THE EFFECT OF PHYSICOCHEMICAL PROPERTIES OF SECONDARY TREATED WASTEWATER … · 2014. 3. 5. · ii...

THE EFFECT OF PHYSICOCHEMICAL

PROPERTIES OF SECONDARY TREATED

WASTEWATER FLOCS ON UV DISINFECTION

By

Yaldah Azimi

A thesis submitted in conformity with the requirements

for the degree of Doctor of Philosophy,

Department of Chemical Engineering and Applied Chemistry

University of Toronto

© Copyright by Yaldah Azimi (2013)

ii

The Effect of Physicochemical Properties of Secondary Treated Wastewater Flocs on UV

Disinfection

Doctor of Philosophy 2013

Yaldah Azimi

Department of Chemical Engineering and Applied Chemistry

University of Toronto

ABSTRACT

The microbial aggregates (flocs) formed during secondary biological treatment of

wastewater shield microbes from exposure to ultraviolet (UV) light, and decrease the

efficiency of disinfection, causing the tailing phenomena. This thesis investigates whether

the formation of compact cores within flocs induces higher levels of UV resistance.

Moreover, it investigates the effect of secondary treatment conditions on the

physicochemical properties of flocs’, effluent quality, and UV disinfection performance.

Compact cores were isolated from the flocs using hydrodynamic shearing. The UV dose

response curves (DRC) were constructed for flocs and cores, and the 53-63 µm cores

showed 0.5 log less disinfectability, compared to flocs of similar size. Based on a

structural model developed for the UV disinfection of flocs, floc disinfection kinetics was

sensitive to the core’s relative volume, their density, and viability.

The UV disinfection and floc properties of a conventional activated sludge (CAS) system,

and a biological nutrient removal (BNR-UCT) system, including both biological nitrogen

and phosphorus removal, was compared. The 32-53 µm flocs and the final effluent from

the BNR-UCT reactor showed 0.5 log and 1 log improvement in UV disinfectability,

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respectively, compared to those from the CAS reactor. The BNR-UCT flocs were more

irregular in structure, and accumulated polyphosphates through enhanced biological

phosphorus removal. Polyphosphates were found to be capable of producing hydroxyl

radicals under UV irradiation, causing the photoreactive disinfection of microorganisms

embedded within the BNR-UCT flocs, accelerating their UV disinfection.

Comparing the UV disinfection performance and floc properties at various operating

conditions showed that increasing the operating temperature from 12 ºC to 22 ºC, improved

the UV disinfection of effluent by 0.5 log. P-Starved condition, i.e. COD:N:P of

100:10:0.03, decreased the average floc size and sphericity, both by 50%. Despite the

higher effluent turbidity of the P-Starved reactor, the final effluent’s UV disinfection

improved by at least 1 log compared to the P-Normal and P-Limited conditions. The

improvement in the floc and effluent disinfectability were accompanied by a decrease in

floc sphericity and a decrease in the number of larger flocs in the effluent, respectively.

iv

FOREWORD

I would like to express my sincere gratitude to the following people and institutions for all the people and institutions that supported me over the course of this research. I especially would like to thank: Prof. Ramin Farnood and Prof. Grant Allen, my supervisors, for their invaluable support and guidance over the past few years. I am thankful for their confidence in my work and their commitment to my education and professional development. Prof. Gideon Wolfaardt and Prof. Yuri Lawryshyn, members of my reading committee, for providing me with their feedback and helpful suggestions. The Natural Sciences and Engineering Research Council of Canada, and Trojan Technologies for financially supporting this project to completion. I would like to specifically thank Dr. William Cairns and Dr. Ted Mao for their constructive feedback The National Water Research Institute (NWRI), Environment Canada, in Burlington. They supported me through providing the facilities required for operating pilot scale experiments. Specifically, I thank Dr. Peter Seto and Scott Dunlop at the Water Science & Technology Directorate. The administrative and technical staff in the Department of Chemical Engineering & Applied Chemistry at the University of Toronto for their friendly support: Gorette Silva, Pauline Martini, Joan Chen, Daniel Tomchyshyn, Julie Mendonca, Anna Ho, Cindy Tam, Isabella Medina, Mary Butera, thank you all. The summer students and 4th year thesis students that helped me throughout my experiments: Lucile Paez, Benoit Barroso, Sara kirchner, Yao (Doria) Wang, Daphne Wilson, many thanks for all your hard work. And finally I would like to thank my friends and fellow lab mates for their moral support throughout my graduate studies.

To my mother. .

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v

TABLE OF CONTENT

1. INTRODUCTION................................................................................... 1

1.1. Ultraviolet Disinfection Resistance of Wastewater Flocs..................................................1

1.2. Hypotheses ............................................................................................................................2

1.3. Objectives..............................................................................................................................3

1.4. Significance ...........................................................................................................................4

1.5. Thesis Outline .......................................................................................................................5

2. LITERATURE REVIEW ........................................................................ 7

2.1. Introduction to Wastewater Treatment .............................................................................7

2.1.1. Conventional Activated Sludge Process............................................................................8

2.1.2. Biological Nutrient Removal Processes ............................................................................8

2.2. Floc Physicochemical Properties.......................................................................................11

2.2.1. Floc Composition ............................................................................................................11

2.2.2. Floc Structure ..................................................................................................................13

2.3. UV Disinfection...................................................................................................................14

2.3.1. UV Disinfection Mechanisms .........................................................................................15

2.3.2. The Effect of Wastewater Flocs on UV Disinfection (Tailing) .......................................16

2.3.3. Modeling the UV Dose Response ...................................................................................18

2.4. The Effect of Secondary Treatment Conditions on Floc Properties ..............................20

2.4.1. The Effect of Oxygen Levels on Sludge and Floc Properties..........................................20

2.4.2. The Effect of Sludge Retention Time (SRT) on Sludge and Floc Properties ..................22

vi

2.4.3. The Effect of Operational Temperature on Sludge and Floc Properties..........................24

2.4.4. The Effect of Influent Phosphorus Content on Sludge and Floc Properties ....................26

2.5. Summary.............................................................................................................................27

2.6. References ...........................................................................................................................27

3. THE EFFECT OF WASTEWATER FLOC STRUCTURE ON UV

DISINFECTION KINETICS ................................................................................... 36

3.1. Abstract...............................................................................................................................36

3.2. Nomenclature......................................................................................................................37

3.3. Introduction ........................................................................................................................38

3.4. A Structural Model for Floc UV Disinfection ..................................................................41

3.5. Materials and Methods ......................................................................................................45

3.5.1. Wastewater Samples........................................................................................................45

3.5.2. Shearing...........................................................................................................................46

3.5.3. UV Disinfection Assay....................................................................................................47

3.5.4. Optical Microscopy and Size Distribution ......................................................................48

3.5.5. Statistical Analysis ..........................................................................................................49

3.6. Results and Discussion .......................................................................................................49

3.6.1. Flocs UV DRC ................................................................................................................50

3.6.2. UV DRC of Compact Cores ............................................................................................54

3.6.3. The Role of Compact Cores in UV Disinfection Kinetics of Flocs.................................60

3.7. Conclusions .........................................................................................................................62

3.8. References ...........................................................................................................................63

vii

4. THE EFFECT OF BIOLOGICAL NUTRIENT REMOVAL ON UV

DISINFECTION KINETICS ................................................................................... 65

4.1. Abstract...............................................................................................................................65

4.2. Introduction ........................................................................................................................65


4.3.1. Sample Collection from the CAS and BNR-UCT Reactors ............................................68

4.3.2. Size Fractionation............................................................................................................70

4.3.3. BNR-UCT and CAS Floc Characteristics .......................................................................70


4.3.5. Mechanical Strength of the BNR-UCT and CAS Flocs ..................................................72

4.3.6. Composition Analysis .....................................................................................................73


4.4.1. Microscopy and Morphological Characteristics ..............................................................73

4.4.2. UV Disinfection Kinetics of the BNR-UCT and CAS Flocs...........................................76

UV Disinfection of BNR-UCT and CAS Mixed Liquor Flocs .................................................76

UV Disinfection of BNR-UCT and CAS Final Effluents .........................................................77

4.4.3. Mechanical Strength and Composition of the BNR-UCT and CAS Flocs ......................79

4.5. Conclusions .........................................................................................................................81

4.6. References ...........................................................................................................................82

5. EVIDENCE OF DISINFECTION BY ADVANCED OXIDATION WITHIN

WASTEWATER FLOCS UNDER UV IRRADIATION........................................... 84

5.1. Abstract...............................................................................................................................84

viii

5.2. Nomenclature......................................................................................................................85

5.3. Introduction ........................................................................................................................86


5.4.1. Sample Collection ...........................................................................................................88


5.4.3. Transmission Electron Microscopy .................................................................................89

5.4.4. Photoreactive Properties of Polyphosphates....................................................................89


5.5.1. Transmission Electron Microscopy .................................................................................90

5.5.2. UV Inactivation of the BNR-UCT Flocs and the Photoactive Role of Polyphosphates ..91

5.6. UV Disinfection Model Including the Effect of Advanced Oxidation............................93

5.6.1. Model Derivation ............................................................................................................93

5.6.2. Comparison with Experimental Data...............................................................................97

5.6.3. The Signifficance of Advanced Oxidation in Floc Disinfection......................................99

5.7. Conclusions .......................................................................................................................100

5.8. References .........................................................................................................................101

6. THE EFFECT OF ACTIVATED SLUDGE PROCESS CONDITIONS

ON UV DISINFECTION: TEMPERATURE, SLUDGE RETENTION TIME,

INFLUENT PHOSPHORUS LEVELS ................................................................. 103

6.1. Abstract.............................................................................................................................103

6.2. Introduction ......................................................................................................................104

6.3. Materials and Methods ....................................................................................................107

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6.3.1. Sample Collection .........................................................................................................107

6.3.2. Size Fractionation..........................................................................................................111

6.3.3. Floc Physicochemical Characteristics ...........................................................................111

Sludge Floc Structural Properties............................................................................................111

Mechanical Stability of Flocs .................................................................................................112

Sludge Floc Chemical Composition........................................................................................112

6.3.4. UV Disinfection Assay..................................................................................................113

6.3.5. Secondary Effluent Turbidity, UV Transmittance and Total Suspended Solids ...........114

6.4. Results and Discussion .....................................................................................................114

6.4.1. Floc Physicochemical Properties...................................................................................115

Size and Sphericity .................................................................................................................115

Mechanical Stability ...............................................................................................................116

Composition............................................................................................................................118

6.4.2. UV Disinfection of Mixed Liquor Flocs .......................................................................118

The Effects of Temperature and Sludge Retention Time........................................................118

The Effect of Influent Phosphorus Content.............................................................................120

6.4.3. Final Effluent Quality and UV Disinfection..................................................................123

The Effects of Temperature and Sludge Retention Time........................................................123

The Effect of Influent Phosphorus Content.............................................................................125

6.4.4. UV Dose Demand of Final Effluent ..............................................................................127

6.5. Conclusions .......................................................................................................................129

6.6. References .........................................................................................................................130

7. OVERALL DISCUSSION & CONCLUSIONS................................... 133

7.1. Identification of Compact Cores as UV Resistant Constituents...................................134

7.2. The Potential of Polyphosphates as Photooxidative Agents in Sludge.........................135

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7.3. Reducing Wastewater Flocs’ UV Resistance by Changing the Secondary Treatment

Operational Conditions ................................................................................................................................136

7.4. Conclusions .......................................................................................................................138

7.5. Recommendations for Future Work...............................................................................141

8. APPENDICES................................................................................... 144

8.1. Appendix 1: Diffusion-Reaction Equation of OH• in the Polyphosphate Accumulating

cells 144

8.1.1. References .....................................................................................................................145

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LIST OF TABLES

TABLE 3. 1. INACTIVATION RATE CONSTANTS (K1 AND K2 ) AND UV-RESISTANT FRACTION

(Β) OF UN-SHEARED FLOC SAMPLES. THE ± VALUES REPRESENT 95% CONFIDENCE

INTERVALS...........................................................................................................................................51

TABLE 3. 2. VOLUME AND NUMBER PERCENT OF CORES EXTRACTED FROM VARIOUS FLOC

SIZE FRACTIONS (VOLUME %=TOTAL CORE VOLUME/ TOTAL VOLUME OF UN-SHEARED

FLOCS, AND NUMBER %= TOTAL NUMBER OF CORES / TOTAL NUMBER OF UN-

SHEARED FLOCS. ................................................................................................................................53

TABLE 3. 3. INACTIVATION RATE CONSTANTS (K1 AND K2 ) AND UV-RESISTANT FRACTION

(Β) FOR CORE SAMPLES. THE ± VALUES REPRESENT 95% CONFIDENCE INTERVALS. .....57

TABLE 4. 1. SOME OF THE OPERATIONAL PARAMETERS COLLECTED OVER THE DURATION

OF OCTOBER 2010 TO JANUARY 2011 (N=26), N IS THE NUMBER OF MEASUREMENTS ....70

TABLE 4. 2. AVERAGE INFLUENT CHARACTERISTICS OF PRIMARY EFFLUENT FROM

SKYWAY WWTP, USED AS INFLUENT TO THE CONTINUOUS REACTORS, MEASURED

THROUGHOUT THE DURATION OF SAMPLING (OCTOBER 2010 – JANUARY 2011).............70

TABLE 4. 3. UV DISINFECTION KINETICS PARAMETERS FOR THE 32-53 µM FLOCS FROM THE

CAS AND BNR-UCT REACTORS .......................................................................................................77

TABLE 4. 4. MEDIAN 10%ILE AND 90%ILE SHEAR STRESS VALUES FOR FLOC BREAKAGE FOR

THE BNR-UCT AND CAS FLOCS (32-63µM FRACTION) ...............................................................79

TABLE 6. 1. OPERATIONAL CONDITIONS OF THE PILOT REACTORS. SRT: SLUDGE RETENTION

TIME, HRT: HYDRAULIC RETENTION TIME, T: TEMPERATURE ............................................109

TABLE 6. 2. THE MICRONUTRIENTS USED IN THE SYNTHETIC FEED USED FOR THE SMALL

REACTORS..........................................................................................................................................109

TABLE 6. 3. PERFORMANCE PARAMETERS OF THE SBR SYSTEM OVER THE PERIOD OF

STABLE OPERATING CONDITIONS. MLSS: MIXED LIQUOR SUSPENDED SOLIDS, VSS:

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VOLATILE SUSPENDED SOLIDS, ESS: EFFLUENT SUSPENDED SOLIDS, SVI: SLUDGE

VOLUME INDEX. ...............................................................................................................................115

TABLE 6. 4. THE EFFECT OF INCREASING SRT, T AND CHANGING FEED PHOSPHORUS LEVEL

ON FLOC SIZE, SPHERICITY, FRACTAL DIMENSION (D1 USING LOGAN & WILKINSON

1991), PROTEIN TO CARBOHYDRATE RATIO, AND MECHANICAL STABILITY (τ 50) ........116

TABLE 6. 5. DOUBLE EXPONENTIAL UV DISINFECTION KINETIC PARAMETERS FOR THE

ISOLATED MIXED LIQUOR FLOCS FROM THE LARGE AND SMALL SBRS ..........................122

TABLE 6. 6. EFFLUENT QUALITY DATA FOR ALL SEVEN PILOT REACTORS..............................124

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LIST OF FIGURES

FIGURE 1. 1. ILLUSTRATION OF A TYPICAL UV DOSE RESPONSE CURVE, NOTE TAILING AT

HIGHER UV DOSES ...............................................................................................................................1

FIGURE 2. 1. GENERAL SCHEMATIC OF WASTEWATER TREATMENT PROCESS, USING THE

ACTIVATED SLUDGE PROCESS AS SECONDARY TREATMENT.................................................7

FIGURE 2. 2. SCHEMATICS OF THE BNR PROCESS CONFIGURATIONS EXAMPLES. FOR MORE

EXAMPLES PLEASE REFER TO WEF & ASCE 2005 .......................................................................10

FIGURE 2. 3. SUGGESTED SCHEMATIC MODELS OF ACTIVATED SLUDGE FLOCS (SHENG ET

AL. 2006, LIAO ET AL. 2002, URBAIN ET AL. 1993) .......................................................................14

FIGURE 3. 1. ILLUSTRATION OF A TYPICAL UV DOSE-RESPONSE CURVE OF WASTEWATER .39

FIGURE 3. 2. SCHEMATIC DIAGRAMS SHOWING MODEL SPHERICAL FLOCS: (A, B) TYPE 1 OR

UV-SUSCEPTIBLE FLOC, (C) TYPE 2 OR UV-RESISTANT FLOC, AND (D) ISOLATED

COMPACT CORE. THE SMALL WHITE CIRCLE INDICATES THE LOCATION OF THE MOST

SHIELDED VIABLE INDICATOR ORGANISM.................................................................................43

FIGURE 3. 3. SCHEMATIC OF THE COUETTE FLOW APPARATUS.....................................................47

FIGURE 3. 4. UV DOSE-RESPONSE CURVES (DRCS) FOR FLOC SAMPLES OF DIFFERENT SIZE

FRACTIONS. SOLID CURVES IN FIGURE REPRESENT THE FITTED DOUBLE-

EXPONENTIAL MODEL. THE DATA PRESENTED INCLUDES REPLICATES OF THE SAME

EXPERIMENT OVER THREE MONTHS WITH AT LEAST FOUR SAMPLINGS. .........................50

FIGURE 3. 5. CORRELATION BETWEEN THE VALUES OF Β AND K1 FOR FLOCS OF VARIOUS

SIZES ......................................................................................................................................................51

FIGURE 3. 6. THE CALCULATED UV DRC USING DOUBLE EXPONENTIAL MODEL FOR

VARIOUS VALUES OF Β. THE INACTIVATION RATE CONSTANTS K1 AND K2 ARE 0.23 AND

0.012 CM2/MJ, RESPECTIVELY. .........................................................................................................53

FIGURE 3. 7. PARTICLE SIZE DISTRIBUTION OF FLOCS (53-63µM AND 75-106µM) AND THE

COMPACT CORES LARGER THAN 45µM EXTRACTED FROM EACH SIZE FRACTION .........54

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FIGURE 3. 8. WHITE LIGHT OPTICAL MICROSCOPY IMAGES OF (A) SINGLE CORE AND FLOC

WITH SIMILAR SIZE, (B) UN-SHEARED FLOC SAMPLE, AND (C) A CORE SAMPLE. NOTE

DIFFERENCES IN THE MORPHOLOGY AND LIGHT TRANSMITTANCE (DENSITY)

BETWEEN CORES AND FLOCS. ........................................................................................................55

FIGURE 3. 9. UV DOSE-RESPONSE CURVES FOR CORE SAMPLES OF DIFFERENT SIZE

FRACTIONS. SOLID CURVES REPRESENT THE FITTED DOUBLE EXPONENTIAL MODEL. 56

FIGURE 3. 10. THE ESTIMATED UV DRC OF THE SPHERICAL CORES WITH DIAMETERS OF 40,

60, 70, AND 80ΜM FOR: A) CORES WITH A VIABLE LAYER OF 20 ΜM, AND B) CORES

WITHOUT VIABLE LAYER. UV ABSORBANCE IS 1000 CM−1......................................................59

FIGURE 3. 11. UV DRC SPHERICAL FLOCS FOR VARIOUS 0%, 10%, 40%, AND 70% CORE

VOLUME FRACTIONS. CORE DIAMETER IS GIVEN IN PARENTHESIS. FLOC DIAMETER IS

60ΜM, AND THE UV ABSORBANCE OF FLOC AND CORE ARE ASSUMED TO BE 500 CM-1

AND 1000CM-1, RESPECTIVELY. .......................................................................................................62

FIGURE 4. 1. UV DRCS OF UN-SHEARED FLOCS AND CORES OF SIMILAR SIZE............................66

FIGURE 4. 2. SCHEMATIC OF THE BNR-UCT (A), AND THE CAS (B) PROCESS................................69

FIGURE 4. 3. TYPICAL OPTICAL MICROSCOPY IMAGES OF THE MIXED LIQUOR FLOCS

COLLECTED FROM THE CAS AND BNR-UCT REACTORS ..........................................................74

FIGURE 4. 4. SLUDGE FLOC SIZE DISTRIBUTION (A) AND SPHERICITY (B) OF THE CAS AND

BNR-UCT FLOCS..................................................................................................................................75

FIGURE 4. 5. NORMALIZED UV DOSE RESPONSE CURVES (DATA POINTS AND FITTED

DOUBLE EXPONENTIAL MODEL) FOR THE 32-53 µM FLOCS FROM THE CAS AND BNR-

UCT FLOCS ...........................................................................................................................................77

FIGURE 4. 6. NORMALIZED (A) AND ACTUAL (B) UV DRCS FOR THE SECONDARY EFFLUENTS

COLLECTED FROM BOTH THE CAS AND BNR-UCT REACTORS. .............................................78

FIGURE 4. 7. BREAKAGE PERCENTAGE OF THE CAS AND BNR-UCT FLOCS (32-63 µM) UNDER

VARIOUS DEGREES OF HYDRODYNAMIC SHEARING WITH THE COUETTE FLOW............80

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FIGURE 5. 1. TRANSMISSION ELECTRON MICROSCOPY (TEM) IMAGES FROM THE FLOCS

COLLECTED FROM THE CAS AND BNR-UCT PROCESSES. ........................................................91

FIGURE 5. 2. DISCOLORATION OF METHYLENE BLUE (MB) UNDER UV LIGHT WITH NA3P3O9

(WITH AND WITHOUT METHANOL (MEOH)), AND TIO2 ............................................................93

FIGURE 5. 3. THE ESTIMATED UV DRC FOR THE SHELL-CORE MODEL WITH ONLY DIRECT UV

INACTIVATION (REPRESENTING CAS FLOCS), AND FOR THE HOMOGENEOUS MODEL

WITH DIRECT UV INACTIVATION AND ADVANCED OXIDATION (REPRESENTING BNR

FLOCS). UV ABSORBANCE OF EPS: 1000 CM−1, FLOC SIZE: 48 ΜM. THE SOLID LINE IS

ESTIMATED UV DRCS AND THE SCATTERED DATA POINTS ARE EXPERIMENTAL DATA.

.................................................................................................................................................................99

FIGURE 5. 4. THE EFFECT OF •OH ON THE DISINFECTABILITY OF HOMOGENEOUS SPHERICAL

FLOCS WITH UV LIGHT, AND THE ESTIMATED EFFECT OF INCREASING •OH

CONCENTRATION. (FLOC SIZE: 48 ΜM).......................................................................................100

FIGURE 6. 1. SCHEMATIC OF THE LARGE-20L (A), AND SMALL-2L (B) REACTORS ....................110

FIGURE 6. 2. PERCENTAGE OF BREAKAGE OBTAINED AT VARIOUS DEGREES OF SHEARING,

FOR (A) THE LARGE (20 L) REACTORS, AND (B) FOR THE SMALL (2 L) REACTORS .........117

FIGURE 6. 3. THE AVERAGED UV DRC OF (A) 32-53 µM AND (B) 75-90 µM ACTIVATED SLUDGE

FLOCS COLLECTED FROM THE SRT20T12, SRT7T12, SRT7T22, AND SRT20T22 REACTORS.

THE LINES REPRESENT THE FITTED DOUBLE EXPONENTIAL MODEL. ..............................120

FIGURE 6. 4. THE AVERAGED UV DRC OF THE 45-75 ΜM ACTIVATED SLUDGE FLOCS

COLLECTED FROM THE P-NORMAL, P-LIMITED, AND P-STARVED REACTORS. THE LINES

REPRESENT THE FITTED DOUBLE EXPONENTIAL MODEL ....................................................121

FIGURE 6. 5. UV DRC DOUBLE EXPONENTIAL MODEL PARAMETERS (Β &K1) FOR MIXED

LIQUOR FLOCS WITH DIFFERENT PERCENTAGES OF FLOCS WITH SPHERICITY LARGER

THAN 0.5. PLEASE NOTE THAT THE DOUBLE EXPONENTIAL PARAMETERS ARE

REPORTED FOR THE 75-90 ΜM, AND 45-75 ΜM FLOCS FOR THE LARGE AND SMALL

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REACTORS, RESPECTIVELY. THE PERCENTAGE OF FLOCS WITH SPHERICITY LARGER

THAN 0.5 IS ESTIMATED FOR 35-120 ΜM FLOCS. ......................................................................123

FIGURE 6. 6. THE UV DRCS OF TREATED EFFLUENTS BASED ON ACTUAL CFU / 100 ML. ......125

FIGURE 6. 7. UV DRCS (ACTUAL CFU/100ML-A, AND NORMALIZED-B) OF FINAL EFFLUENTS

COLLECTED FROM THE P-NORMAL, P-LIMITED, AND P-STARVED CONDITIONS ............127

FIGURE 6. 8. UV DOSE DEMAND OF THE FINAL EFFLUENTS COLLECTED FROM ALL SBRS. .129

FIGURE 7. 1. SUMMARY OF THESIS PROJECT PLAN...........................................................................134

1

1. Introduction

1.1. Ultraviolet Disinfection Resistance of Wastewater Flocs

Ultraviolet (UV) disinfection is a well established technology for disinfecting

secondary wastewater effluents. However, the effectiveness of UV disinfection decreases

in the presence of suspended microbial flocs formed in the activated sludge process (Qualls

et al. 1985, Das 2001). A decrease in the UV disinfection efficiency at high UV doses can

be detected in a typical UV dose-response curve (DRC - a plot of the logarithm of

microbial counts or log surviving fraction vs. UV dose) by a transition from the initial

linear decrease to a near plateau referred to as “tailing” (Figure 1. 1).

Figure 1. 1. Illustration of a typical UV dose response curve, note tailing at higher UV doses

It is believed that the presence of particle-associated bacteria in the effluent causes

the tailing effect, and that increasing particle size has an adverse effect on the efficiency of

UV disinfection (Cerf 1977, Qualls et al. 1985, Loge et al. 1996, Tan 2007). The double

exponential model may be used to describe this effect (Farnood 2005):

2

DkDkee

N

N21)1(

0

−− +−= ββ

Where,

N: number of colonies per 100 mL formed from the survived coliform upon

irradiation with a dose of D

N0: number of colonies per 100 mL of the effluent prior to UV irradiation

D: the delivered ultraviolet dosage, in mJ/cm2

k1: inactivation constant for UV susceptible particles

k2: inactivation constant for UV resistant particles

β: the ratio of UV resistant particles to the total number of viable particles

There is currently a lack of knowledge on the fundamental mechanism of UV

disinfection in the presence of microbial flocs, and the influence of floc structure and

composition on UV disinfectability. More specifically, which constituent of the flocs is the

source of UV resistance, whether or not floc chemical composition affects UV disinfection,

and in what ways the secondary treatment operational conditions affect the UV disinfection

kinetics of flocs and secondary effluents.

1.2. Hypotheses

Several researchers have suggested that wastewater sludge flocs have a two-layer

structure consisting of a compact core surrounded by a loose outer shell (Eriksson et al.

1992, Liao et al. 2002, Yuan 2007). Yuan et al. exposed flocs to hydrodynamic shear in a

Couette flow cell and reported that shear strength of core could be several times greater

than that of the loose shell (Yuan and Farnood 2010). The effect of secondary treatment

conditions on nutrient removal and sludge properties have been well-studied (Wilén et al.

3

1999, Liao et al. 2002, Lodge et al. 2002, Scott et al. 2005, Liao et al. 2006, Liu and Liss

2007, Xie &Yang 2009, Wilén et al. 2010)

The hypotheses of this project are:

1- The presence of mechanically denser “cores” in wastewater flocs can increase

their UV resistance and induce a more severe tailing effect in the UV dose response curve.

2- Polyphosphate accumulation in flocs during enhanced biological phosphorus

removal process can alter the mechanism of UV disinfection, and hence change the UV

inactivation kinetics.

3- Varying the secondary process operational conditions changes the flocs’

sphericity and that will affect their resistance to UV disinfection and the effluent UV dose

demand.

1.3. Objectives

The following objectives were developed to test the hypotheses:

1- To Isolate the compact cores from flocs and compare their UV disinfection

kinetics with that of flocs.

2- To generate flocs under enhanced biological phosphorus removal conditions and

study the UV disinfection mechanisms and kinetics, and compare that to flocs collected

from a conventional activated sludge system. Two parallel pilot-scale primary effluent

treatment systems (a conventional and an enhanced biological nutrient removal system)

were used for this study. Biological nutrient removal (BNR) systems, which incorporate

anoxic and anaerobic conditions, are known to change flocs size distribution and structure

(Crocetti et al. 2000, Martins et al. 2004, Moon et al. 2004).

4

3- To generate flocs of different structure and composition, and compare their UV

disinfectability, mechanical strength, and composition. Manipulation of the secondary

treatment process at lab and pilot scale was used as a tool to vary the floc composition and

structure. A set of four sequencing batch reactors (SBR) was used to investigate the effect

of sludge retention time and the operating temperature. And a set of three SBRs was used

to assess the effect of influent phosphorus levels.

A transition in floc structural properties occurs in the SRT range of 9 to 12 days

(Liao 2000). Flocs formed at higher SRTs settle better and their EPS has a higher protein

to carbohydrate ratio (Liao et al. 2001, Scott et al. 2005, Liao et al. 2006). Increasing

operating temperature from 10 º C to 20 º C improves the compaction and settling

characteristics of sludge. However operating at temperatures higher than 35 º C causes

sludge deflocculation (Morgan-Sagastume & Allen 2003, Wilén et al. 2010). Limiting

flocs on feed phosphorus (COD:N:P of 100:5:0.3) significantly affects structural integrity,

size, and settling properties (Liu & Liss 2007, Dumitrache 2008, Brei 2011).

1.4. Significance

Further understanding of the mechanism of UV disinfection and microorganism

protection against UV light by wastewater particles is useful in designing new

upstream/downstream processes to improve UV treatment. The findings of this work can

be used to design new treatment processes, or to modify existing processes to decrease the

energy usage for UV disinfection, hence securing the eco-effectiveness of utilizing treated

wastewater for reuse applications.

5

1.5. Thesis Outline

This thesis consists of eight chapters:

Chapter 1 provides a brief introduction on tailing in UV disinfection, and the

hypothesis and objectives of the study, in addition to putting the study in context in terms

of industrial significance.

Chapter 2 will first discuss a review of wastewater treatment processes and sludge

flocs’ physicochemical properties. This will be followed by a discussion on the

fundamentals of UV disinfection mechanisms and kinetics, and the impact of wastewater

particles on UV resistance. Moreover, the existing models for predicting UV disinfection

performance will be presented. Finally this chapter will summarize the effects of

wastewater treatment type and operational parameters on floc structure and composition.

Chapter 3 involves comparing the UV disinfection kinetics of flocs and compact

cores; a study intended to identify the main constituent in microbial flocs that induces

tailing.

Chapter 4 focuses on comparing the physicochemical properties and UV

disinfection kinetics of sludge flocs generated in two different types of secondary

wastewater treatment processes. The two processes compared in this chapter are: a

Conventional Activated Sludge (CAS) process, and a Biological Nutrient Removal

(Modified University of Cape Town, BNR-UCT) process.

Chapter 5 will present the effect of in-situ advanced oxidation, triggered by UV

exposure on the UV disinfection kinetics of wastewater flocs. A mathematical model is

presented in this chapter, which accounts for UV disinfection, both by direct DNA damage

and by photoreactive oxidation.

6

Chapter 6 deals with the effect of CAS operational parameters on floc

physicochemical properties, final effluent quality, and UV disinfection kinetics. The

operational parameters studied in this chapter include: temperature, sludge retention time,

and influent phosphorus levels.

Chapter 7 presents an overall discussion of the results, the main conclusions of this

study, and the recommendations for future work.

It should be noted here that the chapters of this thesis are based on papers, which

are either published or to be submitted. Therefore, the references are given in the end of

each chapter. Moreover, for each of the chapters including mathematical modeling, a

specific nomenclature is provided at the beginning of the chapter.

7

2. Literature Review

2.1. Introduction to Wastewater Treatment

Wastewater from municipalities must be treated to remove nutrients and reduce

microbial contamination, which if discharged into the environment could pose a risk to

public health. The treatment process usually involves three steps: primary, secondary, and

tertiary. The primary treatment is typically a physical treatment where the coarse and

floating solids are removed. It sometimes includes phosphorus removal by adding a

coagulant. Secondary treatment is generally a biological process to remove organic matter

through biochemical oxidation. In some cases, the biological process is modified to

remove phosphorus and/or nitrogen. Such a process is referred to as a biological nutrient

removal process. Disinfection of the secondary treatment process’s supernatant is one of

the main tertiary processes. A schematic of a typical wastewater treatment process is

shown in Figure 2. 1.

Figure 2. 1. General schematic of wastewater treatment process, using the activated sludge process as

secondary treatment.

8

2.1.1. Conventional Activated Sludge Process

The activated sludge process is the most common secondary biological treatment

process, in which microorganisms convert carbon into cell biomass and oxidize the organic

matter into CO2 and H2O (Metcalf & Eddy 2003). Through this process two goals are

achieved: (1) oxidation of the biodegradable matter in the aeration tank; (2) flocculation of

the biomass for effective separation from the effluent. The microbial floc mixture formed

in the aeration tank is referred to as activated sludge, and is separated from the effluent in a

clarifier or thickener. A portion of the thickened sludge is usually recycled back into the

reactor to increase the cell concentration in the aerated tank for more effective nutrient

removal.

Many wastewater treatment plants have been required to remove phosphorus and

nitrogen in addition to the carbonaceous nutrients. Secondary biological processes have

been modified, and new processes have been developed to achieve a wider range of

nutrient removal. Decreasing the sludge production and energy usage and enhancing

nutrient removal have been the main drivers for developing new treatment processes.

Aerobic granulation, membrane filtration, trickling filters, rotating biological contactors,

and a wide variety of enhanced biological nutrient removal systems are all examples of the

relatively newly developed processes.

2.1.2. Biological Nutrient Removal Processes

In the past two decades, many wastewater treatment plants have been constructed

for the purpose of biological nutrient removal (BNR) (Trivedi 2009). In a BNR process,

anaerobic and anoxic zones are incorporated with aerobic zones for the simultaneous

9

removal of nitrogen and/or phosphorus, in addition to carbonaceous removal (i.e.,

BOD/COD). There are three processes, one or all of which may be involved in a BNR

system. These processes are: nitrification, denitrification (both processes together are

termed biological nitrogen removal), and biological phosphorus removal. Nitrification is

the oxidation of ammonia to nitrite and nitrate by aerobic autotrophic organisms that are

referred to as Ammonia Oxidizing Bacteria (AOB) and Nitrite Oxidizing Bacteria (NOB).

Operating at higher sludge retention times and higher temperatures promotes the growth of

AOB and NOB (WEF & ASCE, 2005).

NH4+ + 3/2 O2 (AOB) ─> NO2

─ + 2H + + H2O

NOB converts nitrite to nitrate:

NO2─ + 1/2 O2 (NOB) ─> NO3

─

Denitrification follows nitrification for complete nitrogen removal. It occurs under

anoxic conditions where heterotrophic bacteria utilize readily biodegradable organic matter

(for example CH3OH) to reduce nitrate to nitrogen gas:

6 NO3─ + 5 CH3OH ─> 3 N2 (g) + 5 CO2 + 7 H2O + 6 OH

─

Biological phosphorus removal is achieved by cycling between anaerobic and

aerobic conditions to promote the growth of Phosphorus Accumulating Organisms (PAOs)

e.g. Accumulibacter Phosphatis (Gu et al. 2008). In the anaerobic stage, PAOs convert

readily biodegradable organics into energy rich carbon polymers called

polyhydroxylalkenoates (PHAs). The energy required for this process is gained through

breaking polyphosphate bonds, and therefore the phosphate concentration increased during

the anaerobic stage. In the aerobic zone, PAOs oxidize the PHAs to gain energy. More

PAOs are made under the aerobic conditions, and energy reserves are restored through

10

phosphate uptake and polymerization. The phosphorus is then removed from the system

through wasting sludge (WEF & ASCE 2005).

There are many configurations of BNR processes. Some are designed to remove

only phosphorus or nitrogen and others are designed to simultaneously remove both.

Examples of BNR processes include: Bardenpho process (for nitrogen removal), A/O

process (for phosphorus removal), and modified University of Cape Town configuration

(BNR-UCT) (simultaneous removal of phosphorus and nitrogen). Schematics of the BNR

process examples are provided in Figure 2. 2.

Return Activated Sludge

FeedClarifier

Effluent

Waste

Anaerobic Anoxic Anoxic Aerobic

CH3COO ¯ Mixed Liquor Recycle NO3-Recycle


Clarifier

Effluent

Waste

Anaerobic Aerobic

(Modified BNR-UCT Process)

(A / O Process)

Clarifier

Effluent

Waste

Anoxic Aerobic Anoxic Aerobic

Feed

Feed


(Four-stage Bardenpho Process)

NO3-Recyle

Figure 2. 2. Schematics of the BNR process configurations examples. For more examples please refer

to WEF & ASCE 2005

11

2.2. Floc Physicochemical Properties

In biological treatment of wastewater, microbial flocs are formed by flocculation of

the bacteria responsible for removing organic matter. This process occurs as a result of the

formation of extracellular polymeric substances (EPS) network in which the bacteria and

other organic and inorganic particles are embedded (Urbain et al. 1993, Bura et al. 1998,

Liao et al. 2001, Wilén et al. 2008). Filaments from filamentous bacteria serve as a

"backbone", and are required for the formation of larger and denser flocs (Jenkins et al.

1993). In an activated sludge system, a complex microbial ecology is present, which

includes bacteria, protozoa, viruses, and fungi (Bitton 1999, Ayarza et al. 2010). A good

balance between the filamentous and floc forming bacteria favors the formation of large,

dense and strong flocs with good settling characteristics. Overgrowth of filamentous

bacteria can cause filamentous bulking, and scarce growth of the floc-forming bacteria

causes dispersed growth (pin point flocs). Both conditions cause poor settleability and

thickening of sludge, resulting in a very turbid effluent with high concentration of

suspended solids (Jenkins et al. 1993).

2.2.1. Floc Composition

Microbial flocs are complex structures and are held together and mechanically

stabilized by the means of extra-cellular polymeric substances (EPS), salt bridges, polymer

bridging, filamentous backbones, hydrophobic interactions, or a combination of all (Pavoni

et al. 1972, Forster, 1985, Jorand et al. 1995, Liss et al. 1996, Bura et al. 1998, Liao et al.

2006). Through a variety of microscopy techniques, it has been suggested that the

presence of fine polymeric extracellular fibrils (4-6 nm in diameter) bridge the floc

12

components together and play a key role in the flocs structural integrity (Li & Ganczarczyk

1989, Liss et al. 1996).

On average activated sludge flocs consist of 90-98% water, and on a dry basis they

contain about 60-90% cellular organic material (Jenkins 1993), and 10-30% inorganic salts

and particles. The EPS form a three-dimensional, gel-like, hydrated, and often charged

biofilm matrix, in which the microorganisms and other constituents are embedded. They

play a key role in the formation of microbial aggregates, and come from natural secretions

of bacteria, cell lysis, and hydrolysis products (Flemming et al. 2001). EPS are composed

of carbohydrates, proteins, humic compounds, uronic acids and nucleic acids, with the

main constituents being proteins (frølund et al. 1996, Bura et al 1998, Liao et al. 2001).

Divalent cations such as Ca2+ and Mg 2+ are known to be involved in binding the EPS

(Urbain et al. 1992, Peeters et al. 2011).

The ratio of protein to carbohydrates of the sludge EPS is known to significantly

affect sludge settleability, and the flocs mechanical stability. In general, when the

carbohydrate content of sludge decreases and DNA content increases, the flocs have shown

poor settling characteristics (Schmid et al. 2003). The hydrophobicity of sludge flocs

appears to be affected by variations in the protein to carbohydrate ratio, and DNA content

in EPS (Jorand et al. 1998, Liao et al. 2001). Wilén et al. (2008) reported that

mechanically weaker flocs were formed when the sludge protein content and total EPS

increased. It has been suggested by previous researchers that EPS can be divided into two

distinct categories: the slime layer and the capsule layer. The capsular material adheres to

the cell, whereas the slime is either loosely bound to the cell or, unattached (Gehr & Henry

1983, Liss et al. 2005).

13

2.2.2. Floc Structure

Flocs are irregular in shape, highly porous, and have a large range of particle size.

Flocs size range from single bacterial dimensions (1-3 microns) to large aggregate sizes up

to 1mm, and the size distribution is often bimodal (Jorand et al. 1995, Yuan 2007). The

apparent density of sludge flocs is close to that of water (approximately 1.05 g/cm3) (Sears

et al. 2006). Li and Ganczarczyk (1987) estimated activated sludge floc porosity (range of

45-90%) based on settling velocity and size, and reported a negative correlation between

floc size and porosity. Fractal dimension is often positively related to porosity, indicating

that the highly porous flocs are very irregular in shape (Perez et al. 2006). The values of

fractal dimension reported for wastewater flocs lay in the range of 1.5- 2.5 (Bellouti et al.

1997, Guan et al. 1998). Schmid et al. (2003) and Jin et al. (2003) reported that flocs with

higher fractal dimension and lower filament index show improved settleability.

It has been proposed that wastewater flocs have a two-layer structure consisting of a

compact core surrounded by a loose shell (Eriksson et al. 1992, Liao et al. 2002, Sheng et

al. 2006, Yuan 2007). Liao et al. studied the chemical stability of flocs generated at

various sludge retention times (SRT), and suggested that at higher SRTs the flocs’ cores

are more compact and hydrophobic (Liao et al. 2002). By exposing wastewater flocs to

hydrodynamic shear in a Couette flow cell, Yuan and Farnood reported that shear strength

of compact cores could be several times greater than that of their loose shell (Yuan and

Farnood 2010). They suggested that the higher mechanical strength of compact cores may

be due to the densely packed structure of the core where microorganisms are held together

by EPS enmeshment and other intra-floc forces. Sheng et al. (2006) also studied the

stability of flocs under shear and proposed a multilayer structure with two distinct regions

14

(i.e. stability part and dispersible part). The outer part was suggested to be a loosely

entangled readily-extractable EPS fraction that is completely dispersed under shearing,

while the inner part is more stable and could not be dispersed except under extreme

conditions. Suggested floc structure schematics by previous researchers are demonstrated

in Figure 2. 3.

Sheng et al. (2006) Liao et al. (2002)

Urbain et al. (1993)

Figure 2. 3. Suggested schematic models of activated sludge flocs (Sheng et al. 2006, Liao et al. 2002,

Urbain et al. 1993)

2.3. UV Disinfection

Proper disinfection is required by most wastewater treatment plants to reduce

microbial contamination, prior to discharge. Many of the common disinfectants such as

chlorine, ozone, and peracetic acid have disadvantages such as increasing the organic

content of the final effluent, producing toxic by-products, being hazardous to the health of

handlers, or having a high cost (Wojtenko 2001, Gehr et al. 2003, Wang et al. 2003, Kitis

15

2004, Paraskeva et al. 2005). UV is comparatively a cleaner technology as it does not

require the addition of chemicals, and accomplishes microbial inactivation in short contact

time without any known toxicological effects (Blatchley et al. 1997, Matasci et al. 1999,

Asano et al. 2007). Moreover, the ratio of the UV dose required for inactivation of most

resistant bacterial spores and viruses, compared to that of E.coli, is only about 3.5-6,

whereas the same ratio for chlorine is as much as 100 (Cairns et al. 1993). UV has a few

disadvantages including: dependence on effluent transmittance that is affected by the

presence of suspended solids (Severin 1980, Loge et al. 1996), and the possibility of

photoreactivation (Bitton 1999). However, it should be noted here that the presence of

flocs in wastewater decreases the effectiveness of all disinfection methods (Dietrich et al.

2003).

2.3.1. UV Disinfection Mechanisms

UV light, specifically around the germicidal wavelengths (200-300 nm), inactivates

microorganisms by penetrating through the cell wall and altering their DNA (Jagger 1967,

Pfeifer et al. 2005, Piluso & Moffatt-Smith 2006). UV light damages the DNA by

dimerizing adjacent thymine molecules, and quenching further transcription of the cell’s

genetic code, hence preventing reproduction. Microorganisms can sometimes repair and

reverse the destructive effect of UV on their DNA though repair mechanisms known as

photoreactivation and dark repair (Linden et al. 2002, Oguma et al. 2004, Shang et al.

2009). UV light with a wider range of emission (i.e. light emitted by medium-pressure UV

lamps) causes inactivation of the enzymes involved in DNA repair, hence lowering the

chances of photoreactivation and dark repair (Oguma et al. 2002)

16

In 1985, Matsunaga et al. first reported microbial inactivation in water by contact

with TiO2-Pt catalyst under near UV (315-380 nm) exposure of 60-120 min (Matsunaga et

al. 1985). In 1988 they suggested that the intracellular dimerization of coenzyme A was

the root cause of decreasing respiratory activities that lead to cell death. Later on, TiO2 has

been used by many as a photocatalytic agent to enhance the UV disinfection performance

(Watts et al. 1995, Chen et al. 2009). The biocidal action of TiO2 has been attributed to

two major photochemical oxidants: hydroxyl radical (•OH), and reactive oxygen species

(Ireland et al. 1993, Huang et al. 2000). Maness et al. (1999) reported peroxidation of

polyunsaturated phospholipid component of the lipid membrane in the presence of TiO2

photocatalyst. They concluded that the loss of cell membrane structure, destructed by the

release of oxidative species, is the root cause of cell death when the photocatalytic TiO2

particles were outside the cells. Cho et al. (2004) measured the release of •OH in the

UV/TiO2 process and, and found an excellent linear correlation between [•OH]ss and the

rates of E.coli inactivation. These rates of inactivation showed that •OH is one to ten

thousand times more effective for E.coli inactivation than common disinfectants such as

chlorine and ozone.

2.3.2. The Effect of Wastewater Flocs on UV Disinfection (Tailing)

The efficiency of UV disinfection decreases in the presence of wastewater flocs,

which shield microorganisms by scattering and absorbing UV light (Jorand et al. 1998,

Loge et al. 1999, Das 2001). Absorbing and scattering UV light, by the flocs and the

matrix, can decrease the effective UV dose delivery to the microorganisms. The effect of

microorganisms’ particle association on UV disinfection efficacy of wastewater has been

17

well studied (Parker & Darby 1995, Gehr & Nicell 1996, Loge et al. 1996, 1999, 2001,

Emerick et al. 1999, 2000, Nelson 2000, Das 2001, Templeton et al. 2008). Wu et al.

showed that even small goethite particles can protect E.coli and Pseudomonas putida from

UV inactivation, when allowing 24 hr for the attachment of cells to particles (Wu et al.

2005). The effect of wastewater particles on protecting microorganisms is detected by a

decrease in the UV disinfection efficiency at high UV doses. UV DRCs of wastewater

effluents typically show two regions, an initial steep slope followed by a tailing region. In

1977, Cerf suggested that tailing occurs due to the existence of a mixed population with

different degrees of resistance to UV inactivation (Cerf 1977). Blatchley et al. (2001)

reported that when E.coli cells were isolated from the tailing region of the mother culture,

the daughter culture resulted in the same distribution of sensitivities to UV disinfection,

rejecting the hypothesis that UV resistance is genetically acquired.

Most researchers attribute the two regions in a DRC to the presence of free-

swimming microbes (initial steep slope) and particle associated microbes (tailing) (Emerick

et al. 2000, Das 2001). Liu et al. showed that floc particles formed by coagulation and

flocculation lead to lower (more than 1-log) inactivation of E.coli at UV doses ranging

from 10-40 mJ/cm2 compared to non particle associated E.coli at similar turbidity (Liu et al.

2007). Tan found that similar to the UV DRC of whole effluents, a suspension of

wastewater flocs with a narrow size distribution shows the above mentioned distinct

regions (Tan 2007). Since the effect of free-swimming microbes was eliminated in Tan’s

experiments, he attributed the initial steep slope and tailing regions in the UV DRC to the

inactivation of UV-susceptible flocs and the UV-resistant flocs, respectively. It has been

repeatedly reported that larger flocs have a more pronounced worsening effect on the

18

efficiency of UV disinfection (Batch et al. 2004, Blume & Neis 2004, Madge & Jensen

2006, Templeton 2005, Tan 2007, Kollu & Örmeci 2012). Therefore, prefiltration,

sonication, and hydrodynamic shearing have been among suggested methods of decreasing

floc size distribution and improving the efficiency of UV disinfection (Qualls et al. 1985,

Liltved &Cripps 1999, Gibson et al. 2008, Yong et al. 2009, Best 2012).

Several researchers have suggested that besides floc size, other factors such as floc

composition and structure, which affect the inter-floc absorbance of UV light, affect UV

disinfection. For instance, flocs containing humic substances were found to provide more

effective shielding of microorganisms compared to clay minerals and activated sludge flocs;

therefore, showing a stronger decrease in the efficiency of UV disinfection (Bitton et al.

1972, Templeton et al. 2005). Iron is a strong absorber of UV light (Darby et al. 1993,

Kozak et al. 2010); therefore, when embedded in the flocs or free in the solution, it is a

factor that could significantly reduce the efficiency of UV disinfection (Nessim & Gehr

2006). Loge et al. (1999) suggested that target organisms with a direct light pathway to the

bulk solution are easier to disinfect with UV light, which emphasizes the effect of floc

structure on UV disinfection.

2.3.3. Modeling the UV Dose Response

Microorganism populations are exposed to a distribution of UV radiation doses

depending both on their positioning in the UV reactor, and within the floc. The dose

distribution in the reactor affects the received UV dose at the surface of the flocs. However,

the effective dose received by a microorganism that is embedded within the floc is affected

by the flocs’ physiochemical properties.

19

Various kinetics models have been utilized to estimate the impact of wastewater

quality on the reactor performance and for effective reactor design. The one hit model

assumes that a single hit is sufficient to inactivate a microorganism (Zimmer 1961). The

probability of microorganism survival corresponds to the probability that the effective

cross-section of the microorganism escapes the photons. In this case the probability of

survival will correspond to first order kinetics with respect to UV dose:

Dke

N

N0

0

−=

where N is the number concentration of survived bacteria at dose D, and N0 is the number

concentration of bacteria at dose zero, and k0 is the inactivation constant. It should be

noted here that this model is applied in the absence of particulate matter. For particle

associated microorganisms the received dose within the flocs is less than D. Dose is a

product of UV intensity and exposure time, and as there is an exponential decay in UV

light intensity within the flocs (according to the Beer-Lambert’s law), therefore, the

effective dose decreases within the floc.

The double exponential model was later developed by Jagger (1977) that represents

the disinfection of a heterogeneous population with respect to radiation sensitivity. The

population consists of two groups, both inactivated in a single-hit fashion, but one group is

much more sensitive than the other. The model is as follows:

DkDkee

N

N21)1(

0

−− +−= ββ

where D is the delivered ultraviolet dose, N is the number of surviving microorganisms in

the irradiated sample, and N0 is the number of microorganisms in the unexposed sample.

Parameters k1 and k2 are the UV inactivation rate constants; k1 is the inactivation constant

20

for the group of microorganisms much more sensitive to irradiation, and k2 is the

inactivation constant of the less sensitive group of microorganisms. The parameter β

represents the ratio of microorganisms with lower sensitivity to UV disinfection to the total

number of microorganisms.

2.4. The Effect of Secondary Treatment Conditions on Floc

Properties

Altering the secondary treatment process operational conditions have shown to

affect both the flocs physicochemical characteristics, and the particle association of

bacteria (Loge et al. 2002, Scott et al. 2005, Liao et al. 2006). The most well studied

operational parameters in wastewater treatment, in terms of their effects on sludge

characteristics include: oxygen concentration, sludge retention time, operating temperature,

influent phosphorus levels, carbon source, and coagulant type. The effects of each of these

parameters are discussed in the following subsections.

2.4.1. The Effect of Oxygen Levels on Sludge and Floc Properties

Low concentrations of dissolved oxygen (DO) lead to the formation of porous flocs

with poor settling properties (Palmgren et al. 1998, Wilén & Balmér 1999, Martins et al.

2003). The effect is often quantified by measuring sludge volume index (SVI) which is the

volume (mL) that 1g of sludge occupies in suspension after settling by gravity for 30

minutes. Palmgren et al. (1998) showed that dissolved oxygen limitation (DO < 0.1 ppm)

leads to a decrease in cell surface hydrophobicity and the worsening of the solid-liquid

separation, in activated sludge systems. Sponza (2002) reported an increase in municipal

21

flocs’ hydrophobicity under oxygen limited conditions. Martins et al. (2003) attributed the

poor settleability (SVI > 250 mL/g) of sludge formed in low DO (< 1.1 mg/L) to the

overgrowth of filamentous bacteria. They showed that the effect was worsened at high

chemical oxygen demand loading rate, and suggested a DO level of 2.5 mg/L to maintain

good sludge settleability (SVI < 100 mg/L).

Bulking has been identified as a common problem in systems where the operation

shifts in time and/or space between aerobic, anoxic and anaerobic conditions (Eikelboom

1998, Yun et al. 2000, Martins et al. 2004, Vaiopoulou et al. 2007). Wilén & Balmér (1999)

showed that altering aerobic and anaerobic conditions (1-4 h) did not affect the settling

properties to a large extent. However, they reported a significant increase in turbidity and a

gradual increase in the fraction of smaller flocs (2-20 µm) during the anaerobic period.

Moon et al. (2004) used sequencing batch reactors to study the effect of anaerobic

conditions on floc size and concluded that anaerobic flocs are smaller than aerobic flocs.

Martins et al. (2004) concluded that the presence of microaerophilic zones in the anoxic

stages of BNR systems leads to worsening sludge settling characteristics. Tampus et al.

studied the effect of feeding pattern on sludge settleability in denitrifying systems under

anoxic conditions, and concluded that despite the diffused and irregularly shaped structure,

the sludge settled well (SVI < 60 ml/g) (Tampus et al. 2004). Wilén et al. showed that

activated sludge deflocculates under anaerobic conditions and that the deflocculated

fragments are mainly composed of bacteria and EPS (Wilén et al. 2000). They observed

reduced deflocculation when nitrate acted as an electron acceptor (anoxic conditions). It

has also been reported that the presence of PAOs causes the formation of more compact

22

structures and improved sludge settleability (Eikelboom et al. 1998, Crocetti et al. 2000,

Tampus et al. 2004).

Variations in the DO level during wastewater treatment could affect floc structure

by changing the oxidation state of the naturally occurring iron, which impacts floc

aggregation. In the aerated activated sludge systems, where higher levels of DO are

present, iron is mostly in the form of Fe3+ (Rockne 2007). However, in anaerobic and

anoxic environments iron commonly takes the form of Fe2+. Fe2+ is known to be a more

effective flocculent compared to Fe3+ due to its ability in forming stronger ionic bonds

under neutral pH (Oikonomidis et al. 2010). Oikonomidis et al. (2010) reported that Fe3+

forms flocs that are diffuse and irregular, whereas Fe2+ causes the formation of compact

and less filamentous flocs with well-defined boundaries. It has also been shown that

compared to Fe2+, Fe3+ has a high affinity for protein (Murthy et al. 2000), and during

anaerobic conditions as ferric iron converts to ferrous there is a loss in selective binding

between iron and protein, that causes a release of protein (Novak et al. 2003). This can

affect the composition of microbial flocs in aerobic and anaerobic treatments, and their

structural stability.

2.4.2. The Effect of Sludge Retention Time (SRT) on Sludge and Floc

Properties

SRT is the one of the most important operational parameters in biological

wastewater treatment. It is defined as the average time that sludge is retained in the

bioreactor. Liao et al. performed a comprehensive study on the effect of sludge retention

time on sludge flocs composition, morphology, and stability (Liao et al. 2002, 2006). They

23

simulated the wastewater biological treatment process, by operating fully aerated

sequencing batch reactors at 28 ºC with synthetic feed and at SRTs in a range of 4-20 days.

Their results showed that flocs formed at higher SRTs have a more spherical and compact

structure, were less hydrated, and showed higher chemical stability. Investigating the

effect of SRT on the microbial flocs inter-particle interactions, Liao et al. (2002) concluded

that ionic interactions and hydrogen bonds are the two dominant forces that maintain

stability on sludge flocs at lower SRTs, and physical enmeshment and van der Waals

and/or hydrophobic interactions are more important in controlling the stability of sludge

flocs at higher SRTs. A conceptual model was proposed for the floc structure based on

inter-particle interactions, in which flocs generated at higher SRTs form dense cores that

have a lower growth rate than that of the outer layer, due to the limitations of diffusion of

oxygen and nutrients from the bulk solution (Figure 2. 3).

Wang et al. also operated sequencing batch reactors on synthetic feed at 20 ºC and

at SRTs of 5, 10, and 20 days (Wang et al. 2013). They found that the loosely bound EPS

content decreased at higher SRTs. Similar observations were made by Xie and Yang, as

they examined the effect of SRT on flocculation and bacterial communities in wastewater

sludge. They concluded that the loosely bound EPS correlates positively with the sludge

volume index and effluent suspended solids (Xie & Yang 2009). Scott et al. operated

sequencing batch reactors at 27 ºC with synthetic feed at sludge retention times ranging

from 4-20 days to study the effect of SRT and feed source on the physicochemical

properties of sludge flocs (Scott et al. 2005). They found that the composition of EPS

(protein, carbohydrate, and DNA content) did not significantly change by varying SRT.

However, they reported that flocs at higher SRTs are more hydrophobic and a larger

24

fraction of E.coli adheres to them. Therefore, the number of free swimming E.coli

decreases in the final effluent by increasing SRT.

Liss et al. (2002) studied the effect of sludge retention time on microbial floc

structure using various microscopy techniques (optical microscopy, environmental

scanning electron microscopy, transmission electron microscopy). They reported that flocs

at lower SRT are more irregular in shape and have higher carbohydrate content.

Transmission electron microscopy revealed a denser EPS layer formed in flocs at higher

SRT, which protect the interior cells from disruption. In general, flocs generated at higher

SRTs are more stable, hydrophobic and less negatively charged. They suggested that the

hydrophobicity of sludge flocs is affected by the protein to carbohydrate ratio in particular,

and that it could also be related to the differences in the amounts of lipids in the slime and

capsular layers associated with EPS.

Loge et al. took mixed liquor and secondary effluents from eight wastewater

treatment plants, and used in-situ hybridization of a 16S rRNA oligonucleotide probe

specific to the family Enterobacteriaceae (Loge et al. 2002). They concluded that the

fraction of particle associated coliform decreases exponentially with increasing the mean

cell residence time. Emerick et al. (1999) similarly reported that the percentage of floc

association of coliform bacteria was 20, 12, and 4% for sludge retention times of 2.2, 11,

and 16 days respectively.

2.4.3. The Effect of Operational Temperature on Sludge and Floc Properties

Wilén et al. (2010) performed a comprehensive study on the settling and

flocculation of a full-scale activated sludge treatment plant in different seasons. They

25

found that the settling and compaction properties of sludge generally improved during the

summer months (variation of temperature from a minimum of 7 ˚C in the winter to a

maximum of about 20 ˚C in the summer). They also concluded that flocs formed during

the winter are more susceptible to breaking under shear stress, as the bacteria grow in small

colonies in chain-like structures. Tian et al. (1994) found that the hydrolysis rate of

influent particles decreases when the operating temperature decreased from 20 to 8 ºC, and

on average caused a 16% increase in volatile solid production in the colder (8 ºC) reactor.

They suggested that under cold weather the viable organisms will comprise a smaller

percentage of the total activated sludge.

Liao et al. (2011) studied the settling properties of sludge generated in sequencing

batch reactors running at different temperatures (35 ˚C and 55 ˚C) on synthetic feed. The

results showed that the settleability of thermophilic (55 ˚C) sludge was poorer (higher SVI)

than that of the mesophilic (35 ˚C) sludge, due to a higher level of loosely bound EPS in

the thermophilic sludge (32 vs. 17.7 mg/g MLSS for thermophilic vs. mesophilic sludge).

Morgan-Sagastume & Allen (2003) studied the effect of temperature shifts (from 35 ˚C to

45 ˚C) on aerobic sludge and found poor sludge settling at higher temperatures under

steady-state conditions. Krishna and Van Loosdrecht (1999) examined the effect of

operating temperature on sludge settleability using aerobic SBRs treating an acetate

medium. They found a continuous decrease in sludge settleability (SVI increase) when

increasing temperature (15 ˚C, 20 ˚C, 25 ˚C, 30 ˚C, 35 ˚C). They argued that at lower

temperatures of 15 and 20 ˚C larger flocs and many protozoa are found in sludge, whereas

at 30 and 35 ˚C, the flocs become smaller with more zoogloeal overgrowth which led to

higher SVI.

26

2.4.4. The Effect of Influent Phosphorus Content on Sludge and Floc

Properties

The effect of phosphorus limitation on floc structure has been studied by Liu and

Liss (2007). Larger, rounder and more compact flocs, with improved settling

characteristics were formed under the phosphorus limited conditions (100:5:0.3). The

darker regions seen in phase contrast images were attributed to the greater amount of EPS

in the sludge flocs. Dumitrache (2008) also studied the effect of phosphorus limitation on

floc structure. He found that flocs generated at phosphorus limited conditions (100:5:0.33)

were larger and observed darker areas with the optical microscope due to a more densely

packed biomass. Similar observations were made by Brei (2011).

Ericsson & Eriksson (1988) suggested that extreme pre-precipitation of phosphorus

prior to biological treatment, favors the growth of filamentous bacteria, and related that to

the depletion of easily assimilated phosphate. A decrease in the EPS protein content and

surface charge of activated sludge flocs under phosphorus deficiency has been reported by

Sponza (2002). Liu et al. (2006) showed a significant increase in the total polysaccharide

content of sludge and in the acid polysaccharide content of the EPS under the limited

phosphorus conditions.

Other factors in the biological wastewater treatment process, which change floc

structure, composition, and settling characteristics, include adding coagulants in the

primary treatment, varying carbon source, varying organic loading rate, etc. Gehr and

Nicell (1996) reported that replacing FeCl3 with alum significantly improved the efficiency

of UV disinfection and reduced the UV dose requirements. Xie & Yang 2009 showed that

by changing the carbon source from glucose to acetate, at SRTs of 5, 10 and 20 days, the

27

effluent suspended solids decreased and the sludge settled better. The effect of carbon

source on floc composition has been investigated by Scott et al. (2005). They showed that

by changing the main carbon source from glucose to equal portions of skim milk and

glucose, the protein content of the EPS increased.

2.5. Summary

This literature review has indicated that the presence of microbial flocs in final

effluents leads to poor UV disinfection performance, and that larger particles worsen this

effect. Some of the impacts of varying secondary biological treatment conditions on floc

morphological properties and chemical composition have been identified. For example low

oxygen concentrations lead to deflocculation, increase in turbidity, and the formation of

smaller flocs, or that more spherical and compact flocs are generated at higher SRT.

Nevertheless, the mechanism of UV resistance and the source of tailing are not well

understood. Moreover, the effect of biological treatment conditions on UV disinfection

performance and its relation to flocs structure and composition remain unclear. Therefore,

this research aims at identifying the source of UV resistance and relating the flocs

structural properties and chemical composition to UV disinfection kinetics.

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3. The Effect of Wastewater Floc Structure on UV

Disinfection Kinetics1

3.1. Abstract

Ultraviolet disinfection is a physical method of disinfecting secondary treated

wastewaters. Flocs formed during secondary treatment harbor and protect microbes from

exposure to ultraviolet (UV) light, and significantly decrease the efficiency of disinfection

at high UV doses causing the tailing phenomena. However, the exact mechanism of tailing

and the role of floc properties and treatment conditions are not widely understood. The

objective of this work is to test whether a possible mechanism for tailing is that sludge

flocs are composed of an easily disinfectable loose outer shell, and a physically stronger

compact core that is more resistant to UV. In this chapter, hydrodynamic shear stress was

applied to the flocs to peel off the looser outer shell to isolate the cores. Floc and core

samples were fractionated into narrow size fractions by sieving and their UV disinfection

kinetics were determined and compared. The results showed that for flocs, the tailing level

elevated as the floc size increased, showing greater resistance to disinfection. However, for

the cores larger than 45µm, it was found that the UV inactivation curves overlapped, and

exhibited nearly identical inactivation kinetics. When comparing flocs and cores of similar

size fraction, the results consistently showed that cores were harder to disinfect with UV

1 .This Chapter is based on the following article: Azimi, Y., Allen, D. G., Farnood, R. R. (2012). Kinetics of UV Inactivation of Wastewater Bioflocs. Water

Research, 46 (12), 3827-3836.

37

light, and showed a higher tailing level. This study suggests that physical structure of flocs

plays a significant role in the UV inactivation kinetics.

3.2. Nomenclature

N (CFU): Final concentration (i.e. after UV irradiation) of viable flocs

N0 (CFU): Initial concentration of viable flocs

D (mJ/cm²): UV dose delivered to the wastewater. The product of UV intensity (mW/cm²)

and time (s)

δs (µm): The thickness of the loose shell

Deff (mJ/cm²): The effective dose delivered to a certain point inside a floc, which is a

function of δ, A, and floc porosity

δ (µm): The distance of the point of interest in the floc from the flocs surface that varies

between 0 and R

A (1/cm): The UV absorption coefficient of the floc material

Ac (1/cm): The UV absorption coefficient of the core material

AS (1/cm): The UV Absorption coefficient of the loose shell material

R (µm): The spherical radius of a floc

k (cm²/mJ): The inactivation rate constant of embedded coliform bacteria

k1 (cm²/mJ): Inactivation rate constant for the UV-susceptible or type 1 flocs

k2 (cm²/mJ): Inactivation rate constant for the UV-resistant of type 2 flocs

keff (cm²/mJ): The effective inactivation rate constant

f(δ,R): The probability density function of δ

38

β: The fraction of UV resistant flocs, or the ratio of the number of viable UV-resistant flocs

to the total number of viable flocs

δ1 (µm): The distance from the surface of the floc in which the UV inactivation occurs

with a rate constant of k1

δc (µm): The thickness of the viable layer

κ: Ratio of the density of the core to that of the loose shell

DRC: Dose Response Curve

3.3. Introduction

With increased restrictions on the quality of the treated effluents, there has been a

growing interest in applying ultraviolet (UV) disinfection for disinfecting secondary

wastewater effluents. UV disinfection is a physical method, which eliminates much of the

hazards associated with chemical disinfectants (Matasci et al. 1999). UV light, specifically

around the wavelength of 254 nm, can penetrate through the cell wall and alter the

microorganisms DNA and halt their ability to reproduce (Piluso & Moffatt-Smith 2006,

Pfeifer et al. 2005).

Suspended solids in secondary treated wastewater are known to decrease the

efficiency of UV disinfection by absorbing or scattering UV light, and hence protecting the

embedded microorganisms from exposure to UV light (Jorand et al. 1998, Das 2001, Loge

et al. 1999). The suspended solids in the secondary treated effluents are mostly generated

during the biological treatment process, which results in the formation and growth of

microbial flocs. These flocs consist of microorganisms, extracellular polymeric substances

(EPS), organic and inorganic colloidal particles (Sanin & Vesilind 1994, Urbain et al. 1993)

39

all of which are held together and stabilized by various mechanisms, including, EPS

enmeshment, salt bridging, polymer bridging, and hydrophobic interactions (Forster 1985,

Pavoni et al. 1972, Liss et al. 1996). Activated sludge flocs consist of 90-98% water, and

on a dry basis they contain about 60-90% cellular organic material (Jenkins 1993). They

are irregular in shape, vary in size from several microns up to about 1000 µm (Yuan 2007),

and are highly porous (45-90%) (Li & Ganczaraczyk 1987).

Figure 3. 1 is the schematic of a typical UV dose-response curve (DRC) for a

treated effluent. This curve consists of an initial steep slope followed by a tailing region

where a decrease in the rate of inactivation is observed for the applied UV dose. The

tailing phenomenon increases the usage of energy to reach acceptable degrees of

disinfection.

More Easily Disinfected

Lo

g (

N/N

0)

0

Tailing

UV Dose (D)

More Easily Disinfected

Lo

g (

N/N

0)

0

Tailing

UV Dose (D)

Figure 3. 1. Illustration of a typical UV dose-response curve of wastewater

In 1977, Cerf suggested that tailing of UV DRC occurs because of the existence of

a mixed population with different degrees of resistance to UV inactivation (Cerf 1997).

Later, many researchers suggested that the floc-associated microbes were responsible for

tailing (Tan 2007, Loge et al. 1996, Qualls et al. 1985). Recently, Tan studied the effect of

floc size on the degree of disinfectability and on the tailing level of activated sludge flocs

40

ranging in size from 20-120 µm and reported that the required UV dose for disinfection of

wastewater flocs increased with the floc size (Tan 2007). Tan also found that similar to the

UV DRC of whole effluents, the DRC of a suspension of flocs with a narrow size

distribution exhibited two regions, namely, an initial steep slope followed by a tailing

region, and suggested that these two regions represented the inactivation of UV-susceptible

flocs and UV-resistant flocs, respectively.

The kinetics of UV disinfection of wastewater flocs is expected to be a function of

the floc structure. It has been proposed that wastewater flocs have a two-layer structure

consisting of a compact core surrounded by a loose shell (Eriksson et al. 1992, Liao et al.

2002, Yuan 2007, Sheng et al. 2006). Liao et al. studied the stability of flocs generated at

various sludge retention times (SRT), and suggested that at higher SRTs the floc core is

more compact and more hydrophobic (Liao et al. 2002). Recently, by exposing wastewater

flocs to hydrodynamic shear in a Couette flow cell, Yuan and Farnood reported that the

shear strength of cores could be several times greater than that of loose shell (Yuan and

Farnood 2010). They suggested that the higher mechanical strength of cores may be due to

the densely packed structure of the core where microorganisms are held together by EPS

enmeshment and other intra-floc forces. Sheng et al. also studied the stability of flocs

under shear and proposed a multilayer structure for wastewater flocs with two distinct

regions. The outer part was suggested to be a loosely entangled readily-extractable EPS

fraction (Sheng et al. 2006) that is completely dispersed under shearing, while the inner

part is more stable and could not be dispersed except under extreme conditions.

Despite recent advances in the overall understanding of the floc structure, the

relationship between structure and UV disinfection kinetics of wastewater flocs is not well

41

understood. The objective of this study is to investigate the causes of the tailing

phenomenon and its relationship to the structure of wastewater flocs. In particular, it is

hypothesized that the compact cores are more UV-resistant and hence cause the tailing of

UV DRC of wastewater flocs.

3.4. A Structural Model for Floc UV Disinfection

A simplified model is presented here to justify the observed kinetics of UV

disinfection of flocs. In this model, a fraction of the flocs contain a compact ‘core’

surrounded by a loose outer shell; while others have a loose structure and do not contain

any core. Moreover, it is assumed that:

• The UV disinfection of a floc is governed by the inactivation of the most

shielded microorganism embedded within that floc

• The loose and porous outer shell is represented by a continuum spherical shell

layer with an average density and an average UV absorbance

• The compact core can be represented by a continuum spherical material with a

higher density and UV absorbance compared to those of the loose shell

• The probability of finding a microorganism in a given volume is proportional to

the mass of material within that volume

• The UV light intensity in various points of the core and shell may be expressed

by exponential decay (i.e. Beer-Lambert’s law)

• The floc surface is uniformly exposed to UV light

The above model is consistent with the floc structure proposed in the literature

based on floc stability studies; i.e. a double layer structure where a more stable inner core is

42

surrounded by a looser, more easily dispersible outer layer (Yuan et al. 2010, Liao et al.

2002, Sheng et al. 2006).

In the most general case, flocs are represented by a spherical particle of radius R

consisting of a compact core with the radius Rc and a loose shell with thickness δS (Figure

3. 2). Given that the UV disinfection of these model flocs is governed by the location of

the most shielded indicator organism; i.e. an embedded indicator organism that receives the

least amount of UV dose, model flocs can be classified into two types. In Type 1 (or UV-

susceptible) flocs, the most shielded indicator organism is located within the loose material

and near the floc surface. Hence, a Type 1 floc either has or does not have core (Figure 3.

2-a), and the indicator organism is embedded within its loose shell (Figure 3. 2-b). In

contrast, Type 2 (or UV-resistant) flocs are those in which the most shielded indicator

organism is embedded within the compact core and is more protected from UV light.

Compared to Type 1 flocs, Type 2 flocs are harder to disinfect and as will be seen later, are

responsible for the tailing of UV DRC.

Due to the UV absorption by the floc material, the UV dose received by embedded

organisms is less than the estimated average dose delivered to a free floating microbe

within the sample. Assuming that the most shielded indicator organism i.e. fecal coliform

in this study, is located at a depth δ from the floc surface, the effective dose, Deff (δ),

received by this organism can be estimated from the Beer-Lambert law:

Deff (δ) = D e –A(δ) × δ [3-1]

where D is the average UV dose delivered to the sample in mJ/cm2, and A(δ) is the UV

absorption coefficient of the floc material.

43

Type 1 Type 2

Loose shell Compact core

(a) (b) (c) (d)

Core

δSR R

Figure 3. 2. Schematic diagrams showing model spherical flocs: (a, b) Type 1 or UV-susceptible floc, (c)

Type 2 or UV-resistant floc, and (d) isolated compact core. The small white circle indicates the location

of the most shielded viable indicator organism.

It is worth noting that in the standard membrane filtration test, used to assess

wastewater UV disinfection performance (APHA 2001, # 9222 A), every floc that

“survives” UV irradiation leads to a single CFU count, regardless of the number of

surviving coliform bacteria embedded within that floc. This is an inherent limitation of the

membrane filtration technique. However, this technique is currently the industry/regulatory

standard for design and evaluation of disinfection technologies. If the inactivation rate

constant of embedded coliform bacteria, k, is equal to that of free-floating ones, the

inactivation kinetics of a spherical floc with radius R can be expressed by:

N(R, δ) = No(R, δ) e – k Deff

(δ) [3-2]

here, No(R, δ) and N(R, δ) are the initial concentration and the final concentration (i.e. after

UV irradiation) of viable flocs, a ‘viable floc’ being a floc with at least one culturable

embedded coliform bacteria, expressed as the number of colony forming units (CFU) per

44

100 ml. Since in practice the average UV dose is of significance, Equations [3-1] and [3-2]

can be rearranged to obtain:

N(δ, R) = No(δ, R) e – keff

(δ) D [3-3]

where keff (δ) is the effective inactivation rate constant defined by:

keff (δ) = k e −A(δ) × δ [3-4]

Given that δ varies from 0 to R; the UV DRC of ideal spherical flocs with radius R

would be:

N(R) /No(R) = ∫oδ

f (R, δ ) e – keff

(δ) D dδ [3-5]

where, N(R) and No(R) are the initial and final counts of viable flocs with radius R (in

CFU/100 ml), and f (δ,R) is the probability density function of finding a culturable

microorganism at the depth of δ.

Considering the structure of model flocs in Figure 3. 2, and replacing the

inactivation rate constants of Type 1 and Type 2 flocs with their respective average values,

i.e. k1 and k2, Equation [3-5] can be expanded as:

N(R) /No (R) = (∫oδS f (R, δ) dδ) × e – k

1 D + (∫δS

δ

f (R, δ) dδ) × e – k2

D [3-6]

Or:

N(R) /No (R) = (1 – β) e – k1 D + β e – k

2 D [3-7]

where β characterizes the fraction of UV-resistant flocs defined as ratio of the number of

viable UV-resistant flocs to the total number of viable flocs of radius R:

β = ∫δsδ

f (R, δ) dδ [3-8]

45

The first term on the right-hand-side of Equation [3-7] reflects the inactivation of

UV-susceptible or Type 1 flocs while the second term represents UV-resistant or Type 2

flocs.

Equation [3-7], known as the double exponential model, is used in this study to

analyze the inactivation kinetics of wastewater flocs. It is worthwhile to note that the

double exponential model has been used previously to study the UV disinfection of whole

effluents (Farnood 2005). However, in that case k1 is associated with the inactivation of

free-floating organisms and β is the fraction of particle-associated CFU.

3.5. Materials and Methods

3.5.1. Wastewater Samples

Mixed liquor samples were collected from Ashbridges Bay municipal wastewater

treatment facility, located at the eastern end of the city of Toronto. The plant has a capacity

of 818,000 m3/day, and has a suspended solid activated sludge biological system for

secondary treatment. The samples were taken on five different days during a period of

three months (July-October 2009). All experiments were conducted within 48 hours of

sample collection.

To test the effect of floc size on UV disinfection, the mixed liquor sample was

fractionated into narrow size fractions using appropriate sieve trays (U.S.A. Standard

Testing Sieve). The nominal sieve opening sizes were 32, 45, 53, 63, 75, and 90 µm;

therefore, the size fractions obtained were 32-45, 45-63, 53-63, 63-75, and 75-90 µm.

Collected floc fractions were gently washed with milliQ water for 15 minutes and

46

backwashed into a flask. The fractionated samples prepared in this way are referred to as

the “un-sheared floc” or just “floc” samples.

3.5.2. Shearing

In order to isolate the compact cores of the flocs, mixed liquor samples were subject

to hydrodynamic shear in a custom-made Couette flow cell developed by Yuan (Yuan &

Farnood 2010). During this process, the weak loose outer layer of the floc is removed,

leaving behind a mechanically strong compact core that is resistant to breaking under the

applied shear stress.

A schematic drawing of the Couette flow cell device is shown in Figure 3. 3. Ten

milliliters of wastewater floc sample was placed in the gap between the spindle and the cup

and speed of the spindle was set at 3500 rpm for 15 minutes. These shearing conditions

had no significant effect on the culturability of fecal coliforms (data not shown here). The

sheared samples were sieved to the following size fractions: 32-45, 45-63, 53-63, 63-75,

and 75-90 µm. The sieved fractions were gently washed using milliQ water to remove

small fragments and to obtain a narrow size distribution. Collected fractions of compact

cores were re-suspended in milliQ water and stored in separate flasks for further testing. In

this manuscript, these samples are referred to as “sheared” or “core” samples.

To determine the volume fraction and number concentration of compact cores, and

their relation to the floc size, activated sludge samples were fractionated according to the

method described earlier. The collected flocs were analyzed using a Multisizer 3 particle

size analyzer (Beckman Coulter Canada, Mississauga, Ontario, Canada), and diluted using

deionized water to obtain 0.06% suspensions (volume basis). Thirty milliliters of each

suspension was subsequently sheared using the Couette flow device at 3500 rpm for 15

47

minutes. The sheared suspension was filtered through a 45µm mesh sieve to ensure that

small fragments of loose shell (generated during the shearing process) are adequately

removed from the isolated core samples. The collected cores were subsequently analyzed

using a Multisizer 3 particle size analyzer to obtain the number and volume of cores for

each floc size fraction.

Cup

Spindle

Motor

4.4 cm

4 c

m0.2 cm

Cup

Spindle

Motor

4.4 cm

4 c

m0.2 cm

Cup

Spindle

Motor

4.4 cm

4 c

m0.2 cm

Figure 3. 3. Schematic of the Couette flow apparatus

3.5.3. UV Disinfection Assay

A collimated beam apparatus equipped with a low pressure mercury UV lamp was

used to irradiate the samples (VQ 650143 Trojan Technologies, London, Ontario, Canada).

The UV transmittance of the samples was measured at 254 nm using a Lambda 35 UV-Vis

Spectrometer (Perkin-Elmer, Woodbridge, Ontario, Canada). This spectrometer features a

double-beam all-reflecting system, installed with an integrating sphere device: Labsphere

RSA-PE-20 (Perkin-Elmer, Woodbridge, Ontario, Canada). Light intensity was measured

using an IL-1700 radiometer (International Light, Peabody, MA, USA). UV transmittance

and the measured UV intensity were used to estimate the irradiation time required to

48

achieve the desired UV dose according to method described by Bolton and Linden (Bolton

& Linden 2003).

To obtain the UV DRC, 20 mL of resuspended samples was placed in a glass Petri

dish and subjected to a range of UV doses between 10 and 60 mJ/cm2. The irradiated

samples were cultured using the membrane filtration method following the Standard

Methods for the Examination of Water and Wastewater (APHA 2001, # 9222 A). m-FC

agar (VWR, Mississauga, Ontario) was used as the culturing media and the buffer solution

used for rinsing the flocs on the filter (Fisher brand Water Testing membrane Filters, 47mm,

0.45µm, Mississauga, Ontario) was 13.6 g/L KH2PO4 at pH 7.2. The samples were then

incubated at a temperature of 44 ºC for 24 ± 2 hours, and then the colony forming units

were enumerated. The DRC data was analyzed using MathematicaTM v.7 (Wolfram

Research, Champaign, USA) to determine parameters of the double exponential model. UV

bioassay was performed on both “core” and “floc” samples of various size fractions.

Since floc and core samples were prepared by sieving and resuspension of mixed

liquor flocs in milliQ water, it is more meaningful to report the UV DRC data for these

samples as the normalized survival ratio; N/N0. The value of N0 in these samples was

typically between 104-105 CFU/ 100 mL.

3.5.4. Optical Microscopy and Size Distribution

Optical microscopy images of core and floc samples were obtained using a Leica

Optical DM LA microscope (Leica Microsystems, Bannockburn, IL, US) equipped with a

10X objective lens. The field of view at this magnification was 0.870mm × 0.690mm, with

an image resolution of 0.67 µm/pixel.

49

Particle size analyses of core and floc samples were carried out using the Multisizer

3.0 particle analyzer. This device also allows for the measurement of particle concentration,

i.e. # of particles per unit volume of the sample. It should be noted that the Multisizer

operates based on Coulter principal, and hence underestimates the physical size of porous

particles, such as flocs (Kachel 1986).

3.5.5. Statistical Analysis

The statistical significance of the results was tested by performing a student t-test

on the double exponential parameters extracted from each data set. Since each UV DRC

was repeated in triplicates, three values were obtained for k1, k2 and β for each sample.

The student t-test was done in pairs for all size fractions for k1, k2, and β with α = 0.05.

Therefore, if the calculated p-value was smaller than 0.05 the difference between the UV

DRC for any two samples being compared was statistically significant. On the other hand,

if the p-value was larger than 0.05 the two samples were not statistically significantly

different in terms of UV inactivation kinetics.

3.6. Results and Discussion

UV DRC experiments conducted in this study were divided in two sets: 1- UV DRC

of un-sheared flocs; and 2- UV DRC of compact cores. In the first set of experiments,

mechanical sieving was used to isolate flocs of various size fractions. Subsequently, the

UV disinfection assay was performed to obtain the DRC for each floc size fraction. In the

second set of experiments, mixed liquor samples were first sheared and subsequently

sieved to isolate compact cores of various size fractions. The DRC’s of the isolated core

50

samples were then determined using UV disinfection assay. Results of these experiments

are provided and discussed in the following sections.

3.6.1. Flocs UV DRC

In this section, effect of floc size on the kinetics of UV disinfection is examined and

experimental results are compared to those obtained using the double exponential model.

The normalized UV DRC of the un-sheared floc samples are given in Figure 3. 4.

The DRC of each of these samples has a steep initial slope followed by near-plateau tailing

region. As discussed earlier, the initial slope of the UV DRC of flocs is due to the presence

of UV-susceptible (Type 1) flocs in the sample while the tailing region represents the

disinfection of the UV-resistant (Type 2) flocs. It should be pointed out that in this study,

since the floc samples were washed with MilliQ water, the number of free organisms in the

samples was negligible.

0.0001

0.001

0.01

0.1

1

0 20 40 60 80

UV Dose (mJ/cm²)

N /

N 0

75-90µm

53-63µm

32-45µm

Figure 3. 4. UV dose-response curves (DRCs) for floc samples of different size fractions. Solid curves in

figure represent the fitted double-exponential model. The data presented includes replicates of the

same experiment over three months with at least four samplings.

51

Table 3. 1. Inactivation rate constants (k1 and k2 ) and UV-resistant fraction (β) of un-sheared floc

samples. The ± values represent 95% confidence intervals.

Size k1 (cm2/mJ) k2 (cm

2/mJ) β

32-45 µm 0.23 ± 0.01 0.011 ± 0.004 0.04 ± 0.004

45-63 µm 0.22 ± 0.02 0.010 ± 0.006 0.07 ± 0.01

53-63 µm 0.19 ± 0.01 0.014 ± 0.006 0.09 ± 0.01

63-75 µm 0.18 ± 0.02 0.013 ± 0.004 0.12 ± 0.01

75-90 µm 0.17 ± 0.02 0.012 ± 0.003 0.14 ± 0.01

The double exponential model parameters for un-sheared floc samples; i.e. k1, k2,

and β, are given in Table 3. 1. This table shows that the initial inactivation rate constant for

un-sheared flocs, k1, ranged from 0.17 to 0.23 cm2/mJ that was significantly lower than the

inactivation rate constant of free-floating coliform bacteria (~0.67 cm2/mJ) (Wright &

Cairns 1999). This finding supports the assertion that the steep initial slope of the UV

DRC of un-sheared floc samples is not caused by the free-floating organisms (as they were

eliminated in the sieving and washing procedure), but instead is due to the presence of UV-

susceptible flocs. Moreover, k1 is negatively correlated to β, i.e. an increase in the fraction

of UV resistant flocs and hence the tailing level of UV DRC resulted in a lower initial

inactivation rate constant (Figure 3. 5).

y = -1.48x + 0.38

R2 = 0.94

0

0.04

0.08

0.12

0.16

0.16 0.2 0.24

β

k1

Figure 3. 5. Correlation between the values of β and k1 for flocs of various sizes

52

Table 3. 1 also shows that k1 decreased significantly with increasing floc size (p =

0.006). This may be due to an increase in the percentage of flocs that contain compact

cores (see Figure 3. 4), hence, causing an increase in the tailing level of UV DRC with the

floc size. In other words, larger flocs have a larger total volume of cores, hence there is a

higher chance to find an indicator organism embedded in the core, and these flocs would be

harder to disinfect. To illustrate this point, calculated UV DRC’s based on the double

exponential model for various values of β are plotted in Figure 3. 6. Although the initial

inactivation rate constant was kept constant at 0.23 cm2/mJ, the magnitude of the initial

slope of UV DRC decreased with increasing β. The estimated inactivation rate constants

based on the initial slope of the DRC were 0.20, 0.18 and 0.16 cm2/mJ for β = 0.04, 0.09,

and 0.14, respectively.

Contrary to k1, the inactivation rate constant for the tailing region of UV DRC (i.e.

k2) of un-sheared floc samples did not show any statistically significant variation with the

floc size (p= 0.5). This result suggests that the UV inactivation rate constant in the tailing

region, or k2, was independent of the floc size.

Table 3. 1 also shows that larger flocs had higher β values, and hence had more

particles likely to be resistant to UV disinfection (the value of β for 75-90µm fraction was

more than 3 times greater than that of 32-45µm fraction). To better understand this

behavior, the volume percent and the number percent of compact cores extracted from two

floc samples of different size fractions are given in Table 3. 2. These results show that the

ratio of core volume to the total floc volume increased with increasing floc size. Hence,

assuming that viable indicator organisms are randomly distributed within the floc volume,

53

the likelihood of finding an indicator organism embedded within the floc core; hence the

fraction of UV resistance flocs, is expected to increase with the floc size.

0.01

0.1

1

0 20 40 60 80

N/N

o

UV Dose, mJ/cm2

0.04

0.09

β = 0.14

Figure 3. 6. The calculated UV DRC using double exponential model for various values of β. The

inactivation rate constants k1 and k2 are 0.23 and 0.012 cm2/mJ, respectively.

Table 3. 2. Volume and number percent of cores extracted from various floc size fractions (volume

%=total core volume/ total volume of un-sheared flocs, and number %= total number of cores / total

number of un-sheared flocs.

Floc Size Fraction Core volume % Core number %

53-63 µm 38 ± 9 35 ±13

75-106 µm 67 ± 10 91 ± 7

Our results suggest that larger flocs on average not only contained larger cores, but

also a larger number of them contain cores. Figure 3. 7 shows the size distribution of cores

extracted from two different floc samples. Based on this Figure, the average core size

increased from about 27µm for the 53-63µm floc sample to 37µm for the 75-106µm floc

sample.

54

Nu

mb

er

of

Part

icle

s

Particle Diameter (µm)

53-63 µm Bioflocs

Cores > 45µm from 53-63µm Bioflocs

75-106 µm Bioflocs

Cores > 45µm from 75-106µm BioflocsN

um

ber

of

Part

icle

s

Particle Diameter (µm)

53-63 µm Bioflocs


75-106 µm Bioflocs


Figure 3. 7. Particle size distribution of flocs (53-63µm and 75-106µm) and the compact cores larger

than 45µm extracted from each size fraction

3.6.2. UV DRC of Compact Cores

In this section, results of UV disinfection of compact cores with various size

fractions are presented and a structural model is proposed to describe the observed UV

dose response curve of compact cores.

Figure 3. 8 illustrates the morphological differences between typical flocs and

compact cores. This Figure shows that cores were more optically opaque and hence were

likely denser than the original flocs. Cores also appeared to have a more regular

morphology and were more ellipsoidal in shape.

55

(a)

(b) (c)

Figure 3. 8. White light optical microscopy images of (a) single core and floc with similar size, (b) un-

sheared floc sample, and (c) a core sample. Note differences in the morphology and light transmittance

(density) between cores and flocs.

The dose-response curves for core samples are presented in Figure 3. 9. For

comparison, the fitted double exponential models are also plotted. Similar to the un-sheared

floc samples, the UV DRC of cores exhibited an initial steep slope followed by a tailing

region. This behavior suggests the presence of UV-resistant and UV-susceptible cores in

the sample. Hence, Equation [3-7] that was developed for flocs, can be also applied to the

core samples.

According to Figure 3. 9, the kinetics of UV disinfection of core samples larger

than 45 µm were almost identical while the normalized UV DRC for the 32-45µm cores

placed lower than those of larger core size fractions. It should be noted here that the UV

56

DRC data was collected at least five times over the period of June 2009 to September 2009,

and the variations among the DRC of the replicates between different days were within a

third of a log (i.e. well within the experimental error).

The inactivation rate constants of core samples (k1 and k2) and fraction of UV-

resistant cores (β) are given in Table 3. 3. This table shows that k1 did not vary significantly

(p= 0.2) by changing the cores size above 45 µm, while inactivation rate constant for UV-

resistant cores, k2, remained approximately the same for all core samples at about 0.013

cm2/mJ.

0.0001

0.001

0.01

0.1

1

0 20 40 60 80UV Dose (mJ/cm²)

N /

N 0

32 - 53 µm

75 - 90 µm

45 - 63 µm53 - 63 µm

Figure 3. 9. UV dose-response curves for core samples of different size fractions. Solid curves

represent the fitted double exponential model.

57

Table 3. 3. Inactivation rate constants (k1 and k2 ) and UV-resistant fraction (β) for core samples. The

± values represent 95% confidence intervals.

Size k1 (cm2/mJ) k2 (cm

2/mJ) β

32-45 µm 0.21 ± 0.02 0.011 ± 0.003 0.10 ± 0.01

45-63 µm 0.16 ± 0.02 0.014 ± 0.003 0.18 ± 0.02

53-63 µm 0.17 ± 0.03 0.012 ± 0.005 0.17 ± 0.03

63-75 µm 0.16 ± 0.01 0.012 ± 0.003 0.17 ± 0.02

75-90 µm 0.15 ± 0.01 0.012 ± 0.002 0.17 ± 0.02

The observed UV DRC of core samples can be explained based on a non-uniform

spatial distribution of indicator organisms (i.e. fecal coliform bacteria) in the core. If

coliform bacteria were distributed uniformly within the core volume, some cores would

contain coliform bacteria near or at their centres; hence, unlike Figure 3. 9, the tailing level

of UV DRC should have progressively increased with the core size. The results therefore

indicated that viable fecal coliform bacteria were concentrated within a ‘viable layer’ near

the surface of the core. Furthermore, the thickness of the viable layer, denoted by δc, is

independent of the core size such that the effect of core size on the UV inactivation kinetics

is reduced for larger cores. Using Equation [3-4] for k2 = 0.012 and k = 0.67 cm2/mJ, and

assuming the UV absorbance of core material to be equal to that of densely packed cells or

approximately 1000 cm−1 (based on (Jagger 1967) that ~90% of UV light absorbed in 10

µm of cells), δc can be estimated as 18µm. In other words, based on our model, the tailing

region of UV DRC of compact cores was governed by coliform bacteria located within a

layer of roughly 18µm thick. This may contribute to the observation that the 32-45 µm

cores show a lower tailing level than the UV DRCs of larger cores. Another reason may be

the entrapment of the loose outer shell in the sieve with the smallest opening (32 µm), and

58

hence easing the disinfection process and yielding a lower tailing level for the mentioned

size fraction.

A further application of our model can help illustrate the validity of the above

explanation that the thickness of the viable layer of the core is independent of core size. If

we assume that the probability of finding coliform bacteria within a spherical shell of a

core between δ and δ+dδ is proportional to the mass of solids in that shell, then, assuming a

uniform core density, fc(δ) can be expressed by:

fc (δ) = 3 (Rc – δ)2 / (Rc3 – δc

3) 0 < δ < δc [3-9]

= 0 δc < δ < Rc

where Rc is the radius of core, and δ=0 being the surface of the core.

Using the above relationship and assuming a constant UV absorbance of 1000 cm−1,

Equation [3-5] was numerically integrated for various values of core diameter. The results

(Figure 3. 10-a) confirm that for core sizes larger than 60 µm, the UV DRCs would be

almost identical. In fact, according to Equation [3-9], by increasing the core size, f (δ)

approaches unity, and hence the estimated UV DRC will become insensitive to variations

in the core size, Rc. This is not the case for the smallest core size (i.e. 40µm), where δc was

nearly equal to the core radius; hence these small cores resulted in a larger k1 and a lower

tailing level. For comparison, the calculated UV DRCs of spherical cores with a uniform

distribution of coliform bacteria throughout the core; i.e. without a viable layer, is given in

Figure 3. 10-b. As expected, there is a significant difference between the UV DRC of larger

cores sizes in this case.

59

0.01

0.1

1

0 10 20 30 40 50 60 70

N /

No

UV Dose (mJ/cm2)

60μm70μm

80μm

40 μm

(a)

0.01

0.1

1

0 10 20 30 40 50 60 70

N /

No

UV Dose (mJ/cm2)

80 μm

40 μm

60 μm

70 μm

(b)

Figure 3. 10. The estimated UV DRC of the spherical cores with diameters of 40, 60, 70, and 80µm for:

a) cores with a viable layer of 20 µm, and b) cores without viable layer. UV absorbance is 1000 cm−1

.

Figure 3. 10-a and 9.b also emphasize the effect of UV absorbance of the core

material, Ac, on the tailing of UV DRC. In addition, it was found that the predicted UV

60

DRCs of ideal spherical cores were sensitive to variations in δc (results not shown here).

Nevertheless, it is worthwhile to note that the values of k1 and k2, as estimated from Figure

3. 10 were in the order of 10−1 and 10−2, respectively, which are comparable to the values

listed in Table 3. 3.

Comparing the UV dose-response curves for the cores and un-sheared flocs of the

same size fraction, it can be concluded that cores generally had a higher tailing level (i.e. a

larger β value). This observation confirms that compact cores were more UV-resistant than

the flocs of the same size. Moreover, the value of k2 for cores was similar to that of flocs.

In other words, the kinetics of UV disinfection of un-sheared flocs and compact cores in

the tailing region were essentially the same. This finding supports the idea that UV-

resistant flocs likely contained UV-resistant compact cores, and that the presence of

compact cores in suspended flocs was ultimately responsible for the tailing phenomenon of

UV DRC of wastewater.

3.6.3. The Role of Compact Cores in UV Disinfection Kinetics of Flocs

In this section, the structural model presented earlier in this paper (see Section 3.4)

is used to predict the effect of core size on the UV inactivation of flocs.

The UV DRCs of model spherical flocs containing spherical compact cores were

determined using Equation [3-5]. Consider a floc of radius, RF, containing a single

spherical core. The thickness of loose shell and that of viable layer for this floc are δS and

δc , respectively. Assuming that the ratio of density of the core to that of the loose shell is κ,

the probability of finding coliform bacteria in a unit volume of core is assumed to be κ-

times larger than that of the loose shell. Hence, f F(δ) , is given by:

61

f F(δ) = 3 (RF – δ)2 / M 0 < δ < δS [3-10]

= 3 [(RF – δS)2 + κ (RF − δ)2]/ M δS < δ < δS+ δc

= 0 δS+ δc < δ < R

where

M = [RF3 – (RF – δS)3 ] + κ [(RF – δS)3 − (RF – δS– δc)

3 ]. [3-11]

Assuming that core density is twice that of the loose shell, and that the UV

absorbance of cores and loose shells are proportional to their densities, then for Ac = 1000

cm-1 the UV absorbance of loose shell will be AS= 500 cm-1. Using Equations [3-5] and [3-

10] together, the UV DRC’s of model spherical flocs with a diameter of 60 µm were

estimated and plotted in Figure 3. 11. According to this figure, in the absence of any

compact core (identified as ‘No core’ in Figure 3. 11), at high UV doses (> 20 mJ/cm2)

although the inactivation rate slowed down, the kind of tailing phenomenon that was

experimentally observed earlier (see Figure 3. 4) was not evident. With increasing core

diameter, or equivalently core volume, the tailing level increased and the initial slope of the

UV DRC of flocs became less steep. For comparison, the slope of the tailing region of UV

DRC for the flocs containing 40 vol% cores was about 0.027 cm2/mJ which is comparable

to the calculated value for the slope of tailing region of 60µm cores (0.02 cm2/mJ) seen in

Figure 3. 10-a. It is also important to notice the experimental values of 0.01-0.014

(cm2/mJ) found for the slope of tailing region of flocs (Table 3. 1) and isolated compact

cores (Table 3. 3) are of the same order of magnitude of the calculated values. These results

suggest that the loose shell-compact core model of floc structure presented in here

successfully captures qualitative and quantitative features of the UV dose response curve of

flocs. Furthermore, the similarities between the inactivation rate constants of compact

62

cores and loose shells support that the tailing of UV DRC of flocs is controlled by the

presence of compact cores.

0.001

0.01

0.1

1

0 10 20 30 40 50 60 70

N /

No

UV Dose (mJ/cm2)

10% (28 μm)

40% (44 μm)

70% (53 μm)

No core

Figure 3. 11. UV DRC spherical flocs for various 0%, 10%, 40%, and 70% core volume fractions. Core

diameter is given in parenthesis. Floc diameter is 60µm, and the UV absorbance of floc and core are

assumed to be 500 cm-1

and 1000cm-1

, respectively.

3.7. Conclusions

The conclusions of this study are:

• Floc structure has a significant effect on disinfectability with UV light. The

presence of compact cores in flocs reduces the UV penetration and hence

increases their resistance to UV disinfection.

• Cores have a higher tailing level compared to flocs of similar size.

• Larger flocs contain a larger number and a larger volume of cores. Therefore,

they are more likely to be resistant to UV disinfection and hence have a higher

UV DRC tailing level.

63

• In the tailing region, the inactivation rate constant of un-sheared flocs and

compact cores were essentially the same. This finding supports the idea that

compact cores, are responsible for the tailing phenomenon of UV DRC of

wastewater flocs.

• For the cores larger than 45 µm, UV inactivation kinetics are not affected by the

core size. In contrast for flocs, an increase in β was observed with increase in

floc size. This phenomenon was explained by the suggested model of having a

viable layer of a certain thickness for cores of all sizes.

3.8. References

APHA (2001), Standard methods for the examination of water and wastewater. 22nd Ed. Washington DC, USA Bolton, J. R., Linden, K. G. (2003). Standardization of methods for fluence (UV Dose) determination in bench-scale UV experiments. Journal of Environmental Engineering, 129 (3), pp. 209 – 215 Cerf, O. (1977). Tailing of survival curves of bacterial spores. Journal of Applied Bacteriology, 42(1), 1-19. Das, T. K. (2001). Ultraviolet disinfection application to a wastewater treatment plant. Clean Products and

Processes, 3(2, pp. 69-80), August. Eriksson, L., Steen, I., & Tendaj, M. (1992). Evaluation of sludge properties at an activated sludge plant. Water Science and Technology, 25(6), 251-265. Farnood, R., Flocs and ultraviolet disinfection. Chapter 18 in Flocculation in Natural and Engineering

Systems. I.G. Droppo, G.G. Leppard, S.N. Liss, and T.G. Milligan, Eds., CRC Press, Boca Raton, FL, 2005, pp. 385-395. Forster, C. F. (1985). Factors involved in the settlement of activated sludge-II; the binding of polyvalent metals. Water Research, 19(10), 1265-1271. Jagger, J., (1967) Introduction to Research in Ultraviolet Photobiology. Prentice-Hall, Inc. Englewood Cliffs, NJ Jenkins, D. (1993). In Richard M. G., Daigger G. T. (Eds.), Manual on the causes and control of activated

sludge bulking and foaming (2nd Ed.). Boca Raton: Lewis. Jorand, F., Boué-Bigne, F., Block, J. C., & Urbain, V. (1998). Hydrophobic/hydrophilic properties of activated sludge exopolymeric substances. Water Science and Technology, 37(4-5), 307-315. Kachel, V. (1986). Investigations into Coulter Sizing of Biological Particles; Theoretical Background and Instrumental Improvements. Particle Characterization, 3(2), 45-55.

64

Liao, B. Q., Allen, D. G., Leppard, G. G., Droppo, I. G., & Liss, S. N. (2002). Interparticle interactions affecting the stability of sludge flocs. Journal of Colloid and Interface Science, 249(2), 372-380. Li, D., & Ganczarczyk, J. J. (1987). Stroboscopic determination of settling velocity, size and porosity of activated sludge flocs. Water Research, 21(3), 257-262. Liss, S. N., Droppo, I. G., Flannigan, D. T., & Leppard, G. G. (1996). Floc architecture in wastewater and natural riverine systems. Environmental Science and Technology, 30(2), 680-686. Loge, F. J., Emerick, R. W., Heath, M., Jacangelo, J., Tchobanoglous, G., & Darby, J. L. (1996). Ultraviolet disinfection of secondary wastewater effluents: Prediction of performance and design. Water Environment

Research, 68(5), 900-916. Loge, F. J., Emerick, R. W., Thompson, D. E., Nelson, D. C., & Darby, J. L. (1999). Factors influencing ultraviolet disinfection performance part I: Light penetration to wastewater particles. Water Environment

Research, 71(3), 377-381 Matasci, R. Weston, P. Lau, J. Cruver, S. Marek, D. Tomowich (1999), Wastewater Technology Fact Sheet: Ultraviolet Disinfection, United States Environmental Protection Agency, Office of Water, Washington, DC, EPA 832-F-99-064, R. Pavoni JL, Tenney MW, & Echelbreger JR WF. (1972). Bacterial exocellular polymers and biological flocculation. Journal of the Water Pollution Control Federation, 44(3 pt 1), 414-431 Pfeifer, G. P., You, Y., & Besaratinia, A. (2005). Mutations induced by ultraviolet light. Mutation Research -

Fundamental and Molecular Mechanisms of Mutagenesis, 571(1-2 SPEC. ISS.), 19-31. Piluso, L. G., & Moffatt-Smith, C. (2006). Disinfection using ultraviolet radiation as an antimicrobial agent: A review and synthesis of mechanisms and concerns. PDA Journal of Pharmaceutical Science and

Technology, 60(1), 1-16. Qualls, R. G., Ossoff, S. F., & Chang, J. C. H. (1985). Factors controlling sensitivity in ultraviolet disinfection of secondary effluents. Journal of the Water Pollution Control Federation, 57(10), 1006-1011. Sanin, F. D., & Vesilind, P. A. (1994). Effect of centrifugation on the removal of extracellular polymers and physical properties of activated sludge. Water Science and Technology, 30(8 pt 8), 117-127. Sheng, G., Yu, H., Li, X. (2006). Stability of sludge flocs under shear conditions: Roles of extracellular polymeric substances (EPS). Biotechnology and Bioengineering, 93(6), 1095-1102 Tan, T. (2007). Understanding the effect of particle-size on UV disinfection: Kinetics, mechanism and modeling. M.A.Sc. Dissertation # 134, University of Toronto. Urbain, V., Block, J. C., & Manem, J. (1993). Bioflocculation in activated sludge: An analytic approach. Water Research, 27(5), 829-838. Wright, H.B., and Carins W.L. Ultraviolet Water Disinfection, USEPA Workshop on UV Disinfection of

Drinking Water, Arlington, VA, April 1999. Yuan, Y. (2007). Investigating the relationship between the physical structure and shear strength of bioflocs. Ph.D. Dissertation #177, University of Toronto. Yuan, Y., & Farnood, R. R. (2010). Strength and breakage of activated sludge flocs. Powder Technology,

199(2), 111-119.

65

4. The Effect of Biological Nutrient Removal on UV

Disinfection Kinetics2

4.1. Abstract

The microbial aggregates (flocs) generated in the secondary biological treatment

process protect microorganisms from exposure to UV light, decreasing the disinfection

efficiency. In this study the UV disinfection performance of two different secondary

treatment processes were compared. The two processes were: a conventional activated

sludge system (CAS), and a biological nutrient removal (University of Cape Town) (BNR-

UCT). In addition to UV disinfection performance, the physicochemical characteristics of

flocs were also compared. The BNR-UCT flocs were more open and less spherical

compared to the CAS flocs. Both the mixed liquor flocs and the secondary effluents

collected from the BNR-UCT process were easier to disinfect with UV light by 0.5 log, and

1 log, respectively. This study demonstrates the effect of the secondary process conditions

on floc properties and the UV disinfection rates, thereby providing valuable information for

modifying/designing upstream processes to increase disinfection efficiency.

4.2. Introduction

The effectiveness of ultraviolet (UV) disinfection decreases in the presence of

suspended microbial flocs formed in the activated sludge process (Qualls et al. 1983, Das

2 . This Chapter is based on the following article: Azimi. Y., Pileggi. V, Chen. X., Allen. D. G., Farnood. R., Droppo. I., Seto. P. (2013). The Effect of Secondary Biological Treatment Process Conditions on UV Disinfection of Wastewater. Water Science and

Technology, 64 (12), 2719-2723.

66

2001, Tan 2007). Tailing is representative of the disinfection efficiency reduction at higher

UV doses.

It is known that transmittance affects the UV disinfection efficiency. Loge et al.

(1999) suggested the light pathways concept inside a floc, and how target organisms that

have a relatively unobstructed pathway to the bulk medium are disinfected more easily with

UV light. Flocs are mainly composed of extracellular polymeric substances (EPS) which

are high absorbers of UV light (Farnood 2005). As floc density increases there is a higher

volumetric concentration of extracellular polymeric substances. Therefore, according the

Beer-Lamberts law there is a lower intensity of UV light inside the flocs that reduces the

rate of UV disinfection. In the previous Chapter, it was found that the compact core’s UV

DRC has a higher tailing level compared to flocs, implying that they are harder to disinfect

and likely responsible for the tailing of UV DRC (Figure 4. 1).

-3.5

-3

-2.5

-2

-1.5

-1

-0.5

0

0 20 40 60 80UV Dose (mJ/cm²)

Ln

(N

/N0

)

Cores 53-63µm

Un-sheared flocs 53-63µm

Figure 4. 1. UV DRCs of un-sheared flocs and cores of similar size

It is known that floc structure can be changed by varying the activated sludge

process conditions. Flocs generated at higher SRTs have more regular shapes, settle better,

and yield a cleaner secondary treated effluent (Liao et al. 2002, Xie &Yang 2009). Lower

67

DO concentrations in the secondary biological treatment process causes the formation of

porous flocs with poor settling properties (Wilén & Balmér 1999).

Nowadays many wastewater treatment plants have been constructed for the

purpose of biological nutrient removal (BNR). In a BNR process, anaerobic and anoxic

zones are incorporated with aerobic zones for simultaneous removal of nitrogen and/or

phosphorus in addition to carbonaceous removal (i.e., BOD, COD). Bulking is a common

problem in systems where the operation shifts in time and/or space between aerobic, anoxic

and anaerobic conditions (Eikelboom 1998, Yun et al. 2000, Martins et al. 2004,

Vaiopoulou et al. 2007). Moon et al. (2004) used sequencing batch reactors to study the

effect of anaerobic conditions on floc size and concluded that anaerobic flocs are smaller

than aerobic flocs. Martins et al. (2004) concluded that the presence of microaerophilic

zones in the anoxic stages of BNR systems leads to worsening sludge–settling

characteristics. Wilén and Balmér (1999) studied the effect of dissolved oxygen (DO)

concentrations on floc structure and settling properties. They concluded that low DO

concentrations produce sludge with poor settling characteristics mainly due to filamentous

bacterial growth and the formation of porous flocs. Moreover, they reported that by

alternating aerobic/anaerobic conditions, the number of small particles increased during the

anaerobic period and higher turbidities were observed.

Although BNR systems are widely used, there has been hardly any research done

on comparing the UV disinfection efficiency of these systems with that of a conventional

activated sludge (CAS) treatment process. Therefore, the main objective of this study was

to compare the structure and UV disinfectability of flocs generated under CAS conditions

68

and BNR conditions. The BNR was a modified UCT (University of Cape Town) process,

BNR-UCT.


4.3.1. Sample Collection from the CAS and BNR-UCT Reactors

The samples used for this study were collected from two pilot reactors in the

National Water Research Institute (NWRI), Environment Canada, Burlington, Canada.

Both reactors were equipped with identical upflow blanket clarifiers. One of the reactors

was a conventional fully aerated activated sludge system, and the other was a BNR-UCT,

which contained anaerobic, anoxic and aerobic zones. Each reactor had an operating

volume of 350 L with six identical cells in series. Baffles separated the reactor into the six

cells in order to simulate plug flow; mixed liquor flowed from one cell to the next via 25

mm diameter openings. In addition, a heating/cooling jacket was used to keep the reactors

at a temperature of 12 ± 1 ˚C, a common operating temperature at municipal wastewater

treatment plants in Ontario. The mixed liquor samples were collected from cell # 6 (the

last cell in the aerated zone) from both reactors.

The total sludge retention time for the BNR-UCT was 40 days (20 days anaerobic/

anoxic, and 20 days aerobic). From the 6 cells of the BNR-UCT reactor, cell #1 was

anaerobic, cells #2 and #3 were anoxic, and cells #4, #5, and #6 were aerobic, and hence

there was a ratio of 1:2:3 for the anaerobic, anoxic, and aerobic regions respectively. The

total sludge retention time for the CAS system was 20 days. Several operational

parameters for both reactors are listed in Table 4. 1. Schematics of the BNR-UCT and the

CAS reactors are displayed in Figure 4. 2.

69

835

626.25

1670

835

835

626.25

1670

835

(a)

835

626.25

835

626.25

835

626.25

835

626.25

(b)

Figure 4. 2. Schematic of the BNR-UCT (a), and the CAS (b) process

The influent to the pilot reactors was a primary treated sewage from the Skyway

Wastewater Treatment Plant (WWTP) located in Burlington, Ontario, Canada to NWRI.

The average influent characteristics are listed in Table 4. 2. For the BNR- UCT reactor, in

order to provide an easily accessible carbon source in the anaerobic zone for the

phosphorus accumulating organisms (PAOs), sodium acetate (Reagent, A.C.S. Anachemia

Science) was added to reach the VFA level of 120 mg/L.

70

Table 4. 1. Some of the operational parameters collected over the duration of October 2010 to January

2011 (n=26), n is the number of measurements

Reactor CAS BNR-UCT

Mixed Liquor Suspended Solids (g/L) 3.2 ± 0.8 7.5 ± 0.7

Effluent Total Suspended Solids (mg/L) 11.8 ± 6.3 6.42 ± 6

Sludge Volume Index (mL/g) 40.5 ± 4.4 132.9 ± 1.6

Table 4. 2. Average influent characteristics of primary effluent from Skyway WWTP, used as influent

to the continuous reactors, measured throughout the duration of sampling (October 2010 – January

2011)

COD BOD5 NH3-N TKN PO4-P TP-P

Parameter (n=40) (n=15) (n=77) (n=15) (n=36) (n=20)

Value mg/L 244 33 19 26 7 9.5

4.3.2. Size Fractionation

The collected mixed liquor samples were fractionated using sieve trays (U.S.A.

Standard Testing Sieve) with the nominal opening sizes of 32 and 53 microns. Collected

floc fractions were gently washed with UV disinfected RO water and back washed into a

flask. The fractionated samples prepared are referred to as “32-53 CAS” and “32-53,

BNR-UCT”, respectively from CAS and BNR-UCT reactors.

4.3.3. BNR-UCT and CAS Floc Characteristics

The particle size distribution and sphericity of mixed liquor flocs were measured

using Rapid VUE (Beckman Coulter, Miami, USA). This equipment uses image analysis

to calculate particle characteristics in the range of 10-1000µm equivalent area diameter.

Optical microscopy images were taken using an Olympus BX51 Microscope

(LECO Instruments, Mississauga, Ontario, Canada) and the Olympus Cellsens 1.4 software

using a 40X oil immersion objective lens, at a resolution of 0.16 µm.

71



used to irradiate the samples (Trojan Technologies, London, Ontario, Canada). Each

samples UV transmittance was measured at 254 nm using a UV-1700 Pharmaspec UV-Vis

Shimadzu Spectrometer (Shimadzu Scientific Instruments, Maryland, USA). An IL-1400A

radiometer (International Light, Peabody, MA) was used to measure the UV Light intensity.

The samples’ UV transmittance and the measured UV intensity of the lamp were used to

estimate the irradiation time required to achieve the desired UV dose according to the

method described by Bolton and Linden (2003).

To obtain the UV dose response curve (DRC), 100 mL of each sample was placed

in plastic Petri dishes and exposed to a range of UV doses between 10 and 60 mJ/cm2. The

irradiated samples were then cultured using the membrane filtration method following the

Standard Methods for the Examination of Water and Wastewater (# 9222 A, APHA 2001).

m-FC agar (VWR, Mississauga, Ontario) was used as the culturing media to culture the

surviving fecal coliform. The samples were incubated at 44 ºC for 24 ± 2 hours, and then

the colony forming units (CFU) were enumerated. The UV assay was performed on the

mixed liquor, as well as the secondary effluents of both reactors.

All the UV DRC data were fit into the double exponential model, which has been

previously used to characterize the kinetics of UV disinfection for wastewater flocs and

secondary effluents (Farnood 2005):

DkDkee

N

N21)1(

0

−− +−= ββ [4-1]

In this equation, D is the delivered ultraviolet dose, in mJ/cm2, N is the number of

colonies formed from the survived coliform per 100 mL of irradiated sample, and N0 is the

72

number of colonies formed per 100 mL of unexposed sample. Parameters k1 and k2 are the

UV inactivation rate constants; k1 is the inactivation constant for UV susceptible particles

(i.e. initial slope of the DRC). The parameter k2 is the inactivation constant for the UV

resistant particles (i.e. the slope of tailing region). The parameter β is the ratio of UV

resistant particles to the total number of viable particles.

4.3.5. Mechanical Strength of the BNR-UCT and CAS Flocs

In order to compare the mechanical strength of the CAS and BNR-UCT flocs, the

breakage percentage of the flocs was compared under identical hydrodynamic shear stress.

Solutions containing 32-63 µm flocs with similar solids content (0.06 %V) were prepared.

Ten milliliters of each floc sample was placed in the gap between the spindle and the cup in

a Couette flow cell (schematic provided in Figure 3. 3). The spindle speed was adjusted to

achieve the target shear rate and the sample was sheared for 10 minutes. During the

shearing process, flocs that have a higher mechanical strength than the applied turbulent

shear stress will not break.

After shearing at various degrees (controlled with the speed of the spindle ranging

from 2,000-5,000 rpm) the particle size distribution was measured using Multisizer 3

(Beckman Coulter, Miami, USA). Breakage was defined based on the number of flocs that

survive the shearing process (Yuan & Farnood 2010, Yong et al. 2009, Gibson et al. 2009):

% Breakage = (1 – 0P

Ps

N

N) x 100 [4-2]

where NP0 and NPs represent the initial (before shearing) and the final (after shearing)

number of flocs larger than the mode floc size before shearing. Based on this equation, the

73

percentage of breakage was calculated and is reported as a function of the turbulent shear

stress induced by the Couette flow.

4.3.6. Composition Analysis

To extract the EPS from sludge and measure the protein and carbohydrate content,

the cation exchange resin method was applied according to the method developed by

Frølund et al. (1996). Since at the time of this particular test the reactors were no longer in

operation, mixed liquor samples were prepared by resuspending previously freeze dried

sludge samples in milliQ water. To assess the effect of freeze drying the samples on the

measured composition, an extraction followed by composition analysis was done on an

identical sample in both the fresh and frozen dried states and no significant difference was

observed.

The carbohydrate content of the extracts was measured using the phenol-sulfuric

acid method with D-glucose as standard (Masuko et al. 2005), and the protein content was

measured using the BCATM Protein Assay kit and a modified Lowry method (Haff 1978)

using bovine serum albumin (BSA) as standard. Spectrometric readings were measured

using the multiwell-plate reader ThermoU Spectra III A-5082 from SLT-Labinstruments

(TECAN, Durham, NC, United States).


4.4.1. Microscopy and Morphological Characteristics

The optical microscopy images from the BNR-UCT and CAS flocs, taken under

identical illumination are presented in Figure 4. 3. The images show that the CAS flocs

had a higher optical density in comparison to the BNR-UCT flocs suggesting that the CAS

74

process produces denser flocs. Moreover, the BNR-UCT flocs show a more homogeneous

structure compared to the CAS flocs where there are denser regions within the flocs.

Figure 4. 3. Typical optical microscopy images of the mixed liquor flocs collected from the CAS and

BNR-UCT reactors

The BNR-UCT process generates smaller and less spherical flocs (Figure 4. 4).

Possible explanations to having smaller less spherical flocs in the BNR-UCT process

include:

1- The aerobic outer layer built up in the aerated region of the BNR-UCT process,

sloughs off when the flocs enter the anaerobic and anoxic regions. Therefore, the repeated

cycling between aerated and unaerated regions limits floc size. If methane is generated

during the anaerobic reduction of organic matter, it may diffuse out and, as a result of low

solubility in water, an accumulation of methane may occur in the internal zone of the flocs.

This may lead to the formation of bubbles which may ultimately cause sloughing off layers

75

from the flocs (Henze 2002). The formation of bubbles can also contribute to the lower

sphericity of the BNR-UCT flocs.

2- The formation of microaerophilic zones within the flocs in the anoxic region as a

result of nitrogen release during denitrification causes floc shape deformation (Martins et al.

2004). The release of nitrogen bubbles within the BNR-UCT flocs causes them to be less

spherical and more homogeneous than the CAS flocs.

0

2

4

6

8

10

15 24 38 60 96 151

Floc Size (µm)

Pa

rtic

le N

um

be

r %

BNR-UCT

CAS

(a)

0

5

10

15

20

25

0.2 0.4 0.5 0.7 0.8 1.0Sphericity

Pa

rtic

le N

um

be

r %

BNR-UCT

CAS

(b)

Figure 4. 4. Sludge floc size distribution (a) and sphericity (b) of the CAS and BNR-UCT flocs

76

4.4.2. UV Disinfection Kinetics of the BNR-UCT and CAS Flocs

UV Disinfection of BNR-UCT and CAS Mixed Liquor Flocs

In order to eliminate the effect of floc size on UV disinfection kinetics, flocs from

both processes were isolated into similar narrow size fractions (Tan 2007, Madge & Jensen,

2006). Figure 4. 5 demonstrates the normalized UV dose response curves for 32-53 µm

flocs of from both reactors as well as the fitted double exponential equation. The BNR-

UCT flocs’ DRC show a lower tailing level in comparison to the CAS flocs. The steeper

initial slope, and lower tailing level of the BNR-UCT’s UV DRC suggest that the BNR-

UCT flocs are easier to disinfect with UV light (e.g. for the same dose, there is a larger

reduction in CFU). The disinfection kinetics parameters for the data points in this graph

were extracted by fitting the DRC data into the double exponential equation using

MathematicaTM v.7 (Wolfram Research, Champaign, USA) and are listed in Table 4. 3.

The errors represent the standard deviation for each parameter from five different DRCs

constructed on different days.

The data in Table 4. 3 suggests that there is a significant difference in the UV

resistance (β) between CAS and BNR-UCT flocs (student’s t-test, p= 0.01). The double

exponential parameters values are comparable to those reported in Chapter 3.

77

0.00001

0.0001

0.001

0.01

0.1

1

0 20 40 60

UV Dose (mJ/cm²)

N /

N 0

BNR - UCTCAS

32-53 µm CAS

32-53 µm BNR-UCT

Figure 4. 5. Normalized UV dose response curves (data points and fitted double exponential model) for

the 32-53 µm flocs from the CAS and BNR-UCT flocs

Table 4. 3. UV disinfection kinetics parameters for the 32-53 µm flocs from the CAS and BNR-UCT

reactors

Reactor k1 k2 β

CAS 0.17 ± 0.03 0.019 ± 0.003 0.10 ± 0.01

BNR-UCT 0.25 ± 0.02 0.020 ± 0.003 0.06 ± 0.02

UV Disinfection of BNR-UCT and CAS Final Effluents

The normalized UV DRCs for the CAS and BNR-UCT secondary effluents are

shown in Figure 4. 6. The BNR-UCT effluent shows a higher log reduction at every UV

dose compared to the CAS effluent. As demonstrated in Figure 4. 6, at the UV dose of 10

mJ/cm2 the BNR-UCT effluent showed 3.5 log reduction whereas the CAS effluent shows

about 1.5. In terms of actual effluent fecal coliform counts the BNR-UCT had 5 times

more fecal coliform than the CAS (4×105 CFU / 100 mL vs. 8×104 CFU / 100 mL). At the

UV-dose 20 mJ/cm2 the corresponding fecal coliforms CFU / 100 mL were 40 and 1000 for

78

the BNR-UCT and CAS effluents, respectively. At the UV-dose 40 mJ/cm2 the numbers

dropped to 30 and 150 CFU/100ml.

0.000001

0.00001

0.0001

0.001

0.01

0.1

1

0 10 20 30 40 50 60

UV Dose (mJ/cm²)

N/N

0BNR - UCT

CAS

CAS Effluent

BNR - UCT Effluent

(a)

1.E+00

1.E+01

1.E+02

1.E+03

1.E+04

1.E+05

1.E+06

0 10 20 30 40 50 60

UV dose (mJ/cm²)

N

( C

FU

/ 1

00 m

L)

CAS

BNR - UCT

(b)

Figure 4. 6. Normalized (a) and actual (b) UV DRCs for the secondary effluents collected from both

the CAS and BNR-UCT reactors.

79

It should be noted that the UV DRC data was collected at least five times over the

period of December 2011 to March 2012, and the variations among the DRC of the

replicates from different days were within a third of a log (i.e. well within the experimental

error).

4.4.3. Mechanical Strength and Composition of the BNR-UCT and CAS Flocs

Although the BNR-UCT flocs have a more diffused structure and disinfect better,

they are more resistant to breaking under hydrodynamic shear stress (Figure 4. 7). The

median shear stress (τ50

), i.e. the shear stress at which the number of flocs reduced by 50

%, is listed for both the BNR-UCT flocs and the CAS flocs of a 32-63 µm fraction in Table

4. 4. Stronger flocs would have a higher τ50

, as a higher shear stress would be required to

break the flocs. The BNR-UCT flocs had a higher τ50

, implying that they were more

resistant to breakage. Similar results (higher mechanical strength of the BNR-UCT flocs)

were obtained when sonication was applied to break the flocs (not presented here). This is

despite the fact that the BNR-UCT flocs are irregular in shape and have a more open

structure.

Table 4. 4. Median 10%ile and 90%ile shear stress values for floc breakage for the BNR-UCT and

CAS flocs (32-63µm fraction)

Sample τ 50 (Pa) τ 10 (Pa) τ 90 (pa)

CAS 9.4 1.8 16.7

BNR-UCT 20 4 36

80

0

20

40

60

80

100

0 5 10 15 20 25Turbulent Shear Stress (Pa)

% B

rea

ka

ge

BNR-UCT

CAS

Figure 4. 7. Breakage percentage of the CAS and BNR-UCT flocs (32-63 µm) under various degrees of

hydrodynamic shearing with the Couette Flow

In Chapter 3 it was concluded that the compact cores in flocs are mechanically

stronger and harder to disinfect. In this study the BNR-UCT flocs are mechanically

stronger, less integrated and easier to disinfect with UV light. The new findings suggest

that the flocs’ mechanical strength and UV disinfection efficiency are not necessarily

correlated.

The composition analysis of the CAS and BNR-UCT flocs showed that the BNR-

UCT flocs have a significantly lower protein: carbohydrate ratio (2.6 ± 0.3 for the CAS

flocs vs. 1.2 ± 0.6 for the BNR-UCT flocs). This is consistent with the work of Sheng et al.

(2008) who reported that flocs with higher stability have EPS with a lower protein to

carbohydrate ratio. The finding is also in accordance with Wilén et al. (2008), as they

reported that mechanically weaker flocs were formed when the sludge protein content

increased. The lower protein to carbohydrate ratio and higher mechanical stability of the

81

BNR-UCT flocs compared to that of the CAS flocs may be explained with the oxidation

state of iron in both processes. In the BNR-UCT the sludge is cycling between aerobic and

anaerobic phases and iron is present as both ferric and ferrous, whereas in the fully aerated

CAS process ferric is the dominant iron form (Rockne 2007). As ferric iron has a higher

affinity to proteins (Murphy et al. 2000) it could cause a higher accumulation of proteins in

the CAS sludge compared to the BNR-UCT sludge. Moreover, ferrous builds stronger

bonds with EPS due to a higher ionic strength in neutral pH compared to ferric

(Oikonomidis et al. 2010). This may cause the overall sludge floc to have a higher

mechanical strength in the case of BNR-UCT process compared to CAS.

4.5. Conclusions

In this study, the effect of biological nutrient removal (specifically the BNR-UCT

process) on the UV inactivation of sludge flocs and secondary effluents was investigated.

Moreover, the mechanical strength and the composition of the flocs generated under CAS

and BNR-UCT were compared. The results showed that flocs of similar size generated in

the BNR-UCT process are more susceptible to UV disinfection than those generated under

CAS conditions. The BNR-UCT flocs show a more diffused, and less integrated structure

in comparison to the intact structure of the CAS flocs. The presence of the anaerobic and

anoxic zones and cycling between the aerated and unaerated zones in the BNR-UCT

process may be the cause of these differences observed. Although the BNR-UCT flocs

show a more open structure in the microscopy images and are less spherical in shape, they

show a higher mechanical strength in comparison to the CAS flocs. These findings show

82

that altering bioprocess conditions, such as adding extended unaerated zones, may improve

UV disinfectability.

4.6. References

American Public Health Association (APHA) (2001), Standard methods for the examination of water and wastewater. 22nd Ed. Washington DC, USA Bolton, J. R., Linden, K. G. (2003). Standardization of methods for fluence (UV Dose) determination in bench-scale UV experiments. Journal of Environmental Engineering, 129 (3), pp. 209 – 215. Das, T. K. (2001). Ultraviolet disinfection application to a wastewater treatment plant. Clean Products and

Processes, 3(2, pp. 69-80), August. Eikelboom, D. H., Andreadakis, A., Andreasen, K. (1998). Survey of filamentous populations in nutrient removal plants in four European countries. Water Science and Technology, 37(4-5), 281-289. Farnood, R., Flocs and ultraviolet disinfection. Chapter 18 in Flocculation in Natural and Engineering

Systems. I.G. Droppo, G.G. Leppard, S.N. Liss, and T.G. Milligan, Eds., CRC Press, Boca Raton, FL, 2005, pp. 385-395. Frølund, B., Palmgren, R., Keiding, K., Nielsen, P. H. (1996). Extraction of extracellular polymers from activated sludge using a cation exchange resin. Water Research, 30(8), 1749-1758. Gibson, J. H., Hon, H., Farnood, R., Droppo, I. G., Seto, P. (2009). Effects of ultrasound on suspended particles in municipal wastewater. Water Research, 43(8), 2251-2259. Haff, A. C. (1978). Mechanized micro-scale determination of protein in platelet pellet sonicates. Clinical

Chemistry, 24(11), 2031-2032. Henze, M. Ed. (2002). Wastewater treatment : Biological and chemical processes: New York: Springer. Liao, B. Q., Allen, D. G., Leppard, G. G., Droppo, I. G., Liss, S. N. (2002). Interparticle interactions affecting the stability of sludge flocs. Journal of Colloid and Interface Science, 249(2), 372-380. Loge, F. J., Emerick, R. W., Thompson, D. E., Nelson, D. C., & Darby, J. L. (1999). Factors influencing ultraviolet disinfection performance part I: Light penetration to wastewater particles. Water Environment

Research, 71(3), 377-381. Madge, B. A., Jensen, J. N. (2006). Ultraviolet disinfection of fecal coliform in municipal wastewater: Effects of particle size. Water Environment Research, 78(3), 294-304. Martins, A. M. P., Heijnen, J. J., Van Loosdrecht, M. C. M. (2004). Bulking sludge in biological nutrient removal systems. Biotechnology and Bioengineering, 86(2), 125-135. Masuko, T., Minami, A., Iwasaki, N., Majima, T., Nishimura, S., Lee, Y. C. (2005). Carbohydrate analysis by a phenol-sulfuric acid method in microplate format. Analytical Biochemistry, 339(1), 69-72. Murthy, S. N., Novak, J. T., & Holbrook, R. D. (2000). Optimizing dewatering of biosolids from autothermal thermophilic aerobic digesters (ATAD) using inorganic conditioners. Water Environment Research, 72(6-7), 714-721.

83

Oikonomidis, I., Burrows, L. J., & Carliell-Marquet, C. M. (2010). Mode of action of ferric and ferrous iron salts in activated sludge. Journal of Chemical Technology and Biotechnology, 85(8), 1067-1076. Qualls, R. G., Flynn, M. P., Johnson, J. D. (1983). The role of suspended particles in ultraviolet disinfection. Journal of the Water Pollution Control Federation, 55(10), 1280-1285. Rockne, K. J. (2007). Kinetic hindrance of fe(II) oxidation at alkaline pH and in the presence of nitrate and oxygen in a facultative wastewater stabilization pond. Journal of Environmental Science and Health - Part A Toxic/Hazardous Substances and Environmental Engineering, 42(3), 265-275. Sheng, G., Yu, H., Li, X. (2008). Stability of sludge flocs under shear conditions. Biochemical Engineering

Journal, 38(3), 302-308. Tan, T. (2007). Understanding the effect of particle-size on UV disinfection: Kinetics, mechanism and modeling. M.A.Sc. Dissertation # 134, University of Toronto. Vaiopoulou, E., Melidis, P., Aivasidis, A. (2007). Growth of filamentous bacteria in an enhanced biological phosphorus removal system. Desalination, 213(1-3), 288-296. Wilén, B., Balmér, P. (1999). The effect of dissolved oxygen concentration on the structure, size and size distribution of activated sludge flocs. Water Research, 33(2), 391-400. Wilén, B. M., Lumley, D., Mattsson, A., & Mino, T. (2008). Relationship between floc composition and flocculation and settling properties studied at a full scale activated sludge plant. Water Research, 42(16), 4404-4418. Xie, B., Yang, S. (2009). Analyses of bioflocculation and bacterial communities in sequencing batch reactors. Environmental Engineering Science, 26(3), 481-487. Yong, H. N., Farnood, R. R., Cairns, W., Mao, T. (2009). Effect of sonication on UV disinfectability of primary effluents. Water Environment Research, 81(7), 695-701. Yuan, Y., & Farnood, R. R. (2010). Strength and breakage of activated sludge flocs. Powder Technology,

199(2), 111-119.

84

5. Evidence of Disinfection by Advanced Oxidation

within Wastewater Flocs under UV Irradiation3

5.1. Abstract

In this paper, the role of naturally occurring polyphosphate in enhancing the ultraviolet

disinfection of wastewater flocs is examined. It was found that polyphosphate, which

accumulates within the wastewater flocs in enhanced biological phosphorus removal

(EBPR) processes, is capable of producing hydroxyl radicals under UV irradiation and

hence causing the photoreactive disinfection of microorganisms embedded within flocs.

This phenomenon is likely responsible for the improved UV disinfection of the biological

nutrient removal (BNR) effluent compared to that of conventional activated sludge

effluent. A mathematical model is developed that combines the chemical disinfection by

hydroxyl radical formation within flocs, together with the direct inactivation of

microorganisms by UV irradiation. The proposed model is able to explain the observed

improvement in the UV disinfection of the BNR effluents. This study shows that the

chemical composition of wastewater flocs could have a significant positive impact on their

UV disinfection by catalyzing the production of oxidative species.

3 . This Chapter is based on the following article: Azimi. Y., Allen. D. G., Droppo. I., Seto. P. , Farnood. R. (2013). Evidence of Disinfection by Advanced Oxidation within Wastewater Flocs under UV Irradiation. To be submitted

85

5.2. Nomenclature

DRC: Dose Response Curve

BNR: Biological nutrient removal

BNR-UCT: Biological nutrient removal - modified University of Cape Town

CAS: Conventional activated sludge

EPS: Extracellular polymeric substances

PAO: Phosphorus accumulating organisms

EBPR: Enhanced biological phosphorus removal

MLSS (g/L): Mixed liquor suspended solids

VSS (g/L): Volatile suspended solids

COD (mg/L): Chemical oxygen demand

SVI (mL/g): Sludge volume index

TEM: Transmission electron microscopy

RO: Reverse Osmosis

CFU: Colony forming units

TEM: Transmission electron microscopy

EDX: Energy dispersive X-ray

k” (mg.cm2/L.mW): production rate constant of •OH in the presence of polyphosphate and

UV light

k’ (L/mg.min): Inactivation rate constant of free E.coli with •OH

MB: Methylene blue

I (mW/cm2): Average UV intensity at the surface of a floc

kMB, •OH (L/mol. s): Reaction rate constant of Methylene blue and the hydroxyl radical

86

R (µm): Floc radius

δ (µm): The distance of the point of interest from the surface of the floc

Deff (δ) (mJ/cm2): The effective UV dose delivered in the distance of δ from the floc’s

surface

kUV (cm2/mJ): Inactivation rate constant of free E.coli with UV light

k•OH (L/mg.min): Inactivation rate constant of free E.coli with •OH, combined with the

production of •OH

N0(R, δ) (CFU/100mL): Initial concentration of viable flocs

N(R, δ) (CFU/100mL): Final concentration of viable flocs

keffUV(δ) (cm2/mJ): Effective UV inactivation constant at the distance δ from the surface of

the floc inwards

keff•OH(δ) (L/mg.min): Effective •OH inactivation constant at the distance δ from the surface

of the floc inwards

A (cm-1): UV absorbance of EPS

f (R, δ): Probability density function of R and δ

V (µm3): Total floc volume

δs (µm): Shell thickness in the shell-core model used for the CAS flocs

5.3. Introduction

Ultraviolet disinfection is a well established technology for disinfecting secondary

wastewater effluents. UV light, specifically around the germicidal wavelengths (200-300

nm), inactivates microorganisms by penetrating through the cell wall and altering the

microorganism’s DNA (Jagger 1967, Pfeifer et al. 2005, Piluso & Moffatt-Smith 2006).

87

This is widely believed to be the main mechanism for the UV disinfection of wastewater

effluents.

It has been reported that UV light can also cause disinfection through the

production of highly oxidative species such as hydroxyl radicals with the addition of

photocatalysts. Such species could result in the oxidation of cell wall and cause cell death

(Matsunaga et al. 1985, Maness et al. 1999). In particular Cho et al. (2004) measured the

release of •OH in the UV/TiO2 process, and found a linear correlation between the

concentration of hydroxyl radicals and the rate of free E.coli inactivation. Their study

showed that •OH is approximately 3 to 4 orders of magnitude more effective for E.coli

inactivation than common disinfectants such as chlorine and ozone.

Over the past two decades, many wastewater treatment plants have been

constructed for the purpose of biological nutrient removal (BNR) (Trivedi 2009). In a

BNR process, anaerobic and anoxic zones are incorporated with aerobic zones for

biological removal of nitrogen and/or phosphorus in addition to carbonaceous nutrients.

Enhanced biological phosphorus removal (EBPR) is one of the processes involved in some

biological nutrient removal systems, where polyphosphate accumulating organisms (PAOs)

remove phosphorus from wastewater and store it in the form of polyphosphates. In the

aerobic zone of EBPR, the PAOs oxidize the polyhydroxyalkanoates (which were

synthesized by the PAOs in the anaerobic region) and the energy reserves from oxidation

are stored through phosphate uptake and polymerization (WEF & ASCE 2005, Hirota et al.

2010). The total P-content in the sludge can reach 10–15 % (Stowa 2002). Hirota et al.

(2010) heated EBPR sludge to extract the polyphosphates, and applied polyacrylamide gel

electrophoresis to determine the chain length of the released polyphosphate. They

88

concluded that the main compounds stored by the PAOs are polyP (with chain lengths of

100-200 Pi residues) and trimetaphosphate. Lee et al. (2010) studied the positioning of

PAOs in flocs generated in the biological phosphorus removal process and concluded that

they were distributed evenly throughout the floc. However, they reported that phosphate

concentrations increased as the depth increased to the centre of the flocs, which implies that

some of the polyphosphate likely exists within the EPS outside of the PAO cells. As seen in

chapter 4, flocs formed in the BNR-UCT process (where EBPR occurs) are more readily

disinfected with UV light compared to the conventional activated sludge flocs.

In this study it was hypothesized that the phosphorus compounds, accumulated in

EBPR flocs, can catalyze the production of oxidative species within the flocs under UV

irradiation and enhance UV disinfection. The main objectives of this study were to: 1)

explore the effect of polyphosphate storage on the UV disinfection kinetics of PAO-

containing sludge flocs, and 2) develop a mathematical model for the UV disinfection of

PAO-containing flocs that takes into account both the photoreactive disinfection due to the

production of oxidative species within these flocs and the direct UV inactivation of

microorganisms.


5.4.1. Sample Collection

The samples used for this study were collected from two continuous pilot scale

reactors (effective volume of 350 L), treating real municipal wastewater, and operating at

the National Water Research Institute (NWRI), Environment Canada, Burlington, Canada.

Both reactors were equipped with identical secondary clarifiers, and had recycling sludge

streams. One of the reactors was a conventional fully aerated activated sludge system, and

89

the other was a modified University of Cape Town biological nutrient removal system

(BNR-UCT), which contained anaerobic, anoxic and aerobic zones for simultaneous

nitrogen and phosphorus removal. A heating/cooling jacket was used to keep the reactors at

a temperature of 12 ± 1 ˚C, a common operating temperature at municipal wastewater

treatment plants in Canada. The performance and stable operating conditions of the

reactors were determined by monitoring the mixed liquor suspended solids (MLSS),

volatile suspended solids (VSS), chemical oxygen demand (COD), and sludge volume

index (SVI). Details of the influent COD, total nitrogen and phosphorus in addition to the

operating conditions and reactor performance have been described in Chapter 4.


The UV assay was performed as section 4.3.4.

5.4.3. Transmission Electron Microscopy

For transmission electron microscopy (TEM) imaging, the sample preparation

technique was adopted from Liss et al. (1996). The sections made after polymerizing the

samples were observed in transmission mode (TEM) at an accelerating voltage of 80 kV

using a JEOL 1200 EXII TEMSCAN scanning transmission electron microscope.

5.4.4. Photoreactive Properties of Polyphosphates

Trisodium trimetaphosphate (Na3P3O9 > 95%, Sigma Aldrich), was used as the

model compound for this study. This compound is one of the most commonly occurring

derivatives of polyphosphates in natural environments and bacterial cells (Hupfer et al.

2008, Hirota et al. 2010). The photoactivity of this compound was examined by measuring

the discoloration of methylene blue under UV light according to the method described by

90

Zhang et al. (2012). The UV light source used for this experiment was a low pressure

mercury lamp with a main emission line at 254 nm (Trojan Technologies, London, Ontario,

Canada). Several one hundred milliliter oversaturated solutions of trisodium

trimetaphosphate (250 g/L) were prepared, and 10 mg/L methylene blue (0.1% w/v

aqueous solution, Fisher Scientific) was added to each. The solutions were then exposed to

UV light and samples were taken using a 4 mL syringe at various exposure times. These

samples were then passed through a 0.22 µm filter and their absorbance was measured at

665 nm (peak absorbance of methylene blue) using a Lambda 35 UV-Vis Spectrometer

(Perkin-Elmer, Woodbridge, Ontario, Canada). The discoloration percentage of methylene

blue was calculated at each exposure time. For comparison, the test was repeated by

substituting trisodium trimetaphosphate with 1 g/L titanium dioxide TiO2 (>99.5 %,

technical grade, Sigma Aldrich), a well-known photocatalyst. To determine the mechanism

of photoactivity of trisodium trimetaphosphate, the experiment was repeated in the

presence of methanol (ACS grade, BDH) at a concentration of 15 g/L. Methanol is known

to be a strong hole scavenger (Chen et al. 2005).


5.5.1. Transmission Electron Microscopy

The Transmission Electron Microscopy (TEM) images of the CAS and BNR-UCT

flocs show that the BNR-UCT sludge had more clustered colonies of microorganisms than

the CAS flocs (Figure 5. 1). Moreover, an accumulation of polyphosphate crystals in the

BNR-UCT cells was evident, and confirmed by detecting an increased concentration of

phosphorus in the energy dispersive X-ray (EDX) analysis of the intense black spots.

91

Similar images, showing the accumulation of polyphosphates in EBPR sludge have been

reported previously (WEF & ASCE 2005, Hupfer et al. 2008, Hirota et al. 2010).

2µm

CASBNR-UCT

2µm

Poly-Phosphate

Figure 5. 1. Transmission electron microscopy (TEM) images from the flocs collected from the CAS

and BNR-UCT processes.

5.5.2. UV Inactivation of the BNR-UCT Flocs and the Photoactive Role of

Polyphosphates

In Chapter 4, where the UV disinfection kinetics of the flocs produced in the BNR-

UCT and CAS processes were compared, it was concluded that BNR-UCT flocs are easier

to disinfect with UV.

As shown in Figure 5. 2, trisodium trimetaphosphate, a model polyphosphate

compound, showed strong photoreactive properties comparable to titanium dioxide under

exposure to UV light. Note that in the absence of either UV light or trisodium

trimetaphosphate, no significant discoloration of methylene blue was observed. Moreover,

the addition of methanol, which is known to be a hole scavenger (Chen et al. 2005),

inhibited the discoloration of methylene blue.

92

Using the kinetic data for the discoloration of methylene blue, the concentration of

hydroxyl radicals within the reaction media can be estimated. The kinetic equation for the

reaction of methylene blue (MB) and hydroxyl radical is as follows:

•OH + MB ─> Products

]].[[][

, MBOHkdt

MBdOHMB •−= • [5-1]

where OHMBk •, ~ 6 × 1010 L. mol-1 s-1 (Alpert et al. 2010). By rearranging Equation 5-2 and

integration:

=)][

][ln(

0MB

MBtOHk OHMB ×•− • ][, [5-2]

Using the experimental data presented in Figure 5. 2, the concentration of [•OH]

was estimated to be ~ 10-9 mg/L.

These findings suggest that polyphosphates could act as a photocatalyst in

producing highly oxidative hydroxyl radicals under UV irradiation. Therefore, it is quite

likely that similar oxidative species could be formed within the BNR-UCT flocs during

their exposure to UV light causing cell lysis within the flocs and improving their

disinfection. An alternative and complimentary mechanism for improved disinfection

could be the porous nature of the BNR flocs. It has been shown that the release of micro-

bubbles within the BNR-UCT flocs in the anaerobic and anoxic regions gives them a more

open and homogeneous structure, and limits the formation of compact EPS zones

(protecting the fecal coliform from UV light, causing tailing) compared to the CAS flocs

(as discussed in Chapter 4).

93

-5

5

15

25

35

45

55

0 5 10 15

Exposure to UV (minutes)

% D

isc

olo

rati

on

of

MB

MB + UV + Na3P3O9

MB + UV + TiO2

MB + UV + Na3P3O9 + MeOH MB + UV

20 30 40

UV Dose (mJ/cm²)

18 37 50-5

5

15

25

35

45

55

0 5 10 15

Exposure to UV (minutes)

% D

isc

olo

rati

on

of

MB

MB + UV + Na3P3O9

MB + UV + TiO2

MB + UV + Na3P3O9 + MeOH MB + UV

20 30 40

UV Dose (mJ/cm²)

18 37 50

Figure 5. 2. Discoloration of methylene blue (MB) under UV light with Na3P3O9 (with and without

methanol (MeOH)), and TiO2

5.6. UV Disinfection Model Including the Effect of Advanced

Oxidation

5.6.1. Model Derivation

A model is presented here to mathematically describe the observed UV disinfection

kinetics of the BNR-UCT flocs with an embedded photooxidative agent. As discussed

earlier, unlike CAS flocs, disinfection of BNR-UCT flocs occurs by two different

mechanisms namely: (1) direct DNA damage caused by UV irradiation, and (2) cell death

by in-situ advanced oxidation catalyzed by the presence of polyphosphates. In addition, as

discussed earlier, BNR-UCT flocs have a more homogeneous structure compared to CAS

flocs (as discussed in section 4.4.1).

Accordingly, the following assumptions were made in constructing the

mathematical model:

94

• The UV disinfection of floc is governed by the inactivation of the most shielded

indicator organism; i.e. an embedded indicator organism that receives the least

amount of UV dose

• Disinfection of BNR-UCT flocs is due to both DNA damage by UV irradiation

and cell death by advanced oxidation process stemming from the production of

hydroxyl radicals via the photoactivity of polyphosphates. CAS flocs, however,

are only disinfected through the DNA damage alone.

• The BNR-UCT flocs are represented by a continuum sphere with a uniform

density and a uniform UV absorbance. In contrast, the CAS flocs are

represented by a loose and porous outer shell, and a compact core; which are

both represented by continuum spherical shell and core layers, each having an

average density and UV absorbance. The density of compact core is assumed

to be double of that of the loose shell. This is supported by findings of

Hriberšek et al. (2011) & Li and Ganczarczyk (1987) who reported that flocs

with a more compact internal structure have a porosity of about 45% compared

to the 96-99% porosity of loose flocs.

• The probability of finding a culturable microorganism within a floc increases

linearly with distance from the centre of the floc. Therefore, there is a higher

chance of finding a culturable microorganism in the outer layers of a floc.

• The UV light intensity in various points of the floc may be expressed by

exponential decay (i.e. Beer-Lambert’s law).

• Floc surface is uniformly exposed to UV light.

95

• Hydroxyl radicals generated in the vicinity of polyphosphate granules diffuse

through and react with the floc material. The extent of diffusion of •OH can be

as much as 100 µm before [•OH] decreases by one log (see Appendix 1).

Recognizing that BNR-UCT flocs contain multiple polyphosphate granules, for

simplicity, the concentration of polyphosphate within these flocs is considered

to be uniform.

• The rate of disinfection of microorganisms by advanced oxidation process, (dN/

dt)AO, is assume to follow the first order kinetics with the rate constant of k•OH

(in the case of disinfection with •OH) , i.e.:

(d N/ dt)AO = – k•OH N [5-3]

• The rate constant k•OH in the above equation is expected to increase with •OH

concentration and here it was assumed to be proportional to [•OH]:

k•OH = k’ [•OH] [5-4]

The concentration of hydroxyl radicals within the floc [•OH] in turn is a function of

UV intensity and was estimated by the following equation:

[•OH] = k” I [5-5]

where k” is the production rate constant of [•OH] by UV radiation in the presence of

polyphosphate. Hence, the disinfection rate by advanced oxidation is given by:

(d N/ dt)AO = – k’ × k” × I × N [5-6]

or equivalently:

(d N/ dt)AO = – k•OH × N [5-7]

where k•OH = k’ × k” × I. [5-8]

96

The value of k” was calculated from the MB discoloration data. The concentration

of •OH was estimated as 10-9 mg/L, and the UV intensity was calculated according to

Bolton & Linden (2003). Using Equation [5-5] the value of k” was estimated: k”~ 3 x 10-8

mg.cm2/L.mW.

Hence, by combining the effects of advanced oxidation and direct UV disinfection

of microorganisms, the inactivation kinetics of a BNR-UCT floc with radius R can be

expressed by:

N(R, δ) = No(R, δ).)()( δeffOHUV Dkk

e •+− [5-9]

Here, No(R, δ) and N(R, δ) are the initial and final (i.e. after UV irradiation)

concentrations of viable flocs; i.e. flocs with at least one culturable embedded coliform

bacteria, expressed as the number of colony forming units (CFU) per 100 mL. δ is the

distance of most shielded indicator organism from the surface of the floc. The constants

k•OH and kUV are the inactivation rate constants of free E.coli with •OH and UV,

respectively. Deff(δ) is the effective UV dose inside the floc at the distance δ from the

surface. Since in practice the average UV dose at the floc surface, D, is of significance,

Equation [5-4] can be rearranged to obtain:

N(δ, R) = No(δ, R) . Dkk OHeffUVeffe

))()(( ,, δδ •+− [5-10]

where keff,UV(δ), and keff,•OH(δ) are the effective inactivation constants inside the floc at the

distance δ from the surface. Based on the Beer-Lambert’s law:

keff,UV (δ) = kUV e −A(δ) × δ [5-11]

keff,•OH (δ) = k•OH e −A(δ) × δ [5-12]

The parameter A represents the average absorbance of EPS. Given that δ varies

from 0 to R; the UV DRC of BNR flocs with radius R would be:

97

N(R, δ) /No(R, δ) = ∫oR

f (R, δ) e – (keff,UV

(δ) + keff,•OH

(δ))D dδ [5-8]

where f (R, δ) is the probability density function of δ. For the BNR-UCT flocs f (R,

δ) is defined as:

f (R, δ) =V

R

RdRR

)(2)(

3

4)(

3

42

33 δδδπδπ

−××

+−−− [5-9]

The term 2/R2 × (R – δ), accounts for the functionality of microorganism

positioning with the floc radius, and V is total floc volume.

For the CAS flocs, a similar equation can be used for the floc loose shell; however,

for the dense core the effect of increased density must be taken into account:

f (R, δ) =V

R

RRdRR s

)(2)(

3

42)(

3

4)(

3

42

333 δδπδδπδπ

−××

−×++−−−

when δ> δs [5-10]

and

f (R, δ) =V

R

RdRR

)(2)(

3

4)(

3

42

33 δδδπδπ

−××

+−−− when δ< δs [5-11]

5.6.2. Comparison with Experimental Data

To apply the above mathematical model, knowledge of model paramteres, i.e. k•OH ,

kUV , A, and δs , is required. These parameters were obtained, or if necessary estimated,

from the experimental data available in the literature.

Using the experimental data provided by Cho et al. (2004) for the inactivation of

free E.coli, the •OH inactivation rate constant was estimated. By plotting the E.coli

concentration as a function of •OH concentration on a semi-logarithmic scale, and using

linear regression analysis the value of k’ was determined to be 2 × 106 L/mg.min.

98

Assuming that the floc material is saturated with polyphosphates (this is reasonable since

polyphosphates are present in the solid form) the concentration of hydroxyl radicals within

the flocs can be estimated from that of saturated solution of trisodium trimetaphosphate.

Accordingly, using the average intensity of UV light, determined based on Bolton &

Linden (2003), k”, which relates the •OH concentration to UV intensity, was estimated to

be 3 x 10-8 (mg.cm2/L.mW). Hence, the value of k•OH can be calculated as 0.06 cm2/mJ.

For the UV inactivation rate constant of free E.coli the data reported by Wright &

Cairns (1999) was used that results in kUV ~ 0.67 cm2/mJ. The UV absorbance of EPS

material, A, was assumed to be equal to that of densely packed cells or approximately 1000

cm-1 (based on Jagger 1967).

As mentioned previously, the BNR-UCT flocs were modeled as homogeneous

spheres, while the shell-core model (see Chapter 3) was used for the CAS flocs. For

disinfection of CAS flocs, the value of δs ~ 20 μm was used. Using the estimated model

parameters, UV DRCs of 48 µm flocs for both BNR-UCT and CAS processes were

constructed and given in Figure 5. 3. Based on this figure, the homogeneous floc with

advanced oxidation (representing BNR-UCT) showed improved UV disinfectability

compared to the shell-core floc with no advanced oxidation (representing CAS). It is worth

noting that the no “fitting” parameter was used in constructing the predicted UV DRCs of

BNR-UCT and CAS flocs in this figure. In spite of this, the predicted DRCs in Figure 5. 3

agree well with the experimental UV DRCs of the CAS and BNR-UCT flocs.

99

0.0001

0.001

0.01

0.1

1

0 20 40 60

UV Dose (mJ/cm²)

N /

N 0

Core-shell Model with UVDisinfection (CAS)

Homogeneous Model with UV andAdvanced Oxidation Disinfection(BNR-UCT)

Figure 5. 3. The estimated UV DRC for the shell-core model with only direct UV inactivation

(representing CAS flocs), and for the homogeneous model with direct UV inactivation and advanced

oxidation (representing BNR flocs). UV absorbance of EPS: 1000 cm−1

, floc size: 48 µm. The solid line

is estimated UV DRCs and the scattered data points are experimental data.

5.6.3. The Signifficance of Advanced Oxidation in Floc Disinfection

To better illustrate the relative signifficance of advanced oxidation in terms of floc

disifnection, a parametric study was conducted. Using the above mathematical model, the

UV DRC of model BNR-UCT flocs without advanced oxidation and with various

concentrations of hydroxyl radicals were calculated. Figure 5. 4 shows that in the presence

of advacned ocidation, the UV disinfectability of flocs improves significantly; and as

expected, increasing the concentration of •OH further improves the UV disinfection rate

and lowers the tailing level of UV DRC. A closer examination of this figure shows that by

increasing the •OH concentration within the BNR-UCT floc from 0 mg/L to 5×10-9 mg/L

the tailing level decreased by about 2 logs. Furthermore, the predicted value for the UV

dose required for 3 log reduction of CFU decreased by a factor of more than 5, from ~ 55

mJ/cm2 to merely ~10 mJ/cm2.

100

0.00001

0.0001

0.001

0.01

0.1

1

0 10 20 30 40 50 60

UV Dose (mJ/cm²)

N /

N 0

No Advanced Oxidation (AO)

With AO, •OH = 10^-9 mg/L

With AO, •OH = 2 x 10^-9 mg/L

With AO, •OH = 5 x 10^-9 mg/L

Figure 5. 4. The effect of •OH on the disinfectability of homogeneous spherical flocs with UV light, and

the estimated effect of increasing •OH concentration. (floc size: 48 µm)

It should be noted here that the photoactive role of polyphosphate and its effect on

the additional disinfection of BNR-UCT flocs is one explanation for the lower tailing level

observed in the DRC. As discussed earlier, the morphological differences induced by

biological nitrogen removal (lower sphericity) may also affect the flocs’ UV susceptibility.

Further investigation regarding the effects of each process is required to clarify their

relative contributions to the observed improvements in the UV disinfection of BNR flocs.

5.7. Conclusions

This study shows that advanced oxidation due to the naturally occurring

polyphosphate material may play a significant role in improving disinfection of EBPR

effluents under UV irradiation. Polyphosphate, naturally formed within the flocs during

the EBPR processes, is proposed to act as a photocatalyst to produce hydroxyl radicals

under UV irradiation and accelerate disinfection through cell lysis. This finding provides

101

new and significant opportunities in terms of optimizing upstream treatment processes to

improve the UV disinfection performance of effluents.

5.8. References

Alpert, S. M., Knappe, D. R. U., Ducoste, J. J. (2010). Modeling the UV/hydrogen peroxide advanced oxidation process using computational fluid dynamics. Water Research, 44(6), 1797-1808. Bolton, J. R., Linden, K. G. (2003). Standardization of methods for fluence (UV Dose) determination in bench-scale UV experiments. Journal of Environmental Engineering, 129 (3), pp. 209 – 215. Chen, Y., Yang, S., Wang, K., Lou, L. (2005). Role of primary active species and TiO2 surface characteristic in UV-illuminated photodegradation of acid orange 7. Journal of Photochemistry and Photobiology A:

Chemistry, 172(1), 47-54. Cho, M., Chung, H., Choi, W., Yoon, J. (2004). Linear correlation between inactivation of E. coli and OH radical concentration in TiO2 photocatalytic disinfection. Water Research, 38(4), 1069-1077. Hirota, R., Kuroda, A., Kato, J., Ohtake, H. (2010). Bacterial phosphate metabolism and its application to phosphorus recovery and industrial bioprocesses. Journal of Bioscience and Bioengineering, 109(5), 423-432. Hriberšek, M., Žajdela, B., Hribernik, A., & Zadravec, M. (2011). Experimental and numerical investigations of sedimentation of porous wastewater sludge flocs. Water Research, 45(4), 1729-1735. Hupfer, M., Glöss, S., Schmieder, P., Grossart, H. (2008). Methods for detection and quantification of polyphosphate and polyphosphate accumulating microorganisms in aquatic sediments. International Review

of Hydrobiology, 93(1), 1-30. Jagger, J., (1967) Introduction to Research in Ultraviolet Photobiology. Prentice-Hall, Inc. Englewood Cliffs, NJ Lee, W. H., Bishop, P. L. (2010). In situ microscale analyses of activated sludge flocs in the enhanced biological phosphate removal process by the use of microelectrodes and fluorescent in situ hybridization. Journal of Environmental Engineering, 136(6), 561-567. Li, D., & Ganczarczyk, J. J. (1987). Stroboscopic determination of settling velocity, size and porosity of activated sludge flocs. Water Research, 21(3), 257-262. Liss, S. N., Droppo, I. G., Flannigan, D. T., & Leppard, G. G. (1996). Floc architecture in wastewater and natural riverine systems. Environmental Science and Technology, 30(2), 680-686. Maness, P., Smolinski, S., Blake, D. M., Huang, Z., Wolfrum, E. J., & Jacoby, W. A. (1999). Bactericidal activity of photocatalytic TiO2 reaction: Toward an understanding of its killing mechanism. Applied and

Environmental Microbiology, 65(9), 4094-4098. Matsunaga, T., Tomoda, R., Nakajima, T., & Wake, H. (1985). Photoelectrochemical sterilization of microbial cells by semiconductor powders. FEMS Microbiology Letters, 29(1-2), 211-214. Pfeifer, G. P., You, Y., & Besaratinia, A. (2005). Mutations induced by ultraviolet light. Mutation Research - Fundamental and Molecular Mechanisms of Mutagenesis, 571(1-2 SPEC. ISS.), 19-31.

102

Piluso, L. G., & Moffatt-Smith, C. (2006). Disinfection using ultraviolet radiation as an antimicrobial agent: A review and synthesis of mechanisms and concerns. PDA Journal of Pharmaceutical Science and

Technology, 60(1), 1-16 Stowa, . (2002). Biological phosphorus removal. London: IWA Trivedi, H. K. (2009). Simultaneous nitrification and denitrification. In Advanced biological Treatment

Processes; Wang, L. K., Shammas, N. K., Hung, Y. T., Eds.; Humana Press: Totowa, NJ, pp. 186-189. Water Environment Federation., & Environmental and Water Resources Institute (U.S.). (2006). Biological

nutrient removal (BNR) operation in wastewater treatment plants. New York: McGraw-Hill. Wright, H.B., and Carins W.L. Ultraviolet Water Disinfection, USEPA Workshop on UV Disinfection of

Drinking Water, Arlington, VA, April 1999. Zhang, Q., Li, C., Li, T. (2012). Rapid Photocatalytic Degradation of Methylene Blue under High Photon Flux UV Irradiation: Characteristics and Comparison with Routine Low Photon Flux. International Journal

of Photoenergy, vol. 2012, Article ID 398787, 7 pages.

103

6. The Effect of Activated Sludge Process Conditions

on UV Disinfection: Temperature, Sludge Retention Time,

Influent Phosphorus Levels4

6.1. Abstract

The effect of sludge retention time (SRT), temperature and influent phosphorus levels on

the ultraviolet (UV) disinfection of effluent was investigated using a pilot-scale sequencing

batch reactor (SBR). Increasing the operating temperature from 12 ºC to 22 ºC improved

the tailing level of UV dose response curve (DRC) of the final effluent by 0.5 log. A

similar improvement was observed for the 75-90 µm flocs that were collected and

fractionated from the mixed liquor. The improvement in the floc disinfectability was

accompanied by an increase of 0.1-0.2 in the fractal dimension. Introducing a feed with

phosphorus limitation, i.e. COD:N:P of 100:10:0.3, caused the formation of rounder flocs

with improved settling characteristics, however these flocs were more resistant to UV

disinfection resulting in a higher tailing level of UV DRC (by 1 log). Starving phosphorus

conditions, i.e. COD:N:P of 100:10:0.03, decreased the average floc size and sphericity,

both by 50%, and improved the final effluent UV disinfection by at least 1 log, compared

to the normal and limited phosphorus conditions. The results also showed that the UV

disinfectability of the mixed liquor flocs and that of the final effluent did not change

4 This Chapter is based on the following article: Azimi. Y., Allen. D. G., Farnood. R., Seto. P.. The Effect of Activated Sludge Process Conditions on UV Disinfection: Temperature, Sludge Retention Time, Influent Phosphorous Levels. To be submitted

104

significantly when increasing the SRT from 7 to 20 days. Similarly, no clear trend was

observed between either of the flocs’ protein to carbohydrate ratio or the flocs’ mechanical

stability with UV disinfectability.

6.2. Introduction

Ultraviolet disinfection is a well-established, environmentally friendly technology

for effluent disinfection. However, the effectiveness of UV disinfection decreases in the

presence of suspended microbial flocs formed in the activated sludge process (Qualls et al.

1983, Das 2001). In a typical UV dose response curve (DRC) this decrease in the

efficiency is detected at higher UV doses, referred to as tailing.

The tailing phenomenon is generally attributed to the absorbance of UV light by

the floc material and hence shielding of the embedded microorganisms. The reduction of

UV transmittance within flocs is due to the presence of extracellular polymeric substances

(EPS), which are strong absorbers of UV light (Kalisvaart 2004, Farnood 2005). Hence,

UV disinfectability of flocs is expected to be a function of floc structure and composition.

Loge et al. (1999) suggested that the transmittance of UV light within a floc would

affect its UV disinfectability, and proposed the light pathways concept. They suggested that

the target organisms with a relatively unobstructed pathway to the bulk medium are

disinfected more easily. In Chapter 3 it was suggested that the formation of compact EPS

regions could create two types of flocs: 1) flocs with a homogeneous loose structure, and 2)

flocs with a double layer structure of a compact core surrounded by a loose shell.

Accordingly, they developed a structure-based model to describe the UV disinfection

kinetics of wastewater flocs.

105

It is expected that as floc compactness increases, the UV transmittance within the

floc decreases, and hence the effective dose delivered to embedded microorganisms within

the floc decreases. Floc compactness is expected to be related to its chemical and

mechanical stability (Liao et al. 2002, Yuan & Farnood 2010, Sheng et al. 2008). Flocs

generated at higher SRTs were found to have higher protein content, settle better, and have

a higher chemical stability (Liao et al. 2002, Xie and Yang 2009). Xie and Yang (2009)

reported that the loosely bound EPS content decreased as SRT increased and was the cause

of good settleability and low sludge volume index. However, it has been reported that

neither the EPS composition (protein, carbohydrate, and DNA content) nor the UV

disinfectability of the flocs was significantly affected by varying SRT from 4 to 20 days

(Scott et al. 2005). Xie and Yang (2009) as well as Li and Yang (2007) showed that

changing the carbon source from glucose to acetate had a significant effect on improving

sludge settleability. Li and Yang (2007) studied the effect of SRT ranging from 5 to 20

days on the loose, and tightly bound EPS content, and the settling properties of sludge flocs.

They found that sludge at lower SRTs had a higher loosely bound EPS content which

caused a higher sludge volume index. However, the tightly bound EPS content was

independent of the SRT in the range tested. Liss et al. (2002) investigated the effect of SRT

on microbial floc structure using various microscopy techniques and reported that flocs

formed at higher SRTs were more stable, more hydrophobic and less negatively charged.

Liao et al. (2002) examined the chemical stability of wastewater flocs and found that flocs

at higher SRTs are more stable, and have a dense core associated with their structure that

due to the limitations of diffusion of oxygen and nutrients has a growth rate lower than that

of the outer layer. In addition to the effect of SRT on floc structure, increase in SRT has

106

been reported to result in an exponential decrease in the fraction of flocs with associated

coliform (Loge et al. 2002).

Similarly, earlier studies have shown that the operating temperature of the

wastewater treatment process could have a significant impact on the floc composition and

structural characteristics. Liao et al. (2011) found that aerobic flocs generated under

mesophilic (35ºC) conditions settle better compared to those generated under thermophilic

(55ºC) conditions, due to a higher level of filaments. Morgan-Sagastume and Allen (2005)

reported that temperature shifts from 30˚ to 45 ˚C causes deflocculation and an increase in

the final effluent’s suspended solids. Tian et al. (1994) examined the effect of operating

temperature from 20 ˚C to 8 ˚C and suggested that under cold weather conditions the viable

organisms will comprise a smaller percentage of the total activated sludge. Wilén et al.

(2010) showed that flocs generated in the summer have more and larger microbial colonies

and correlated this observation to the metabolic activity differences which affects the

growth pattern of the microbial species.

The effect of phosphorus limitation on floc structure has been studied by Liu and

Liss (2007). The response of the activated sludge process to the phosphorus limited

conditions (COD:N:P, 100:5:1 switching to 100:5:0.3) was the production of flocs that

were larger, rounder, and more compact, which improved their settling characteristics. In a

more recent study, Dumitrache (2008) found that flocs generated at phosphorus limited

conditions (100:5:0.33) were larger with more densely packed biomass, observed as darker

clusters in optical microscopy images. Similar observations were made by Brei (2011).

Although the above literature suggest that wastewater treatment process conditions

could have a significant effect on the UV disinfectability of final effluent, very little

107

systematic study has been done to better quantify such effects. Therefore, the main

objective of this chapter was to compare the UV disinfectability of flocs generated under

various SRTs, temperatures, and nutrient (phosphorus) conditions in relation to floc

characteristics. Results from this study could provide a basis for a better and holistic

design and operation of wastewater treatment processes for improved disinfection

performance of final effluents.


6.3.1. Sample Collection

The samples used for this study were collected from two sets of pilot sequencing

batch reactors (SBR), operating at Environment Canada’s National Water Research

Institute (NWRI), located in Burlington, Canada. Schematics of the reactors are presented

in Figure 6. 1.

The first set of reactors consisted of four SBR systems, each having an operating

volume of 20L and using primary treated sewage from the Skyway Wastewater Treatment

Plant (WWTP) located in Burlington as the feed. In this chapter, these reactors are referred

to as the “large reactors”. The COD, BOD5, TKN and TP values for the influent to the

reactors were: 195 ± 80 mg/L, 125 ± 60 mg/L, 43 ± 7 mg/L, and 6 ± 1.3 mg/L,

respectively. Each reactor had a heating/cooling jacket keeping it either at 22˚C or 12 ˚C,

the pH was kept between 7.3-7.5 by adding sodium bicarbonate (ACS grade, Anachemia),

and the dissolved oxygen concentration was maintained between 2-2.5 mg/L, through

aeration in the reactor media. These large reactors were operated to study the effect of SRT

and temperature (T) on UV disinfection. From the four reactors, two were operating at an

108

SRT of 20 days and the other two at an SRT of 7 days. One reactor at each operating SRT

was at 12 ºC and the other operated at 22 ºC.

The second set of SBR reactors consisted of three reactors, each having an

operating volume of 2L. In this document, these reactors are referred to as the “small

reactors”. These small reactors were used to study the effect of phosphorus deficiency (i.e.

the influent’s chemical oxygen demand (COD) to phosphorus ratio (COD:P)) on UV

disinfection performance. To better control the nutrient ratio, a synthetic feed composed of

glucose and inorganic salts was used as the influent. The carbon loading of the synthetic

feed was 468.6 mg/L of glucose (Bioreagent grade, Sigma Aldrich) which was equivalent

to a COD of 500 mg/L. Nitrogen and phosphorus were added in the form of (NH4)2SO4,

KH2PO4, K2HPO4. For the starving phosphorus conditions (P-Starved), concentrations of

the above micronutrients were 236 mg/L, 0.33 mg/L, and 0.5 mg/L; respectively, to create

a COD:N:P of 100:10:0.03. For the limited (P-Limited) and normal (P-Normal)

phosphorus conditions, additional potassium phosphate was added to reach COD:N:P of

100:10:0.3 and 100:10:1 ratios, respectively. The concentrations of these micronutrients

are summarized in Table 6. 2. The pH was kept at 7.2-7.5 by adding sodium bicarbonate

(ACS grade, Anachemia). The operating SRT for these reactors was kept at 17 to 20 days,

and the operating temperature was 19 ± 2 ºC. The synthetic feed was refrigerated at 4 ºC,

and a water bath was used to warm up the feed to 20 ˚C before entering the reactors. A

continuous air flow of 56 L/min was supplied for each reactor. The operational conditions

for the SBR reactors are listed in Table 6. 1.

The sequencing batch operation for all seven reactors was controlled by a

programmable logic controller (EATON EZ-719-AC module with an EZ 618 AC

109

expansion module). The filling, reaction, waste, settling, and decanting durations were 30,

200, 5, 70, 55 minutes, respectively. Hence, the duration of a complete cycle was 6 hours

that translated to a total of 4 cycles per day. Peristaltic pumps (Cole-Parmer Instrument Co.,

Montreal, Quebec) were used to transfer fluids through the reactors.

Table 6. 1. Operational conditions of the pilot reactors. SRT: sludge retention time, HRT: hydraulic

retention time, T: temperature

Reactor

SRT

(days)

HRT

(h)

Effective

Volume (L)

Feed T

(˚C)

Large Reactors

C (SRT 20 T 12) 20 12

B (SRT 20 T 22) 20 22

D (SRT 7 T 12) 7 12

E (SRT 7 T 22) 7

9

20

Primary Effluent

(Skyway)

22

Small Reactors

P-Normal COD:P, 100:1

P-Limited COD:P, 100:0.3

P-Starved

19

9

2

COD:P, 100:0.03

19

Table 6. 2. The micronutrients used in the synthetic feed used for the small reactors

Micronutrient Concentration (mg/L)

MgSO4.7H2O 5.07 (0.5 Mg)

CaCl2.2H2O 2.0 (0.5 Ca)

Na2MoO4.2H2O 0.01 (0.004 Mo)

Fe (III). EDTA 3 (0.5 Fe)

MnCl2.4H2O 0.36 (0.1 Mn)

FeSO4.7H2O 0.5 (0.1 Fe)

CuSO4.5H2O 0.39 (0.1 Cu)

ZnSO4.7H2O 0.44 (0.1 Zn)

CoCl2.6H2O 0.41 (0.1 Co)

110

C

E

B

Ch

ille

r at

12

˚C

Wate

r B

ath

at

22˚C

Feed

D

Feed

Feed

FeedDrain

Sk

yw

ay F

eed

DO Probe

pH Probe

Deca

nt

Drain

Was

teWaste

CC

EE

BB

Ch

ille

r at

12

˚C

Wate

r B

ath

at

22˚C

FeedFeed

DD

FeedFeed

FeedFeed

FeedFeedDrain

Sk

yw

ay F

eed

DO Probe

pH Probe

Deca

nt

Deca

nt

Drain

Was

teW

as

teWasteWaste

(a)

P Starved P Limited P Normal

Feed Mixture With a Starving

Dose of Phosphorous

Water Bath

at 22˚CAdditionalP

AdditionalP

Decant

Dra

in

Feed

P Starved P Limited P Normal

Feed Mixture With a Starving

Dose of Phosphorous

Water Bath

at 22˚CAdditionalP

AdditionalP

DecantDecant

Dra

in

FeedFeed

(b)

Figure 6. 1. Schematic of the large-20L (a), and small-2L (b) reactors

The mixed liquor samples were collected during waste phase, and secondary

effluent samples were collected during decanting phase.

111

6.3.2. Size Fractionation

To isolate flocs with a narrow size distribution, mixed liquor samples collected

during waste phase were fractionated using sieve trays (U.S.A. Standard Testing Sieve).

Collected floc fractions were gently washed with RO water and back washed into a flask.

The fractionated samples are identified by the name of reactor operating condition and the

floc size fraction, for example SRT20T22 (32-53) refers to the 32-53 micron size fraction

collected from the reactor running at an SRT of 20 days and a temperature of 22 ºC.

6.3.3. Floc Physicochemical Characteristics

Sludge Floc Structural Properties

Mixed liquor and secondary effluent particle size distributions were measured using

Rapid VUE (Beckman Coulter, Miami, USA). This equipment uses image analysis

techniques to determine particle characteristics in the range of 10-1000 µm. The value of

% flocs larger than 50 µm, % flocs larger than 30 µm, and the number of particles per unit

volume (in #/mL); were determined and are reported in Table 6. 4 & Table 6. 6. For flocs

with a size range of 35 to 120 µm, the average perimeter and longest length as well as the

% flocs with sphericity larger than 0.5, were measured using FlowCAM (Fluid Imaging

Technologies, Yarmouth, USA). The two-dimensional fractal dimension of flocs was

calculated based this data as follows (Logan & Kilps. 1995, Bellouti et al. 1997):

1DlP α [6-1]

where P is the floc perimeter, l is the longest floc dimension, size and D1 is the fractal

dimension. In general, flocs with greater fractal dimensions are more irregular in shape

and are expected to be more porous. It should be noted that the fractal dimension data

112

presented in Table 6. 4 was determined using the data obtained from the FlowCAM particle

analyzer and hence was limited to the 35 to 120 µm floc fraction.

Mechanical Stability of Flocs

In order to compare the mechanical strength of the flocs collected from each set of

reactors, the breakage of various floc samples were compared under identical

hydrodynamic shear stress using a Couette flow device according to the method described

elsewhere (Yuan & Farnood 2010, Yong et al. 2009, Gibson et al. 2009). For the large

reactors (i.e. reactors B, C, D, and E), suspensions of wastewater flocs with a size fraction

of 63-106 µm were prepared for strength measurements, while for the small reactors, the

45-75 µm floc size fraction was used. In both cases, the solid content of the suspensions

was kept constant at about 0.06 vol%. Ten milliliters of the floc sample was placed in the

gap between the spindle and the cup of the Couette flow device. Samples were exposed to

various shear stress levels by adjusting the spindle speed for duration of 10 min. The

breakage percentage was calculated according to:

% Breakage = ( 1 – 0P

Ps

N

N) × 100 [6-2]

where NP0 and NPs represent the initial and the final number of flocs larger than the mode

of the floc size distribution before shearing.

Sludge Floc Chemical Composition

The EPS was extracted from mixed liquor flocs according to the cation exchange

resin method developed by Frølund et al. (1996). For the large reactors (i.e. reactors B, C,

D, and E), suspensions of wastewater flocs with a size fraction of 63-106 µm were

prepared, while for the small reactors, the 45-75 µm floc size fraction was used. The

113

carbohydrate content of the extracts was measured using the phenol-sulfuric acid method

with D-glucose as standard (Masuko et al. 2005), and the protein content was measured

using the BCATM Protein Assay kit and a modified Lowry method (Haff 1978) using

bovine serum albumin (BSA) as standard. Spectrometric readings were measured using the

multiwell-plate reader ThermoU Spectra III A-5082 from SLT-Labinstruments (TECAN,

Durham, NC, United States).



used to irradiate the samples (Trojan Technologies, London, Ontario, Canada). The UV

transmittance of each sample was measured at 254 nm using a UV-1700 Pharmaspec UV-

Vis Shimadzu Spectrometer (Shimadzu Scientific Instruments, Maryland, USA). An IL-

1400A radiometer (International Light, Peabody, MA) was used to measure the UV light

intensity. UV transmittance of the sample and the UV intensity were used to estimate the

irradiation time required to achieve the desired UV dose according to method described by

Bolton and Linden (2003).

To obtain the UV DRC, 100 mL of each sample was placed in a plastic Petri dish

and exposed to the desired UV dose. The irradiated samples were then cultured using the

membrane filtration method following the Standard Methods for the Examination of Water

and Wastewater (# 9222 A, APHA 2001). m-FC agar (VWR, Mississauga, Ontario) was

used as the media to culture the surviving fecal coliform. The samples were incubated at

44 ºC for 24 ± 2 hours, and then the colony forming units (CFU) were enumerated. The

UV assay was performed on the effluent decanted from the SBR reactors as well as the

fractionated mixed liquor samples.

114

The UV DRC data were fit into the double exponential model as reported earlier

(Farnood 2005):

DkDkee

N

N21)1(

0

−− +−= ββ [6-3]

In this equation, D is the delivered ultraviolet dose in mJ/cm2, N is the number of

colonies formed from the survived coliform bacteria per 100 mL of irradiated sample, and

N0 is the number of colonies formed per 100 mL of the sample prior to the UV irradiation.

Parameters k1 and k2 are the UV inactivation rate constants; k1 is the inactivation constant

for UV susceptible particles (i.e. initial slope of the DRC), and k2 is the inactivation

constant for the UV resistant particles (i.e. the slope of tailing region). The parameter β is

the ratio of UV resistant particles to the total number of viable particles, and is indicative of

the susceptibility of the sample to UV irradiation (Tan. 2007).

6.3.5. Secondary Effluent Turbidity, UV Transmittance and Total Suspended

Solids

The effluent turbidity was measured using a 2100Q Portable Turbidimeter (Hach,

Mississauga, Ontario, Canada). The UV transmittance was measured at 254 nm using a

UV-1700 Pharmaspec UV-Vis Shimadzu Spectrometer (Shimadzu Scientific Instruments,

Maryland, USA). Total suspended solids were determined following the Standard Methods

for the Examination of Water and Wastewater (# 2540, APHA 2001).


The operational parameters for the SBR reactors are listed in Table 6. 3. In the

subsequent sections the characteristics of flocs collected from these reactors will be

discussed.

115

Table 6. 3. Performance parameters of the SBR system over the period of stable operating conditions.

MLSS: mixed liquor suspended solids, VSS: volatile suspended solids, ESS: effluent suspended solids,

SVI: sludge volume index.

Reactor

MLSS

(g/L)

VSS

(g/L)

ESS

(mg/L)

SVI

(mL/gr)

Influent

COD

(mg/L)

Effluent

COD

(mg/L)

Influent

NH4+

(mg/L)

Effluent

NH4+

(mg/L)

C

(SRT20T12) 2.5 ± 0.5 2.1 ± 0.4 9 ± 4 100 ± 30 191 ± 85 32 ± 16 22 ± 6 5.5 ± 5

D

(SRT7T12) 2.0 ± .04 1.6 ± 0.3 5 ± 3 91 ± 46 191 ± 85 38 ± 22 23 ± 6 15 ± 6

B

(SRT20T22) 2.0 ± 0.6 1.7 ± 0.6 4 ± 3 83 ± 18 191 ± 85 28 ± 12 24 ± 6 5 ± 5

E

(SRT7T22) 1.5 ± 0.4 1.3 ± 0.3 3 ± 2 98 ± 30 191 ± 85 34 ± 16 25 ± 6 3 ± 3

P-Normal 3.2 ± 0.6 3.1 ± 0.6 24 ± 10 93 ± 25 500 22 ± 9 28 20 ± 6

P-Limited 3.8 ± 0.5 3.7 ± 0.6 10 ± 10 103 ± 23 500 20 ± 6 28 18 ± 9

P-Starved 1.7 ± 0.5 1.5 ± 0.6 17 ± 12 128 ± 20 500 35 ± 19 28 19 ± 8

6.4.1. Floc Physicochemical Properties

Size and Sphericity

The morphological characteristics of the flocs generated in the SBR reactors are

listed in Table 6. 4. The results suggest that flocs formed at the lower operating

temperature of 12 ºC were smaller and more spherical compared to those generated at 22 ºC,

at both 7 days and 20 days SRT. Reactors operating at 22 ºC produced larger flocs with

higher fractal dimension and improved settleability characteristics; i.e. lower ESS and SVI

as seen in Table 6. 3. By increasing SRT from 7 days to 20 days at both 12 ºC and 22 º) the

average mixed liquor floc size increased.

116

In general, limiting or starving the SBR reactors on phosphorus caused the

formation of smaller flocs. Moreover, under the P-Starved conditions flocs were notably

less spherical and more irregular in shape (higher fractal dimension).

Table 6. 4. The effect of increasing SRT, T and changing feed phosphorus level on floc size, sphericity,

fractal dimension (D1 using Logan & Wilkinson 1991), protein to carbohydrate ratio, and mechanical

stability (τ 50)

Reactor

% Flocs with

Sphericity > 0.5

(35-120 µm)

% ML Flocs

> 50 µm

D1

(35-120 µm)

ML Protein:

Carbohydrate

τ 50*

SRT 20 T 12 20 28 1.25 2.5 ± 0.04 9.0 ± 0.4

SRT 7 T 12 15 17 1.3 1.8 ± 0.02 9.4 ± 0.3

SRT 20 T 22 8 22 1.4 1.9 ± 0.01 9.9 ± 0.5

SRT 7 T 22 11 23 1.5 1.5 ± 0.02 10.0 ± 0.6

P-Normal 12 27 1.36 3.2 ± 0.04 11.7 ± 0.5

P-Limited 14 20 1.3 1.0 ± 0.03 14.6 ± 0.3

P-Starved 5.5 12 1.4 3.7 ± 0.01 9.4 ± 0.4

*: Calculated for the 63-106 µm floc fraction in the large reactor and 45-75 µm floc fractions in the

small reactors.

Mechanical Stability

The breakage curves of flocs as a function of shear stress for the SBR reactors are

plotted in Figure 6. 2. The median shear stress (τ50

), i.e. the shear stress at which the

number of flocs reduced by 50 %, is reported for flocs of narrow size fractions in Table 6. 4.

Mechanically stronger flocs would have a higher τ50

, as a higher shear stress would be

required to break the flocs. Although there is significant variability in the data, increasing

the operating temperature from 12 to 22 oC caused a decrease in floc breakage, i.e.

produced somewhat stronger flocs. No clear relationship was found between SRT and the

flocs’ mechanical stability. Phosphorus limitation on the other hand, had a noticeable

effect on floc mechanical stability (Figure 6. 2-b). At similar shear stress levels, P-Limited

flocs showed a lower breakage tendency (i.e. higherτ50

), indicating that these flocs were

117

mechanically stronger. This is consistent with the previous studies that reported flocs

under phosphorus limitation condition tend to have a more compact and spherical structure,

with more densely populated biomass and better settleability characteristics (Liu & Liss

2007, Dumitrache 2008).

0

20

40

60

80

100

6.7 8.6 10.6 12.6

Turbulent Shear Stress (Pa)

%B

rea

ka

ge

C SRT20T12 (63-106 µm)

D SRT7T12 (63-106 µm)

B SRT20T22 (63-106 µm)

E SRT7T22 (63-106 µm)

(a)

0

20

40

60

80

6.7 8.6 10.6 12.6

Turbulent Shear Stress (Pa)

%B

reak

ag

e

P-Normal (45-75 µm)

P-Limted (45-75 µm)

P-Starved (45-75 µm)

(b)

Figure 6. 2. Percentage of breakage obtained at various degrees of shearing, for (a) the large (20 L)

reactors, and (b) for the small (2 L) reactors

118

Composition

The protein to carbohydrate ratio of the EPS samples extracted from the SBR

reactors are shown in Table 6. 4. The P-Limited flocs, which had significantly higher

mechanical strength, also showed a lower protein to carbohydrate ratio compared to the P-

Normal and P-Starved conditions. At both SRTs of 7 & 20 days, increasing the operating

temperature from 12 to 22 ºC decreased the protein to carbohydrate ratio and increased the

mechanical strength. The calculated protein to carbohydrate ratio is negatively correlated to

mechanical strength.

6.4.2. UV Disinfection of Mixed Liquor Flocs

As discussed in pervious chapters, the tailing of the UV DRC is controlled by the

presence of larger flocs in the effluent. Hence, it is expected that the physicochemical

characteristics of the flocs affect the level of tailing and the resistance of the effluent to UV

disinfection. For this reason, it is important to asses the UV disinfectability of isolated

flocs in relation to the wastewater treatment conditions.

The Effects of Temperature and Sludge Retention Time

The UV DRC of 32-53 µm and 75-90 µm flocs, collected from the large SBRs are

displayed in Figure 6. 3. The double exponential model parameters for these samples are

also provided in

Table 6. 5. For the 32-53 µm flocs (Figure 6. 3- a), regardless of variations in SRT

and temperature, the kinetics of UV disinfection were nearly the same. However, in the

case of 75-90 µm flocs (Figure 6. 3-b), increasing the operating temperature from 12 to 22

119

ºC at both SRTs resulted in a lower tailing level of the UV DRC by 0.5 log, while SRT had

no significant effect on the tailing level.

Higher operating temperatures are known to favor the growth of filamentous

bacteria over the floc-forming bacteria in flocs (Eckenfelder. 2000, Rossetti et al. 2002).

The presence of filamentous bacteria is known to give the flocs a more open structure (Liu.

2006, Tian & Su. 2012). Liu et al. (2005) showed that the abundance of filamentous

bacteria increased when increasing the operating temperature from 17 ºC to 25 ºC.

Furthermore, filamentous networks direct floc growth into shapes other than spherical

(Jenkins et al. 2004). The morphological data presented in Table 6. 4 also show that flocs

formed at higher temperatures were less spherical. Higher irregularity in floc surface (or

higher fractal dimension) is generally associated with higher porosity (Perez et al. 2006),

and would allow for better penetration of UV light. Elongated filamentous bacteria (that

could be hundreds of microns in size) are more likely to be embedded in larger flocs. This

could be the reason that temperature showed a noticeable effect on the UV disinfection of

the larger floc fraction (75-90 µm) but had no influence on the tailing level of the 32-53 µm

floc fraction.

120

0.0001

0.001

0.01

0.1

1

0 10 20 30 40 50 60

UV Dose (mJ/cm²)

N /

N 0

SRT 20 T 12 (32-53 µm)

SRT 7 T 12 (32-53 µm)

SRT 20 T 22 (32-53 µm)

SRT 7 T 22 (32-53 µm)

(a)

0.001

0.01

0.1

1

0 10 20 30 40 50 60

UV Dose (mJ/cm²)

N /

N 0

SRT 20 T 12 (75-90 µm)

SRT 7 T 12 (75-90 µm)

SRT 20 T 22 (75-90 µm)

SRT 7 T 22 (75-90 µm)

(b)

Figure 6. 3. The averaged UV DRC of (a) 32-53 µm and (b) 75-90 µm activated sludge flocs collected

from the SRT20T12, SRT7T12, SRT7T22, and SRT20T22 reactors. The lines represent the fitted

double exponential model.

The Effect of Influent Phosphorus Content

The averaged double exponential UV DRCs of 45-75 µm mixed liquor flocs,

collected from the small SBRs are displayed in Figure 6. 4. The flocs isolated from the P-

Limited reactor showed significantly higher resistance to UV disinfection. However, the P-

121

Starved flocs were easier to disinfect and showed a lower tailing level. The total

phosphorus in the mixed liquor (TP/MLSS) was estimated for the P-Normal, P-Limited,

and P-Starved reactors were 1.3, 0.34, and 0.02 mg/g, respectively. At lower feed

phosphorus content, there was less accumulation of phosphorus in the mixed liquor.

P-Starved flocs were more irregular in shape with poor settling characteristics

compared to the P-Limited and P-Normal flocs (Table 6. 4 &Table 6. 3). The UV DRC

also showed a lower tailing level by 1 log and 0.5 log compared to the P-Limited and P-

Normal flocs, respectively (Figure 6. 4). The lower tailing level of the P-Starved flocs

suggests that these flocs were less likely to contain viable coliform within their structure.

In fact it has been reported that coliform bacteria were less readily present under

phosphorus deficiency (An et al. 2002, Davies et al. 1995). Therefore, a lower percentage

of flocs grown under these conditions would harbor and shield coliform bacteria, and hence

they are less likely to be UV resistant (i.e. have lower tailing level).

0.00001

0.0001

0.001

0.01

0.1

1

0 10 20 30 40 50 60

UV Dose (mJ/cm²)

N /

N 0

P-Normal 45-75

P-Limited 45-75

P-Starved 45-75

Figure 6. 4. The averaged UV DRC of the 45-75 µm activated sludge flocs collected from the P-Normal,

P-Limited, and P-Starved reactors. The lines represent the fitted double exponential model

122

The UV DRC double exponential kinetic parameters are listed in Table 6. 5.

Higher values of β and lower values of k1 correspond to higher UV resistance. For a given

SBR system, the UV resistance (i.e. β) of flocs was positively correlated with the

percentage of flocs with a sphericity > 0.5. This correlation was particularly strong in the

case of larger reactors (r2= 0.99). Sphericity affects the path length of UV light within the

floc and hence could alter the average dose delivered to the embedded microorganisms

within the floc. More elongated (less spherical) flocs could provide a larger average dose

and hence would be less UV resistant. It appears that higher UV resistance occurred when

the percentage of particles with a sphericity > 0.5 increased for all of the SBR floc samples

(Figure 6. 5).

Table 6. 5. Double exponential UV disinfection kinetic parameters for the isolated mixed liquor flocs

from the large and small SBRs

Large Reactors (75-90 µm) k1 k2 β

SRT 20 T 12 0.14 ± 0.02 0.01 ± 0.005 0.21 ± 0.03

SRT 20 T 22 0.16 ± 0.04 0.01 ± 0.006 0.12 ± 0.03

SRT 7 T 12 0.15 ± 0.02 0.009 ± 0.005 0.18 ± 0.04

SRT 7 T 22 0.16 ± 0.03 0.012 ± 0.007 0.15 ± 0.02

Small Reactors (45-75 µm)

P-Normal 0.2 ± 0.05 0.02 ± 0.006 0.09 ± 0.02

P-Limited 0.12 ± 0.01 0.01 ± 0.006 0.06 ± 0.03

P-Starved 0.26 ± 0.01 0.02 ± 0.003 0.05 ± 0.07

123

0

0.05

0.1

0.15

0.2

0.25

0.3

8 11 15 20 5.5 12 14

% Flocs with Sphericity > 0.5

β o

r k

1

β

k1

Large Reactors Small Reactors

Figure 6. 5. UV DRC double exponential model parameters (β &k1) for mixed liquor flocs with

different percentages of flocs with sphericity larger than 0.5. Please note that the double exponential

parameters are reported for the 75-90 µm, and 45-75 µm flocs for the large and small reactors,

respectively. The percentage of flocs with sphericity larger than 0.5 is estimated for 35-120 µm flocs.

6.4.3. Final Effluent Quality and UV Disinfection

The Effects of Temperature and Sludge Retention Time

The final effluent quality parameters from all seven reactors are listed in Table 6. 6.

For the large reactors, operating at the higher temperature (22˚C) improved the overall

effluent quality (i.e. lower ESS, lower turbidity, and lower percentage of flocs larger than

50µm). There was no clear relationship between SRT and the effluent quality parameters.

For the small reactors, the P-Limited condition showed improved effluent characteristics

(i.e. lower turbidity, lower TSS and higher UVT) as well as improved settleability

compared to the P-Normal and P-Starved conditions. However, the quality of final effluent

worsened significantly in the P-Starved condition. It should be noted here that the

percentage of particles larger than 50 µm plays a significant role in the UV disinfectability

of the final effluent, as larger particles are harder to disinfect (Tan 2007, Yong et al. 2009,

124

Gibson et al. 2008). Effluents with a significantly lower percentage of particles larger than

50 µm included the P-Starved and the large reactors operating at 22 ºC. The P-Starved

effluent showed a higher turbidity and lower UVT compared to the P-Limited and P-

Normal conditions.

Table 6. 6. Effluent quality data for all seven pilot reactors

Reactor

Turbidity

(NTU)

ESS

(mg/L)

UVT% at

254nm

#Particle/m

L

% of

particles

> 30µm

% of

particle

s >

50µm

C (SRT20T12) 2.8 ± 1.3 9 ± 4 67 ± 3 108 ± 10 54 9

D (SRT7T12) 3.9 ± 0.2 5 ± 3 71 ± 3 102 ± 10 54 9

B (SRT20T22) 1.8 ± 0.2 4 ± 3 72 ±3 82 ± 10 50 4

E (SRT7T22) 1.4 ± 0.6 3 ± 2 75 ± 2 73 ± 10 57 5

P-normal 3.7 ± 0.3 24 ± 10 73 ± 4 110 64 17

P-limited 2.3 ± 0.3 10 ± 10 84 ± 3 87 58 13

P-starved 6.9 ± 0.6 17 ± 12 64 ± 4 100 53 6

The average UV DRCs of the secondary effluents collected from all reactors are

shown in Figure 6. 6 & Figure 6. 7. These DRCs are displayed both based on the

concentration of colony forming units (N) in CFU / 100 mL as well as the normalized

(N/N0) scale.

Based on Figure 6. 6, operating at higher temperature (22 ºC) improved the UV

disinfection performance of final effluent (i.e. lower CFU / 100mL at each UV dose).

Table 6. 6 shows that the secondary effluents collected from the 22 ºC reactors contained a

significantly lower percentage of particles smaller than 50 µm. According to Figure 6. 3-b,

larger flocs at higher operating temperatures were found to be more susceptible to UV

irradiation. Therefore, it is expected that the secondary effluents collected from the

reactors operating at 22 ºC would show an improved disinfectability. It should be

125

emphasized that the above effects are limited to temperatures examined in this study (i.e.

12 ºC and 22 ºC).

10

100

1000

10000

100000

1000000

0 20 40 60

UV Dose (mJ/cm²)

N

(CF

U /

10

0 m

L) B SRT 20 T 22

E SRT 7 T 22

10

100

1000

10000

100000

1000000

0 20 40 60

UV Dose (mJ/cm²)

N

(CF

U /

10

0 m

L) C SRT 20 T 12

D SRT 7 T 12

10

100

1000

10000

100000

1000000

0 20 40 60

UV Dose (mJ/cm²)

N (C

FU

/ 1

00

mL

) C SRT 20 T 12

B SRT 20 T 22

10

100

1000

10000

100000

1000000

0 20 40 60

UV Dose (mJ/cm²)

N

(CF

U /

10

0 m

L) D SRT 7 T 12

E SRT 7 T 22

a - the effect of SRT at high T b - the effect of SRT at low T

c - the effect of T at high SRT d - the effect of T at low SRT

10

100

1000

10000

100000

1000000

0 20 40 60

UV Dose (mJ/cm²)

N

(CF

U /

10

0 m

L) B SRT 20 T 22

E SRT 7 T 22

10

100

1000

10000

100000

1000000

0 20 40 60

UV Dose (mJ/cm²)

N

(CF

U /

10

0 m

L) C SRT 20 T 12

D SRT 7 T 12

10

100

1000

10000

100000

1000000

0 20 40 60

UV Dose (mJ/cm²)

N (C

FU

/ 1

00

mL

) C SRT 20 T 12

B SRT 20 T 22

10

100

1000

10000

100000

1000000

0 20 40 60

UV Dose (mJ/cm²)

N

(CF

U /

10

0 m

L) D SRT 7 T 12

E SRT 7 T 22

a - the effect of SRT at high T b - the effect of SRT at low T

c - the effect of T at high SRT d - the effect of T at low SRT

Figure 6. 6. The UV DRCs of treated effluents based on actual CFU / 100 mL.

The Effect of Influent Phosphorus Content

The UV DRCs of final effluents collected from P-Normal, P-Limited, and P-

Starved conditions are given in Figure 6. 7- a. This figure shows the P-Starved effluent

reached about 10 CFU/ 100 mL at a dose of 15 (mJ/cm2) whereas the P-limited and P-

normal effluents reached the same target at a UV dose of about 30 (mJ/cm2). At higher

126

doses, however, there was no significant difference among the UV DRC of the effluents

collected from the small reactors. In contrast, in their normalized form (Figure 6. 7-b), UV

DRC of the P-Starved effluent had a notably lower tailing level (about 1 log). The lower

tailing level of P-Starved effluent was likely a combination of two factors: 1) the P-starved

effluent contained a lower percentage of larger flocs (Table 6. 6), which are known to be

harder to disinfect, and 2) the P-Starved flocs were more susceptible to UV disinfection

compared to the P-Normal and P-Limited flocs as discusses in the previous section. In

contrast, the final effluent of the P-Limited reactor showed a higher tailing level implying

that although the effluent had a lower number of flocs, the flocs were harder to disinfect (in

accordance with the results in Figure 6. 3).

This information may be applied in the selection of upstream processes. For

example, it is known that the P-limited condition produces fewer flocs in the final effluent,

but with much higher UV resistance compared to the P-Starved and P-Normal conditions.

If filtration is the subject upstream process of interest, it would be more effective to apply

P-Limited conditions due to the presence of fewer particles in the effluent. Whereas, if

sonication or hydrodynamic shearing is the upstream process of interest, a process should

be selected that produces flocs with lower mechanical stability (P-Starved). However, in

the selection of upstream processes, the changes in UV dose demand should also be

accounted for.

127

0.1

10

1000

100000

10000000

0 20 40 60

Dose (mJ/cm²)

N (

CF

U /

10

0 m

L)

P-Normal

P-Limited

P-Starved

(a)

0.000001

0.0001

0.01

1

0 20 40 60

Dose (mJ/cm²)

N /

N 0

P-Normal

P-Limited

P-Starved

(b)

Figure 6. 7. UV DRCs (actual CFU/100ml-a, and normalized-b) of final effluents collected from the P-

Normal, P-Limited, and P-Starved conditions

6.4.4. UV Dose Demand of Final Effluent

In all of the UV DRCs presented so far in this study, the UV dose displayed on the

x axis was calculated based on the UV transmittance, and hence the effect of effluent

quality (turbidity, UVT) on UV dose was eliminated. However, for design purposes, it is

128

necessary to determine the total dose demand (i.e. energy demand) that includes the effect

of effluent quality parameters. To observe the effect of effluent quality in the UV DRC,

the fecal coliform survival ratio must be plotted against UV energy delivered to the sample.

The effluents’ UV dose demand for the large and small SBR reactors are displayed in

Figure 6. 8. UV dose demand in this section is defined as the time for achieving 3 log

reduction for the large reactors and 4 log reduction for the small reactors. The results

showed that for 3 log reduction of the large reactors’ effluents, SRT20T22 condition

needed almost half the UV dose than that of the reactors operating at lower temperature, or

lower SRT. The P-Starved condition effluent also showed a significantly lower (less than

half) UV dose demand to reach 4 log reduction compared to the P-limited and P-Normal

conditions.

In Figure 6. 8, the P-Starved effluent had lower effluent quality parameters

compared to the other small reactors, i.e. lower UVT and higher turbidity. However, the P-

Starved effluent had a much lower UV dose demand due to much higher susceptibility of

the flocs. This represents a case in which the conventionally measured effluent quality

parameters do not correlate with the UV resistance of the effluent. These results further

confirm that operating an activated sludge system at 22 ºC compared to 12 ºC, and P-

Starved compared to P-Limited and P-Normal produced effluents with significantly lower

UV dose demand.

129

10

20

30

40

50

SRT20

T12

SRT7

T12

SRT20

T22

SRT7

T22

P-

Normal

P-

Limited

P-

Starved

UV

Do

se D

em

an

d (

mJ /

cm

²)

3 log Reduction 4 log Reduction

Figure 6. 8. UV dose demand of the final effluents collected from all SBRs.

6.5. Conclusions

In this study the effects of SRT, temperature, and influent phosphorus levels on UV

disinfection kinetics were investigated. All reactors operated aerobically, and the results

showed that in general, more irregularly shaped flocs with lower sphericity and higher

fractal dimension were susceptible to UV disinfection. Increasing the operating temperature

from 12 ºC to 22 ºC, at both 7 & 20 days SRT, induced the formation of less spherical flocs

with lower resistance to UV disinfection (0.5 log lower tailing for the 75-90 µm flocs).

The 22 ºC final effluent showed 40% reduction in the number of particles larger than 50

µm compared to that of the 12 ºC reactors. Varying the SRT did not significantly change

the UV dose response of the sludge flocs and secondary effluents. Mechanical strength was

negatively correlated to UV resistance when varying temperature and SRT, but no

130

correlation was found when changing the influent phosphorus levels. P-Limited condition

(COD:N:P of 100:10:0.3) causes the formation of more spherical flocs with better

settleability and higher UV resistance compared to the P-Normal and P-Starved conditions.

However, P-Starved condition promotes the formation of small and irregularly shaped flocs

with poor settling characteristics. Despite having a lower effluent quality (high ESS and

high turbidity), the P-Starved condition showed improved UV disinfection performance, i.e.

less than half the UV dose required for 4 log removal of fecal coliform, compared to the P-

Limited and P-Normal effluents.

6.6. References

American Public Health Association (APHA) (2001), Standard methods for the examination of water and wastewater. 22nd Ed. Washington DC, USA. An, Youn-Joo, Donald H. Kampbell, and G. Peter Breidenbach. (2002). Escherichia coli and total coliforms in water and sediments at lake marinas. Environmental Pollution 120, 771-778. Bolton, J. R., Linden, K. G. (2003). Standardization of methods for fluence (UV Dose) determination in bench-scale UV experiments. Journal of Environmental Engineering, 129 (3), pp. 209 – 215. Bellouti, M., Alves, M. M., Novals, J. M., & Mota, M. (1997). Flocs vs granules: Differentiation by fractal dimension. Water Research, 31(5), 1227-1231. Brei, E. (2011). Bacterial Adhesin Proteins Associated with Microbial Flocs and EPS in Activated Sludge. Thesis (Ph.D)- University of Toronto Das, T. K. (2001). Ultraviolet disinfection application to a wastewater treatment plant. Clean Products and Processes, 3(2, pp. 69-80), August. Davies, Cheryl M., Julian A. H. Long, Margaret Donald, and Nicholas J. Ashbolt. 1995. Survival of fecal microorganisms in marine and freshwater sediments. Applied and Environmental Microbiology 61, 1888-1896. Dumitrache, R. G. (2008). Characterization of a hybrid sequencing batch reactor system for treatment of wastewater using suspended microbial aggregates and biofilms. Thesis (M.A. Sc.)- Ryerson University Eckenfelder WW. Industrial water pollution control. Singapore: McGraw-Hill; 2000. Farnood, R., Flocs and ultraviolet disinfection. Chapter 18 in Flocculation in Natural and Engineering

Systems. I.G. Droppo, G.G. Leppard, S.N. Liss, and T.G. Milligan, Eds., CRC Press, Boca Raton, FL, 2005, pp. 385-395.

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133

7. Overall Discussion & Conclusions

This research provides a comprehensive understanding on the effects of wastewater

flocs’ structure and composition on UV disinfection kinetics. The objectives of this work

were to investigate the causes of tailing, and to evaluate the effect of secondary biological

treatment conditions on flocs physicochemical characteristics and UV disinfection kinetics.

Figure 7. 1 shows a summary of the project plan.

The three significant contributions of this study are listed below:

(1) The formation of compact cores in wastewater flocs significantly increases their

resistance to UV disinfection. A conceptual model for UV disinfection of flocs was

proposed that takes into account the floc structure;

(2) The flocs’ chemical composition could significantly affect their UV disinfection

kinetics. Finding of this study suggest that the presence of naturally occurring

polyphosphates in flocs collected from biological removal processes was likely inducing

additional UV disinfection of this type of effluent. This effect was attributed to the

photoreactive release of oxidative species that could results in cell lysis and accelerate UV

disinfection. On this basis, an invention disclosure was filed with the University of

Toronto (Disclosure # 10002579, Confidential Intellectual Property Center at the

University of Toronto);

(3) The effect of activated sludge process operational temperature, sludge retention

time, and phosphorus levels on floc structure and UV disinfection was investigated.

These topics are discussed in more detail in the following sections.

134

Figure 7. 1. Summary of thesis project plan

7.1. Identification of Compact Cores as UV Resistant Constituents

This study showed that activated sludge flocs may contain mechanically strong and

compact cores that exhibit significantly higher resistance (i.e. higher tailing level in DRC)

to UV disinfection. The volume and number of cores and their viability increased as floc

size increased. Therefore larger flocs were more likely to be resistance to UV (larger β)

and showed higher tailing level in the UV DRCs, as indicated in Chapter 3. The worsening

effect that the compactness of microbial aggregates has on UV disinfection and increasing

the UV dose demand (measured by comparing DRCs of flocs and cores of similar size), is

comparable to the worsening effect of increasing the average floc size by 20 µm.

135

Based on a novel structural model for the UV disinfection of floc, it was found that

floc disinfection kinetics was sensitive to the relative volume, the density (UV absorbance),

and the viability of dense core. However, the exact mechanisms and conditions that control

the formation of such cores are not widely understood, and additional work is required.

Gaining such fundamental understanding would be an important step towards designing

more effective wastewater treatment processes.

7.2. The Potential of Polyphosphates as Photooxidative Agents in

Sludge

This study showed that flocs harvested from BNR-UCT process exhibited improved

UV disinfection kinetics compared to the conventional activated sludge. It was suggested

that the origin of such differences was due to the chemical composition and structure of

flocs under BNR conditions. This finding suggests that through the holistic design of

wastewater treatment facilities it is possible to manipulate the physicochemical properties

of activated sludge flocs and improve their UV disinfection kinetics. Manipulating the

physicochemical properties of sludge flocs by integrating enhanced biological nutrient

removal processes improves the UV disinfection performance. The main contribution was

reporting evidence of in-situ advanced oxidation when disinfecting biological phosphorus

removal sludge with UV light, due to the photoreactive properties of polyphosphates in

releasing oxidative species under UV exposure.

The causes of higher UV susceptibility of the BNR-UCT flocs were proposed to be:

(1) the release of oxidative hydroxyl radicals within the flocs due to the photoreactive

effect of polyphosphates (embedded in BNR flocs), which can cause cell lysis; (2) the

release of micro-bubbles within the BNR-UCT flocs in the anaerobic and anoxic regions

136

gives them a more open and homogeneous structure, and limits the formation of compact

EPS zones, which further protect fecal coliform from exposure to UV.

Mathematical modeling showed that the UV disinfectability of flocs, as expected,

was quite sensitive to the generation of •OH within flocs. Formation of hydroxyl radicals

could improve the disinfection rate and significantly lower the tailing level of UV DRC.

7.3. Reducing Wastewater Flocs’ UV Resistance by Changing the

Secondary Treatment Operational Conditions

This study showed that the UV dose demand of final effluent, which is related to

the effluent quality and sludge flocs’ UV susceptibility, can be significantly altered by

changing the operational conditions of the activated sludge system. A strong correlation

was found between sphericity and UV resistance. As sphericity decreases, flocs are more

susceptible to UV disinfection. This finding was consistent among all the samples tested

over the course of this study. Compact cores had higher resistance to UV and were more

spherical compared to flocs; the BNR-UCT flocs were less spherical and had lower UV

resistance compared to CAS flocs; the flocs formed at 22 ºC were less spherical and more

susceptible to UV compared to those formed at 12 ºC; and the P-Staved flocs were less

spherical and more UV susceptible compared to the P-Limited and P-Normal flocs.

Moreover, it was found that compared to UVT, the percentage of flocs larger than

50 µm in the effluent correlated better to the UV disinfection performance. Lower UV

dose demand was estimated for two effluents, one with a lower turbidity (22 ºC) and the

other with higher turbidity (P-Starved). However, both effluents with low UV dose

137

demand had a lower percentage of flocs larger than 50 µm compared to the other

conditions.

Mechanical strength did not show a consistent trend with UV disinfection. The

cores were mechanically stronger than the flocs and showed higher UV resistance

compared to flocs; the BNR-UCT flocs were mechanically stronger with higher UV

susceptibility; the flocs produced at 22 ºC appeared to be mechanically stronger and

showed a lower UV resistance; and the P-Starved flocs which were mechanically less

stable showed high susceptibility to UV disinfection. Therefore, no conclusions can be

drawn regarding the effect of mechanical strength on UV resistance. However, the protein

to carbohydrate ratio has a strong negative correlation with mechanical strength. In Chapter

3, it was found that the mechanical strength of the compact cores was higher than that of

the flocs, and it was found that they contained a lower protein to carbohydrate ratio (Wang

2012). Moreover, the BNR-UCT flocs that were mechanically more stable had a lower

protein to carbohydrate ratio (see Chapter 4) than the CAS flocs. In Chapter 6, the flocs

formed at 22 ºC or at the P-Starved condition were stronger with lower protein to

carbohydrate ratio.

The addition of upstream processes that cause an increase in floc porosity or break

larger flocs would increase UV disinfection efficiency. Breaking the flocs by sonication or

hydrodynamic shearing prior to UV disinfection are suggested methods to improve the

flocs disinfectability. Floc breakage causes a size distribution shift towards smaller flocs,

while increasing the number of flocs. This increases effluent turbidity and decreases UV

transmittance, which adversely affect UV disinfection performance. Therefore, it is

suggested that processes that reduce sphericity and increase the surface roughness of the

138

microbial aggregates without breaking them would be more effective in improving their

UV disinfectability.

7.4. Conclusions

Considering the conducted research in this thesis, the following conclusions can be

made:

1- The formation of compact cores within the flocs contributes significantly to

their UV resistance. The compact cores were isolated from the flocs by shearing and

sieving, which peels off the loose outer layer. When the UV DRCs of flocs and cores of

similar size were compared, the tailing level was 0.5 log higher for the cores.

2- Enhanced biological nitrogen/phosphorus removal increases the UV

susceptibility of the sludge flocs. Denitrification was suggested to give the flocs a more

porous structure due to the internal release of nitrogen bubbles, hence increasing UV

transmittance within the flocs. Moreover, the accumulation of polyphosphate as a result of

EBPR causes significant changes in the mechanism of UV disinfection. In addition to

direct DNA damage from UV light, the photoreactive effect of polyphosphate causes the

release of •OH, which ruptures cell membrane and causes death. The UV DRCs showed

that at a UV dose of 15 mJ/cm2 the BNR-UCT final effluent shows 4 log disinfection,

while the CAS final effluent showed 2 log disinfection. At the UV dose of 60 mJ/cm2 the

BNR-UCT final effluent shows a lower tailing level than the CAS (1 log).

3- Polyphosphates, which occur naturally in sludge during the enhanced

biological phosphorus removal (EBPR) process, induce in-situ advanced oxidation

when exposed to UV light. For the first time it has been shown in this study that

139

photoactive disinfection occurs within the flocs, not by adding a photocatalyst, but merely

by changing the type of wastewater treatment process and the accumulation of

polyphosphates in the sludge.

4- Floc sphericity is directly correlated to UV resistance. It was consistently

reported in this study that where flocs are less spherical, the embedded fecal coliform had a

lower chance of surviving UV disinfection.

5- The UV dose demand is dependent on secondary process conditions. Among

the conditions studied, two of them significantly reduced the UV dose demand of

secondary effluents: 1) the BNR-UCT process compared to conventional activated sludge

systems; and 2) conventional activated sludge system operating at COD:N:P of 100:10:0.03

(i.e. P-Starved condition) compared to 100:10:1 (i.e. P-Normal condition).

6- P-Limited conditions induce the formation of rounder and more compact

flocs with higher UV resistance, and higher mechanical stability. The P-Limited

condition appears to also improve the effluent characteristics such as decreasing turbidity,

decreasing ESS, and increasing the UVT. However, the flocs were more compact and

showed higher resistance to UV disinfection.

7- P-Starved conditions cause the formation of irregular flocs with high UV

susceptibility. The 45-75 µm mixed liquor flocs generated under P-Starved conditions

show improved UV disinfectability compared to the P-Limited (2 log improvement) and

the P-Normal (1 log improvement) flocs of similar size. The UV Dose demand of the P-

Starved effluent was also significantly lower than that of the P-Limited and P-Normal

conditions.

140

8- Operating a conventional activated sludge system at 22 ºC, compared to 12

ºC, decreases floc sphericity and the UV resistance of the final effluent. Flocs formed

at 22 ºC are more irregular in shape, and show improved settling properties compared to

those formed at 12 ºC (at both SRTs of 7 and 20 days). The final effluent collected from

the 22 ºC reactor had lower ESS, lower turbidity, higher UVT, and a lower percentage of

particles larger than 50 µm compared to the one collected from the 12 ºC reactor.

Therefore, UV disinfection improves when operating conventional activated sludge

systems at higher temperature (comparing 12 & 22 ºC). It should be noted here that the

above effects are limited to temperatures examined in this study (i.e. 12 ºC and 22 ºC). It is

known that higher operating temperatures (e.g. above 35 ºC) result in deflocculation, which

decreases secondary effluent quality. Therefore, the UV disinfection kinetics of flocs

formed at temperatures higher than 22 ºC remains unknown.

9- For the final effluents, the percentage of particles larger than 50% seems to

be a more relevant parameter to UV resistance compared to UVT and turbidity. In all

of the conditions studied, it appears that the effluent’s percentage of particles larger than 50

µm is more consistently related parameter to UV resistance. From the two operating

conditions that had the lowest UV dose demand, one had a comparatively low turbidity and

high UVT and the other had a higher turbidity and lower UVT.

10- Changing the sludge retention time from 7 to 20 days does not significantly

affect the UV disinfectability of the mixed liquor flocs and final effluents. Increasing

the sludge retention time slightly increases the mixed liquor size distribution, but no impact

on UV disinfectability was observed.

141

7.5. Recommendations for Future Work

1- In this study, the enhanced biological nutrient removal process was found to

generate flocs with higher susceptibility to UV disinfection. It was suggested that a

combination of polyphosphate’s photoreactivity and the increase in floc porosity by the

release of nitrogen gas was the origin of higher UV susceptibility. Polyphosphates are

produced in the enhanced biological phosphorus removal process, and the release of

nitrogen gas occurs during the denitrification (part of the enhanced biological nitrogen

removal) process. However, the effect of each of the enhanced biological

nitrogen/phosphorus removal on floc properties and UV disinfection remains unknown.

Assessing the effects of each process on its own would clarify the contribution of porosity

increase, and the presence of polyphosphates on UV disinfectability.

2- Studying the effect of mixing a pure culture of polyphosphate accumulating

organisms such as Accumulibacter Phosphatis on the UV disinfection kinetics of non-

particle associated fecal coliform. In this study it was shown that polyphosphate

accumulating organisms have a significant effect on UV disinfection mechanism, due to the

photoreactive effect of polyphosphates and the release of oxidative species. Moreover, it

would be helpful to systematically study the effect of polyphosphate accumulation

(different amounts of accumulated polyphosphate) on the UV disinfectability of e.coli.

3- Assessing the photoreactivity of polyhydroxylalkenoates (PHAs). In this study

the photoreactivity of polyphosphate was assessed. However, the BNR-UCT sludge also

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contains PHA in large quantities and they may also be contributing to the accelerated

disinfection in the BNR-UCT flocs. More specifically, the photoreactivity of

polyhydroxybutyrate (PHB) should be studied.

4- Assessing the effect of nitrogen limitation on the amount of extracellular

polymeric substances and proteins, effluent quality and UV disinfectability. In this study

the effect of phosphorus limitation on some floc physicochemical characteristics was found,

and the UV disinfectability of mixed liquor flocs and final effluents were compared for

various degrees of phosphorus limitation.

5- Understanding the effect of feeding pattern on the flocs structural properties,

viability and UV disinfection kinetics. Throughout this study, although not reported,

observations were made that feeding pattern can significantly impact the mechanical

strength and viability-culturability of the microbial flocs. It is suspected that the feeding

pattern would change the response of microbial aggregates to UV disinfection.

6- In this study, composition was a general protein to carbohydrate ratio

measurement. However, further characterization of the type of sugars and proteins in the

flocs, their UV absorbance and contribution to floc UV resistance is needed.

7- UV disinfection performance in the industry and in this research is assessed

using the membrane filtration method. In this method, if a floc is cultured and produces

one colony, this is independent of whether the floc has one or a thousand viable-culturable

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microorganisms. It would be of value to develop methods to assess the total number of

fecal coliform in the flocs.

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8. Appendices

8.1. Appendix 1: Diffusion-Reaction Equation of OH• in the

Polyphosphate Accumulating cells

The diffusion and reaction of hydroxyl radical produced at the surface of inter-

cellular polyphosphate in the polyphosphate accumulating organisms is considered the

problem of Fickian diffusion including a first order irreversible reaction. The system can

be described as the following in one-dimensional coordinates:

x

CCk

x

CD r

∂

∂=−

∂

∂2

2

The solution to this problem can be found in Basmadjian & Farnood (2010) as

following:

−+

−−= tk

tD

xerfcDkxtk

tD

xerfcDkx

C

Crrrr

2)/exp(

2

1

2)/exp(

2

1

0

where C0 is the initial concentration of OH•, and C is the concentration of OH• at distance

x from the polyphosphate granule, at time t. D is the diffusion coefficient of OH• in the

cytoplasm, and kr is and the reaction rate constant of OH• with cell material.

The value of D of OH• in water is estimated by Campo & Grigera (2005) as 7.1E-9

m2s-1. The value of kr is extracted from Liu et al. (2007), where they reported the

degradation rate of proteins and carbohydrates in the presence of hydroxyl radicals as

1.25E-4 s-1. Substituting the values of D and kr and t as 60 seconds (a smaller value than

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the exposure time for the lowest UV dose), the distance for a 1 log reduction of OH• can be

estimated as 100 µm.

8.1.1. References

Basmadjian, D., Farnood, R., & Basmadjian, D. (2007). The art of modeling in science and engineering with Mathematica. Boca Raton: Chapman & Hall/CRC. Campo, M. G., & Grigera, J. R. (2005). Classical molecular-dynamics simulation of the hydroxyl radical in water. The Journal of Chemical Physics, 123(8), 084507. Liu, Y., Li, J., Qiu, X., Burda, C. (2007). Bactericidal activity of nitrogen-doped metal oxide nanocatalysts and the influence of bacterial extracellular polymeric substances (EPS). Journal of Photochemistry and

Photobiology A: Chemistry, 190(1), 94-100.

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