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PHYSICOCHEMICAL STUDIES OF MICROBIAL FLOCS
BAOQIANG LIA0
A tbesis submitted in conformity with the requirements
for the Degree of Doctor of Philosophy,
Graduate Department of Chernical Engineering & Applied Chemistry,
University of Toronto
O Copyright by Baoqiang Liao 2000
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PHYSICOCHEMICAL STUDIES OF MICROBIAL FLOCS
Baoqiang Liao, Ph.D.
Department of Chernicd Engineering & Applied Chemistry, University of Toronto
June 2000
ABSTRACT
This thesis investigates the idluence of sludge retention tirne (SRT) on surface properties
and interparticle interactions of sludge flocs, and the role of surface properties and interparticle
interactions in bioflocculation, compaction and stability. Four laboratory-scale sequencing
batch reactors (SBRs) were operated in parallel at different SRTs (4 to 20 days) for two years.
Surface properties of sludge flocs were examined by a number of state-of-the-art techniques,
including physicochemical extraction, chemical analyses and ~Itrafiltration of extracellular
polymeric substances (EPS), contact angle measurement (hydrophobicity) and colloidal
titration (surface charge). Interparticle interactions (electrostatic, ionic interactions and
hydrogen bonds) and the stability of sludge flocs at differmt SRTs were evaluated in batch
expenments by manipulating the water chemistry of the surrounding solutions.
At al1 SRTs, proteins and carbohydrates were the dominant components of EPSI followed by
a small amount of DNA. Acidic polysaccharides were not detected in the EPS. Molecular
weights of proteins, carbohydrates and DNA in the EPS covered a broad range, fkom less than
1,000 daltons to more than 100,000 daltons. The majority (>85%) of EPS components had
molecular weights larger than 10,000 daltons. The total amount of EPS was independent of the
SRT. The ratio of proteins to carbohydrates in the EPS, however, increased fiom 1.5 (c0 .9) at an
SRT of 4 days to 5.1 ( 2 1.5) at an SRT of 12 days, and then reached an almost constant value of
4.5 (I 1.9) at SRTs from 16 to 20 days. A larger arnount of EPS was associated with a higher
sIudge volume index (SVI), but no significant correlation was found between the amount of EPS
and effluent suspended solids (ESS).
Sludge surfaces were more hydrophobic (a higher value of water contact angle) and less
negatively charged at higher SRTs (16 and 20 days) than those at lower SRTs (4 and 9 days).
The contact angle values were strongly correlated to the surface charge density. A change in the
proportion of EPS components provides partial explanations for changes in hydrophobicity and
surface charge with respect to the SRT. A higher contact angle and lower surface charge were
associated with a lower level of ESS. There was no correlation, however, either between the SV1
and contact angle or between the SV1 and surface charge.
Changes in dissociation constants of sludge flocs in batch experiments indicated that
electrostatic interactions were involved in disruptinp the stability of sludge flocs, and ionic
interactions and hydrogen bonds were present to compensate for the negative influence of
electrostatic interactions on the stability of sludge flocs and to keep flocs together. Ionic
interactions and hydrogen bonds were two dominant forces that maintained the stability of sludge
flocs at lower SRTs: while other mechanisms, such as physicai enmeshment and hydrophobic
interactions. were likely more important than ionic interactions and hydrogen bonds in
controlling the stability of sludge flocs at higher SRTs. Sludge flocs at higher SRTs (16 and 20
days) were structurally more stable than those at lower SRTs (4 and 9 days).
In summary, the results from this study demonstrate that it is possible to operationally and
chemically manipulate surface properties and interparticle interactions of sludge flocs for
effective floc separation. A transition in floc surface properties \vas found to occur in the SRT
range of 9 to 12 days. This transition in floc surface properties was linked to irnproved
bioflocculation and a more stable fioc structure. It is the physicochemical properties of sludge
flocs, such as hydrophobicity and surface charge, rather than the quantity of EPS, that govem the
flocculating ability. In contrast, the total EPS content plays a more important role in determining
the compressi bi lity of sludge flocs than hydrophobicity and surface charge. A conceptual mode1
is proposed to describe the floc structure of floc-forming microorganisms at different SRTs.
1 am grateful to my supervisors. Prof Steven N. Liss and Prof. D. Grant Allen, for introducing me into this interdisciplinary and challenging area -"Biofilm Egineenng and Particie Technology" and for their inspiration, direction and encouragement throughout the course of this study and preparation of this thesis. The financial support arranged by Prof. Steven N. Liss through NSERC strategic grants for my graduate studies and attending conferences was highly appreciated. Special thanks also go to both supervisors for their advice and patience in improving my English skills.
1 would like to express my sincere acknowledgments to the following:
-ProE Michael V. Sefton (Chemicat Engineering)? Prof. A. Wilhelm Neumann (Mechanical Engineering) and Prof. Gary G. Leppard (Biology, McMaster University) for their interest. time and guidance in my graduate studies and for serving on my reading committee.
-Dr. Ian G. Droppo of the National Water Research Institute, Environment Canada for his helpful advice throughout this project.
-Prof. David M. Bagley (Civil Engineering) for teaching me the fundamentals of biological wastewater treatrnent, his thoughtful cornments and discussion in revising my manuscripts, and his interest and time in participating in my examination committee.
- Prof. L. Diosady (Chemical Engineering) for his interest and time in participating in my examination committee.
- Prof. D. Berk (Chemical Engineering) at McGill University for his interest and tirne served as the external examiner of my thesis.
-Prof. D. Reeve (Chemical Engineering) and the Pulp and Paper Centre at the University of Toronto for providing me with tremendous opportunities to improve my communication skills and Ieadership ability.
-Mrs. Zenka Policova of the Applied Surface Therrnodynamic Laboratory at the University of Toronto for her advice and help in the measurement of contact angles.
-Members of the Environmental Biotechnology Laboratory at Ryerson Polytechnic University: Dr. K. G. Mo, Mee Cheung, Jennine C. Finlayson, Valerie Famgia, Renata Bura, Sonia Na and Min-jin Kim for their help, friendship and advice.
-Members of the Bioprocess Environmental Laboratory at the University of Toronto: Christina and Chandra for their friendship and advice; particularly K. H. Chee, Livia Lau and Raymond Chin for their technical assistance throughout this work.
Last but not least, 1 would like to thank my wife, Jing Li, for her patience, understanding and support during the long-term "baby-sitting" of the bioreactors. To her, this thesis is dedicated.
TABLE OF CONTENTS
. . .................................................................................................. ABSTRACT 11
................................................................................ ACICIYO WLEDGMENTS iv
............................................................ TABLE OF CONTENTS ............... ... v
......................................................................................... LIST OF TABLES ix
........................................................................................ LIST OF FIGURES x
. . ....................................................................................... NOMENCLATURE mi
.................................................................... CHAPTER 1 INTRODUCTION 1
1.1 Biosolids-liquid Separation Problems .................................................. -1
1.2 A Rationai Approach to Improving Biosolids-liquid Separation ..................... 2
1 -3 Motivation of the Present Study ......................................................... -4
1.4 Research Goal and Objectives ........................................................... -6
.............................................................................. 1.5 Thesis Outline -6
CHAPTER 2 LITERATURE SURVEY ............................................................ 8
2.1 Activated Sludge Flocs ..................................................................... 9 .................................... 2.2 Bioflocculation, Settleability and Cornpressibility 14
2.3 Interfacial Forces lnvolved in Microbial Floc Formation ........................... -16
2.4 Mechanisms of Sludge Floc Formation ................................................ -18
2.5 Process-related Factors Affecting the Formation and Properties of
Sludge Flocs ............................................................................ - 2 5
2.5.1 Microbiological factors ........................................................ 25
.......................................................... 2.5.2 Sludge retention time 25
2.6 Measurement and Analytical Methods for Sludge Surface Characterization ...... 31
2.7 Summary of Literature Survey ......................................................... -38
2.8 Significance of this Study ............................................................... -39
........................ CHAPTER 3 EXPERIMENTAL MATEMALS AND MEITHODS 42
................................................................. 3.1 Experimental Approach -42 3 -2 Experirnental System ................................................................... -48
3.3 Measurement and Analytical Methods ................................................. 53
3 3 . 1 Standard wastewater analyses ............................................. 3 3 - 3 ................................................................. 3 -3 -2 Microbiology - 5 5
3.3.2.1 Microscopic observation ............. .. .......................... 55
3 .3.2.2 Phenotypic fingerprinting .......................................... 55
3.3.3 Extraction and Chemical Analyses of Extracellular
........................................................ Polymeric Substances -56 ............................................................ 3 -3 3 . 1 Extraction 56
3 -3 -3 -2 Chemical analyses .................................................. -57
3.4 Moiecular Weight Distribution of Extracellular Polymeric Substances ............ 60
................................. 3.5 Contact Angle Measurement ......................... .. -61
........................................................... 3.6 Surface Charge Determination 64 . . .............................................................................. 3.7 Stability Test -65
. . ............................................................ 3.7.1 Dissociation of flocs 66
......................................................... 3 .7.2 Dissociation constant -66
......................................................................... 3 -8 Statistical Methods 67 C) ................................................... 2.8.1 Analysis of variance 68
............................................................. 3.8.2 Correlation 68
CHAPTER 4 OPERATIONAL PERFORMANCE OF SEQUENCING
BATCH REACTORS .............................................................. 69
4.1 General Description ..................................................................... -70
4.2 Sludge Volume Index and Effluent Suspended Solids ............................. -77
................................ 4.3 Effluent Chemical Oxygen Demand ... ............... -80
..................................................... 4.4 Calculated Sludge Retention Time 81 ............................................................................ 4.5 Microbiology - 3 2
................................................................................... 4.6 Summary 84
CHAPTER 5 EXTRACELLULAR POLYhIERIC SUBSTANCES .................... ... 85
5 . I Influence of Sludge Retention Time on the Production and
Composition of Extracellular Polymeric Substances ............................... ..85
5.2 Molecular Weight Distribution of Extracellular Polymeric Substances .......... -90 5.3 Proposed Mechanism and Sources of the Production and Composition of
.................................................... Extracellular Polymeric Substances 92
5.4 Influence of Extracellular Polymeric Substances on Flocculating Ability . . . ..................................................................... and Compress~b~lity -95
5.5 Reiationship Behveen Sludge Volume Index and Effluent Suspended Solids ... 99
................................................................................. 5.6 Summary 100
........................ CHAPTER 6 HYDROPHOBICITY AND SURFACE CHARGE 102
6.1 InHuence of Sludge Retention Time on Hydrophobicity and Surface Charge ... 102
6.2 Relationship Between Hydrophobicity and Surface Charge ....................... 107
6.3 Influence of Extracellular Polymeric Substances on Hydrophobicity and
.......................................................................... Surface Charge -109
6.4 Influence of Hydrophobicity and Surface Charge on Flocculating Ability and . . . ......................................................................... Compressibility -1 12
........................................................ ...................... 6.5 Summary .. -1 17
CHAPTER 7 INTERPARTICLE INTERACTIONS AND COLLOIDAL ....................................................................... STABILITY -119
................................................................................... 7.1 Results -120
7-1.1 Effect of the nunber of sequential washings on . . ...................................................... accumulative nirbidity -120
..................................................... 7.1 -2 Effect of pH on stabili t).. 121
................. 7.1.3 Effects of ionic strength and cation valency on stability 122
7.1.4 Effect of EDTA concentration on stability ................................. 124
.................................... 7.1.5 Effect of urea concentration on stability 125
................................................................................. 7.1 Discussion 126
..................................................... 7.2.1 Electrostatic interactions 126 . . .............................................................. 7.2.2 Ionic interactions 129
.............................................................. 7.2.3 Hydrogen bonds -131
7.2.4 Influence of sludge retention time on stability and
...................................................... interparticle interactions 132
............................................... 7.2.5 Proposed floc structure mode1 134
................................................................................... 7.3 Summary 137
........................... CHAPTER 8 CONCLUSIONS AND FUCCOMMENDATIONS 139
.............................................................................. 8.1 Conclusions -141
........................................................................ 8.2 Recornrnendations 142
............................................ CHAPTER 9 ENGINEERING SIGNIFICANCE 145
vii
REFERENCES ......................................................................................... -149
APPENDICES
. A Mixed Liquor Suspended Solids Data ............................................................ 163
. B S ludge Volume Index Data .......................................................................... 166
. C Effluent Suspended Solids Data ..................................................................... 169
D . Effluent Chernical Oxygen Demand Daîa ......................................................... 172
E- 1 . Extracellular Polymeric Substances Data ......................................................... 173
E.2.Molecular Weight Distribution of EPS Compownts ............................................ 179
F- 1 .Contact Angle Data ................................................................................... 183
F-2 Contact Angle and Extracellular Polymenc Substances ........................................ -185
F-3 Correlations between Contact Angle and Surface Charge, Sludge Volume index,
and Effluent Suspended Solids .................................................................... -186
. G- 1 S d a c e Charge Data ................................................................................ -187
G-2 Surface Charge and Extracellular Polymeric ~tibstances ....................................... 191
. G-3 .Surface Charge Sludge Volume Index and Effluent Suspended Solids .................... -193
. . . H Dissociation Constant Data ........................................................................ -195
Table Page
Causes and categories-of biosolids-liquid separation problems .............................. 2 ............................................ Chernicd composition of the synthetic wastewater 43
...................................... Typical physical . chemical properties of the inoculurn 46 Chernical composition of extracellular polymenc substances of the inoculum ............ 46
...................................................... Operating conditions of the SBR system -52 General operating characteristics of the SBR system at stable operation .................. 69 Pearson's coefficients of linear correlation between the SV1 or ESS and the
.................................................................... quantity of EPS components -96 Pearson's coefficients of linear correlation between hydrophobicity or surface charge . . ............................................................................. and EPS composition 1 IO
Figure
LIST OF FIGURES
Page
Approaches to explaining and controlling biosolids-liquid separation problems 3 (modified fiom Forster. 1996) ................................................................... ..J
Effects of physical, chemical and biological conditions on sludge floc formation and properties (rnodified fiom Atkinson and Baoud. 1976) ..................... -9 Schematic diagram of an activated sludge floc (modified fiom Urbain et al., 1993) ..... 10 Process flow diagram in the operating sequence of the SBR ................................ .44 Pararnetric classifications for mixed liquor characterization ................... .. ............ 47 (a) Laboratory-scale sequencing batch reactor set-up; (b) Schematic diagram of the Iaboratory-scale SBR system ...................................................................... 49 Schematic diagram of a laboratory-scale SBR ................................. - ............. - 3 1 Schematic diagram of the ultrafiltration set-up for molecular weight distribution . . Determrnation ....................................................................................... 61 Schematic diagram of an ADSA-CD set-up (modified fiom Neumann et al., 1996) ..... 62 Effect of biomass concentration with measuring time on contact angle measurernents.64 Variation in the surface charge density of mixed liquor with respect to biomass concentration ....................................................................................... -65 A plot of the accumulative turbidity versus the number of sequential washings ......... -67 Mixed liquor suspended solids versus experimental time .............................. ... .. -72 Sludge volume index versus experimentai time ................................................ 74 Effluent suspended solids versus experimental time ........................................... 76 Microscopie pictures of (a) non-settleable fine flocs and (b) settleable large flocs ....... 79 Changes in emuent COD with respect to reaction tirne in one cycle (data collected on May 1 O. 1998) ..................................................................... -80 Calculated SRT values with respect to time ..................................................... 81 Principal component analysis of al1 BiologTM data fiom the SBR system .................. 82 (a) Effect of SRT on the production of EPS components under stable operating conditions; (b) Effect of SRT on the total EPS content and ratio of proteins to carbohydrates under stable operating conditions ..................................... .... ...- -89 Molecular weight distributions of EPS components ........................................... 91 Effect of SRT on the total carbohydrate content of whole sludge .......................... -93 Correlation between DNA and proteins in the EPS .......................................... -94 Relationship between total EPS content and sludge volume index ........................ -98 Relationship between the SV1 and ESS ........................................................ 100 Effect of SRT on contact angle values of sludge ............................................. 103 Influence of pH on the surface charge of sludge at different SRTs ........................ 105 Effect of SRT on the surface charge of sludge ................................................ 106 (a) Relationship between surface charge and water contact angle; (b) Relationship between zeta potential and water contact angle (data derived fiom Pere et al., 1993) .. 108 Relationship between water contact angles and effluent suspended solids ............... 113 ReIationship between the surface charge and effluent suspended solids .................. 113 Relationship between water contact angles and sludge volume index ..................... 116 Relationship behveen the surface charge and sludge volume index ....................... 116
Effect of the number of sequential washings on accumulative turbidity under different water chemistry conditions .................................................. -121 Effect of pH on colloidal stability .............................................................. 122 (a) Effect of ionic strength (KCl) on colloidal stability; (b) Effect of ionic strength (CaC12) on stability ..................................................................... -123 Effect of cation valency under different ionic strength conditions on . . stability ............................................................................................. 124 Effect of EDTA concentration on stabili ty.. .................................................. 125 Effect of urea concentration on stability ........................................................ 126 Effect of SRT on stability ...................................................................... -134 A schematic mode1 of a two-layer sludge floc ................................................ 136 A conceptual model of surface interactions that maintain the stability
of the outer layer of sludge flocs ................................................................ 137 Changes in sludge properties with respect to the SRT ........................................ 140
NOMENCLATURE
ADSA-CD
BOD5
CAM
CSTR
COD
DLVO
DNA
DO
DS
E
EDTA
EPS
ESS
FCP
F/M
HRT
1
K d
MLSS
P K ~
PVSK
Q QE
Qw
so SBR
SRT
SV1
v
Axisymmetric drop shape anaiysis-contact diameter.
Biochemical oxygen demand in 5 days [mg/L]
Contact angle measurement [degreesj
Continuous stirred tank reactor
Chernical oxygen demand [mg/L]
De jaguin and Landau (1 94 1) and Verwey and Overbeek (1 948)
Deoxyribonucleic acid
Dissolved oxygen Ipprn]
Dried biornass
COD removal efficiency
Ethy fenediaminetetraacetate
Extracellular polyrneric substances
Effluent suspended solids [mg/L]
Folin-Ciocalteau's phenol
Food (kg COD) /Microorganism (kg DS. Day)
Hydraulic retention time pour]
ionic strength [mol/L]
Decay coefficient [day-']
Mixed liquor suspended solids [m&]
Ionization constant
Polyvinyl sulfate potassium IN]
Influent wastewater flow rate &/min]
The voIume of discharged effluent CL] The arnount of mixed liquor wasted per day [L/day]
Influent COD concentration [mg/L]
Sequencing batch reactor
Sludge retention time [day]
Sludge volume index [mL/g]
Aeration tank volume IL]
xii
VO
VSS
X
XE
Yy icld
The volume of each SBR IL] Volatile suspended solids [mg/L]
Biomass concentration in the aeration tank prior to settling [mg/L]
Total suspended solids concentration in treated effluent [mg/L]
Yield coefficient of biomass [mg VSS/mg COD]
AGnoccuieion lnterfacial free energy of bioflocculation [ergs/cm2]
A Absorbante of the aqueous phase at 400nrn before mixing with hydrocarbons
a Absorbance of the aqueous phase at 400nm afier phase sepmation
YB Surface tension of sludge surfaces [ergs/cm2]
YL Surface tension of the treated effluent [ergs/cm2]
YBL Interfacial tension between sludge surfaces and the treated effluent [ergs/cm2]
CHAPTER I
INTRODUCTION
1.1 Biosolids-liquid Separation Problems
Biosolids-liquid separation by gravity settling is one of the most important unit operations
in the activated siudge process. The effective separation of biomass from treated effluent is crucial
for the quality of the receiving water. The loss of biomass in treated effluent causes oxygen
depletion in the aquatic system and also transports heavy metals and potential pathogens into the
receiving water. National surveys indicate that 70Y0 to 90% of activated sludge systems have
experienced various biomass separation problems in the settling tank (Wagner, 1982; Jenkins et al.,
1993: Blackbeard et al., 1986; Seviour et al., 1990: Blackball et al.. 1991; Pujoi et al., 1991 and
1992: Wanner et al., 1998). It is therefore not surprising that the control of flocculating ability,
settleability. and compressibility of sludge flocs has k e n one of the most important tasks of the
Activated Sludge Population Dynamics Specialist Group of the International Association on Water
Quality (IAWQ) (Wanner, 1994a).
Biosolids-liquid separation problems in the secondary clarifier have received great attention
since the early stages of the activated sludge process (Morgan and Beck, 1928; Townsend, 1932).
However. separation problems, such as bulking and foaming, have plagued this process for almost
a century (Forster, 1996). Currently, empirical methods and experïence are still the major tools for
the controi of these separation problems. They have been classified according to cause and effect
into four major categones by Jenkins et al. (1993) and Wanner (1994a and b), as shown in Table
1.1. Two types of biosoiids-liquid separation problems are thought to be associated with the
structure and surface properties of sludge flocs, while the other two types are related to the excess
growth of filarnentous microorganisms in activated sludge systems.
Table 1.1 Causes and Categories of Biosolids-liquid Separation Problems
(modified fiom Jenkins el ai., 1993)
Foaming
- --
Related to Floc Structure?
Not well u n d e r ~ t d ! Sufiace properties are thought to be crucial.
Not well understood! Surface properties may play an important role.
Filaments extend from the surfaces of flocs and therefore prevent close contacts of flocs. Loose structure.
Ca tegories of Problems
D ispersed growth
Pin-point floc
Fi lamentous bu lking
1.2 A Rational Approach to Improving Biosolids-liquid Separation
r - - - -
Causes and Effets o f Problems
Single cells or fine clumps of cells in dispersed state. No settling and high turbidity.
Small, compact microflocs. Weak structure of the macroflocs. Smaller flocs settle slowly.
Overabundance of filamentous microorganisms. Large buoyancy hinders settling.
Presence of non-bidegradable surfactants andor m icroorganisms producing surfactants.
In the pas, considerable efforts have been made to understand filamentous bulking and
foaming problems. The phenornena of filamentous buking and foaming are basically well
rinderstood at present, and strategies for addressing hem are usually available. However, bulking
and foaming problems caused by sorne special filarnentous microorganisms still challenge the
present control strategies (Jenkins et al., 1993; Wanner, 1994a and 1994b; Forster, 1996). In
contrast, the structure, surface chemistry, mechanisms of sludge floc formation, and disintegration
of sludge fiocs (e.g., dispersed growth and pin-point floc formation) have not been investigated to
The presence of both hydrophobic m icroorganisms and surfactants.
the same extent (Keiding et al.. 1993; Wanner, 1994a; Liss et al., 1996). Factors causing the
changes in surface chemistry of sludge flocs, such as extracellular polymeric substances (EPS),
hydrophobicity and surface charge, are still not well understood (Figure 1.1). IdentiQing the
changes which bring about differences in surface chemistry would be of great significance for
controlling biosolids-liquid separation,
Figure 1 - 1 Approaches to explaining and controlling biosolids-liquid separation problems
(Modified fiom Forster, 1996)
In the activated sludge process, physical characteristics of sludge flocs, such as size and
density, are important in determining the performance of biosolids-liquid separation. This is
because the efficiency of separation is related to the floc size and density. Large and dense flocs
with good water permeability are always desirable in gravity settling. The floc structure depends on
cell-ce11 interactions. For instance, the porosity of sludge flocs is related to the interparticle
4
interactions arising fiom sludge surfaces. The interactions are prirnarily govemed by the surface
chemistry of sIudge flocs, in particuiar EPS, a thin layer of biopolymers attached to sludge
surfaces. Consequently, more accurate fiindamental Iliformation about the molecular architecture of
sludge surfaces, the mechanisms of sludge floc formation, the forces that keep flocs together, and
the ways to manipulate the surface properties is essential for a better understanding of the
bio flocculation processes and the optimal design and operation of the activated sludge process.
1.3 Motivation of the Present Study
Although the importance of surface properties of sludge flocs in controlling gravity settling
has been recognized for some time, current means for controlling the surface properties of sludge
Rocs for effective biosolids-liquid separation is the addition of polyelectrolytes (inorganic and
polymeric flocculants). But polyelectrolytes are expensive and can potentially cause a second hand
contamination of treated effluent. One attractive option would be the control of the physiological
status of sludge flocs to manipulate the surface properties for effective floc separation.
Unfortunately, there is relatively Iittle known about the molecular control of natural aggregation.
Finding microbiologicaVphysiological-based approaches for manipulating the surface properties of
sludge flocs is therefore of great significance.
The physiology of sludge flocs is affected by a number of factors, such as the operating and
environmental conditions and the reactor configuration. Among them, sludge retention time (SRT)
(Le., the average residence time of sludge in a bioreactor) is one of the most important parameters
in controlling the microbial comrnunity and its physiology. Practically, the control of activated
sludge systems is achieved through the adjustment of SRT or organic loading intensity (F (food)
5
M (microorganism) ratio). It is therefore not surprishg that considerable efforts have been made to
investigate the effect of SRT or F/M on the separation efficiency of sludge flocs. However,
previous studies have mainly concentrated on macroscopic properties, such as effluent suspended
solids (ESS) (flocculating ability), zone settling velocity (ZSV) (settleability), and sludge volume
index (SVI) (compressibility) (Bisogni and Lawrence, 197 1 ; Chao and Keinath, 1979; Lovett et al.,
1983; Barahona and Eckenfelder, 1984; Andreadakis, 1993). Al1 these studies suffer from a lack of
detailed and reliable information about the microscopic properties of sludge fiocs. and about how
the fundamental information will correlate to the macroscopic characteristics: ESS, ZSV and Sm.
In other words, what causes the changes in the ESS, ZSV and SV1 under different environmental
and operating conditions is not well understood.
Previous studies (Forster, 1985; Urbain et al., 1993; Nielsen et al., 1996; Higgins and
Novak, 1997) on the role of surface properties in controlling biosolids-liquid separation usually
used sludge samptes from full-scale activated sludge systems. A drawback these studies suffered
was the poorly controlled conditions associated with full-scale activated sludge systems and a lack
of detailed information on the role of specific molecules of EPS and particular surface properties in
biosolids-liquid separation. The true correlations between surface properties and the performance
of biosolids-liquid separation by gravity settling might be masked in these studies.
Based on a cornprehensive literature survey (Chapter II), two basic hypotheses were tested
under weli-controlled laboratory-scale studies in this thesis: 1) surface properties and interparticle
interactions of sludge flocs can be manipulated by the SRT; and 2) flocculating ability and
compressibility of sludge flocs are controlled by particular surface properties.
1.4 Research Goal and Objectives
The primary goal of this study was to develop an improved understanding of the effect of
sludge retention tirne (SRT) on the rnicroscopic properties of sludge flocs in activated sludge
systems. It is expected that this research would improve the understanding of sludge floc formation
and the optimal design and operation of biosolids-liquid separation in the activated sludge pmcess.
Of particular interest were the surface properties (EPS, hydrophobicity, and surface charge) and
interparticle interactions (electrostatic, ionic interactions, and hydrogen bonds) of sludge flocs
mder different physiological status, and their dependent variables: bioflocculation, compressibility,
and stability. Specific objectives were:
1) to sîudy the effect of sludge retention time (SRT) on the production, composition and
molecula. weights of extracellular polymeric substances (EPS).
2) to investigate the influence of SRT on the hydrophobicity and surface charge of sludge flocs.
3) to evaluate the physicochemical nature of interparticle interactions that maintain the stability
of sludge flocs and its relationship to SRT.
4) to understand the relationships between surface properties and bioflocculation,
cornpressibility as well as the stability of sludge flocs.
1 -5 Thesis Outline
The motivation, pnmary goal and specific objectives of this research are stated in Chapter 1.
Chapter 11 sumrnarizes what we know about the properties of sludge flocs through a comprehensive
Iiterature survey on the mechanisms, interparticle interactions involved in sludge floc formation,
7
and process-related factors that affect floc formation and properties. Experimental materials and
methods are described in detail in Chapter III. The resuits and discussion are presented in Chapters
IV. V. VI and VIL The general performance of the sequencing batch reactors (SBRs) is described
in Chapter IV, including the flocculating ability, compressibility and microbio~ogy. The
production. composition and molecular weights of EPS at different SRTs and the role of EPS in
bioflocculation and compaction are presented in Chapter V. Chapter VI deais with the
hydrophobicity and surface charge of sludge flocs at different SRTs and their role in
bioflocculation and compaction. The interparticle interactions and the stability of sludge flocs at
different SRTs are presented in Chapter VII. The general conclusions fiom this snidy and
recommendations for future studies are surnmarized in Chapter VIII. Finally, Chapter IX discusses
the engineering significance of this study in tenns of consequences for the design and operation of
biosolids-liquid separation systems.
LITERATURE SURVEY
The Iiterature describes a number of physical, chernical and microbiological factors that
govem the flocculating ability, settleability, and compressibility of sludge flocs. The properties
of sludge flocs are the result of combined interactions of physicochemical environrnents and
microbiological conditions in the activated sludge system (Figure 2.1). Due to the complexity of
the relationships, it is difficult to precisely predict the behaviour and properties of sludge flocs.
Nevertheless. studies on these factors in both laboratory-scale and Ml-scale activated sludge
systems have greatly improved our understanding of the formation and properties of sludge flocs
and the efficiency of biosolids-liquid separation-
The purpose of this literature survey was to review what is known about the mechanisms
and interfacial forces invotved in sludge floc formation, and what is known about the coupling
between the microbiological, process conditions and the formation, development and properties
of sludge flocs.
Figure 2.1 Effects of physicôl, chernical and biological conditions on sludge
floc formation and properties (modified from Atkinson and Baoud, 1976)
2.1 Activated Sludge Flocs
Sludge flocs or activated sludge flocs are usually considered to be particulate aggregates
consisting of microorganisms, extracellular polymeric substances (EPS), organic and/or
inorganic colloidal particles (Unz, 1987; Li, 1992; Urbain et al., 1993; Sanin and Vesilind, 1994
and I996), as illustrated in Figure 2.2. They are formed through a dynamic process involvuig
EPS from either ce11 lysis, production or adsorption. The integrity and mechanical stability of
sludge flocs are maintained by either EPS, salt bridges, fiIamentous backbones (Pavoni et al.,
1972; Sezgin et al., 1978; Forster, 1985), or a combination of al1 three.
0 PO,"- C O 0
d OH - Hydvophobic area > N+
Inorganic particles
-V- Extracellular polymers
d+~ iva l en t cations
Figure 2.2 Schematic diagram of an activated sludge floc (modified fiom Urbain et al., 1993)
Phvsical Pro~erties: Sludge flocs are usuaily not spherical but possess highly irregular shapes
characterized by a large range of partide sizes, from single bacterial dimensions (ca. 1 to 3 pm)
to large aggregate sizes of more than I O00 prn (Li, 1992; Jorand et al., 1995; Droppo et al.,
1998). A bimodal floc size distribution is ofien found (Parker et al., 1970; Jorand er al., 1999, in
which single cells form microflocs, which then form macroflocs. Sludge flocs have an extremely
loose structure owing to the loose contact of cells. The high porosity of sludge flocs as observed
by microscopes (Liss et al., 1996) and as demonstrated by the measurement of flow through flocs
(Li and Ganczarczyk, 1988) is also responsible for the low strength of colloidal stability. The
apparent density of sludge flocs (about 1.03 g/cm3) (Li, 1992; Lee, 1994) is close to the density
of water. Consequently, the smali difference in densities between flocs and bulk water leads to a
low settling velocity in the secondary clarifier by gravity sedimentation.
11
Chetnical Pmperties: The mass of sludge flocs consists of about 60 to 90% cellular organic
material (Jenkins ef al., 1984) on a dry basis and large amounts of water (90 to 98% of the mas),
due to the hydrophilic nature of EPS and most aerobic bacteria (Dugan. 1987; Schmitt and
Fleming, 1999). Carbohydrates and proteins are the dominant components of sludge flocs. DNA,
RNA and lipids are also components of sludge flocs, but they are present in lower
concentrations. Carbohydrates and proteins account for 540% and 20-50% cf the volatile
suspended solids (VSS), respectively (Forster, 1971; Barber and Veenstra, 1986 Andreadakis,
1993; Frdund CC al., 1994). Inorganic salts or particies are another important part of sludge flocs,
constituting i O to 30% of the dry mass of sludge (Barber and Veenstra, 1986).
EIcrraceIhlar Polvmeric Substances: The third most cornmon component of studge flocs is EPS,
following microorganisms and water (Li and Ganczarçzyk, 1990). The amount of EPS strongly
depends on the loading of both organic substrates and biornass concentration, the influent
wastewater composition, and the operating conditions of the activated sludge system (Eriksson
and Alm, 199 1 ; Nielsen et al., 1996). EPS have long been associated with the formation and
development of sludge flocs as well as a number of properties of sludge flocs, such as water
binding capacity, settleability and dewaterability (Pavoni el al., 1972; Pere ef al., 1993; Urbain et
al., 1993). The current state of knowledge of EPS is described below.
EPS Constittrents EPS constituents of sludge surfaces have complicated composition. A
number of studies have found that EPS contain different types of biopolymers, including
proteins, humic acids, carbohydrates, acidic polysaccharides, DNA, RNA, and lipids (Pavoni ef
al., 1972; Kiff. 1978; Sakka and Takahashi, 1982; Goodwin and Forster, 1985; Urbain, et al.,
1993; Frdund et al., 1996; Palmgren and Nielsen, 1998). Proteins and carbohydrates are
12
believed to be the dominant components (Forster and Dallas-Newton, 1980; Forster, 1 985;
Eriksson and Alm, 1991 ; Urbain et al., 1993; Fralund et al., 1996) in the EPS. A large arnount of
humic acids was usually found in municipal and industrial sludge samples (Eriksson and Aim,
199 1 ; Urbain et al., 1993 Fralund et ai., 1996). DNA, RNA and Lipids comprise a relatively
smali proportion of EPS (Pavoni et al., 1972; Brown and Lester, 1980; Sakka and Talcahashi,
1982; Palmgren and Nielsen, 1998).
Further separation of carbohydrates in the EPS using gel chromatography indicates that
many different monosaccharides, such as rhamnose, fùcose, ribose, arabinose, xylose, mannose,
galactose and glucose, were present in EPS carbohydrates (Forster, 1971 and 1976; Horan and
Eccles. 1 986; Dignac et al., 1 998). Chromatographie separation of proteins demonstrates that
both hydrophobic amino acids (including alanine, leucine, glycine: valine, proline, isoleucine and
phenylalanine) and hydrophilic amino acids (such as lysine, threonine, arginine, serine, tyrosine,
histidine and methionine) existed in proteins extracted from sludge surfaces (Higgins and Novak,
1998; Dignac et al., 1998).
Characîerization of EPS Components In addition to their complex chernical composition,
the physical structure of EPS is also complicated and variable results are present in the literature.
Ultrafiltration and gel chromatography are the most common techniques for the molecular weight
characterization of EPS components (Forster, 1976; Frdund et al., 1994; Higgins and Novak,
1 998). Ultrafiltration studies (Forster, 1976; Goodwin and Forster, 1989) indicate that molecular
weights of EPS components covered a broad range, from less than 500 daltons to larger than
1 00,000 daltons, and the majority of EPS components had molecular weights under 500 daltons.
In contrat, Fralund et al. (1994) and Jahn et al. (1997) found a molecular weight distribution
13
(MWD) fiom 10,000 daltons to 2,000,000 daltons in a senes of studies using size exclusion
chromatographic columns. More recently, by using chromatographic separations Higgins and
Novak (1998) found the molecular weight of proteins in the EPS was relatively u n i h at
15,000 daltons.
These results suggest that the MWD of EPS components is complicated and may depend
on the techniques for molecular weight characterization. A combination of different techniques
can give a more comprehensive description of the physical configuration of EPS components.
Sources and Classîjication of EPS There are three sources that contribute to EPS: 1) metabolic
synthesis, 2) ce11 lysis, and 3) adsorption of polIutants from wastewaters. The EPS content from
metabolic synthesis is strongly related to the environmental, operating and microbiological
conditions. The contribution of ce11 lysis is associated with the physiological conditions and EPS
extraction methods. The adsorption of pollutants fiom wastewaters is related to the composition
and properties of wastewaters. At present, it is dificult to distinguish the relative importance of
the contributions of the three mechanisms to EPS content.
Basically, EPS can be divided into two distinct categories: the slime layer and the capsule
Iayer (Tuntoolavest, 1980; Gehr and Henry, 1983). The slime iayer is unattached or loosely
attached to sludge surfaces, while the capsule layer is attached directly to the exterior of ce11
waIls. Most of the previous studies on EPS have focused on the capsule layer of EPS, while the
slime layer of EPS has not been studied extensively.
Hvdronhobicitv: Generall y, sludge flocs are naturall y hydrated, due to the presence of large
numbers of hydroxyl, carboxyl and phosphate groups. However, sludge surfaces are known to
14
possess hydrophobic areas (Magnusson, 1980; Urbain et al., 1993). For example, the
hydrophobic side chains in amino acids, the methyl groups in polysaccharides, and the long-
shain carbon groups in lipids al1 contribute to the hydrophobic properties of sludge flocs. The
coexistence of hydrophobic and hydrophilic microorganisms in sludge flocs has been
demonstrated by Singh et al. (1987) and Jorand et al. (1994). Foaming problems in activated
sludge systems are also associated with the presence of very hydrophobic cells (Jenkins et al.,
1993; Wanner, 1994b). It is believed that the hydrophobic-hydrophiiic balance plays an
important role in governing the behaviour of sludge flocs (Eriksson and Axberg, 198 1 ; Urbain et
al., 1993; Jorand et al., 1994 and 1998).
Surface Charae: Sludge flocs are negatively charged under neutral pH conditions. The presence
of ionizable groups such as carboxyl, phosphate and amino groups, in the EPS and ce11 surfaces
is responsible for the density of surface charge. The zeta potential of sludge flocs in the effluent
is usually in the range of -1 O to -30 mv (Forster, 1968 and 197 1 ; Valin and Sutherland, 1982).
The properties of sludge flocs descnbed above are central to the behaviour and efficiency
of the activated sludge process (such as flocculating ability, settleability and compressibility).
The key to effective floc separation is to design the flocs with the desired properties.
2.2 Bioflocculation, Settleability and Compressibility
Biosolids-liquid separation by gravity settling involves three different steps: 1) sludge
floc formation and growth; 2) settling; and 3) compaction.
15
Bioflocculation is a process involvhg the aggregation of single cells or fine flocs to form
larger flocs. In order to f o m larger sludge flocs, individual cells or fine flocs must first corne
together through collision and then adhere to each other. During this process, the size of flocs
increases with time, and the settling velocity of formed flocs increases as the result of
bioflocculation. Following this phase, sludge flocs settle as a zone, and the individual particles
remain in relatively the sarne position to each other. The zone settling velocity (ZSV) is
independent of floc size. At the bottom of the settlhg tank, the sludge density increases gradually
and the ZSV decreases simultaneously. The water is squeezed out in the settled layer.
Compaction of loose sludge flocs at the bottom of the settling tank is often the final stage in
gravity settling.
It appears that bioflocculation is linked to settling through an increase in floc size.
However, a quantitative evaluation of bioflocculation is the flocculating ability of sludge, which
is defined as the amount of non-settleable fuie flocs after a certain time of gravity settling. A
higher turbidity or effluent suspended solids (ESS) indicates a poorer flocculating ability. On the
other hand, compaction reflects the behavior of settleable larger flocs at the bottom of a settling
tank. Compaction is usually evaluated by gravity settling one liter of mixed liquor for 30 minutes
in a one liter graduated cylinder. The volume afier 30 minutes divided by the mixed liquor
suspended solids is called sludge volume index (SVI). A higher SV1 value is related to a poorer
compaction of siudge.
This study focused on a better understanding of fùndamental links between surface
properties and the flocculating ability as well as the compressibility of sludge. The degree of
16
bioflocculation was evaluated by the effluent suspended solids (J3S). The compressibility of
sludge was estimated by the SVI.
2. 3 Interfacial Forces Involved in Microbial Floc Formation
Before considering the specific mechanism of sludge floc formation in the activated
sludge process, it is usehl to consider the forces goveming the stable dispersion and aggregation
of sludge flocs (Matsuo et al., 1981; Wame and Bowden, 1987). In general, the forces acting on
and between particles can be grouped into two categories: external forces, which include
gravitational, hydrodynamic (drag) forces and thermal energy; and interfacial forces. There are
four types of primary non-covalent interfacial forces involved in sludge floc formation: van der
Waals forces; electrostatic double-layer forces; hydrophobic/hydrophilic forces; and steric forces.
Ionic bonds are also involved in controlling sludge floc formation and its stability. A detailed
description of al1 forces involved in sludge floc formation is beyond the scope of this study. Only
the interfacial forces are summarized here.
Van der Waals Forces: Van der Waals forces are always present between the interacting
bacteria. They are attractive forces and depend on the geometry and nature of the interacting
bacteria. Typically energies of van der Waals interactions Vary with the inverse sixth power o f
the particle sepmation. They contain three ternis: permanent dipole-permanent dipole
interactions; permanent dipole-induced dipole interactions, and induced dipole-induced dipole
interactions. A number of methods have been proposed for evaluating the values of van der
Waals forces between the interacting bacteria (Valin and Sutherland, 1982; Gregory, 1993).
17
Ekctrostatic Double-laver Forces: Charged bacteria in an aqueous medium attract oppositely
charged ions fiom the solution leading to the establishment of an electrostatic double-layer at the
solid-liquid interface. Electrostatic interactions arke as the electrical double-layer of two
approaching bactena overlaps. They are usually repulsive forces and are responsible for the
stable dispersion of charged particles. Under neutrd pH conditions, bacterial surfaces are
negatively charged due to the dissociation of anionic groups associated with the bacterial surface.
Therefore. bacteria have a tendency to be dispersed due to double-layer interactions of charged
bacterial surfaces. The electrostatic double-layer forces depend on the geometry and electrical
behavior of the charged bacteria and the solution properties (Hogg, 1989; Gregory, 1993).
HvdroDhobic/Hvdro~hific Forces: The hydrophobic forces between bacterial surfaces depend in
large part on the unique properties of the thin layer of water associated with the surfaces and
forces at a short distance (< 2 nm). They are attractive forces, while their counterparts are usually
referred to as hydrophilic forces, which are repulsive. The hydrophobic/hydrophilic forces are of
a polar nature and may be two orders of magnitude higher than the van der Waals and
electrostatic forces (van Oss, 1994) at short distances (<2nm). Therefore, negIecting the
hydrophobic/hydrophilic interaction at the surfaces of bacteria, as treated in classic DLVO theory
(De jaguin and Landau, 1941 ; Verwey and Overbeek, 1948), is probably not reasonable in
understanding sludge floc formation and floc strength. At present, there are several methods
available for estimating hydrophobic forces (van Oss, 1994). Typically, energies of hydrophobic
interaction decrease exponentiaily with the particle separation distance.
S~eric Forces: This type of force anses between polyme~ coated surfaces; it is believed to be
relevant in biological systems. In the sludge floc matrix, sludge surfaces are covered by a layer of
18
macromolecules called EPS. The presence of EPS on sludge surfaces may prevent the approach
of sludge flocs, allowing only a loose contact. The situation is even more complicated for
charged polymer-coated surfaces. Generally, it is dificult to quanti6 steric forces in biological
systems on account of their complexity (Oliveira, 1992). More recently, Jucker et al. (1999) have
proposed an approach for evaiuating the magnitude of stenc forces arising fiom polymer
interactions, but it is still in its infancy.
In order to form sludge flocs, the aggregation process of sludge rnicroorganisms must
overcome barrien f?om the repulsive forces (repulsive electrostatic, hydrophilic and steric
forces).
2.4 Mechanisms of Sludge Floc Formation
ï h e mechanisms of sludge floc formation have received great attention since the 1950s
(McKimey, 1956; Parvoni et al., 1972; Campbell, 1972; Eriksson and Alm, 1991) due to their
importance in biological wastewater treatrnent. Although several mechanisms have been
proposed for describing the phenornena of sludge floc formation, a considerable debate about the
causes persists, and the exact reasons for biofloccutation have not yet been well elucidated.
The proposed mechanisms for sludge floc formation can be classified into £ive types: 1)
charge neutralization; 2) hydrophobic interaction; 3) polyrner bridging; 4) salt bridging; and 5)
the surface thermodynarnic approach. The polymer bridging mechanism has received the greatest
attention; it involves the entanglement and adsorption of microorganisms by the EPS. Al1 these
rnodels emphasize the importance of surface properties in siudge floc interactions. The
19
differences lie in which of the surface properties is considered to be most important, and how a
particular parameter is affected by nutritional and environmental conditions.
Charge Neutralization: Based on size (yeast: 10-13 Pm; bacteria: 1-3 Pm; ce11 debris: 0.2 -0.5
p m diameters) and charge properties, microorganisms are often considered as biocolloids.
Consequentl y, the aggregation of microorganisms has been described in the literature (Marshall,
1992) mostly in terms of DLVO theory, which was originaliy developed to explain the stability
of hydrophobic inorganic colloid suspensions. According to this theory, the sludge interaction
potential is the sum of van der Waals (attractive) and electrostatic (repulsive) interaction
potentials. The stability of sludge flocs in an aqueous medium depends on the results of
cornpetition between van der Waals and electrostatic forces at a given floc separation distance.
Addition of oppositely charged polyelectrolytes results in a decrease in the surface charge
density. Le., the repulsive forces between the interacting particles are reduced. Therefore, the
particles can get sufficiently close to each other at a distance in which the van der Waals forces
are effective, and then flocculate.
However, the validity of DLVO theory to explain microbial aggregation has also been
questioned, when both anionic and neutral polyelectrolytes were found to facilitate microbial
aggregation ( U n , 1987). For some microorganisms, no fiocculation was observed even at the
iso-electric point (Temey and Stumm. 1965). The failure of DLVO theory to explain microbial
aggregation under certain conditions may be due to the complexity of biological systems. The
microbial ce11 surface is not smooth, and hydrophobic/hydrophilic groups and EPS are also found
at the ce11 surface (hydrophobic and steric forces) (Warne and Bowden, 1987). Since DLVO
theory assumes smooth sudaces on the interacting particles and considers only the van der Waals
20
and electrostatic forces, a consideration of other physical forces, such as hydrophobic
interactions, may extend its application.
W o ~ h o b i c hteractiom: The presence of a shell layer of bound water near the ce11 surface
accounts for hydrophobichydrophilic interactions. When two similar surfaces approach each
other at a short distance, the bound water layers surrounding the surfaces will overlap, fonning a
displacement of the bound water layer into the bulk water. This leads to a decrease in the k e
ener,oy (Le., attraction) in the case of a hydrophobic surface, but to an increase in the fkee energy
(Le., repulsion) in the case of a hydrophilic surface (Absolom et al., 1983). Therefore,
hydrophobic cells favor aggregate formation, while hydrophilic cells are found in a dispersed
state.
According to this theory, any alteration of the ce11 surface which increases the surface
hydrophobicity will enhance the formation of flocs. However, only in recent years, has there
been a realization that hydrophobic interactions play an important role in sludge floc formation
(Singh and Vincent, 1987; Urbain et al., 1993; Jorand et al., t 994 and 1998).
Valin and Sutherland (1982) used the contact angle measurements to correlate the
flocculation performance of sludge to its hydrophobic characteristics. Singh and Vincent (1987)
observed a positive correlation between the hydrophobicity of sludge flocs, or bacteria isolated
from sludge, and their flocculation performance. Overmarm and Pfernig (1992) noted that
floccuiation of Amoebobacrer purpureus was linked to increased hydrophobicity of ce11 surfaces,
probably via a mechanism mediated by a surface protein. Furthermore, Urbain et of. (1993)
found that interna1 hydrophobic bonding was involved in the compressibility of sludge, and
21
Jorand et al. (1994) suggested that sludge floc formation was controlled by the ratio between
hydrophilic EPS, in which bactena were embedded, and hydrophobic interactions. and
demonstrated the coexistence of hydrophobic and hydrophilic bacterial strains in sludge flocs.
Al1 these results lead to the conclusion that the role of hydrophobic interactions in sludge
floc formation should receive more attention. Therefore, additional studies are desirable to gain a
better understanding of the function of hydrophobic interactions.
Polvmer Bridxins: Sludge floc formation usually involves high molecular weight polymers.
Cationic, anionic, and non-ionic polymers have al1 been found to be effective in enhancing
microbial aggregation (Warne and Bowden, 1987). The polymer bridging rnechanism has long
been used to interpret microbial aggregation by McKinney (1956), Busch et ai. (1968), Ries er
al. (1 968), Friedman et al. (1 968), Pavoni er al. (1 Wî), and Eriksson and Hardin (1 984).
On the one hand, Peter and Wuhrmarn (1970) used bacterial cells as dispersed solids and
found that polymer bridging was the predominant mechanism for floc formation, and humic
acids were proposed as flocculating agents. Alternatively, Deinerna and Zevenhuizen (1971)
showed that the EPS that played an important role in sludge floc formation were cellulose fibrils.
In contrast, Sakka and Takahashi (1982) found that the natural flocculation of dispersed cells
involved a bacterial strain which possesses a specific DNA binding activity on its ce11 surface.
The bacterial strain spontaneously aggregated in the stationary phase of growth following the
accumulation of' long chains of DNA in the medium. In this case, the released DNA acted as a
natural flocculant. At present, the reai rnolecular determinants for bioflocculation have not been
well understood.
22
Pavoni et al. (1972) argued that sludge flocculation was related to the change in EPS,
which include carbohydrates, proteins, DNA and RNA produced in the decay phase of biomass.
An increase in the EPS content in the decayphase led to a decrease in emuent turbidity. More
recently, Eriksson and Hardin (1984) proposed a unifying mode1 of bioflocculation. They
suggested that EPS are responsible for bridging the distance between electrostatically stabilïzed
cells to form a weak, elongated floc during the initiation of flocculation. Up to a certain level,
M e r polysaccharide synthesis produces stronger flocs by bridging cells more f d y . M e r
that, polysaccharides will have a dispersing effect due to steric forces.
Based on studies on the role of EPS in sludge floc formation, it is logical to conclude that
EPS are linked to sludge floc formation. Although it has been claimed that some chemical
components in EPS are responsible for the initiation of sludge floc formation, whether the
amount or the specific composition of EPS, or both, is crucial to sludge floc fonnation is still not
well understood. At present, the general hypothesis is that the greater the amount of EPS, the
better the bioflocculation.
Salr Brid~ing: It is known that both the bacterial surface and EPS provide negatively charged
adsorption sites. Moreover, the lack of structural strength of EPS raises the question of how EPS
molecules can bind cells into a floc. A bridging agent is probably needed to bind the bacterial
surface and EPS together to maintain floc stability. Inorganic ions, such as divalent cations (Ca2+
and MgL*), have been found to be strongly associated with the chemical structure of sludge flocs,
as shown by a number of studies (Forster, 1972 and 1985; Eriksson and A h , 1991 ; Bruus et al.,
1992; Urbain et al., 1993; Sanin and Vesilind, 1996).
23
Forster et al. (1972 and 1985) and Eriksson and A h (1991) studied the effect of
polyvalent cations on bioflocculation. They found that polyvalent metals were involved in the
floc matrix through the binding of metal ions with either EPS or the surface of cells. Addition of
EDTA, which removes polyvalent cations fiom the floc matrix by ionic chelating reactions, led
to the disintegration of sludge flocs and thus Sec ted their settling velocities (Eriksson and Alm,
199 1 ). Bmus et al. (1992) also demonstrated that the removal of Ca2+ fiom sludge flocs increased
the disintegration of flocs and found that some 50% of ca2' was associated with EPS. This
suggests that the stability of sludge flocs was aEected by ionic binding forces.
More recently, Sanin and Vesilind (1996) found that only limited adsorption of alginate, a
polysaccharide, was observed on polystyrene latex particles similar in size to bacteria when no
polyvalent metals were present in the solution. The adsorption of alginate significantly increased
with the addition of Ca2- ions. This suggests that the formation of polymer-metal complexes and
polymer gelation were important in bioflocculation.
It is clear that the presence of polyvalent metah is helpîùl in enhancing sludge floc
formation and maintaining the stability of sludge floc structure, but which anionic groups take
part in the polymer-metal reaction is still not clear.
Swfuce Thermodvnamics Approach: - In addition to the applications of interfacial interactions
and the DLVO theory to explain microbial floc formation, the initiation of microbial floc
formation can also be interpreted in ternis of system fiee energy. The basic concept of the
thermodynamic mode1 is that the system fiee energy is minimized at equilibrium. Consequently.
the process of microbial floc formation will be thermodynamically favored if the process itself
24
causes the system free energy to decrease. Assuming that the electrostatic interactions and other
specific bindings can be ignored, the fiee energy of the interaction between two identical
bacterial cells (B), irmnersed in liquid (L) can be described as follows:
where AGfl,,Ia,i, is the free energy of floc formation and y,, is the interfacial tension for the
bacteria-liquid interface. If the total free energy of a system is reduced (AG,,,,, < O) by ce11
interactions, then microbial floc formation will be thermodynamically favored (Neumann et al.,
1974a; Neumann et al., 1980; Absolom et al., 1983).
According to Neumann's equation-of-state, the interfaciai tension y,, is correlated to the
surface tension terms of bacterial cells y, and suspended solution y, as follows (Neumann et al.,
1974b and 1980):
yBL = (Jy , - J y j 2 / (1- 0.01 5 JyB JyJ.
Substituting Equation (2-2) into Equation (2-l), the fiee energy of floc formation cm be
calculated if the water contact angle on ce11 layers and water surface tension are known
(Neumann et al., 1980).
The surface fiee energy concept has not yet been used to describe the phenornena of
sludge floc formation. Although the correlation between the hydrophobicity of cells and
flocculation performance (Valin and Sutherland, 1982), and the influence of liquid surface
tension on the granulation and stability of anaerobic granular sludge (Thaveesri, 1995) were
25
reported, the vaiidity of surface thermodynamics to explain sludge floc formztion is not well
understood.
2.5 Process-related Factors Aff'ting the Formation and Properties of Sludge Flocs
The process environments are basically classified into intemal and externai environments
(Celleja, 1984). Interna1 environments deal with the cell itself, while extemal environments
include the physical, chernical and hydraulic conditions, including organic loading, dissolved
oxygen (DO) concentration, environmental temperature and pH, ionic strength, surface tension of
the solution, organic substrate chemistry and mechanical shear forces. A comprebensive
literature survey in this subject is beyond the scope of this study. Only the factors related t o this
thesis are reviewed and discussed below.
2.5.1 Microbiological factors
It has long been known that not al1 microbes aggregate. Even within a species, strains
have different aggregative capacities. Some strains are more aggregative. some less, whereas
others are not aggregative at d l (Caileja, 1984). In an activated sludge system, a complex
microbial ecology is present, which includes bacteria, protozoa, viruses and many other types of
organisms (Hanel, 1988). Arnong the various bacteria in sludge, some are floc-forming
microorganisms which are important in sludge floc formation. They include cellulose (or
ce1 lulose-li ke) producing bactena such as Pseudomonas, Aerobacter, Ag-obacrerium,
Azotobacter and Zooglea ramigera (Wanner, 1994b and 1995). Sludge flocs are formed when
such bacteria are ernbedded in the EPS rnatrix. Another type of floc-forming bacteria, 2.
ramigera 1 15 (Friedman and Dugan, 1968; Easson et al., 1987), forms sludge flocs by producing
26
capsular polysaccharides enclosing large packets of cells. In contrast, a certain amount of
filarnentous microorganisms are thought to be necessary for the formation of larger, denser flocs
(backbone theory) (Segzin et al., 1978). However, the overgrowth of filamentous
microorganisms is generally responsible for filarnentous bulking (Jenkins et al., 1993; Wanner,
1994b and 1998).
Among the physiological conditions, the physiological age of cells is one of the primary
factors related to the formation and properties of sludge flocs. Experimentai evidence indicates
that large amounts of EPS were accumulated in the endogenous phase, and sludge floc formation
depended on the physiological age of cells (McKinney, 1956; Parsons et al., 197 1 ; Pavoni et al..
1972). Sludge microorganisms in wastewater treatment plants did not flocculate during the log
growth phase. Flocculation started during the stationary phase and was optimal in the
endogenous phase (McKimey, 1956; Pavoni et al., 1972). Pavoni et al. (1972) suggested that
EPS excised during the endogenous phase were rnainly responsible for sludge floc formation and
development. However. results relating the EPS production to physiological status are very
contradictory. For instance, the results fiom continuous bioreactors (Kiff, 1978; Gulas et al.,
1979; Pere et al., 1993) indicate that EPS were also produced in rapid-growth situations, which is
not consistent with the findings of Pavoni et al. (1972).
2.5.2 Sludge retention time: Sludge retention time (SRT) and Food/Microorganisms (FM) ratio
are two of the key process variables that, in practice, can be adjusted to optimize the operation of
the activated sludge process. The SRT is defined as the average residence time of sludge within a
bioreactor, while the F/M is defined as the ratio of the feeding intensity of organic pollutants to
the microorganisms in the bioreactor (g. COD/g. biomasdday). These two parameters are not
independent, but at steady state are correlated through the following equations:
where Y,,,, is the yield coefficient of biomass; FIM is the organic loading intensity; E is the
COD removal efficiency; K, is the decay coefficient; S, is the influent COD concentration; Q is
the influent wastewater flow rate; V is the aeration tank volume, and X is the concentration of
biornass in the aeration tank. In general, a lower SRT is related to a higher F M .
Following the work of many others in this area (Bisogni and Lawrence, 1971 ; Pitman,
1975; Chao and Keinath, 1979; Gulas et al., 1979; Saunders and Dick, 198 1 ; Lovett er al., 1983;
Whalberg, 1992; Andreadakis, 1993), the SRT was chosen in preference to FA4 as the parameter
used to describe the operating state of an activated sludge system in this study. The following
paragraphs present a brief discussion of what we know about the knowledge of the influence of
SRT or F/M on sludge properties.
EKect of SR T on Microbial Communitv: In general, low SRTs are associated with rapid rates
of microbial growth and high rates of sludge production and wastes; high SRTs are related to
slow growth and low rates of sludge production. Based on kinetic selection principles, a change
in the microbial composition with respect to the SRT is expected. A high SRT will favor the
accumulation of slow-growing microorganisms, while a low SRT enhances the domination of
fast-growing microorganisms (Hanel, 1988). Unfortunately, little information is available about
28
changes in the microbial cornmunity structure in terms of the SRT. Most previous studies have
focused on the effect of SRT on the @OH& of filaments (Jenkins et al., 1993; Wanner, 1994b
and 1998). For instance, different types of filamentous microorganisms were observed at
di Rerent S RTs (Wanner, 1 995). Ford and Eckenfelder (1 967) and Sezgin et al. (1 978) also noted
that a higb organic loading (i.e., a low SRT) would promote the overgrowth of filamentous
microorganisms, and the overgrowth of these filamentous rnicroorganisms was responsible for
poor settleability of sludge flocs.
Further characterization of the microbial community structure is necessary to understand
the influence of SRT on the rnicrobial comrnunity.
Effect of S R T on Flocctrlatina Abiliîy, Settleabilitv and Com~ressibilitv: The most common
parameters used to quanti@ the flocculating ability. settleability and compressibility of sludge
are the effluent suspended solids (ESS), zone settling velocity (ZSV), and sludge volume index
(SVI), respectively.
Considerable efforts have k e n made to understand the flocculating ability and
settleability of sludge flocs at different SRTs. However, the information from previous studies
was contradictory, especially for settling properties. Both Englande and Eckenfelder (1973) and
Pitman (1975) observed that the SV1 decreased as the SRT increased. In contrast, Barahona and
Eckenfelder (1984) found that the zone settling velocity (m/s) of sludge increased as SRT
decreased. On the other hand, Bisogni and Lawrence (1971), Heddie (1977), Chao and Keinath
(1979) and Lovett et a1.(1983) observed that the SV1 decreased with an increase in the SRT,
following an initial increase in the SV1 with the SRT. It appears that it is difficult to establish a
29
generally valid rule to predict the settleability with respect to the SRT. This is not surpriskg, as
the SV1 is affected by a number of factors, including wastewater composition, and
environmentaï, operational and microbiological conditions.
Results about the flocculating ability of sludge flocs with respect to the SRT are highly
consistent with previous studies. The flocculating ability of sludge flocs generaily increases with
an increase in the SRT. Al1 previous studies Bisogni and Lawrence, 971; Pitman, 1975; Heddle,
1977; Chao and Keinath, 1979; Gulas et al., 1979; Lovett et al., 1983; Whalberg, 1992;
Andreadakis, 1993) observed that better bioflocculation of sludge flocs (less ESS) is associated
with a higher SRT. However, an extended aeration with an extremely high SRT may disintegrate
sludge flocs into pin-point flocs (Jenkins et al., 1984).
Effecr o f SRT on Mkrosco~ic Siudae Pro~erties: As discussed above, most studies about the
SRT have concentrated on the influence of SRT on macroscopic properties, such as the SVI and
ESS. These studies lacked detailed and reliable information about the fhdarnental structure and
properties of sludge flocs in terms of the SRT, so, what causes the change in the flocculating
ability and settleability of sludge flocs at different SRTs is still unknown. Only in recent years,
have a few studies addressed the fundamental structure and properties of sludge flocs at different
SRTs.
Floc s i x Two recent studies have Iooked at the effects of F/M on the particle size distribution of
sludge flocs. Li and Ganczarczyk (1993) performed an extensive study to evaiuate the influence
of operating conditions on the floc size distribution in full-scale activated sludge systems. They
found that the F/M was one of the most significant factors influencing the floc size distribution,
30
and a larger median floc size was associated with a higher FM. Barbusinski and Koscielniak
(1995) also demonstrated that FIM strongly infiuenced the floc size distribution, and m e r
indicated that long-term loading changes caused larger disturbance to the floc size distribution
than more rapid and shorter changes. Generally, an overload of organic substrates would lead to
the breakup of flocs, but sludge would also exhibit poor floccuIation at very low F/M conditions
(Eckenfelder and Musteman, 1995). A suitable FiM ratio is aiways desirable for effective
operation of each activated sludge system.
Andreadakis (1993) aiso found that the floc size distribution was related to the SRT. With
the exception of an SRT of 1.1 day, more than 85 % of flocs (SRTs fiom 4.2 to 17.4 days) were
in the range 10 to 70 Pm, with a median size between 35 and 45 Pm. The flocs at an SRT of 1.1
day were significantly smaller, with a median value of around 20 Pm.
EPS: Only a few studies have investigated the influence of SRT on the EPS production. Chao
and Keinath (1979) using a chemostat found that the carbohydrate content of EPS increased with
an increase in the SRT. In contrast, Kiff (1978) and Gulas (1979) observed that the total EPS
production was almost independent of the SRT, and even more EPS was extracted at very low
SRTs (1-4 days). A more recent study (Pere el al., 1993) with sludge fiom a pulp and paper mil1
effluent treatment plant aiso indicated that a larger arnount of EPS was recovered at lower SRTs.
However. all these studies focused on the total EPS content. Further characterization of both the
chernical composition and physical configuration of EPS were not part of these studies.
It has long been thought that a larger mount of EPS is responsible for better
bioflocculation, a conclusion based on results fiom batch reactors (Pavoni et al., 1972; Sheintuch
3 1
et al.. 1986; Sheintuch, 1987). Results obtained fiom continuous reactors (Kiff, 1978; Gdas et
al.. 1979; Pere et al., 1993), however, do not support the explanation of polymer bridging
mechanisms derived fiom batch studies (Pavoni et al., 1972; Sheintuch et al., 1986; Sheintuch,
1987). Therefore, the use of total EPS content as a parameter has limited value in understanding
bioflocculation. Further characterization of the chernical composition and physical codiguration
of EPS and other surface properties, such as hydrophobicity and surface charge, is essential for
providing a scientific explanation why the flocculating abiiity of sludge flocs changes with
respect to the SRT.
2.6 Measurement and Analytical Methods for Sludge Surface Characterization
The measurement and anaiytical methods used to study sludge surfaces are important for
understanding the effects of physical, chernical and rnicrobiological conditions on rnicroscopic
properties of sludge flocs, and for relating rnicroscopic properties of sludge flocs to their
floccdating ability and settleability. Unfortunately, at present, few standard methods with good
reproducibility are available, although various physical, chernical and microbiological
measurernent and analytical techniques have been developed and used in sludge floc research.
Arnong these, Li and Ganczarczyk (1986) give a comprehensive review of the methods and
techniques for the measurement of physical characteristics. Because of the importance of EPS
and surface properties, some basic methods for the extraction of EPS and the determination of
hydrophobicity and surface charge of siudge are surnmarized and discussed below.
Extraction o f Extracellular Polymeric Substances: Usually, EPS are associated with cells,
inorganic particles and organic colloids in the sludge floc matrix in wastewater treatment. A
32
nurnber of methods have been investigated and applied to separate EPS fiom sludge flocs. They
include physical (centrifugation, sonication and thermal extraction), chernical (hydroxide
addition, acidic stripping, ion exchange), and combined phy sicochemical methods.
Centrifùgal stripping for EPS recovery in sludge was once considered to be an important
technique in understanding the role of EPS in sludge floc formation. Pavoni et al. (1972)
assumed that at a force of 35,000 x g for 15 minutes, EPS could be quantitatively extracted fiom
sludge flocs, while different centrifûge speeds were used by difTerent r-ch groups. However,
since the early 1980s, a number of studies on centrifbgation have demonstrated that it is not an
effective method for the extraction of EPS fkom sludge flocs. Both Brown and Lester (1 980) and
Novak and Haugar (1981) found that even high-speed centrifugation does not effectively strip
EPS f?om sludge flocs. Sanin and Vesilind (1994) concluded that high-speed centrifugation c m
remove only some of the EPS. Based on these investigations, it has k e n suggested that the EPS
collected by centrifugation stripping may not corne fiom the surface of sludge flocs but fiom the
supernatant (Sanin and Vesilind, 1994).
Sonication is another physicai method for extracting EPS fiom sludge flocs. Brown and
Lester (1980) found that ultrasonication released low concentrations of EPS and was not
effective as an EPS extraction method in the bacterial cultures examined. since ultrasonication
caused no significant dismption of cellular surfaces. However, they suggested that
ultrasonication may be usefbl as a pretiminary treatment in conjunction with other extraction
methods. More recently, King and Forster (1 99 1) studied the effect of sonication on sludge flocs-
They found that sonication reieased a certain amount of soluble carbohydrates and proteins fiom
33
sIudge flocs and hypothesized that EPS components in the floc matrix may have different
strengths of attachment.
Chernical methods involve the addition of chemical agents into sludge samples to leach
EPS from the floc matrix. Many chemical methods have been tested or used to extract EPS fkom
different bacterial cultures, including ammonium hydroxide, sodium hydroxide, EDTA, and
sulphuric acid (Brown and Lester, 1980). Usually chemical rnethods yield much higher
concentrations of EPS than physical methods. However, due to harsh conditions, chemical
methods may cause a significant disruption of cells and hydrolyze polymenc molecules (Gehr
and Henry, 1983).
The most effective and practical methods for EPS extraction fiom sludge flocs are
combined physicochemical methods. Typically, methods for EPS extraction fiom sludge flocs
are composed of four or five distinct steps (Gehr and Henry, 1983), depending on the purpose: 1)
pretreatrnent of the sludge sarnple; 2) extraction of the capsular part of EPS from microbial cells;
3) precipitation and collection; and 4) purification.
A more recently developed method involves the use of an ion exchange resin (Dowex
resin) for EPS extraction fiom sludge flocs (Frdund et al., 1994 and 1996) at a low temperature
(4°C). This method yields a higher protein concentration than the widely used thermal extraction
- solvent precipitation method, and prevents the lysis of cells at high temperatures.
Currently, different research groups use different methods for EPS extraction. No standard
extraction method is availabie, but thermal extraction-solvent precipitation technique is the most
34
widely used. The ion exchange resin method was w d in this study, due to its advantages of high
extraction efficiency and less ce11 lysis during extraction.
Measurement of Hvdro~hobicihi: Most of the techniques for hy dropho bic evaluation can be
grouped into two categories. The first measures the hydrophobic properties of the outer ce11
surface as a whole; the second measures the colonial hydrophobici ty of multicellular aggregates
(Rosenberg and Doyle, 1990). Among the numerous methods proposed, three techniques are
typically used in studge floc research: microbid adhesion to hydrocarbons (MATH); contact
angle measurement (CAM); and salt aggregation test (SAT).
1) Microbiai Adhesion to Hydrocarbons: This technique was originally developed by
Rosenberg et al. (1 980). They found that various bacterial strains possessing hydrophobic surface
characteristics adhered to liquid hydrocarbons, whereas hydrophilic strains did not. Based on
different affinities of bacterial strains to hydrocarbons, the adherent microorganism with a
hydrophobic surface will be observed adhering at the oil-water interface. MATH has been
proposed as a simple and general technique for studying ceil surface hydrophobicity. In recent
years. Kocianova et al. (1992), Jorand et al. (1994) and Kertey and Forster (1995) have used this
method to measure the hydrophobicity of sludge flocs.
This simple experimental approach is based on mixing washed ce11 suspensions with test
hydrocarbons (n-hexadecane, n-octane, p-xylene) for a given time and then measuring adhesion
simply as the decrease in the turbidity in the aqueous phase after separating the two phases. For
application in sludge samples, a slight modification is rcquired. The sludge sarnple is washed and
resuspended in a phosphate-urea magnesium (PUM) buffer (Rosenberg et al., 1980) to give an
35
initial absorbance of 1.5 at 400nm. Then, the mixture of the microbial ce11 suspension and
hydrocarbon with the sarne volume is shaken for 2 minutes. The relative hydrophobicity is
calculated fiom:
where a is the absorbance of the aqueous layer after phase separation; A is the initial absorbance
of the aqueous phase at 400 nm before m f i g with hydrocarbons.
Although MATH has been demonstrated to be a simple and effective method for the
measurement of bacterial cell surface hydrophobicity, attention and care must be paid to the
possibility of cell clurnping during the assay, which can resutt in a huge reduction in absorbance.
Changes in the initial ce11 density c m also affect the measurement. Thus, in any comparison
among closely related strains or treatrnents, the initial ce11 density must be comparable to permit
valid conclusions regarding the relative hydrophobicity (Rosenberg, 1984).
2) Contact Angle Measurement: At present, CAM is one of the most common techniques
for the measurement of hydrophobicity of bacterial ce11 surfaces because the surface fiee enerw
of these cells can be estimated fiom the measurement (Absolom et al., 1983; Busscher et al.,
1984). Although difierent apparatus may be used, al1 the measurements involve the preparation
of a thin bacterial lawn through the vacuum filtration of a bacterial suspension and the
determination of sessile drop contact angles on the bacterial lawn, either by using a
telegoniorneter or by projecting a magnified image system. The most recent advance is the
application of axisymmetric drop shape analysis (ADSA), which is believed to overcome some
of the problems inherent in contact angle measurements on biological cells (Duncan-Hewitt et
36
ol., 1989; Neumann et al., 1996). CAM has been used to evaluate the hydrophobicity of sludge
flocs by Valin and Sutherland (1 982) and Pere et al. (1 993).
3) Sait Aggregation Test: SAT is an extremely simple technique for studying the
aggregative behaviour of celis by increasing concentrations of salting-out agents. This technique
was developed by Lindahl et al. (198 1 ) for testing the hydrophobicity of pure bacterial strains.
The principle behind SAT is based on the aggregation of cells by saits. The order in which cells
are aggregated and settied is a measure of their surface hydrophobicity. ï h e most hydrophobic
cells are first aggregated and settled at a low salt concentration. During the evaluation, different
concentrations of ammonia sulphate solutions are mixed with an equal volume of bacterial
suspensions. A reaction causing optimal aggregation is regarded as positive. A reaction giving
only a few aggregates or none at al1 is regarded as negative. Al1 readings are compared to the
reaction at the highest rnolarity of s d t (positive control). Bacterial suspensions mixed with a
0.002M sodium phosphate (pH 6.8) without adding salts are used as a negative control. An
improved SAT method, in which a drop of methylene blue is added to enhance the visualization
of the aggregates, was developed by Rozgoni et al. (1 985). The improved SAT technique is very
rapid and sensitive; the reaction is easily read with the naked eye. Recently, Urbain et al. (1993)
used this method to study the intemal hydrophobicity of sludge flocs.
The SAT technique has several limitations. First, many hydrophobic bacterial cells will
clump in the absence of any added ammonium sulphate. Second, it provides only a qualitative
estimation of the relative rank of hydrophobicity. Finally, the electrostatic interaction may affect
the results of SAT more than other hydrophobic measwement techniques (Rosenberg and Doyle,
1990).
37
At present, the MATH, CAM and SAT methods are used in the study of sludge flocs, but
it appears that the CAM is the most suitable method with high reproducibility in evaluating the
hydrophobicity of sludge surfaces (Pere et al., 1993) and bactena (van Loosdrecht et al., 1987).
The CAM was selected to investigate the hydrophobicity of sludge in this study.
Determination of Surface Charne: A variety of methods have k e n developed to determine the
microbial surface charge. These include colloid titration (Watanabe and Takesue, 1976),
attachment to charge-modified polystyrene, fluorescent probe ion exchange resin, and
electrophoretic mobility @enyer et al., 1993). The surface charge of sludge flocs has long been
measured using either electrophoretic mobility or colloidal titration.
Electrophoretic mobility is particularly valuable because it allows the estimation of the
zeta potential. However, the zeta potential must be calculated fiom the migration velocity of cells
in an electric field, which is strongly related to the water chemistry of the suspension solution,
and the type, size, and shape of the cells. The microorganisms and flocs in sludge have different
size distributions. The average migration velocity is therefore required for calculating the zeta
potential. An accurate measurement of this type is difficult, and requires skiIl, experience. and
time. along with the use of specialized apparatus.
In contrast, the colloid titration method involves the application of the standard positively
and negatively charged polymers. For sludge samples, an excess amount of positively charged
polymer is added to the sludge suspension, then the excess amount of positively charged
polymers is titrated by a negatively charged polymer; the surface charge density of sludge flocs
can be calculated from the titration. This is an easy, convenient method widely used in the
38
determination of dosage in tlocculation processes. Currently, çoNoidal titration is the most
common rnethod for determining the surface charge density of sludge flocs (Morgan et al., 1990;
Mikkelsen et al., 1996).
Considenng the complexity of sludge flocs, in this study, colloidal titration was chosen
for the evaluation of surface charge of sludge flocs.
2.7 Summary of Literature Suwey
Based on a literature s w e y of past and recent publications, conclusions about the present
state of knowledge of sludge flocs can be summarized as follows:
1) A number of flocculation mechanisrns have been proposed to explain sludge floc
formation. These models are effective under certain conditions, but their validity is often
challenged by unexplained experimental phenomena. At present, polymer bridging seems to be
the most acceptable mechanism. Information about hydrophobic interactions of cells is still
scarce. The dominant forces governing sludge floc formation have not yet been clearly identified,
especially the mechanisms of dispersed growth and disintegration of sludge flocs.
2) The presence of EPS on sludge surfaces is traditionally considered to induce and
enhance the formation of flocs-polymer bridging. However, this conclusion has been questioned
in recent years. Experimental studies have demonstrated the negative effect of EPS on the
formation of compact flocs under certain conditions. It seems that the presence of EPS on sludge
surfaces has complicated eiTects on studge floc formation. Emphasis on not only the arnount but
39
also on the composition and physical configuration of EPS should be useful in clwifying and
understanding the precise role of EPS in floc separation and dewatering.
3) A number of contradictory conclusions with respect to the role of sludge swfaces and
operating variables, such as SRT, in bioflocculation and settling exist in previous studies. This
confusion may be due to the influence of a number of uncontrolled factors with sludge samples
from poorly controlled fuli-scale activated sludge systems. It would be desirable to understand
the precise influence of environmentai and operating variables on the performance of activated
sludge processes using ngorously controlled conditions.
3) Reliable and reproducible measurement techniques are essential. Based on past and
recent work. there is still much progress to be made toward developing measurement techniques
for sludge flocs.
5) Although there are some publications on either the physical or the chemical properties
of sludge flocs. there is limited integrated research on both the physical and the chemical
properties. The relationships between the physicochernical properties and the flocculating ability,
settleability and compressibility of sludge flocs are still not clear.
2.8 Significance of this Sîudy
Based on the literature survey, it is clear that the phenomena of sludge floc formation and
floc structure are complicated, and affected by a number of physical, chemical and
microbioIogical conditions. A better control of sludge floc formation to get the desired
physicochemical properties for effective floc separation depends on an improved understanding
40
of the mechanisms of sludge floc formation and floc properties. Therefore, a comprehensive
study on the causes of sludge floc formation and the variations in surface properties and
interparticle interactions of sludge flocs under rigorously controlled conditions is of primary
importance. This researc h enhances understanding of the surface pro perties, and interparticle
interactions of sludge flocs at. different SRTs, and of the role of surface properties in
biofloccuIation, compression, and stability.
Cornpared with the results of previous studies, the research as described in this thesis
advances the knowledge of sludge floc formation and properties in the folIowing ways:
1) The effect of the operating variable - sludge retention time (SRT) on sludge floc
formation and its mechanical stability was studied in rigorously controlled laboratory-scale
sequencing batch reactors (SBRs). The possible influence of other dominant factors was
minimized.
2) Not only the arnount of EPS, but also the composition and molecular weight of EPS in
response to the change in the SRT were investigated using rigorously controlled laboratory-scale
SBRs.
3 ) The role of surface properties, such as hydrophobicity and surface charge, of sludge
flocs in bioflocculation and settling and their responses to the change in the SRT were
systematically investigated.
41
4) The influence of interparticle interactions on the stability of sludge flocs and theu
responses to the change in the SRT were evaiuated in batch experiments by suspending sludge
flocs under difiierent water chemistry conditions.
5) The potentiai correlation between the surface properties and bioflocculation,
compression, and stability of sludge flocs is evaluated and discussed.
CHAPTER III
EXPERDMENTAL MATERIALS AND METHODS
Described in this chapter are the experimental materials and methods used in the present
study. The subsections are outlined in the following sequence: experimental approach;
expenmental system; microbiology; extraction and chemical analyses of extracellular polymenc
substances (EPS); characterization of EPS; contact angle measurement; surface charge
determination; colIoidal stability test; and statistical methods.
3.1 Experimental Approach
Rational: A laboratory-scale sequencing batch reactor (SBR) system fed with a synthetic
wastewater (Table 3.1) was selected for this study. This system consisted of four SBRs operated
in parallel at different sludge retention times (SRTs). The SBRs were nin for a long-term period
(two years) so that tme stable conditions were reached to generate biomass for characterization.
This approach pmvided the following advantages: 1) the specific factor, such as SUT, was
studied in isolation under well-controlled conditions; 2) compared to a conventional reactor, the
SBR had a smaller reactor volume, was relatively easy to control, and was operated in a cyclic
manner to create a substrate gradient for minimizing filamentous bulking; 3) the type of substrate
and levels of nutrients (organic and inorganic) were easily manipulated. The low concentrations
of cations (0.1-0.5 mgL) in the feed (Table 3.1) were used to minimize the effect of chemical
flocculation; and 4) sludge samples were collected and characterized imrnediately. It is well
known that the transportation and storage of siudge samples has a significant impact on sludge
properties (Jenkins et al., 1993; Nielsen et al., 1996; Bura et al., 1998).
Table 3.1 Chernical Composition of the Synthetic Wastewater
(COD:N:P=100:5: 1; COD fiom glucose; N fiom NH,Cl and P from KH,PO,)
Cvcf ic Ooerution of the Seauencin~ Batch Remtors: The SBR is a fill-and-draw activated sludge
(COD = 140 - 540 mg/L)
system. Al1 SBRs have four operating steps in each cycle that are carried out in a time sequence
r
Chernicals
MgS04.7H20
FeSOq.7H20
Na2Mo04.2HsO
?#inS04.4H20
CuSO4
ZnS04.7H20
NaCl
CaS04-2H20
CoC12.6H20
as follows: 1) fill; 2) reaction (aeration); 3) sedimentation/clarification; and 4) draw. The
operating sequence is illustrated in Figure 3.1. Sludge samples were collected for
* Glucose, NH,CI and KH,PO, obtained fiom BDH Chemicals were of analytical grade.
Concentration in the Fe& (m%L)
5.07 (0.5 M ~ Z + )
2.49 (0.5 ~ e 2 + )
1.26 (0.5 MO@)
0.3 1 (0.1 ~ n 2 + )
0.25 (0.1 c d + )
0.44 (0.1 ~ n 2 + )
0.25 (O. 1 ~ a + )
0.43 (0.1 ~ a 2 + )
0.41 (0.1 CO^+)
characterization at the end of the reaction phase.
Grade and Source o f Cbemicrils
Analytical (Fisher Scientific)
Analytical (Fisher Scientific)
Analytical (Fisher Scientific)
Analytical (ACP Chemicals)
Analytical (Fisher Scientific)
Analytical (BDH chemicals)
Analytical (BDH chemicals)
Analytical (BDH chemicals)
Analytical (Fisher Scientific)
Oxygen m d stin-ing: on On Off Off
Fil1 of substrate Clarification Effluent and siudge discharge
Figure 3.1 Process flow diagram in the operating sequence of the SBR
Control ofSlttdre Retention Time: Changes in the SRT were achieved by controlling the
arnount of sludge wasted per day as well as by taking into account effluent suspended solids
(ESS) lost in treated effluent. Mixed liquor suspended solids (MLSS) and ESS were monitored
every 3 to 14 days. These measurements gave the information necessary for calculating the SRT
and determining the change in the arnount of sludge wasted per day to achieve better control of
the SRT. The equation for calculating the SRT is as follows:
Sludge in Reactor (gVSS) SRT = - - vo x
Sludge waste (glday) (Qw x + QE XE) 7
where SRT is the sludge retention time, [days]; Vo is the volume of each SBR, [LI; X is the
mixed liquor suspended solids (MLSS) concentration measured at the end of one cycle (prior to
settling), [mg/L]; Qw is the arnount of sludge wasted per day from each SBR, IL]; QE is the
45
amount of discharged effluent per day from each reactor, [LI (5.4 L/day in this study); and XE is
the effluent suspended solids (ESS) concentration in treated effluent, [mg/L].
Experimental Procedures: At the start of this study, the SBR system was inoculated with
biomass (Tables 3.2 and 3.3) fiom the aeration tank of an activated sludge plant treating
municipal wastewater (Main Treatment Plant, City of Toronto). Initially, al1 four SBRs were
operated at an SRT of 6 days. Following a five-week stabilization period, three of the SBRs
(SBR-2, -3 and -4) were initially switched to an SRT of 12, 16 and 20 days, respectively. SBR-1
was kept at an SRT of 6 days for another six weeks and then switched to an SRT of 4 days. SBR-
2 was switched to an SRT of 9 days fiom 12 days after day 124 of the experimental period. The
influence of biomass concentration on the flocculating ability and compressibility of sludge flocs
in the SBRs was minimized by achieving a similar level of MLSS (about 2000 mg/L) in each
reactor. A similar level of MLSS with changing SRTs was maintained by changing the influent
COD. as described in Chapter IV.
Stable operation of the SBRs at each SRT was determined by monitoring some selected
parameters, such as MLSS, ESS, sludge volume index (SVI), effluent soluble chernical oxygen
dernand (COD) and the carbon substrate oxidation profiles of the microbial community.
Considering the fact that the process of population shifi is much slower than their growth
(Sheintuch. 1987; Characklis, 1990), at least 5 times the SRT was allowed for SBRs at lower
SRTs (4 and 9 days) to establish stable operation. For SBRs at higher SRTs (12,16 and 20 days),
at least 3 times the SRT were allowed to build stable operation. Only when al1 the monitored
parameters showed relatively constant levels, was the SBR considered to be at stable conditions.
On several occasions, sludge in SBR-1, -2, -3 expenenced modest non-filamentous bulking (100
46
< SV1 < 200 mL/g MLSS) and dispersed growth (ESS > 50 mg/L), but the bulking and dispersed
growth phenornena were usually diminished wvithin two weeks.
A long-term study over two years was carried out to study the influence of SRT on sludge
properties. This is because al1 previous studies as discussed in Chapter II (pages 25-29) were
conducted in a short-term period (a few weeks to 2-3 months), and resuits fiom these studies
were contradictory. One of the reasons for explaining these differences might be that no stable
operation was really reached in a short-term study. Therefore, a long-term study is desirable to
test this hypothesis and make sure the results reflect the tme response of sludge flocs to new
conditions. Surface properties of sludge flocs at different SRTs were the focus of this study in the
first one and half years. Interparticle interactions and the colloidal stability of sludge flocs at
different SRTs were evaluated in the last half a year of the operation of the SBRs.
Table 3.2 Typical Physical-Chernical Properties of the Inoculum
1 MLSS (rng/L) 1 vss (mw) 1 SVI (mL/g) 1 ~urfncc Charge
I I I 1 (meq./g VSS)
Table 3.3 Chernical Composition of Extracellular Polymeric Substances of the Inoculum
Composition
Concentration ( m g k VSS)
Carbohydrates
12.7
Proteins
96.6
DNA
6.5
Acidic Polysaccharide
4.8
47
Characteriaz~ion of the Mked Liauor: Characterization of the mixed liquor fiom the SBRs
included the following items: standard wastewater analyses; microbiology of sludge; and
c haracterization of surface properties and the colloicial stability of sludge flocs. The parameters
used for characterization are illustrated in Figure 3.2. Al1 the measurement and analyticai
methods used in this study are described in M e r detail in Section 3.3 of this Chapter.
1 Standard Wastewater 1 1 Analyses 1
Microbiologicrl Floc Cbarscterization
=iMixed liquor suspended solids =Quantification of filaments -Volatile suspended solids *Chernical oxygen demand =Dissolved oxygen =Sludge volume index Surface Analyses Colloidrl Stability =Effluent suspended solids *Total carbohydrate content
=Extraction of EPS *Turbidity =Chernical analyses o f EPS *Dissociation constant
=Total carbohydrate .PH *Protein =Ionie strengih -DNA Cation valency =.4cidic polysaccharide -EDTA
aHydrophobicity =Urea *Surface charge
Figure 3.2 Parameteric classifications for mixed liquor characterization
3.2 Experimental System
The expenmental system, as illustrated in Figure 3.3, is composed of the followîng
elements: four SBRs operated in parallel for treating wastewaters and producing biomass; a
refi-igerator for storing the synthetic feed at 4OC; a preheater tank that increases the temperature
of the synthetic feed fiom 4 ' ~ to 28OC before it enters the SBRs; a water bath that circulates
water at a constant temperature (2g°C) through the jacket of each SBR; four on-line pH
controllers (one per SBR); and timers for controlling the cyclic operation of each SBR.
Figure 3.3a Laboratory-scale Sequencing Batch Reactor Set-up
Feed Staoge
Figure 3.3b Schematic diagram of the laboratory-scale SBR system
A general description of the laboratory-scale SBR system is presented in the following
sections:
Feed Sforaae Refii~eraior: The feed was prepared by diluting the stock solutions (glucose,
NH,Cl + KH2P0,, inorganic salt solutions) to the desired concentration in four autoclavable
rectangular polypropylene carboys (votume: 9 L each) (Naigene Company. Rochester, NY) and
then stored in a bar sized refiigerator (W. C. Wood Co. Ltd, Guelph, ON). The purpose of using
a refngerator was to minimize the potential biodegradation of the feed by maintaining the feed at
a low temperature (4°C) during storage (< 24 hours).
50
Preheater Units: The preheater units are composed of four holding tanks (95 mm I.D. x 340 mm
length, 2 L capacity) which are suspended in a preheater water tank made of plexiglas (340 mm x
360 mm x 500 mm). The holding units are custom-made cylindrical glass tanks with flanged
nms and outlet ports at the bottom. The preheater water tank is made of acrylic (hci teR) and
maintained at 28OC using an aquarium type immersion heater (Thermal Compact Pre-set
submersible Aquarium Heater, Rolf C. Hagan Inc., Saint Laurent, PQ). Each of the four fixed-
speed peristaltic pumps is attached with a level controller (Single Point Controller, IMA
Industries, Plainville, CT, USA) with an adjustable height polypropylene float switch (Madison
Co., Branford, CT, USA) to control the transfer of 0.9 L feed fiom the refiigerator to each
holding unit in each cycle- The purpose of the preheater system was to increase the feed
temperature fiom 4OC to 2S°C before the feed entered the SBRs.
Se~uencinn Butch Reactors: Each SBR was made of glass with an operating diarneter of 127 mm
and a height of 340 mm. It was enclosed within a 25.4-mm annular thickness glass jacket for
temperature control and had a volume of 2 L (Figure 3.4). Five outlet ports were positioned on
each SBR. corresponding to volumes of 0, 0.4, 0.5, 0.8 and 1 L fiom the bottom. The 0.4 L port
was used for discharging the treated effluent, and the 0.5 L port was designed for collecting
mixed liquor samples. The other ports were sealed with rubber septa. The aeration, feed and pH
buffer tubes, and pH probe to the SBR were held in a rubber stopper with a hollow ring made of
acrylic ( ~ u c i t e ~ ) at the top of each SBR.
The sludge suspension was continuously stirred in the reaction phase with a magnetic
stirring bar (0.7 cm x 0.7 cm x 5 cm) placed at the bottom of each SBR. The mixing intensity
was maintained by appropriate adjustment of each magnetic stirrer and by controlling the flow
rate of air. The air, which served as an oxygen supply, was introduced in the SBR through a
stone air diffiser positioned at a level correspondhg to approximately the 0.4 L reactor volume.
The SBRs were housed and secured in a wooden h m e (Figure 3.3a).
I L Port
0.5 L Port probe
Figure 3.4 Schematic diagram of a laboratory-scale SBR
Tern~eraîure Controller: The operating temperature of the SBRs was maintained at 28OC by
circulating 28OC water through the SBRs jacket. The water bath was maintained at 2g°C. The
water was circulated through the jacket of the SBRs using a variable speed peristaltic pump
driving with four pump heads (Masterflex Standard Pump Drive and Masterflex L/S size 18
Pump head, Cole-Parrner Instrument, CO., Niles, Illinois, USA).
On-lhe PH Controikr and Buffer: Control of the pH was achieved by immersing a pH electrode,
which was connected to an on-line pH controller (LED pWORP controller, Cole-Parmer
Instnunent Co., Niles, Illinois, USA) at the 0.4 L reactor volume level in each of the four SBRs.
52
A 0.02-0.05 N NaOH solution was used as a buffer and was intmcîuced in the SBRs by the on-
line pH controllers to maintain the pH at the desired value (7.0 f 0.2).
Thers : Control of the cyclic operation of the SBRs was achieved through the use of four on-
and-off programmable timers (Sper Scientific Mode1 810030, Sper Scientific Ltd., Scottsdale,
Arizona USA). The four programmable timers were used to control the transfer of the cold feed
to the preheater unit, the transfer of the warmed feed to the SBRs, the aeration and mixing tirne,
and the discharge of the treated effluent, respectively. Control was achieved by switching the
electric equipment on and off at predetermined times. The operathg conditions of the SBR
system are shown in Table 3.4.
Table 3.4 Operating Conditions of the SBR System
Parameters Units Value
Sludge retention time
Hydraulic retention t h e
Effective volume of each SBR
Cycle length
Filling period
Aeration (reaction) perïod
Compaction penod
Wi thdrauing period
Operating temperature
Operational pH
Days
Minutes
Minutes
Minutes
Minutes
53
A11 polyvinyl tubes for transfening the feed to the SBRs were cleaned with detergents
every two weeks, in order to prevent the accumulation of biomass in the tubes.
3 -3 Measurement and Analytical Methods
Al1 the measurement and anaiytical methods as described in Figure 3.2 are described in
fùrther detail in the following paragraphs,
3 -3.1 Standard wastewater analyses
General characterization of wastewater and biomass followed standard methods described
by the American Public Health Association (APHA) (1992).
M x e d and Volatile Liuuor Susvended Soli&: Mixed and volatile suspended solids (MISS and
VSS) were measured at the end of the reaction phase in accordance with Standard Methods
(APHA. 1992).
Chernical O m e n Demand The closed reflux colorimetric method (Section 5220D. APHA,
1992) was used to determine chemical oxygen demand (COD) of the feed and the treated
effluent. The treated effluent was filtered through a 0.45 Pm pore size filter paper (Gelman
Sciences Filter Paper, diameter 25mm) before COD measurement. Cultwe tubes with desired
effluent (2.5 mL). digestion solution (KCr20, + HgSO, + H2S04), and reagents (Ag,SO, +
H,SO,) were heated in a Hach COD reactor (Mode1 45600-00, Hach Co., Loveland, CO, USA)
for 2 hours at 1 50°C. The cooled samples were then measured spectrophotometrically (Bausch &
Lornb Spectronic 20D with Hach 19230-00 Adapter) at 600 nm. Potassium hydrogen phthalate
54
(KHP) was used as a COD standard. Al1 chernical reagents used for COD measurements were
fiom BDH Chernicals Inc. and were of analytica! grade.
Dissolved c)Xy~en: The dissolved oxygen (DO) levels in the SBRs were fiequently monitored
using a DO meter (Mode1 600 Oxygen Analyzer, Engineered Sysîems & Designs, Newark. DE,
USA). The DO levels during the reaction phase were maintained at 2.5 to 5.5 ppm in each SBR.
Sl&e Volume Index: The cornpressibility of sludge was evaluated by the sludge volume index
(SVI). The SV1 is the volume in mL occupied by one gram of MLSS afier 30 minutes of
compaction. The SV1 measurement was carried out by placing the well-mixed liquor in a 250 mL
graduated cylinder. The sludge samples for the SV1 measurement were directly taken fiom the
mixed liquor at the end of the cyclic operation in the SBRs. The concentration of the mixed
liquor for the SV1 measurement was usually in the range of 2000 +. 300 mg& where the effect of
sludge concentration on the SV1 is not important (Lovett et al., 1983).
SV1 = VoIurne in rnL afier 30 minutes of cornriaction in a 250 mL nraduated cvlinder x 4, (3-2)
Biornass concentration (g/L)
E f i e n r Suspended Solids: The flocculating ability of sludge flocs in the SBR system was
evaluated by the effluent suspended solids (ESS) measurement as described in Standard Methods
(Method 2540D) (APHAJ992). The level of ESS was detennined after a 40-minute compaction
of the mixed liquor.
Total Carbohvdrafe Content of the m o l e Sludkg: The total carbohydrate content of sludge
samples at each SRT was determined by the Anthrone method as described by Gaudy (1962).
The Anthrone method is described in Section 3.3.3 of this chapter.
3 -3.2 Microbiology
The general microbiology of sludge was examined both by microscopie observations to
check the presence of filamentous microorganisms, and by the phenotypic fingerprinting to
monitor the carbon substrate oxidation profiles of the microbial comrnunity structure in each
SBR.
3.3 -2.1 Microscollic Observarion: The morphology and structure of sludge flocs were extensively
examined with a light microscope (Olympus, BH2-RFCA) at a magnification of X 400. The
number of filaments was classified into levels 1 to 6 according to Jenkins et al. (1993). A smaller
score corresponds to a lower level of filaments.
3 -3 -2.2 Phenoh/oic Finpervrintinq: Characterization of the microbial community structure was
conducted according to phenotypic fingerprinting, which uses carbon substrate oxidation profiles
to describe the comrnunity level microbiology. Information fiom phenotypic fingerprinting
indicates whether microbes fiom different sources, andor operating and environmental
conditions. are functionally similar as a community, based on their carbon source utilization
characteristics. The basic procedures are descnbed below:
A 2 mi, sample of mixed liquor fiom the SBR at the end of the cyclic operation was
diluted in a Waring blender to 100 mL with deionized distilled water to make a biomass
concentration of about 100 mg/L. One drop (10 PL) of TWEEN 80 and one drop of sodium
pyrophosphate were added in sequence to the blender. The suspension was then homogenized
using the blender for 30 seconds. The recovered biomass was washed three times with saline
(0.85% NaCI) and centrifùged at 15,000 x g for 15 minutes each time to remove contaminating
56
organic materials that had accumulated on sludge surfaces and in the liquid phase of the biomass
suspension. M e r treatment and washing, aliquots of 150 pL of biomass suspension were
transferred into each well (96 wells) of the Biologm GN microplates (BIOLOGTH Inc.,
Hayward, CA, USA) and incubated at 37°C for 24 hours. Before and after inoculation at 37"C,
the pattern of absorbances of the microplates was spectrophotometrically measured at 590 nm
with a Biologm microplate reader (BIOLOGW Inc., Hayward, CA, USA) in conjunction with
Biolog- Microlog software, version 3.5. Duplicate GN microplates were used for each sample.
The resulting multivariate data set was subjected to the principal component analysis
(PCA) (STATISTICA Software, Statsoft, Tulsa, OK, USA) in order to establish associations
between carbon sources and microbial communities in each of the SBRs (Vittorio et al., 1996;
Schneider et al., 1998). Relationships were observed by retaining the first N o principal
cornponents (PC 1 and PC 2) and plotting in two dimensions.
3 -3 -3 Extraction and chemical analyses of extracellular polymeric substances
3.3.3.1 Extraction The extraction of EPS fiom sludge flocs was performed by a Dowex
Cation Ion Exchange Resin method (Fralund et al., 1996; Bura et al., 1998). Sludge samples
collected from the SBRs at the end of the reaction phase were first concentrated to about 3.0 - 5.0
g/L by gravity sedimentation. Then concentrated sludge samples were washed twice with
extraction buffer solution (Fralund et al., 1996) at pH=7 and centriîüged at 2000 x g for 5
minutes each time. Washed samples were resuspended in about 60 rnL of extraction buffer
solution. A measured amount of Dowex cation ion exchange resin (90 g residg MLSS) was then
added to the extractor assembly (I.D. 1Ocm; height: 10cm) with washed sample. Next, the
57
extractor assembly was kept in an ice-water bath and stirred at 600 rpm for 2 hours. The mixture
fiom the extractor assembly was then transferred into two centrifiigal tubes (50 mL each) and
centrifûged twice at 12,000 x g for 15 minutes each. Finally, the supernatant was separated f?om
the sludge pellets; it contained the aqueous EPS fi-actions. A 2-hour extraction time was used in
this study, as suggested by Frdund et al. (1 W6), to minimize ce11 lysis.
3.3.3.2 ChemicalAnalvses Chemical analyses of EPS composition were conducted by
colorimetric methods. Spectrometry (carbohydrates, proteins and acidic polysaccharides) and
fluorometry (DNA) were carried out by a SP6-500 UV spectrophotometer (PYE UNICAM) and
a fluororneter (Mode1 110, G. K. Turner Associates, USA), respectively. Deionized distilled
water (MiIlipore Water) was used as the blank solution in each measurement. Al1 chernicals used
in this study were obtained fiom Sigma-Aldrich Chemical, Inc. and were of analytical grade.
Four components (carbohydrates, proteins, acidic polysaccharides and DNA) were measured in
the EPS. Al1 EPS component concentrations were converted into mg (standard equivalent) /g
VSS.
Tora2 Carhohydrates: The detemination of the total carbohydrate concentration was performed
by the Anthrone method (Gaudy, 1962) using glucose as a standard. The Anthrone solution was
prepared fiesh for each day's analysis at least 2 hours before use by dissolving 0.2 g of Anthrone
reagent in 100 rnL of 95% sulphuric acid. The standard solution was made fiom 100 mg/L stock
glucose solution stored in the refiigerator at 4°C for no more than two weeks. A series of
standard solutions containing 5-100 mgL glucose were prepared to calibrate the standard c w e
when analyzing each batch of EPS solution.
58
A known volume (2.5 mL) of the extracted EPS supernatant and of each of the standards
was pipetted into the HACH test tubes (10 mL each), and an aliquot (5 mL) of the cold fiesh
Anthrone reagent was added to each HACH test tube. Then the HACH test tubes were capped
with rubber stoppers and mixed thoroughly on a vortex mixer for a few seconds. After mixing,
the HACH test tubes were placed into a boiling water bath for 15 minutes. At the end of the
heating penod, the HACH test tubes were taken out of the boiling water bath and then cooled
down in an ice-water bath- The intensity of colour developed in the solution at room temperature
was rneasured as the absorbance at 625 nrn. The total carbohydrate concentration (glucose
equivalent) in the EPS solution was then calculated fiom the standard curve (mg/L) and finally
converted into mg/g VSS, based on the biomass concentration for EPS extraction.
Proreins: The protein concentration was measured by Lowry's method (Lowry et al., 1% 1) wïth
bovine serum albumin (BSA) solution (50-200 m&) used as a standard. Three reagents were
required in the Lowry's method: 1) Reagent A was a 2% (wh) Na,CO, and 0.1 N NaOH
solution: 2) Reagent B was prepared fresh by dissolving 0.5 g CuS0,.SH20 in 100 mL of a 1%
(w/v) aqueous solution of sodium tartrate; and 3) Reagent C was prepared by mixing Reagents A
and B at a ratio (vh) of 50 to 1 just before use. Aliquots (1 mL) of blank, standard and sample
solutions were placed into each HACH test tube, and 5 mL of Reagent C was added into each
HACH test tube and mixed well. The mixture was allowed to stand for 10 minutes at room
temperature. Then 0.5 mL of diluted Folin reagent (Folin-Ciocalteau/Deionized distilled water =
18/90 (v/v)) was added to the mixture and mixed immediately. The mixture was again allowed to
stand for 30 minutes at room temperature to develop colour. The absorbance of the solutions was
59
then measured at 750 nm using a spectrophotometer. The protein concentration was calculated
fiom the standard curve and converted into mg (BSA equivalent) / g VSS.
A cidic Polysaccharides: A modi fied su1 famate/m-hy droxydipheny 1 sulfùric acid method was
used to detennine the acidic polysaccharide concentration; it was adapted fiom Filisetti-Cozzi
and Carpita (1 99 1). D-glucuronic acid was used as a standard. Aiiquots (0.8 mL) of blank,
standard (1 -10 mgL) and sample solutions were pipetted into the HACH test tubes, and 80 pL
of 4M su1 famic acid-potassium suifamate (pH =1.6. adjusted with NaOH solution) was added to
these tubes and mixed thoroughly. Anaiytical grade H2S04 (96%) containing 75 mM sodium
tetraborate (4.8 mL) was then added, and the mixture was stirred vigorously with a vortex mixer.
The HACH test tubes were capped with rubber stoppers and heated in a boiling water bath for 20
minutes. At the end of the heating period, the HACH test tubes were chilled in an ice-water bath
to cool the solution down quickly to room temperature. m e r cooling, 160 PL of 0.15% (w/v) rn-
hydroxydiphenyl in 0.5% (wh) NaOH was added and then stirred vigorously in a vortex mixer.
The pink colour was developed in 5 to 10 minutes. The absorbance was measured at 525 nm. The
acidic polysaccharide concentration was calculated fiom the standard curve and converted into
mg (glucuronic acid equivalent) / g VSS.
DN4: The DNA concentration in the EPS was rneasured by the DAP1-method (Brunk et al.,
1979) and salmon testes DNA (Sigma) was used as a standard. Aliquots (200 PL) of blank,
standard and sarnple solutions were added to the HACH test tubes; 5mL of DAPI reagent
(O.2ppm DAPI in 100 mM NaCl, 10 mM EDTA, 10 m M Tris solution, pH = 7.0) was added to
each tube and mixed vigorously. The mixture was then allowed to stand for 10 minutes at room
temperature to develop fluorescence. After that, the solution was transferred to a cuvette to
60
measure the fluorescence directly (excitation at 360 nm and emission at 450 nm) by the
fluorometry .
Total EPS Content: The sum of the total carbohydrate, protein and DNA content in the EPS was
considered as the total EPS content, since carbohydrates, proteins and DNA are the dominant
cornponents of EPS (Forster, 1976 and 1985; Frdund et al., 1996; Bura et al.. 1998).
3.4 Molecular Weight Distribution of Extracellular Polymeric Substances
The molecular weight distributions (MWDs) of EPS components were evaluated using
ultrafiltration. The set-up for ultrafiltration is illustrated in Figure 3.5. The pressure ultrafiltration
ceII used in this study had an effective volume of 180 mL (Amicon, USA) and was equipped
with a stirrer assembly. The ultrafiltration membranes, used in sequence, had normal molecular
weight exclusions of 100,000, 10,000 (low protein binding), and 1,000 (low protein binding)
daltons (Millipore Corporation, USA). Some 100 mL of clear EPS supematant was passed
through each of these membranes in sequence. The driving force of the nitrogen for each
membrane was 20, 40, and 70 psig, respectively. In order to minimize decomposition of EPS
components, the ultrafiltration ce11 and the filtrate collecting beakers were kept in an ice-water
bath. The filtrates together with the original EPS supematant were measured for total
carbohydrates, proteins and DNA.
Figure 3.5 Schematic diagram of the ultrafiltration set-up for rnolecular weight distribution
determination (1- Magnetic stirrer; 2- Ice-water bath; 3- Ultrafiltration chamber; 4-
Ultrafiltration membrane; 5-S timng assembly)
3 -5 Contact Angle Measurement
The hydrophobicity of sludge flocs was evaluated by water contact angle measurements.
which were carrïed out using a modified Axisymmetric Drop Shape Analysis-Contact Diameter
(ADSA-CD) technique (Duncan-Hewitt et al., 1989; Neumann et al., 1996). A schematic
diagram of the ADSA-CD set-up is shown in Figure 3.6. This method involves two separated
steps: preparation of sludge layers on the membrane. and contact angle measurement using
images.
ADSA-CD MicroscopeICCD Carnera
Digitizer - Computer
Incident Illuminator ,, k-e7- Liquid drop
Light Sludge Cake on a 2% Agar Plate source
Figure 3.6 Schematic diagram of an ADSA-CD setup (modified from Neumann et al., 1996)
Sludge samples fiom the SBRs were washed with deionized distilled water twice using a
centrifuge at 2000 x g for 5 minutes each time. The washed sludge samples had concentrations of
2200 + 200 mg/L; they were dispersed using a vortex mixer at level 10 for one minute. Then 10
mL of the dispersed sludge sample was transferred to a g l a s filter holder (Millipore) equipped
with a cellulosic membrane (E3lack MSI Microsep*, pore size 0.45 pm). Under suction filtration
using a vacuum of 400 mmHg, the dispersed sludge was deposited on the membrane surface. The
sludge layers were washed twice with double distilled water during the filtration process. The
sludge cake was filtered until moist and there were no signs of excess water that could be sucked
out.
63
The membrane with the deposited sludge cake was then carefully tramferreci to the
surface of a fkeshly prepared 2% solidified agar plate to preserve the moisture in the cake. M e r
waiting about 15-20 minutes to equilibrate the moisture between the agar and the sludge cake,
the agar plate was placed in the ADSA-CD system. A drop of double distilled water was placed
on the sludge cake using a Gilmont micrometer syringe (accuracy 0.022 PL) equipped with a
stainless steel needle. A video image system was used to view the sessile drop fkom the top. The
drop shape was captured atter no M e r shrinking of the water drop was observed (5-7 seconds
after the sessile drop was placed on the surface of the sludge cake). The images of the sessile
drops were used to estirnate contact angle values using the ADSA-CD software (Neumann et al..
i 996).
In the early stages of using the ADSA-CD technique for the evaiuation of sludge
hydrophobicity. preliminary studies were conducted to h d suitable conditions for the contact
angle measurement. It was found that the ADSA-CD technique is highly reproducible for contact
angle measurements of sludge sarnples. The advaritages of this method are: 1) compared to the
telegoniorneter, arbitrariness in contact angle measurements is elirninated; 2) the contact angle
can be evaluated in a partly hydrated state; and 3) the drying time of the sludge cake has no
significant effect on the measured contact angles, as shown in Figure 3.7.
O 10 20 30 40 50
Time (minutes)
Figure 3.7 Effect of biomass concentration with measuring time on contact angle
measurements (-O- 12 mg sludge, drying 5 minutes; -a- 12 mg sludge, drying 15
minutes; -A- 16 mg sludge, drying 20 minutes)
3 -6 Surface Charge Determination
The surface charge of sludge was deterrnined by colloidal titration. Al1 the solutions were
adjusted to either neutrality or the expected pH values prior to titration. Polybrene (0.002 N) and
the potassium salt of polyvinyl sulphate (PVSK) (0.001N) ( d l chernicals from Sigma, analyticîl
grade) were used as the cationic and anionic standards, respectively. A known arnount of sludge
sample (2 mL, about 2000 mgR) was diluted with deionized distilled water with an adjusted pH
(7.0) and mixed with an excess amount of Polybrene standard solution. The PVSK standard
solution was then used to titrate against the excess amount o f polybrene using toluidine blue as
an indicator. An equal volume of Polybrene standard solution diluted with the same amount of
deionized distilled water for each titration series was used as a blank.
65
Prelirninary studies were conducted to test the effect of biomass concentration on the
surface charge rneasurernent. Figure 3.8 shows that the total charge density of the titrated
solution increased linearly with increasing biomass concentration. The linear behaviour of the
charge density plotted against biomass concentration indicates that biomass concentration did not
interfere with the surface charge measurement in the absolute uni& (meq./g VSS).
Biomass Concentration (g/L)
O , 1
-3.5 l Figure 3.8 Variation in the surface charge density of mixed liquor with respect to
biomass concentration
3.7 Stability Test
The stability of sludge flocs from different SRTs was characterized by the dissociation
constant as described by Zita and Hermansson (1994). nie dissociation constant is defined here
as unit absorbance/g dried biomasslper washing, Le., the number of particles or the mass of
particles removed per washing, based on one gram dried biomass. The basic procedures for the
stability test and the calculation of dissociation constants are descnbed as follows:
3 -7.1 Dissociation of flocs
Sludge flocs fiom different SRTs were taken fiom the SB& at the end of the reaction
phase. A 15 mL sludge simple (MLSS 2000 mg/L) was placed in a 50 mL test tube and
concentrated by gravity settling for 15 minutes. Afier concentration, 5 mL of supernatant was
removed fiom the test tube. The concentrated sludge was treated witb chemically manipulated
aqueous solutions and diluted to 30 mi,. The manipulations were: (a) adjutment of pH through
addition of 1N HC1 or 1N NaOH solution; (b) addition of different concentrations of KC1 and
CaClz to obtain different ionic strengths; and (c) addition of urea and EDTA. The diluted sludge
sarnples were gently mixed in a shaker (Model G25 Incubator Shaker, New Brunswick Scientific
Inc.. NJ, USA) at a speed of 200 RPM for 15 minutes. After 15 minutes of sedimentation
following the mild mixing procedure, 15 rnL of the supernatant was gently removed fiom the top
of the test tube and the turbidity was measured spectrophotometrically a 420 nm (SP6-500 W
Spectrophotometer. PYE, UNICAM, England). Then more solution was added to the remaining
sludge to the original volume (30mL) in the test tube. Subsequently, samples were gently mixed
for 15 minutes and settled for 15 minutes. These procedures were repeated 5 to 8 times.
3.7.2 Dissociation constant
The dissociation constant for each treatrnent was detennined fiom the slope of the
calculated straight line (using standard Iinear regression) of the accurnulative turbidity against the
number of sequential washings, as shown in Figure 3.9, in which the dissociation constant is
0.07 1 turbidity/g MLSS/washing.
O 1 2 3 4 5 6
Number of washings
Figure 3 -9 A plot of the accumulative turbidity venus the number of sequential washings
(sludge sarnples fiom an SRT of 4 days, treated with deionized distilled water)
3 -8 Statistical Methods
The Statistica (version 6.0) s o h a r e package (Statsoft, Tulsa, OK, USA), run on a
personal cornputer, was used for al1 statistical analyses. Basically, two types of statistical
analyses were used in this study: analysis of variance (ANOVA) and correlation. The
significance of the influence of sludge retention time (SRT) on the surface properties and
interparticle interactions of sludge flocs and effluent quality was evaiuated by ANOVA, while
the significance of correlations between surface properties and bioflocculation and compaction
properties was tested by Pearson's product-momentum correlation (Mendenhail and Sincich,
1 992; Pagano and Gauvreau, 1993).
3.8.1 Analysis of variance
Analysis of variance (ANOVA) is a method for testing two or more treatments to
detemine whether their sample means could have been obtained from populations with the same
true mean. This is done by estimating the amount of variance within treatments and comparing it
to the variance between treatments. The Type 1 error rate was set at 0.05 for al1 tests performed in
this study. Once differences between means were identified by ANOVA, a least significant
difference test was performed to estimate the means and to determine the magnitude of the
di fferences.
3 -8.2 Correlation
The potential correlation between two variables was evaluated by the Pearson's product
moment correlation coefficient (rJ. Experimental data measured within three days was used for
correlation analyses (Barber and Veenstra, 1986; Jenkins et al., 1993). The statistical significance
of a caiculated correlation coefficient was determined with the t-test. The correlations between
variables were considered to be significant at a 95% confidence level. The relationships were
checked graphically in order to avoid situations where dispersion around the regression line w-as
high.
CHAPTER N
OPERATIONAL PERFORMANCE OF SEQUENCING BATCH REACTORS
The purpose of this study was to investigate the effect of sludge retention time (SRT)
on surface properties and interparticle interactions of sludge flocs under well-controlled
conditions. Accordingly, considerable efforts were made to control the operation of laboratory-
scale sequencing batch reactors (SBRs) to reach the desired SRTs. This chapter describes the
general performance of four paralle1 SBRs and the strategies for controlling a stable SRT.
Parametric evaluations with time were performed in order to provide knowledge of the
physical and physiological statu of siudge in each SBR. To detennine whether stable
conditions were attained at a given SRT, the pedormance of the SBRs was monitored and
evaluated by m e a s u ~ g the mixed liquor suspended solids (MLSS), sludge volume index
(SVI), effluent suspended solids (ESS) and chernical oxygen demand (COD) removal and by
monitoring the carbon substrate oxidation profiles of microbial communities. The average
stable operating conditions of the SBRs are summarized in Table 4.1.
Table 4.1 General Operating Characteristics of the SBR System at Stable Operation
Reactor
SBR-1
SBR-2
SBR-2
SBR-3
SB R-4
- Resi 1 I I I I l
Its are expressed as the average + one standard deviation. No. of observations >3
Target SRT da^)
4
Actual SRT* da^)
Influent COD (ml@)
Average MLSS* (m%L)
4 r 0.5 2250 + 150 540
ESS* (mg/L)
F/M(kg CODkg MLSSId ay)
30 510
Emueat soluble COD* (mg/L)
0.95 50 I 15
4.1 General Description
In Figures 4.1 to 4.3, the MLSS, SV1 and ESS values are plotted against time. These
parmeters varied greatly in the early stage (up to 2 - 3 months) of the experiment. This is
perhaps not surprising, as the inoculum had to adapt to the synthetic wastewater and experience
a stabilization period. It is also interesting to note that the proportion of inorganic solids in the
sludge decreased dramatically with time, from about 35% in the inoculum to about 10% in
sludge after about two months of operation. The large fluctuations in the compressibility (SVT)
and flocculating ability (ESS) of sludge flocs in the early stage are generally in agreement with
the observations of Forster and Dallas-New~on (1980), Lovett et al. (1983) and Sheintuch
(1987) in that a long-term stabilization period was necessary to reach relatively stable
operation. SBR-1 at an SRT of 6 days did not reach stable operation (Figures 4.1 to 4.3) within
80 days. Therefore, no data at an SRT of 6 days is available for comparison at stable operation.
Afier the stabilization period, the operation of the SBRs was more stable, and relatively
small variations in sludge characteristics were observed. On several occasions. sludge in SBR-
1. SBR-2 and SBR-3 experienced modest buking (100 mL/g <SVIQ00 mL/g). Microscopic
examination of the modest bulking sludge samples revealed that filamentous microorganisms
were not noticeable. The absence of visible filamentous microorganisms indicates that the high
SV1 values reflected non-filamentous bulking. On several occasions, the biomass in SBR-I and
SBR-2 flocculated poorly (dispersed growth) with relatively higher ESS values. Both
phenomena (modest non-filamentous bulking and dispersed growth) usually diminished within
two weeks. The self-correcting phenomena made this stuày possible over a long-tenn period,
without restarting the test with new inoculwn.
SDR- I
-SRT= 6 d ays (before day 80), 4 days (after day 80)
S R T = 6 days (before day 30), 12 days (between day 30 and day124), 9 days (after day 124)
- - - ~
Experiiiiciiiiil 'I'iiiie (düys)
Figure 4.1 M ixed liqiior suspeiided solids versiis experiii~eiiiiil t irne
The operating strategy used to maintain a similar MLSS level at each SRT was to change
the influent COD, as shown in Table 4.1. The reason for operating the SBRs with a similar
MLSS level (2000 mg/L), as shown in Table 4.1 and Figure 4.1, was to rninllnize the effect of
particle interactions caused by different biomass concentrations on parametric measwements,
such as flocculating, compacting, and surface properties of sludge flocs. In practice, it is well
known that the MLSS level has a significant impact on the compressibility of sludge (Lovett et
al-, 1983).
From a long-tem study point of view, the SBR system was well operated at a stable
operating cycle, although occasional bulking and dispersed growth in SBR-1, SBR-2 and SBR-3
occurred. A stable condition was assumed to be reached when the fluctuations of the relevant
parameters (MLSS, SVI. ESS, COD and carbon substrate oxidation profiles of the microbial
cornrnunities) were at a minimum (k 15%) within two weeks, and no bulking or no dispersed
growth occurred.
4.2 Sludge Volume Index and Emuent Suspended Solids
SZudae Volume Index: Figure 4.2 and Table 4.1 show the influence of SRT on the compressibility
of sludge as measured using the SVI. There was a statistically significant difference in the SV1
between SRTs in the range of 4-9 days and those in the range of 12-16-20 days (ANOVA, p<
0.05). Relatively larger SV1 values (poorer compaction) and a higher fkequency of non-
filamentous bulking situations were observed at an SRT of 4 days. From an engineering point of
view, however, the differences in the SV1 at different SRTs were not large enough to show any
significance, as al! sludge flocs generally settled well at different SRTs. Taken together with the
78
contradictory findings of previous studies (Ford and Eckenfelder, 1967; Bisogni and Lawrence,
197 1 ; Chao and Keinath, 1979; Lovett et al., 1983; Andradekas. 1993), this suggests that it is
difficult to establish a valid relationship between the SRT and the SV1 for general applications.
Effcluent Sus~ended S O U : The effect of SRT on the level of effluent suspended solids (ESS)
is shown in Figure 4.3 and Table 4.1. The statistical analysis indicates that there was a significant
difference in the level of ESS at different SRTs. A much higher concentration of ESS existed in
the treated effluent at lower SRTs (4, 9 and 12 days) than that at higher SRTs (16 and 20 days)
(ANOVA. p <0.05). Dispersed jgowth was also occasionally observed at lower SRTs (4 and 9
days). These results are generally in agreement with previously reported results (Bisogni and
Lawrence, 1971 ; Chao and Keinath, 1979; Lovett et al., 1983; Wahlberg, 1992; Andreadakis,
1993) in that a lower levei of ESS was associated with a higher SRT. The results from this study
together with literature findings confirm that the flocculating ability of sludge flocs changes with
respect to the SRT, and M e r impIy the possibility of controlling bioflocculation by using a
suitable SRT.
In most situations, the sludge layer at the bottom of the SBR after 40-minute settiing was
well below the effluent port; only small clurnps of microorganisms and non-settleable fine flocs
were washed out of the SBRs. In this study, the small clumps and non-settleable fine flocs were
classified as pin-point flocs. A cornparison of particle sizes of the non-settleable fine flocs and
well flocculated flocs (Figure 4.4) indicates that the majority of non-settleable fine flocs had a
floc size less than 10pm, while the well-fiocculated flocs usually had a floc size larger than
100pm.
Figure 4.1 ~1icroscopic picnires of (a) non-sedable fine flocs (about 3-8 um) and
cb) settleable large flocs (about t O O p m ) . The mapification is 400X.
4.3 Effluent Chemical Oxygen Demand
Figure 4.5 shows the decrease in effluent chernical oxygen demand (COD) with reaction
time in one cycle. Most of the consumable COD was removed within either a one-hour period for
higher SRTs (9.16 and 20 days) or a two-hour period for lower SRT (4 days). The COD removal
efficiency was about 85 to 90%. As glucose is an easily biodegradable compound and should be
totally removed within three hours, the presence of 20 to 50 mgL COD in the treated effluent
strongly suggests that compounds other than glucose were responsible for the COD in the treated
effluent. These compounds could be from ce11 lysis or soluble microbial products (Saunders and
Dick, 198 1).
e S R T = 4 days
4 S R T = 9 days
-SRT= 16 days
-SRT= 20 days
O 50 100 150 200
Reaction Time in one Cycle (minute)
Figure 4.5 Clianges in effluent COD with respect to reaction time in one cycle (samples collected
on May 10,1998)
4.4 Calculated Sludge Retention Time
From Figures 4.1 and 4.3 and Equation (3-l), it is possible to determine the actual SRT in
each SBR over time by taking into account the amount of wasted sludge and discharged ESS per
day. The actual SRT in each SBR over tirne is illustrated in Figure 4.6. These results indicate that
the SRT in each SBR was well controlled during the experimental period. As shown in Table 4.1,
the actuaI SRT values were quite close to the expected SRT values. A more stable SRT was
obtained at lower SRTs (4 and 9 days) than that at higher SRTs (1 6 and 20 days), as shown in
Figure 4.6. This is because a small change in the ESS or the amount of sludge wasted per day
would have a more significant impact on the calculated SRT (Equation 3.1) at higher SRTs than
at lower SRTs.
O 200 400 600 800
Experimental Time (days)
Figure 4.6 Calculated SRT values with respect to time (-O-SBR-1; -c-SBR-2;
-&-SBR-3; -0-SBR-4)
4.5 Microbiology
Plzenotvpic Finner~rintina of the Micro bial Comrnunity : Carbon substrate oxidation pro fi les of
the microbial community in each SB R were monitored iising phenotypic frngerprinting. Figure
4.7 shows the plot of two principal compnents. These results indicate that carbon substrate
oxidation profiles of the microbiai community in each SBR clustered together in a narrow area
(Figure 4.7), implying that the microbial community was functionally similar at different SRTs,
and relatively stable over time. It is interesting to note that carbon substrate oxidation profiles of
the microbial community at lower SRTs (4, 6 days) were slightly more variable, as reflected by
the greater scattering of data show in Figure 4.7. This is perhaps not surprising, as the rapid
growth rate of sludge flocs at lower SRTs could Iead to large population fluctuations, as observed
by Pere el a1.(1993).
YS-B-R:1(S1RX6 da y s ) SBR-2 (SRT=9, 12 days) SBR-3 (SRT=16 days) S B R 4 (SRT=20 days) O , Inoculum sludge -
O 0.2 0.4 0.6 0.8 1
Principal Component 1
Figure 4.7 Principal cornponent anaiysis of Biologm data fiom the SBR system
(data was collected fiom Feb. 1997 to Dec. 1998)
Ouantifkation of Filaments: In addition to the monitoring of carbon substrate oxidation profiles
of the microbial community using phenotypic fingerprïnting, sludge samples were also
fiequently examined using a light microscope to reveaI the presence of any unusuai
microorganisms. During the experimental period, the quautity of filamentous microorganisms in
any of the four SBRs was always in the range of level O to 1 (a smaller number is associated with
less filaments), according to Jenkins' ciassification (levels 0-6) (Jenkins et al., 1993). The control
of filamentous bulking was important as the main purpose of this study was to understand the
causes. of non-filamentous bulkïng problems, such as disintegration and pin-point flocs, which
involve diRerent mechanisrns than filamentous bulking. Strategies for preventing the potential
overgrowth of fitamentous microorganisms are listed below:
1) Minimize the number of filamentous microorganisms fiom the activated sludge seed.
The inoculum fiom an activated sludge p h t treating municipal wastewaters (Main Treatment
Plant. City of Toronto) contained only a few filamentous microorganisms, which were classified
in the 0-1 level of Jenkins' classification (Jenkins et al., 1993).
3) Create a substrate concentration gradient to suppress the growth of fitamentous
microorganisms in the bioreactor (Jenkins et al., 1993; Wanner, 1994a and 1994b). It is well
known that the plug-flow reactor (substrate concentration gradient dong the length of the
reactor) has fewer settling problems than the cornpletely mixed flow bioreactor in activated
sludge plants. The SBRs are operated in a cyclic manner and therefore create a substrate
concentration gradient in each cycle.
84
3 ) Prevent a very low FM (<O. 1 ) at higher SRTs. A very low FM ratio could favor the
overgrowth of slowly growing filaments (Jenkins et al., 1993; Wanner, 1994a and 1994b). The
minimum FM, which is related to an SRT of about 20 days, was about 0.25 in this study.
4) Provide enough DO concentration in the aeration tank (Jenkins et al., 1993). It is
generally believed that a DO concentration larger than 2 ppm is necessary to prevent filamentous
bulking. The DO concentration in this study was maintained within a range of 2.5- 5.5ppm.
4.6 Summary
The time needed to adapt to a new environment and to reach stable operation \vas up to
two or three months, and a long-term expriment is therefore essential to obtain rneaningful
results at a truly stable situation. Unfortunately, this requirement was usually ignored in previous
studies investigating the influence of operating and environmental conditions on sludge
properties. From an engineering point of view, the operating characteristics, as shown in Figures
4.1 to 4.6 and Table 4.1, suggest that SBRs were well controlled during the experimental period.
The results reflect the influence of SRT on the properties of floc-forming microorganisms in
sludge with a functionally similar microbial community over two years. A detailed description
and discussion of the surface properties and interparticle interactions of sludge flocs at different
SRTs and their role in bioflocculation, compaction, and stability are the key points of this thesis
and are the subjects of the following chapters (Chapters V. VI and VII).
CHAPTER V
EXTRACELLULAR POLYMERIC SUBSTANCES
This chapter presents expenmental resuits concerning the production and composition of
extracellular polymeric substances (EPS) at different sludge retention times (SRTs). The results are
divided into four sections in the followùig sequence: influence of SRT on the production and
composition of EPS; characterization of EPS components by ultrafiltration; the sources and
proposed mechanism of the production and composition of EPS; and the role of EPS in
bioflocculation and compaction.
5.1 Influence of Sludge Retention Time on the Production and Composition of Extracellular
Polymeric Substances
EPS Constiruenrs: Figure 5.1 shows the concentrations of EPS components (carbohydrates,
proteins. DNA) with respect to the SRT under stable operational conditions. Similar to previous
studies (Urbain et al., 1993; Frdund et al., 1994 and 1996; Bura et al., 1998), protein was the
dominant component found in the EPS, followed by carbohydrate and a small proportion of DNA
at higher SRTs (>9 days). The carbohydrate content was significantly greater (ANOVA, p c 0.05)
and the protein content significantly lower (ANOVA, p< 0.05) at lower SRTs (4 and 9 days) than
those at higher SRTs (>9 days). There were, however, no significant differences (ANOVA. p>
0.05) in the protein content of the EPS for sludge either fiom the region of lower SRTs (4 and 9
days) or fiom the region of higher SRTs (12,16 and 20 days). The total carbohydrate content
decreased with an increase in the SRT fiom 4 days to 9 days, and then plateaued between 12 days
and 20 days.
86
The ratio of proteins to carbohydrates increased as the SRT increased fiom 4 days (1 -5 2
0.9) to 1 2 days (5.1 I 1 3, and then leveled out (4.5 r 1.9) at SRTs of 16 and 20 days (Figure
5.1 b). The change in the ratio of proteins to carbohydrates in the EPS is similar to the findings of
Jahn and Nielsen (1 996) and Famgia (1999), in which the relative protein content in the EPS
increased with increasing age of biofilrns. This M e r implies potential changes in the
physicochemical properties of sludge surfaces at different SRTs, as the hydrophobicity and surface
charge of sludge are believed to be related to the chernical composition and physical structure of
EPS.
Many studies indicate that acidic polysaccharides are the main components of EPS in the
bacteria (Kenne and Lindberg, 1983; Figueroa and Silverstein, 1989); acidic polysaccharides in the
EPS. however. were detected (0.5 to 1.5 mg/g VSS) only in the early stages of the experirnent.
These levels were significantly lower than the levels of acidic polysaccharides found in the
inoculum (4.8 mg/g VSS) and in sludge fiom other full-scale activated sludge systems (Urbain et
al.. 1993: Fralund et al., 1994 and 1996, Forster, 1996; Bura et al., 1998). Under stable operating
conditions of the SBRs, acidic polysaccharides were not detectable in the EPS of al1 sludge. The
absence of acidic polysaccharides in the EPS is similar to the findings of Jahn et al. (1997)
invoIving a chemostat fed with a synthetic wastewater. This may suggest that the synthetic feed
may not have been conducive to the biosynthesis of acidic polysaccharides or that the feed selected
against the major acidic polymer producers as observed in the inoculurn.
Although Forster and Dallas-Newton (1980) suggested that acidic polysaccharides were the
main ionogenic material that maintains the colloidal stability of sludge flocs, it is likely that the
contribution of acidic polysaccharides to salt bndging was minor in this study, owing to the non-
87
detectable level of acidic polysaccharides in the EPS. This M e r suggests that functional groups
(carboxyl. hydroxyl. phosphate and amino groups) fiom other EPS components, such as proteins
and DNA. could be more important in coctrolling bioflocculation and colloidal stability of sludge
flocs in the present research.
The DNA content in the EPS at different SRTs was highly variable (a high ratio of standard
deviation to average value called the coefficient of variation: 0.3-0.6), as shown in Figure 5. la
DNA constituted about 245% of the EPS content in the sludge samples studied. There were,
however. no signifiant differences in the DNA content with respect to the SRT (ANOVA,
p>0.05). The accumulation of DNA on sludge surfaces suggests that DNA contributed to the
colloidal stability of sludge flocs (Urbain et al., 1993; Palmgren and Nielsen, 1997). hdeed, the
importance of DNA in bioflocculation was observed by Sakka and Takahashi (1 982), who found
tliat the natural flocculation of bacteria involved the specific binding activity of DNA on ce11
surfaces and the accumulation of long chah DNA in the aqueous medium acted as a natural
flocculant.
S f a b i l i ~ o fEPS Production: A relatively larger fluctuation (a larger coefficient of variation: 0.4 -
0.55) of EPS concentration, especially the two dominant components (proteins and carbohydrates),
was associated with lower SRTs (4 and 9 days), as compared to that at higher SRTs (12, 16 and 20
days) (Figure 5.la). These results are consistent with the literature findings of Gulas et al. (1979)
and Pere et QI. (1993), who also observed a relatively higher variability in the EPS production at
lower SRTs. The variation in EPS production observed in the present study could be related to the
relative stability of the functional abilities of the microbial community. As illustrated in Figure 4.7,
the slightly higher variability in carbon substrate oxidation profiles of the microbial comrnunity at
lower SRTs (4 and 6 days) could partially explain the fluctuations in the EPS.
Total EPS: The total amount of EPS was independent of the SRT (Figure 5. l b) (ANOVA p>
0.05). The results obtained ftom this study, together with the findings of Kiff (1978) and Gulas et
al. (1 979), who also used continuous reactors, indicate that the production of EPS was not limited
to the stationary and endogenous phases of sludge associated with very high SRTs (very slow
growth rates). A large amount of EPS was also extracted from sludge at lower SRTs (rapid growth
rates) in these shidies. These results are not consistent with sorne of the literature fmdings fiom
batch reactors (Pavoni et al., 1972; Chao and Keinath, 1979; Sheintuch et al., 1986; Sheintuch,
1987). It has been thought that a larger arnount of EPS would be produced in the starving or
endogenous phases of sludge than in the exponential growth phase (Pavoni et al., 1972; Chao and
Keinath, 1979; Sheintuch et al., 1986; Sheintuch, 1987). However, in the present long-terrn study
the most significant change in the EPS with respect to the SRT was the ratio of proteins to
carbohydrates. but not the total EPS content. The results from this study strongly suggest that the
pattern of EPS production fiom batch reactors can not be referred, in a simple way, to those
involving the use of continuous or semi-continuous reactors.
4 9 12 16 20
S RT (day s)
Figure 5.la Effect of SRT on the production of EPS components under stable operating conditions. Results are expressed as the average I one standard deviation. No. of observations is shown in Appendix E-1.
- a Total E -PSxRat io f proteins to carbohydrates
-
9 12 SRT (days)
Figure 5.lb Effect of SRT on totai EPS content and ratio of EPS components under stable operating conditions. Results are expressed as the average 5 one standard deviation. No. of observations is shown in Appendix E-1.
5.2 Molecular Weight Distribution of Extrncellular Potymeric Substances
Ultrafiltration was used to characterize the molecular weight distribution (MWD) of EPS
components. In early experiments, it was found that polar interactions between the EPS
components and the membrane surface were significant. As mwh as 40 to 60% of the EPS content
could be adsorbed on the membrane surfaces or pore surfaces, and the recovery efficiency of each
EPS component was as low as 50% without washings in each separation step. For each membrane,
a tsvo- or three-step washing with deionized distilled water significantly improved the recovery
efficiency of the EPS components by up to 85 to 95%. All results reported here are based on a
minimum of at least a 75% mass recovery.
As s h o w in Figure 5.2, the MWDs of al1 EPS components covered a broad spectrum, fiom
less than 1,000 daltons to more than 100,000 daltons, but a significant portion (8590%) of al1 EPS
components had molecular weights larger than 10,000 daltons. These results are generdly
consistent with the literanire findings (Forster, 1976; Goodwin and Forster, 1985) with sludge
samples from Full-scale activated sludge systems. It is interesting to note the presence of a small
portion of EPS components having molecular weights less than 1.000 ddtons; EPS with such a
small molecular weight fraction might suggest that simple sugars, disaccharides, fiee amino acids,
and fragments of DNA are present in the EPS. Although Goodwin and Forster (1985) found that a
large proportion (40%) of EPS components had molecular weights less than 500 daltons, in this
study the proportion of each EPS component with molecular weights less than 1,000 daltons was
smaller (2-8%). A direct comparison may be difficult due to different extraction methods. But
thermal extraction used in the earlier study might decompose EPS to smaller molecules. It appears
that M WDs of EPS components are related to operating and environmental conditions.
cl O00 1 O00 - 10.000 10,000 - 100,000 '1 00.000
Molecular Weig ht Fraction (dalton)
Figure 5.2 Molecular Weight Distributions of EPS cornponents (U- SRT=4 days; 4 - SRT=9days-
-1- SRT=l6 days; - 0 - SRT=20 days). Results are expressed as the average 2 one
standard deviation. No. of observations is sumrnarized in Appendix E-2.
92
Statistical analyses indicate that there were no significant differences in the MWDs of EPS
components in terms of the SRT. A cornparison of the MWD of EPS components at different SRTs
might be dificult owing to the nature of the ultrafiltration analysis (semiquantitative). It appearç
that either the MWDs of EPS fiom different SRTs were similar or a cutoff of the MWD with the
ultrafiltration technique could not distinguish the potential differences in the MWDs of EPS
components at different SRTs. A combination of different techniques, including electrophoresis,
gel chromatography, and ultrafiltration, may be able to provide more comprehensive information
about the MWDs of EPS components at different SRTs. For example, ultrafiltration can provide a
range of molecular weight fractions with a volume sufficient for further biochemical analyses,
while gel chromatography enables the observation of a continuous distribution of mûlecufar
weiglits (Fralund et al-,1994). Further characterization of the MWDs of EPS components should
combine different techniques.
5.3 Proposed Mechanism and Sources of the Production and Composition of Extracellular
Polymeric Substances
. The change in the ratio o f proteins to carbohydrates with respect to the SRT could be
related to differences in both the growth rate and the microbial community at different SRTs. The
total carbohydrate content of whole sludge was generaily higher (>200 mglg MLSS) at an SRT of
4 days (Figure 5.3) and decreased to 150 mglg MLSS or less at an SRT of 20 days. It is likely that
sludge at lower SRTs (higher organic loadings) did not consume the carbon source available for
growth. Excess carbon substrates could have been converted into intracellular storage granules and
exopolysaccharides that attached as EPS (Harris and Mitchell, 1973 and 1975; Dugan, 1987;
Andreadakis, 1993). At higher SRTs, carbon source (COD) became limiting for sludge growth, and
93
the level of storage polymers and exopolysaccharise declined. This explains the decrease in the
total carbohydrates in the EPS with respect to an increase in the SRT.
O 5 10 15 20 25
SRT (days)
Figure 5.3 Effect of SRT on the total carbohydrate content of whole sludge (Pearson's
coefficient r,= -0.85, p<O.05) (Data was collected fiom Jun 1 997 to Dec. 1 997)
The presence of DNA in the EPS strongly suggests that naturally occurring ce11 lysis within
the floc assemblage in the SBRs contributed to the composition of EPS. A positive correlation
between the protein content and the DNA content in the EPS (Figure 5.4) fiirther indicates that ce11
lysis in the SBRs was also one of the mechanisms contnbuting to the presence of proteins in the
EPS. The ratio of proteins to carbohydrates (0.5-5) in the EPS, however, was quite sirnilar to the
ratio of proteins to carbohydrates (3-4) generally found on ce11 surfaces (Fralund et al.. 1994),
implying that both secretion as well as the accumulation of ce11 lysis could contribute to the protein
content in the EPS. The relative importance of protein accumulation due to secretion and ce11 lysis
of microbial cells is not known. It is clear, at least, that a comparatively Iarger amount of ce11 lysis
occurred at higher SRTs, due to lower microbial growth rates and higher levels of endogenous
94
metabolism (Gulas et al., 1979). It is likely that the increase in the relative protein content with
respect to the SRT was, in some way, related to the accumulation of proteins fiom cell iysis in the
aeration tank.
O 5 10 15 20 25 30
Proteins (mg/g VSS)
3 . --
Figure 5.4 Correlation between DNA and proteins in the EPS (Pearson's coefficient r, = 0.65, p
<0.05)
2.5
Another factor that might have contributed to the change in the ratio of proteins to
carbohydrates in the EPS was the potential differences in the microbial cornmunity at different
O SRTa days a S R T z 6 w - ASRT=9days ,SRT=l2days
- ,SRT= 16days .SRT=20days
SRTs. as EPS components can Vary according to microorganisms (Costerton et al., 1934).
However. the carbon substrate oxidation profiles (Figure 4.7) indicate that microbial cornmunities
were functionally similar (sharing the same area of the PCA plot) at different SRTs, when the
SBRs were operated under normal conditions. This may suggest that the change in the ratio of
proteins to carbohydrates in the EPS was caused mainly by changes in the growth rate or
physioloçical status of fbnctionaily similar microbial communities at different SRTs.
The biosorption of macromolecules on sludge surfaces, such as cellulose and hurnic
substances fiom wastewaters, is important in municipal and industrial wastewater treatrnents
(Forster and Dallas-Newton, 1980; Eriksson and A h , 1991). Its possible contribution to the
arnount and composition of EPS was eliminated in this study by using a synthetic glucose-based
wastewater. The only potential component contributed to biosorption was glucose in the feed.
However, ultrafiltration of EPS components demonstrated that the proportion of carbohydrates in
the EPS with molecular weights Iess than 1,000 daltons was only about 2 to 4% of the total
carbohydrates in the EPS. This result suggests that the contribution of biosorption of glucose to the
EPS content was small, and the majority of carbohydrates in the EPS originated fiom rnicrobid
cells.
5.4 Influence of Extracellular Polymeric Substances on Flocculatiog Ability and
Compressibility
EPS have long been believed to be important in controlling the flocculating ability (ESS)
and compressibility (SVI) of sludge (Forster, 1985; Eriksson and A h , 1991; Frdund et al., 1996;
Bura el al., 1998; Higgins and Novak, 1998). However, there is only little direct evidence to
support this hypothesis. The relative importance of different EPS components has not k e n
established. An in-depth investigation on the influence of EPS components on the flocculating
ability and compressibility of sludge was conducted in this research; and the results are shown in
Table 5.1. It appears that the quantity of EPS components had a different role in controlling the
flocculating ability and compressibility o r sludge. The quantity of EPS components played a more
important role in goveming the compressibility of sludge than in controlling bioflocculation.
Table 5.1 Pearson's Coefficient of the Linear Correlation Between the SV1 or ESS and the Quaatity
of EPS Components'
Pearson's Coefficient (rd'
EPS components SVI ESS
Total carbohydrate
Protein
DNA
Total EPS
I No. of observations: 98; * Marked coefficients are significant at a 95% confidence level.
Cornpressibilitv: The compressibility of sludge was evduated by the sludge volume index
(SVI). A higher SV1 is related to a poorer compressibility. As shown in Table 5.1, no significant
correIation was found in this study between the total carbohydrate in the EPS and the SVI. This is
consistent with the findings o f Chao and Keinath (1979) in that the arnount of extracted
carbohydrates was not related to the SVI. However. the protein and DNA contents showed a
moderate correlation (rp=0.53, 0.5, respectively) to the SVI, implying their negative influence on
the SVI. It appears that different EPS components may play different roles in governing the
compressibility of sludge flocs. At present, an appropriate hypothesis has not been fomulated to
explain the diEerent roles of EPS components, but this may related to the physical and chernical
properties of specific EPS molecules.
Arnong the measured EPS components, the totai EPS content showed the strongest
correlation (rp=0.64) with the SVI, as shown in Table 5.1. A higher SV1 was associated with an
increase in the totai EPS content (Figure 5.9, suggesting the totai EPS content had a negative
97
effect on the compressibility of sludge. However, in previous studies the EPS content was related
in inconsistent ways to the SVI. For example, Urbain et al. (1993) found that the amount of EPS
was proportional to the SV1 using sludge samples from Ml-scale activated sludge systems. On the
contrary, Goodwin and Forster (1985) observed an opposite result with sludge samples fiom full-
scale activated sludge systems, i.e., a higher SV1 was related to a srnaller amount of EPS. These
contradictory results from earlier studies were not surptising because a number of possibly
dominant factors affecting the compressibility of sludge c m coexist in full-scale activated sludge
systems and the sample size used in the earlier studies was small (3 - 16 samples). Consequently,
the importance of the correlations fiom earlier studies may be limited. The results from this study
cover a long-term period and are supported by extensive data (Figure 5 .9 , which suggests that the
presence of a large amount of EPS does have a negative effect on the compressibility of sludge.
The reason a larger arnount of EPS produces a higher SV1 c m be explained by the steric force
arising fiom the EPS. The EPS molecules extend out from the ce11 surfaces and therefore physically
prevent the cells from coming into close contact. The EPS also form a dense gel which resists the
expression of water from gel pores. Ultrastructural studies employing correlative microscopy
reveaied a complex and hydrated EPS within the floc matrix, indicating a larger capacity to retain
water (Liss et al., 1996).
T S R F 4 d a y s &RT=6 days A SRT=S days 0 SRT=12 days , SRT=16 days x x
SRT=20 days O 8
- --- - O
O. O 5.0 10.0 15.0 20.0 25.0 30.0 35.0
Total EPS (mgig VSS)
Figure 5.5 Relationship between total EPS content and sludge volume index (Pearson's
coefficient rp=0.64, p ~ 0 . 0 5 )
Bioflocculation: As s h o w in Table 5.1, there was only a very weak correlation between the total
carbohydrate content and the level of ESS. However, the contents of protein, DNA and total EPS
had no significant correlation with the level of ESS. These results suggest that the use of the
quantity of EPS components as a measuring parameter has lirnited value in understanding
bioflocculation.
Traditionally, polymer bridging has been considered to be the most popular mechanism of
microbial floc formation, and the quantity of EPS is of signifiant importance in the polymer
bridging mechanism. Previous studies showed that turbidity was related to the amount of
extractable EPS. For exarnple, Pavoni el al. (1972) found a direct correlation between the EPS
content and the flocculating ability of sludge in batch reactors. Chao and Keinath (1979) also
obsewed a direct correlation between the level of ESS and the amount of extractable carbohydrates
99
in a chemostat study. In contrast, Kiff (1 978) and Gulas et al. (1 979) observed that the presence of
a larger number of pin-point flocs was associated with greater EPS production at lower SRTs.
These contradictory results suggest that the EPS content can not be correlated to the flocculating
ability of sludge in a simple way, and m e r fundamentai knowledge of EPS is still necessary for a
better understanding of the role of EPS in biofloccuiation.
The lack of a significant correlation between the quantity of EPS and the ESS, as shown in
TabIe 5.1. is not unexpected, considering that bioflocculation depends on not only the quantity but
also on the type of biopolymer. Harris and Mitchell (1975) found that some types of biopolyrners,
suc11 as dextran, secreted fiom certain bacteria, indeed, inhibited bioflocculation, implying the
importance of biopotymer composition in microbial floc formation. It is probably the
physicochemical properties of EPS, rather than the quantity of EPS, that govern bioflocculation.
Thess specific physicochemical properties of EPS will be a function of the specific kinds of
proteins and carbohydrate polymers in the floc matrix; relating specific kinds of EPS molecuIes to
floc cliaracteristics, however. is a field in its idancy.
5.5 Relationship Between Sludge Volume Index and Effiuent Suspended Solids
As shown in Figure 5.6, there was no correlation between the SV1 and the ESS. This result
suggests that the compressibility and flocculating ability are two different phenomena related to
biomass separation problems. The level of ESS is a measure of the degree of flocculation of sludge,
while the SV1 is actually a reflection of the compaction of sludge. Unfortunately, these two
separation problems have long been intertwined. The results shown in Figure 5.6 indicate that the
utilization of good or poor "flocculation" can not simultaneously explain the coexistence of a
1 O0
higher levei of ESS and a lower SV1 and/or a Iower Ievel of ESS and a higher SVI. Therefore, it is
important to distin-pish these two separation problems. An indepth investigation may suggest
different causes and develop different strategies for controlling them.
x SRT=t 6 days 0 SRT=20 days
60
O 50 3 0 0 1 50 2 0 250
Svi (mUg MLSS)
50
Figure 5.6 Relationship between the SV1 and ESS @ >0.05)
O SRT=4 days O n SRT=6 days
A SRT=Qdays O SRT=? 2 days
5.6 Summary
Based on long-term experiments using laboratory-scale SBRs. the following conclusions
are made related to the production, composition, and molecular weights of EPS and their role in
bioflocculation and settling:
1 ) The production and composition of EPS were affected by the SRT. Protein and carbohydrate
were the dominant components, followed by a small amount of DNA, in the EPS. Acidic
polysaccharides were not detectable in the EPS fkom sludge fed a glucose- and mineral salts-based
synthetic wastewater.
101
2) The total amount of EPS was independent of the SRT. The ratio of proteins to carbohydrates,
however. increased with an increase in the SRT fiom 4 to 12 days and then reached an almost
constant value for SRTs from 12 to 20 days. These results suggest that the EPS were heterogeneous
and the proportion of EPS components changed with respect to the SRT.
3) The molecular weight distributions (MWDs) of EPS components (carbohydrates, proteins and
DN.4) covered a broad spectrum, ranging from less than 1,000 daltons to more than 100,000
daltons. A significant portion ( 3 5 % ) of EPS components had molecular weights larger than
10.000 daltons. Based on the results fiom ultrafiltration, no significant differences in the pattern of
the MWDs of EPS components were observed in terms of the SRT.
4) A higher SV1 (poorer compressibility) was strongly associated with a larger amount of total
EPS. This result suggests that the presence of a large amount of total EPS has a negative effect on
the cornpressibili~ of sludge.
5 ) The poor correiation or absence of a correlation between the EPS content and the level of ESS
suggests that the utilization of the quantities of EPS components as a parameter contributed little to
an understanding of bioflocculation. This may imply that other properties, such as physicochemical
properties of specific EPS molecules, may be more important than the quantity of EPS in goveming
bioflocculation.
CHAPTER VI
HYDROPHOBICITY AND SURFACE CHARGE
As shown in Chapter V, the quantity of extracellular polymeric substances (EPS) plays an
important role in controlling the compressibility of sludge flocs, but has limited value for
understanding bioflocculation. In other words, factors that govem the flocculating ability of fine
flocs are still unknown. One hypothesis for explaining the change in the level of effluent
suspended solids (ESS) in terms of the sludge retention tirne (SRT) could be related to the
physicochemical properties of sludge surfaces: such as the hydrophobicity and surface charge of
sludge flocs. This chapter presents experimental results for the hydrophobicity and surface charge
of sludge flocs at different SRTs and the role of hydrophobicity and surface charge in
bioflocculation and compaction. The subsections of this chapter are organized as follows:
influence of SRT on hydrophobicity and surface charge; the relationship between hydrophobicity
and surface charge; relationships between hydrophobicity or surface charge and EPS production
and composition; and the role of hydrophobicity and surface charge in bioflocculation and
compaction.
6.1 Influence of Sludge Retention Time on Hydrophobicity and Sudace Charge
Hvdro~l~obicirv: The hydrophobicity of sludge surfaces was evaiuated by the contact angle
measurement (CAM). A more hydrophobic surface was related to a larger water contact angle.
Figure 6.1 shows the results of water contact angles on sludge cakes fiom different SRTs. Sludge
at an SRT of 9 days had the smallest water contact angles versus sludge at other SRTs (4, 12, 16
and 20 days) (ANOVA, p <O.OS). There were no significant differences in water contact angles
between an SRT of 16 and 20 days (ANOVA, p >0.05). However, sludge at higher SRTs (12, 16
and 20 days) had significantly larger water contact angles than sludge at lower SRTs (4 and 9
days) (ANOVA, p<0.05). It appears that a transition in the hydrophobicity occurred in the SRT
range of 9-12 days. The observed change in the hydrophobicity with the SRT is in accordance
with previous results from batch experiments where sludge (data not shown) and Escherichia coli
(Allison et al., 1990) in the stationary phase were more hydropbobic than those in the exponentid
growth phase, since higher SRTs are related to older average physiological ages of the microbial
cornmunity. Pere et al. (1993) also observed that sludge from a pulp and paper effluent treatment
plant was generally more hydrophobic at a lower F/M (or a higher SRT). More recently, Frralund
er al. (1994) found that the EPS fiom a füll-scale activated sludge system operated at an SRT of
36 days were more hydrophobic than those at an SRT of 8 days. These results indicate that the
SRT or the average physiological age is important in determining the hydrophobicity and that a
more hydrophobic surface is related to a higher SRT.
4 9 12 16
SRT (days)
Figure 6.1 Effect of SRT on contact angle values of sludge. Results are expressed as the average
2 one standard deviation. No. of observations is shown in Appendix F-1.
Surface Charaet A number of midies have shown that sludge surfaces are negatively charged
under neutral pH conditions due to the presence of anionic fünctional groups, such as carboxyl,
hydroxyl and phosphate groups on sludge surfaces (Forster, : 908 and 197 1). The influence of pH
on the surface charge is illustrated in Figure 6.2. The change in the surface charge was caused by
variations in the degree of dissociation of weakly ionizable groups on sludge surfaces with respect
to the pH. No significant change in the surface charge was observed in the pH range of 5.6 to 9.3.
At pH values below 4.6, however, the negative surface charge value decreased dramatically. A
positive surface charge was observed for al1 sludge samples in the pH range of 1.8 to 2.6. This
indicates that isoelectric points of sludge flocs Lay between a pH value 2.6 and 3.5, at which point
no cationic polymer would likely bind to the sludge matrix. This is consistent with previous
studies (Forster. 1968 and 1 Wl), which reported the isoelectric points of sludge flocs in the pH
range of 2 to 3.5. The main anionic groups contributing to the surface charge might be the
carboxylic groups from proteins and phosphate groups fiorn DNA in the EPS, because the
contribution of acidic polysaccharides was probably minor as indicated by the non-detectable
levels of acidic polysaccharides in the EPS. The role of amino groups in deterrnining the surface
charge is more complicated; and related to the solution pH. At a low pH value, amino groups (R-
NH?) tend to make the surface more positively charged by absorbing H' to form R-NH,', and this
esplains why the net surface charge of sludge is positive when the pH is less than 2.6.
, SRT=20 days
0.6 -,
Figure 6.2 Influence of pH on the swface charge of sludge at different SRTs
0.4 5 V) >
From a practical viewpoint, it is important to know the magnitude of the surface charge of
SRT=4 days
- e a SRT=9 days
A SRT=16 days
sludge under approximately neutral pH conditions, because fiill-scale activated sludge systems are
usually operated at a pH at or near 7. The effect of SRT on the surface charge of sludge under
neutral pH conditions is shown in Figure 6.3. A similar surface charge value was observed at
SRTs of 4 and 9 days or 16 and 20 days (ANOVA, p >0.05), but a significantly less negative
charge value was observed at higher SRTs (16 and 20 days) than at lower SRTs (4 and 9 days)
(ANOVA, ~ 4 . 0 5 ) . This intermediate SRT range for changes in the surface charge is consistent
with that for changes in the proportions of EPS components (Figure 5. lb) and hydrophobicity
(Figure 6.1).
1 O6
The relatively constant surface charge in SRTs of 4 to 9 days is consistent with the
literature findings (Chao and Keinath, 1979), in which the zeta potential did not change
significantly with respect to the SRTs (1.1 to 1 1.5 days). Udortunately, the SRTs investigated in
the earlier study did not cover the range of SRTs larger than 12 days. The significant decrease in
the surface charge at SRTs of 16 and 20 days is aiso consistent with the observation of Pere et al.
(1993) in that sludge from a pulp and paper mil1 treatment plant at higher FM ratios (Le., lower
SRTs) was more negatively charged than at lower FM ratios. The results fiom the present and
earlier (Pere et al., 1993) study indicate that experimental investigation must cover a broad range
of SRTs to understand potentiai changes in the surface charge in different SRT ranges.
SRT (days)
Figure 6.3 Effect of SRT on the surface charge of sludge. Results are expressed as
the average value I one standard deviation. No. o f observations is shown
in Appendix G- 1.
6.2 Relationship Between Hydrophobicity and Surface Charge
A cornparison of the water contact angle and surface charge (Figure 6.4a) indicates that
there was a strong inverse correlation between the surface charge and hydrophobicity. This is
because the surface charge is related to the ionizable groups on sludge surfaces; it increases the
polar interactions of EPS with water molecules. Therefore, the more charged the sludge surfaces,
the lower the hydrophobicity of the sludge surfaces (i-e.. the lower the contact angle value).
Moreover, a careful examination of the experirnentai data published by Pere et al. (1 993) revealed
that an inverse correlation existed between the zeta potential and the water contact angle in full-
scale sludge sarnples (Figure 6.4b). The correlation, however, was weaker because other possible
factors that affect the properties of sludge flocs may coexist in full-scaie activated sludge systems
(Barber and Veenstra, 1986). The correlation between the water contact angle and surface charge
as shown in Figure 6.4 suggests that the surface charge plays an important role in the measured
hydrophobicity, and the measurement of surface charge by simple colloid titration may be a useful
method for indirectly evaluating the relative hydrophobicity.
SRT=Gdays A SRT=9 days
SRT=l2 days
, SRT=16 days x 0 &RT=20 days O
O
O 10 20 30 40 50
Contact Angle (degrees)
Figure 6.4a Relationship between surface charge and water contact angle
coefficient r,, =0.85, p ~ 0 . 0 5 )
0 F/M= 3.3, controlled 0 F/M=3.3, oxidative conditioning O
I)
A F/M=2.6, controlled O O F/M=2.6, oxidative conditioning X FIM=l.2, controlled 9 OF/M=l.2 , oxidative conditioning + O t x e
+ F/M=0.8, controlled al - F/M=0.8, oxidative conditioning - ----
O X e
10 15 20 25 30 35 40
Contact Angle (degrees)
(Pearson' s
Figure 6.4b Relationship between zeta potential and water contact angle (Pearson's coefficient
r, = 0.62, p <0.05, data denved fiom Pere et al., 1993)
6.3 Influence of ExtraceIlular Polymeric Substances on Hydrophobicity and Surface Charge
Due to the complex composition and structure of sludge surfaces, molecular determinants
for the hydrophobicity and surface charge of siudge flocs have not yet been well established.
However. it has been hypothesized that the hydrophobicity and surface charge are related to the
production, composition and physical configuration of EPS. A statisticai analysis of linear
correlations between the water contact angle or surface charge and EPS composition is presented
in Table 6.1. It appears that the proportions of EPS cornponents @roteindcarbohydrates andor
proteinsl(carbohydrates + DNA)) were more important than the quantities of individual EPS
components in controlling the hydrophobicity and surface charge. Different EPS components may
have different roles, depending on the physicai and chemical properties of EPS. For instance, the
amount of total carbohydrates had a negative influence on the hydrophobicity and surface charge
(rp= -0.50. -0.38' respectively). In contrast, the protein content had a positive influence on the
surface charge (rp=0.43). On the other hand. the total EPS content was not related to the
hydrophobicity and surface charge.
Table 6.1 Pearson's Coefficients of Linear Correlation Between Clydrophobicity or Surface Charge and EPS Composition
Pearson's Coefficient (r,,))'
EPS Components Contact Angle
(32)'
SurCace Charge
(3 8)'
Total carbohydrates -0.50*
Proteins 0.27
DNA
Total EPS content
Proteins~Total carbohydrates OS8*
Proteins/(Total carbohydrates 0.52*
+ DNA)
* ' Marked Pearson's coefficients (r,) are significant at a 95% confidence level.
Bracketed figures indicate the nurnber of observations.
It is reasonable that the change in the hydrophobicity of sludge flocs with the SRT was
affected by variations in the protein, total carbohydrate, and DNA content in the EPS. The
positive role of proteins in the EPS on the hydrophobicity is supported by a recent study by
Jorand el al. (1998), who found that the hydrophobic fraction of EPS was made up only of
proteins and not carbohydrates. ïheir findings suggest that amino acids with hydrophobic side
groups play a dominant role in determining the hydrophobicity. Furthemore, the presence of a
large amount of carbohydrates at lower SRTs (4 and 9 days) and the accumulation of DNA on
sludge surfaces could aiso be expected to decrease the hydrophobicity, due to the hydrophilic
nature of carbohydrates (mainly neutml carbohydrates in this study) and the polar interactions
between phosphate groups in DIVA and water molecules. Consequently, the hydrophobicity
expressed by the CAM reflected the presence of both hydrophobic and hydrophilic groups in the
EPS. and may depend on the relative density of hydrophobic and hydrophilic sites (proteinshotal
carbohydrates) on sludge surfaces. This explanation is supponed by the fuidings of Daffonchio ef
al. (1995), who observed that the hydrophobicity of a mixed culture measured by the water
contact angle was apparently an average of the hydrophobicity of its components.
The importance of the ratio of proteins to carbohydrates in determining the surface charge
could be related to the unique charge properties of proteins. The amino groups in proteins carry
positive charges. and can neutralize some of the negative charge fiom carboxyl and phosphate
groups and therefore decrease the net negative surface charge of sludge flocs. This result is
consistent with the findings of Morgan el al. (1990) in that the proportions of EPS components
(proteins/totaI carbohydrates) were more important than the quantities of individual EPS
components in determining the surface charge of both anaerobic and aerobic sludge flocs.
Ho~vever, the weak and moderate correlations between the proportions of EPS components and
the hydrophobicity or surface charge may further suggest that some other factors and EPS
components which are still unknown are involved in determining the hydrophobicity and surface
charge.
At present, little is known about relating the hydrophobicity and surface charge to specific
molecuiar deterrninants. A better understanding of the hydrophobici ty and surface charge requires
more fundamental information about the EPS composition and the physical nature of specific EPS
molecules.
6.4 Influence of Hydropbobicity and Surface Charge on Flocculating Ability and
Compressibility
Bioflocculation: The degee of bioflocculation was determined by the level of effluent suspended
solids (ESS). As shown in Figures 6.5 and 6.6, the level of ESS was strongly correlated to the
water contact angle, as well as to the surface charge. The role of hydrophobicity in controlling the
level of ESS is consistent with the findings of Valin and Sutherland (1982) in that a more
hydrophobie surface led to better biofloccuIation. Similar to the hydrophobicity, the correlation
between the surface charge and the level of ESS suggests that electrostatic interactions fiom
sludge surfaces were involved in controlling bioflocculation. A more negatively charged surface
prevents bioflocculation. Al1 the experimental results indicate that bioflocculation strongly
depends on surface interactions arising from sludge flocs. Furthermore, the correlations between
contact angle or surface charge and the level of ESS (Figures 6.5 and 6.6), together with the
results as shown in Table 5.1, demonstrate that it is the physicochemical properties of sludge
surfaces. such as hydrophobicity and surface charge, rather than the quantity of EPS, that control
the flocculating ability of sludge flocs.
--- - O
O SRT=4 days
O SRT=6 days , SRT=9 days
A " O SRT=l2 days
A O , SRT=16 days A
A . SRT=20 days A O
0 0 O O O ex x O
O 10 20 30 40 50
Contact Angle (degrees)
Figure 6.5 Relationship between contact angle and effluent suspended solids (Pearson's
coefficient r,= -0.70. p < 0.05)
-.
a 0 , S-RT=4 days
QRT=B days O A
a , S RT=9 days
A SRT=12 days A A A
Oa A 2~ S RT= 16 days A O
A o n - - SRT=20 days O O O A O O
O -
-CO , O X x
X rn O x
-0.7 -0.6 -0.5 4 . 4 4 . 3 6.2 -0.1 O
Surface Charge (meq./g VSS)
Figure 6.6 Relationship between suface charge and effluent suspended solids
(Pearson's coefficient r,, = -0.74. p c0.05)
From an engineering vieupoint, it is interesting to note that the hydrophobicity and
surface charge varied with respect to the SRT, as shown in Figures 6.1 and 6.3. These fmdings
cIari@ the role of the EPS content in bioflocculation and provide a new scientific explanation for
changes in the flocculating ability of sludge flocs in terms of the SRT. The larger incidence of
pin-point flocs at lower SRTs, as reported in the literature (Bisogni and Lawrence, 1971 ; Chao
and Keinath. 1979; Gulas et al., 1979; Lovett el al., 1983; Wahlberg, 1992; Andreadakis, 1993)
can be explained by the less hydrophobic and more negatively charged properties of sludge
surfaces. but is not caused by a decrease in the EPS production at Iower SRTs. The intemediate
range of SRTs for changes in surface properties (EPS, hydrophobicity and surface charge) found
in this study is close to the design window (usualIy SRT=15 days) of SRT used for the extended
aeration and conventional operation of the activated sludge process, and is consistent with the
suggestions in the literatwe (Chao and Keinath, 1979) for better bioflocculation.
Cumoressibilihr: The compressibility of sludge flocs was evaluated by the sludge volume index
(SVI). Figure 6.7 shows that there was no correlation between the water contact angle and the
SVI. This is consistent with the findings of Urbain et al. (1993), who found that there was no
direct correlation between the internai hydrophobicity and the SV1 in a study of full-scale
activated sludge sampfes. I t appears, therefore, that the hydrophobicity of sludge flocs is not
related to variations in the SVI.
In previous studies, Forster (1968 and 1971) and Steiner et al. (1 976) found that the
electrophoretic mobility of sludge was directly related to the SVI. In contrast, Magara et al.
(1976) demonstrated an inverse correlation between the surface charge and the SVI, and no
correlation was observed by Chao and Keinath (1979) and Rarber and Veenstra (1986). The
results from the present study suggest that there was no correlation between the surface charge
and the SV1 (Figure 6.8). A direct comparison of the results fiom different studies may be
dificult. due to changes in the physical, chernical, and microbiologicat conditions in different
studies. However. the idea that surface charge is related to the compressibility of sludge flocs is
doubtful. According to the DLVO (Derjaguin and Landau, 1941 ; Verwey and Overbeek, 1948)
theory. less charged, fine flocs will favor the formation of large and dense flocs and thus reduce
the SVI. However, the presence of filamentous microorganisms and heterogeneity of floc sizes in
sludge sarnples make it dificult when applying the DLVO theory to explain the change in the
compressibility of sludge flocs, as the DLVO theory is only most important to explain the
behaviour of colloidal particles (less than a few pm).
The results from the present study indicate that both hydrophobicity and surface charge
had different roles in controlling the level of ESS and the SV1 (Figures 6.5 to 6.8). This could be
esplained by the relative importance of interparticle and gravity forces in relationship to particle
sizes. As shown in Figure 4.4, the non-settleable particles (Figure 4.4a) in the effluent
(contributing to ESS) were much smaller in size than the settleable sludge tlocs (Figure 4.4b)
(contributing to the SVI). A direct comparison of different kinds of interaction forces acting on
flocs indicates that the van der Waals and electrostatic interactions dominate over the floc mass
for sizes less than about 10-20 Pm, while the gravitational force is more significant for flocs
larger than 50 p m (Hogg, 1989). This may explain why hydrophobicity and surface charge were
crucial in controlling the level of ESS (fine flocs), but were not important in determining the SV1
of sludge (iarger flocs).
-- S-RTg43ays
ci SRT=6 days X
A SRT= 9 days SRT=12 days SRT=l6 days O
X
SRT=20 days -- O
O 10 20 30 40 50
Contact Angle (degrees)
Figure 6.7 Relationship between water contact angles and sludge volume index ( p >0.05)
O SRT=4 day s SRT=6 days
A SRT=9 days O SRT=12 days x SRT=16 days 0 SRT=20 days
- . - - - - . -
-0.7 -0.6 -0.5 -0.4 -0.3 -0.2 -0.1 O
Surface Charge (meq./g VSS)
Figure 6.8 Relationship between the surface charge and sludge volume index ( p>O.OS)
6.5 SUMMARY
Based on the experimental results fiom a long-terni study in laboratory-scale sequencing
batch reactors (SBRs), the conclusions related to the hydrophobicity and surface charge at
different SRTs and their role in bioflocculation and compaction can be sumrnarized as follows:
1) Physicochemical properties of sludge surfaces, such as hydrophobicity and surface charge,
were affected by the SRT. Sludge surfaces at higher SRTs (16 and 20 days) were less negatively
charged and more hydrophobic (larger water contact angle) than those at lower SRTs (4 and 9
days). A transition in the hydrophobicity and surface charge of sludge flocs occurred in the SRT
range of 9-1 2 days.
2) A strong inverse correlation existed between the water contact angle and the surface charge of
sludge. This suggests that the surface charge measured by simple coltoid titration may be a usefùl
measure of the relative hydrophobicity of sludge surfaces.
3) The strong inverse correlations either between the water contact angle and the level of ESS or
between the surface charge and the level of ESS indicate the importance of surface interactions in
goveming bioflocculation. It is the physicochemical properties (hydrophobicity and surface
charge) of sludge surfaces, rather than the quantity of EPS alone, that govem bioflocculation. The
observed influence of SRT on the hydrophobicity and surface charge suggests that it is possible to
use the SRT to alter the swface nature of sludge flocs to maximize the flocculating ability of
sludge for effective floc separation.
4) The absence of correlations between the physicochemical properties (hydrophobicity and
surface charge) and the SVI indicates the limited potential of using hydrophobicity and surface
charge in understanding the compressi bility of sludge. The results suggest that the cornpressi bility
of sludge may be controlled mainly by the physical structure of flocs and surfaces, such as the size
and EPS content, but not by the physicochernical properties of sludge surfaces.
5 ) The proportions of EPS components (proteins/carbohydrates and /or proteins/(carbohydrates +
DNA)) were more important thm the quantities of individuai EPS components in controllhg the
hydrophobicity and surface charge, and contributed, at Ieast partially, to the changes in
physicochemical properties (hydrophobicity and surface charge) of sludge surfaces. It is likely
that it is the specific molecular composition and structure, such as proteins, rather than the
quantity of the total EPS content, that determine the hydrophobicity and surface charge.
CHAPTER VII
INTERPARTICLE INTERACTIONS AND STABILITY
Based on the experimental results in Chapter VI, it is clear that the hydrophobicity and
surface charge of sludge flocs are crucial for determinhg the flocculating ability of fine flocs; they
also provide a scientific explanation for the change in the flocculating ability of fine flocs with
respect to the sludge retention tirne (SRT). This may in tum suggest that interparticle interactions
associated with the hydrophobicity and surface charge are important in controlling the stability of
sludge flocs. The stability of sludge flocs at different SRTs may Vary owing to changes in the
hydrophobicity and surface charge of sludge flocs in terrns of the SRT. However. more direct
evidence is still necessary to support these hypotheses. For better controI of the stability of sludge
flocs. a further study is essential to chrifi the dominant interparticle interactions keeping flocs
together. It is also of interest to understand the feasibility of biologically rnanipulating these
interparticle interactions to minimize the number of non-settleable fine flocs in the treated effluent.
One wa>- to evaluate floc properties, and thus to predict how flocs will interact with one
another is to observe how flocs respond to extemally imposed environmental conditions.
Accordingly, for sludge flocs with different SRTs, the environmental conditions were chemically
manipulated in batch experiments by varying pH, ionic strength, cation valency, and urea and
ethylenediaminetetraacetate (EDTA) concentrations. The effects of pH, ionic strength and cation
valency on stability were used to test electrostatic interactions. The presence of ionic interactions
\vas studied by adding a strong chelating agent (EDTA) to break flocs linked by salt bridges. Urea,
a chaotropic agent, was applied to test the potential existence of hydrogen bonds between flocs.
120
The stability of sludge flocs was evaluated by the dissociation constant, which was
calculated from the linear dope of the accumulative turbidity in sequential washings versus the
nurnber of sequential washings (Figure 3.9). The dissociation constant has a unit of turbiditylg
biomass per washing. A poorer stability is associated with a larger dissociation constant.
7.1 Results
The expenmental results are outiined in the following sequence: influence of pH, ionic
strength, and cation valency on stability; effects of EDTA and urea on stability; influence of SRT
on interparticle interactions and the stability of siudge flocs. Each data point of dissociation
coristants in the following figures represents the average value (at least three different
measurements at different experimental times) I one standard deviation.
7.1. I Effect of the number of sequential washings on accumulative turbidity
Figure 7.1 shows the influence of the number of sequential washings on the accumulative
turbidity. The increase in the accurnulative turbidity with respect to the number of sequential
washings was almost linear within 5 or 6 washings. However, the magnitude of changes in the
accumulative turbidity significantly decreased after 5 or 6 washings under neutraI pH and lower
ionic strength conditions. In contrast, the influence of the number of sequential washings on the
slope of the accurnulative turbidity was not significant within 8 washings with urea and EDTA
treatments. The different behaviour of sludge flocs with different water chernistries may suggest
that the relative importance of different interparticle interactions varied with respect to the distance
from the surface to the core in a floc.
rn SRT=4 days, EDTA (1 00 rrg/L) A SRT= 4 days. Uea (8 M) 2
No. of Washings
Figure 7.1 Effect of the number of sequential washings on the accumulative
turbidity under different water chemistry conditions
7. i .2 Effect of pH on stability
The influence of pH on the stability of sludge flocs is shown in Figure 7.2. Sludge flocs
were generally more stable (smafler dissociation constant) in the pH range of 4.5 to 9.5 than those
in the pH range less than 2. For sludge flocs fiom different SRTs, a minimum dissociation constant
occurred in the pH range of 2.6 to 3.6, which is the pH range for the isoelectric point (Figure 6.2).
Dissociation constants of sludge flocs at lower SRTs (4 and 9 days) were larger than those at higher
SRTs (1 6 and 20 days), in the pH range of 4.5 to 9.5 (ANOVA, p< 0.05).
---- -SRT=4 da* 1 e S R T = 9 days
*SRT=16 days -SRT=ZO days
Figure 7.2 Effect of pH on stability. Results are expressed as
the average t one standard deviation.
7.1 -3 Effects of ionic strength and cation valency on stability
The effects of ionic strength and cation valency on the stability of sludge flocs are
illustrated in Figures 7.3 and 7.4. Dissociation constants of sludge flocs decreased with an increase
in the ionic strength (1) of bo t . KCl and CaClz solutions in the range from O to 0.025 molL. No
significant changes in dissociation constants were observed for sludge flocs fiom SRTs of 4. 9 and
16 days. when the ionic strength was larger than 0.025 mol/L (Figure 7.3). Following an initial
decrease under lower ionic strength conditions (O to 0.025 mol/L), the dissociation constants,
however, increased with an increase in the ionic strength for sludge flocs from an SRT of 20 days
under higher ionic strength conditions (0.025 to 0.6 moi&) (Figure 7.3). Under lower ionic strength
conditions (O to 0.025 mollL), the dissociation constants of sludge flocs in the CaCl, solution were
smalter than those in the KCI solution, as shown in Figure 7.4.
SRT=9 days
.- SRT=16 days -- SRT=20 day s
Figure 7.3a Effect of ionic sbength (KCl) on stability. Results are
expressed as the average +. one standard deviation.
i-rh,-ST=.l day s 4 SRT=9 day s - SRT=16 day s T -- ,-=-. SûT=20 day s
Log (1) (mol1L)
Figure 7.3b Effect o f ionic strength (CaClJ on stability. Results are
expressed as the average rt one standard deviation.
4 -3 -2 -1 O
Log (1) (mol/L) Figure 7.4 Effect of cation valency at different ionic strength conditions on stability
(Samples fiorn SRT= 4 days). Results are expressed as the average t one
standard deviation.
7.1 -4 Effect of EDTA concentration on stability
Figure 7.5 shows the influence of EDTA concentration on the stability of sludge flocs. No
statistical 1 y signi ficant changes in the dissociation constants were observed at each SRT. when the
EDTA concentration was less than 40 mg/L or greater than 100 mg/L. It appears that there was a
criticai EDTA concentration at about 100 mgL, at which there was a significant increase in
dissociation constants. The dissociation constants of sludge flocs at Iower SRTs (4 and 9 days)
were greater than those at higher SRTs (16 and 20 days) (ANOVA, p <0.05).
0.00 ! a 1 a I 1 a 1 I 1 a J O 100 200 300 400 500 600
EDTA (mglL)
,&, SRT=4 day s
Figure 7.5 Effect of EDT.4 concentration on stability. Results are
expressed as the average + one standard deviation
7.1.5 Effect of urea concentration on stability
d S R l = 9 days
-SRT=l6 days
-. SRT=20 days
The influence of urea concentration on the stability of sludge flocs is shown in Figure 7.6.
"
Generally. the addition of urea broke larger flocs into fine flocs. A sharp change in the dissociation
constants of sludge flocs was observed when urea (2M) was added with SRTs of 4 and 9 days; but
no statistically significant differences in the dissociation constants were found in the range of urea
concentration from 2 to 11 M (ANOVA, p >O.OS). On the other hand, dissociation constants of
sludge flocs with SRTs of 16 and 20 days increased slowly with an increase in urea concentration
from O to 8 M, and then reached an almost constant value in the urea concentration range of 8 to 1 1
M. Dispersed flocs reflocculated again when the urea was removed by washing sludge flocs with
deionized distilled water.
,, SRT=4 days ,-=-, S RT= 9 day s -SRT=16 days ,- SRT=20 days
Figure 7.6 Effect of urea concentration on stability. Results are
expressed as the average + one standard deviation.
7.2 Discussion
The work described here is concerned with elucidating the physicochemical nature and
relative importance of interparticle interactions that bind sludge flocs together, and the influence of
SRT on the stability and interparticle interactions of sludge flocs. The experimental results are
discussed in the following sequence: electrostatic interactions, ionic interactions, hydrogen bonds,
and influence of SRT on the stability and interparticle interactions. Next, a conceptual mode1 is
proposed to describe the floc structure.
7.2.1 Electrostatic interactions
/iH. Al1 sludge rnicroorganisms in aqueous suspensions cany a surface charge which can arise
from the ionization of hc t iona l groups on sludge surfaces. The presence of a net surface charge on
sludge surfaces creates repulsive electrostatic interactions which prevent close contact of sludge
microorçanisms. The change in dissociation constants with respect to pH, as shown in Figure 7.2,
provides strong evidence that repulsive electrostatic interactions are involved in disrupting the
stability of sludge flocs. The variation in dissociation constants with pH could be explained by the
presence on sludge surfaces of different functional groups, such as carboxyl, arnino and phosphate
groups with different ionization constants @Ka). A pKa range of 2.9 to 5.0 for sludge swfaces
(ionic strength I=0.004 mol&) was reported by Forster (1 968 and 197 1) for sludge flocs with
different compressibilities. Since EPS were composed mainly of proteins and carbohydrates with
only a srnaIl amount of DNA, and acidic polysaccharides were not detectable in this study, it is
reasonable to believe that the contribution of acidic polysaccharides and DNA to the surface charge
was minor. The major source of surface charge was the ionization of carboxyl and amino groups in
proteins.
It is known that the carboxyl and amino groups of amino acids have pKa values of 1.71 to
3.0 and 8.2 to 1 1, respectively (Smith and Martell, 1976; Lehninger et al., 1993). Here, the
repulsive electrostatic interaction was minimized in the pH range of 2.6-3.6 for the isoelectric point
(Figure 6.2) sa that fine flocs could approach each other closely, and the dissociation constants of
sludge flocs were at a minimum. The sharp increase in the dissociation constants at a pH less than
the isoelectric pH could be related to the adsorption of H+ by amino groups (R-NH,) to form R-
NH,- under low pH conditions. The relatively stable dissociation constants, when the pH was
higher than 4.3, could be attributed to the combined influence of the ionization of amino and
carboxyl groups on sludge surfaces. The naturally occumng arnino groups might form a stable five-
member ring chelated with divalent cations, such as Ca2+! Mg2+, fkom nutrients, and therefore
neutralize the amino groups (Huang et al., 1999). At least 99% of the carboxyl groups of amino
acids will ionize into -COOe groups when pH = pKa + 2 (i. e., pH = 4-S), so that the magnitude of
128
the surface charge is relatively stable, which further explains the stability of the dissociation
constants in the pH range of 4.5 to 9.5.
h i c Srrenath and Cafion Vafencv. The effect of electrostatic interactions on the stability of sludge
flocs is further supported by other experiments, where sludge flocs were suspended in solutions
with different ionic strengths and cation valencies, as shown in Figures 7.3 and 7.4. The variations
in dissociation constants with respect to changes in ionic strength and cation valency could be
explained by the change in the electrical double-layer thickness under lower ionic strength
conditions (O - 0.025 mol/L). According to the DLVO theory (Dejaguin and Landau. 1941;
Venvey and Overbeek. 1948), an increase in ionic strength has two effects on the electrical double
layer interaction. First, the corresponding increase in the Debye-Hukel parameter causes a decrease
in the range of repulsion so that fine flocs may approach more closely. Also. added salt causes a
decrease in the zeta potential of fine flocs, which reduces the repulsion at a given distance. The
dccrease in the dissociation constants was sharper in solutions with a higher cation valency (Ca2+)
under lower ionic strength conditions (O - 0.025 mol/L) (Figure 7.4), because both of these effects
are more pronounced with more highly charged ions (Overbeek, 1980).
Following an initial decrease under lower ionic strength conditions (O - 0.025 mol/L). the
dissociation constants of sludge flocs from SRTs of 4, 9 and 16 days remained the same regardless
of changes in ionic strength fiom 0.025 to 0.6 mol/L. These results indicate that a criticai
concentration of electrolytes existed for maintaining the stability of sludge flocs. Figures 7.3a and
7.3b show that the critical concentration of KCI and CaCI, for a minimum deflocculation of sludge
flocs was about 30 rnM and 2 mM, respectively. The critical concentration of CaCl, at 2mM is
129
consistent with the critical concentration (about 1 rnM) of CaCl, for the destabilization of a pure
culture in aqueous solutions, as reported by Eriksson and Hardin (1 987).
The U shape of the dissociation constant curve for sludge flocs with an SRT of 20 days is
generally in accordance with the findings of Zita and Hermansson (1994) with sludge flocs fiom a
full-scaie activated sludge system. It is interesting to note that the dissociation constants of sludge
flocs increased with an increase in ionic strength in the range of 0.025 to 0.6 rnoi/L, implying
deflocculation of sludge flocs under higher ionic strength conditions. Although Zita and
Hermansson (1994) concluded that the DLVO theory could not account for the increase in the
dissociation constant with an increase in ionic strength above 0.1 mol& a reversa1 in the zeta
potential on both sides of the isoelectric point could provide another explanation. At higher salt
concentrations (KCl and CaCII), ionic adsorption ont0 sludge surfaces might cause a charge
reversal from a negative to a positive value, so that the repulsive electrostatic interactions could
disrupt sludge flocs again under higher ionic strength conditions. Forster (1968 and 1971) reported
that a reversal of the surface charge of sludge occurred at an ionic strength of about 0.2-0.3 moVL
at the neutral pH. For this reason, DLVO theory might still be applicable in explaining the stability
of sludge flocs under higher ionic strength conditions in Zeta and Herrnansson's (1 994) and in the
present study.
7.2.2 Ionic interactions
It has long been thought that the functional groups, such as the amino, carboxyl, and
phosphate groups on sludge surfaces can incorporate divalent cations from nutrients into the floc
matrix and contnbute to its stability (Eriksson and Alm, 1991; Bruus et al., 1992; Higgins and
Novak. 1997). An indirect study of the salt bridging mechanisrns can be performed by adding
130
strong chelating agents, such as EDTA, to remove divalent cations firom the floc matrix (Eriksson
el al., 1989; Eriksson and Aim, 1991). As shown in Figure 7.5, the results indicate that the stability
was disrupted in the presence of EDTA and further imply that ionic interactions through salt
bridging were involved in maintainhg the stability of sludge flocs. This is generaily consistent with
the observations of Eriksson and Alm (1991), who found a sirnilar influence of EDTA on the
turbidity of sludge flocs from full-scale activated sludge systems. The presence of a cntical EDTA
concentration (about 100 mg/L) may suggest the combined influence of Na- and [HIEDTA] " ions
on the stability of sludge flocs. On the one hand, the [H2EDT~IL' ions could remove the divalent
cations from the floc matrïx and therefore break floc structures linked by salt bridging. On the other
hand. the addition of Na+ and [H,EDTA]" ions could compress the electrical double layer thickness
under lower ionic strength conditions, and further enhance the stability of sludge flocs. These
opposite effects cornpensated for each other, and therefore no significant changes in dissociation
constants would be expected at lower EDTA concentrations (0-40 mg/L). As was the case in this
study. a significant increase in the dissociation constants at higher EDTA concentrations
( r 100mgL) suggests that the disruption of ionic interactions by removing divalent cations was
more important than the compression of electrical double Iayer thickness. The results fkom this
study îùrther supports the salt bridging structure models proposed in previous studies (Novak and
Haugan. 1 978; Bruus et al-, 19923 Higgins and Novak, 1997).
Although Forster and Dallas-Newton (1 980) suggest that, for sludge samples from full-scale
activated sludge systems, cations were involved in bridging sludge flocs through acidic
polysaccharides, the results fiom this study indicate that salt bridges through acidic polysaccharides
were minor for al1 sludge samples at different SRTs, due to the non-detectable levels of acidic
polysaccharides in the EPS. This M e r suggests that carboxyl groups from amino acids in proteins
131
and phosphate groups fiom DNA were the dominant fùnctional groups involved in salt bridging,
which is consistent with recent findings of Higgins and Novak (1997) that proteins were important
in the salt bridging of sludge flocs.
7.2.3 Hydrogen bonds
Urea has a strong ability to fonn hydrogen bonds with water, biopolyrners and itself, and
therefore is widely used as a chaotropic agent in breaking d o m the hydrogen bonds between
biopolymers (Gordon and Jencks, 1963; Engelborghs, 1992); this effect increases with an increase
in the urea concentration.
In the sludge floc matrix. it has k e n hypothesized that hydrogen bonds are involved in
sludge floc formation and its stability, because a nurnber of patterns of hydrogen bonds can be
formed arnong carbohydrates, proteins and DNA in the EPS. This hypothesis was indirectly
verifred in the present study by the addition of urea, a chaotropic agent. An increase in the
dissociation constants after wea treatments (Figure 7.6) indicates that urea acted in disrupting the
stabi lity. The direct influence of urea on hydrogen bonds formed between sludge surfaces could be
at least part of the explanation. Similar phenornena were found conceming the influence of urea
and guanidiniurn chloride on the flocculating ability of yeast (Mill, 1964; Calleja, 1974; Kamada
and Murata. 1 984) and bacteria (Nesbitt er ul., 1 982).
Since turbidity decreased to the initial level once the urea solution was removed and
replaced with deionized distilled water, it is clear that fine flocs regained their flocculating ability
wi th this replacement. This observation also suggests that urea primarily acted by disrupting
structure of interparticle interactions, but not structures on fine particle surfaces.
7.2.4 Influence of sludge retention time on the stability and interparticle interactions
In practice, al1 interparticle interactions occur simultaneously. The relative importance of
these interactions varies with changes in particle properties and with changes in the environmental
conditions of the surrounding solutions.
The variations in the stability of sludge flocs with respect to pH, ionic strength, and cation
valency (Figures 7.2 and 7.3) indicate that larger repulsive electrostatic interactions arose fiom
sludge surfaces at lower SRTs (4 and 9 days). This is not surprising, as sludge surfaces were more
negatively charged at lower SRTs (4 and 9 days) than at higher SRTs (16 and 20 days) (Figure 6.3),
and further increased repulsive electrostatic interactions at lower SRTs (4 and 9 days). This could
also explain the presence of larger dissociation constant at lower SRTs (4 and 9 days) than those at
higher SRTs (16 and 20 days) (Figure 7.5) (ANOVA, p<0.05), because it is reasonable that the
larger the magnitude of the surface charge. the more divalent cations cm be adsorbed in the floc
matrix and therefore the greater the influence of EDTA on stability.
From Figure 7.6, it was found that the stability of sludge flocs at SRTs of 4 and 9 days was
quite sensitive to the addition of urea; in contrast, sludge flocs at higher SRTs (16 and 20 days)
w-ere relatively persistent to the treatment of urea even at a urea concentration of 1 1M. These
resuIts suggest that hydrogen bonds formed between sludge surfaces at an SR?' of 4 days played a
more important role in controlling the stability of sludge flocs than those at higher SRTs (1 6 and 20
days).
In surnmary, it appears that hydrogen bonds and ionic interactions played a dominant role in
binding flocs together by overcoming the larger repulsive electrostatic interactions at lower SRTs
(4 and 9 days). The persistence of sludge flocs at higher SRTs (16 and 20 days) to EDTA and urea
133
solutions (Figures 7.5 and 7.6) suggests that ionic interactions and hydrogen bonds were less
important as compared to those at lower SRTs (4 and 9 days). This in turn indicates the
involvement of other dominant physical forces in controlling the stability of sludge flocs at higher
SRTs (16 and 20 days). The potential involvement of physical enmeshment of EPS and
hydrophobic interactions could explain the stability of sludge flocs in the presence of either EDTA
or urea at higher SRTs (16 and 20 days). The presence of a "skin" layer biopolymer on the surface
of microcoIonies at higher SRTs (16 and 20 days) (pictures not shown) could protect the interior
celIs from disruption by changes in external environments and therefore make them physically
more stable. In addition, a more hydrophobic surface at higher SRTs (1 6 and 20 days) (Figure 6.1)
could provide much stronger van der Waals and hydrophobic interactions which govern the
stability of sludge flocs.
Figure 7.7, together with Figures 7.2 to 7.6. siiows that sludge flocs at higher SRTs (16 and
20 days) were generally more stable (smaller dissociation constants) than those at lower SRTs (4
and 9 days). The results from this study are consistent with the findings of previous studies (Parker
et d.. 1 97 1 : hlalberg, 1992) with sarnples from full-scale activated sludge systerns. However,
previous studies (Parker et al., 1971; Whalberg, 1992) assumed that the weaker interaction or
poorer stability of sludge flocs at lower SRTs was caused by lower levels of EPS production.
Unfortunately, the measurement of EPS was not part of the study of Parker et al. (1971), and only
acidic polysaccharides were detected in the study of Whalberg (1992). The results from this study
(Figure 5.1 b). together with the findings of Kiff (1978) and Gulas et al. (1979): indicate that EPS
production was not limited to the stationary and endogenous phases of sludge growth which are
associated with much higher SRTs. In these studies, a large amount of EPS was also extracted from
sludge at lower SRTs. Therefore. the poorer stability of sludge flocs could not be explained by
134
assuming a lack of EPS production at Iower SRTs. The variations in the stability of sludge flocs in
terms of the SRT were caused by changes in the physicochemical properties (hydrophobicity and
surface charge) of floc surfaces and in the dominant interparticle interactions.
SRT (days)
Figure 7.7 Effect of SRT on stability (sludge flocs in deionized distilled water).
Results are expressed as the average r one standard deviation. No. of
observations is 17, 15, 15 and 15 at an SRT of 4. 9, 16 and 20 days, respectively.
7.2.5 Proposed floc structure mode1
Based on the expenmental results fiom this study, a conceptual model is proposed to
describe the floc structure of floc-forming microorganisms. Unlike the structural model proposed
by Eriksson er al. (1992), which was focused mainly on the rnorphology of flocs at different SRTs,
the structural model proposed here emphasizes the interparticle interactions between flocs.
135
~Macrostrucrure: The observation of a significant decrease in turbidity after several
sequential washings (usually 5- 6 washings) under lower ionic strength conditions or with
deionized distilled water (Figure 7- 1) is consistent with the findings of Zita and Hermansson
(1994). These results suggest that sludge flocs can be divided into two layers, as proposed by
Eriksson and Alm (1991): an outer layer and an interior layer (Figure 7.8). The outer layer of
sludge flocs has a loose structure kept together by relatively weak interactions; it is therefore more
sensitive to changes in physical and chernical environments. The interior layer is much stickier and
has a stronger cell-ce11 adherence to resist the change in e.uternal environrnents. This mode1
esplains well the observed change in turbidity with the number of sequential washings, and
suggests that different dominant mechanisms may be involved in different positions of a floc.
It is postulated that the interior layer is more hydrophobic than the outer layer. First, the
hydrophobic surfaces have a nafural tendency to avoid hydrated environments and therefore to
aggregate. Second, Ganaye et al. (1 997) observed the presence of highly pronounced hydrophobic
zones inside sludge flocs based on spectroscopic observations. Third, the interior layer of sludge
flocs has a lower growth rate than the outer layer, due to the limitation of dif is ion of oxygen and
nutrients to the interior layer from the bulk solution (Atkinson and Daoud, 1976; Eriksson et al.,
1 992). A lower growrh rate was associated with more hydrophobic surfaces (Figure 6.1).
Outer layer
Interior Iayer
Fine fragment washed away
Oxygen, nutricnts, and growth rate profiles
Interior layer u
A
Hydrophobicity pro fi Ie /
Figure 7.8 A schematic mode1 of a two-layer sludge floc
O
.Lficrosrructure: The colloidal sbbiiity of the outer layers of sludge flocs is maintained by the
surface interactions between sludge flocs, as s h o w in Figure 7.9. As demonstrated in this study,
electrostatic and ionic interactions and hydrogen bonds are invoived in governing the colloidal
stabiIity of this layer.
w i Position
4
Outer iayer
Centre
a'!!
lonic interactions Hydrophobic interactions
f Electrostatic 1 interactions A O CH3
'c-O- I I O HZ
CH3
Hydrogen bonds
Figure 7.9 A conceptual mode1 of surface interactions that maintain the stability of the outer layer of sludge flocs
nie interior layer of sludge flocs is held strongly by the EPS and physical enrneshment. The
more hydrophobic nature of the interior layer provides stronger van der Waals and hydrophobic
interactions than those in the outer layer; these contribute to the mechanical strength of the interior
layer. It is logical to assume that the van der Waals and hydrophobic interactions are two potential
dominant mechanisms controlling the stability of the interior layer of sludge flocs.
7.3 Summary
The interparticle interactions and their influence on the stability of sludge flocs at different
SRTs were studied by suspending sludge flocs in chemically manipulated solutions. The
conclusions are presented as follows.
138
1) The strong influence of pH, ionic strength, and cation valency on the deflocculation of sludge
flocs shows that repulsive electrostatic interactions were involved in dismpting the stability of
sIudge flocs.
2 ) Deflocculation of sludge flocs with a urea concentration fiom 2 to 5 M suggests that hydrogen
bonds were important in maintaining the stability of sludge flocs. Deflocculation of sludge flocs
with EDTA treatrnents indicates that ionic interactions were also involved in governing the
stability.
3) The stability of sludge flocs can be manipulated by varying the SRT. Floc strength was stronger
3t higher SRTs than at lower SRTs.
4) The relative importance of interparticle interactions changed with respect to the SRT. Ionic
interactions and hydrogen bonds were two dominant interparticle interactions that maintained the
stability of sludge flocs at lower SRTs (4 and 9 days); while other mechanisms, such as physical
enmeshment and hydrophobic interactions, were likely more important than ionic interactions and
hydrogen bonds in controlling the stability of sludge flocs at higher SRTs (1 6 and 20 days).
5 ) A mode1 of sludge floc structure is proposed to describe the differences within different layers
of a floc. It is suggested that hydrophobic interactions play an important role in the initial stage of
bioflocculation by overcoming the repulsive electrostatic interactions and bringing the
microorganisms close enough to form specific interactions. Then sludge flocs are linked through
salt bridging (ionic interactions), and the structure is M e r stabilized by hydrogen bonds and
physical enmeshment formed between sludge surfaces.
CONCLUSIONS GND RECOMMENDATIONS
The contributions of this thesis resuit fiom an integrated and in-depth evaluation with
new information and insights on the surface properties and interparticle interactions of sludge
flocs at different sludge retention times (SRTs), and on their role in bioflocculation, compaction
and stability.
Figure 8.1 summarizes these new findings. An improved understanding has k e n
achieved in four aspects: 1) production, composition and molecular weight of extraceliular
polymeric substances (EPS); 2) physicochemical properties (hydrophobicity and surface charge)
of sludge surfaces; 3) interparticle interactions governing the stability of sludge flocs; and 4)
clarification of the role of swface properties in biofiocculation and compaction. A transition in
floc properties occurred at an intennediate SRT range of 9 to 12 days. The results from this study
provide new explanations for changes in the flocculating ability of sludge flocs at different SRTs.
This study is the first to demonstrate expenmentally the changes in the proportion of EPS
components and physicochemical properties of sludge flocs with respect to the SRT, and provide
preliminary information about the molecular weight distribution (MWD) of EPS components. An
integrated study not only on the production, composition and molecular weight of EPS but also
on the hydrophobicity and surface charge of sludge flocs provides an improved understanding of
the role of particular surface properties in the flocculating and compaction properties of sludge
flocs.
The utilization of chemically manipulated solutions to challenge stability provides a more
comprehensive and improved understanding of interparticle interactions that keep flocs together.
I t appears that electrostatic, ionic interactions and hydrogen bonds are al1 involved in goveming
stability. The results fiom this study demonstrate that interparticle interactions can be
manipulated to control the stability of sludge flocs by the SRT.
Physicochemical < * Hydrophobicity properties Surface charge
EPS
Interparticle interactions
Carbohydrates
Proteins
DNA
Total EPS
13 Electrostatic interactions Ionic interactions
\ + Hydrogen bonds
ColIoidal stability
ESS
SRT
Figure 8.1 Changes in sludge properties with respect to the SRT. The arrow direction means
either an increase in the magnitude of the measurement (physicochemical properties
and EPS) or an increase in the relative importance of interparticle interactions in
terrns of SRT. No arrow in the line means no change.
8.1 Conclusioas
Based on the results fiom a long-term and well-controlled laboratory study, a more
comprehensive and improved understanding has been obtained in surface properties and
interparticle interactions as well as their correlations to the flocculating ability, compressibility,
and stability of sludge flocs. The specific conclusions from this study can be surnmarized as
follows:
1 ) The production and composition of EPS were affected by the SRT. Proteins and carbohydrates
were the dominant components, followed by a smali amount of DNA in the EPS. Acidic
polysaccharides were not detectable in the EPS fiom a laboratory-scale SBR system fed a
glucose-based synthetic wastewater. The molecular weight distribution ( M W ) of proteins.
carbohydrates, and DNA in the EPS covered a broad spectrum, fiom less than 1,000 daltons to
more than 100,000 daltons, and the majority (>85%) of EPS components had moIecuIar weights
larger than 10.000 daltons. The total arnount of EPS was independent of the SRT. The ratio of
protein to carbohydrate, however, increased with an increase in the SRT fiom 4 to 12 days, and
then reached an almost constant value for SRTs fiom 12 to 20 days.
2 ) Physicochemical properties of slirdge sw$aces, such as hydrophobicity and surface charge,
rc7crc. uflected by the SRT. Sludge surfaces at higher SRTs (16 and 20 days) were less negatively
charged and more hydrophobic (larger contact angle) than those at lower SRTs (4 and 9 days).
3 ) The proportions of EPS componenfs played a more important r-ole in defermining the
physicochemical properties (hydrophobicity and surface charge) of sludge surfaces than the
overaZl quantity of EPS components.
142
4) It is the physicochemical properties (hydrophobicity and surface charge) of sludge surfaces,
rather than the quantity of EPS, that govern biojoccuZation In contrust* the EPS content plays a
more important role in governing the compressibiZity of sludge.
5 ) The relative importance of interparticle interactions changed with respect to the SRT. lonic
interactions and hydrogen bonds were two dominant forces maintainhg the stability of sludge
flocs at lower SRTs, but were less important at higher SRTs. This irnplies that other mechanisms,
such as physical enrneshrnent and tiydrophobic interactions, were iikely more important than
ionic interactions and hydrogen bonds in controlling the stability of sludge flocs at higher SRTs.
6 ) The stability could be rnanipulated by varying the SRT. Flocs were stnicturalty stronger at
Iiigher SRTs than at lower SRTs. The stability of sludge flocs at lower SRTs was much more
sensitive to changes in environmental conditions than that at higher SRTs.
7 ) Chmges in the proportions of the EPS components, hydrophobicity, szrrface charge, and
srability of sludge jlocs occurred ut an in termediate SR T range o f9 ru 12 days. This suggests that
it is possible to control the surface properties and interparticle interactions for effective floc
separation by designing an appropnate SRT.
8. 2 Recommendations for Future Studies
Although this thesis provides new fimdamental information and insights about the
iniprovement of floc separation, a number of questions raised by this study still require fùrther
investigation. Specific recommendations for fiiture studies are outlined below.
143
1) Novel methods are desirable for determining the surface composition of sludge flocs without
the disruption of cells. The amount of EPS is an operationally defined parameter and the
definition of EPS is strongly dependent on the techniques used to detect them. Classical methods
used to determine the amount and composition of EPS require extraction procedures to separate
EPS fiorn ce11 walls. Consequently, the amount and composition of EPS are strongly related to
the extraction eficiency. At present, a number of extraction methods, including thermal
extraction, centrifugation and chernical treatments, are used by different research groups to
assess the role of EPS in bioflocculation, compaction and dewatering. The results tiom different
studies can not be compared owing to the different extraction methods used. It is suggested that
X-ray photoelectron spectroscopy (XPS) is useful in determining the surface composition of
sludge flocs without the disruption of cells. However, care must be taken to eliminate the
contamination of sludge flocs in using the XPS method. The XPS method should be applicable to
sludge sarnples from laboratocy -SC& bioreactors receiving synthetic wastewaters.
2) Further characterization of the composition and physical configuration of EPS cornponents is
necessary for a better understanding of sludge floc formation. The results obtained in this study
indicate that hydrophobicity and surface charge appear to be important in bioflocculation.
However, what causes changes in these properties is still not well understood. Understanding
what induces these variations to occur and deveIoping the strategies for controlling these changes
are essential for optimizing floc separation. In particular, it is recornmended that the MWD be
exarnined in more detail. A combination of high performance liquid chromatography (HPLC),
gas chromatography/mass spectroscopy (GUMS), ultrafiltration, and electrophoresis will
provide more comprehensive information about the structure of specific EPS molecules.
3) The development of mathematical models to predict the intemediate SRT range is of
engineering significance. Althou& there is an intermediate SRT range for better flocculation, as
found in both this study and previous investigation (Bisogni and Lawrence, 1971; Chao and
Keinath, 1979; Lovett et al., 1983; Knocke and Zentkovich, l986), very little has k e n done to
understand why there is an intermediate SRT range, and how to predict this range.
4) The settleability (zone settling velocity) of sludge at different SRTs was not a focus of this
study. No work has k e n done to link fundamental floc properties, such as surface properties,
with practical settling techniques (Zheng and Bagley, 1998). An effort in this area is essential for
a better understanding of the settleability of sludge at different SRTs.
5) A more comprehensive study on the microbial comunity structure at different SRTs is
necessary for a better understanding of the contributions of both the growth rate and the
microbial cominunity to the changes in the surface properties of sludge flocs. A combination of
carbon substrate oxidation profiles (phenotypic fingerprinting) and other techniques, such as
DNA sequence analyses. will provide new insights into the microbial community structure with
respect to the SRT.
CHAPTER IX
ENGINEERING SIGMFICANCE
Effective separation of sludge flocs fiom the treated effluent has been a tremendous
challenge for activated sludge systems for alrnost a century. An estimate of the flocculating
efficiency of sludge flocs indicates that as much as 25% of newly grown microorganisms fail to
aggregate to f o m large flocs for effective gravity separation (Bossier and Versmete, 1996).
Most of effluent suspended solids (ESS) are smaller clumps or fine flocs. The efficiency of
gravity separation of sludge flocs is dependent on the nature of sludge surfaces and the
hydraulic conditions in the aeration tank and secondary clarifier. This study focused on
developing an improved understanding of the nature of sludge surfaces to optimize biosolids-
1 iquid separations.
A significant conclusion fiom this study is that the physicochemical properties
(hydrophobicity and surface charge) of EPS are more important than the quantity of EPS
extracted fiom sludge surfaces in goveming the flocculating ability of sludge flocs. This
clarifies a misunderstanding about the role of the total EPS in bioflocculation, and is of practicd
guidance to wastewater treatment professionals. Using the EPS content as a measuring
parameter makes a limited contribution to understanding bioflocculation. Indeed, the larger
arnount of EPS can have a negative influence on the compressibility o f sludge, and still not
minimize the nwnber of non-settleabte fine flocs in the treated effluent. The results fiom the
present study also suggest that selecting the right type of polyrner with suitable physicochernical
properties is more important than adding more polymers in bioflocculation and compaction.
146
Addition of large amounts of polymers can even stabilize the sludge suspension and cause the
opposite results. due to the sîeric forces arising fiom the adsorbed polymers on sludge surfaces.
Accordingly, there should be optimal arnounts of polymen for bioflocculation and compaction.
Changes in the hydrophobicity, surface charge and colIoidal stability of sludge flocs
with respect to the SRT suggest that an intermediate SRT range exists in the activated sludge
process to produce the desired surface properties of sludge flocs for effective floc separation and
dewatering. Sludge flocs had better flocculating ability at higher SRTs (16 and 20 days).
However. Jenkins et al. (1984) and Eriksson et al. (1992) hypothesized that an extended
aeration with extremely high SRTs could disintegrate sludge flocs into pin-point flocs.
Therefore, an optimal SRT range for effective floc separation should exist in the activated
sludge process. The identification of the intermediate SRTs as found in this study is important,
because this allows the design of an activated sludge plant at a maximum eficiency of the
bioreactor, and such information would help to predict the potential floc separation problems in
the secondary clarifier for a given influent. A pilot-scale study on a given influent is
recommended, as it would enable wastewater treatrnent professionals to find out the
intermediate SRT range for optimal floc separation and maximum e=ciency of bioreactors.
The development of an improved hndamental understanding of interparticle interactions
that govern the colloidal stabiiity of sludge flocs provides scientific explmations for the
deterioration of the treated effluent during heavy rainfalls, spring thaws or at high temperattues.
This is because the ionic strength in the influent is significantly lower dwing rainfalls and
thaws. and high temperatures disrupt hydrogen bonds fonned between sludge flocs. The results
from this study suggest that addition of inorganic salts or polymeric flocculants is a useful
method for controlling the deterioration of treated effluents when necessary. Thennophilic
biological treatments of wastewaters may require the incorporation of membrane filtration to
reduce the level of ESS.
Another significant contribution of this study to the optimal design and operation of
sludge floc separations is to suggest promising directions for developing analytical tools that
can be used for the operational management of activated sludge systems. As the activated
sludge process is extrernely complicated, no one management tool can achieve ail the goals of
quality control. A combination of different management tools, such as microbiological. physical
and chemical tools, is necessary for effective monitoring and managing the optimal operation of
activated sludgc: processes. The results obtained from this study indicate that hydrophobicity
(contact angle) and surface charge (colloidal titration) are two of the dominant factors governing
the flocculating ability of sludge flocs. While contact angle measurement and surface charge
detemination do not solve the problems of disintegration and pin-point flocs, such information
does allow operators, managers, and engineers of activated sludge plants to monitor the
potential of sludge disintegration and the presence of pin-point flocs. The routine measurement
of contact angles at an activated sludge plant may be dificult, due to the cost for installing
contact angle measurement apparatus and the need for highly skilled technicians. However, the
strong correlation between surface charge and contact angle (Figure 6.4a) suggests that surface
charge is an adequate panmeter for estimating the potential of pin-point floc formation and
dispersed çrowth. Surface charge determination is easy and fast to perform. Therefore, it is
suggested that the routine measurement of surface charge should be considered a management
toot for activated sludge systems. However, it is emphasized that the application of surface
148
charge determination will be no use in situations where bulking is caused by the overgrowth of
f i I arnents because filamentous bulking involves completely di fferent mec hanisms.
Bioflocculation and compaction of activated sludge are complex phenornena. The results
from this study develop an in-depth understanding of the structure, surface properties. and
interparticle interactions of sludge flocs and the particular role of these properties in controlling
biofloccuIation. compaction, and colloidal stability at different SRTs. With this understanding,
the operators, managers, and engineers of activated sludge systems will be able to manage the
efficiency of floc separation and improve the quality of treated effluent more eficiently.
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Appendix A Mixed Liquor Suspended Solids Data (Unit: mg/L)
Date D ~ Y SBR-1 SBR-2 SBR-3 SBR4
StatisticaI Results
(mg/L) ( W L ) SRT (days) Ave. S D
M L S S 4 2250 150 9 2200 130
12 2000 130 16 2070 160 20 2000 170
* SD - - Standard Deviation
Date D ~ Y 23-Fe b-97 25-Feb-97 27-Fe b-97 06-Mar-97
15-Mar 1 9-Mar-97 3 1 -Mar-97 04-Apr-97 08-Apr-97 1 2-Apr-97 1 5-Apr-97 18-Apr-97 24-Apr-97 28-Apr-97
04-May-97 09-May-97 12-May-97 19-May-97 23-May-97 29-May-97 02-Jun-97 09-Jun-97 14-Jun-97 1 8-Jun-97 23-Jun-97 27-Jun-97 05-J uI-97 ? O-Jul-97 15-Jul-97 19-Jul-97 29-J uI-97
05-Aug-97 08-Aug-97 14-Aug-97 18-Aug-97 29-Aug-97 04-Sep-97 10-Sep-97 16-Sep-97 26-Sep-97 06-Oct-97 1 9-Ott-97
Appendu B Sludge Volume Index Data (Unit: mL/g DS)
Statistical Results:
SUMMARY SR T(days) Count
4 56 12 19 9 51
16 66 20 64
ANOVA Source of Variation SS
Between Groups 8567.922 Within Groups 94476.44 Total 103044.4
Sum Average Variance S. deviation 4788.1 85.501 79 481 .SM2 21 -9453
1260 66.31579 141 -1903 71.88235 41 55.2 81.47451 432.4787 20.79612 4965.4 75.23333 433-0075 20.80883 4654.1 72.72031 248.854 15.7751 1
Date D ~ Y 23-Feb-97 28-Feb-97 03-Mar-97 06-Mar-97 10-Mar-97 1 7-Mar-97 20-Mar-97 23-Mar-97 27-Mar-97 29-Mar-97 3 1 -Mar-97 04-Apr-97 05-Apr-97 08-Apr-97 12-Apr-97 15-Apr-97 20-Apr-97 29-Apr-97
04-May-97 09-May-97 12-May-97 15-May-97 19-May-97 23-May-97 29-May-97 02-Jun-97 09-Jun-97 1 8-Jun-97 23-Jun-97 27-Jun-97 05-Jul-97 15-Jul-97 19-Jul-97 29-JuI-97
08-Aug-97 18-Aug-97 04-Sep-97 1 0-Sep-97 1 6-Sep-97 18-Sep-97 22-Sep-97 30-Sep-97
Appendix C Effluent Suspended Soüds Data (Unit: mgL)
05-Oct-97 13-Oct-97 20-Ott-97 27-Ott-97 03-NOV-97 1 1 -Nov-97 25-NOV-97 1 0-Dec-97 26-Dec-97 10-Jan-98 22-Jan-98 31 -Jan-98 06-Fe b-98 09-Fe b-98 1 1 -Feb-98 16-Feb-98 23-Feb-98
29-Feb-98 07-Mar-98 12-Mar-98 15-Mar-98 24-Mar-98 3 1 -Mar-98 08-Apr-98 14-Apr-98 23-Apr-98 1 1 -May-98 27-May-98 30-May-98
Jun. 15, 98 29-Jun-98 06-Jul-98 1 9-JuI-98 28-JuI-98
10-Aug-98 24-Aug-98 08-Sep-98 20-Sep-98 04-Oct-98 20-Ott-98 01 -Nov-98 09-NOV-98 1 4-NOV-98 21 -Nov-98 O1 -0ec-98 10-Dec-98 17-Dec-98 22-Dec-98 27-Dec-98
Statistical Results:
Anova: Single Factor
SUMMARY SRT (days) Count Sum Average Variance
4 67 1994 29.761 19 69.00271 12 66 1949 29.5303 53.14522 9 14 459 32.78571 26.02747
16 81 1227 15.1481 5 25.47778 20 81 1112 13.7284 35.22531
ANOVA Source of Variation SS df MS F P-value F crit
Between Groups 1874 1 -26 4 4685.316 107.8779 4.27E-57 2.401 343 Within Groups 13203.22 304 43.43165 Total 31 944.49 308
Appendu D Emuent Chernical Oxygen Demaad Data (Unit: mg COD&)
Date SBR-1 SBR-2 SBR-3 S8R4
Appendix E-1 Extraceilular Poiymeric Substances Data (EPS unit: mg/g VSS)
SRT=6 days (SBR-1.2.3 anâ4) Date of sample TCHO Protein DNA AP ProteiniTCHO Total EPS SV1 (mUg) E
1 2-1 SApr-97 3.3 25.0 1.5 7.6 29.8 1 99 14-1 5Apr-97 3.4 17.6 2.0 5.1 23.0 181
1 û-Apr-97 4.4 18.3 1.9 0.8 4.1 24.7 1 58 01 -02-May-97 3.5 19.8 1.6 1.5 5.7 24.9 145 08-09May-97 3.4 21.9 1.7 1.2 6.5 26.9 102 1 2-1 4-May-97 3.7 18.9 1.2 1.5 5.1 23.8 94
02-Ju~97 2.8 12-7 0.9 1.3 4.5 16.4 32 05-Jun-97 2.9 13.6 1.0 1.4 4.6 17.6 31
1 4- 1 5June-97 3.3 6.6 0.5 0.7 2.0 10.4 35 17- Jun-97 2.7 7.5 0.6 2.8 10.8 37
22-23 June-97 4.7 8.4 0.5 1.8 13.6 41 24-Apr-97 7.0 20.5 1.1 2.9 28.6 178 30-Apr-97 8.0 21.0 1.2 2.6 30.2 150
Statistical Resuiîs SRT=6 days
Mean 4.01692 1629û769 1.1984 1.2014 4.401706891 2 1 . m Standard Error Median Mode Standard Deviation Sarnple Variance Kurtosis S kewness Range Minimum Maximum Sum Count
S R T 4 days (SBR-1) Date TCHO Protein DNA AP
Statistical Results SRT* days
ProteinBCHO Total EPS SV1 (rnUg) ESS (m)
Mean Standard Error Median Mode Standard Deviation Sample Variance Kurtosis Skewness Range Minimum Maximum Sum Count
SRT=9 days (SBR-2) Date TCHO Protein DNA AP
0 1 AU^-97 2 5 7.9 0.4 1 2- 13-Aug-97 2.3 6.4 0.4 1 7-1 8-Aug-97 5.6 5.2 0.3 25-26-Aug-97 4.4 3.7 0.4 07-08-Sept-97 4.1 12.7 1.5
1 &Sep-97 3.3 10.5 1.1 26-Sep-97 2.4 7.6 1.0 06-Od-97 3.7 13.7 1.6 1 7-06-97 3.2 10.8 1.3 27-Oct-97 2.3 5.8 0.7 03-NOV-97 3.4 6.2 0.8 1 &Jan-98 6.5 13.5 1.3 18-Feb-98 5.6 9.7 0.8 04-Mar-98 4.9 124 1.1 31 -Mar-98 5.1 8.8 0.8 13-Apr-98 6.7 14.9 1.4 2 1 -Apr-98 5.2 11.7 1.3
27-May-98 6.4 7.4 1.2 01 -Feb99 4.9 7.6 1.1
Statistical Results SRT=9 days
Protein/TCHO Total EPS SV1 (mifg) ESS (mgrl) 3.2 10.8 46 38 2.8 9.1 49 30 0.9 11.1 51 28 0.9 8.5 64 27 0.9 11.1 58 34 3-2 14.9 63 35 3.2 11.1 53 41 3.6 19.0 68 35 3.3 15.3 59 27 2 5 8.8 67 29 1.8 10.4 68 28 2.1 21 -3 108 38 1.7 16-1 90 50 2 5 18.4 88 35 1-7 14.7 74 21 2.2 23.0 108 13 2.3 18.2 95 17 1.2 14.9 79 33 1.5 13.6 95 30
Mean Standard Error Median Mode Standard Deviatior Sample Variance Kurtosis S kewness Range Minimum Maximum Sum Count 19 19 19 O 19 19
SRT=12 days (SBR-2) Date TCHO Protein DNA AP ProteirVTCHO Total EPS SV1 (mUg) ESS (m)
1 3- 1 SApr-97 2.9 21.9 1.7 7.5 26.4 59 40 16-18-Apr-97 2.8 15.8 0.8 0.7 5.6 19.4 58 39
1 1 - 1 3-May-97 2.9 15.5 0.6 1.5 5.4 19.1 61 37 23-25-May-97 4.2 20.3 1.1 4.8 25.6 75 30 1 4-1 5 June-97 4.6 13.7 1.0 0.8 3.0 19.2 61 35
17-Jun-97 2.4 14.0 1.1 6.0 17.5 70 31 22-23-June-97 3.4 11.9 0.9 3.5 16.2 63 35
Statistical Results
Mean 3.3 16.155714 f -0379 0,9967 5.1 17473708 20.49357 Standard Error 0.30205 1.3691096 0.125 0.2302 0.57568863~ 1.494269 Median 2.9 15.52 1 0.84 5.351724138 1 9.22 Mode 2.9 #NIA M A #NIA W A M A Standard Deviation 0.7991 5 3.6223236 0.3308 0.3988 1 S23128957 3.953463 Sample Variance 0.63863 13.12'1229 0.1094 0.1 59 2.31 9921819 15.62987 Kurtosis -0.80449 -0.725978 1.9364 #ûlV/û! -0.1 10280367 4.88861 9 Skewness 0.72541 0.760716ô 1.1295 1.495 0.092204313 0.871932 Range 2.2 9.92 1.03 0.75 4.530087154 10.235 Minimum 2.35 11.93 0.64 0.7 3.004395604 16.185 Maximum 4.55 21.85 1-67 1.45 7.534482759 26.42 Sum 23.1 1 13.09 7.265 2.99 35.82231 596 143.455 Count 7 7 7 3 7 7
SRT=16 days (SBR-3) Date TCHO Protein DNA AP ProteiriTTCHO Total EPS SV1 (mUg) ESS (mgil)
Statistical Results SRT=16 days Mean Standard Error Median Mode Standard Deviation Sample Variance Kurtosis Skewness Range Minimum Maximum Sum
SRT=20 days (SBR4) Date TCHO Protein DNA AP ProteintTcHo Total EPS SV1 (mUg) ESS (mgk)
Statistical Results SRT=20 days Mean 3.24305 12.940143 1 . l a 1.425 4.35273558 17.31929 Standard Error 0.1 8108 0.6142638 0.1069 0.135 0.427138447 0.59061 Median 3.1 12.6 1.16 1.425 4 17.481 Mode 3 #N/A 1.8 #NIA W A #NIA Standard Deviation 0.82981 2.8149105 0.4898 0.1 909 1.957394267 2.706515 Sample Variance 0.68859 7.9237212 0.2399 0.0365 3.831 39231 8 7.325224 Kurtosis 1 -99899 0.0839465 -0.522 #û IVIO! 6.954937854 -0.329637 S kewness 0.80848 0.2703009 -0.243 #OIVtû! 2.194095925 -0.324954 Range 3.74 11.2 1.73 0.27 9.579545455 10.89 Minimum 1.76 8.3 0.07 1.29 1.5 11.53 Maximum 5.5 19.5 1.8 1.56 11 .O7954545 22.42 Sum 68.1 W 271 -743 23.958 2.85 91.40744719 363.705 Count 2 1 21 21 2 21 21
Appendix E-2 Molecular Weight Distribution of EPS Components
(Ultrafiltration Results)
SRT=4days Date MWD(daltons)
06-Feb-98 <1,000 1,000-1 0,000 10,000-100,000 ~100,000
Statistical Results SRT=4days
MWD(dal1ons)
%Proteins 2
11 7
47
8 21 22 57
O 3 4 93
5 9
12 73
O h Proteins Average S. D. Average S.D. Average S.D.
2 3 4 4 6 4 11 7 11 8 11 6 9 4 12 8 8 4
SRT=9 days
Date MWD(daltons) %TCHO 1 2-Jan-98 4,000 11
1,000-1 0,000 14 10,000-1 00,oo 10 >100,OOO 41
Statistical Results SRT=9 days MWD(da1tons) %TCHO
% Proteins 5
12 9
42
12 16 20 29
O 2 6
92
4 17 14 65
%Proteins Average S.D. Average S.D. Average S. D.
c l ,000 9 8 5 5 7 6 1,000-1 0,000 14 3 12 7 10 5 10,000-1 00,OO 10 4 12 6 13 11 >100,000 56 23 57 28 77 12
SRT=16 days Date MWD(daltons) %TCHO
1 3-Feb-98 4,000 O 1,000-1 0,000 8 10,000-1 00,000 11 ~100.000 47
% Proteins 7
24 22 36
% DNA 10 21 15 32
Statistical Results SRT=16 days MWD(dalt0ns) % TCHO %Proteins %DNA
Average S.D. Average S.D. Average S.D. c l O00 1 3 4 3 6 3 1 O00 - 10, O00 8 2 12 8 11 7 10,000 - 100,000 10 3 19 5 10 5 >100,000 72 18 62 18 69 25
SRT=20 days Dtae MWD (daitons) %TCHO
07-Jan-98 4 ,000 7 1,000-1 0,000 13 10,000-1 00,000 12 ~100,000 39
Statistical Results SRT=20 days %TCHO %Proteins %DNA
MWD(daltons) Average S. O. Average S. D. Average S.D. c l 000 3 3 5 4 6 2 1000 - 10,000 11 2 14 6 11 4 10,000 -1 00,OO 14 3 14 3 10 3 >100.000 65 18 63 18 73 11
* S. D. - Standard Deviation.
Appendix F-1 Contact Angle Data (unit: degrees)
Date SRT (day) 16-Sep-97 22-Sep-97 04-NOV-97 05-NOV-97 21 -Jan-98 28-May-98 02-Feb-99
Contact Angle 4 23.9 4 26.2 4 20.7 4 22.2 4 29 4 29 4 25
Sta tistical Results SRT=4 days Mean 25.14285714
Standard Error 1.203537069 Median 25 Mode 29 Standard Deviation 3.1 84261 894 Sample Variance 10.1 3952381 Kurtosis -1 -26509569 S kewness O.OW103316 Range 8.3
- Minimum 20.7 Maximum 29 Sum 1 76 Count 7 Confidence Leve1(95.0%) 2.944953228
Date SRT (day) 17-Sep-97 24-Sep-97 05-NOV-97 22- J a n-98 20-Fe b-98 28-May-98 29-May-98 02-Fe b-99
Contact Angle 16 38 16 29.3 16 31 16 37.3 16 26.3 16 39.1 16 38 16 47
Statistical Results SRT=16 days Mean 35.75
Standard Error 2.329086271 Median 37.65 Mode 38 Standard Deviation 6.587650784 Sample Variance 43.39714286 K U ~ ~ O S ~ S -0.0909988 Skewness O. 1 78900968 Range 20.7 Minimum 26.3 Maximum 47 Sum 286 Count 8 Confidence Leve1(95%) 5.56321 451 2
Date SRf (&Y) Contact Angle 16--97 9 22.4 22-Sep97 9 20 WNOV-97 9 15.5 21 -Jan-= 9 14.5 28-May-98 9 11.6 02-Feb99 9 15
Statistical Results SRT* days Mean 16.5
Standard Error 1.61 5755757 Median 15.25 Mode W / A Standard Deviation 3.9577711 54 Sample Variance 1 5.664 Kurtosis 4.65726349 S kewness 0.572 1 30279 Range 10.8 Minimum 11.6 Maximum 22.4 Sum 99 Count 6 Confidence Leve1(95.0%) 4.1 5342561 3
Date SRT (day) 17-Sep97 24-Sep-97 05-NOV-97 20-Feb98 21 -Jan-98 29-May-98 02-Feb-99
Contact Angle 20 37 20 3 1 20 32 20 30.8 20 45 20 40 20 42
Statistical Results Mean 36.828571 43
SRT=20 days Standard Error 2.167038168 Median 37 Mode #N/A Standard Deviation 5.733444074 Sample Variance 32.87238095 Kurtosis -1 -81 8361 14 Skewness 0.230448834 Range 14.2 Minimum 30.8 Maximum 45 Sum 257.8 Count 7 Confidence Leve1(95.0%) 5.302555253
Date SRT (days) 27-Mar-97 12-May-97 09-May-97 1 2-May-97
Statisticai Results
Contact Angle 6 21.5 6 23
12 30 12 29
SRT (days) Average Contact Angle SD 4 25 9 17
12 30 16 36 20 37
'SC)-- Standard deviation
Anova: Single Factor
SUMMARY Groups Count Sum Average Variance
SRT=4 days 3 83 27.66667 5.333333 SRT=Sdays 2 26.6 13.3 5.78 SRT=16 days 4 150.5 37.625 72.9225 SRT=20 davs 3 127 42.33333 6.333333
ANOVA ource of Variafio SS df MS F P-value F crit Between Groups 1201 -808 3 400.6028 12.92888 0.007954 4.06618 Within Groups 247.8808 8 30.9851
Total 1449.689 11
Appendu F-2 Contact Angle and Extracellular Polymeric Substances (CA: Contact Angle, degrees; EPS unit: mg/g VSS)
Date SRT(days) CA TCHO Rotan DNA TacaiEPS RaeiriTTcHO Rateiril(rCW+ONA)
Correlations (test(bio).sta) Marked correlations are significant at p < -05000 N=32 (Casewise deletion of missing data)
TCHO Protein DNA Total EP ProterVTCHO ProteinI(TCH0 +DNA) CA -0.5 0.27 0.1 5 0.04 0.58 0.53
e W H
Appendix F-3 Correlations between Contact Angle and Surface Charge, Sludge Volume Index, and Effluent Suspended Solids
(C A-Contact Angle, degrees; SC-Surface Charge. meqJg VSS ; SVI-S ludge Volume Index, mL/g MLSS; ESS-Ef3iuent Suspended Solids, m d )
Date SRT (days) CA SC SV1 ESS
Correlations (data-sta) Marked correlations are significant at p c .O5000 N=31 (Casewise deletion of missing data)
SC SV1 ESS CA 0.8662 0.045 -0.71 191 8
* H
Appendix G-1 Surface Charge Data
Date SRT (day) SC (meq-lg VSS) Date SRT (day) SC (meq.1g VSS) 04-Jun-97 4 -0.434 23-Mar-97 6 -0.208 23-Jun-97 4 -0.287 29-Mar-97 6 -0.312 27-Jun-97 4 -0.37 05-Apr-97 6 4 - 4 4 29-Jul-97 4 -0.543 29-Apr-97 6 -0.45
18-Sep-97 4 -0.483 1 5-May-97 6 4-52 23-Sep-97 4 -0.489 30-Sep-97 4 -0.302 04-Nov-97 4 -0.5 Statistical Resuits of Sudace Charge 21 -Jan-98 4 -0.26 SRT=6 days 1 0-Fe b-98 4 -0.25 Mean -0.386 1 6-Fe b-98 4 -0.25 Standard Enor 0.055735088 26-Fe b-98 4 -0.265 Median -0.44 24-Mar-98 4 -0.372 Mode #NIA 1 9-May-98 4 -0.435 Standard Deviation O. 124627445 28-May-98 4 -0.351 Sample Variance 0.01 5532 30-Jun-98 4 -0.497 Ku rtosis -0.841447718 O 1 -Fe b-99 4 -0.3 Skewness 0.692953797
Range 0.31 2 Statistical Results of Surface Charge Minimum 4-52
SRT=4days Maximum -0.208 Mean -0.3690625 Sum -1 -93 Standard Error 0.025341 742 Count 5 Median -0.3605 Mode -0.25 Standard Deviation 0.101 366969 Sample Variance 0.01 0275263 Kurtosis -1.378041 775 S kewness -0.358274044 Range 0.293 Minimum -0.543 Maximum -0.25 Sum 4.423 Count 17
Date SRT (day) SC (meq./g VSS) Date SRT (day ) SC (meq./g VSS) 29-Jul-97 9 -0.421 29-Mar-97 12 -0.185
1 8-Sep-97 9 -0.534 23-Mar-97 12 -0.203 23-Sep-97 9 -0.534 05-Apr-97 12 -0.361 30-Sep-97 9 -0.348 29-Apr-97 12 -0.348 04-Nov-97 9 -0.6 1 S-May-97 12 -0.312 21 -Jan-98 9 -0.64 04-J un-97 12 4.389 28-Jan-98 9 -0.64 23-Jun-97 12 -0.295 06-Feb-98 9 -0.38 27-Jun-97 12 -0.35 10-Feb-98 9 -0.32 16-Feb-98 9 -0.37
26-Feb-98 9 -0.252 Saît;sticai Results of Surface Charge 24-Mar-98 9 -0.34 SRT=12 days 19-May-98 9 -0.416 Mean -0.305375 28-May-98 9 -0.503 Standard Enor 0.02639598 30-Jun-98 9 -0.341 Median -0.33 0 1 -Fe b-99 9 -0.41 Mode #NIA
Standard Deviation 0.0746591 06 Statistical Results of Surface Charge Sampfe Variance 0.005573982
days Mean -0.434333333 Standard Error 0.031 340627 Median -0.41 Mode -0.64 Standard Deviation 0.121381728 Sample Variance 0.014733524 Kurtosis -0.74926443 1 Skewness Range Minimum Maximum Surn Count
Kurtosis
Skewness Range Minimum Maximum Sum Count
Date SRT (day)
23-Mar-97 29-Mar-97 05-Apr-97 29-Apr-97 15-May-97 04-Jun-97 23-Jun-97 27-Jun-97 29-Jul-97
18-Sep-97 23-Sep-97 30-Sep-97 04-NOV-97 21 -Jan-98 1 2-Fe b-98 26-Feb-98 24-Mar-98 19-May-98 28-May-98 30-Jun-98 0 1 -Feb-99
SC (meq-/g VSS) Date SRT (day) 16 -0.155 23-Mar-97 16 -0.1 73 29-Mar-97 16 -0.344 05-Apr-97 16 -0.325 29-Apr-97 16 -0.164 1 5-May-97 16 -0.1 95 04-Jun-97 16 -0.284 23-Jun-97 16 -0.427 27-Jun-97 16 -0.138 29-Jul-97 16 -0.1 8 18-Sep-97 16 -0.167 23-Sep97 16 -0.212 30Sep-97 16 -0.195 04-NOV-97 16 -0.18 21-Jan-98 16 -0.27 09-Feb-98 16 -0.246 f2-Feb-98 16 -0.284 24-Mar-98 16 -0.242 1 9-May-98 16 -0.188 28-May-98 16 -0.096 30-Jun-98 16 -0.165 01-Feb-99
SC (meq./g VSS) 20 -0.159 20 -0.207 20 -0.317 20 -0.261 20 -0.272 20 -0.39 20 -0.351 20 -0.392 20 -0.354 20 4.2 20 -0.1 73 20 4.2 20 -0.21 20 4.13 20 -0.13 20 -0.16 20 -0.09 20 -0.1 35 20 -0.109 20 4.143 20 -0.145
Statistical Results of Sudace Charge Statistical Results of Surface Charge SRT=16 days SRT=20 days
Mean -0.2225 Mean -0.21 64 Standard Enor 0.01 7948684 Standard Error 0.02 1856879 Median -0.195 Median -0.1865 Mode -0.195 Mode -0.13 Standard Deviation 0.080268956 Standard Deviation 0.097746934 Sample Variance 0.0064431 O5 Sample Variance 0.009554463 Kurtosis 0.781 851 358 Kurtosis -0.93967507 S kewness -0.931 41 0738 Skewness -0.6678881 34 Range 0.331 Range 0.302 Minimum -0.427 Minimum -0.392 Maximum -0.096 Maximum -0.09 Sum 4.53 Surn 4.528 Count 21 Count 21
Anova: Single Factor
SUMMARY
Groups Count Sum Average Vadance SRT=4 days 17 4.423 -0.369û6 0.010275 SRT=9 days 16 -7.045 -0.43433 0.014734 S RT= 1 2 days 8 -2.443 -0.30538 0.005574 SRT=l6 days 21 -4.53 -0.2225 0.006443 SRT=20 days 21 4.528 4.2164 0.009554
ANOVA
Source of Variation SS df MS F P-value F crït Between Groups 0.607087 4 0.1 51 772 1 5.96757 1 -82E-09 2.495391 Within Groups 0.70337 74 0.009505
Total 1 -31 0457 78
Appendix G-2 Surface Charge and Extracellular Poiymeric Substances
(SC-Surface charge, meq./g VSS; EPS unit: mglg VSS)
Date SRT(days) SC TCHO Prote DNA AP TCHO Proteinl Protein Total EPS ins +DNA /
Note: EPS data with dectable APS was not used for statistical analysis Correlations (datasta) Marked correlations are significant at p < .O5000 N=38 (Casewise deletion of missing data)
TCHO Protein DNA TCHO +DNA Proteinl Protein/ Total EPS (TCHO+DNA) TCHO
SC -0.38 0.43 O. 16 4.32 0.53 0.55 0.16 + t t +
Appendix G-3 Surface Charge, Sludge Volume Index and Effluent Suspended Solids (SC-Surface Charge, meq./g VS S; ES S-Emuent Suspended Solids, mg/'; SVI- Sludge Volume Index, rnL& ~ L s s )
Date SUT (days) SC ESS
Correlations (data.sta) Marked correlations are significant at p < -05000 N=75 (Casewise deletion of missing data)
ESS SV1 SC -0.72395 NSC
Appendix H Dissociation Constants (Unit: turbiditylg DShvashing; Sampling period: January-February, 1999)
H-1: Effect of pH on Dissociation Constants
R I (SRT=4 days) PH
1.77 2.7
3.59 4.29
5.9 7
8.1 9.31
R3(SRT=16 days) PH
1.77 2.7
3.59 4.29
5.9 7
8.1 9.31
* DC-Dissociation Constant; SD-Standard Deviation.
H-2: Effects of Ionie Strengtb and Cation Valency on Dissociation Constants
R1 (SRTddays) KCI (mM)
O 0.5 5 25 50 300 600
R2 (SRT=9 days) KCi (mM)
O 0.5 5 25 50 300 600
R3 (SRT=16 days) KCI (mM)
O 0.5
5 25 50
300 600
R4 (SRT=ZOdays) KCI (mM)
O 0.5
5 25 50
300 600
Statistical Results-Ionic Strength and Cation Valency
DC-Dissociation Constant; SD-Standard Deviation; 1- Ionic Strength (moüi,).
ES-3: Effect of EDTA concentration on Dissociation Constants
R i (SRT=4days) EDTA (mg/L)
O 20 40 100 300 600
R2 (SRT=Sdays) EDTA (mg/L)
O 20 40 100 300 600
R3 (SRT=lGdays) EDTA (mg/L)
O 20 40
1 O0 300 600
R4 (SRT=20days) EDTA (mg/L)
O 20 40 100 300 600
S tatistical Results-EDTA SRT=4day
EDTA BC SD
* DC-Dissociation Constant; SD-Standard Deviation.
H-4: Effect of Urea Concentration on Dissociation Constants
RI (SRT=4 days) Urea (M)
O 2 5 8
11
R 2 (SRT=Sdays) Urea (M)
O 2 5 8
11
R3 (SRT=lGdays) Urea (M)
O 2 5 8
11
R4(SRT=20 days) Urea (M)
O
Statistical Results-Urea
SRT=4days SRT=9 days Urea (M) OC SD DC SD
SRT=l6 days SRT=20days DC SD DC SD
* DC-Dissociation Constant; SD-Standard Deviation.
