 94 ± 2 85 ± 2

For Zetag 7651, a concentration of 5, 7.5 and 10 mg L-1 was used and the maximum decrease

of VSS and turbidity was 97 ± 0 and 95 ± 1 % respectively, obtained at a concentration of 10

mg L-1 Zetag 7651 (Table III.4).

Table III.4: Decrease in VSS and turbidity (%) for Zetag 7651 (n=2)

Zetag 7651 (mg L-1 ) Decrease in VSS (%) Decrease in turbidity (%)

5 87 ± 8 90 ± 4

7.5 88 ± 2 80 ± 1

10 97 ± 0 95 ± 1

For all three coagulants, the highest removal in terms of VSS corresponded to the highest

removal of turbidity. Both parameters can thus be used as a measure of flotation efficiency.


3.3. DAF performance

The DAF performance was evaluated for all three coagulants by means of removal efficiency

of tCOD, sCOD and VSS (Table III.5). All three coagulants achieved high removal (> 80 %)

of TSS and VSS, AlC3.6H2O and Zetag 7651 achieved high removal (> 80%) of COD, while

only Zetag 7651 achieved removal of sCOD higher than 80%. The best average removal of

tCOD, sCOD, TSS and VSS was 84%, 81% 94% and 94%, respectively for AlC3.6H2O, Zetag

7651 and FeCl3.6H2O. Remarkably, these optimal removal values were achieved with a

different coagulant for every parameter. Looking at overall performance, Zetag 7651 could be

selected as the best performing coagulant.

The VSS removal efficiency strongly depended on whether a new or older batch of SYNTHES

wastewater was used. This could be visually observed; newer batches of SYNTHES led to much

clearer effluent than older batches (Figure III.7). This could also explain why the VSS removal

efficiencies obtained with the same concentration of coagulant, gave better results in Part III,

section 3.2, compared with the results in this experiment (Table III.5).

Table III.5: Removal efficiencies for all three coagulants in terms of tCOD, sCOD (n=3), TSS and VSS (n=2)

tCOD sCOD TSS VSS

Removal

efficiency

(%)

AlC3.6H2O 84 ± 7 24 ± 2 89 ± 12 88 ± 12

FeCl3.6H2O 75 ± 2 39 ± 6 94 ± 1 94 ± 1

Zetag 84 ± 1 81 ± 1 93 ± 0 91 ± 2


It was not possible to calculate the removal efficiencies of the A-stage coupled to DAF in batch

mode. New inoculum was used during every cycle, therefore no adaptation of the bacteria to

conversion and adsorption of the sCOD present in the SYNTHES wastewater into biomass

occurred. Also, by running the system in batch mode, no selection of biomass able to float the

sludge occurred (as it typically happens in a continuous A-stage system coupled with a settler

run in continuous mode that selects for settling biomass). Every cycle, 250 mL of A-sludge

(used as inoculum) was introduced, that contained a significant amount of extra COD, in respect

to the COD already present in the SYNHTES wastewater. This amount of COD also needed to

be oxidized during the contact phase, contributing to a higher initial COD than solely the

influent COD. When operated in continuous mode, the inoculum of the SBR would after a

certain time period have oxidized this COD, achieving steady-state conditions, which did not

occur due to batch mode operation.

3.4. Comparison DAF with settling experiments

For the optimal concentrations determined for settling (Part II, section 2.1) and flotation ( Part

III, section 3.2), the residual turbidity of the effluent can be compared (Table III.6). It has to be

taken into account that the settling experiments were conducted on effluent obtained from the

CSTR, while the DAF experiments were conducted on A-stage MLVSS, but since the

Figure III.7: Sludge separated through flotation with Zetag 7651 as coagulant, left: older batch of SYNTHES, right: new batch of SYNTHES


characteristics of the CSTR mimicked fairly the A-stage (average VSS and COD was 2.02 g L-

1 4.96 g O2 L-1), this comparison is possible.

The DAF achieved lower residual turbidity when employing AlC3.6H2O and Zetag as

coagulant, and achieved a slightly higher residual turbidity when FeCl3.6H2O was used as

coagulant. The optimal concentrations of coagulant used in DAF experiments, are more than

ten times lower for AlC3.6H2O and FeCl3.6H2O, and four times lower for Zetag than the optimal

concentrations in the settling experiments (AlC3.6H2O: 1 g L-1 versus 50 mg L-1, FeCl3.6H2O:

2 g L-1 versus 75 mg L-1, Zetag: 40 mg L-1 versus 10 mg L-1 using a settler and a DAF system,

respectively).

Table III.6: Lowest turbidity achieved with optimal concentrations in settling experiments and with DAF experiments

Settling (NTU) DAF (NTU)

(NTU) AlC3.6H2O 159.1 ± 37.0 120.8 ± 36

FeCl3.6H2O 129.67 ± 37.5 131.5 ± 15

Zetag 186.0 ± 26.85 118.5 ± 0.1

3.5. Sludge characterization

It was not straightforward to pump out the floated sludge layer without pumping out a

significant amount of water with it. This was due to the small water layer that was always

present on top of the floated sludge layer, as well as the dense thick sludge flocs that are not

easy to pump out. To correct for this, and to obtain a concentrated sludge that was similar to

the concentrated sludge in the reactor after flotation, the sludge that was pumped out was left

to settle for ± 1 hour, and the supernatant was removed from the sludge. On larger scale this

would not be necessary, since a scraping mechanism can be used to remove the floated layer

and drainage of interstitial water from this floated sludge layer can be executed (Bratby &

Marais, 1976).

All values in Table III.7 are measurements done on concentrated sludge where the water layer

was removed prior to the measurement. The sCOD, TSS and VSS of sludge obtained with all


three coagulants was very similar, only in terms of tCOD a difference could be observed. This

makes the comparison in fermentation and digestion behaviour in further experiments for all

three coagulants possible.

Table III.7: tCOD, sCOD, TSS and VSS of floated sludge coagulated with AlC3.6H2O, FeCl3.6H2O and Zetag (n=3)

AlC3.6H2O FeCl3.6H2O Zetag

tCOD (g O2 L-1) 16.9 ± 0.4 23.8 ± 2.6 21.6 ± 1.6

sCOD (g O2 L-1) 1.3 ± 0.0 1.3 ± 0.1 1.4 ± 0.2

TSS (g L-1) 16.8 ± 2 16.3 ± 1 16.1 ± 2.3

VSS (g L-1) 13.8 ± 2 13.6 ± 0.3 13.0 ± 1.9

Batch tests 60

4. Batch tests

4.1. Fermentation batch test

Fermentation batch tests using floated sludge coagulated with AlC3.6H2O, FeCl3.6H2O and

Zetag were carried out to determine the amount of VFA that can be produced from the floated

A-sludge, as well as to assess the influence of the different coagulants on the fermentation

behaviour. Sludge characteristics can be found in Part III, section 3.5.

The first fermentation batch test was conducted at pH 7, with sludge coagulated with

AlC3.6H2O and an inoculum acclimated to A-sludge as substrate at mesophilic conditions, and

a HRT of 4 days. Remarkably, the autofermentation yield, without inoculum, was higher than

the fermentation yield (Figure III.8). The maximum yield of VFA occurred at day 7 both for

autofermentation and fermentation with inoculum, and reached 337 ± 16 mg COD g-1 CODfed

and 278 ± 13 mg COD g-1 CODfed, respectively.

Figure III.8: Fermentation and autofermentation VFA yield (mg COD g-1 CODfed) for sludge

coagulated with AlC3.6H2O

050

100150200250300350400450

0 2 4 7 9

VFA

(mg

CO

D g

-1C

OD

fed

)

Time (days)

Fermentation Autofermentation

Batch tests 61

Propionate was the most abundant VFA (139 ± 8 mg COD g-1 CODfed for fermentation, 179 ±

16 mg COD g-1 CODfed for autofermentation at day 7). A very low methane production was

observed and (3 ± 0 mg COD g-1 CODfed for autofermentation and for fermentation on day 7,

respectively). The inoculum used in this experiment did not prove to be efficient in producing

VFA, so for further fermentation batch tests a different inoculum was selected.

Fermentation batch tests using sludge coagulated with FeCl3.6H2O, AlC3.6H2O and Zetag were

conducted at pH 10 using an inoculum acclimated to fermentation of A-sludge, at alkaline

conditions (pH 10) and HRT of 23 days.

The batch test using sludge coagulated with FeCl3.6H2O was also conducted at pH 7 to

investigate the influence of pH on VFA production (Figure III.9). At pH 7, the maximum VFA

yield occurred on day 14 and reached 184 ± 22 mg COD g-1 CODfed, while maximum VFA

yield at pH 10 occurred on day 9, reaching 341 ± 18 mg COD g-1 CODfed. A methane production

of 6 ± 1 mg COD g-1 CODfed was observed at pH 7 on day 14, while no methane production

was observed at pH 10 on day 9. Since the maximum at pH 7 was only reached on day 14, the

inoculum possibly needed time to adapt to the new pH. Overall, VFA yield at pH 10 was much

larger than at pH 7, so it was concluded to execute the following fermentation batch tests with

sludge coagulated with AlC3.6H2O and Zetag 7651 at pH 10.

Figure III.9: Fermentation VFA yield (mg COD g-1 CODfed) for sludge coagulated with FeCl3.6H2O at pH 10 and pH

0

50

100

150

200

250

300

350

400

2 4 7 9 14

VFA

(mg

CO

D g

-1C

OD

fed

)

Time (days)

VFA pH 10 VFA pH 7

Batch tests 62

Figure III.10 shows the VFA yields (mg COD g-1 CODfed) for all three coagulants at pH 10.

Maximum yield for sludge coagulated with AlC3.6H2O was reached on day 2, and obtained a

value of 401 ± 14 mg COD g-1 CODfed, while for FeCl3.6H2O and Zetag 7651 the maximum

yield was reached on day 9, and obtained a value of 341 ± 18 mg COD g-1 CODfed and 405 ±

23 mg COD g-1 CODfed, respectively. The maximum yield obtained for Zetag 7651 was thus

the highest, while remarkably the maximum yield obtained with AlCl3.6H2O was already

reached at day 2.

Figure III.10: Fermentation VFA yield (mg COD g-1 CODfed) for sludge coagulated with AlC3.6H2O, FeCl3.6H2O and Zetag at pH 10

On the peak production day, namely day 2, 9 and 9 for AlC3.6H2O, FeCl3.6H2O and Zetag 7651

respectively, the maximum VFA yield (mg COD g-1 CODfed) is shown in Figure III.11.

0

100

200

300

400

2 4 7 9 14

VFA

(mg

CO

D g

-1C

OD

fed)

Time (days)

AlCl₃ FeCl₃ Zetag

Batch tests 63

Figure III.11: Maximum VFA yield (mg COD g-1 CODfed ) on peak production day, namely day 2, 9 and 9 for AlC3.6H2O, FeCl3.6H2O and Zetag 7651

Moreover, VFA speciation was calculated on the day of maximum production. A similar VFA

speciation was found for AlC3.6H2O and FeCl3.6H2O (Figure III.12). In case of the AlC3.6H2O

coagulated sludge, propionate was the most abundant VFA (60 ± 5 %), besides 12 ± 6 % acetate

and 11 ± 0 % butyrate. In case of the FeCl3.6H2O coagulated sludge, propionate was again the

most abundant VFA, but in a lower abundance than the AlC3.6H2O coagulated sludge (49 ± 2

%). Moreover, a higher abundance of acetate (29 ± 2 %) was present. For the Zetag 7651

coagulated sludge however, the most abundant VFA was acetate (50 ± 2 %), besides 21 ± 1 %

propionate and 6 ± 2 % butyrate.

0

50

100

150

200

250

300

350

400

450


VFA

(mg

CO

D g

-1C

OD

fed)

Batch tests 64

Figure III.12: VFA specificity (%) for sludge coagulated with AlC3.6H2O, FeCl3.6H2O and Zetag 7651 at pH 10

4.2. BMP test

The sludge obtained from flotation with all three coagulants was tested for anaerobic

biodegradability and methane production. Characteristics of the sludge can be found in Part III,

section 3.5. The BMP test was stopped when a plateau in cumulative biogas production was

reached.

The floated sludge originating from coagulation with FeCl3.6H2O obtained a higher digestion

efficiency than the floated sludge originating from coagulation with AlC3.6H2O and Zetag 7651

(Figure III.13). The FeCl3.6H2O coagulated sludge was converted for 39 % into methane, while

the AlC3.6H2O coagulated sludge was only converted for 29% into methane and the Zetag 7651

coagulated sludge only for 24%.

0%

20%

40%

60%

80%

100%


Isocaproate

Valerate

isoValerate

Butyrate

isoButyrate

Propionate

Acetate

Batch tests 65

Figure III.13: Digestion efficiency (% CODCH4 COD-1fed) for sludge coagulated with AlC3.6

H2O, FeCl3.6 H2O and Zetag 7651

0

5

10

15

20

25

30

35

40

45

0 5 10 15

Dig

estio

n ef

ficie

ncy

(% C

OD

CH

4C

OD

-1fe

d)

Time (d)

FeCl₃.6H₂O AlCl₃.6H₂O Zetag 7651

67

Part IV Discussion

A-stage: state of the art 69

1. A-stage: state of the art

In view of the paradigm shift in modern society, from regarding wastewater as waste to

recognizing wastewater as an opportunity of resource recovery, the AB-system can play a

substantial role (Angenent et al., 2004; Meerburg et al., 2014). The low SRT and high organic

loading rate in the A-stage result in highly biodegradable sludge with limited loss of organic

matter due to extensive oxidation, and thus maximizing the incorporation of COD into biomass

(Nansubuga et al., 2015). Currently, the focus lies on anaerobic digestion of the A-sludge, since

the high methane yield of the A-sludge offers a high potential for energy recovery via biogas

production. However, methane has a low economic value, so more promising could be the

fermentation of the A-sludge into short chain VFA and subsequent conversion of these VFA

into high-volume fuels, alcohols, bioplastics (Polyhydroxyalkanoates (PHA)) or industrial

solvents (Agler et al., 2011; Kleerebezem et al., 2015; Lee et al., 2014).

Traditionally, A-sludge is separated from clean water by means of gravity settling, one of the

most critical operations in the activated sludge process (Jin et al., 2003). Due to the low SRT

of the A-sludge, filamentous bulking can prevail leading to poor settling and ultimately

resulting in increased effluent suspended solids and poorly thickened sludge (Ramalho, 2012;

Urbain et al., 1993). Therefore, further thickening/settling of the wastes sludge is typically

needed prior to anaerobic digestion or fermentation to increase the organics concentration of

the sludge. This leads to loss of organic matter, which defies the purpose of the A-stage.

A possible improvement in the performance of the activated sludge process could be realized

by implementing a more innovative separation process, increasing the effectiveness of solids

separation and concentration of the sludge (Javaheri & Dick, 1969). Dissolved air flotation

(DAF) is a possibility of a more innovative separation technique, by means of which higher

solids content in the floated sludge can be obtained (Wang et al., 2010). Higher solids content

could possibly eliminate the need of thickening the sludge before being used in an anaerobic

digester or fermenter.

In this thesis, the innovative combination of an A-stage with dissolved air flotation for solids

separation was investigated to combine the advantages of both processes, namely the maximum

incorporation of organic matter in sludge and more concentrated separated sludge, maximizing

the recovery of energy and materials present in wastewater. The effect on fermentation and

A-stage: state of the art 70

digestion of the floated A-sludge was also evaluated more in detail, as well as the possible effect

of using different coagulants for proper floc formation.

CSTR 71

2. CSTR

2.1. CSTR with concentrated SYNTHES

The goal of running a CSTR fed with concentrated SYNTHES wastewater was to achieve a

MLVSS with similar composition as the MLVSS concentration in a real A-stage. Since the

average VSS of the reactor was 2.02 g L-1, and the VSS concentration of an A-stage reactor is

on average between 2 and 4 g L-1, the obtained MLVSS did mimic the composition of a real A-

stage, and could thus be used for the coagulation jar testing in further experiments.

The COD removal efficiency obtained a value of 31 %, resulting in an average COD of 4.96 g

O2 L-1, and a COD:VSS ratio of 2.5. Since the goal of the experiment was to obtain a MLVSS

similar to an A-stage, this low COD removal efficiency contributed to the similarity of the

obtained MLVSS and a real A-stage. The average COD:VSS ratio of an A-stage operated at a

SRT of 1 day is 1.48 (Jimenez et al., 2015), thus the COD:VSS ratio in this experiment was

considerably higher. This could be attributed to the lack of retention of biomass in the reactor,

since the reactor was operated as a continuous stirred tank reactor and HRT was equal to SRT

(1 day). This results in less biomass in the reactor and less COD removal.

The conversion yield (g VSS produced g-1 COD removed) of CAS is 0.45 due to aerobic

conversion of COD to biomass (van Haandel & van der Lubbe, 2012). In an A-stage reactor,

the conversion yield can amount to 0.8 due to the adsorption and bioflocculation, inherent to

the A-stage process, taking place. Since only 10% of the COD is lost to CO2 due to unavoidable

mineralisation, a real A-stage achieves a total yield of 0.7 (Meerburg et al., 2014). If a

continuous A-stage reactor with biomass retention would have been run in this experiment, it

could be assumed that a total yield of 0.7 g VSS produced g-1 COD consumed would be

obtained.

2.2. Coagulation jar testing

Based on turbidity measurements an optimal concentration of all three coagulants could be

obtained. As reported in literature (Duan & Gregory, 2003), and also noticed in these

experiments, excess dosing of coagulants has a negative effect on the residual turbidity of the

CSTR 72

supernatant. Dosing higher concentrations than the optimum is the reason for this negative

effect, since charge reversal of the particles causes restabilisation.

Flocculation time also influenced the residual turbidity; an optimal flocculation time could be

found for AlC3.6H2O and Zetag 7651. For FeCl3.6H2O no effect of flocculation time on residual

turbidity was observed, however, this could also be attributed to the use of an older FeCl3.6H2O

solution, where extensive hydrolysis in water already had occurred, and negatively charged

hydrolysis products were formed instead of the desired positively charged hydrolysis products

(Crittenden et al., 2012).

Although, the concentrations determined in this experiment were optimal for sedimentation,

they are not necessarily the optimal ones for flotation. Possibly lower concentrations are more

suited for flotation, since less heavy flocs are formed. Similar experiments with the A-stage

DAF system are needed to determine optimal concentrations for flotation, but the

concentrations determined in this experiment could be used as general guidelines for the

flotation.

Development and optimization of A-stage DAF 73

3. Development and optimization of A-stage DAF


In the development phase, 7 operational parameters were investigated for their role in

solid/liquid separation by DAF. From these 7 parameters, 6 were not the reason why the sludge

could not be separated successfully. Nonetheless, these still played an important role in

solid/liquid separation by flotation, and need to be taken into account when optimal flotation of

A-sludge is desired.

I. Type of A-sludge

The MLVSS obtained from the CSTR did not float, therefore the first hypothesis was that the

sludge originating from the CSTR did not have the correct properties for flotation, such as

sufficient floc strength after coagulation and the presence of hydrophobic spots on the surface

for bubble attachment (Edzwald, 2010). Inoculum attained from a real A-stage was employed,

but again no flotation occurred.

This leads to the conclusion that the type of sludge used in the A-stage was probably not the

reason why the flotation was unsuccessful, since sludge obtained with synthetic wastewater as

well as real A-stage sludge did not give rise to flotation.

II. Type of coagulant/ flocculant

With the use of an anionic and cationic flocculant, the hypothesis that the formed flocs were

not strong enough for flotation was tested. If the flocs were not strong enough, they could

possibly be broken when the bubbles were released in the liquid. The use of polymeric

flocculants as coagulant aid to bridge the particles, that were first coagulated by the use of

aluminium and iron salts, can increase the toughness and size of flocs. Additionally, polymers

can restore floc shearing caused by overly vigorous contact of air bubbles with the flocs in the

DAF process (Bolto & Gregory, 2007).

Floc formation was not the main cause of unsuccessful flotation, since all measures undertaken

in this experiment would have led to strong flocs susceptible to flotation. However, adequate


aggregation of particulate matter into flocs is essential for efficient flotation, making

coagulation an essential parameter in the flotation process (Klute et al., 1995).

III. pH of the MLVSS

The pH control during coagulation with AlC3.6H2O is of crucial importance, since it controls

the precipitation of metal hydroxides, as well as the charges of the flocs (Binnie et al., 2002).

Since the largest pH drop after addition of AlC3.6H2O was 3.7 ± 0.1, manual pH control was

necessary.

Such a low pH can be detrimental to the bacteria in the sludge, and has an influence on the

aluminium species that are present in the MLVSS. Solubility of aluminium hydroxide is

dependent on the type of water and temperature, and the theoretical value can be derived from

solubility diagrams. Aluminium solubility is lowest at pH 6.2, so when sweep coagulation is

the preferred mechanism of coagulation, neutral pH values should be applied. However, below

pH 6.2, positively charged aluminium species dominate, aiding in charge neutralisation as the

destabilisation process (Binnie et al., 2002). The choice of optimal pH is, thus, not

straightforward, and depends on the circumstances wherein the coagulation should take place.

Since sweep coagulation generally leads to considerably improved particle removal than solely

destabilisation by charge neutralisation, neutral pH values are preferred. (Duan & Gregory,

2003).

It was chosen to conduct the experiment at neutral pH values, since no flocculant aid was used

and charge neutralisation alone would perhaps not be sufficient for good floc formation.

IV. Influent COD

The hypothesis behind trying both concentrated and non-concentrated SYNTHES wastewater,

was the possible insufficient growth of the bacteria in the A-stage, leading to insufficient flocs

and unsuccessful flotation. When using a higher influent COD concentration of 7 g L-1, it could

be assumed that more growth of sludge would occur, generating more flocs after coagulation,

as well as a denser sludge blanket to be floated. A higher amount of flocs could result in more

contact with the generated bubbles and a higher probability of flotation.

Influent COD was not the main cause of unsuccessful flotation. Since the goal was to create an

A-stage DAF that could be scaled up, non-concentrated SYNHTES that mimics real wastewater

was used in further experiments.


V. MLVSS

The problem of unsuccessful flotation was not to be attributed to the amount of sludge and the

related lack of flocs in the reactor, since the use of gravity concentrated sludge did not give rise

to flotation.

VI. Amount of bubbles

The hypothesis was that by increasing the amount of liquid transferred to the DAF, more

bubbles would be generated, resulting in increased contact between bubbles and sludge flocs.

It can be concluded that the volume of liquid that was send to the DAF to be pressurized is not

the most important parameter to take into account for sludge flotation. Question still remained

if the bubbles generated by the DAF were sufficient to provoke flotation of the sludge blanket.

VII. Air dissolution into the liquid

The last hypothesis to determine why the flotation of A-sludge had been unsuccessful so far,

was that the air dissolution into the liquid was insufficient, and subsequently not enough

microbubbles were generated. This appeared to be the main problem causing the unsuccessful

flotation.

Thus, the physical process of dissolving air into liquid is of critical importance. By improper

dissolution, the formed microbubbles are unable to determine sludge flotation.

A summary of all tested parameters and their relative importance in successful flotation of A-

sludge flocs can be found in Table IV.1.

Table IV.1: Summary of all 7 tested parameters and their importance in flotation

Tested parameters Importance in flotation

Type of A-sludge Sludge flocs should have sufficient hydrophobic spots for

bubble attachment.

Type of

coagulant/flocculant

Coagulation should result in strong flocs with low

electrostatic forces.


pH Since sweep coagulation is preferred, neutral pH values are

important when AlC3.6H2O is chosen as coagulant.

Influent COD

Sufficient COD should be present in the wastewater to ensure

sufficient growth of the bacteria, and consequently sufficient

flocs for formation of a sludge blanket.

MLVSS Sufficient sludge flocs should be present for flotation.

Amount of bubbles

Enough bubbles should be generated during pressurization to

ensure a large surface area and sufficient contact between

sludge flocs and air bubbles.

Air dissolution

Air dissolution is critical to generate enough small

microbubbles, preferably of 75 µm (Wang et al., 2010) , again

to ensure a large surface area for attachment air bubbles to

particles.

3.2. DAF: Benefits compared to conventional settling

To make a fair comparison between the efficiency of a lab-scale A-stage DAF that was used in

all experiments and a more conventional gravity settler, the two processes should have been

run in parallel with each other, making it possible to sample both and compare all characteristics

of the obtained effluent and sludge. Due to time constraints this was not possible, thus, to

evaluate the ability of the lab-scale DAF to separate A-sludge and clear effluent, and the

potential of implementing this process on a full scale in the AB-system, literature values of full

scale AB-systems will be reported and compared with the lab scale experimental values

obtained in this thesis.

In terms of turbidity, a comparison can, however, be done between the A-stage DAF and the

coagulation jar tests (Table III.6). Lower residual turbidity values were obtained when

AlCl3.6H2O and Zetag 7651 were used as coagulants rather than FeCl3.6H2O. Moreover, the

dosages required were lower (4 times lower for Zetag 7651 and 20 times lower for AlCl3.6H2O).

However, AlCl3.6H2O was used in combination with a cationic polymer for the flotation, while


this was not the case for the settling tests. The cost of the coagulants will be significantly lower

when DAF would be implemented for solid/liquid separation. Moreover, the metal salts used

as coagulant are non-biodegradable, ending up in the sludge and causing secondary pollution

(Zhao et al., 2000). When lower dosages are required, less secondary pollution occurs,

especially when polymers, such as Zetag, are chosen as coagulants, since these are

biodegradable.

The best removal efficiencies, in terms of COD, sCOD, TSS and VSS, obtained by using DAF

for solid/liquid separation were higher than 80 %. Comparing the TSS removal efficiency

values with full scale efficiencies of gravitational settlers for A-sludge in four Dutch AB-

WWTPs (Table IV.2), it can be concluded that the DAF performed better. The highest removal

efficiency achieved in WWTP Dokhaven (Rotterdam) was 68%, while the DAF achieved a

removal efficiency of 94% when FeCl3.6H2O was used as coagulant. The less efficient settling

of the A-sludge can be attributed to the high F/M ratio employed in A-stage reactors, as well as

the low DO level present (Ramalho, 2012). DAF does not suffer from these drawbacks; high

F/M ratio and low DO level do not seem to influence the separation efficiency of DAF. It always

has to be taken into account however that the comparison is done between lab-scale and full-

scale values.

Table IV.2 Removal efficiencies (%) of suspended solids of the gravitational settler in full-scale AB-WWTP’s (De Graaff & Roest, 2012)

WWTP SS removal efficiency (%)

Nieuwveer, Breda 59

Dokhaven, Rotterdam 68

Gamerwolde, Groningen 67

It was also reported by De Graaff & Roest (2012) that strong differences occurred in the wash-

out of suspended material from the gravitational settler of the A-stage to the B-stage of the

process, due to instability of settler operation. Since the use of flotation requires shorter

retention times, in the range between 15 to 40 minutes instead of 2 - 3h in settling tanks, higher


hydraulic loadings can be treated, and less instability would occur when fluctuating hydraulic

loadings need to be treated (Wang et al., 2010). These shorter retention times imply another

great advantage over the use of a settler, namely the much lower space requirement of DAF.

Therefore, the combination of DAF with the activated sludge tanks can allow to reduce space

requirements to only 15 % of that of a settler (Wang et al., 2010).

When comparing sludge characteristics from a real AB-WWTP (Nieuwveer) with a settler (in

terms of tCOD and TSS), and the lab-scale A-stage DAF, the better performance of DAF in

sludge concentration can be noticed. While the average tCOD and suspended solids of the

settled sludge in WWTP Nieuwveer (Breda, The Netherlands) was 11 g O2 L-1 and 7.4 g L-1

respectively, the highest average tCOD of the floated sludge was 23.8 ± 2.6 g O2 L-1 and 16.8

± 2 g L-1 total suspended solid. It can be hypothesized that using DAF instead of a settler in

full-scale can also achieve the same concentrated sludge characteristics, reducing the need for

further thickening of the sludge before fermentation or digestion. In this way, the concentrated

sludge can be sent directly to the fermenter or digester, without need for transport and avoiding

extra loss of COD during this transport.

Implementing DAF instead of the conventional gravity settling has great potential in terms of

more concentrated sludge, better suspended solids removal and less environmental impact.

However, lab scale results need to be confirmed on larger scale before definite conclusions can

be taken.

3.3. Optimal conditions for flotation

The use of Zetag 7651 as coagulant had the best results in terms of tCOD and sCOD removal,

while for FeCl3.6H2O the removal was better in terms of TSS and VSS. Since high

concentrations of COD are desirable for fermentation and anaerobic digestion, Zetag 7651 as

coagulant might be the best choice regarding the further application of the floated sludge.

Especially since also TSS and VSS removal are very high (93 and 91% respectively), Zetag

7651 can be considered as the best performing coagulant. Not only performance of the

coagulant should be taken into account, also the costs associated with the use of these

coagulants is an important aspect. The average price of polymers when bought in bulk volumes

is 3 euro kg-1 (personal communication with Bart De Gusseme), while the average price of alum

and iron chloride is 0.15 euro kg-1 (personal communication with Korneel Rabaey). A simple

economic comparison between the use of FeCl3.6H2O as coagulant and Zetag 7651 can be done.


Since an optimal concentration of 75 mg L-1 FeCl3.6H2O in combination with 3 mg L-1 polymer

gave the best flotation, and this concentration was used for treating 750 mL of SYNTHES

wastewater, this would result in a cost of 0.027 euro m-3 treated wastewater. The use of Zetag

7651 as coagulant would result in a cost of 0.04 euro m-3, for the optimal concentration of 10

mg L-1.

Despite the high cost of polyelectrolytes as primary coagulant in comparison with metal salts,

polyelectrolytes are biodegradable and this could compensate for the higher cost by higher

energy production as CH4 or higher VFA production. A trade-off between performance of

flotation and cost of coagulation should be made, including the possible positive or negative

effect of the different coagulants on the fermentation and digestion behaviour of the floated

sludge, in order to maximize organics incorporation in the sludge while still being economically

feasible on large scale.

Since dissolved air flotation has not yet been applied in combination with AB-system, no

comparison of the optimal concentrations of coagulant with similar systems can be carried out.

Coagulant concentrations for DAF as a tertiary treatment, after an aerobic treatment of the

wastewater and secondary clarifier, are, however available (personal communication with Bart

De Gusseme). In tertiary treatment, an optimal dose of 175 mg L-1 polyaluminium chloride in

combination with 0.65 mg L-1 anionic polymer was found. In comparison with the optimal

concentration in A-stage DAF, these concentrations are clearly higher (for example, optimal

dose AlCl3.6H2O is 75 mg L-1 in combination with 3 mg L-1 cationic polymer), regarding the

fact that A-stage MLVSS has a higher amount of COD and TSS than the secondary effluent.

To determine if lower doses of coagulants are indeed necessary when DAF is employed instead

of a conventional settler, coagulant doses of an A-stage combined with a settler are necessary.

These values could, however, not be found in literature, since metal salts are primarily dosed

in WWTPs for coagulation and phosphorus removal, making a fair comparison not feasible,

since phosphorus removal was out of the scope of this research.

80

4. Batch tests

4.1. Fermentation batch test

Since the biological nutrient removal (BNR) process, and, thus, removal of phosphorus and

nitrogen, is highly dependent on adequate supply of readily biodegradable COD, mainly in the

form of VFA in the feed wastewater, the production of VFA from A-sludge as an additional

carbon source can be a sustainable option for both utilization of waste and ensuring sufficient

BNR. It has been reported that 6-9 mg of VFA is required for biological removal of 1 mg of

phosphorus, hence, VFA production from A-sludge can strongly contribute to this when a high

enough amount of VFA can be produced (Zhang et al., 2009)(Q. Yuan et al., 2009).

Conventionally, fermentation of primary sludge is the main source of VFA, however, the

process control and reliability of VFA generation are often not adequate to always ensure

efficient removal of both nitrogen and phosphorus (Ahn & Speece, 2006). Fermentation of the

highly biodegradable A-sludge could be a more interesting source, since higher concentrations

of easily biodegradable organic polymers could entail shorter hydrolysis and fermentation times

(Q. Yuan et al., 2009).

The first batch test at pH 7 with AlCl3.6H2O coagulated sludge reached a 36% conversion of

COD to VFA. Since in literature it is reported that alkaline fermentation of secondary sludge

can increase the production of VFA over 3 times the production at acidic or uncontrolled pH, it

was chosen to conduct the following fermentation batch tests at pH 10 (H. Yuan et al., 2006).

Alkaline conditions enhance sludge hydrolysis, resulting in higher soluble substrate

concentration available for production of VFA. Alkaline fermentation also inhibits the

methanogens from converting the VFA into methane, since their growth and activity is inhibited

at pH > 8.5 (Lee et al., 2014).

The batch tests at pH 10 obtained the lowest conversion of COD to VFA (35%) for sludge

coagulated by using FeCl3.6H2O, while a 40 % conversion occurred for sludge coagulated by

using AlCl3.6H2O and a 40 % conversion by using Zetag 7651 to coagulate the sludge. The

influence of aluminium on fermentation has not yet been investigated, but Cabirol et al. (2002)

reported that accumulation of VFA instead of CH4 production was observed during anaerobic

digestion when sludge coagulated with aluminium was used as substrate. Similar to what was

81

observed in the experiments conducted on the floated sludge coagulated using AlCl3.6H2O,

mainly propionic acid was accumulated, confirming a possible positive role of aluminium in

fermentation.

Overall, conversion efficiencies were relatively high (> 30 %) in all conducted experiments in

comparison with values reported in literature, where primary or secondary sludge and A-sludge

were fermented (Cagnetta et al., 2016; Lee et al., 2014). Typically, VFA yields are reported in

literature in terms of g CODVFA g-1 VSSfed, thus to make a fair comparison, VFA production

and yield values were converted in the same unit. This conversion results in 0.58 ± 0.02 g

CODVFA g-1 VSSfed, 0.49 ± 0.03 g CODVFA g-1 VSSfed and 0.63 ± 0.03 g CODVFA g-1 VSSfed for

FeCl3.6H2O, AlCl3.6H2O and Zetag 7651, respectively. In studies regarding fermentation of

primary or secondary sludge, production between 0.2 – 0.368 g CODVFA g-1 VSSfed are reached,

these are considerably lower than the values obtained with the floated sludge (Ahn & Speece,

2006; P. Zhang et al., 2009). Regarding the fermentation of A-sludge, Cagnetta et al. (2016)

obtained a 30 % conversion of COD to VFA using thermophilic conditions at pH 7, while in

this study higher conversion efficiencies were obtained using mesophilic conditions at pH 10.

A-sludge could, thus, be a more suitable source for on-site VFA fermentation, always ensuring

sufficient BNR.

Based on these VFA yields, a theoretical calculation could be made to verify if the VFA

production from the floated A-sludge would be sufficient to remove the phosphorus present in

domestic wastewater to a concentration that meets the regulatory standards. As example, the

WWTP in Utrecht was considered, where 74 455 m3 day-1 of wastewater was treated in an AB-

system, with an influent phosphorus concentration of 7 mg L-1, and an A-sludge production of

9 378 kg VSS day-1 in the year 2011 (De Graaff & Roest, 2012). To remove the phosphorus

biologically to an effluent concentration of 1 mg L-1, 3780 kg VFA day-1 would be necessary

(70 000 m3 day-1 * 6 mg L-1 phosphorus that needs to be removed equals 420 kg P day-1 and

assuming 9 mg L-1 VFA is necessary per mg P). If all the A-sludge produced per day would be

fermented, and a VFA yield of 0.63 g CODVFA g-1 VSSfed is assumed, which was the VFA yield

obtained for the sludge coagulated with Zetag 7651, a VFA production of 5908 kg day-1 would

be obtained, which is 1.5 times the amount that would be necessary to remove the phosphorus.

This proves the potential of fermenting the floated A-sludge to provide VFA to ensure sufficient

biological phosphorus removal.

82

Remarkably, in the autofermentation batch test (no inoculum) of sludge coagulated with

AlCl3.6H2O carried out at pH 7, the COD conversion into VFA reached a 34 % efficiency. This

is relatively high, considering that by using an adapted inoculum, the VFA production

efficiency was 40 % when the fermentation was carried out at pH 10. Although pH 10 gave

better results in terms of VFA production with sludge coagulated with FeCl3.6H2O, pH 7 might

be better suited for sludge coagulated with AlCl3.6H2O. The fermentation batch test at pH 7

might reach even higher conversion efficiencies when an adapted inoculum would be used. This

creates the incentive to search for a better suited inoculum, and the optimization of the

experimental conditions for the fermentation batch tests.

4.2. BMP test

To recover the maximum amount of energy possible, the sludge obtained by flotation should

be highly biodegradable anaerobically. A BMP test was conducted to quantify the methane

production and anaerobic biodegradability of the floated sludge, as well as to investigate the

possible influence of the different coagulants on methane production.

It is observable that the methane production is influenced by the type of coagulant used (Figure

III.13). The digestion efficiency was highest for the sludge coagulated with FeCl3.6H2O (39%).

This corresponds with what is reported in literature, namely that iron plays an important role in

several enzymatic reactions occurring during methanogenesis, resulting in the possible

stimulation of methane production by the addition of iron (Zandvoort et al., 2006). The

digestion efficiency of the floated sludge coagulated with Zetag 7651 obtained the lowest

digestion efficiency of 23 %, and showed the slowest kinetics. Although it is demonstrated that

the polyelectrolyte partially could be digested anaerobically, it can also decrease the methane

production, presumably due to the much greater floc size, causing higher resistance to mass

exchange within the sludge flocs (Chu et al. , 2003). The VFA production of the Zetag 7651

coagulated sludge, however, was over 40 %. Therefore, it is unlikely that resistance to mass

exchange is the reason for the low digestion efficiency, since this mass resistance would also

result in lower VFA production. The digestion efficiency of sludge coagulated with AlCl3.6H2O

was 29 %. It is reported in literature that aluminium at concentrations higher than 1000 mg L-1

can have an inhibitory effect on the methanogenic activity, but that this inhibitory effect can be

overcome on long term (Cabirol et al., 2003). Since only 75 mg L-1 of AlCl3.6H2O was used to

float the sludge, and the BMP batch test only lasted for 15 days, inhibition did not occur and

83

rather the FeCl3.6H2O boosted the methane production instead of concluding that there was any

inhibitory effect of AlCl3.6H2O.

Overall, digestion efficiency based on g COD g-1 COD basis was low for all coagulants (< 40

%). This was not expected, since the conversion efficiency of A-sludge to CH4 is typically 50 -

70 %. This might have several reasons: (i) the methane production was influenced by the use

of the different coagulants; (ii) the use of synthetic wastewater instead of real wastewater leads

to less biodegradable sludge; (iii) the BMP test failed because VFA were produced, and pH

decreased; (iv) the inoculum was not suited for the floated sludge. Since iron is normally

boosting methane production, and the used coagulant concentrations are low, the coagulants

are probably not the reason for low digestion efficiency. No VFA production was observed in

the experiments, thus, the BMP test did not fail. No positive control of settled A-sludge obtained

with synthetic wastewater was included, thus, it cannot be verified whether the synthetic

wastewater resulted in a type of sludge that is less biodegradable than A-sludge. The VFA

production, however, was relatively high (> 35%), higher than that reported in literature from

Cagnetta et al., (2016) (30 %). Thus the sludge should not be less biodegradable than normal

A-sludge. A possible reason for the low CH4 production is the inoculum that is not adapted to

the floated sludge or the use of an inoculum which was too old, and this could be verified when

a different inoculum is tested.

85

5. General conclusions

In this thesis, the feasibility of combining an A-stage with dissolved air flotation for solid/liquid

separation was demonstrated. High removal efficiencies of tCOD, sCOD, TSS and VSS were

obtained, higher than 80 %, depending on which coagulant was used for proper floc formation.

A crucial parameter for flotation of sludge was the proper dissolution of air into the liquid,

needed to create microbubbles that were able to attach to the sludge flocs and provoke their

flotation.

The choice of coagulants influenced the removal efficiencies. The highest removal of COD and

sCOD was achieved by using Zetag 7651 (84 % and 81 % respectively), while highest removal

efficiency of TSS and VSS was achieved with FeCl3.6H2O (94 and 94 % respectively). Overall,

the removal efficiency of TSS and VSS with Zetag 7651 was not considerably lower than with

FeCl3.6H2O, thus, it could be concluded that Zetag 7651 was the best choice of coagulant.

The floated sludge characteristics proved that a concentrated sludge layer in terms of COD,

TSS and VSS could be obtained through flotation (for example with FeCl3.6H2O, a sludge layer

with 24 g L-1 COD and 16 g L-1 TSS was obtained). Since the highly biodegradable nature of

the A-sludge resulted in opportunities for resource recovery in the form of anaerobic digestion

and fermentation, it is important that the solid/liquid separation is highly efficient as possible

to exploit this potential. When compared with full-scale values of settled A-sludge, the COD

and TSS of the floated sludge were approximately twice as high, demonstrating the capability

of DAF to obtain a more concentrated sludge layer, and possibly resulting in a higher potential

of resource recovery.

In light of the carboxylate platform, fermentation of A-sludge in VFA could be a more valuable

and interesting approach than anaerobic digestion to CH4. These VFA are building blocks for

the production of valuable products, such as biopolymers, medium or long chain fatty acids or

biofuels, and have a higher economic value than CH4. A relatively high production of VFA was

obtained (40 % conversion of COD into VFA when AlCl3.6H2O and Zetag 7651 coagulated

sludge was fermented), proving the significance of obtaining a concentrated sludge layer after

solid/liquid separation. This suggests that implementing DAF instead of a conventional settler

would be more beneficial in view of resource recovery. Only FeCl3.6H2O seemed to influence

the fermentation behaviour in a negative way, since the lowest conversion efficiency was

86

obtained (35 %), making the choice of coagulant in the DAF important for further application

of the sludge.

Although it was expected that high digestion efficiencies would be obtained with the floated

sludge, CH4 production was low, even for the FeCl3.6H2O coagulated sludge where the highest

efficiency was expected (only 35 % digestion efficiency). However, VFA production and yield

values were relatively high, proving that the floated A-sludge is indeed biodegradable. Possibly

the BMP test failed due to the use of an unsuited inoculum, and, consequently, the BMP test

should have to be repeated to investigate why the CH4 was lower than expecte

Future perspectives 87

6. Future perspectives

Based on the results obtained in this thesis, several research possibilities come forward. First of

all, the A-stage DAF should be run in continuous mode instead of in batch mode, mainly

because the MO in the sludge can then adapt to the synthetic wastewater, and to flotation instead

of sedimentation. This adaptation is important, since it can further improve the efficiency of

solid/liquid separation by DAF and improve the COD removal in the A-stage itself. When the

A-stage DAF is run continuously, an overall removal efficiency of the A-stage in combination

with the DAF can be calculated and compared to overall efficiencies of A-stage settlers. In

addition, a COD balance could be made to investigate the yield of the A-stage and the amount

of COD present in the influent that can be harvested as sludge. As such, the real potential of

implementing DAF could be investigated more in detail, especially the innovative combination

of A-stage and DAF.

To make a fair comparison between the DAF and a conventional settler, both in terms of

removal efficiency and fermentation and digestion efficiency, a parallel reactor should be run

with a conventional settler next to the A-stage DAF. This would result in unbiased conclusions

about the possible better implementation of DAF in combination with an A-stage. Also, the

harvesting of the floated sludge should be optimized, since pumping out the sludge does not

seem to be an efficient option. A possible solution is the use of an overflow mechanism that

might improve the harvesting. As mentioned before, this would not be of importance in larger

scale applications, since specialized scrapers are already available that are efficient in scraping

of the sludge without taking out too much water with it.

Since the BMP test did not give the expected results, these tests should be conducted again,

possibly with a young inoculum that is already adopted to digestion of floated sludge. This

might result in a higher digestion efficiency, and further improve the resource recovery potential

of flotation of A-sludge.

Further on, upscaling the combination A-stage DAF and both fermentation and digestion of the

floated sludge are necessary for the process to be implemented on larger scale applications. Not

only the harvesting of the sludge will be an important parameter to take into consideration, but

also the creation of the correct microbubbles, as well as the coagulation will play a significant

role on larger scale. All of these parameters will need to be optimized in the upscaling, as they

88

were optimized on lab scale. The optimization of fermentation and digestion by the use of an

inoculum adapted to floated sludge might also prove to be of importance to achieve the

maximum possible resource recovery.

89

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