ANAEROBIC - AEROBIC TREATMENT OF DOMESTIC SEWAGE CONTAMINANTS

ANAEROBIC - AEROBIC TREATMENT OF DOMESTIC SEWAGE CONTAMINANTS
1 ANAEROBIC - AEROBIC TREATMENT OF DOMESTIC SEWAGE
FOR THE REMOVAL OF CARBONACEOUS AND NITROGENOUS
CONTAMINANTS
by
Qais Hisham BaniHani
Copyright © Qais Hisham BaniHani 2009
A Dissertation Submitted to the Faculty of the
DEPARTMENT OF CHEMICAL AND ENVIRONMENTAL ENGINEERING
In Partial Fulfillment of the Requirements
For the Degree of
DOCTOR OF PHILOSOPHY
WITH A MAJOR IN ENVIRONMENTAL ENGINEERING
In the Graduate College
THE UNIVERSITY OF ARIZONA
2009
2 THE UNIVERSITY OF ARIZONA
GRADUATE COLLEGE
As members of the Dissertation Committee, we certify that we have read the dissertation
prepared by Qais Hisham Banihani, entitled Anaerobic – Aerobic Treatment of Domestic
Sewage for the Removal of Carbonaceous and Nitrogenous Contaminants, and
recommend that it be accepted as fulfilling the dissertation requirement for the Degree of
Doctor of Philosophy.
________________________________________________Date: 11/16/2009
James A. Field
________________________________________________Date: 11/16/2009
Robert G. Arnold
________________________________________________Date: 11/16/2009
Maria R. Sierra Alvarez
________________________________________________Date: 11/16/2009
Thomas Meixner
Final approval and acceptance of this dissertation is contingent upon the candidate’s
submission of the final copies of the dissertation to the Graduate College.
I hereby certify that I have read this dissertation prepared under my direction and
recommend that it be accepted as fulfilling the dissertation requirement.
_______________________________________________ Date: 11/16/2009
Dissertation Director: James A. Field
3 STATEMENT BY AUTHOR
This dissertation has been submitted in partial fulfillment of requirements for an
advanced degree at the University of Arizona and is deposited in the University Library
to be made available to borrowers under rules of the Library.
Brief quotations from this dissertation are allowable without special permission, provided
that accurate acknowledgment of source is made. Requests for permission for extended
quotation from or reproduction of this manuscript in whole or in part may be granted by
the copyright holder.
SIGNED: Qais Hisham BaniHani
4 ACKNOWLEDGEMENTS
I would like to express my greatest gratitude to my advisor Dr. James Field for his
guidance, caring and patience. Without his encouragement this dissertation would not
have been possible. Special thanks are extended to my committee members, Dr. Robert
Arnold, Dr. Reyes Sierra-Alvarez and Dr. Thomas Meixner for their invaluable help and
advice on this research work.
I would like to honor my parents for their unconditional love, support, encouragement
and faith during my academic endeavors.
I am grateful to my wife, Shoroog Shunnag for her endless love. She always stood by my
side and sheered me up through hard times. She is the most precious gift that God gave
me. This dissertation is mine as much as it is hers.
I would like to recognize colleagues who have helped me throughout the years: Otakuye,
Irail, Pete, Kim and Carry. In addition, I would like to recognize my undergraduate
students Mark and Miranda.
Finally, I would like to thank USGS and UA Water Sustainability program for their
financial support.
5 DEDICATION
To my proud parents Hind and Hisham
And
To my beloved wife Shoroog
6 TABLE OF CONTENTS
LIST OF TABLES……………………………………………………………………….12
LIST OF FIGURES……………………………………………………………………...14
ABSTRACT……………………………………………………………………………...20
OBJECTIVES……………………………………………………………………………22
DISSERTATION OVERVIEW………………………………………………………….24
CHAPTER 1: INTRODUCTION……...………………………………………………..28
1.1 Anaerobic Treatment of Domestic Wastewater………………………………28
1.1.1 Domestic wastewater sources…………………………………………….31
1.1.2 Domestic wastewater characteristics……………………………………..31
1.2 Microbiology and Biochemistry of Anaerobic Digestion…………………….34
1.2.1 Hydrolysis………………………………………………………………..37
1.2.2 Acidogenesis……………………………………………………………..38
1.2.3 Acetogenesis……………………………………………………………..39
1.2.3.1
Obligate hydrogen producing acetogens…………………………...39
1.2.3.2 Homoacetogenic bacteria…………………………………………..43
1.2.4 Methanogenesis………………………………………………………….43
1.2.4.1
Acetoclastic methanogens………………………………………….44
1.2.4.2
Hydrogen utilizing methanogens…………………………………...46
1.3 High Rate Anaerobic Systems………………………………………………..47
1.3.1 UASB process description……………………………………………….52
1.3.2 Effects of suspended solids in domestic wastewater……………………53
1.3.3 Applicability of UASB technology for treating domestic wastewater….54
7 TABLE OF CONTENTS - Continued
1.3.4 Post treatment of anaerobic processes…………………………………..57
1.4 Fundamentals of Biological Nitrogen Removal………………………….......61
1.4.1 Nitrification…………………………...………………………………….61
1.4.2 Denitrification…………………………………………..………………..64
1.4.3 Anaerobic ammonium oxidation (Anammox)…………………………..65
1.5 References……………………………………...…………………………......71
CHAPTER 2: METHANOGENIC INHIBITION BY INORGANIC FLUORIDE..…...84
2.1 Abstract…………………………………………………..…………………...84
2.2 Introduction………………………………………..……………………...…..86
2.3 Materials and Methods…………………………………………..…………....90
2.3.1
Chemicals……………………………………………..…………………90
2.3.2
Sludge sources…………………………………………………..……….90
2.3.3
Basal media……………………………..………………………………..91
2.3.4 Methanogenic toxicity assays…………………………….……………...91
2.3.5
Analytical methods………………………………….…………………...93
2.4 Results………………………………….……………………………………..94
2.4.1 Toxicity to acetoclastic methanogens………………………….……......94
2.4.2 Toxicity to Hydrogenotrophic methanogens……………………….…..103
2.5 Discussion………………………………………………………………..…..111
2.6 Conclusions..………………………………………………………………...115
2.7 References…………………………………………………………………...116
8 TABLE OF CONTENTS - Continued
CHAPTER 3: TREATMENT OF HIGHT STRENGTH SYNTHETIC SEWAGE
IN A LABORATORY- SCALE UPFLOW ANAEROBIC SLUDGE
BED WITH AEROBIC ACTIVATED SLUDGE POST
TREATMENT………………………………………………………….121
3.1 Abstract……………………………………..……………………………….121
3.2 Introduction……………………………………………..…………………...123
3.3 Materials and Methods………………………………………………………130
3.3.1
Chemicals………………………………………………………..……..130
3.3.2
Experimental setup……………………………………………………..130
3.3.3
Seed sludge………………………………………………………..……134
3.3.4
Analytical methods……………………………………………………..134
3.4 Results…………………………………………………...…………………..136
3.4.1 Performance of combined UASB-AS reactor system for organic
removal…………………………………………………………………138
3.4.1.1 Total COD removal………………………………..……………....138
3.4.1.2 Total VFA removal………………………………………….…….143
3.4.1.3
3.4.2
Protein removal……………………………………………….…...149
Performance of combined UASB-AS system for nitrogen removal…...152
3.4.2.1
Ammonium removal in AS reactor………………………….….....152
3.4.2.2 NOx Species………………………………………………………..154
3.4.2.3 Nitrogen balance in AS…………………………………………....156
3.5 Discussion……………………………………………………………….…..159
3.6 Conclusion…………………………………………………………………...163
9 TABLE OF CONTENTS - Continued
3.7 References…………………………………………………………………...164
CHAPTER 4: NITRATE AND NITRITE INHIBITION OF METHANOGENESIS
DURING DENITRIFICATION IN GRANULAR BIOFILMS AND
DIGESTED DOMESTIC SLUDGES……….………………………….170
4.1 Abstract…………..……………………………………………………….….170
4.2 Introduction……………………………………..…………………………...172
4.3 Materials and Methods…………………………………………………..…..178
4.3.1
Chemicals………………………………………………………….…...178
4.3.2
Sludge sources……………………………………………………….....178
4.3.3
Basal media………………………………………………………….....179
4.3.4 Methanogenic toxicity assays……………………………………….….179
4.3.5
Analytical methods………………………………………………….….181
4.4 Results…………………………………………………………………….....182
4.4.1
Methanogenic inhibition…………………………………………….....182
4.4.2 Electron balance………………………………………………………..192
4.4.3
Nitrate and nitrite metabolism………………………………………....195
4.5 Discussion…………………………………………………………………...197
4.6 Conclusions……………………………………………………………….....204
4.7 References…………………………………………………………………...205
10 TABLE OF CONTENTS - Continued
CHAPTER 5: ENRICHMENT OF ANAEROBIC AMMONIUM OXIDIZING
(ANAMMOX) BACTERIA FROM WASTEWATER SLUDGE FOR
BIOLOGICAL NUTRIENT NITROGEN REMOVAL..........................209
5.1 Abstract……………………………………………………………………... 209
5.2 Introduction………………………………………………………………….211
5.3 Materials and Mehtods………………………………………………………217
5.3.1
Sludge sources………………………………………………………….217
5.3.2
Basal media……………………………………………………………..218
5.3.3 Screening assay………………………………………………………....219
5.3.4
Analytical methods……………………………………………………..220
5.4 Results and Discussion……………………………………………………....221
5.4.1
Methanogenic granular sludges………………………………………...221
5.4.2
Municipal WWTPs sludges…………………………………………….226
5.4.2.1 Return activated sludge……………………………………………226
5.4.2.1.1
Stage I…………………………………………………………226
5.4.2.1.2 Stage II………………………………………………………...229
5.4.2.1.3
5.4.2.2
Stage III………………………………………………………..230
Oxidation ditch sludge……………………………………………..234
5.5 Conclusion…………………………………………………………………...237
5.6 References…………………………………………………...........................238
CHAPTER 6: CONCLUSIONS………………………………………………………..243
6.1 Methanogenic Inhibition by Fluoride…………………………………….....243
6.2 Treatment of High Strength Synthetic Sewage in a Laboratory-Scale
Upflow Anaerobic Sludge Bed (UASB) with Aerobic Activated Sludge (AS)
Post Treatment…………………………………………................................245
11 TABLE OF CONTENTS - Continued
6.2.1 Organic matter removal………………………………………………..245
6.2.2 Nitrogen removal………………………………………………………246
6.3 Nitrate and Nitrite Inhibition of Methanogenesis During Denitrification
in Granular Biofilms and Digested Domestic Sludge………........................247
6.4 Enrichment of Anaerobic Ammonium Oxidizing (Anammox) Bacteria
From Wastewater Sludge for Biological Nutrient Nitrogen Removal……..248
SUMMARY……………………………………………………………………………250
12 LIST OF TABLES
CHAPTER 1
Table 1.1
Main advantages and constraints of anaerobic sewage treatment……….30
Table 1.2
Composition of raw domestic wastewater in different cities……….........33
Table 1.3
Stoichiometry and change in free-energy of the reactions for
catabolism of propionate and butyrate by H2-producing acetogens
in pure culture with H2-utilizing methanogens…………………………..41
Table 1.4
Benefits and drawbacks of various high rate anaerobic reactor types…...50
Table 1.5
Treatment performance of the first full scale UASB plants treating
domestic wastewater. COD refers to total COD of the raw wastewater…56
Table 1.6
Half-reactions and enzymes involved in denitrification process………...65
Table 1.7
Gibbs free energy of several reactions involved in dentrification
and ammonium oxidation…………………………………………….…..69
CHAPTER 2
Table 2.1
Inhibitory effect of fluoride on the acetoclastic and hydrogenotrophic
methanogens……………………………………………..……………….99
CHAPTER 3
Table 3.1
Composition of synthetic domestic sewage…………………………….133
Table 3.2
Periods and Operating Conditions of the UASB and AS Reactors……..136
Table 3.3
COD Mass Balance for UASB reactor………………………………….141
Table 3.4
Summary of Process Performance of UASB unit, AS unit and
Combined System for Organic Constituent Removal…………………..142
13 LIST OF TABLES - Continued
Table 3.5
Summary of Process Performance of AS unit for Nitrogen Removal….156
CHAPTER 4
Table 4.1
Specific rates of NOx- metabolism in anaerobic sludge with no prior
enrichment………………………………………………………………196
CHAPTER 5
Table 5.1
Origin, sample and VSS of each inocula………………………………218
Table 5.2
Summary results of all the experiments……………………………....224
14 LIST OF FIGURES
CHAPTER 1
Figure 1.1
Metabolic steps and microbial groups involved in anaerobic
treatment………………………………………………………………..36
Figure 1.2
Effect of hydrogen partial pressure on the free energy of conversion
of ethanol, propionate, acetate and hydrogen during methane
fermentation…………………………………………………………….42
Figure 1.3
Monod growth curves of the acetotrophic methanogens Methanosarcina
spp. and Methanosaeta spp……………………………………………...45
Figure 1.4
Various configurations of anaerobic reactors………………………...…51
Figure 1.5
Schematic diagram of UASB reactor……………………………...……53
Figure 1.6
Input and output of compounds in an anaerobic reactor treating
domestic sewage………………………………………………...……...58
CHAPTER 2
Figure 2.1
Time curse of cumulative methane production for mesophilic acetoclastic
methanogens present in Eerbeck sludge in the presence of increasing
fluoride concentrations………………………………………………….95
Figure 2.2
Time curse of cumulative methane production for mesophilic acetoclastic
methanogens present in Ina Rd sludge in the presence of increasing
fluoride concentrations………………………………………………….96
Figure 2.3
Inhibitory effect of fluoride on specific activity of mesophilic acetoclastic
methanogens present in Eerbeek……………………………………......97
15 LIST OF FIGURES - Continued
Figure 2.4
Inhibitory effect of fluoride on specific activity of mesophilic acetoclastic
methanogens present in Ina Rd sludge………………………………...98
Figure 2.5
Time curse of cumulative methane production for thermophilic
acetoclastic methanogens present in Hyperion sludge in the presence
of increasing fluoride concentrations…………………………………101
Figure 2.6
Inhibitory effect of fluoride on specific activity of thermophilic
acetoclastic methanogens present in Hyperion sludge……..................102
Figure 2.7
Time curse of cumulative methane production for mesophilic
hydrogenotrophic methanogens present in Eerbeck sludge in the
presence of increasing fluoride concentrations……………………….104
Figure 2.8
Time curse of cumulative methane production for mesophilic
hydrogenotrophic methanogens present in Ina Rd sludge in the presence
of increasing fluoride concentrations…………………………………105
Figure 2.9
Inhibitory effect of fluoride on specific activity of mesophilic
hydrogenotrophic methanogens present in Eerbeck………………….106
Figure 2.10
Inhibitory effect of fluoride on specific activity of mesophilic
hydrogenotrophic methanogens present in Ina Rd sludge……………107
Figure 2.11
Time curse of cumulative methane production for thermophilic
hydrogenotrophic methanogens present in Hyperion sludge in the
presence of increasing fluoride concentrations……………………….109
16 LIST OF FIGURES - Continued
Figure 2.12
Inhibitory effect of fluoride on specific activity of thermophilic
hydrogenotrophic methanogens present in Hyperion sludge…………110
CHAPTER 3
Figure 3.1
Schematic diagram of the experimental set-up of the combined system
reactor………………………………………………………………….132
Figure 3.2
Time course of organic loading rate (OLR) for (●) UASB and (∆) AS
reactors………………………………………………………………...137
Figure 3.3
Time course of influent COD (●), UASB effluent COD (○), AS effluent
COD (∆)……………………………………………………………….139
Figure 3.4
Time course of total COD removal efficiency by: (●) UASB and (□)
combined UASB-AS system………………………………………….140
Figure 3.5
Time course of influent VFA (●), UASB effluent VFA(○), AS effluent
VFA (∆)………………………………………………………………..145
Figure 3.6
Time course of influent acetic acid (●), UASB effluent acetic acid (○),
AS effluent acetic acid (∆)……………………………………………146
Figure 3.7
Time course of influent propionic acid (●), UASB effluent propionic acid
(○), AS effluent propionic acid (∆)…………………………………...147
Figure 3.8
Time course of influent butyric acid (●), UASB effluent butyric acid (○),
AS effluent butyric acid (∆)…………………………………………..148
Figure 3.9
Time course of influent protein (●), UASB effluent protein (○),
AS effluent protein (∆)………………………………………………..150
17 LIST OF FIGURES - Continued
Figure 3.10
Time course of protein removal efficiency by: (●) UASB and (□)
combined UASB-AS system. The first dashed line indicates the start
of AS operation……………………………………………………….151
Figure 3.11
Time course of influent NH4+ (●), UASB effluent NH4+ (○) and AS
effluent NH4+ (∆)……………………………………………………...153
Figure 3.12
Post aerobic treatment effluent NO2--N (●) and NO3--N (○)
concentrations………………………………………………………...155
Figure 3.13
Nitrogen balance on AS. Symbols: (♦) NH4+ removed, (◊) effluent
NO2- -N and NO3--N formed………………………………….……….158
CHAPTER 4
Figure 4.1
Impact of NO3--N concentration on methane formation and
denitrification by Mahou granular sludge with acetate as the electron
donor………..........................................................................................183
Figure 4.2
Impact of NO3--N concentration on methane formation and
denitrification by Mahou granular sludge with hydrogen gas as the
electron donor........................................................................................184
Figure 4.3
Impact of NO2--N concentration on methane formation and
denitrification by Mahou granular sludge with acetate as the electron
donor……..……………………………………………………………186
Figure 4.4
Impact of NO2--N concentration on methane formation and
denitrification by Mahou granular sludge with hydrogen gas as the
electron donor........................................................................................187
18 LIST OF FIGURES - Continued
Figure 4.5
Impact of NO3--N concentration on methane formation and
denitrification by Ina municipal digester sludge with hydrogen
gas as the electron donor…………………………...............................189
Figure 4.6
Relative methanogenic activity as a function of NOx--N concentration.
Panel A. Mahou granular sludge with added NO3-……………………191
Figure 4.7
The balance of electron equivalents used for methanogenesis and NO3removal as a function of increasing NO3- addition in the assay with
Mahou sludge and acetate………………………………………….....193
Figure 4.8
The balance of electron equivalents used for methanogenesis and NO3removal as a function of increasing NO3- addition in the assay with Ina
sludge and hydrogen……………………………………………..……194
Figure 4.9
Inhibition of methanogenic activity during the exposure as a function of
the maximum NO2--N concentration measured………………………200
CHAPTER 5
Figure 5.1
Time course of nitrite and ammonium for Mahou methanogenic granular
sludge…………………………………………………………………225
Figure 5.2
Rapid enrichment of an anammox culture from return activated sludge
inoculum obtained at a biological nutrient removal plant treating
municipal wastewater (Ina Road)……………………………………..228
Figure 5.3
The time course of the conversion ratio of nitrite to ammonium for return
activated sludge inoculum obtained at a biological nutrient removal plant
treating municipal wastewater (Ina Road)…………………………….233
19 LIST OF FIGURES - Continued
Figure 5.4
Rapid enrichment of an anammox culture from oxidation ditch
inoculums obtained at a biological nutrient removal plant treating
municipal wastewater (Green Valley)……………...............................236
20 ABSTRACT
Domestic wastewater is the most abundant type of wastewater. Direct discharge of
untreated domestic wastewater has environmental and public health risks due to the
presence of organics, nutrients and pathogens. Application of anaerobic processes for the
treatment of domestic sewage, which at present is largely treated by aerobic processes,
has drawn considerable attention recently. Anaerobic processes can be applied for the
removal of organic matter (methanogenesis) and nitrogen (anaerobic ammonium
oxidation (Anammox)).
The toxicity of fluoride to methanogenesis was investigated. The results indicate that
acetoclastic were more susceptible to fluoride than hydrogenotrophic methanogens. The
concentration of fluoride causing 50% inhibition (IC50) to acetoclastic ranged from 18.1
to 155.7 mg L-1 while for hydrogenotrophic methanogens was > 400.0 mg L-1.
The feasibility of a combined system consisting of anaerobic up-flow anaerobic sludge
blanket (UASB) followed by aerobic activated sludge (AS) reactor for removal of
carbonaceous and nitrogenous contaminants from strong synthetic sewage (2.5 g
chemical oxygen demand (COD) L-1) was also studied. The average combined removal
of total COD, volatile fatty acids (VFA) and protein was higher than 89.0%, 99.0% and
97.0%; respectively. Extensive nitrification (96.0%) was observed when dissolved
oxygen (DO) concentration was > 2.0 mg L-1. In contrast, only partial nitrification
21 occurred when the AS received high organic loads and/or the DO level was below 2.0 mg
L-1.
The inhibitory effect of nitrite and nitrate on methanogenesis was evaluated.
Methanogenic activity was inhibited by the presence of NOx- compounds (i.e., nitrite and
nitrate). The inhibition imparted by nitrate was not due to the nitrate itself, but rather to
its reduced intermediate, nitrite. The toxicity of NOx- to methanogens was found to be
reversible after all the NOx- were reduced during denitrification.
Moreover, the development of Anammox enrichment cultures was evaluated. Anammox
cultures were successfully developed using sludge samples collected from municipal
wastewater treatment plants (WWTPs) as inocula but not from methanogenic granular
sludges. Return activated sludge (RAS) collected from WWTP operating for biological
nitrogen removal had the highest intrinsic level of Anammox activity. RAS Anammox
culture was developed rapidly within 40 days with a doubling time of 6.8 days.
22 OBJECTIVES
The objectives of this Ph.D. research work are divided into four parts:
1. Study the inhibitory effects of inorganic fluoride on methanogenesis in granular
biofilms and digested domestic sludges.
a. Evaluate the inhibitory effect of fluoride on acetoclastic and
hydrogenotrophic methanogens.
b. Evaluate the effect of fluoride on methanogens present in granular biofilm
and in anaerobically digested sludge.
c. Evaluate the impact of long term exposure to fluoride on methanogens.
2. Investigate the feasibility of a combined system consisting of a (pre) anaerobic
treatment up-flow anaerobic sludge blanket (UASB) followed by aerobic
treatment activated sludge (AS) for removal of carbonaceous and nitrogenous
contaminants from strong synthetic domestic wastewater.
a. Study the organic matter removal in UASB, AS and the combined system.
b. Study the ammonium nitrogen removal in AS system.
c. Study the effect of dissolved oxygen concentration on the nitrification
process.
23 d. Study the effect of dissolved oxygen on the nitrogen speciation (nitrate vs.
nitrite).
3. Evaluate inhibitory effects of nitrite and nitrate on methanogenesis in granular
biofilms and digested domestic sludges.
a. Assess the inhibitory effect of nitrate and nitrite on methanogens present
in granular and anaerobically digested sludge.
b. Assess effects of substrate (acetate vs. hydrogen) on the inhibition of
methanogens.
4. Determined which sources of inoculum available at municipal wastewater
treatment plants has the highest level of intrinsic anammox bacteria, and would
therefore be most suitable as an inoculum to accelerate the start-up of the
anammox process.
a. Determine the background level of Anammox activity in different sludges.
b. Compare sludges collected from different operating units in municipal
wastewater treatment plants to determine under which conditions the
Anammox bacteria exist.
24 DISSERTATION OVERVIEW
This dissertation work is divided into six chapters. A brief overview of the content of
each chapter is given below.
CHAPTER 1.
INTRODUCTION
This chapter gives an overview on the anaerobic treatment of domestic wastewater
including its biochemistry and microbiology, the high rate anaerobic reactors and the
need for post treatment. The sources and the characteristics of the domestic wastewater
were discussed. Also, it covers the fundamentals of biological nitrogen removal such as
nitrification, denitrification and anaerobic ammonium oxidation (Anammox).
CHAPTER 2.
METHANOGENIC INHIBITION BY FLUORIDE
This chapter evaluates the inhibitory effect of fluoride on methanogenesis present in
granular biofilms and digested domestic sludges. To answer this question, methanogenic
batch toxicity assays were conducted using two different mesophilic inocula (methanogic
granular sludge and anaerbically digested sludge) and one thermophilic anaerobically
digested. The assays were supplemented with acetate and hydrogen for each inocula to
evaluate the effects of fluoride on both types of the methanogens. The results from this
25 study was published in a paper entitle “Toxicity of fluoride to microorganisms in
biological wastewater treatment systems: by Valeria Ochoa-Herrera, Qais Banihani,
Glendy León, Chandra Khatri, James A. Field and Reyes Sierra-Alvarez . Water
Research (2009) 43, 3177-3186.
CHAPTER 3.
TREATMENT OF HIGH STRENGTH SYNTHETIC SEWAGE IN A
LABORATORY-SCALE UPFLOW ANAEROBIC SLUDGE BED
(UASB) WITH AEROBIC ACTIVATED SLUDGE (AS) POSTTREATMENT
The main objective of this study is to investigate the feasibility of UASB-AS integrated
system for the removal of carbonaceous and nitrogenous contaminants from strong
synthetic domestic wastewater. The organic matter removal in both UASB and AS and
the ammonium nitrogen removal in the AS post treatment were monitored for about 627
days. In addition, the effect of dissolved oxygen concentration on the nitrification process
in order to promote partial nitrification in the AS post treatment was studied.
26 CHAPTER 4.
NITRATE AND NITRITE INHIBITION OF METHANOGENESIS
DURING DENITRIFICATION IN GRANULAR BIOFILMS AND
DIGESTED DOMESTIC SLUDGES
This study answers a question about the inhibitory effect of nitrate and nitrite on
methanogens in anaerobic bioreactors that can support simultaneous microbial processes
of denitrification and methanogenesis. To answer this question, toxicity batch assays
were conducted using methanogenic granular sludge and anaerobically digested sludge.
Two different electron donors, acetate and hydrogen, were added to study the effect of
these electron donors on toxicity of nitrate and nitrite. The results from this study was
published in a paper entitle “Nitrate and Nitrite Inhibition of Methanogenesis During
Denitrification in Granular Biofilms and Digested Domestic “ by Qais Banihani, Reyes
Sierra-Alvarez and James A. Field . Biodegradation (2009) 20, 801-812.
CHAPTER 5.
ENRICHMENT OF ANAEROBIC AMMONIUM OXIDIZING
(ANAMMOX) BACTERIA FROM WASTEWATER SLUDGE FOR
BIOLOGICAL NUTRIENT NITROGEN REMOVAL
The aim of this study is to determine which sources of inoculum available at municipal
wastewater treatment plants have the highest background level of anammox bacteria.
This is of particular importance in order to determine the sources that are most suitable as
an inoculum to accelerate the start-up of the anammox process.
27 CHAPTER 6.
CONCLUSIONS
The main conclusions from this dissertation are highlighted in this chapter.
28 CHAPTER 1
INTRODUCTION
1.1
Anaerobic Treatment of Domestic Wastewater
Domestic wastewater in terms of the quantity is considered the most abundant type of
wastewater on earth (vanLier et al., 2008). Direct discharge of untreated domestic
wastewater has a negative environmental impact and poses serious health concerns on
humans due to the presence of organic substances, nutrients and pathogens. The main
incentive of treating domestic wastewater is to reverse such consequences (Metcalf and
Eddy, 2003). Moreover, the worldwide population is growing rapidly and there is an
increase in the scarcity of drinking water resources (Verstraete et al., 2009). Therefore, a
better management of water resources is needed. One of the goals of environmental
protection and resource conservation concepts is the re-use of treated wastewater and its
valuable treatment byproducts (Lettinga et al., 2001; Yi, 2001). Consequently, by
implementing such goals, a wastewater such as domestic wastewater besides being
sanitized, it can be an important source of re-usable water, fertilizer, soil conditioner and
energy.
Anaerobic treatment is considered as one of the oldest methods used for the treatment of
sewage. It has been used for over 100 years (McCarty, 1982). Anaerobic treatment
29 involves the degradation of organic matter into useful end product methane in the
absence of free molecular oxygen (Grady et al., 1999). The application of anaerobic
processes for the treatment of low strength wastewater such as domestic sewage has
drawn considerable attention recently, which at present is largely treated by aerobic
processes. There are two main concerns related to aerobic treatments: high operating cost
associated with the aeration and high sludge production (Seghezzo et al., 1998).
Compared to the conventional aerobic methods and in the light of implementing
sustainable methods, anaerobic wastewater treatment offers unique characteristics
(McCarty, 2001). There is no need for adding electron acceptors such as oxygen or nitrate
in order for the process to work. The organic matter itself or the carbon dioxide produced
from its destruction serve this purpose. Also, the oxygen mass transfer limitations are not
involved; thus; much higher organic loading rate can be applied to the anaerobic reactors
compared to the aerobic one. This would result in a smaller reactor size. During the
anaerobic treatment useful energy in form of methane is produced instead of consuming
energy as in the aerobic methods (Verstraete and Top, 1992). Moreover, only few percent
of the chemical oxygen demand (COD) is converted into biomass; thus, the volume of the
sludge produced is significantly lower (Aiyuk et al., 2006). These produced biosolids are
well stabilized and can be used as soil conditioners (McCarty, 2001). From these
perspectives, anaerobic treatment of domestic wastewater appears to be a sustainable
option and economically more attractive. The main advantages and constraints of
anaerobic sewage treatment are listed in table 1.1.
30 Table 1.1 Main advantages and constraints* of anaerobic sewage treatment (Adapted
from van Lier et al., 2008).
Advantages
•
Sustainable savings, reaching 90%, in operational costs and no energy is required for
aeration.
•
4-60% reduction in investment cost as less treatment units are required.
•
If implemented at appropriate scale, the produced CH4 is of interest for energy recovery
or electricity production.
•
The technologies do not make use of high –tech equipment, except for main headwork
pumps and fine screens. Treatment system is less dependent on imported technologies.
•
The process is robust and can handle periodic high hydraulic and organic loading rates.
•
Technologies are compact with average HRTs between 6 and 9 hrs and are; therefore,
suitable for application in the urban areas, minimizing conveyance costs.
•
Small scale applications allow decentralization in treatment, making sewage treatment
less dependent on the extent of the sewerage network.
•
The excess sludge production is low, well stabilized and easily dewatered so it does not
require extensive post treatment.
•
The valuable nutrients (N and P) are conserved which give high potential for crop
irrigation
•
A well designed UASB** filters Helminth’s eggs from the influent, a prerequisite prior to
agricultural reuse
Constraints
•
Anaerobic treatment is a partial treatment, requiring post-treatment for meeting the
discharge or reuse criteria
•
The produced methane is largely dissolved in the effluent (depending on the influent
COD concentration). So far no measure are taken to prevent methane escaping to the
atmosphere
•
The collected methane is often not recovered nor flared
•
There is little experience with full-scale application at moderate to low temperatures
•
Reduced gases like H2S, that are dissolved in the effluent may escape causing odor
problems
* Compared to activated sludge processes; **UASB: Up-flow anaerobic sludge blanket reactor.
31 1.1.1
Domestic wastewater sources
According to van Haandel and Lettinga (1994), domestic sewage refers to the wastewater
that is produced by a community. Such wastewater may originate from three different
sources. The first source is the domestic wastewater generated from toilet wastes (black
water) and household wastewater, i.e. wastewater produced from kitchen and bathroom.
The second source is the industrial wastewater originated from industries using the same
sewage system for their effluents. Although some industries treat and discharge their
effluents directly into surface water, the most common procedure is that the effluents are
discharges into the sewage system after (pre) treatment. The third source is the rain-water
especially in the case of a combined sewer system, i.e. sewer system constructed for both
wastewater and storm water. The sewage flow and composition depends on several
factors such as economic aspects, social behavior, type and number of industries located
in the collection area etc. The main component of the domestic sewage is the domestic
wastewater and is often used as a synonym (Seghezzo et al., 1998).
1.1.2
Domestic wastewater characteristics
Domestic wastewater has a low organic concentration with a typical COD concentration
ranging from 250 to 1000 mg L-1 (Metcalf and Eddy, 2003). Despite being low-strength
wastewater, the domestic wastewater is characterized as a complex type of wastewater. It
contains relatively large fraction of suspended solids in addition to fatty compounds,
32 proteins, detergents, among other barely known compounds (Aiyuk et al., 2006). Raw
domestic wastewater contains about 50-65% of the total COD in the form of suspended
solids (vanLier et al., 2008). The composition and biodegradability of domestic
wastewater varies depending on the location (Henze and A., 2001; Mahmoud, 2002). It
has been reported in literature that the biodegradability of the domestic sewage ranges
from 64% to 79% depending on the temperature and the level of fractionation
(Elmitwalli, 2000; Mgana, 2003). Table 1.2 shows the most important constituents of
domestic wastewater in three cities where anaerobic treatment systems of UASB type are
in operation.
33 Table 1.2 Composition of raw domestic wastewater in different cities (Adapted from van
Haandel and Lettinga, 1994).
Constituents (mg L-1)
Pedregal (Brasil)
Cali (Colombia)
Bennekom (The Netherlands)
8.5
--
--
Total
429.0
215.0
--
Ash
177.0
106.0
--
Volatile
252.0
107.0
--
BOD*
368.0
95.0
231.0
COD**
727.0
267.0
520.0
Nitrogen (as N)
44.0
24.0
45.0
Organic
10.0
17.0
--
Ammonia
34.0
17.0
--
Total
11.0
1.3
18.0
Orthophosphate
8.0
--
14.0
Organic
3.0
--
4.0
4 * 107
--
--
Sulfate
18.0
--
15.0
Chloride
110.0
--
--
Alkalinity
388.0
120.0
350.0
Calcium
110.0
--
4.0
Magnesium
105.0
--
2.0
Maximum
26.0
27.0
20.0
Minimum
24.0
24.0
8.0
Settleable solids (mL L-1)
Suspended solids
Phosphorus
Escherichia Coli
(number in 100 mL)
Temperature (○C)
*
BOD: Biological Oxygen demand; **COD: Chemical oxygen demand.
34 1.2
Microbiology and Biochemistry of Anaerobic Digestion
Anaerobic digestion (AD) refers to the fermentation process in which the organic matter
is degraded and biogas composed mainly of methane and carbon dioxide is produced in
anaerobic environment that receives a limited supply of oxygen, sulfate and nitrate. This
process occurs in natural ecosystems where the organic matter is available and the redox
potential is low (no molecular oxygen) such as rivers, lakes, swamps, soils, sediments
and digestive systems of ruminant animals (Novaes, 1986; vanLier et al., 2008). Similar
processes have also been utilized in anaerobic treatment (digestion) of wastes and
wastewaters containing organic matter, thus reducing their organic contents (Grady et al.,
1999).
The anaerobic digestion is a result of a complex interaction among different bacterial
groups (Novaes, 1986). Anaerobic degradation of organic matter is a multi step process
of series and parallel biological reactions during which the products from one group of
microorganisms serve as a substrates for the next microbial group (Pavlostathis and
Giraldo-Gomez, 1991). Thus, the microbial groups form a chain- anaerobic food chainin which organic complex compounds are degraded into a more simple compounds as
they are passed along the food chain (Gerardi, 2003). A simplified model of the anaerobic
degradation of the organic matter indicating the main metabolic stages is shown in Figure
1.1. Anaerobic digestion is carried out by four successive steps: hydrolysis, acidogenesis,
acetogenesis and methanogenesis. During these four steps the microbial groups from two
biological kingdoms, the Bacteria and the Archaea, cooperate sequentially in order to
35 achieve the degradation of a variety of polymeric and monomeric substrates (Dugba and
Zhang, 1999; Veeken and Hamelers, 2000).
In order to achieve a stable digestion
process, it is important that the biological conversion processes remain sufficiently
coupled during this process in order to avoid the accumulation of intermediates in the
process which results in the failure of this process (Kaseng et al., 1992). Figure 1.1 shows
the metabolic and microbial groups involved in anaerobic treatment.
36 COMPLEX ORGANIC COMPOUNDS
(Carbohydrates, Proteins, Lipids)
1
HYDROLYSIS
SIMPLE ORGANIC COMPOUNDS
(Sugars, Amino acids, Peptides)
1
ACIDOGENESIS
LONG CHAIN FATTY ACIDS
(Propionate, Butyrate, etc)
2
H2, CO2
4
METHANOGENESIS
ACETOGENESIS
3 ACETOGENESIS
CH4, CO2
ACETATE
5
METHANOGENESIS
Figure 1.1 Metabolic steps and microbial groups involved in anaerobic treatment: (1)
Fermentative (hydrolytic) bacteria, (2) H2-producing acetogenic bacteria, (3) H2consuming acetogenic or homoacetogenic bacteria, (4) CO2-reducing methanogen, and
(5) acetoclastic methanogen. (Adapted from Novaes, 1986).
37 1.2.1
Hydrolysis
The first step of the anaerobic digestion consists of the hydrolysis of the complex
insoluble compounds such as particulates and colloidal organic matter. These complex
organic matters are degraded through the action of extracellular enzymes excreted by the
hydrolytic bacteria to produce simpler, dissolved compounds usually monomers or
dimmers which in turn are readily available for uptake by the bacterial cells (MataAlvarez, 2002). According to Pavlostathis and Giraldo-Gomez (1991) the three organic
groups that are considered as the major components of the complex organics are
polysaccharides, proteins and lipids. During the enzymatic hydrolysis process the
polysaccharides are hydrolyzed to simple sugars, the proteins to amino acids and the
lipids to long chain fatty acids (LCFA) (Zeikus, 1979). Hydrolysis in most cases,
especially with wastewaters containing high suspended solids, is considered the ratelimiting step for the overall anaerobic digestion process. The hydrolysis process is merely
a surface phenomenon; therefore, the availability of free accessible surface area of the
solid particles and the overall structure of the solids affects the hydrolysis step (Chandler
et al., 1980; van Lier et al., 2008). Moreover, the hydrolysis step is also affected by other
factors such as the availability of hydrolytic enzymes, pH, temperature and the solid
retention time (Pavlostathis and Giraldo-Gomez, 1991; Sanders et al., 2000). The design
of the anaerobic digesters for wastewaters containing high solid contents, particularly in
cold regions, is usually based on the hydrolysis step (van Lier et al., 2008).
38 1.2.2 Acidogenesis
During the acidogenesis step, the hydrolysis products (simple sugars, amino acids and
LCFA) are used as substrates by acidogenic bacteria to produce a wide variety of
fermentation end-products including organic acids (e.g., acetic, propionic, butyric and
other short chain fatty acids) formate, methanol, H2 and CO2 (Gujer and Zehnder, 1983;
Fang et al., 2002). The conversion of organic nitrogen compounds such as amino acids
and proteins are of particular importance since such compounds yield ammonium
nitrogen. At high temperature and alkaline pH, a large portion of the ammonium is then
present in unionized form (i.e., ammonia, pKa = 9.25) which is toxic and inhibitory
(Angelidaki and Ahring, 1993). McCarty and McKinney (1961) reported that an
ammonia concentration of about 100.00 mg L-1 caused inhibition in acetate-fed anaerobic
system.
The acidogenesis reactions are performed by a large group of hydrolytic and nonhydrolytic microorganisms including many different fermentative genera and species
(O’Flaherty et al., 2006). Among them are Clostridium, Bacteroides, Ruminococcus,
Desulfobacter and Escherichia (Anderson et al., 2003). The facultative members of the
acidogenic bacteria consume the oxygen that may enter with the feed; thus, they protect
the oxygen-sensitive methanoges (Anderson et al., 2003).
The acidogenesis is the most rapid conversion step in the anaerobic digestion processes
(Mosey and Fernandes, 1989). The growth rate of acidogenic bacteria is ten to twenty
folds higher than methanogeneis (Anderson et al., 2003). For this reason, the anaerobic
39 digesters are subjected to souring (i.e., sudden pH drop) when the digesters are
overloaded or perturbed by toxic compounds (Gerardi, 2003). This is because the
alkalinity will be consumed by the volatile organic acids that are produced from the
acidogenesis step and the pH will start to drop. This results in increasing the unionized
forms of the volatile fatty acids leading to a more severe inhibition to methanogens (van
Lier et al., 2008).
1.2.3 Acetogenesis
The main function of the acetogic bacteria in the anaerobic digestion is the production of
acetate, hydrogen gas and carbon dioxide (Grady et al., 1999). This is of a particular
importance for the successful production of biogas since these products are the only
substrate that can be metabolized by the methanogens in the final stage of the anaerobic
digestion (Novaes, 1986). The acetogenic bacteria can be distinguished into two groups
based on their metabolism.
1.2.3.1
Obligate hydrogen producing acetogens
Obligate hydrogen producing acetogens (OHPA) further oxidize the products of the
acidogenic step into acetate, hydrogen gas and carbon dioxide. OHPA are slow-growing
bacteria and their metabolism is inhibited by the hydrogen (van Lier et al., 2008). Studies
carried out on the acetogenic reactions have elucidated the requirement of a syntrophic
40 association between the OHPA and the hydrogen-consuming methanogens in order to
regulate the hydrogen level in their environment. An explanation for this syntrophic
relationship has been presented by Bryant (1979) and is summarized in Table 1.3. Under
standard conditions, the catabolism of the propionate (reaction A) and the butyrate
(reaction B) which are substrate for the acetogenic bacteria are endergonic reactions and
the reactions would not proceed alone as indicated by the positive free Gibbs energy
(ΔG0’) in Table 1.3. However, the hydrogen utilization reaction by hydrogen consuming
methanogesn (reaction C) is exergonic reaction and would proceed because of its
negative ΔG0’ (Table 1.3). When the OHPA is in syntrophic relationship with the
hydrogen consuming methanogens (sum of reactions A+C and sum of reactions B+C),
the combined reactions (reactions D and E) become energetically favorable (i.e. negative
ΔG0’) and the acetogenic reactions proceed. Thus, OHPA bacteria must always grow in
syntrophy with the hydrogen-consuming methanogens, sulfate-reducing bacteria or
homoacetogenic bacteria in order to facilitate the interspecies hydrogen transfer and gain
energy from the growth on the acidogenic products (McCarty, 1982; Schink, 1992).
According to Daniels (1984), in order for the propionate and butyrate reactions to
proceed in the acetogenesis phase, the hydrogen partial pressure must be less than 10-4
and 10-3 atm
, respectively. This means that propionate OHPA bacteria are more
sensitive to hydrogen pressure than butyrate OHPA bacteria. Figure 1.2 shows the
relationship between the hydrogen partial pressure and the Gibbs free energy available
for the OHPA bacteria. The OHPA include i) valerate- and butyrate-degrading bacteria,
41 e.g. Syntrophomonas wolfei and ii) propionate degrading bacteria, e.g. Syntrophobacter
wolinii (Anderson et al., 2003).
Table 1.3 Stoichiometry and change in free-energy of the reactions for catabolism of
propionate and butyrate by H2-producing acetogens in pure culture with H2-utilizing
methanogens (Adapted from Bryant, 1979).
A. Propionate-catabolizing acetogenic bacterium
CH3CH2COO- + 3H2O→ CH3COO- + HCO3- + H+ + 3H2
ΔG0’= + 18.2 Kcal/reaction
B. Butyrate-catabolizing acetogenic bacterium
CH3CH2CH2COO- + 2H2O→ 2CH3COO- + H+ + 2H2
ΔG0’= + 11.5 Kcal/reaction
C. H2-utilizing methanogenic bacterium
HCO3- + 4H2 + H+ → CH4 + 3H2O
ΔG0’= - 32.4 Kcal/reaction
D. Sum A+C. Syntrophic association
4CH3CH2COO- + 3H2O→ 4CH3COO- + HCO3- + H+ +
3CH4
E. Sum B+C. Syntrophic association
2CH3CH2CH2COO- + HCO3- + H2O→ 4CH3COO- + H+ +
CH4
ΔG0’: change in Gibbs free energy ΔG0’= - 24.4 Kcal/reaction
ΔG0’= - 9.4 Kcal/reaction
42 Figure 1.2 Effect of hydrogen partial pressure on the free energy of conversion of
ethanol, propionate, acetate and hydrogen during methane fermentation (Adapted from
McCarty, 1981).
43 1.2.3.2 Homoacetogenic bacteria
Homoacetogenic bacteria are capable of autotrophic and heterotrophic growth catalyzing
the formation of acetate from hydrogen/carbon dioxide or multi-carbon compounds. The
most
common
autotrophic
genera
of
the
homoacetogens
are
Clostridium,
Acetobacterium, Acetogenium and Butribacterium (Dolfing, 1988). The homoacetogenic
bacteria also participate in the interspecies hydrogen transfer which maintains low
hydrogen pressure required by the OHPA (O’Flaherty et al., 2006). However, their
importance relative to the hydrogen consuming methanogens is still unclear (Anderson et
al., 2003).
1.2.4 Methanogenesis
During this final stage of the anaerobic degradation of organic matter, the Archaea
methanogens utilize mainly the hydrogen/carbon dioxide and acetic acid to form methane
and carbon dioxide. The methanogens can also produce methane by utilizing other
substrates such as methanol, methylamines, alcohol and formate (Hwang et al., 2001).
According to Bryant (1979) the ultimate break down of the organic matter would not be
complete without the methanogens due to the accumulation of the organic acids produced
by the acetogenesis phase which have an energy content almost equal to the original
organic matter. The methanogens are known to be truly distinct from the bacteria,
Eubacteria, and they are belong to a separate kingdom, Archchaea (Zeikus, 1979; MataAlvarez, 2002). Methanogenesis is considered the slowest step in the anaerobic digestion
44 process (Oremland, 1988). Acording to their substrate specificity, methanoges are
divided into acetoclastic methanogens and hydrogen consuming methanogens.
1.2.4.1
Acetoclastic methanogens
Acetate is considered as the most important precursor of the methane production and the
source of approximately 70% of the methane produced in the anaerobic digestion process
(Oremland, 1988; Solera et al., 2002). The rest of the methane generates mainly from the
hydrogen/carbon dioxide. The acetoclastic methanogens have a slow growth rate with a
doubling time of several days or even more (van Lier et al., 2008). For this reason, the
start-up of the anaerobic reactors with un-adapted seeding material require a long time up
to 7 months. It has been reported in literature that there are only two methanogenic
genera contain species that are able to use the acetate. These are Methanosaeta (formely
known as Methanothrix) and Methanosarcina.
The two types of the acetoclastic methanogens have different kinetic and morphological
characteristics. Methanosarcina spp. are characterized by a coccoid shape and have a
relatively wide substrate spectrum including acetate, hydrogen/carbon dioxide,
methylamines, methanol and formate (Smith et al., 1980). They have a high growth rate
with a doubling time of 1 day and a low substrate affinity for acetate of about 300 mg
COD L-1 (Zinder, 1984). Methanosaeta spp. are filamentous organisms and they can only
use acetate as substrate (Smith et al., 1980). Their growth rate (doubling time of 4-9 days)
is lower than Methanosarcina spp. but they have a higher substrate affinity for acetate
45 (Ks of 30 mg COD L-1) (Lafitte-Trouque and Forster, 2000). Despite their relatively low
growth rate, Methanosaeta spp. are the dominant acetoclastic in high rate anaerobic
bioreactors (Rittmann and McCarty, 2001). This is attributed to the fact that the aim of
the wastewater treatment systems is to achieve the lowest possible effluent
concentrations. The substrate concentrations inside the anaerobic granules or biofilms
approaches zero when the bulk liquid concentrations are low. Thus, Methanosarcina spp.
have a kinetic advantages under these conditions as can be seen in Figure 1.3.
Figure 1.3 Monod growth curves of the acetotrophic methanogens Methanosarcina spp.
and Methanosaeta spp. (Adapted from van Lier et al., 2008).
46 1.2.4.2
Hydrogen utilizing methanogens
A significant quantity of methane, up to 30% of the total methane produced in anaerobic
digestion, is produced by hydrogen utilizing methanogens (Zinder, 1984).
These
methanogens reduce carbon dioxide, formate, methanol and methylamine using the
hydrogen that was produced in the earlier stages of the anaerobic digestion processes
(Anderson et al., 2003). Thus, they are important in keeping the hydrogen partial pressure
below 10-4 to 10-6 atm in the anaerobic environment (Oremland, 1988). Unlike the
acetoclastic methanogens, nearly all the methanogens can produce methane from
hydrogen/carbon dioxide (Gujer and Zehnder, 1983). The hydrogen pathway is more
energy yielding (ΔG0’ = -131 KJ mol-1) than the acetate pathway (ΔG0’= -31 KJ mol-1);
therefore, it is normally not rate limiting like the acetate pathway (Bryant, 1979). The
hydrogen consuming methanogens have a much higher growth rate than acetoclastic
methanogens with a doubling time of 4 to 12 hours (vanLier et al., 2008).
47 1.3 High Rate Anaerobic Systems
One of the major successes in the development of anaerobic treatment was the
introduction of high rate anaerobic systems in 1960s and 1970s (Barber and Stuckey,
1999; van Lier et al., 2001). The introduction of these reactors was pioneered by Perry
McCarty who introduced the anaerobic filter (AF) (Young and McCarty, 1969). The high
rate anaerobic reactors are flexible systems, feasible for treating various types of
wastewater at different environmental conditions (Lettinga, 2001). Several anaerobic
reactors have been developed during the past decades including anaerobic
expanded/fluidized bed reactor (Switzenbaum and Jewell, 1980), anaerobic rotating
biological reactor (Tait and Friedman, 1980; Blank et al., 1983), anaerobic baffled reactor
(Bachmann et al., 1985) and up-flow anaerobic sludge blanket (UASB) (Lettinga et al.,
1980). Unlike the aerobic processes, the maximum permissible load in anaerobic
processes is not limited by the supply of any reagent such as oxygen (Lettinga, 1995).
The common feature among the high rate anaerobic reactors is that they operate at long
sludge retention time (SRT) under relatively short hydraulic retention time (HRT) due to
the retaining high concentrations of biomass. This means that high organic loading rate
can be applied to relatively small size reactors. Table 1.4 and Figure 1.4 show the
benefits and drawbacks of various high rate anaerobic processes as well as various
configurations of the different systems, respectively.
The main concept of the high rate anaerobic reactors is based on three fundamental
aspects (Iza et al., 1991):
48 a) High retention of viable biomass which also allows the retention of slow growing
microorganisms. This can be achieved by ensuring that the SRT is much higher than
HRT.
b) Improved contact between biomass and incoming wastewater. This overcomes the
diffusion problems of the substrates and products from the bulk liquid into biofilms or
granules.
c) Enhanced activity of the biomass due to the adaptation and growth of the active
biomass.
According to Hulshoff and Letinga (1986), retaining high viable biomass in the modern
high rate anaerobic reactors can be accomplished by some mode of bacterial
immobilization which can be achieved by one of the following methods:
•
Formation of highly settleable sludge aggregates combined with gas separation
and sludge settling such as in the UASB and anaerobic baffled reactors.
•
Bacterial attachment to carrier materials such as in the fluidized bed react and
anaerobic expanded bed reactor.
•
Entrapment of sludge aggregates between backing material supplied to the reactor
such as in the anaerobic filter and up-flow anaerobic filter.
Within the spectrum of high rate anaerobic technologies, UASB reactor is considered as
the most widely and successful applied technology with over 200 installation worldwide
(Singh et al., 1996; Aiyuk et al., 2006). It was first introduced to optimize the anaerobic
49 treatment of agro-industrial wastes (van Haandel et al., 2006). The UASB reactor has
been applied for treating different types of wastewater such as potato starch (Field et al.,
1987), sugerbeets (Lettinga et al., 1976) and pulp and paper wastes (Lettinga and Pol,
1991). The application of the system for the treatment of domestic wastewater was also
found to be feasible. The reactor offers great promise especially in developing countries
in tropical regions of the world where artificial heating can be avoided. The process has
been applied at full scale for domestic wastewater in many countries such as Brazil,
Colombia, Cuba, Uruguary and Argentina (Maaskant et al., 1991; Seghezzo et al., 1995;
deSousa and Foresti, 1996; Chernicharo and Nascimento, 2000; Florencio et al., 2000;
Rodriguez et al., 2000; Torres and Foresti, 2000).
50 Table 1.4 Benefits and drawbacks of various high rate anaerobic reactor types (Adapted
from Kaksonen, 2004).
Reactor Type
Anaerobic contact process
(ACP)
Anaerobic filter reactor
(AFR)
Benefits
•
Better retention of
biomass as compared to
CSTR
•
•
Low shear forces
•
Longer sludge retention
time
•
Possibility
to
utilize
gravitation in down-flow
mode
Channeling of the flow
possible
Pressure gradient can be
large
•
Energy needed for carrieir
fluidization
Shear forces can detach
biomas
Less volume available for
biomass compared to the
UASB reactor due to the
inert biomass carrieir
•
•
•
Fluidized bed reactor
(FBR)
•
•
•
•
•
•
Upflow anaerobic blanket
bed reactor (UASB)
•
•
•
•
•
•
Anaerobic hybrid reactor
(AHR)
Drawbacks
•
•
Large surface area for
biofilm formation due to
fluidized carrier material
High retention of biomass
on the carrier
Efficient mass transfer
Small pressure gradients
No channeling of the flow
Influent concentrations
diluted due to recycle flow
No clogging
Selects for microbes with
low Km values
No channeling of flow
No compacting of sludge
No costs due to biomass
carrier
No clogging
Possibility to obtain high
treatment rates
Less
susceptible
to
clogging compared to
AFR
Sludge removal easier
than in AFR
Biomass retention better
than in UASB
•
•
•
•
•
Pumping of biomass
breaks down flocks and
sludge
Biomass
flush
out
common during process
failures
More
susceptible
to
changes in influent quality
compared to AFR
51 Figure 1.4 Various configurations of anaerobic reactors. Symbols: I: Wastewater
influent; E: Wastewater effluent; D: Gas (Adapted from Kaksonen, 2004).
52 1.3.1 UASB process description
The UASB system was described in detail in literature (Lettinga et al., 1980; Lettinga and
Pol, 1991). The UASB reactor consists of a circular or rectangular tank in which
wastewater flows upward through a sludge bed or blanket of settled microorganisms. The
sludge bed occupies about half the volume of the reactor and consists mainly of highly
settleable granules or flocks. The main advantage of the system is that there is no support
material required for retaining high density anaerobic sludge (Elias et al., 1999). During
the passage of wastewater through the sludge bed of active biomass the treatment process
takes place by solid entrapment and organic matter conversion into sludge and biogas
consisting mainly of methane and carbon dioxide. The produced biogas provides
hydraulic mixing. The upper portion of the reactor has a three-phase separator used to
separate sludge granules from effluent from biogas. The UASB reactor does not require
expensive and energy consumption of pumps for recirculating the effluent (Wentzel et al.,
1994; Rajeshwari et al., 2000). The high settleability of the sludge (60-80 m h-1) allows
UASB to operate at high upflow velocities without loss of granules (Lettinga and Pol,
1991; Zoutberg and Eker, 1999). Figure 1.5 shows a schematic diagram of a UASB
reactor.
53 Figure 1.5 Schematic diagram of UASB reactor.
1.3.2
Effects of suspended solids in domestic wastewater
It is well documented that the suspended solids affect anaerobic treatment process
negatively (Kalogo and Verstraete, 1999; Zeeman and Lettinga, 1999). The impact of the
suspended solids depends on the type of inoculums used (Sayed and Fergala, 1995). In
the case of flocculent sludge, the accumulation of the suspended solids leads to an
increase in the sludge bed height. When the accumulated solids are not completely
digested, this can lead to diluting the active biomass. As a consequence, the sludge
activity decreases gradually. In granular sludge reactor, the suspended solids form a
barrier around the granular sludge. This barrier leads to the disintegration of the granules
under long term feeding conditions. Also, the accumulation of the suspended solids in the
reactor decreases the COD conversion efficiency. Thus, one of the biggest challenges for
54 the treatment of domestic wastewater in high rate anaerobic reactors is the high level of
suspended solids present in wastewater. Generally, 50% of the COD in domestic
wastewater consists of suspended solids (Mahmoud et al., 2003). The biodegradability of
the suspended solids has been estimated to range from 55 to 85% (Elmitwalli et al., 2003;
Alvarez et al., 2004). The SRT in the high rate anaerobic reactors needs to be sufficient in
order to ensure that the biodegradable fraction of the suspended solids becomes
stabilized. The required SRT depends on the temperature since the hydrolysis rate of the
suspended solids is dependent on the temperature (Ferreiro and Soto, 2003). For effective
treatment of raw sewage, it was suggested that the ratio of soluble COD to volatile
suspended solids (CODs/VSS) of the wastewater should be >10 (Vieira, 1988; Elmitwalli
et al., 1999) . The primary clarifier has been applied in some studies to remove the
suspended solids from raw sewage before feeding the UASB reactors. However, it can
take >10 hrs to sediment the major part of non biodegradable fraction of the suspended
solids (Elmitwalli et al., 1999).
1.3.3 Applicability of UASB technology for treating domestic wastewater
The literature review shows that UASB reactors have been successfully applied as a
primary unit in systems designed to treat domestic wastewater at temperatures higher
than 200C. According to Schellinkhout and Collazos (1992) applying the UASB as
primary unit for treating domestic sewage results in a reduction of total HRT by a factor
of 4 to 5 compared to pond systems. Satisfactory COD removal efficiencies of 65% to
80% have been reported in literature at an organic loading rate (OLR) lower than 3 Kg
55 COD m-3 d-1 and HRT ranging from 6 to 10 hrs (Van Haandel and Lettinga, 1994;
Chernicharo and Nascimento, 2001; Torres and Foresti, 2001). The HRT below 6 hrs has
been also practiced without significant impairment on the reactor’s performance in
tropical regions where the temperature is close to or higher than 250C (Foresti, 2002).
Table 1.5 lists some of the results of full scale UASB reactors under tropical conditions.
The start-up period has been considered as an important step for the stable operation of
the anaerobic reactors. One of the main points for starting-up anaerobic reactors is the
need of inoculating the reactors with high quality methanogenic sludge. Sludge of poor
quality can also be used (Rodriguez et al., 2000) and recent case studies on full scale
anaerobic reactors treating domestic wastewater have shown that inoculation can be
neglected in case of domestic wastewater treatment (Passig et al., 2000). Many reactors
started up without being inoculated at all, either at pilot scale (Barber and Stuckey, 1999)
or full scale (Draaijer et al., 1992; Schellinkhout and Collazos, 1992). Nevertheless, the
start-up period in the case of self-inoculation can last up to six months (Foresti, 2002).
In non-tropical conditions, the treatment of domestic wastewater by the UASB reactors is
still feasible. However, the design needs to be modified to improve the removal and
stabilization of the suspended solids. One approach is to use heated digester in
conjunction with the UASB reactor (Mahmoud et al., 2004). In this approach the excess
suspended solids accumulate in the UASB are brought to the digester to be stabilized.
The stabilized sludge is then returned back to the UASB reactor to improve the
performance of the reactor since the stabilized sludge becomes enriched with the
methanogens. This approach has shown to enable the treatment of domestic wastewater at
56 temperature of 15oC with 67% removal of the total COD using HRT of 6 hours. A COD
balance revealed that only 15% of the influent COD was accounted for by waste sludge.
Another approach is to install an anaerobic filter either before or after the UASB in order
to capture the suspended solids (Elmitwalli et al., 2002; Halalsheh et al., 2005). The
anaerobic filter will have a long SRT to enable the stabilization of the captured suspended
solids. This approach has shown that the treatment of domestic wastewater is feasible at
130C with total COD removal of 70% in a combined system of anaerobic filter with HRT
of 4 hours followed by anaerobic hybrid reactor with HRT of 8 hours. A COD balance
revealed that 60% of the influent COD was converted to methane (Elmitwalli et al.,
2002).
Table 1.5 Treatment performance of the first full scale UASB plants treating domestic
wastewater. COD refers to total COD of the raw wastewater (Adapted from van Haandel
and Lettinga, 1994).
Country
Volume Temperature HRT
m3
○
Influent
COD
Effluent
COD
Removal
COD
C
hrs
mg L-1
mg L-1
%
Colombia
64.0
24-26
4-6
267
110
65.0
Colombia
6600
25.0
5.2
380
150
60-80
Brazil
120.0
23.0
4.7-9
315-265
145
50-70
Brazil
67.5
23.0
7.0
402
130
74.0
Brazil
810.0
30.0
9.7
563
185
67.0
India
1200
20-30
6.0
563
146
74.0
57 1.3.4 Post treatment of anaerobic processes
It is well known that the effluent quality of the existing anaerobic reactors treating
domestic sewage can rarely comply with the discharge standards established by the
environmental agencies (Foresti, 2002; Aiyuk et al., 2006). Besides the remaining
fraction of particulate and soluble organic matters in the effluent, the pathogen removal in
the anaerobic reactors is inadequate. Moreover, reduced inorganic compounds such as
ammonium, sulfide and phosphate present in the effluent of such reactors (Figure 1.6).
The effluent quality of anaerobic reactors cannot be defined strictly since it varies widely
depending on several factors such as local conditions, influent characteristics, reactor
design and operational parameters (Foresti et al., 2006). Published data from literature
reveals that the effluent from anaerobic treatment of domestic sewage normally contains
COD ranges from 100 to 200 mg L-1; total suspended solids ranges from 50 to 100 mg L-1
(Van Haandel and Lettinga, 1994) and ammonia ranges from 30 to 50 mg L-1 (Torres and
Foresti, 2001).
58 Figure 1.6 Input and output of compounds in an anaerobic reactor treating domestic
sewage.
Taking into consideration the intrinsic limitation of anaerobic reactors treating domestic
sewage, it is important to include an adequate post treatment method in order to comply
with the permissible limits for discharge. The main role of the post treatment is to further
polish the remaining organic matter, as well as to remove the nutrients and the pathogenic
organisms. Several aerobic treatment processes have been proposed as a post treatment
including polishing ponds, trickling filter systems, activated sludge systems, submerged
59 aerated filters and rotating biological contactors (Goncalves et al., 1998; Chernicharo and
Nascimento, 2001; von Sperling et al., 2001; Cavalcanti, 2003; Tawfik et al., 2004). In
Brazil, most of the domestic sewage treatment plants include a combined system of
UASB reactor followed by submerged aerated filters, activated sludge or stabilization
ponds (Foresti et al., 2006). The choice of the post treatment method depends on the
characteristics of the anaerobic effluent and on the standards set by environmental
agencies for direct discharge to the receiving water or reuse of the treated effluent
(Kujawa-Roeleveld and Zeeman, 2006).
Aerobic and anaerobic treatments of domestic sewage offer several advantages and
disadvantages relative to each other. They have several drawbacks. Firstly, the treatment
costs are high due mostly to high energy inputs required for aeration. Secondly,
approximately half of the COD is converted to biosolids. These biosolids need to be
treated and disposed of safely at a great expense. The anaerobic treatment processes offer
several advantages over the aerobic processes. McCarty (1964) listed these advantages
as i) high degree of waste stabilization, ii) low production of waste biological sludge, iii)
low nutrient requirement, iv) no oxygen requirement, and v) production of methane as a
useful end product.
Thus, by combining both treatment processes in series, the
disadvantages of one process are offset by the advantages of the other process.
The characteristics of the anaerobic effluent are quite different than the original influent
sewage. The effluent has low BOD to COD ratio (poor biodegradability) and high
ammonium (Metcalf and Eddy, 2003). Thus, when the aim of the post treatment besides
60 polishing the organic matter is good nitrogen removal via nitrification and denitrification
processes, the anaerobic treatment should be used to treat only a part of the influent
sewage (about 50-70%) and the remaining should be directed towards the post treatment.
In this way, there is no need for adding external carbon source to support the
denitrification step. The great advantage of using anaerobic treatment is this case is to
receive and stabilize the sludges from the post treatment; therefore, eliminating the need
for anaerobic sludge digester (Chernicharo, 2006).
61 1.4 Fundamentals of Biological Nitrogen Removal
Domestic wastewater contains organic-nitrogen compounds and ammonium nitrogen
ions. Nitrogen in domestic wastewater originates from protein metabolism in the human
body. In domestic wastewater, about 60% of the nitrogen is in organic form (e.g. amino
acids, proteins and urea) and the rest of the nitrogen exists in inorganic form (e.g.
ammonium ions) (Gerardi, 2003). The concentrations of organic nitrogen and ammonium
in domestic wastewater are 8-15 and 12-25 mg N L-1 (Metcalf and Eddy, 2003).
Inappropriate discharge of nitrogenous compounds has adverse effects on the aquatic
systems such as eutrophication, toxicity to aquatic organisms and depletion of dissolved
oxygen in receiving water bodies (Klees and Silverstein, 1992). Thus, the removal of
such compounds is of particular importance. Several biological methods have been
applied for nitrogen removal such as nitrification-denitrification and anaerobic
ammonium oxidation (anammox).
1.4.1 Nitrification
Nitrification is the biological conversion of ammonium (NH4+) to nitrite, and eventually
to nitrate (NO3-) under oxic conditions. This is achieved in two sequential oxidative
stages. The first stage is the ammonia oxidation which involves the conversion of
ammonia to nitrite by ammonium oxidizing bacteria (AOB) of the genus Nitrosomonas
(Eq.1). However, several other AOB genera such as Nitrosococcus, Nitrosopira,
62 Nitrosovibrio, and Nitrosolobus are also able to oxidize ammonia to nitrite (Bock et al.,
1986; Burrell et al., 1999). The second stage is the nitrite oxidation which involves the
conversion of nitrite to nitrate by nitrite oxidizing bacteria (NOB) such as Nitrospira,
Nitrospina, Nitrococcus, and Nitrocystic (Eq. 2). However, the most famous NOB genus
is Nitrobacter (Abeliovich, 1992). Both nitrifying bacteria are chemolithoautotrophs that
use ammonia or nitrite as an energy source, molecular oxygen as an electron acceptor,
and carbon dioxide as a carbon source (Rittmann and McCarty, 2001). Nitrifiers have a
low growth rate with a cell yield (Y) of 0.15 g biomass g-1 NH3-N for AOB and 0.02 g
biomass g-1 NO2- -N for NOB (Wong et al., 2003). In practice, the kinetic of the
nitrification process is limited by AOB and the nitrite is rapidly converted to nitrate
(Metcalf and Eddy, 2003).
NH4+ + 1.5 O2 → NO2- + 2 H+ + H2O
[Eq. 1]
NO2- + 0.5 O2 → NO3-
[Eq. 2]
The performance of the nitrification process is affected by a number of environmental
factors such as dissolved oxygen concentration (DO), temperature, influent ammonium
and nitrite concentrations and pH. Studies show that at low DO concentration (<0.5 mg
L-1) the nitrification rate is greatly inhibited and a minimum DO concentration of 2.0 mg
L-1 is required for a complete nitrification to take place (Reddy, 1998). The low DO
inhibition effect has been found to be greater for NOB than AOB since NOB has a lower
affinity to DO than AOB (Wang and Yang, 2004). In such case, partial nitrification takes
place with increased nitrite concentration in the effluent. Nitrifying bacteria are very
63 sensitive to temperature and the optimum nitrification rate is only achieved at
temperature higher than 15○C (Weon et al., 2004). However, a stable nitrification can be
maintained at 5○C or lower as long as the SRT is high enough in order for nitrifyers to
grow (Grady et al., 1999). Nitrification is inhibited by the free ammonia (NH3) and unionized nitrous acid (HNO2) (Cheng et al., 1997). The inhibition effects depend on the
total nitrogen species concentration, temperature and pH. It was found that at 20oC and
pH 7.0, the ammonium nitrogen concentrations of 100.0 and 20.0 mg N L-1 inhibits AOB
and NOB, respectively, and nitrite nitrogen concentrations of 280.0 mg N L-1 inhibits
NOB (U.S.EPA, 1993).As for most microorganisms, the nitrifying bacteria are sensitive
to pH and the nitrification rate decline significantly at pH below 6.8. The optimum pH for
the nitrification process is in the range of 7.5 to 8.0 (Skadsen, 2002).The nitrification
process consumes about 7.1 g of alkalinity (as CaCO3) per g of nitrogen removed. So, for
the domestic wastewater that generally contains 30.0 to 40.0 mg L-1 of nitrogen, almost
300.0 mg L-1 of alkalinity (as CaCO3) can be destroyed. Thus, the alkalinity needs to be
added for wastewaters that do not have enough alkalinity to maintain acceptable range of
pH that support nitrification processes (U.S.EPA, 1993).
64 1.4.2 Denitrification
Denitrification is the biological reduction of oxidized nitrogen compounds such as nitrate
or nitrite to nitrogen gas under anoxic conditions (Chiemchaisri et al., 1992). The process
is carried out by various chemoorganotrophs and lithoautotrophs that can utilize organic
(e.g. methanol and acetic acid) and inorganic (e.g. reduced sulfur and hydrogen) electron
donors, respectively (Wong et al., 2003). All denitrifiers are facultative aerobes that can
shift to nitrate and nitrite respiration under low DO conditions (Wong et al., 2003).
Typical denitrifiers include the species Pseudomonas, Alcaligenes, Thiobacillus and
Paracoccus (Zumft, 1992). The growth yield of heterotrophic denitrifiers is about 0.4 g
Biomass g-1 COD which is much higher than nitrifiers (Y= 0.15 g biomass g-1 NH3-N)
(Metcalf and Eddy, 2003). Denitrifiers are commonly found in soils, sediments, surface
waters, ground waters, and wastewater treatment plants due to their metabolic diversity
(Rittmann and McCarty, 2001).
Denitrification process takes place in a stepwise manner in which nitrate is sequentially
reduced to nitrite, nitric oxide (NO), nitrous oxide (N2O) and ultimately to nitrogen gas
(N2). The half-reaction of each step and the enzyme catalyzing it are shown in Table 1.6.
One point of concern in the denitrification process is the release of nitric oxide and
nitrous oxide from the treatment process into the environment since they are green house
gases (Vonschulthess et al., 1994; Czepiel et al., 1995). It is clear that low concentrations
of electron donor limit the supply of electron to derive the reductive half reactions, thus,
these intermediates can accumulate (Rittmann and McCarty, 2001) . Also, high DO
65 concentrations repress the nitrite and nitric oxide reductases before the nitrate reductase is
repressed (Korner and Zumft, 1989; Otte et al., 1996; van Benthum et al., 1998).
Table 1.6 Half-reactions and enzymes involved in denitrification process (Adapted from
Rittmann and McCarty, 2001).
Half-reaction
Enzyme name
NO3- + 2 e- + 2 H+ Æ NO2- + H2O
Nitrate Reductase
NO2- + e- + 2 H+ Æ NO + H2O
Nitrite Reductase
2 NO + 2 e- + 2 H+ Æ N2O + H2O
Nitric Oxide Reductase
N2O + 2 e- + 2 H+ Æ N2 (g) + H2O
Nitrous Oxide Reductase
1.4.3
Anaerobic ammonium oxidation (Anammox)
In theory, the ammonium can be used as an electron donor for the denitrification process.
The free energy associated with this reaction (Table 1.7, Eq.3) is nearly as favorable as
the aerobic nitrification (Table1.7, Eq. 4). Based on these thermodynamic calculations,
Broda (1977) postulated the existence of two chemolithoautotrophic microorganisms
capable of oxidizing ammonium to dinitrogen gas with nitrate two decades ago. The
existence of these microorganisms was not demonstrated at that time. The microbial
process of the anaerobic ammonium oxidation (Anammox) is the missing link in the
nitrogen cycle and was recently discovered in a denitrifying reactor in Delft, The
66 Netherlands (Mulder et al., 1995). The Anammox process involves the oxidation of
ammonium with nitrite as electron acceptor to produce dinitrogen gas under anoxic
conditions (Eq. 3) (Strous et al., 1998). The main product of the Anammox reaction is
dinitrogen gas, but about 10% of the influent nitrogen (NH4+ and NO2-) is converted to
nitrate. Van de Graaf (1996) found that the overall nitrogen balance in fluidized bed
reactor gave an ammonium to nitrite conversion ratio of 1:1.31 ± 0.06 and nitrite
conversion to nitrate production ratio of 1:0.22 ± 0.022. The excess 0.31 mole of nitrite is
anaerobically converted to nitrate.
NH4+ + 1.32 NO2- + 0.066 HCO3- + 0.13 H+ → 1.02 N2 + 0.26 NO3- + 0.066 CH2O0.5N0.15 + 2.03 H2O
[Eq. 3]
The Anammox reaction is carried out by the Planctomycetes genus Candidats “Brocadia
anammoxidans” and “Kuenenia stuttgartiensis” as well as species of “Scalindua”
(Schmid et al., 2003). The Anammox bacteria have a slow growth rate with a doubling
time of 10 to 12 days (van der Star et al., 2007). Despite their slow growth rate, their
specific activity is high (> 1.0 g N g-1 biomass d-1) (Strous et al., 1998). Anammox
bacteria are very sensitive to the oxygen and nitrite. It was found that oxygen
concentrations higher than 0.6 mg L-1 and nitrite concentrations of 150 mg NO2- -N L-1
inhibit the Anammox bacteria completely but reversibly (Jetten et al., 1998; Jetten et al.,
2001). They are also sensitive to the presence of some organic compounds. For example,
Anammox bacteria are very susceptible to alcohol especially methanol which can cause
complete and irreversible inhibition at concentrations as low as 40 mg L-1 (Guven et al.,
67 2005). This fact is of a particular importance very since methanol is added as an
inexpensive carbon source in post-denitrification (Metcalf and Eddy, 2003). On the other
hand, the organic acids are not inhibitory to the Anammox bacteria. Propionate and
acetate were found to be substrate to the Anammox bacteria. Propionate was oxidized by
the Anammox bacteria with the nitrite and or nitrite as an electron acceptor with
simultaneous anaerobic oxidation of ammonium (Guven et al., 2005).
Anammox process is a novel, promising and cost-effective alternative to the conventional
nitrification-denitrification processes for nitrogen removal; particularly, for wastewater
with deficiency in organic matter (Vandegraaf et al., 1995). The process can reduce the
cost of aeration and additional of organic matter as is required for the conventional
nitrification-denitrification process. Moreover, the low yield of the Anammox bacteria
(0.07 ± 0.01 mol C mol-1 NH4+-N) results in low sludge production (Liao et al., 2008). As
a consequence, the Anammox process can save up to 90% of the operational cost as
compared to the conventional nitrification-denitrification processes (Jetten et al., 2001).
There are two main ways for obtaining high nitrogen removal rate through Anammox
process. These are two -reactor systems and one-reactor systems.
68 A) Two-Reactor systems
Anammox reaction requires an influent containing an almost equal amount of ammonium
and nitrite. Thus, a first partial nitrification step is required in which the complete
nitrification is avoided and around 50% of the ammonium is converted to nitrite. There
are several strategies can be applied to achieve the partial nitrification such as
manipulating the temperature, hydraulic retention time and temperature using SHARON
process (Single reactor System for High Ammonium Removal Over Nitrite), inhibition of
NOB by free ammonia (Anthonisen et al., 1976) or by lowering the DO concentration to
avoid the oxidation of NO2- to NO3- due to the higher affinity of AOB to DO than NOB
(Gerardi, 2003). Once the adequate mixture of ammonium and nitrite is obtained, the
effluent of the first reactor is fed to the second reactor where the Anammox process takes
place. Different reactor configurations have been used on a laboratory and pilot scale
such as fixed bed reactor, fluidized bed reactor, sequencing batch reactor and UASB
reactor (Jetten et al., 1997; Strous et al., 1997; van Dongen et al., 2001; Sliekers et al.,
2003; Ahn et al., 2004; Fux et al., 2004; Schmidt et al., 2004). Among these studies, the
highest removal rate (8.9 Kg N m3 d) was reported by Sliekers et al. (2003) on a
laboratory and pilot scale using a gas lift reactor with granular sludge.
B) One-Reactor systems
In this system both partial nitrification and Anammox occur in the same reactor. This can
be achieved by using a biofilm system in which classical nitrification is developed by the
69 AOB in the outer aerobic layer and anaerobic oxidation takes place in the deeper zones of
the biofilm under low DO concentrations (Helmer-Madhok et al., 2002; Egli et al., 2003).
The process was developed by different research groups and has been given different
names: “Aerobic/Anoxic Deammonification”, “Onland” and “CANON” (Fux and
Siegrist, 2004). Rotating biological contactors and moving bed systems have been used
for landfill leachate, sludge liquor and synthetic wastewater. The applied nitrogen loading
rate ranges between 1.4 and 4.8 Kg N m-2 d-1. The highest nitrogen removal rate was
reported to be higher than 60% (Siegrist et al., 1998; Hippen et al., 2001; Pynaert et al.,
2002).
Table 1.7 Gibbs free energy of several reactions involved in dentrification and
ammonium oxidation (Adapted from Mulder et al., 1995).
Equation no.
Reaction
ΔG0’
1
2 NO3- + 5 H2 + 2 H+ Æ N2 + 6 H2O
-1121 KJ/reaction
2
8 NO3- + 5 HS- + 3 H+ Æ 4 N2 + 5 SO42- + 4
H2O
-3721 KJ/reaction
3
3 NO3- + 5 NH4+ Æ 4 N2 + 9 H2O + 2 H+
-297 KJ/mol NH4+
4
5 NH4+ + 2 O2 Æ NO3- + H2O + 2 H+
-349 KJ/mol NH4+
5
8 NH4+ + 6 O2 Æ 4 N2 + 12 H2O + 8 H+
-315 KJ/mol NH4+
70 It seems from literature review that a sustainable approach towards domestic wastewater
treatment is to use anaerobic high rate reactor followed by aerobic post treatment. Most
of the organics will be removed in anaerobic reactor. The remaining of organics will be
polished in the aerobic post treatment. For nitrogen nutrient removal, the post aerobic
treatment will operate under low dissolved oxygen conditions to promote the partial
nitrification. The produced nitrite can be either re-circulated back to the anaerobic reactor
for complete nitrogen removal via denitrification and anammox processes or can be
directed towards an anammox reactor for complete nitrogen removal.
71 1.5
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84 CHAPTER 2
METHANOGENIC INHIBITION BY INORGANIC FLUORIDE
2.1 Abstract
The inhibitory effect of fluoride toward acetoclastic and hydrogenotrophic methanogens
was investigated. In this study, two mesophilic sludges (granular and flocculent) and one
thermophilic sludge were tested. The results indicate that the sensitivity to fluoride varied
depending on methanogenic population examined. Acetoclastic methanogens present in
mesophilic and thermophilic sludges were sensitive to fluoride. The concentration of
fluoride causing 50.0% inhibition (IC50) to acetoclastic methanogenic activity in
mesophilic granular sludge, mesophilic flocculent sludge, and thermophilic flocculent
sludge was 155.7, 34.5 and 18.1 mg L-1, respectively. The toxicity of fluoride to
mesophilic acetoclastic methanogens increased with time of exposure such that in a
second acetate feeding the IC50 values were 29.9, and 25.8 mg L-1, respectively; for
granular and flocculent sludge. The high initial IC50 value for granular sludge may have
been due to initial attenuation of fluoride by calcite present in granule. In contrast, the
hydrogenotrophic methanogens in mesophilic and thermophilic sludges tolerated
relatively high concentration of fluoride. The IC50 values for hydrogenotrophic
methanogens were higher than 400.0 mg L-1. The toxicity of fluoride to both mesophilic
85 and thermophilic hydrogenotrophic methanogens decreased with increasing time of
exposure. The results of this study indicate that untreated industrial effluent containing
fluoride can potentially be toxic to both acetoclastic and hydrogenotrophic methanogens.
86 2.2
Introduction
There is a growing concern regarding the fluoride as an environmental contaminant in
various parts of the world (Prakasini et al., 2003; Miretzky et al., 2008). The presence of
fluoride in water has been the focus of research for several years. Community water
fluoridation has been known to be effective in preventing the dental caries and has been
used in US for the past 50 years (Browne et al., 2005). The Centers of Disease Control
and Prevention (CDC) regarded the fluoridation as one of the major public health
measures of the last century (CDC, 1999). The allowable concentration of fluoride in
drinking water is 1.5 mg/L (WHO, 2004). Despite its beneficial effects, fluoride can be
detrimental to human health at concentration exceeding 1.5-2.0 mg L-1. Fluoride at
concentration above the allowable concentration is known to cause dental and/or skeletal
fluorosis (WHO, 2004). For example, in India there are 25 million people in 15 states that
are affected by fluorosis (Naoki et al., 1996).
Unpolluted surface water generally contains less than 0.2 mg L-1 fluoride (Dave, 1984).
However, high concentration of fluoride as high as 67.0 mg L-1 was reported in some
natural waters (Faust and Aly, 1987). Two major sources contribute to such high elevated
concentration of fluoride through natural and human activity. In nature, fluoride exists
through primary and secondary sources. Primary sources of fluoride include geochemical
characteristics of the earth’s crust that result from volcanic and plutonic activities. This
87 primary source is the reason why some regions in the world contain high fluoride
concentration. The secondary source of fluoride includes water and flood that derive
fluoride from the earth’s crust (Mariappan and Vasudevan). For example, hot springs and
geysers of Yellowstone National Park contain 25.0 to 50.0 mg L-1 of fluoride (Neuhold
and Sigler, 1960). Several fluoride compounds have been extensively used in various
industries such as semiconductor, metal processing, fertilizers, and glass manufacturing
(Buffle et al., 1985; Parthasarathy et al., 1986; Huang et al., 1999; Chuang et al., 2002).
The untreated effluent of such industries contains high concentration of fluoride which
also contributes to fluoride pollution in receiving water body. For instance, typical
wastewater from the semiconductor in Taiwan contains 1,000-3,500 mg L-1 of fluoride
(Chou et al., 1994).
The antimicrobial effects of fluoride on oral bacteria and dental plaque are well
documented in literature. Fluoride can interfere with the initial adhesion (Mannerberg,
1968; Rolla, 1977), aggregation (Svatun et al., 1977) and affect bacterial metabolism
(Hamilton, 1977). Fluoride is known to inhibit the activity of various microbial enzymes
such as glycolytic enolase and membrane-associated proton-pumping ATPase (Marquis
et al., 2003; Sutton et al., 1987). Several mechanisims have been proposed on the
inhibition action of fluoride to microbial cells. Fluoride or HF can bind directly to many
enzymes such as heme-containing enzymes and other metalloenzymes therefore affecting
metabolism. In addition, fluoride can form complexes with metals such as aluminum and
beryllium and such complexes can mimic phosphate which adversely affects variety of
88 enzymes and regulatory phosphatases (Marquis et al., 2003). Furthermore, fluoride can
diminish ΔpH across the cell membrane by acting as a trans-membrane proton conductor
in the acidic environment. Fluoride is a weak acid (pKa = 3.15) and its undissociated
form (i.e., HF) can bring protons across cell membrane leading to acidification of
cytoplasm of the cell (Eisenberg and Marquis, 1980; Iwami et al., 1995; Marquis et al.,
2003). In one study, it was shown that the sensitivity of a range of oral bacteria to
fluoride increased 3 to 50 fold when the minimum inhibition concentration (MIC) was
determined at pH of 5 rather than 7 (Bradshow et al., 1990).
Anaerobic bioreactors are an alternative wastewater treatment technology in which both
pollution control and energy recovery can be achieved. They have been widely used for
the stabilization of municipal wastewater sludges and municipal solid wastes (Rittmann
and McCarty, 2001). In addition, anaerobic digestion of municipal wastewater sludges
results in inactivation of pathogens that is usually required prior to the ultimate disposal
of biosolids (Grady et al., 1999). Anaerobic process involves the decomposition of
organic matter in a series of microbial events that results in the production of methane,
carbon dioxide and new bacterial cells (Geradi et. al., 2003). These events occur in four
stages. The first stage involves the hydrolysis of solids which results in the production of
soluble organic compounds. The second stage is the acidogenesis which involves the
conversion of soluble organic compounds into volatile fatty acids, ethanol, hydrogen and
carbon dioxide. The end products of the acidogenesis are further oxidized to acetate,
hydrogen and carbon dioxide in the third stage which is referred to as acetogenesis. The
final stage is the methanogenesis which involves the conversion of the acetogenesis end
89 products into methane and carbon dioxide (O’Flaherty et al., 2006).
Without the
methanogenesis, the ultimate breakdown of the organic matter would not take place due
to the accumulation of the actogenesis end products (Anderson et al., 2003).Methanogens
are strictly anaerobic Archaea that can be divided into two groups: i) hydrogen utilizing
methanogens which forms methane by the reduction of hydrogen/carbon dioxide and ii)
acetoclastic methanogens which produce methane by the conversion of acetate (Van
Andel and Breure, 1984). Acetoclastic methanogens are considered the most important
methanogenic species since 70% of the methane produced is generated via this pathway
(Lettinga, 1995).
However, anaerobic bioreactor technology can experience operational problems that
result in disruption of its performance and potential failure. Such operational problems
include introduction of inhibitory substances. Methanogenesis is considered to be the
slowest (rate limiting) step in the anaerobic process (Anderson et al., 2003). There is a
lack of information in literature on the inhibitory effect of inorganic fluoride on the
methanogens present in anaerobic digester. Therefore, the objective of this research is to
evaluate the inhibitory effect of inorganic fluoride toward acetoclastic and
hydrogenotrophic methanogens in the anaerobic digesters
90 2.3
2.3.1
Materials and Methods
Chemicals
Sodium fluoride (99%) and sodium acetate were purchased from Sigma Chemical Co.
(St. Louis, MO, USA). All chemicals were used as received.
2.3.2
Sludge sources
Two different mesophilic inocula were used in this investigation including granular
sludge (Eerbeek) and anaerobically digested sludge (Ina Road sludge). Eerbeek sludge
was obtained from industrial anaerobic treatment plants treating recycle paper wastewater
(Industriewater, Eerbeek, The Netherland). The sludge was washed and sieved to remove
fine particles before use in the toxicity assays. Ina Road sludge was obtained from an
anaerobic digester at a local municipal wastewater treatment plant (Ina Road wastewater
treatment plant, Tucson, Arizona). The content of volatile suspended solids (VSS) in
Eerbeck and Ina Road sludge was 12.9% and 1.23%, respectively. A thermophilic
anaerobically digested sludge (Hyperion sludge) was also used in this study. The
Hyperion sludge was obtained from municipal wastewater treatment plant (Hyperion
wastewater treatment plant, Los Angles, California). The content of VSS in Hyperion
sludge was 1.55%.
91 2.3.3
Basal media
The basal mineral medium was prepared using ultrapure water (Milli-Q system;
Millipore) and contained (in mg L-1): KH2PO4 (37); CaCl2•2H2O (10); MgSO4•7H2O
(10); MgCl2•6H2O (78.1); NH4Cl (668.6); NaHCO3 (3003); Yeast Extract (20), and trace
element solution (1.1 ml L-1). The trace element solution contained (in mg L-1): H3BO3
(50); FeCl2•4H2O (2000); ZnCl2 (50); MnCl2•4H2O (50); (NH4)6Mo7O24•4H2O (50);
AlCl3•6H2O
(90);
CoCl2•6H2O
(2000);
NiCl2•6H20
(50);
CuCl2•2H2O
(30);
NaSeO3•5H2O (100); EDTA (1000); resazurin (200) and 36% HCl (1 mL L-1). The final
pH of the basal medium was adjusted to 7.2 with HCl.
2.3.4
Methanogenic toxicity assays
Methanogenic toxicity assays were conducted in glass serum flasks (160 mL) at 30 ± 2oC
(mesophilic assays) and 55 ± 2oC (thermophilic assays) in an orbital shaker (75 rpm).
Eerbeek granular sludge, Ina Road sludge and Hyperion sludge were added at a
concentration of 1.5 g volatile suspended solids (VSS) L-1, 10 % (v/v) and 10% (v/v);
respectively, to serum flasks containing 25 mL basal medium. All flasks were sealed with
butyl rubber stoppers and aluminum crimp seals and their headspace was flushed with
N2/CO2 (80:20, v/v) for 5 minutes to create an anaerobic condition. Both acetate and
hydrogen were used as an electron donor. For acetoclastic assays, the acetate was
supplemented at a final concentration of 2 g COD L-1. For hydrogenotrophic assays, the
H2/CO2 was supplied by pressurizing the flasks to 1 atm (80:20, v/v). The flasks were
92 pre-incubated overnight to ensure that the sludge was adapted to medium conditions. On
the following day, fluoride was added at different concentrations then the headspace of
the flasks for acetoclastic assays was flushed again with N2/CO2 (80:20, v/v). For the
hydrogenotrophic assays, the headspace of the flasks was flushed with N2/CO2 (80:20,
v/v) and then pressurized with H2/CO2 (80:20, v/v) to 1 atm. Then, the flasks were
incubated at least two hours prior to determination of methane content in the headspace
of each flask. The methane content of the headspace of each flask was measured several
times during the first day of the experiment and periodically during the subsequent days
until 80% of the substrate (acetate or H2) was depleted.
In the second feeding, the acetate was supplemented at a final concentration of 2 g COD
L-1 for acetoclastic assays. Then the headspace of the flasks was flushed with N2/CO2
(80:20, v/v) to remove methane from headspace. For the hydrogenotrophic assays, the
headspace of the flasks was flushed with N2/CO2 (80:20, v/v) and then pressurized with
H2/CO2 (80:20, v/v) to 1 atm. the flasks were incubated at least two hours prior to
determination of methane content in the headspace of each flask. The methane content of
the headspace of each flask was measured several times during the first day of the second
feeding and periodically during the subsequent the substrate (acetate or H2) was depleted.
The maximum specific methanogenic activities (mg CH4-COD/ (g VSS d)) were
calculated from the slope of the cumulative methane production (%) versus time (hours)
and the biomass concentration as the mean value of the triplicate assays. The fluoride
93 concentration that causes 20%, 50% and 80% reduction in activity compared to
uninhibited control were referred to as IC20, IC50 and IC80, respectively.
2.3.5
Analytical methods
The methane concentration from the headspace of the serum flasks was determined by
gas chromatography (GC) using an HP5290 series II system (Agilent Technologies, Palo
Alto, CA) equipped with a flame ionization detector (GC-FID). The GC was fitted with a
DB-FFAP column (J&W Scientific, Palo Alto, CA) capillary column. The temperature
of the column, the injector port and the detector was 140, 180 and 250°C, respectively.
The carrier gas was helium at a flow rate of 9.3 mL min-1 and a split flow of 32.4 mL
min-1. Samples for measuring methane content (100 µL) in the headspace were collected
using a pressure-lock gas syringe.
Volatile suspended solids and other analytical parameters were determined according to
Standard Methods for Examination of Water and Wastewater (APHA, 2005).
94 2.4
Results
2.4.1 Toxicity to acetoclastic methanogens
In this study we investigated the inhibitory effects of fluoride to mesophilic acetoclastic
methanogens. The results indicate that fluoride was toxic to acetoclastic methanogens.
The time course of methane production rate with different fluoride concentration for the
mesophilic Eerbeek granular sludge and Ina Rd flocculent sludge are shown in figures
2.1, and 2.2, respectively. The graphs clearly show that the methane production rate
decreases with increasing fluoride concentration. In addition, Figures 2.3, and 2.4
revealed a sharp decrease of the normalized activity of Eerbeek sludge and Ina Rd;
respectively, with the fluoride concentration. The IC50 values for Eerbeek granular sludge
and Ina Rd sludge were 155.7 and 34.5 mg L-1, respectively (Table 2.1). During the first
feeding, acetoclastic methanogens present in Eerbeek sludge appeared to be more
resistant to fluoride toxicity than in Ina Rd sludge.
A second feeding of substrate to the bioassay was supplemented once the initially
supplied substrate was 80.0% removed. The reason behind the second feeding was to
evaluate the inhibitory effect of a longer term exposure to fluoride on acetoclastic
methanogens. The results revealed that the toxicity of fluoride to acetoclastic
methanogens increased as the time of exposure increased. Figures 2.1 and 2.2 showed a
decrease in methane production rate with increasing time of exposure to fluoride. Also,
the decrease in specific activity of acetoclastic methanogens was steeper in the case of the
95 second feeing compared to first feeding. Among the two acetoclastic methanogens tested,
the inhibitory effects of long term exposure to fluoride was similar for Eerbeek sludge
and Ina Rd sludge as can be evidenced by the similar IC50 values in the second feeding
(Table 2.1), ranging from 25.0 to 30.0 mg L-1.
14
CH 4 production (%)
12
10
8
6
4
2
0
0
50
100
150
200
250
300
350
Time (hrs)
Figure 2.1 Time curse of cumulative methane production for mesophilic acetoclastic
methanogens present in Eerbeck sludge in the presence of increasing fluoride
concentrations (in mg L-1): (♦) 0, (○) 20, (Δ) 100, (▲) 250, (●) 500. Dashed line
represents the starting of the second feeding.
96 14
CH 4 production (%)
12
10
8
6
4
2
0
0
100
200
300
400
500
600
Time (hrs)
Figure 2.2 Time curse of cumulative methane production for mesophilic acetoclastic
methanogens present in Ina Rd sludge in the presence of increasing fluoride
concentrations (in mg L-1): (♦) 0, (□) 20, (●) 40, (▲) 100, (●) 250, (◊) 500. Dashed line
represents the starting of the second feeding.
97 100
Activity (% of Control)
90
80
70
60
50
40
30
20
10
0
0
100
200
300
400
500
600
Fluoride Concentration (mg/L)
Figure 2.3 Inhibitory effect of fluoride on specific activity of mesophilic acetoclastic
methanogens present in Eerbeek : (♦) first feeding, (□) second feeding.
98 100
Activity (% of Control)
90
80
70
60
50
40
30
20
10
0
0
100
200
300
400
500
600
Fluoride Concentration (mg/L)
Figure 2.4 Inhibitory effect of fluoride on specific activity of mesophilic acetoclastic
methanogens present in Ina Rd sludge: (♦) first feeding, (○) second feeding.
99 Table 2.1 Inhibitory effect of fluoride on the acetoclastic and hydrogenotrophic
methanogens. IC20, IC50 and IC80 are the fluoride concentration causing 20%, 50%, and
80% inhibition in activity of methanogens, respectively.
1st Feeding
Inoculum
2nd Feeding
Substrate
IC20
IC50
IC80
IC20
IC50
IC80
-------------------------------mg L-1------------------------------Mesophilic Methanogens
Eerbeek
Acetate
38.5
155.7
>500.0
15.0
29.9
58.1
Eerbeek
H2
814.3
>814.3
>814.3
520.0
645.0
795.0
Ina Rd
Acetate
17.5
34.5
93.0
12.4
25.8
35.7
Ina Rd
H2
31.1
77.8
396.2
806.3
1006.3
1127.6
Hyperion
Acetate
7.2
18.1
39.2
29.5
42.6
62.9
Hyperion
H2
218.6
432.6
>600.0
287.2
>600.0
>600.0
Thermophilic
Methanogens
For the sake of comparison, the inhibitory effect of fluoride to thermophilic Hyperion
sludge was studied. Fluoride was also found to be toxic to Hyperion sludge. Figure 2.5
shows a decrease in the methane production rate with the fluoride concentrations.
Moreover, the normalized activity of Hyperion acetoclastic methanogens was decreased
with the fluoride concentrations (Figure 2.6). Unlike the mesophilic sludges, the long
term exposure to fluoride did not result in an increase in the inhibitory effects of fluoride
on the Hyperion acetoclastic methanogens. Figure 2.5 shows that the methane production
100 rate increased with increasing time of exposure. Also, the specific activity of acetoclastic
in Hyperion was higher in the second feeding than in the first feeding (Figure 2.6). The
IC50 values calculated in the first and second feeding were 18.1 and 42.6 mg L-1 (Table
2.1). The increase in the activity may suggest that the acetoclastic methanogens present in
thermophilic sludge might have developed some adaptation to fluoride toxicity or a more
fluoride tolerant strain of acetoclastic methanogens became enriched as time of exposure
increased.
101 35
CH 4 production (%)
30
25
20
15
10
5
0
0
100
200
300
400
500
600
Time(hrs)
Figure 2.5 Time curse of cumulative methane production for thermophilic acetoclastic
methanogens present in Hyperion sludge in the presence of increasing fluoride
concentrations (in mg L-1): (♦) 0, (□) 25, (Δ) 100, (◊), 150, (●) 300. Dashed line
represents the starting of the second feeding.
102 100
Activity (% of Control)
90
80
70
60
50
40
30
20
10
0
0
50
100
150
200
250
300
350
Fluoride Concentration (mg/L)
Figure 2.6 Inhibitory effect of fluoride on specific activity of thermophilic acetoclastic
methanogens present in Hyperion sludge: (○) first feeding, (▄) second feeding.
103 2.4.2 Toxicity to Hydrogenotrophic methanogens
The inhibitory effect of fluoride to mesophilic hydrogenotrophic methanogens was also
investigated. The results indicate that fluoride was toxic to mesophilic H2-utilizing
methanogens at only relatively high fluoride concentrations. Figures 2.7 and 2.8 show the
time course of the methane production rate with different fluoride concentrations for the
mesophilic H2-utilizing methanogens in Eerbeek granular sludge and Ina Rd
anaerobically digested sludge, respectively. The graphs clearly show that fluoride
resulted in a decrease in the methane production rate only when it was present at high
concentrations. Moreover, only high concentrations of fluoride caused a sharp decrease in
the normalized activity of H2-utilizing methanogens (Figures 2.9 and 2.10). However, the
H2-utilizing methanogens present in mesophilic Ina Rd were more sensitive to inhibitory
effect of fluoride than the ones present in Eerbeek sludge as demonstrated by the IC20
values of 31.3 and 814.3 mg L-1, respectively (Table 2.1).
The inhibitory effect of a long term exposure to fluoride on hydrogenotrophic
methanogens present in the mesophilic sludges was evaluated by supplementing a second
feeding to bioassays. Unlike the acetoclastic methanogens, the H2-utilizing methanogens
were more tolerant to high fluoride concentration during the second feeding (Figures 2.9
and 2.10). For example, The IC50 values for the H2-utilizing methanogens present in Ina
Rd during the first and second feeding were 77.8 and 1006.3 mg L-1, respectively. This
indicates that the inhibitory effect of fluoride was decreasing with the exposure time.
104 18
CH 4 production (%)
16
14
12
10
8
6
4
2
0
0
50
100
150
200
250
Time (hrs)
Figure 2.7 Time curse of cumulative methane production for mesophilic
hydrogenotrophic methanogens present in Eerbeck sludge in the presence of increasing
fluoride concentrations (in mg L-1): (♦) 0, (□) 135.7, (●) 542.9, (Δ) 814.3. Dashed line
represents the starting of the second feeding.
105 30
CH4 production (%)
25
20
15
10
5
0
0
100
200
300
400
500
600
700
Time (hrs)
Figure 2.8 Time curse of cumulative methane production for mesophilic
hydrogenotrophic methanogens present in Ina Rd sludge in the presence of increasing
fluoride concentrations (in mg L-1): (♦) 0, (□) 40, (▲) 900, (●) 1500. Dashed line
represents the starting of the second feeding.
106 140
Activity (% of control)
120
100
80
60
40
20
0
0
200
400
600
800
1000
Fluoride concentration (ppm)
Figure 2.9 Inhibitory effect of fluoride on specific activity of mesophilic
hydrogenotrophic methanogens present in Eerbeck : (♦) first feeding, (○) second feeding.
107 100
Activity(% of Control)
90
80
70
60
50
40
30
20
10
0
0
200
400
600
800
1000
1200
1400
1600
Fluoride Concentration (mg/L)
Figure 2.10 Inhibitory effect of fluoride on specific activity of mesophilic
hydrogenotrophic methanogens present in Ina Rd sludge: (♦) first feeding, (○) second
feeding.
108 The behavior of the H2-utilizing methanogens present in thermophilic sludge (Hyperion
anaerobically digested sludge) was similar to H2-utilizing methanogens present in
mesophilic sludge. Only high fluoride concentration caused a decrease in the rate of
methane production during the first feeding (Figure 2.11). The presence of high fluoride
concentration of 600.0 mg L-1 resulted in sharp decrease in normalized activity of H2utilizing methanogens in the Hyperion sludge compared to low fluoride concentration of
75.0 mg L-1 during the first feeding (Figure 2.12). The toxicity of fluoride to H2-utilizing
methanogens in the Hyperion sludge appeared to decrease with the prolonged exposure to
fluoride. For instance, the IC50 values determined during the first and second feedings
were 432.6 and > 600.0 mg L-1, respectively (Table 2.1).
109 30
CH4 production (%)
25
20
15
10
5
0
0
20
40
60
80
100
120
140
160
Time (hrs)
Figure 2.11 Time curse of cumulative methane production for thermophilic
hydrogenotrophic methanogens present in Hyperion sludge in the presence of increasing
fluoride concentrations (in mg L-1): (♦) 0, (Δ) 75, (●) 300, (□) 600. Dashed line represents
the starting of the second feeding.
110 Activity(% of Control)
100
80
60
40
20
0
0
100
200
300
400
500
600
700
Fluoride Concentration (mg/L)
Figure 2.12 Inhibitory effect of fluoride on specific activity of thermophilic
hydrogenotrophic methanogens present in Hyperion sludge: (○) first feeding, (▪) second
feeding.
111 2.5
Discussion
The results of this study indicate that acetoclastic methanogens were more susceptible to
inhibitory effects of fluoride than hydrogenotrophic methanogens. This was evident by
the decreased rate of methane production resulting from exposures to relatively low
fluoride concentrations. The observed low inhibitory fluoride concentrations to
acetoclastic methanogens are of particular importance. Maintaining an active acetoclastic
population in anaerobic digester is critical for stable performance. Without the
acetoclastic methanogens, the ultimate breakdown of organic material in anaerobic
digester would not take place. Acetoclastic methanogens are relatively slow-growing and
require long time to replicate themselves to levels to achieve stable operation of digester.
The doubling times of Methanosarcina spp. and Methanosaeta spp are 24 hours and 3.59 days, respectively (Anderson et al., 2003).This means that any upset to anaerobic
digester due to toxicity of fluoride would result in a long recovery time.
The toxicity mechanism of fluoride toward methanogens is not well understood. Fluoride
is suspected of inhibiting certain inorganic pyrophosphatase (PPase) enzymes. Soluble
PPase is found in almost all living cells (Young et al., 1998) including methanogens
(Jetten et al, 1992). This enzyme is essential for life (Parfenyev et al., 2001) since it plays
a crucial role in energy metabolism of various organisms. PPase catalyses the hydrolysis
of inorganic pyrophosphate (PPi), formed as a by-product of numerous metabolic
reactions, to inorganic phosphate (Pi) (Chen et al., 1990; Lundin et al., 1991). The PPase
112 enzymes are grouped into three classes designated types A, B and C. The three types of
PPase differ in their sensitivity to fluoride ion. Classes A and B are known to be strongly
inhibited by fluoride ions (Marquis et al., 2003). Roth and Bachofem (1994) were able to
isolate PPase enzyme from the methanogen, Methanobacterium thermoautotrophicum.
Fluoride at concentrations of 1 and 10 mM was found to inhibit the activity of isolated
PPase by over 95.0%. On the other hand, Type C PPase is not inhibited by fluoride
(Marquis et al., 2003). Jetten et al. (1992) isolated PPase from another methanogen,
Methanothrix soehngenii. The activity of PPase enzyme was not inhibited by fluoride. In
our study, the different sensitivities of acetoclastic and hydrogenotrophic methanogens to
fluoride could potentially be explained by the variation of sensitivity of different PPase
enzyme types to fluoride. Another explanation could also be the exposure of PPase
enzymes to manganese (Mn+2) and/or cobalt (Co+2) that are present in trace element
solution that was added to mineral medium. Mn+2 ions reduce the suppressive effect of
fluoride on the PPase enzymes that are sensitive to fluoride (Parfenyev et al., 2001). In
one study, prior exposure of PPase enzyme isolated from the methanogen,
Methanocaccus jannaschii, to Mn+2 or Co+2 ions provided a protection against inhibition
by fluoride while the enzyme without prior exposure is very sensitive (Kuhn et al., 2000).
While there is little known about fluoride toxicity to methanogens, the toxicity of organic
fluoride compounds, such as methyl fluoride (CH3F) is well documented. Methyl fluoride
is an inhibitor that is selective against acetoclastic methanogens (Conrad and Klose,
1999; Janssen and Frenzel, 1997). Studies with pure cultures of different methanogens
showed that acetoclastic methanogens were inhibited at much lower concentration of
113 CH3F than hydrogenotrophic methanogens (Janssen and Frenzel, 1997). Methanosarcina
barkeri, a methanogen that can grow on either acetate or hydrogen, was inhibited by
CH3F only when it was growing on acetate (Janssen and Frenzel, 1997). The inhibitory
effect of CH3F toward acetoclastic and hydrogenotrophic methanogens are in agreement
with our observation in this study that acetoclastic methanogens were more susceptible to
inorganic fluoride at low concentration than hydrogenotrophic methanogens.
Anaerobic granular sludges developed in anaerobic bioreactors such as up-flow anaerobic
sludge blanket (UASB) are a special type of biofilms (Gonzalez-Gil et al., 2001). These
spherical biofilms are formed by auto- immobilization of anaerobic bacteria (Lettinga,
1995). It is believed that multivalent positive ions especially calcium plays an important
role in the granulation process (Schmidt and Ahring, 1996). Extensive accumulation of
calcite has been found in anaerobic granules (Wang et. al., 2007). Calcite can accumulate
either inside or on the surface of the granules (Dofling et. al., 1985). Therefore, a possible
explanation for increasing the sensitivity of Eerbeek granular sludge to fluoride with
increasing exposure time could be due to the adsorption of fluoride to calcite. This
adsorption capacity eventually gets saturated after which the fluoride toxicity increases.
The tolerance of hydrogenotrophic methanoges in Ina Road to fluoride increased with the
exposure time. The IC50 values for the first and second feedings were 77.8 and 1006.3 mg
L-1, respectively. A possible explanation to such increase in tolerance to fluoride could be
due to the physiological adaptation of hydrogenotrophic methanogens to fluoride.
Another explanation could also be the long exposure to fluoride resulted in a shift in
114 microbial population which resulted in the dominance of hydrogenotrophic methanogens
that were less sensitive to fluoride. This is consistent with the different tolerances of
PPases to fluoride. The concentrations of fluoride in untreated industrial wastewater effluent vary widely. A
large amount of fluoride-containing wastewater is generated in rinsing and cleansing
operations from semiconductor industry (Chuang et al., 2002). Fluoride concentrations as
high as 500.0-2000.0 mg L-1 have been reported in the effluent from semiconductor
operations in Taiwan (Huang et al., 1999; Hu et al., 2005). The commonly used method
for the removal of fluoride in such wastewaters is the physical-chemical processes that
involve the precipitation of fluoride as calcium fluoride (CaF2) through the addition of
calcium salt (Fan et al., 2003). However, the precipitation of CaF2 only reduces the
fluoride concentration to around 20.0 mg L-1 due to the solubility of CaF2 (i.e., solubility
of CaF2 is 17.0 mg L-1 at 25oC in water) (Toyoda and Taira, 2000; Shen et al., 2003).
Comparison of the fluoride concentration in the effluent of semiconductor industry after
physical-chemical treatment with the IC50 values for the methanogens determined in this
study suggests that only the thermophilic acetoclastic methanogens (IC50 value of 18.1
mg L-1) could be potentially affected by fluoride level in the pretreated semiconductor
effluents. Nonetheless, the fluoride concentrations in municipal wastewaters receiving
wastewater-containing fluoride are expected to be lower than 20.0 mg L-1 due to the
dilution with residential and other industrial sources.
115 2.6
Conclusions
Fluoride imposes inhibitory effect on acetoclastic and hydrogenotrophic methanogens in
both
mesophilic
and
thermophilic
anaerobic
digesters.
However,
acetoclastic
methonagens are more susceptible to inhibitory effect of fluoride. Only high
concentration of fluoride imposes inhibitory effect on hydrogenotrophic methanogens.
The inhibitory effect of fluoride to acetoclastic increases with the exposure time. On the
other hand, the inhibitory effect of fluoride to hydrogenotrophic bacteria decreases with
the time of exposure. Finally, fluoride concentrations in municipal wastewater that
receives industrial effluents containing fluoride are expected to be low due to the dilution
of such effluents with wastewater from residential and other industrial sources.
116 2.7
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121 CHAPTER 3
Treatment of High Strength Synthetic Sewage in a Laboratory-Scale Upflow
Anaerobic Sludge Bed with Aerobic Activated Sludge Post-Treatment
3.1 Abstract
Feasibility of a combined system consisting of an anaerobic up-flow anaerobic sludge
blanket (UASB) pre-treatment followed by aerobic activated sludge (AS) post-treatment
for the removal of carbonaceous and nitrogenous contaminants at an average temperature
of 25○C was investigated. The combined system was fed with high strength synthetic
sewage having a chemical oxygen demand (COD) of 2,500 mg COD L-1. The organic
loading rate (OLR) of the UASB reactor was increased gradually from 1.1 to 3.8 g COD
Lr-1 d-1 by decreasing the hydraulic retention time (HRT) from 56.0 to 15.4 hours. At
steady state conditions, the UASB reactor achieved removal efficiency up to 83.5 % of
the total COD, 74.0% of volatile fatty acid (VFA) and 94.0% of protein. The combined
system performed an excellent organic removal pushing the overall removal efficiency of
the total COD, VFA and protein to 91.0%, 99.9% and 98.2%, respectively. When the
OLR of the UASB increased to 4.4 g COD Lr-1 d-1, the effluent quality of the UASB
deteriorated (i.e., UASB was overloaded). Despite the effort of decreasing the OLR, the
UASB reactor did not recover completely. At the end of this study, the total COD, VFA
122 and protein removal efficiency of the UASB was 70.7%, 53.0% and 96.4%, respectively.
The main reason for the deterioration in the UASB performance was disintegration of the
granular biofilms structure causing loss of biomass.
With respect to the nitrogen removal, the formation of NH4+-N in UASB reactor
accounted for the majority of gelatin-N loss (NH4+-Nformed/protein-Nremoved = 0.9),
indicating that gelatin was highly mineralized to NH4+. Both partial nitrification and
complete nitrification took place in aerobic post treatment. When the dissolved oxygen
(DO) concentration was > 2.0 mg L-1, complete nitrification (period B) occurred with an
average nitrification efficiency of 96.2%. Partial nitrification occurred due to high OLR
to AS during the overloading event (period A) and when the DO concentration was
maintained below 2.0 mg L-1 in periods without organic overloading (period C). The
maximum accumulated nitrite concentration in periods A, B and C were 90.0, 0.9 and
75.8 mg NO2- -N L-1, respectively. The nitrogen balance results of periods A and C
indicated that there was a discrepancy between the amount of ammonium nitrogen
removed and the amount of oxidized nitrogen formed. This suggests the occurrence of
simultaneous nitrification/denitrification (SND) in aerobic post treatment.
123 3.2
Introduction
The worldwide population is growing rapidly and there is an increase in the scarcity of
drinking water resources (Aiyuk et al., 2006). A better management of water resources is
therefore needed. One of the goals shared by environmental protection and resource
conservation concepts is the re-use of treated wastewater and its valuable treatment
byproducts (Lettinga et al., 2001; Yi, 2001). Consequently, by implementing such goals,
a wastewater such as domestic wastewater besides being sanitized, can become an
important source of re-usable water, fertilizer, soil conditioner and energy.
Anaerobic treatment is the degradation of organic matter into useful end product methane
in the absence of free molecular oxygen (Geradi, 2003). The anaerobic degradation
occurs by sequential reactions of a number of different microbial groups that cooperate
sequentially to achieve degradation of a variety of polymeric and monomeric substrates
(Ferry 1999). The process is initiated by the hydrolytic bacteria whose enzymes facilitate
the hydrolysis of complex organic molecule such as proteins and polysaccharides to
simple soluble compounds. Further fermentation of the simple compounds results in the
generation of a wide variety of fermentation end products including volatile fatty acids,
ethanol, hydrogen and carbon dioxide. This fermentation process is known as
acidogenesis. The products of acidogenesis are further oxidized to acetate, hydrogen and
carbon dioxide in a process referred to as acetogenesis. Under standard conditions, the
oxidation of the acidogenesis end products is an endergonic reaction. It is only when the
124 hydrogen partial pressure is lowered by the hydrogen utilizing methanogens the reaction
becomes exergonic. Therefore, the acetogenic bacteria have to grow in syntrophy with
the hydrogen utilizing methanogens in order to facilitate the interspecies hydrogen
transfer and gain energy on the products of the acidogenesis phase. The final step in the
anaerobic treatment is methanogenesis. Methanogens are strictly anaerobic Archaea that
can be divided into two groups: i) hydrogen utilizing methanogens which form methane
by the reduction of hydrogen/carbon dioxide and ii) acetoclastic methanogens which
produce methane by the conversion of acetate (Van Andel and Breure, 1984).
Acetoclastic methanogens are considered the most important methanogenic species since
70% of the methane produced is generated via this pathway (Lettinga, 1995).
Application of anaerobic processes for the treatment of low strength wastewater such as
domestic sewage has drawn considerable attention recently, which at present is largely
treated by aerobic processes. There are several drawbacks associated with aerobic
processes especially the activated sludge processes. The treatment costs are high due
mostly to high energy inputs required for aeration. Also, approximately half of the
organic pollutants measured as chemical oxygen demand (COD) are converted to sludge
biosolids. The biosolids need to be treated and disposed of safely which results in high
costs. On the other hand, anaerobic treatment processes offer several advantages over the
aerobic processes. They result in a net production of useful energy in the form of
methane. While the aerobic treatment requires input of 1.0 kWh kg-1 organic removed,
the anaerobic treatment results in conversion of about 77.0% of the organic waste to
methane which corresponds to the production of 2.75 kWh kg-1 organic removed. Also,
125 the biosolids produced from anaerobic treatment is 6.0 to 12.0 fold lower than aerobic
treatment (Rittmann and McCarty, 2001). From these perspectives, anaerobic treatment
appears to be economically more attractive.
Unquestionably, one of the major successes in the development of anaerobic treatment
was the introduction of high rate reactors in which solid retention time (SRT) and
hydraulic retention time (HRT) are uncoupled (McCarty, 1982; Van Lier et al, 2008). In
such reactors, the maximum permissible load is governed by the amount of viable
anaerobic bacteria. High biomass concentrations are obtained by physical retention and or
immobilization of anaerobic biofilms (Alphenaar, 1994). Retaining high biomass
concentration in such reactors allows long SRT to be achieved while keeping the HRT
short. Thus, these configurations allow the application of high organic loading rate in
small reactor volumes. Several anaerobic reactors have been developed including
anaerobic filter (Young and McCarty 1969), anaerobic expanded/fluidized bed reactor
(Switzenbaun and Jewel 1980), anaerobic rotating biological reactor (Blanc et. al., 1983;
Tait and Friedman, 1980), anaerobic baffled reactor (Bachmann et. al., 1985) and up-flow
anaerobic sludge blanket (UASB) (Lettinga et al., 1980).
Among all the high-rate anaerobic reactor configurations, UASB have been widely
adopted for sewage treatment due to their superior performance (Van Haandel and
Lettinga, 1994). The reactor offers great promise especially in developing countries in
tropical regions of the world. It is a robust technology without moving parts; therefore,
limiting both capital and operating costs (Schellinkhout and Collazos, 1992). Like many
126 high-rate reactors, the UASB retains high concentration of biomass in the form of a dense
granular sludge. The UASB design is a simple, low cost and easy to operate. Recent
reports shows that in Argentina (Seghezzo et al., 1995), Brazil (Sousa and Foresti, 1996),
and Colombia (Maaskant et al., 1991; Mora and Sterling, 1996) about 45.0% of all
anaerobic reactors are currently treating domestic sewage, including the biggest UASB
reactor ever built (Monroy et al., 1996) (Seghezzo et al, 1998). COD removal efficiencies
of 65.0-80.0% have been achieved at organic loading rate lower than 3.0 kg COD m-3 d
and HRT ranging from 6.0-10.0 hours (Rodriguez et al., 2001). HRT can be reduced
depending on the quality of the effluent accepted by post treatment units and more
favorable thermal conditions. In tropical regions where the mean temperatures are around
25○C or higher, HRT less than 6.0 hours has been practiced without significant
impairment on the reactor performance (Foresti, 2002).
Although the UASB reactor and other anaerobic reactors are efficient for removing
biodegradable organics, their effluent quality often does not meet the permissible limits
for discharge (Souza, 1986). Furthermore, the nitrogen removal in the UASB is
practically negligible with the transformation of only the organic nitrogen into
ammonium nitrogen (Verstraete and Vandevivere, 1999; Verstraete et al., 2002). The
effluent quality of anaerobic processes vary widely depending on several factors such as
local conditions, influent characteristics, reactor design, operational parameters, etc
(Foresti et al., 2006). Published data from literature reveals that the effluent from
anaerobic treatment of domestic sewage normally contains COD ranges from 100 to 200
mg L-1; total suspended solids ranges from 50 to 100 mg L-1 (Passig et al., 2000; Vieira,
127 1998; Van Haandel and Lettinga, 1994) and ammonia ranges from 30 to 50 mg L-1
(Kobayashi et al., 1983; Torres and Foresti, 2001). Therefore, an adequate post treatment
system is necessary for further polishing the remaining organic pollutants and for the
removal of nitrogenous compounds. The conventional biological method for removing
ammonium nitrogen is the nitrification/denitrification process in which ammonium
nitrogen is converted to dinitrogen gas. Several aerobic treatment units can serve as a
polishing and nitrogen removal step including polishing pond (Cavalcanti, 2003),
trickling filter systems (Chernicharo and Nascimento, 2001), submerged aerated filters
(Goncalves et al., 1998), activated sludge systems (Passig et al., 2000; Von Sperling et
al., 2001), rotating biological contactors (Tawfic et al., 2004; Castilho et al., 1997) and
sequencing batch reactor (Torres and Foresti, 2001; Sousa and Foresti, 1996). The
aerobic step converts ammonium nitrogen to nitrate through nitrification process.
Depending on the ultimate use of the reusable wastewater, further removal of nitrate may
be needed. In such situations, nitrate can be recycled back to the UASB reactor for the
removal of nitrate through denitrification process (An et al., 2008; Tai et al., 2006).
In comparison to the conventional wastewater treatment plant consisting of primary
sedimentation tank followed by aerobic biological treatment in which primary and
secondary sludges are passing through a thickener and digested in anaerobic digester, a
treatment consisting of UASB as a pretreatment followed by aerobic post treatment as a
polishing step in which secondary sludge is digested in the UASB reactor itself offers
several advantages (Alem Sobrinho and Jordao, 2001). Firstly, the primary sedimentation
tank, the sludge thickener and the anaerobic digester as well as their equipments are
128 replaced with the UASB reactor. In this configuration, the UASB reactor functions as a
main treatment as well as a thickener and digester for the produced aerobic sludges.
Secondly, the power consumption for aeration in aerobic treatment unit is substantially
reduced compared to the conventional aerobic treatment without anaerobic pretreatment.
Thirdly, the volume of sludge to be disposed of from the combined system of
anaerobic/aerobic is much lower than those of aerobic alone. This is due to the lower
sludge production from anaerobic processes and their better dewaterability. Pontes (2003)
reported that the total sludge production from a combined system of UASB/Trickling
filter is about 30.0-50.0% lower than in a conventional trickling filter system. Fourthly,
the construction cost of a wastewater treatment plant with a combined system of UASB/
aerobic biological treatment is about 50.0-80.0% of the cost of a conventional treatment
plant. Also, saving in the operation and maintenance costs are usually in the range of
40.0-50.0% compared to the conventional treatment plant (Von Sperling and
Chernicharo, 2005). Therefore, the objective of this research is to investigate the removal
of carbonaceous and nitrogenous contaminants from synthetic high strength sewage by
combined UASB-AS system.
The goal of this study was to investigate the removal of organic and nitrogen compounds
from high strength synthetic sewage in a sequential laboratory-scale UASB and AS
system. The specific objectives were to i) elucidate the removal of most of the organic
compounds in the UASB reactor; ii) polish the remaining organic compounds in AS
reactor; iii) promote both full and partial nitrification in AS reactor for nitrogen
129 ammonium removal; iv) evaluate the overall performance of the integrated UASB-AS
system for combined removal of organics and nitrogen.
130 3.3
Materials and Methods
3.3.1 Chemicals
Gelatin (food grade) was obtained from Research Organics (Cleveland, Ohio, USA). Dglucose anhydrous was purchased from (Fisher Scientific, UAS). Sodium nitrite (>
99.0%), acetic acid (>99.7 %), propionic acid (≥ 99.5%) and butyric acid (≥ 99.0%) were
purchased from Sigma Chemicals Co. (St. Louis, MO, USA). Potassium nitrate (>
99.0%) was obtained from Fisher Scientific (Fair Lawn, New Jersey, 204 USA). All
chemicals were used as received.
3.3.2
Experimental setup
Figure 3.1 depicts the schematic diagram of the experimental set up, consisting of UASBAS post treatment reactor system. The total volume and the working volume of the
UASB reactor were 1.8 and 1.6 L, respectively. The post treatment system (AS reactor)
consisted of an aeration tank and a settler with working volumes of 1.1 and 0.6 L,
respectively. Both UASB and AS reactors were manufactured from Plexiglass and
connected in series. The study was conducted at an average room temperature of 24○C.
The UASB reactor was fed with a synthetic high strength domestic sewage according to
the medium described in Table 3.1. The chemical oxygen demand (COD) strength of the
synthetic sewage was approximately 2.5 g COD L-1. The synthetic sewage was pumped
131 into the UASB reactor from a refrigerator and maintained under an anaerobic
environment (with N2 gas) to prevent the spoiling of the synthetic medium. The initial
hydraulic retention time (HRT) of the UASB reactor was averaged 56.0 hours,
corresponding to organic loading rate (OLR) of 1.1 g COD L-1reactor d-1. The OLR for the
UASB reactor was increased gradually whenever the COD removal efficiency was higher
than 80.0%.
The biogas from the UASB reactor was collected and passed through two 2-liter
Erlenmeyer flasks designed to remove CO2 from the biogas stream. The first flask was
empty, preventing any backflow of sodium hydroxide (NaOH) into the UASB reactor,
while the second flask was filled with 3.0% (m/v) NaOH, which scrubs the CO2.
Following the scrubbing of the biogas, the remaining CH4 gas flow was passed through a
wet-type precision gas meter manufactured by Schlumberger Industries.
The aerobic post treatment was connected to the UASB reactor on day 222. The aerobic
post treatment unit received the effluent directly from the UASB reactor. The solid
retention time (SRT) in the AS reactor was maintained at 25.0 days by wasting a
designated amount of excess sludge from the aeration tank. The aeration tank of the AS
reactor was aerated with an aquarium pump. The dissolved oxygen (DO) concentration in
the aeration tank was maintained higher than 4.0 mg L-1 for the operation period of 222 to
521. On day 522, the DO concentration in AS reactor was reduced below 2.0 mg L-1 to
study the impact of DO on nitrification.
132 Figure 3.1 Schematic diagram of the experimental set-up of the combined system
reactor: (1) UASB reactor, (2) aeration tank, (3) settler.
133 Table 3.1 Composition of synthetic domestic sewage.
Medium Component
Concentration
Concentration
-1
*
(g L )
(g COD L-1)
Gelatin (Protein)
1.116
1.25
D-Glucose
0.469
0.50
VFA* Mixture Solution
2.500 ml L-1
0.75
K2HPO4
0.171
NaH2PO4·H2O
0.227
CaCl2·2H2O
0.020
MgSO4·7H2O
0.035
NH4Cl
0.095
NaHCO3
2.500
Yeast extract
0.020
Trace element solution
1.000 ml
Mixture of Volatile Fatty Acids (acetate, propionate and butyrate).
0.025
134 3.3.3
Seed sludge
The UASB reactor was seeded with anaerobic granular sludge collected from full-scale
UASB reactor treating potato processing wastewater (Aviko, Steenderen, The
Netherlands). The sludge was washed and sieved to remove fine particles before seeding
the UASB reactor. The content of volatile suspended solids (VSS) in Aviko sludge was
11.5%. The reactor was initially seeded with 15.0 g VSS L-1.
The AS reactor was seeded with return activated sludge (RAS) collected from local
municipal wastewater plant (Randolph, Tucson, AZ). The mixed liquor suspended solids
(MLSS) and mixed liquor volatile suspended solids (MLVSS) concentrations were 4778
and 3800 mg L-1, respectively. The aeration tank of the AS reactor was seeded with 10%
(v/v) of its working volume.
3.3.4
Analytical methods
The concentration of VFA (i.e, acetate, propionate and butyrate) in liquid samples was
determined by gas chromatography (GC) using an HP5290 Series II system (Agilent
Technologies, Palo Alto, CA) equipped with a flame ionization detector (GC-FID). The
GC was fitted with a DB-FFAP capillary column (J&W Scientific, Palo Alto, CA). The
temperature of the column, the injector port and the detector were 140, 180 and 275°C,
respectively. Samples collected for VFA measurements were centrifuged at 10,000 rpm
for 10 min.
135 Nitrite and nitrate were analyzed by suppressed conductivity ion chromatography (IC)
using a Dionex 3000 system (Sunnyvale, CA, USA) fitted with a Dionex IonPac AS18
analytical column (4 x 250 mm) and an AG18 guard column (4 x 40 mm). The column
was maintained at 35.0ºC. The eluent used was 10.0 mM KOH at a flow rate of 1.0 ml
min-1. The injection volume was 25.0 μl. Before measurement, all samples were passed
through a membrane filter (0.45 μm).
The total chemical oxygen demand (CODt) was determined by oxidation with dichromate
and analyzed using the colorimetric micro-method described in Standard Methods for
Examination of Water and Wastewater (APHA, 2005). Protein was determined by
Bradford method using Coomassie plus protein assay kit (Thermo scientific, USA). The
gelatin was used to prepare the protein standards. Samples collected for protein
measurements were centrifuged at 10,000 rpm for 10 minutes. Ammonium was
determined using Mettler Toledo SevenMulti ion selective meter with Mettler Toledo
selective ammonium electrode. The dissolved oxygen (DO) concentration in the aeration
tank was measured using Orion DO probe (Orion 97-08, thermo scientific, USA). The pH
was determined immediately after sampling with an Orion model 310 PerpHecT pHmeter with a PerpHecT ROSS glass combination electrode. Other analytical parameters
(e.g., MLSS, MLVSS etc.) were determined according to Standard Methods for
Examination of Water and Wastewater (APHA, 2005).
136 3.4
Results
The feasibility of a combined UASB-AS system treating synthetic sewage for
carbonaceous and nitrogenous contaminant removal was investigated. The operating
conditions of the combined UASB-AS system are shown in Table 3.2 and Figure 3.2. The
operating conditions were divided into six different periods based on organic loading rate
(OLR) of UASB. In periods I-III, the average OLR was 1.1, 2.2 and 3.8 g COD Lr-1 d-1;
respectively. In period IV, the OLR was further increased to 4.4 g COD Lr-1 d-1. The
increase in OLR in period IV lead to biomass washed out from the UASB reactor. The
OLR was, therefore, reduced to 2.2 g COD Lr-1 d-1 for period V. In period VI, the OLR
was further reduced to 1.7 g COD Lr-1 d-1.
Table 3.2 Periods and Operating Conditions of the UASB and AS Reactors.
UASB
*
†
AS*
Period
Days
OLR
(g COD/Lr d)
HRT (d)
OLR
(g COD/Lr d)
NLR
(g N/Lr d)
HRT (d)
I
0-69
1.1
56.0
NM†
NM
NM
II
70-156
2.2
26.0
NM
NM
NM
III
157-296
3.8
15.4
0.9
0.4
0.45
IV
297-313
4.4
13.9
1.8
0.4
0.41
V
314-563
2.2
25.5
1.0
0.2
0.77
VI
564-627
1.7
33.4
0.7
0.2
1.01
AS reactor was started on day 222.
NM: Not measured because AS not yet operational.
137 Periods:
6
I
II
III
IV
V
VI
OLR (g COD/Lr d)
5
4
3
2
1
0
0
100
200
300
400
500
600
700
Time (days)
Figure 3.2 Time course of organic loading rate (OLR) for (●) UASB and (∆) AS
reactors. The dashed line indicates the start of AS operation. The AS OLR was calculated
from the measured UASB effluent COD and the HRT applied to the AS.
138 3.4.1 Performance of combined UASB-AS reactor system for organic removal
3.4.1.1 Total COD removal
The time course of the total COD influent, UASB effluent and AS effluent as well as
removal efficiencies by UASB and by the combined system are depicted in Figures 3.3
and 3.4; respectively. After achieving steady state on day 40, the average total COD
removal of UASB in periods I-III were 86.4±3.2%, 80±4.1% and 83.5±4.9%;
respectively. High biomass washout occurred in UASB between days 297-313 (period
IV) due to increase of the OLR and as a consequence the average total COD removal
efficiency in this period decreased to 66.3±10.9%. Despite lowering the OLR in periods
V and VI, the UASB reactor did not recover completely and the COD removal efficiency
was less than 72.0%. The COD mass balance for UASB for the entire period of operation
is shown in Table 3.3. The CODmethane production was in the same order of magnitude as
the COD removed for the entire period of operation. The slight overestimation of
methane production in periods II, III, V and VI could be due to the experimental error in
measuring COD and/or biogas produced. A plausible explanation could be an incomplete
scrubbing of CO2 from the biogas by the NaOH.
139 Periods:
3000
I
II
III
IV
V
VI
2500
COD (mg/L)
2000
1500
1000
500
0
0
100
200
300
400
500
600
700
Time (days)
Figure 3.3 Time course of influent COD (●), UASB effluent COD (○), AS effluent COD
(∆). The first dashed line indicates the start of AS operation. The second dashed line
indicates lowering dissolved oxygen concentration in AS below 2 mg L-1.
140 Periods:
100
I
II
III
IV
V
VI
90
CODt removal (%)
80
70
60
50
40
30
20
10
0
0
100
200
300
400
Time (days)
500
600
700
Figure 3.4 Time course of total COD removal efficiency by: (●) UASB and (□)
combined UASB-AS system. The first dashed line indicates the start of AS operation.
The second dashed line indicates lowering dissolved oxygen concentration in AS below 2
mg L-1.
141 Table 3.3 COD Mass Balance for UASB reactor.
CODin
CODout
CODCH4
CODout +
CODCH4
(CODout +
CODCH4)/CODin
CODCH4/(CODin
- CODout)
Period
(gCOD/Lr d)
(gCOD/Lr d)
(gCOD/Lr d)
(gCOD/Lr d)
(%)
(%)
I
1.6 (0.2)*
0.2 (0.1)
1.2 (0.1)
1.5 (0.1)
89.3
85.7
II
3.3 (0.3)
0.7 (0.2)
2.9 (0.3)
3.6 (0.3)
108.3
111.5
III
6.0 (0.5)
0.9 (0.3)
5.6 (0.9)
6.5 (1.0)
108.6
109.8
IV
7.8 (0.8)
2.5 (0.8)
4.7 (1.7)
7.2 (1.7)
91.3
88.7
V
3.5 (0.7)
1.1 (0.5)
2.5 (0.4)
3.6 (0.5)
103.0
104.2
VI
2.7 (0.2)
0.8 (0.3)
2.0 (0.4)
2.8 (0.2)
105.5
105.3
*
( ) standard deviation
The aerobic post treatment was responsible for additional COD removal (Figures 3.3 and
3.4). In periods III and IV, the aerobic treatment was accounted for a combined total
COD removal of 90.7±3.7 and 91.7±2.4; respectively. During period V, the total COD
removal of the combined system increased gradually to 100.0%. On day 522, the
dissolved oxygen concentration of AS was lowered below 2.0 mg L-1 (to evaluate its
impact on N removal) and; consequently, the combined total COD removal decreased to
86.0 ± 2.7%. During period VI, the average combined COD removal efficiency increased
again to around 89.1 ± 3.7 % most likely due to a decrease in the OLR applied to the
UASB (Table 3.4).
142 Table 3.4 Summary of Process Performance of UASB unit, AS unit and Combined System for
Organic Constituent Removal.
AS‡
UASB
Combined system
Period
COD
VFA
Protein
COD
VFA
Protein
COD
VFA
Protein
------------------------------------ Removal (%) ------------------------------------
*
I
86.4
(3.2)*
NA†
94.2
(0.8)
NA†
NA
NA
NA
NA
NA
II
80.0
(4.1)
67.9
(3.2)
89.3
(3.2)
NA
NA
NA
NA
NA
NA
III
83.5
(4.9)
73.9
(11.2)
94.1
(1.8)
56.7
(13.4)
99.8
(0.5)
83.4
(19.3)
90.7 (3.7)
99.9 (0.1)
98.2 (2.1)
IV
66.3
(10.9)
53.5
(21.6)
85.0
(13.4)
77.1
(10.2)
99.2
(0.5)
94.1
(9.1)
91.7 (2.4)
99.3 (0.4)
98.1 (2.9)
V
68.7
(10)
47.0
(20)
92.8
(4.3)
73.5
(18.1)
99.1
(1.1)
81.3
(26.6)
91.9 (5.9)
99.1 (1.3)
97.1 (7.6)
VI
70.7
(11)
53.0
(14.6)
96.4
(3.1)
60.6
(14.1)
98.8
(0.7)
88.5
(14)
89.1 (3.7)
99.4 (0.3)
99.6 (0.5)
( ) standard deviation
AS removal = [(UASB effluent-AS effluent)/UASB effluent]*100
†
NA=not applicable, because aerated activated sludge not operational, or VFA not measured yet ‡
143 3.4.1.2 Total VFA removal
Starting on day 98, the volatile fatty acids (VFA) were monitored (Figure 3.5). The
effluent VFA concentration from UASB declined gradually during periods II and III with
an average removal of 67.9±3.2% and 73.9±11.2%; respectively (Table 3.4). The acetic
acid, propionic acid and butyric acid removal efficiency during period III were
54±16.0%, 71±10.9% and 97±4.6%, respectively (Figures 3.6, 3.7 and 3.8). However, the
VFA effluent concentration increased sharply when the UASB was overloaded in period
IV. The VFA in the effluent was even higher than the influent at the end of period IV and
start of period V because gelatine and glucose in the influent was converted to VFA that
accumulated. This was also evident by the accumulation of acetic acid and propionic acid
during the same period of time. As can be seen in Figure 3.6, the acetic acid
accumulation was the highest among the VFA components. However, due to reduction in
the OLR during period V, the VFA accumulation in UASB effluent started to decrease
from 1210.0 mg L-1 COD and stabilized around 434.0 mg L-1 COD. In period VI, the
average VFA removal was 53.0±14.7%. During periods V and VI, the concentration of
acetic acid in the UASB effluent was nearly equal to its concentration in the UASB
influent (Figure 3.6). The propionic acid removal efficiencies during periods V and VI
were 46.5±18.1% and 42.2±20.0%, respectively (Figure 3.7). However, the removal of
butyric acid was almost 100.0% during periods V and VI (Figure 3.8).
With regards to VFA removal, the combined system was not affected by UASB
overloading event. The aerobic post treatment was capable of metabolizing excess VFA
144 from the UASB effluent, pushing the overall efficiency of the combined system to be
higher than 99.0% VFA removal efficiency during the entire period of operation (Figure
3.5 and Table 3.4). The removal efficiency of the combined system for the acetic acid
was nearly 100.0% during the entire period of operation except during the overloading
event which was higher than 90.0% (Figure 3.6). The removal efficiency of the combined
system for the propionic and butyric acids was not affected by the overloading effect and
the removal was almost 100.0% during the entire period of operation (Figures 3.7 and
3.8)
145 Periods:
I
II
III
IV
V
VI
1200
VFA (mg COD/L)
1000
800
600
400
200
0
0
100
200
300
400
Time (days)
500
600
700
Figure 3.5 Time course of influent VFA (●), UASB effluent VFA (○), AS effluent VFA
(∆). The first dashed line indicates the start of AS operation. The second dashed line
indicates lowering dissolved oxygen concentration in AS below 2 mg L-1.
146 Periods:
I
II
III
IV
V
VI
Acetate Conc. (mg COD/L)
1200
1000
800
600
400
200
0
0
200
400
600
Time (days)
Figure 3.6 Time course of influent acetic acid (●), UASB effluent acetic acid (○), AS
effluent acetic acid (∆). The first dashed line indicates the start of AS operation. The
second dashed line indicates lowering dissolved oxygen concentration in AS below 2 mg
L-1.
147 Periods:
I
II
III
IV
V
VI
Propionate Conc. (mg COD/L)
1200
1000
800
600
400
200
0
0
200
400
600
Time (days)
Figure 3.7 Time course of influent propionic acid (●), UASB effluent propionic acid (○),
AS effluent propionic acid (∆). The first dashed line indicates the start of AS operation.
The second dashed line indicates lowering dissolved oxygen concentration in AS below 2
mg L-1.
148 Periods:
I
II
III
IV
V
VI
Butyrate Conc. (mg COD/L)
1200
1000
800
600
400
200
0
0
200
400
600
Time (days)
Figure 3.8 Time course of influent butyric acid (●), UASB effluent butyric acid (○), AS
effluent butyric acid (∆). The first dashed line indicates the start of AS operation. The
second dashed line indicates lowering dissolved oxygen concentration in AS below 2 mg
L-1.
149 3.4.1.3
Protein removal
The time course of protein influent, UASB effluent and AS effluent as well as the protein
removal by UASB and by combined UASB+AS system is shown in Figures 3.9 and 3.10
as well as summarized in Table 3.4. The UASB removal efficiency for periods I, II and
III were higher than 89.0%. During period IV, the protein removal efficiency of UASB
was decreased to around 85.0% due to effects of UASB overloading. However after
decreasing the OLR, the UASB protein removal efficiency recovered quickly and the
average removal efficiency for periods V and VI increased to 92.8 and 96.4%;
respectively. The aerobic post treatment resulted in additional protein removal,
accounting for combined removal efficiency higher than 97.0% for the entire period of
operation except during the UASB overloading event in period IV (Figure 3.9).
150 Periods:
1600
I
II
III
IV
V
IV
Protein (mg/L)
1200
800
400
0
0
100
200
300
400
Time (days)
500
600
700
Figure 3.9 Time course of influent protein (●), UASB effluent protein (○), AS effluent
protein (∆). The first dashed line indicates the start of AS operation. The second dashed
line indicates lowering dissolved oxygen concentration in AS below 2 mg L-1.
151 Periods:
100
I
II
III
IV
V
VI
90
Protein removal (%)
80
70
60
50
40
30
20
10
0
0
100
200
300
400
Time (days)
500
600
700
Figure 3.10 Time course of protein removal efficiency by: (●) UASB and (□) combined
UASB-AS system. The first dashed line indicates the start of AS operation. The second
dashed line indicates lowering dissolved oxygen concentration in AS below 2 mg L-1.
152 3.4.2 Performance of combined UASB-AS system for nitrogen removal
3.4.2.1
Ammonium removal in AS reactor
During the operation of UASB, mineralization of gelatin-N to ammonium (NH4+-N) was
monitored. The production of NH4+-N due to the mineralization of protein as well as the
AS effluent NH4+-N concentration is presented in Figure 3.11. On day 221, the AS
reactor was started and during the start up period the nitrification efficiency increased
with the time. By the end of the period III, a nitrification efficiency of about 96.0% was
achieved. Nevertheless, the NH4+-N concentration started to increase in AS effluent
during UASB overloading event in period IV and reached 126.0 NH4+-N L-1. During
period V, the AS recovered from overloading shock and the average nitrification
efficiency was reestablished at 96.2%. On day 522, the dissolved oxygen concentration
(DO) in AS reactor was reduced below 2.0 mg L-1 to study the impact of DO on
nitrification and as a consequence the NH4+-N removal fluctuated somewhat.
153 Periods:
250
I
II
III
IV
V
VI
+
NH4 -N (mg N/L)
200
150
100
50
0
0
100
200
300
400
Time (days)
500
600
700
Figure 3.11 Time course of influent NH4+ (●), UASB effluent NH4+ (○) and AS effluent
NH4+ (∆).The first dashed line indicates the start of AS operation. The second dashed line
indicates lowering dissolved oxygen concentration below 2 mg L-1 in AS.
154 3.4.2.2 NOx species
Time course of nitrogen species in post aerobic treatment is shown in Figure 3.12.
Initially, partial nitrification to nitrite (NO2-) occurred in AS reactor (period A: days 222381). The average concentrations of NO2- and NO3- were 90.0 and 0.5 mg N L-1;
respectively (Table 3.5). On day 382, the UASB started to stabilize and its organic load to
post aerobic treatment started to decrease and as a consequence the complete nitrification
took place (period B: days 382-521). After the complete nitrification reached steady state
condition (day 417), the average concentrations of NO2- and NO3- were 0.9 and 162.8 mg
N L-1; respectively (Table 3.5). In order to achieve partial nitrification, the DO
concentration in AS was lowered below 2.0 mg L-1 (period C: days: 522-627). Such
reduction in DO concentration led to increase NO2- concentration and decrease NO3- in
post aerobic treatment. However, due to difficulty in maintaining DO below 2.0 mg L-1,
the NO2- and NO3- concentrations were fluctuating. During period C, the average
concentrations of NO2- and NO3- were 75.8 and 31.5 mg N L-1; respectively (Table 3.5).
155 Periods:
500
A
B
C
N-Species (mg N/L)
400
300
200
100
0
220
270
320
370
420
470
520
570
620
670
Time (days)
Figure 3.12 Post aerobic treatment effluent NO2--N (●) and NO3--N (○) concentrations.
The solid vertical line indicates the reduction of organic load to aerobic post treatment
(day 382). The dashed line represents lowering dissolved oxygen below 2 mg L-1 in postaerobic treatment. Periods A, B and C represent partial, complete and partial nitrification;
respectively.
156 Table 3.5 Summary of Process Performance of AS unit for Nitrogen Removal.
*
Periods Parameter A B C OLR (g COD/Lr d)
0.9
0.5
0.6
DO concentration (mg/L)
NM*
> 2.0
< 2.0
NH4+-N in (mg N/L)
178.9
179.0
163.8
NH4+-N eff (mg N/L)
56.0
6.8
14.8
NH4+ removal (%)
69.4
96.2
91.5
NO2- -N formed (mg N/L)
90.0
0.9
75.8
NO3--N formed (mg N/L)
0.5
162.8
31.5
(NO2- -N + NO3--N)/ NH4+-N
removed (%)
73.6
95.1
72.0
NM = not measured
3.4.2.3
Nitrogen balance in AS
Nitrogen balance on AS reactor is shown in Figure 3.13. During the period of
overloading event in the UASB (period A), the amount of ammonium nitrogen removed
(NH4+-Nremoved) was greater than the amount of oxidized nitrogen formed (NO2--Nformed
and NO3--Nformed). The formation of NO2--N and NO3--N accounted for 73.6 % of NH4+-N
removed (Table 3.5). The whole nitrogen balance suggests that maybe denitrification was
occurring in the AS, which is conceivable considering the high organic load the AS was
receiving. During period B, the UASB reactor performance started to stabilize and thus
157 the OLR to the AS started to decrease, as a result, the amount of ammonium nitrogen
removed was close to the amount of oxidized nitrogen formed. The formation of NO2--N
and NO3- -N accounted for 95.1 % of NH4+-N removed (Table 3.5). Therefore during this
period there is no evidence that denitrification occurred.
During the period of
purposefully lowering the dissolved oxygen concentration in AS (period C), the
formation of NO2--N and NO3- -N accounted for 72.0 % of NH4+-N removed (Table 3.5).
Again the whole nitrogen balance is suggestive that some denitrification was occurring in
the AS reactor.
158 N concentration (mg N/L)
Periods:
250
A
B
C
200
150
100
50
0
200
300
400
500
Time (days)
600
700
Figure 3.13 Nitrogen balance on AS. Symbols: (♦) NH4+ removed (◊) effluent NO2- -N
and NO3--N formed. The solid vertical line indicates the reduction of organic load to
aerobic post treatment (day 382). The dashed line represents lowering dissolved oxygen
below 2 mg L-1 in post-aerobic treatment. Periods A, B and C represent partial, complete
and partial nitrification; respectively.
159 3.5
Discussion
The results indicate the feasibility of UASB for (pre) treatment of high strength synthetic
domestic wastewater (2.5 g COD L-1) for organic matter removal down to a HRT of 15.4
h. The average total COD removal efficiency was above 80.0% with an average total
COD concentration in UASB effluent of 371.0 mg L-1 (period III). The average VFA
removal was 74.0% with the residual mostly composed of acetic acid. The average
protein removal efficiency of 94.0% was attained in UASB. In addition, the formation of
NH4+-N accounted for the majority of gelatin-N loss (NH4+-Nformed/protein-Nremoved =
0.9), indicating that gelatin was highly mineralized to NH4+. When the HRT was lowered
to 13.9 h; however, the effluent quality of UASB deteriorated. The VFA started to
accumulate in the UASB effluent and reached 1210.0 mg COD L-1. Such accumulation
of VFA resulted in drop of pH from 7.33 to 7.0. The accumulation of VFA indicates that
the methanogens were the limiting step during the overloading event. Total COD and
protein removal efficiencies decreased from 83.5 to 66.3 % and 94.1 to 85.0%;
respectively. Despite the efforts of correcting the problem by lowering the OLR in
periods V and VI, the UASB did not completely recover. The total COD removal
efficiency was less than 72.0% in both periods. One explanation for the inability of te
UASB to recover UASB could be due to the high biomass wash out that occurred during
the overloading event. Such biomass wash-out lowered the loading capacity of UASB.
Another explanation could be the effects of gelatin (protein) on the stability of anaerobic
160 granule biofilms structure. The effects of proteinaceous compounds on granules are well
documented in literature. Fang et al. (1994) reported that degradation of protein causes
several operational problems such as foaming and biomass floatation, scum layer
formation in the gas-liquid-solid separator of the UASB reactors. In addition, anaerobic
granules in protein rich media exhibited lower hydrophobicity which in turn slowed down
the anaerobic granulation process (Thaveesri et al., 1995) and reduced granular biomass
yield (Vanderhaegen et al., 1992). In this study the reactor was started up with 24.0 g
volatile suspended solids (VSS), at the end there was only 19.1 g VSS left, indicating a
net loss of 20.0% of the biofilms, when in a well-functioning system there should be
biofilms growth.
The combined UASB-AS system performed excellently with respect to organic matter
removal. The average total COD removal efficiency was consistently higher than 89.0%
during the entire period of operation. During the overloading event, the average
combined total COD removal was 91.7%. This high removal efficiency is mainly
attributed to ability of post aerobic treatment to absorb a high OLR from UASB when the
VFA accumulated. The average combined removal of VFA and protein were higher than
99.0 and 97.0%; respectively, during the entire period of operation.
The aerobic post treatment exhibited not only excellent organic removal, but also high
nitrification efficiency. Both partial and complete nitrification took place in aerobic post
treatment. During period A (Figure 3.8), the partial nitrification occurred due to high
OLR to the post aerobic treatment reactor. Increasing organic matter leads to nitrite
161 accumulation due to oxygen deficiency (Leu et al., 1998). Heterotrophs compete with the
nitrifiers for oxygen in the presence of organic matter and generally they outcompete the
nitrifiers due to their higher oxygen affinity (Zhang et al., 1995). Complete nitrification
took place in period B when the OLR from UASB stabilized and the DO concentration
was maintained above 2.0 mg L-1. The minimum DO concentration recommended for
design of AS for nitrification is 2.0 mg L-1 (Metcalf and Eddy). The average nitrification
efficiency achieved during period B was 96.2% (Table 3.4). During period C, the DO
concentration in aerobic post treatment purposefully was lowered to test the hypothesis
that low DO would promote partial nitrification to nitrite. The oxygen affinity constants
for ammonia-oxidizing bacteria (AOB) and nitrite oxidizing bacteria (NOB) are 0.36 and
1.1 mg L-1; respectively (Wiesmann, 1994). Because of higher oxygen affinity constant
for NOB, they are more sensitive to low DO concentration than AOB, causing nitrite
accumulation (i.e., partial nitrification) under low DO concentrations (Garrido et al.,
1997; Joo et al., 2000; Bernet et. al., 2001; Ruiz et al., 2003). The critical values of DO
concentration causing nitrite accumulation are generally in agreement in literature. In one
study, 75.0% of nitrite accumulation with 95.0% of ammonia removal took place in an
AS reactor at a DO concentration of 1.4 mg L-1 (Ciudad et al., 2005). Garrido et al.
(1997) found that the maximum ammonium oxidation and nitrite accumulation was
achieved at DO concentration of 1.5 mg L-1. On the other hand, Ruiz et al. (2003)
observed that 65.0% of the influent ammonium was converted into nitrite at DO
concentration of 0.7 mg L-1.
162 The nitrogen balance results indicate that there was a discrepancy between the amount of
ammonium nitrogen removed and the amount of oxidized nitrogen formed in periods A
and C (Figure 3.8). This discrepancy is attributed to the occurrence of a simultaneously
occurring denitrification in anoxic zones or microniches in the AS. The occurrence of
simultaneous nitrification/denitrification (SND) during aerobic treatment is well
documented in literature (Fuerhacker et al., 2003; Guo et al., 2005; Gupta et al., 1994).
Guo et al. (2005) illustrated the occurrence of simultaneous nitrification/denitrification in
airlift reactor. Also, Holman and Wareham (2005) showed that 75% of nitrogen removed
was via SND process in two laboratory sequencing batch reactors. According to Pochana
and Keller (1999), SND is a physical phenomenon, governed by oxygen diffusion into
flocs of activated sludge. SND requires aerobic zone in the outer part of floc for
nitrification and an interior anoxic zone for denitrification. Pachana and Keller (1999)
reported three important factors affect the SND: availability of sufficient organic carbon,
DO concentration and floc size. Availability of organic carbon is necessary to accomplish
denitrification since it is needed as a carbon and energy source for denitrifiers. Low
dissolved oxygen (< 2.0 mg L-1) and large flocs are essential for optimal SND (Munch et
al. 1996; Third et al., 2003). When the DO concentration decreases, the oxygen
penetration depth decreases and; therefore, the anoxic volume inside the floc increases
(Dekreuk et al., 2005).
163 3.6
Conclusions
The feasibility of a combined UASB-AS system for carbon and nitrogen removal from
high strength synthetic domestic wastewater (2.5 g COD L-1) was demonstrated in this
study. The combined system performed excellently with respect to organic matter
removal. The average total COD removal efficiency was always higher than 89.0%
during the entire period of operation even during the overloading event. This high
removal efficiency is mainly attributed to ability of post aerobic treatment to absorb a
high OLR from UASB when the VFA accumulated.
The aerobic post treatment exhibited not only excellent organic removal, but also high
nitrification efficiency. Both partial nitrification and complete nitrification took place in
the aerobic post treatment. Partial nitrification is an important process in nitrogen
removal since it can be combined with either conventional denitrification which is also
called nitrogen removal via nitrite route or with anaerobic ammonium oxidation
(ANAMMOX). In either way, partial nitrification results in savings in aeration. The
results of this study showed that the partial nitrification can be achieved through
maintaining DO concentration below 2.0 mg L-1.
164 3.7
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170 CHAPTER 4
NITRATE AND NITRITE INHIBITION OF METHANOGENESIS DURING
DENITRIFICATION IN GRANULAR BIOFILMS AND DIGESTED DOMESTIC
SLUDGES 4.1 Abstract
Anaerobic bioreactors that can support simultaneous microbial processes of
denitrification and methanogenesis are of interest to nutrient nitrogen removal. However,
an important concern is the potential toxicity of nitrate (NO3-) and nitrite (NO2-) to
methanogenesis. The methanogenic toxicity of the NOx- compounds to anaerobic
granular biofilms and municipal anaerobic digested sludge with two types of substrates,
acetate and hydrogen was studied. The inhibition was the severest when the NOxcompounds were still present in the media (exposure period). During this period, 95.0%
or greater inhibition of methanogenesis was evident at the lowest concentrations of added
NO2- tested (7.6 to 10.2 mg NO2--N L-1) or 8.3 to 121.0 mg NO3--N L-1 of added NO3-,
depending on substrate and inoculum source. The inhibition imparted by NO3- was not
due directly to NO3- itself, but instead due to reduced intermediates (e.g,. NO2-) formed
during the denitrification process. The toxicity of NOx- was found to be reversible after
the exposure period. The recovery of activity was nearly complete at low added NOx-
171 concentrations; whereas the recovery was only partial at high added NOx- concentrations.
The recovery is attributed to the metabolism of the NOx- compounds. The assay substrate
had a large impact on the rate of NO2- metabolism. Hydrogen reduced NO2- slowly such
that NO2- accumulated more and as a result, the toxicity was greater compared to acetate
as a substrate. The final methane yield was inversely proportional to the amount of NOxcompounds added indicating that they were the preferred electron acceptors compared to
methanogenesis.
172 4.2
Introduction
Ammonium nitrogen, formed from the decomposition of organic matter, is an important
contaminant in a large variety of wastewaters. Some of the better known examples
include municipal domestic wastewater (Henze and Comeau 2008; Tchobanoglous et al.
2003), animal wastes (Knight et al. 2000; Mallin and Cahoon 2003) and landfill leachate
(Kjeldsen et al. 2002). Currently environmental regulations on nutrient nitrogen
discharges are becoming stricter, and consequently biological nutrient nitrogen removal
(BNNR) is progressively becoming an integral part of wastewater treatment. The
overwhelming majority of nutrient nitrogen removal options are based on the microbial
processes of nitrification and denitrification (Tchobanoglous et al. 2003; Ekama and
Wentzel 2008). Nitrification (Eq. 1) is carried out by two sequential reactions of obligate
aerobic chemolithoautotrophic bacteria, the ammonia oxidizing bacteria (AOB); and the
nitrite oxidizing bacteria (NOB).
NH4+ + 2 O2 → NO3- + 2 H+ + H2O
[Eq. 1]
Denitrification is carried out by a broad diversity of mostly heterotrophic as well as a few
autotrophic bacteria that utilize many types of organic and reduced inorganic compounds
as their energy source in the absence of elemental oxygen. Nitrate is reduced to inert
environmentally benign dinitrogen gas. Examples of heterotrophic and autotrophic
denitrification reactions are shown below in Eqs. 2 and 3 with acetic acid and hydrogen
gas as the electron donating substrates, respectively:
173 NO3- + H+ + 0.625 CH3COOH → 0.5 N2 + 1.25 CO2 + 1.75 H2O
[Eq. 2]
NO3- + H+ + 2.5 H2 → 0.5 N2 + 3 H2O
[Eq. 3]
BNNR is achieved by sequential nitrification in the presence of O2 followed by
denitrification driven by organic substrates (or reduced inorganic substrates) in the
absence of O2. As a technology, this can be staged in sequential reactors, a nitrification
reactor followed by a denitrification reactor. Or alternatively, the processes can be
separated temporally in sequencing batch reactors. Both processes can also occur
simultaneously in a single oxygen limited aerobic reactor (Tchobanoglous et al. 2003;
Ekama and Wentzel 2008; Rittman and McCarty 2001).
The microbial process of anaerobic ammonium oxidation (Anammox) is the missing link
of the nitrogen cycle discovered in The Netherlands in the early 1990’s (Mulder et al.
1995; Vandegraaf et al. 1995). Anammox involves the reaction of ammonium (NH4+)
with nitrite (NO2-) to produce dinitrogen gas (N2) (Jetten et al. 2005; Kuenen 2008; van
der Star et al. 2007) as summarized in the simplified equation shown below:
NH4+ + NO2- →N2 + 2 H2O
[Eq. 4]
The thermodynamically favorable reaction (ΔG’° = -257 kJ) utilizes NH4+ as the electron
donor
and
NO2-
as
the
electron
acceptor.
Anammox
is
catalyzed
by
chemolithoautotrophic aquatic bacteria belonging to the order planctomycetes (Strous et
al. 1999). Anammox adds to the repertoire of possible microbial reactions that can be
174 utilized in BNNR during wastewater treatment. To realize anammox with a waste stream
rich in NH4+, it is necessary to oxidize approximately half of the NH4+ to NO2-. This is
achieved by partial nitrification (nitritation) with AOB followed by anammox (van
Dongen et al. 2001) or partial denitrification to nitrite followed by anammox (Kalyuzhnyi
et al. 2008; Sumino et al. 2006).
All the processes of BNNR share a common feature; nitrogen oxides (NOx-), NO3- and/or
NO2-, are brought into contact with microbial active consortia under anaerobic conditions
to promote their conversion to N2 gas. Since anaerobic reactors relying on
methanogenesis to convert organic carbon to methane are utilized at many wastewater
treatment plants (WWTPs), the question arises as to whether they can serve the dual
function of NOx removal. At municipal WWTPs, anaerobic reactors are commonplace for
sludge stabilization (anaerobic digesters) of waste activated sludge (Appels et al. 2008).
Additionally, high rate anaerobic biofilm bioreactors are applied for the treatment of
industrial effluents (Kassam et al. 2003; Frankin 2001) and domestic wastewater (van
Haandel et al. 2006).
Several studies have evaluated combined denitrification and methanogenesis in anaerobic
biofilm reactors, where nitrate is introduced into the anaerobic bioreactor. Denitrification
occurred in upflow anaerobic sludge blanket (UASB) reactors fed with nitrates and
synthetic feeds composed of volatile fatty acids (VFA) (Hendriksen and Ahring 1996;
Lee et al. 2004). Steady state denitrification rates up to a of loading of 336.0 to 600.0 mg
NO3-N L-1 d-1 and 3,300 to 6,600 mg chemical oxygen demand (COD) L-1 d-1 with more
175 than 99.0% removal of both nitrate and carbon was achieved in these studies. COD that
was not utilized for denitrification was converted to methane. The specific denitrification
activity of the biofilms reached values as high as 1.1 to 4.3 g NO3--N g-1 VSS·d-1.
During wastewater treatment, a common strategy proposed for combined C and N
removal involves a sequential treatment of the wastewater in an anaerobic reactor
followed by an aerobic post-treatment with up-front recycle of the final effluent to the
anaerobic reactor to promote denitrification of NOx- formed during post-treatment as
shown in Figure 1. This approach has been tested in several types of systems. In one
study, a synthetic wastewater containing sucrose and peptone was treated with a UASB
followed by an airlift reactor (Tai et al. 2006). The denitrification efficiency of the
integrated system was 86.0%, when the recycle ratio was 400.0%. A brewery wastewater
was treated in a baffled granular sludge bed reactor with five compartments followed by a
nitrification unit. With an effluent recycle of 200.0%, the inlet NOx- of the anaerobic
reactor was 70.0 mg L-1 and the NOx- concentration dropped to zero in the second
compartment of the baffle reactor (Baloch et al. 2006). A pre-settled piggery wastewater
was treated in a UASB – activated sludge (AS) system in which nitrified effluent of the
AS reactor was recycled to the UASB at a recycle ratio of 300.0% (Huang et al. 2007).
The combined reactor system achieved efficient removal of COD (96.0–97.0%) and total
nitrogen removal of (54.0–77.0%). Methanogenesis occurred with nearly-complete
denitrification in the UASB reactor (Huang et al. 2007). The up-front recycle approach
has also been employed with post-treatment nitritation, where NO2- is the predominant
NOx- recycled back to the anaerobic reactor. Landfill leachate was treated in a two-stage
176 UASB reactor followed by post treatment in an anoxic/oxic baffled reactor for nitritation.
Simultaneous methanogenesis and denitrification of NO2- (denitritation) were observed in
the first UASB reactor; where a maximum NO2--N removal rate of 3.0 g N L-1 d-1 was
obtained (Peng et al. 2008). A combined UASB and aerobic membrane reactor for
nitritation was utilized with a synthetic wastewater (glucose, VFA, meat extract and
peptone) and an effluent recycle of up to 800.0% was employed (An et al. 2008). In the
UASB nitrite was converted to N2 gas and excess organic carbon was simultaneously
converted to methane. Organic carbon and total nitrogen removal achieved in the
combined system was 98.0 and 83.0%, respectively.
Recycling of NO2- back to the anaerobic reactor can also result in the anaerobic reactor
becoming enriched with anammox bacteria since the NO2- will occur simultaneously with
NH4+-N. In this regard an expanded granular sludge bed (EGSB) reactor was evaluated
for simultaneous anammox, denitrification and methanogenesis (Zhang 2003). NO2removal efficiencies of 97.0 to 100.0% were achieved with removal rates as high as 800.0
mg N L-1 d-1. NO2- formed by partial denitrification of NO3- in a UASB reactor can also
favor the co-occurrence of anammox bacteria and methanogens (Sumino et al. 2006).
The reports of simultaneous methanogenesis with denitrification and anammox are
intriguing because there is also evidence that NOx- compounds are inhibitory to
methanogenesis. As early as 1941, Barker reported that high concentrations of nitrate
inhibited a pure culture of the methanogen, Methanobacterium omelianskii (Barker
1941). Complete growth inhibition was observed at 140.0 mg NO3--N L-1 and higher.
177 Autotrophic
methanogenesis
by
pure
cultures
of
Methanobacterium
thermoautotrophicum and Methanobacterium formicicum were partially inhibited by
130.0 mg L-1 NO3--N and severely to completely inhibited in the presence of 13.0 mg L-1
NO2--N (Balderston and Payne 1976). Autotrophic methanogenesis by Methanosarcina
barkeri was inhibited 50.0% by 42.0 mg NO3--N L-1 and 0.1 mg NO2--N L-1 (Cluber and
Conrad 1998). Acetoclastic methanogenesis by Methanosarcina mazei was only partially
inhibited by high concentrations of NO3-, ranging in concentrations from 200.0 to 1,000.0
mg L-1 NO3--N but severely inhibited by only 2.5 mg L-1 NO2--N (Clarens et al. 1998).
The results taken as a whole suggest that NO3- inhibition of methanogenesis is low to
moderate whereas the inhibition caused by NO2- is very severe.
The objective of this study is to determine how toxic NOx- compounds are to mixed
microbial consortia found in anaerobic bioreactors. The methanogenic inhibition and
denitrification of NO3- and NO2- were studied with anaerobic granular biofilms from a
UASB reactor and a municipal digested sludge from a digester. The impact of two
substrates, H2/CO2 and acetate on NOx- inhibition and denitrification was also evaluated
for each microbial consortium.
178 4.3
Materials and Methods
4.3.1 Chemicals
Potassium nitrate (> 99.0% purity) was obtained from Fisher Scientific (Fair Lawn, New
Jersey, USA). Sodium nitrite (> 99.0%) and sodium acetate were purchased from Sigma
Chemical Co. (St. Louis, MO, USA). All chemicals were used as received.
4.3.2
Sludge sources
Two different methanogenic inocula were used in this investigation including granular
sludge (Mahou) and anaerobically digested sludge (Ina Road sludge). Mahou sludge was
obtained from a full-scale UASB treating wastewater at Mahou beer brewery in
Guadalajara, Spain. The sludge was washed and sieved to remove fine particles before
use in the toxicity assays. Ina Road sludge was obtained from a municipal anaerobic
digester at a local municipal wastewater treatment plant (Ina Road wastewater treatment
plant, Tucson, Arizona). The content of volatile suspended solids (VSS) in Mahou and
Ina Road sludge was 8.13% and 1.23%, respectively, of the wet weight. The maximum
specific activity of the Mahou sludge was 409.0 and 207.0 mg chemical oxygen demand
(COD) as CH4 g-1 VSS d-1, respectively with acetate and hydrogen as the substrates. The
maximum specific activity of the Ina sludge was 185.0 and 214.0 mg CH4-COD as CH4
g-1 VSS d-1, respectively with acetate and hydrogen as the substrates.
179 4.3.3 Basal media
The basal mineral medium was prepared using ultrapure water (Milli-Q system;
Millipore) and contained (in mg L-1): KH2PO4 (37.0); CaCl2•2H2O (10.0); MgSO4•7H2O
(10.0); MgCl2•6H2O (78.1); NH4HCO3 (987.9); NaHCO3 (3,003.0); Yeast extract (20.0),
and trace element solution (1.1 ml L-1). The trace element solution contained (in mg L-1):
H3BO3
(50.0);
FeCl2•4H2O
(2000.0);
ZnCl2
(50.0);
MnCl2•4H2O
(50.0);
(NH4)6Mo7O24•4H2O (50.0); AlCl3•6H2O (90.0); CoCl2•6H2O (2000.0); NiCl2•6H20
(50.0); CuCl2•2H2O (30.0); NaSeO3•5H2O (100.0); EDTA (1000.0); resazurin (200.0)
and 36% HCl (1.0 ml L-1). The final pH of the basal medium was adjusted to 7.2 with
HCl.
4.3.4
Methanogenic toxicity assays
Methanogenic toxicity assays were conducted in glass serum flasks (160.0 ml) at 30±2○C
in an orbital shaker (75 rpm). Mahou granular sludge and Ina Road sludge were added at
a concentration of 1.5 g VSS l-1 and 1.2 g VSS L-1; respectively, to serum flasks
containing 70.0 mL basal medium. All flasks were sealed with butyl rubber stoppers and
aluminum crimp seals and their headspace was flushed with N2/CO2 (80:20, v/v) for 5.0
min to create an anaerobic condition. Both acetate and hydrogen were used as an electron
donor. For the acetoclastic assays, the acetate was supplemented at a final concentration
of 2.0 g COD L-1. For hydrogenotrophic assays, the H2/CO2 was supplied by pressurizing
the flasks to 1.5 atm (80:20, v/v). The flasks were pre-incubated overnight to ensure that
180 the sludge was adapted to medium conditions. On the following day, either nitrite or
nitrate was added at different concentrations then the headspace of the flasks for
acetoclastic assays was flushed again with N2/CO2 (80:20, v/v). For the hydrogenotrophic
assays, the headspace of the flasks was first flushed with N2/CO2 (80:20, v/v) and then
pressurized with H2/CO2 (80:20, v/v) to 1.5 atm, the final partial pressure of H2 was 1.2
atm, corresponding to 1.0 g COD L-1liq. Then, the flasks were incubated at least two hours
prior to determination of methane content in the headspace of each flask. The methane
content of the headspace of each flask was measured several times during the first day of
the experiment and periodically during the subsequent days until all the substrate (acetate
or H2) was depleted. In addition, several liquid samples were taken from each flask over
the course of the experiment for the determination of nitrite and nitrate. Triplicate
controls were included in each assay where no toxicant was added.
The specific methanogenic activities (mg CH4-COD g-1 VSS d-1) were calculated from
the slope of the cumulative methane production versus time and the initial biomass
concentration as the mean value of the triplicate assays. The activities of treatments
spiked with NO3- or NO2- were normalized to the activity in a control lacking NOxcompounds
181 4.3.5
Analytical methods
The methane concentration from the headspace of the serum flasks was determined by
gas chromatography (GC) using an HP5290 series II system (Agilent Technologies, Palo
Alto, CA) equipped with a flame ionization detector (GC-FID). The GC was fitted with a
DB-FFAP column (J&W Scientific, Palo Alto, CA) capillary column. The temperature
of the column, the injector port and the detector was 140, 180 and 250°C, respectively.
The carrier gas was helium at a flow rate of 9.3 mL min-1 and a split flow of 32.4 mL
min-1. Samples for measuring methane content (100 µL) in the headspace were collected
using a pressure-lock gas syringe.
Nitrite and nitrate were analyzed by suppressed conductivity ion chromatography using a
Dionex 3000 system (Sunnyvale, CA, USA) fitted with a Dionex IonPac AS18 analytical
column (4 mm x 250 mm) and AG18 guard column (4 mm x 50 mm). The column was
maintained at 35.0ºC. The eluent used was 10.0 mM KOH at a flow rate of 1.0 mL min-1.
The injection volume was 25.0 μL. Before measurement, all samples were centrifuged at
10,000 rpm for 10.0 min.
VSS and other analytical parameters were determined according to Standard Methods for
Examination of Water and Wastewater (APHA 2005).
182 4.4
4.4.1
Results
Methanogenic inhibition
The impact of NO3- on methanogenesis by granular anaerobic sludge (Mahou) is shown
in Figures 4.1 and 4.2 in assays with acetate and hydrogen. In the assay with both acetate
and hydrogen, the denitrification lag phase lasted for approximately 1 day, during which
time there was no large impacts on methanogenesis. As soon as denitrification started,
inhibition of methanogenesis was evident, being more severe at the higher initial NO3concentrations. The inhibition was markedly higher in assays with hydrogen.
The
initiation of denitrification was paralleled by an accumulation of NO2-. This accumulation
was low in the case of acetate and high in the case of nitrite. At the peak, of the
accumulation, NO2- was recovered with an 8 and 88% molar yield of the added NO3- in
the assay with acetate and hydrogen, respectively. An additional consequence was the
sludge was exposed to NO2- for a longer period in the hydrogen assay.
The
methanogenesis resumed after all the NOx- species were consumed. During the period
when NOx- species were present degree of inhibition was markedly less in the acetate
assay compared to the hydrogen assay. At the two highest NO3- concentrations tested, the
inhibition was close to 100% in the hydrogen assay. The methanogenic activity was
completely to partially reversible depending on the initial NO3- concentration and
electron-donor used. The final yield of methane was inversely proportional to the NO3initially added.
183 70
120
100
50
40
80
30
60
20
40
10
20
-
A
60
-
CH4 Conc. Headspace (%)
NO3 Conc. (mg/L NO3 -N)
0
NO2 Concentration (mg/L NO2 -N)
0
0
50
30
100
Time (h)
-
25
20
150
B
15
10
5
-
0
0
50
100
150
Time (h)
Figure 4.1 Impact of NO3--N concentration on methane formation and denitrification by
Mahou granular sludge with acetate as the electron donor. Panel A. Headspace methane
concentrations with initial NO3--N concentrations of: (▲), 0; (●), 16; (■), 46; (♦), 121 mg
N L-1. NO3--N concentrations with initial NO3--N concentrations of: (○), 16; (□), 46; (◊),
121 mg N L-1. Panel B. NO2--N concentrations with initial NO3--N concentrations of: (○),
16; (□), 46; (◊), 121 mg N L-1.
184 30
120
A
25
100
20
80
15
60
10
40
5
20
-
CH4 Conc. Headspace (%)
-
NO3 Conc. (mg/L NO3 -N)
0
0
0
100
200
300
Time (h)
120
B
-
NO2 Concentration (mg/L NO2 -N)
100
80
60
40
-
20
0
0
100
200
300
Time (h)
Figure 4.2 Impact of NO3--N concentration on methane formation and denitrification by
Mahou granular sludge with hydrogen gas as the electron donor. Panel A. Headspace
methane concentrations with initial NO3--N concentrations of: (▲), 0; (●), 6; (■), 45; (♦),
109 mg N L-1. NO3--N concentrations with initial NO3--N concentrations of: (○), 6; (□),
45; (◊), 109 mg N L-1. Panel B. NO2--N concentrations with initial NO3--N
concentrations of: (○), 6; (□), 45; (◊), 109 mg N L-1.
185 The inhibition of methanogenesis by NO2- was also evaluated with the Mahou sludge
using acetate and hydrogen as the electron donating substrate (Figures 4.3 and 4.4). In
both assays the methanogenesis was inhibited nearly completely from time 0 as long as
the NO2- was present. The methanogenesis started once the NO2- was consumed. This
occurred earlier in the acetate assay because the NO2- was consumed faster compared to
the hydrogen assay. The methanogenic activity was more severely impacted by the
highest NO2- concentration in the hydrogen assay, most likely because it was exposed to
NO2- for a longer period. As was observed with NO3-, the final yield of methane was
inversely proportional to the addition of NO2-.
186 60
160
-
50
120
40
30
80
20
40
-
NO2 Concentration (mg/L NO 2 -N)
CH4 Conc. Headspace (%)
10
0
0
50
100
150
0
200
Time (h)
Figure 4.3 Impact of NO2--N concentration on methane formation and denitrification by
Mahou granular sludge with acetate as the electron donor. Headspace methane
concentrations with initial NO2--N concentrations of: (▲), 0; (●), 8; (■), 62; (♦), 153 mg
N L-1. NO2--N concentrations with initial NO2--N concentrations of: (○), 8; (□), 62; (◊),
153 mg N L-1.
187 35
30
140
25
120
100
20
80
15
60
10
40
-
-
NO2 Concentration (mg/L NO2 -N)
160
CH4 Conc. Headspace (%)
5
20
0
0
0
100
200
300
400
500
Time (h)
Figure 4.4 Impact of NO2--N concentration on methane formation and denitrification by
Mahou granular sludge with hydrogen gas as the electron donor. Headspace methane
concentrations with initial NO2--N concentrations of: (▲), 0; (●), 8; (■), 68; (♦), 162 mg
N L-1. NO2--N concentrations with initial NO2--N concentrations of: (○), 8; (□), 68; (◊),
162 mg N L-1.
188 Anaerobic digester sludge (Ina) was also tested to determine if the trends observed were
consistent with a completely different source of inoculum. Ina sludge was assayed with
NO3- and hydrogen as electron acceptor as shown in Figure 4.5. Unlike, Mahou sludge,
denitrification started without a lag phase. This coincided with evidence for nearly
complete inhibition of methanogenesis at the start of the experiment. Denitrification was
rapid with only a low peak NO2- concentration (molar yield of 14% compared to added
NO3- added). When all the NOx- species were consumed after 2 days methanogenesis
started and the activity was only partuially inhibited compared to the untreated control.
As was observed in the Mahou sludge, the final yield of methane was inversely
proportional to the addition of NO3-.
189 30
-
120
25
20
80
15
10
40
5
0
0
160
-
A
35
-
CH4 Conc. Headspace (%)
40
NO3 Conc. (mg/L NO3 -N)
NO2 Conc. (mg/L NO2 -N)
50
150
200
0
250
150
200
250
100
Time (h)
50
B
40
30
20
-
10
0
0
50
100
Time (h)
Figure 4.5 Impact of NO3--N concentration on methane formation and denitrification by
Ina municipal digester sludge with hydrogen gas as the electron donor. Panel A.
Headspace methane concentrations with initial NO3--N concentrations of: (▲), 0; (●), 8;
(■), 72; (♦), 170 mg N L-1. NO3--N concentrations with initial NO3--N concentrations of:
(○), 8; (□), 72; (◊), 170 mg N L-1. Panel B. NO2--N concentrations with initial NO3--N
concentrations of: (○), 8; (□), 72; (◊), 170 mg N L-1.
190 The relative activity of all the experiments are summarized in Figure 4.6, including
additional experiments conducted with Ina sludge. For the most part, the methanogenic
inhibition was nearly complete during exposure to NOx- compounds. An exception was
the lower NO3- concentration treatments in the Mahou sludge assay with acetate.
However, after metabolism of the NOx-, methanogenic activity was partially restored. The
degree of restoration was greater at lower NOx- concentrations. The restoration of
methanogenic activity was greater when acetate was the electron donor compared to
hydrogen.
191 100
60
40
A
80
20
Relative Activity (%)
Relative Activity (%)
0
50
100
B
80
60
40
20
0
150
-
80
C
60
40
20
0
100
150
-
NOx Conc. (mg/L NOx -N)
Relative Activity (%)
100
50
-
-
NOx Conc. (mg/L NOx -N)
100
D
80
60
40
20
0
0
100
0
0
Relative Activity (%)
50
-
100
150
-
NOx Conc. (mg/L NOx -N)
0
50
-
100
150
-
NOx Conc. (mg/L NOx -N)
Figure 4.6 Relative methanogenic activity as a function of NOx--N concentration. Panel
A. Mahou granular sludge with added NO3-. Panel B. Mahou granular sludge with added
NO2-. Panel C. Ina municipal digester sludge with added NO3-. Panel D. Ina municipal
digester sludge with added NO2-. Relative activities during the period of NOx- exposure:
(●), acetate as electron-donor; (▲), hydrogen as electron donor. Relative activity after
metabolism of NOx-: (○), acetate as electron-donor; (Δ), hydrogen as electron donor. The
activities are normalized to a parallel incubated control with no added NOx-. The highest
NO2- treatment in panel D for Ina sludge with hydrogen could not be plotted for the
period after metabolism, because NO2- was not completely metabolized.
192 4.4.2 Electron balance
A balance of the electron equivalents is shown in Figure 4.7 for the assay with Mahou
sludge spiked with NO3- and acetate. The balance shows that as the NO3--N
concentrations were progressively increased, the decrease in methane yield coincided
with electrons required to reduce NO3- to N2. All the other assays had a different pattern
as shown in a typical example with Ina sludge spiked with NO3- in Figure 4.8. This
example shows that as the NO3--N concentrations were progressively increased, the
decrease in methane yield could not be totally accounted for by the electrons required to
reduce NO3- to N2. Instead, the hole in the balance could be accounted for if the NO3were assumed to have undergone dissimilatory nitrate reduction to ammonium (DNRA).
193 0.25
e equivalents/L
0.20
0.15
0.10
0.05
0.00
0
6
46
121
-
NO3 -N Conc. (mg/L N)
Figure 4.7 The balance of electron equivalents used for methanogenesis and NO3removal as a function of increasing NO3- addition in the assay with Mahou sludge and
acetate. The black bars indicate methane production, the spotted bars indicate loss of
NO3-, assuming conversion to N2.
194 0.16
e equivalent/L
0.12
0.08
0.04
0.00
0
8
72
170
-
NO 3-N Conc. (mg/L N)
Figure 4.8 The balance of electron equivalents used for methanogenesis and NO3removal as a function of increasing NO3- addition in the assay with Ina sludge and
hydrogen. The black bars indicate methane production, the spotted bars indicate loss of
NO3-, assuming conversion to N2, the white bars indicate the additional electron
equivalents beyond denitrification if the NO3- were reduced by dissimilatory nitrate
reduction to ammonia (DNRA).
195 4.4.3 Nitrate and nitrite metabolism
Figures 4.1-4.5 indicate that both NO3- and NO2- were readily metabolized during the
methanogenic inhibition assays. The denitrification or DNRA was initiated in most cases
with a lag phase of 1 d or less. The results indicate that anaerobic sludges are abound
with bacteria capable of denitrification or DNRA. The maximum specific NOx- removal
activity is summarized in Table 1. The highest specific rates of NOx- metabolism were
observed with NO3- and hydrogen in Ina sludge as well as NO3- and acetate in Mahou
sludge. The lowest specific rates were observed with NO2- and H2 regardless of the
inoculum source. The low rate probably accounts for the high accumulation of NO2- in
the experiment with Mahou sludge and hydrogen. The slow metabolism of NO2- in the
experiment with Ina sludge and H2 is due to inhibition of NO2- metabolism by high NO2concentrations. Compared to the rates recorded at the lower concentrations, the NO2removal activity at 180 mg NO2--N L-1 was 40% inhibited (results not shown).
196 Table 4.1. Specific rates of NOx- metabolism in anaerobic sludge with no prior enrichment
Sludge
NOx-
Electron Donor
Specific Activity
(mg NOx--N g-1 VSS d-1)
Mahou
NO2-
acetate
717
Mahou
NO2-
H2
250
Mahou
NO3-
acetate
1036
Mahou
NO3-
H2
680
Ina
NO2-
acetate
302
Ina
NO2-
H2
77
Ina
NO3-
H2
1938
197 4.5
Discussion
The results of this study indicate that NOx- compounds are highly toxic to
methanogenesis and they are also readily metabolized by the anaerobic sludges. The
inhibition was the severest when denitrification was active and the NOx- compounds were
still present in the media (exposure period). During this period, 95.0% or greater
inhibition of methanogenesis was already evident at the lowest NO2- concentrations
tested (7.6 to 10.2 mg NO2--N L-1) in all assays of this study. In the case of NO3-, 95.0%
or greater inhibition of methanogenesis during the exposure period was evident at
concentrations of 8.3 to 121.0 mg NO3--N L-1, depending on substrate and inoculum
source.
The inhibition imparted by NO3- does not appear to be directly due to NO3- itself, but
instead due to reduced intermediates formed during the denitrification process. There are
multiple lines of evidence to support this hypothesis. Firstly, during the first day of the
experiments with the granular biofilm sludge (Mahou), there is a lag phase prior to the
initiation of denitrification, and during this period there is no noteworthy inhibition of
methanogenesis. Secondly, the lag phase was followed by a temporal accumulation of
NO2-, a known intermediate of denitrification and during this period, there was a
complete cessation of methane production. Thirdly, when NO2- was completely
consumed, the methane production resumed again. These findings are in agreement with
previous studies exploring the inhibition NO2- and NO3- to methanogenesis where it has
198 been consistently observed that NO2- is markedly more toxic to methanogenesis. This has
been evident in pure cultures (Clarens et al. 1998, Balderston and Payne 1976, Kluber
and Conrad 1998) as well as natural mixed- and enrichment cultures (Balderston and
Payne 1976, Tugtas and Pavlostathis 2007a). Likewise in this study, the inhibitory impact
of NO2- to Mahou during the period of NOx- exposure was greater than that of NO3-.
Fourthly, the process of denitrification itself has been associated with a greater toxicity of
NO3- to methanogenesis. Co-culture experiments showed that the methanogen,
Methanosarcina mazei, growing with 210 mg NO3--N L-1, produced methane from
acetate until the denitrifying bacterium Pseudomonas stutzeri was introduced into the
culture and NO3- denitrification began Clarens et al. (1998), suggesting that reduced
products of NO3- were responsible for the toxicity. While exploring different electron
donor substrates for denitrification in a methanogenic enrichment culture, it was observed
that the substrates causing a greater accumulation of reduced denitrification intermediates
(e.g., NO2- and nitric oxide (NO)) were associated with a greater toxicity caused by added
NO3- (Tugtas and Pavlostathis 2008).
The pattern of methane formation with the Mahou sludge and added NO3- followed a
staircase pattern with an initial upward slope – followed by a plateau – followed by a
resumption of the upward slope – followed by a final plateau. The first plateau
corresponded to the period of NO2- accumulation; the second plateau corresponded to the
exhaustion of substrate. This pattern was distinct with the Ina sludge and added NO3-,
which lacked the initial period of methane production and the methane production only
started after intermediate NO2- had disappeared. The difference in behavior may possibly
199 be attributed to the more rapid initiation of denitrification in the Ina sludge experiments
as evidenced from an early start of the NO3- removal without any lag phase. The lowest
NO3- concentration tested was inhibitory initially even though hardly any measurable
NO2- was formed in that case (Figure 4.5). Other denitrification products aside from NO2, such as NO may have accumulated as well. Nitric oxide is known to be one of the most
toxic intermediates of denitrification to methanogenesis (Balderston and Payne 1976,
Tugtas and Pavlostathis 2007a).
Figure 4.9 summarizes the inhibition data of this study as a function of the maximum
NO2- concentration, irregardless of whether the NO2- was added initially or formed form
the biotransformation of added NO3-. The figure illustrates a relationship between the
observed inhibition during the exposure period and the maximum NO2- concentration in
that period. The NO2- was very toxic, such that at most concentrations nearly complete
inhibition was observed. A few points in the range of 0.3 to 0.6 mg L-1 NO2--N provided
partial inhibition data enabling estimates of the 50% inhibiting concentration (IC50). The
IC50 of acetoclastic and hydrogenotrophic methanogenesis in the Mahou sludge was
estimated to be 0.83 and 0.38 mg L-1 NO2--N, respectively; indicating that the
hydrogenotrophic
methanogens
were
more
sensitive
to
NO2-
toxicity.
The
hydrogenotrophic methanogenesis of the Ina sludge was inhibited by 98.5% at 0.38 mg L1
NO2--N, corresponding to an estimated IC50 of 0.19 mg L-1 NO2--N. The higher toxicity
of the latter is most likely attributable to dispersed biomass in the Ina sludge which was
more exposed to NO2- toxicity compared to methanogens in biofilm granules of the
Mahou sludge.
200 100
60
40
40
100
20
0
0
2
4
NO2
6
(mg/L
8
10
NO2 -N)
80
120
160
Conc. NO2- (mg/L NO2--N)
B
80
Inhibition (%)
40
0
60
Conc.
80
0
20
Inhibition (%)
100
100
60
Inhibition (%)
Inhibition (%)
A
80
40
80
60
40
20
0
20
0
2
Conc.
4
NO2
6
(mg/L
8
10
NO2 -N)
0
0
40
80
120
160
Conc. NO2- (mg/L NO2--N)
Fig. 4.9 Inhibition of methanogenic activity during the exposure as a function of the
maximum NO2--N concentration measured. Panel A. Mahou granular sludge. Panel B.
Ina municipal digester sludge. Legend: (▲), acetate as electron-donor in assays with
added NO3-; (●), acetate as electron-donor in assays with added NO2-; (Δ), hydrogen as
electron donor in assays with added NO3-; (○), hydrogen as electron donor in assays with
added NO2-.
201 The toxicity of NOx- was found to be reversible after the exposure period. The reversal is
attributed to the metabolism of the NOx- compounds and intermediates. The recovery of
activity was complete at low NOx- concentrations; whereas the recovery was only partial
at high NOx- concentrations. The pattern of inhibition recovery to NOx- exposure has
been observed in other studies. In a methanogenic enrichment culture, with added NO2--N
additions ranging from 17 to 500 mg L-1, the recovery at 17 to 50 mg NO2--N L-1 was
nearly complete; whereas, the activity recoveries at 250 to 500 mg NO2--N L-1 were only
a fraction of the original activity (Tugtas and Pavlostathis 2007a). In a mesophilic sulfate
adapted sludge, complete recovery from NO2- exposure was observed at added
concentrations ranging from 5 to 150 mg NO2--N L-1, the only difference was that the lag
phase for the recovery increased with increasing concentrations (O'Reilly and Colleran
2005). The results taken as a whole seem to suggest that exposure to very high NO2concentrations (and possibly other reduced denitrification intermediates) are damaging to
the methanogens; whereas low concentrations only cause a largely reversible inhibition.
The electron donor of the experiment also impacted the toxicity. When H2 was used as
the substrate, reduction of NO2- was slower compared to acetate. In the Mahou sludge
160 mg NO2--N L-1 was completely denitritated in 70 hours; with acetate; whereas it took
240 hours to completely denitritate the same concentration with H2 as electron donor. The
resulting impact on the residual methanogenic activity was noticeable. The recovery of
activity with acetate was 28%; whereas the longer exposure to NO2- with H2 as an
electron donor resulted in a recovery of only 14% of the methanogenic activity. The
electron donor also impacted the temporal accumulation of NO2- from NO3-. With acetate
202 and Mahou sludge, the maximum accumulation of NO2--N from 120 mg NO3--N added L1
was 10 mg L-1; whereas with H2, the accumulation was 96 mg L-1, which was almost a
stoichiometric conversion. The difference on the recovery of the methanogenic activity
following NOx- exposure was noteworthy with the 60 and 27% recovery in the acetate
and hydrogen experiment, respectively. The greater accumulation of NO2- with H2 versus
acetate as an electron donor was also noted in a methanogenic enrichment culture, and
that difference also corresponded to a markedly enhanced toxicity as well (Tugtas and
Pavlostathis 2008).
The electron balances in this study indicate that denitrification to N2 gas occurred in
Mahou sludge with added NO3- and acetate as electron donor. However, most of the other
assays could only be properly balanced if the NOx- compounds were assumed to have
undergone DNRA. DNRA is known to occur in very reduced environments with either
sulfide or high chemical oxygen demand to NO3- ratios (Tugtas and Pavlostathis 2007b,
Akunna et al. 1992). Injection of nitrate into an anaerobic digester treating municipal
solid waste was associated with ammonium accumulation and DNRA was suspected
(Vigneron et al. 2007). In this study the most noteworthy distinguishing feature between
cultures suspected of carrying out DNRA compared to denitrification, is that the latter
was less inhibited by NOx- compounds. Therefore when methanogenesis is not
functioning properly there is a greater need for a sink of electrons and DNRA may be that
sink.
203 The NOx- compounds were the preferred electron acceptors compared to methanogenesis.
In most cases, once denitrification commenced, the electron equivalents were exclusively
directed towards denitrification or DNRA. The sludges used were not previously enriched
for denitrification, yet in all cases the NOx- removal was fully underway after one day.
Specific rates of nitrate removal ranged up to 1938 mg NO3--N g-1 VSS d-1; and specific
rates of denitritation ranged up to 717 mg NO2--N g-1 VSS d-1. These rates are
comparable to rates obtained in highly enriched biofilms granules (Hendriksen and
Ahring 1996, Lee et al. 2004).
While this study confirms that NOx- compounds are highly toxic to methanogenesis this
observation is not necessarily inconsistent with the numerous reports demonstrating
simultaneous denitrification and methanogenesis in anaerobic biofilms reactors
(Hendriksen and Ahring 1996, Lee et al. 2004, Huang et al. 2007, Peng et al. 2008, An et
al. 2008). The lessons learned from this study indicate that denitrification intermediate
concentrations in the anaerobic bioreactors have to be maintained at very low
concentrations. Additionally, biofilms may develop zones where methanogens are
protected from denitrification products by spatial separation from denitrifiers.
204 4.6
Conclusions
The aim of this study is to evaluate the inhibitory effects of nitrite and nitrate on
methanogenesis in granular biofilms and digested domestic sludges. The results indicated
that nitrate and nitrite are highly toxic to the methanogenesis. Nitrite is more toxic to
methanogenesis than nitrate and the inhibition of nitrate to methanogenesis is not related
to nitrate itself but it is due to the formation of its intermediate (e.g. nitrite) during
denitrification. As long as the NOx- compounds present in the media, the methanogenesis
is inhibited. Once the NOx- compounds are completely reduced by denitrification, the
methane production is resumed. In this study, acetate is a better electron donor to
denitrification than hydrogen when nitrite is used as electron acceptor. This is evident by
the slow reduction of nitrite when hydrogen is used compared to acetate.
205 4.7
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Science Biotechnology 5:21-38.
Vigneron, V., Ponthieu, M., Barina, G., Audic, J.M., Duquennoi, C., Mazeas, L., Bernet,
N., and Bouchez, T. (2007) Nitrate and nitrite injection during municipal solid waste
anaerobic biodegradation. Waste Management 27:778-791.
Zhang, D.J. (2003) The integration of methanogenesis with denitrification and anaerobic
ammonium oxidation in an expanded granular sludge bed reactor. Journal Environmental
Science-China 15:423-432.
209 CHAPTER 5
Enrichment of Anaerobic Ammonium Oxidizing (Anammox) Bacteria from
Wastewater Sludge for Biological Nutrient Nitrogen Removal
5.1 Abstract
Sludge samples from different sources such as municipal wastewater treatment plants
(WWTPs) and methanogenic granular sludges were screened in batch cultures to
determine which had the highest intrinsic level of Anammox activity. The occurrence of
Anammox activity in such cultures was demonstrated when the nitrite consumption was
significantly enhanced in treatments containing ammonium compared to controls lacking
ammonium. Anammox enrichments were readily developed from municipal WWTP
sludges, but not from methanogenic granular sludges. Among the municipal WWTP
sludges, the return activated sludge (RAS) obtained from Ina Road WWTP operated for
biological nitrogen nutrient removal (BNNR) had the highest intrinsic level of Anammox
activity. The culture developed rapidly within 40.0 days and thereafter the culture was
enriched further. The doubling time of RAS culture was estimated to be approximately
6.8 days. The development of Anammox culture from RAS in such short period of time
210 was due to operating the batch culture in non-toxic range with respect to nitrous acid and
ammonia as well as good pH conditions.
211 5.2
Introduction
Nitrogenous compounds are a common contaminant of a variety of wastewaters.
Inappropriate discharge of these compounds has adverse effects on aquatic systems such
as eutrophication, toxicity to aquatic organisms and depletion of dissolved oxygen in
receiving water bodies (Klees and Silverstein, 1992). Thus, the removal of nitrogenous
compounds is of increasing importance. Several biological and physical-chemical
methods have been applied for nutrient nitrogen removal (Dapena-Mora et al., 2004a).
Among them, biological nitrogen nutrient removal (BNNR) has been proven to be a
simple, reliable and cost effective technology (Hanaki et al., 1990; Tai et al., 2006).
BNNR is generally achieved by means of nitrification followed by denitrification
processes (Tchobanoglous et al., 2003). Nitrification process is carried out by two
sequential reactions of aerobic chemolithoautotrophic bacteria, the ammonium oxidizing
bacteria (AOB) and nitrite oxidizing bacteria (NOB). First, the ammonium (NH4+) is
oxidized to nitrite (NO2-) by AOB such as Nitrosomonas or Nitrosospira (Wullenweber
and Koops, 1980). This step is followed by the oxidation of NO2- to nitrate (NO3-) by
NOB such as Nitrobacter-like organisms (Abeliovich, 1992). During the subsequent
denitrification step, NO2- and NO3- are further reduced to dinitrogen gas (N2) by
denitrifying bacteria. Denitrification is carried out by a wide spectrum of mostly
heterotrophic bacteria that utilize wide variety of organic compounds as an energy source
in the absence of elemental oxygen (Kuenen and Robertson, 1994).
212 Complete nitrogen removal can be achieved from wastewater containing a total chemical
oxygen demand to nitrogen (TCOD: TKN) ratio as low as 7:1 (Third et al., 2005).
Nevertheless, several wastewaters are characterized by high ammonium concentration
and low organic matter content, resulting in low TCOD: TKN ratio. Some examples of
such wastewaters include sludge liquor (Trigo et al., 2006), old landfill leachate (Liao et
al., 2007) and effluents from manure treatment (Wiesmann, 1994) and fish canning
industry (Soto et al., 1991). Sludge liquor from sludge dewatering processes in municipal
wastewater treatment plants contains high levels of ammonium (300-2000 mg NH4+ -N L1
) which account for up to 25% of the total nitrogen load in the raw sewage even though
it contributes only 1-3% of the total influent flow (Janus and vanderRoest, 1997; Fux et
al., 2006; Ahn and Choi, 2007; Jung et al., 2007). Treatment of such wastewaters by
conventional nitrification and denitrification processes involve high oxygen demand due
to the oxidation of NH4+ coupled with adjustment for changes in alkalinity for
nitrification step. Moreover, the denitrification step requires addition of external organic
source such as methanol to sustain the activity of the denitrifying bacteria (Thuan et al.,
2004). These requirements would result in an increase in operational cost. In wastewater
treatment plants designed for nitrogen removal, the aeration cost accounts for about 50%
of the total electrical power consumption of the plant (Ferrer et al., 1998). Therefore,
there is a need to explore alternative process with reduced aeration and external carbon
requirement for treating wastewater rich with ammonium and contains low carbon to
nitrogen ratio.
213 The microbial process of the anaerobic ammonium oxidation (Anammox) is the missing
link in the nitrogen cycle which was initially discovered in a denitrifying reactor in Delft,
The Netherlands (Mulder et al., 1995). The Anammox process involves the oxidation of
ammonium with nitrite as electron acceptor to produce dinitrogen gas under anoxic
conditions. The mass balance of the Anammox reaction has been studied carefully in
sequencing batch reactors (Strous et al., 1998). Based on these studies, the actual
stoichiometry of the Anammox reaction accounting for the cell yield is shown in Eq.1.
The main product of the Anammox reaction is the N2, but a small amount of NO3- is
formed as a byproduct which accounts for about 10.0% of the total fed nitrogen (NH4+
and NO2-). The overall nitrogen balance gives a stoichiometric molar ratio of NO2- to
NH4+ of 1.3. However, minor variations have been reported in literature using various
reactor types. In one study with Anammox gas lift reactor, the nitrite to ammonium molar
ratio was found to be 1.28 (Dapena-Mora et al., 2004a). Strous et al., (1997) reported that
the molar ratio of nitrite to ammonium was 1.4-1.5 and 1.0-1.2 in fludized bed and fixed
bed reactors, respectively.
NH4+ + 1.32 NO2- + 0.066 HCO3- + 0.13 H+ → 1.02 N2 + 0.26 NO3- + 0.066 CH2O0.5N0.15 + 2.03 H2O
[Eq. 1]
Anammox process is a novel, promising and cost-effective alternative to the conventional
nitrification-denitrification processes for nitrogen removal; particularly, for wastewater
with deficiency in organic matter (Vandegraaf et al., 1995). The process requires less
aeration and no addition of external carbon source. Moreover, the low yield of the
Anammox bacteria (0.07 ± 0.01 mol C mol-1 NH4+-N) results in low sludge production
214 (Liao et al., 2008). As a consequence, the Anammox process can save up to 90.00% of
the operational cost as compared to the conventional BNNR processes (Jetten et al.,
2001).
In order to apply the Anammox process for a waste stream rich in ammonium, Anammox
should be combined with a previous partial nitrification process in which around 57% of
the ammonium should be converted to nitrite. This can be achieved either by
manipulating the temperature, hydraulic retention time and temperature using SHARON
process (Single reactor System for High Ammonium Removal Over Nitrite), inhibition of
NOB by free ammonia (Anthonisen et al., 1976) or by lowering the dissolved oxygen
(DO) concentration to avoid the oxidation of NO2- to NO3- due to the higher affinity of
AOB for DO than NOB (Garrido et al., 1997; Wang and Yang, 2004).
It is well documented that Anammox is catalyzed by chemolithoautotrophic bacteria
belonging to the phylum Planctomycetes (Strous et al., 1999b). Common fresh water
species of Anammox bacteria include the frequently encountered, Kuenenia
stuttgartiensis and Brocadia anammoxidans, which are detected in the sludge or biofilms
of anammox bioreactors (Chamchoi and Nitisoravut, 2007; Innerebner et al., 2007; van
der Star et al., 2007). Different Anammox microorganisms have been detected by
molecular probing techniques such as PCR, phylogenetic analysis or FISH in both natural
and engineered systems all over the world (Vandegraaf et al., 1995; Kuypers et al., 2003;
Dalsgaard et al., 2005; Third et al., 2005). These microorganisms have been detected in
several wastewater treatment plants in Netherlands, Switzerland, UK, Germany,
215 Australia, and Japan (Jetten et al., 1998). Egli et al. (2001) reported enrichment of
Anammox culture from rotating biological contactor treating ammonium-rich leachate.
Also, Anammox activity was detected in trickling filter treating wastewater (Schmid et
al., 2000).
The practical application of the Anammox technology is limited by its long start-up
period of the Anammox bioreactor and the availability of the appropriate seed sludge
containing Anammox activity. Anammox bacteria have a slow growth rate. Doubling
times have been reported to be 10.00 to 12.00 days (Strous et al., 1998) and a biomass
yield of 0.13 g dry weight g-1 NH4+-N oxidized (Chamchoi et al., 2008). Thus, they are
difficult to cultivate. The start-up of Anammox process can take months to a year in
laboratory scale reactors (Trigo et al., 2006). The first full scale application of Anammox
process treating sludge digester effluent is in Rotterdam wastewater treatment plant with
a start-up period of approximately two years (Kuenen, 2008). During the reactor start-up,
it is critical to have a source of inoculum in which anammox bacteria are present and to
provide appropriate conditions for their growth in the bioreactor. Conditions during
municipal wastewater treatment create environments where NH4+ and NO2- coexist.
Therefore, it is conceivable that sludge from different unit operations at municipal
wastewater treatment plants contain low levels of intrinsic anammox bacteria. This would
make the application of Anammox process feasible in situations where the Anammox
biomass is not directly accessible. The specific objective is to determine which sources of
inoculum available at municipal WWTPs has the highest level of intrinsic anammox
216 bacteria, and would therefore be most suitable as an inoculum to accelerate the start-up of
the anammox process
217 5.3
Materials and Methods
5.3.1
Sludge sources
Five different inocula were investigated in this study (Table 1). Three of the suspect
inocula were collected from municipal wastewater treatment plants (WWTPs) including
Ina Road (Tucson, AZ) and Green Valley (Green Valley, AZ). The inocula were
collected from different operating unit such as return activated sludge (RAS),
anaerobically digested sludge (ADS) and oxidation ditch. Also, two methanogenic
granular sludges were investigated: Nedalco granular sludge which was obtained from
UASB treating alcohol distillery wastewater (Royal Nedalco BV, Bergen op Zoom, The
Netherlands) and Mahou granular sludge which was obtained from a full-scale UASB
treating brewery wastewater (Mahou beer brewery, Guadalajara, Spain).
218 Table 5.1 Origin, sample and VSS of each inocula
Sample
VSS* (%)
RAS
0.25
ADS
1.21
Green Valley WWTP
Oxidation Ditch
0.42
UASB** reactors
Mahou granular sludge
8.13
Nedalco granular sludge
6.35
Origin
Ina Road WWTP
*
VSS: Volatile Suspended Solids (% of wet weight)
UASB: Up-flow anaerobic sludge blanket
**
5.3.2 Basal media
The basal mineral medium was modified from Dapena-Mora et al., (2004). The medium
was prepared using ultrapure water (Milli-Q system; Millipore) and contained (in g L-1):
NH4HCO3 (0.21); NaNO2 (0.25); NaH2PO4•H2O (0.06); CaCl2•2H2O (0.100);
MgSO4•7H2O (0.20); NaHCO3 (1.05); trace element solution 1 (1.00 ml L-1) and trace
element solution 2 (1.00 ml L-1). The trace element solution 1 contained (in g L-1): FeSO4
(5.00) and Ethylenediamine-tetraacetic acid (EDTA) (5.00). The trace element solution 2
contained (in g L-1): EDTA (15.00); ZnSO4•7H2O (0.43); CoCl2•6H2O (0.24); MnCl2
(0.63);
CuSO4•5H2O
(0.25); Na2MoO4•2H2O
(0.22);
Na2SeO4•10H2O (0.21); H3BO3 (0.01); NaWO4•2H2O (0.05).
NiCl2•6H20
(0.19);
219 5.3.3
Screening assay
The screening assays were conducted in glass serum flasks (250.0 mL) at 30±20C. The
granular sludges (Mahou and Nedalco sludges) and the WWTP sludges were added at 1.5
g VSS L-1 and 10% (v/v); respectively, to serum flasks containing 200.0 mL medium.
The medium contained molar stoichiometric mixture of NO2- and NH4+ of 1.3:1 as well
as basal nutrients, and inorganic carbon (NaHCO3) of 1.0 g L-1. All the flasks were sealed
with butyl rubber stoppers and aluminum caps. The headspace of all the bottles was
flushed with helium/CO2 (80:20%, v:v) for 10 minutes to create an anaerobic
environment and periodically monitored for NH4+ and NO2-. The bottles were spiked with
NO2- and/or NH4+ each time the substrate was consumed. The occurrence of anammox
was confirmed by simultaneous decrease in NH4+ and NO2-. Control bottles containing
inoculum incubated with NH4+ but without NO2- and vice versa were also included for
each inoculum. In addition, abiotic treatment (no active biomass) was included. In
enrichment cultures, the occurrence of anammox can be demonstrated when the NO2consumption is significantly enhanced in treatments containing NH4+ compared to
controls lacking NH4+. All the bottles were incubated in dark to prevent the inhibition by
light.
220 5.3.4
Analytical methods
Nitrite and nitrate were analyzed by suppressed conductivity ion chromatography using a
Dionex 3000 system (Sunnyvale, CA, USA) fitted with a Dionex IonPac AS18 analytical
column (4 x 250 mm) and AG18 guard column (4 mm x 50 mm). The column was
maintained at 35ºC. The eluent used was 10.0 mM KOH at a flow rate of 1.0 mL min-1.
The injection volume was 25.0 μL. Before measurement, all samples were centrifuged at
10,000 rpm for 10.0 min.
Ammonium was determined using Mettler Toledo SevenMulti ion selective meter with
Mettler Toledo selective ammonium electrode (Mettler Toledo, UAS). The pH was
determined immediately after sampling with an Orion model 310 PerpHecT pH-meter
with a PerpHecT ROSS glass combination electrode. Volatile suspended solids (VSS)
was determined according to Standard Methods APHA (2005).
221 5.4
Results and Discussion
Sludge samples from different sources were screened and used as inocula for enrichment
of Anammox bacteria in batch cultures. In enrichment cultures, the occurrence of
Anammox activity was demonstrated when the nitrite consumption was significantly
enhanced in treatments containing ammonium compared to controls without ammonium.
The two methanogenic sludges had similar performance and; therefore, only the Mahou
granular sludge results will be discussed. Table 5.2 summarizes the results of all the
conducted experiments.
5.4.1
Methanogenic granular sludges
Two methanogenic granular sludges, Mahou and Nedalco granular sludges, were
screened for the existence of Anammox bacteria. The nitrite profile of the abiotic (no
inoculums added) treatment of Mahou sludge suggests that the nitrite removal in the
batch culture was due to the biological activity contributed by inoculated sludge (Fig.
5.1a). Neither of the sludges showed any sign of Anammox reaction during the entire
experimental period of 212 days since there was no loss of ammonium during this period
of time (Fig. 5.1b). As can be seen in Fig. 5.1a, the similar removal of nitrite in the
treatment lacking ammonium and the full treatment was due to endogenous denitrifying
activity caused by electron donor released from the decay of inoculated biomass. The
222 ammonium concentration did not decrease but increased in the treatment lacking
ammonium and the full treatment from 5.1 to 20.4 and from 38.3 to 59.3 mg NH4+ -N L-1
in 139 days, respectively (Fig. 5.1b). This increase in ammonium concentration was due
to the biomass decay. The nitrite concentration decreased rapidly in the initial 14 days
(Fig. 5.1a). The maximum denitrification rate was 1.9 mg NO2- -N L-1 d-1 for the
treatment lacking ammonium and 2.1 mg NO2- -N L-1 d-1 for the full treatment. The
denitrification rate of Mahou sludge was gradually declined with the time due to the
exhaustion of endogenous substrate (Fig5.1a). The denitrification rates of the treatment
lacking ammonium and the full treatment were 0.6 and 0.5 mg NO2- -N L-1 d-1 on day
157, respectively. On day 159, the Mahou sludges in all treatment bottles were
transferred to a fresh media. The reason behind this transfer is to ensure that the long
period of incubation in the same media did not cause negative impact on developing the
Anammox activity (such as accumulation of inhibitors like NH3). Also, the initial nitrite
and ammonium concentrations in the new fresh medium were cut in half in case the
initial nitrite concentration was toxic to Anammox bacteria. However, this is most
unlikely since the reported inhibitory level of nitrite to Anammox bacteria reported in
literature is around 100.0 mg NO2- -N L-1 (Strous et al., 1999a) and the initial nitrite
concentration in our study was 50.0 mg NO2- -N L-1. In the next 53 days, there was also
no sign of Anammox reaction. The ammonium concentration did not change in both the
treatment lacking ammonium and the full treatment (Fig. 5.1b). The nitrite concentration
continued to decrease at a lower rate by the endogenous denitrification. The
denitrification rate was measured to be 0.4 mg NO2- -N L-1 d-1 for the treatment lacking
223 ammonium and 0.4 mg NO2- -N L-1 d-1 for the full treatment (Fig. 5.1a). After 212 days,
this batch culture was discarded.
Several literature studies have reported the enrichment of Anammox bacteria from
methanogenic granular sludge (Imajo et al., 2004; Thuan et al., 2004; Liao et al., 2007;
Tang et al., 2009). Planctomycetes were found in anaerobic granular sludge (Sekiguchi et
al., 1998). Tang et al. (2009) reported the start-up of Anammox process in Upflow
biofilm reactor seeded with anaerobic granular sludge collected from a full-scale UASB
reactor treating paper mill wastewater. Our unsuccessful attempt to develop Anammox
enrichment culture from methanogenic granular sludge might be due to the following
reasons. According to Strous et al., (1999) the stagnant condition of the batch culture is
not suitable for the Anammox bacteria. This is because high concentrations of organics or
toxic compounds in the initial seed or generated from the culture are not washed out or
diluted. Therefore, the heterotrophic denitrifires are selected over the Anammox bacteria.
However, the presence of high concentration of organic or toxic compounds in the initial
seed is most unlikely since the seed was washed three times before it was incubated in the
batch culture. Another reason is the low intrinsic level of Anammox bacteria in the
granular sludge. Thus, a very long period of time is needed to enrich such bacteria.
224 Table 5.2 Summary results of all the experiments
Sludge
Type
Source
Duration of
experiment
Days to first
evidence of
Anammox
Fastest
Rate of
nitrogen
removal
d
d
mg N/L.d
NO2- -Nremoved/
NH4+ -Nremoved
Stoichiometric
Ratio
RAS
Ina Rd
BNNR
85.0
38.0
54.6
1-1.2
ADS
Ina Rd
BNNR
268.0
175.6
6.5
1.8
Oxidation
Ditch
Green
Valley
220.0
125.7
13.8
1.5
Mahou
UASB
212.0
No evidence
2.1
--
Nedalco
UASB
212.0
No evidence
4.6
--
225 60
A
NO2- Conc. (mg-N/L)
50
40
30
20
10
0
0
50
100
150
200
250
200
250
Time (days)
100
B
90
NH4+ Conc (mg‐N/L)
80
70
60
50
40
30
20
10
0
0
50
100
150
Time (days)
Figure 5.1 Time course of nitrite and ammonium for Mahou methanogenic granular
sludge. A. Nitrite-N. B. Ammonium-N. Legend: (○), full treatment containing inoculum,
nitrite and ammonium; (□), lacking ammonium; (♦), lacking nitrite; (▲), lacking
inoculum. In plot A, dashed vertical lines indicate respiking of nitrite in full treatment. In
plot B, dashed vertical line indicate transfer of Mahou sludge to new fresh medium.
226 5.4.2 Municipal WWTPs sludges
5.4.2.1 Return activated sludge
The three samples collected from municipal WWTPs showed to be a potential source for
developing Anammox enrichment culture. Figure 5.2 shows that Anammox enrichment
culture can be developed rapidly from RAS obtained from Ina Road WWTP operated for
biological nitrogen nutrient removal (BNNR). Figure 5.2 clearly shows that Anammox
activity can be developed within 40 days and thereafter the culture was enriched further.
Based on the ammonium removal performance (Fig. 5.2b) the experimental period can be
divided into three stages: stage I (day 1-14), stage II (day 14-46) and stage III (46-85).
5.4.2.1.1
Stage I
Initially NO2- was rapidly removed which is truly a biological removal because there was
no loss of NO2- in abiotic control (Fig. 5.2a). NO2- removal was predominantly due to
endogenous denitrifying activity and no evidence of Anammox activity appeared during
this stage since there was no ammonium loss during this time period. The results in Fig.
5.2a show that RAS has strong denitrification activity in this stage. This was evident by
the sharp decrease in nitrite concentrations for both full treatment and treatment lacking
ammonium (Fig.5.2a). The nitrite (50.4 mg NO2- -N L-1) was almost removed completely
in the initial 14 days. The maximum nitrite removal rate for the full treatment and
227 treatment lacking ammonium were 3.9 and 3.4 mg NO2- -N L-1 d-1, respectively. Also, the
ammonium concentration did not decrease but increased during this stage (Fig.5.2b).
These nitrite and ammonium results indicate that the nitrite removal was caused mainly
by denitrification which must have been driven by endogenous substrate in the inoculum.
Endogenous denitrification phenomenon was also reported in literature (Toh et al., 2002;
Jung et al., 2007; Tang et al., 2009). According to Chamchoi and Nitisoravut (2007) the
change of the seed sludge might cause the turnover of the bacteria and thus the former
dormant bacteria might be killed. This results in breakdown of organic nitrogen to
ammonium nitrogen. Also, the experiments conducted by Imajo et al., (2001) showed
that ammonium was produced during the adaptation due to cell decay and starvation.
Ammonium concentration for the full treatment, treatment lacking ammonium and
treatment lacking nitrite increased from 36.5 to 53.6, 3.4 to 12 and 40.2 to 56.0 mg NH4+N L-1, respectively (Fig.5.2b). Also, the dead bacteria release organic matter which is
expressed as chemical oxygen demand (COD) (Dapena-Mora et al., 2004c). This COD
can be used by heterotrophic denitrifyers as an organic carbon source and electron donor
to support their denitrification activity. In one study conducted to enrich Anammox
bacteria from activated sludge in MBR, the COD released from the cell lysis ranged from
87.0 to 191.0 mg L-1 during the first week of operation (Wang et al., 2009). This view
was supported by the fact that no organic matter was added to the mineral medium.
228 70
A
Stage I
NO2- Conc. (mg-N/L)
60
Stage III
Stage II
50
40
30
20
10
0
0
20
40
60
80
100
Time (days)
70
B Stage I
Stage III
Stage II
NH4+ Conc (mg‐N/L)
60
50
40
30
20
10
0
0
20
40
60
80
100
Time (days)
Figure 5.2 Rapid enrichment of an anammox culture from return activated sludge
inoculum obtained at a biological nutrient removal plant treating municipal wastewater
(Ina Road). A. Nitrite-N. B. Ammonium-N. Legend: (○), full treatment containing
inoculum, nitrite and ammonium; (□), lacking ammonium; (♦), lacking nitrite; (▲),
lacking inoculum. In plot A, dashed vertical lines indicate respiking of nitrite in full
treatment, arrows indicate respiking of nitrite in treatment lacking ammonium. In plot B,
dashed vertical lines indicate respiking of ammonium in full treatment.
229 5.4.2.1.2 Stage II
On day 14, the full and ammonium lacking treatments were spiked with nitrite. The idea
behind the spiking is to maintain enough substrate (NO2- and NH4+) for the Anammox
bacteria in order to favor their growth over the other bacteria. During this stage, the
nitrite was still being removed by denitrification activity. However, the activity of the
denitrifying bacteria decreased. This was evident by the decrease of denitrification rate
between days 14 and 36 of both full and lacking ammonium treatments to 1.4 mg NO2- N L-1 d-1 for both (Fig.5.2a). A possible explanation for such decrease could be that the
organic substrate released from the breakdown of the inoculated sludge started to deplete.
On day 38, the biomass showed the first sign of Anammox reaction. As can be seen in
Fig.5.2a, the nitrite removal rate for the full treatment was higher than for the lacking
ammonium treatment. The nitrite removal rates for full treatment and the treatment
lacking ammonium were 2.7 and 1.4 mg NO2- -N L-1 d-1, respectively. In addition, the
ammonium removal at the end of this stage was about 10.6% (Fig.5.2b). The difference
in nitrite removal rate between full and ammonium lacking treatment suggests that it is
most likely that both denitrifiers and Anammox bacteria were responsible for nitrite
removal in the full treatment. It is possible that the decrease in denitrification rate in the
full treatment during this stage led to more availability of nitrite to Anammox bacteria.
Thuan et al., (2004) reported that incomplete denitrification due to low carbon to nitrogen
ratio led to the availability of excess nitrite in the up-flow anaerobic sludge blanket
reactor that could be used by Anammox bacteria. Since the heterotrophic denitrifiers
grow much faster than autotrophic Anammox bacteria; denitrifiers bacteria predominate
230 in the early stages when the flux of endogenous substrate from biomass decay is high.
This phenomenon is also well documented for of slow-growing aerobic nitrifiers being
outcompeted by heterotrophic denitrifiers in the presence of organic matter (Tijhuis et al.,
1994).
5.4.2.1.3 Stage III
From day 42 and onward, the Anammox activity was the dominant process. Fig.5.2
demonstrates clear evidence that the NO2- removal in full treatment was greatly enhanced
by the presence of ammonium. Likewise, ammonium removal was dependent on the
addition of nitrite. Also, comparing the nitrite removal rate in full treatment and the
treatment lacking ammonium, clearly shows that nitrite removal in full treatment is
mainly due to the Anammox reaction since the endogenous nitrite removal had decreased
significantly with the time during this stage. The decrease in the endogenous
denitrification rate was probably due to the depletion of endogenous substrate. The
maximum nitrite, ammonium and total nitrogen removal rates in full treatment were 39.1
mg NO2- -N L-1 d-1, 15.5 mg NH4+ -N L-1 d-1 and 54.6 mg N L-1 d-1, respectively. In
comparison to stages I and II, the ratios of NO2- -Nremoved/ NH4+ -Nremoved decreased to the
range of 1.0-1.2 at the end of this stage (Fig. 5.3). This ratio is slightly lower than the
previous reported values of 1.32 (Strous et al., 1998) and 1.4 (Helmer et al., 2001).
The observed doubling time of the Anammox bacteria was estimated to be 6.8 days.
There is a variation in literature regarding the reported doubling times of Anammox
231 bacteria. The typical doubling times in Anammox bioreactors reported in literature are
10-30 days (Schmidt et al., 2003; Fux et al., 2004; van Niftrik et al., 2004). A doubling
time of 11.0 days was reported by Strous et al. (1998) working with biomass highly
enriched in Anammox bacteria in sequencing batch reactor (SBR). Dapena-Mora et al.
(2004a) estimated the doubling time of Anammox bacteria in gas-lift reactor and SBR
inoculated with enriched Anammox granular sludge as 19.0 and 15.0 days, respectively.
However, by quantitative PCR, the doubling time of Anammox bacteria during the
exponential growth phase in shake flasks was estimated to be 3.6-5.4 days (Tsushima et
al., 2007). Also, van der Star et al. (2008) reported a doubling time of 8.3 days using
granular Anammox sludge in MBR. Such variation in the reported doubling time could
be explained by inefficient retention of biomass in bioreactors except for the case of
MBR. The MBR is well known for its efficiency in retaining the biomass (Wang et al.,
2009). A continuous wash out of small amount of biomass via the effluent leads to a
significantly longer observed doubling time especially in the case of slow-growing
microorganisms such as Anammox (van der Star et al., 2008).
A few studies have been conducted on using the conventional sludge as a seeding
material to start-up the Anammox bioreactors. The choice of the bioreactor is very
important. The bioreactor used to carry out the Anammox process should guarantee
almost complete biomass retention especially during the start-up phase due to the slow
growth rate of the Anammox bacteria (Dapena-Mora et al., 2004a). Dapena-Mora et al.
(2004b) enriched Anammox biomass from municipal activated sludge in SBR. In one
study conducted by Third et al. (2005), Anammox culture was enriched from activated
232 sludge collected from conventional WWTP in western Australia. It is well documented in
literature that the start-up of most Anammox bioreactors using activated sludge as an
inoculum was around four months or more (Toh et al., 2002; Third et al., 2005;
Chamchoi and Nitisoravut, 2007). This long start-up of Anammox bioreactor compared
with our results (within 40.0 days) is probably due to the insufficient biomass build up in
the bioreactor due to the continuous loss of small amount Anammox bacteria via effluent.
Wang et al. (2009) attributed the short start-up (2 months) of the Anammox process in
MBR inoculated with activated sludge to the complete retention of biomass by the MBR.
For successful enrichment of Anammox culture, it is important to inoculate appropriate
seed to batch culture as described by other researchers (Egli et al., 2001; Toh et al.,
2002). The key, therefore, for shortening the long start-up of Anammox bioreactors is to
start the bioreactor in a batch mode to retain all the biomass in order to ensure successful
development of Anammox culture in a short period of time.
After about three months of enrichment, the color of the biomass changed from brown to
red. This phenomenon is well documented in many previous studies on the enrichment of
Anammox bacteria (Jetten et al., 1998; Imajo et al., 2001; Thuan et al., 2004). According
to Third et al. (2005) the red color is attributed to the presence of many hemes in the
hydroxylamine oxidoreductase (HAO) enzyme that is essential for the Anammox
reaction.
233 3.5
Conversion ratio: NO2‐/NH4+
3.0
2.5
2.0
1.5
1.0
0.5
0.0
0
10
20
30
40
50
60
70
80
90
Time (days)
Figure 5.3 The time course of the conversion ratio of nitrite to ammonium for return
activated sludge inoculum obtained at a biological nitrogen nutrient removal plant
treating municipal wastewater (Ina Road). The dashed horizontal line represents the
theoretical ratio of 1.32 based on Anammox reaction.
234 5.4.2.2 Oxidation ditch sludge
Initially the nitrite was rapidly removed which is truly a biological removal since there
was no nitrite removal in abiotic control (Figure 5.4a). The removal of nitrite was most
likely due to the endogenous denitrification. The reason it is not initially due to the
anammox reaction is because no ammonium removal took place (Figure 5.4b). The
ammonium concentration increased in all treatments that had inoculum. The ammonium
concentration increased from 37.9 to 47.8 mg NH4+-N L-1 for the full treatment and from
1.9 to 12.2 mg NH4+-N L-1 for the treatment lacking ammonium (Figure 5.4b). The
maximum nitrite removal rate for full treatment and treatment lacking ammonium were
6.1 and 5.8 mg NO2- -N L-1 d-1, respectively (Figure 5.4a). On days 15 and 24 the full
treatment and treatment lacking ammonium were spiked with nitrite in order to maintain
nitrite to favor the growth of anammox bacteria. The ammonium concentration in both
full treatment and treatment lacking ammonium continued to increase with time which
indicates that endogenous denitrification was still the predominant process responsible
for nitrite removal (Figure 5.4b). On day 31.6, the full treatment and treatment lacking
ammonium were over-spiked with nitrite (81.0 mg NO2--N L-1) by mistake. Such overspiking appeared to have no effects on nitrite removal since the nitrite concentration
continued to decrease with the time. The first evidence of anammox reaction was on day
125.7. The ammonium concentration of in both full treatment and treatment lacking
ammonium decreased from 58.2 to 17.1 mg NH4+-N L-1 and from 20.9 to 2.4 mg NH4+-N
L-1, respectively. Anammox reaction took place in both treatments since both have high
nitrite and ammonium concentrations that are readily available to anammox bacteria to
235 double with the time. From day 125.7 and onward, the anammox reaction was the
predominant process for removing the nitrite in full treatment. Figure 5.4 clearly shows
that the NO2- removal in full treatment was greatly enhanced by the presence of
ammonium. Likewise, ammonium removal was dependent on the addition of nitrite. In
addition, by comparing the nitrite removal rate in full treatment and in treatment lacking
ammonium it is obvious that the anammox reaction was the only process responsible for
the nitrite removal in full treatment since the removal of nitrite due to the endogenous
denitrification in treatment lacking ammonium was almost completely stopped. The
reason for nitrite removal in treatment lacking ammonium was the exhaustion of the
endogenous substrate in the inoculums that support the denitrification process. The
maximum total nitrogen removal rate in full treatment was 13.8 mg N L-1 d-1 (Table 3.2).
The average ratio of NO2- -Nremoved/ NH4+ -Nremoved was 1.5 (Table 3.2).
236 100
NO2- Conc. (mg-N/L)
A
80
60
40
20
0
0
50
100
150
200
250
Time (days)
NH4+ Conc (mg-N/L)
80
B
70
60
50
40
30
20
10
0
0
50
100
150
200
250
Time (days)
Figure 5.4 Rapid enrichment of an anammox culture from oxidation ditch inoculum
obtained at a biological nutrient removal plant treating municipal wastewater (Green
Valley). A. Nitrite-N. B. Ammonium-N. Legend: (○), full treatment containing
inoculum, nitrite and ammonium; (□), lacking ammonium; (♦), lacking nitrite; (▲),
lacking inoculum. In plot A, dashed vertical lines indicate respiking of nitrite in full
treatment, arrows indicate respiking of nitrite in treatment lacking ammonium, dashed
dotted line indicates transfering the inoculums to new medium (day 189). In plot B,
dashed vertical lines indicate respiking of ammonium in full treatment, dashed dotted line
indicates transferring the inoculums to new medium (day 189).
237 5.5
Conclusions
The aim of this study was to screen sludges from different environmental sources to
determine which have the highest intrinsic level of Anammox activity. The results
reported here, show that Anammox cultures can be developed from municipal WWTP
sludges. RAS from treatment plants operated for BNNR had the highest intrinsic level of
Anammox activity. The culture can be developed rapidly in a short period of time
compared to what was reported in literature. The key strategy for the development of
Anammox culture is operating the batch culture in non-toxic range with respect to nitrous
acid and ammonia as well as good pH conditions.
238 5.6
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243 CHAPTER 6
CONCLUSIONS
6.1
Methanogenic Inhibition by Fluoride
•
Fluoride imposes inhibitory effect on acetoclastic and hydrogenotrophic
methanogens in both mesophilic and thermophilic anaerobic digesters.
•
Sensitivity to fluoride varies depending on methanogenic population. Acetoclastic
methanogens are more susceptible to fluoride toxicity than hydrogenotrophic
methanogens. Fluoride is toxic to hydrogenotrophic methanogens at only
relatively high fluoride concentrations.
•
Long-term exposure to fluoride increases the toxicity of fluoride to acetoclastic
methanogens, but not to hydrogenotrophic methanogens. However, some
acetoclastic methanogens either adaptat to fluoride toxicity or more fluoride
tolerant strain of acetoclastic may become enriched over time which results in
decreasing sensitivity to fluoride as in the case of thermophilic acetoclastic
methanogens.
244 •
Higher tolerance of hydrogenotrophic population to fluoride toxicity might be due
to the physiological adaptation to fluoride or the long exposure resulted in a shift
in microbial population which resulted in dominance of hydrogenotrophs that are
less sensitive to fluoride.
•
Initially, acetoclastic methanogens present in granular sludge appear less sensitive
to fluoride toxicity than in the flocculent sludge possibly due to an initial
attenuation of fluoride by calcite present in granular sludge. •
The observed low inhibitory fluoride concentrations to acetoclastic methanogens
is of particular importance since maintaining active acetoclastic population in
anaerobic digesters is critical for stable performance.
245 6.2 Treatment of High Strength Synthetic Sewage in a Laboratory-Scale Upflow
Anaerobic Sludge Bed (UASB) with Aerobic Activated Sludge (AS) PostTreatment
6.2.1
•
Organic matter removal
Combined system of UASB-AS performed an excellent organic removal from
high strength synthetic domestic wastewater (2.5 g COD L-1) with total COD,
VFA and protein removals of 91.0%, 99.9% and 98.2%, respectively.
•
UASB reactor was feasible for (pre) treatment of organic matter from high
strength synthetic domestic wastewater up to HRT of 15.4 hrs and OLR of 3.8 g
COD Lr-1 d-1 (total COD removal was 83.5%).
•
When UASB reactor was overloaded, complete recovery was not achieved due to
loss of biomass from reactor (20% of biomass was lost)
•
Gelatine was highly mineralized to ammonium nitrogen in UASB (NH4+Nformed/protein-N removed = 0.9)
246 •
The aerobic post treatment was able to absorb the high OLR from UASB during
overloading rate (average combined total COD, VFA and protein removals were
91.7%, 99.0% and 97.0%, respectively.
•
Methanogenesis was the limiting step during the overloading event as was evident
by the accumulation of VFA
6.2.2 Nitrogen removal
•
Complete and partial nitrification took place in aerobic post treatment
•
Aerobic post treatment performed well in regard to ammonium removal when DO
concentration exceeded 2 mg L-1 and the OLR was not high (nitrification removal
efficiency was 96.2%).
•
Partial nitrification in aerobic post treatment can be accomplished by either
increasing organic matter concentration or lowering DO concentration below 2
mg L-1. However, controlling DO below 2 mg L-1 is difficult.
•
Nitrogen balance during partial nitrification periods revealed that simultaneous
nitrification/denitrification (SND) occurred; thus, in this study, the occurrence of
partial nitrification is linked to the SND. During the occurrence of partial
247 nitrification the formation of NO2--N and NO3- -N accounted for only 72 % of
NH4+-N removed.
6.3
Nitrate and Nitrite Inhibition of Methanogenesis During Denitrification in
Granular Biofilms and Digested Domestic Sludges
•
NOx- compounds are highly toxic to the methanogenesis.
•
Nitrite is remarkably more toxic to methanogenesis than nitrate and the inhibition
of nitrate to methanogenesis is not related to nitrate itself but it is due to the
formation of its intermediate (e.g. nitrite) during denitrification.
•
As long as the NOx- compounds were present in the media, the methanogenesis
was inhibited. Once the NOx- compounds were completely reduced by
denitrification, the methane production resumed.
•
In this study, acetate was a better electron donor to denitrification than hydrogen
when nitrite was electron acceptor. This was evident by the slow reduction of
nitrite when hydrogen was used compared to acetate.
248 •
The denitrification process out-competed methanogenesis as indicated by the fact
that NOx- compounds were the preferred electron acceptors.
6.4
Enrichment of Anaerobic Ammonium Oxidizing (Anammox) Bacteria from
Wastewater Sludge for Biological Nutrient Nitrogen Removal
•
Anammox bacteria exist in municipal wastewater treatment sludges and
they can be enriched in batch cultures in less than 200 days.
•
Anammox bacteria were not enriched in methanogenic granular sludges
during the time period of the experiment (approx. 200 days).
•
The time required to enrich Anammox bacteria in batch cultures varies
depending on the source of the sludge.
•
The best source at municipal wastewater treatment plant for inoculating
anammox bioreactors to promote a rapid start up is from operating units
where bacteria are exposed to oxic/anoxic conditions such as activated
sludge operating for biological nutrient nitrogen removal.
249 •
A key strategy to enrich Anammox bacteria rapidly is to operate in a nontoxic range with respect to HNO2 and NH3 as well as optimal pH
condition.
•
An enrichment culture of Anammox bacteria was developed 40.0 days
with a doubling time of 6.8 days.
250 SUMMARY
Domestic wastewater in terms of quantity is globally the most abundant type of
wastewater. Direct discharge of untreated domestic wastewater has environmental and
public health risks due to the presence of organic substances, nutrients and pathogens.
Treatment of domestic wastewater by conventional aerobic treatment plants offer
promising results in terms of water quality. However, this technology has several
drawbacks such as high sludge production and high operating cost associated with
aeration. On the other hand, anaerobic treatment offers several inherent advantages over
the aerobic treatment such as low sludge production, low energy consumption and
production of useful energy in the form of methane. Thus, instead of aerobic treatment, a
more cost-effective and sustainable treatment system would involve an anaerobic (pre)
treatment step to remove the majority of the organic pollution at a lower cost and a lower
production of surplus biosolids. An aerobic post treatment step (including pre-existing
aerobic facilities) could then be used to polish the effluents to acceptable levels of water
quality.
Several wastewaters are characterized by high ammonium concentration and low organic
matter content such as sludge liquor from anaerobic sludge digestion. Such wastewaters
are generally treated by means of nitrification followed by denitrification processes
which are costly and troublesome. The recently discovered anaerobic ammonium
oxidation (anammox) process provides considerable savings in cost and energy during
251 nitrogen nutrient removal. Anammox bacteria are autotrophic and catalyze the conversion
of ammonium and nitrite to dinitrogen gas in the absence of oxygen. Although the
specific activity of these microorganisms is high, they have very low growth rates
(doubling time ≈ 10-12 days), which is problematic during reactor start up. Moreover, the
inoculum to start up the anammox bioreactors is not readily available.
The main objectives of this research were:
•
To investigate the feasibility of a combined wastewater treatment system
consisting of a (pre) anaerobic treatment up-flow anaerobic sludge blanket
(UASB) followed by post aerobic treatment activated sludge (AS) for removal of
carbonaceous and nitrogenous contaminants from strong synthetic domestic
wastewater.
•
To evaluate the inhibitory effects of inorganic fluoride, nitrite and nitrate on
methanogenesis in granular biofilms and digested domestic sludges.
•
To determined which sources of inoculum available at municipal wastewater
treatment plants have the highest level of intrinsic anammox bacteria, and would
therefore be most suitable as an inoculum to accelerate the start-up of the
anammox process.
The results indicate the feasibility of UASB for (pre) treatment of high strength synthetic
domestic wastewater (2.5 g COD L-1) for organic matter removal with a HRT of 15.4 h
and organic loading rate (OLR) of 3.8 g COD Lr-1 d-1. , the UASB reactor achieved
removal efficiencies of 83.5 % for the total COD, 74% for volatile fatty acid (VFA) and
94% for protein. The combined system (UASB coupled to post treatment with AS)
252 provided an excellent organic removal pushing the overall removal efficiencies of the
total COD, VFA and protein to 91%, 99.9% and 98.2%, respectively. When the OLR of
the UASB increased to 4.4 g COD Lr-1 d-1, the UASB was overloaded and the effluent
quality of the UASB deteriorated. Despite the effort of decreasing the OLR, the UASB
reactor did not recover completely. At the end of this study, the UASB removal
efficiencies for total COD, VFA and protein were 70.7%, 53% and 96.4%, respectively.
The average combined removal of total COD, VFA and protein were higher than 89%,
99% and 97%; respectively, during the entire period of operation. This high removal
efficiency of the combined system is mainly attributed to ability of post aerobic treatment
to absorb a high OLR from UASB when the VFA accumulated during overloading
events.
With respect to nitrogen removal, both partial nitrification and complete nitrification took
place during aerobic post treatment. The complete nitrification took place when dissolved
oxygen (DO) concentration was > 2 mg L-1 (Period B). During this period, the average
nitrification efficiency achieved was 96.2%. The partial nitrification occurred during
overloading event (period A) and when the DO concentration was lowered below 2 mg L1
(period C). The maximum accumulated nitrite concentration in periods A, B and C were
90, 0.9 and 75.8 mg NO2- -N L-1, respectively. This corresponds to nitrite yield (as
percentage of NH4+ removed) of 73.2% for period A, 0.5 % for period B and 50.9 % for
period C. The nitrogen balance during partial nitrification (periods A and C) revealed that
there was a discrepancy between the amount of ammonium nitrogen removed and the
amount of oxidized nitrogen formed. The amount of oxidized nitrogen formed (NO2--
253 Nformed and NO3--Nformed) for periods A and B accounted for 73% and 72% of the NH4+-N
removed,
respectively.
This
suggests
the
occurrence
of
simultaneous
nitrification/denitrification (SND) in aerobic post treatment.
With regard to fluoride toxicity to methanogenesis, it was found that the sensitivity of
methanogenesis to fluoride varied depending on methanogenic population examined.
Acetoclastic methanogens present in mesophilic and thermophilic sludges were sensitive
to fluoride. The sensitivity of acetoclastic methanogens in granular sludge to fluoride
increased with the time of exposure. In contrast, the hydrogenotrophic methanogens in
mesophilic and thermophilic sludges tolerated relatively high concentration of fluoride.
The toxicity of fluoride to both mesophilic and thermophilic hydrogenotrophic
methanogens decreased with increasing time of exposure.
Methanogenic activity was inhibited by the presence of NOx- compounds (i.e. nitrite and
nitrate). The inhibition imparted by nitrate was not due to the nitrate itself, but instead
due to its reduced intermediates (e.g. nitrite). The toxicity of NOx- to methanogens was
found to be reversible after all the NOx- was reduced during denitrification. The recovery
of the methanogenic activity was nearly complete at low NOx- concentrations; while the
recovery was only partial at high NOx- concentration. It was also found that the assay
substrate has a significant impact on the nitrite metabolism rate. Hydrogen reduced nitrite
slowly and thus the toxicity of nitrite was greater to methanogens compared to when the
acetate was used as a substrate.
254 Anammox enrichment cultures were developed from samples collected from municipal
wastewater treatment plants. However, none of the methanogenic granular sludges
collected from UASB reactors showed any sign of Anammox activity. It was found that
return activated sludge (RAS) from Ina Road wastewater treatment plant (Tucson, AZ)
which was operated for biological nitrogen removal had the highest intrinsic level of
Anammox activity. RAS enrichment culture developed rapidly within 40 days with a
doubling time of approximately 6.8 days.
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