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BIOLOGICAL TREATMENT OF HIGH SALINITY WASTEWATER USING YEAST
AND BACTERIAL SYSTEMS
by
Nguyen Phuoc Dan
A dissertation submitted in partial fulfillment of the requirements for the degree of Doctor of
Engineering
Examination Committee: Prof. C. Visvanathan (Chairman)
Prof. Chongrak Polprasert (Co-chairman)
Prof. Nguyen Cong Thanh
Dr. Josef Trankler
Dr. Sudip K. Rakshit
External Examiner: Prof. Ronald E. Simard
Nationality: Vietnamese
Previous Degree: Bachelor of Engineering (Civil)
Hochiminh City University of Technology (HUT)
Hochiminh City, Vietnam
Master of Engineering (Environmental Engineering)
AIT, Thailand
Scholarship Donor: Swiss Development Cooperation (SDC)
Asian Institute of Technology
School of Environment, Resources and Development
Bangkok, Thailand
December 2001
ii
Acknowledgements
The author wishes to deeply express his gratitude to his advisor, Prof. C. Visvanathan
for kindly giving valuable guidance, suggestions and encouragement through his study in
AIT. He would like to express his appreciation to his co-advisor, Prof. Chongrak Polprasert
for his valuable comments and suggestions provided throughout the research work. The
author wishes to express deepest sincere thanks to Prof. Nguyen Cong Thanh, Dr. Josef
Trankler, Dr. Sudip K. Rakshit and Dr. A. Sathasivan for their valuable comments, critical
ideas and serving as members of examination committee.
A special thank is addressed to Prof. Ronald E. Simard for kindly accepting to serve as
External Examiner His constructive and professional comments are highly appreciated.
The author gratefully acknowledges Swiss Development Cooperation (SDC)-
EPFL,IGE/GS for his financial support. Grateful acknowledgement is also extended to
Nishihara ERSC. for supporting partially experimental equipment. Also acknowledgement is
given to the SERD school for financial support on attendance of Conference in Malaysia.
The author is very grateful to Mdm. Visvanathan, Mr. Jonathan Shaw and Mr. Basu
for providing comments and editing in English language.
A special thank is extended to Lab Supervisors, Mr. Suwat, Ms. Salaya, Mr. Peter and
Mr. Chai, technicians Khun Verin, Khun Tam and others. The author would like to thank his
friend, Master student, M.M. Cho, for co-operation of thesis works.
The author is most grateful to his family and CEFINEA’s Director, Prof. L.M. Triet,
for mental support during study in AIT.
iii
Abstract
This study aimed to compare the performance of aerobic treatment using wild mixed
yeast and bacterial culture for high salinity wastewater. The operating conditions of yeast
treatment under high salinity such as pH, sludge retention time (SRT) and dissolved oxygen
(DO) were examined. The comparative evaluation is based on determination of biokinetic
coefficients using the respirometric method and treatment efficiency of long-term operation of
two laboratory-scale membrane bioreactor systems.
The biokinetic experiments reveal that yeast culture has a lower observed maximum
specific grow rate ( obs) at low salt content (20g/L) than that of bacteria. But obs of yeasts at
higher salt contents (above 30 g/L) did not decline dramatically and had higher value than that
of bacteria. The osmotolerant yeast mixture was able to tolerate a wider pH range than
bacterial culture. The chemical oxygen demand (COD) removal rate of the yeast mixture was
highest at pH values 5.0-5.5.
Two laboratory-scale membrane bioreactor systems were investigated to treat high
salinity wastewater containing high organic (5,000 mg/L COD) and salt content (32 g/L
NaCl), namely: the Yeast Membrane Bioreactor (YMBR), and Yeast pretreatment followed
by Bacterial Membrane Bioreactor (BMBR). In the YMBR system, experimental runs were
conducted with a mean biomass concentration of 12 g MLSS/L. Here, the maximum COD
removal rate of 0.93 g COD/g MLSS.day was obtained at F/M of 1.5 g COD /g MLSS.d,
whereas the BMBR system was operated with a biomass concentration of up to 25 g
MLSS/L, resulting in maximum COD removal rate of 0.32 kg COD /kg MLSS.day at F/M
ratio of 0.4. In comparison the BMBR, the YMBR could obtain higher COD removal rate at
higher organic loading, indicating the potential of the yeast reactor system to treat high
salinity wastewater containing high organic concentration.
Transmembrane pressure in the BMBR was progressively increased from 2 to 60 kPa
after 12d, 6 d and 2 d at hydraulic retention time (HRT) of 14h, 9 h and 4h, with average
biomass concentration of 6.1, 15 and 20 g MLSS/L respectively. By contrast, the
transmembrane pressure in YMBR was only increased from 2 to 60 kPa only after 76 days of
operation, with an average biomass concentration of 12 MLSS/L and an operating HRT range
of 5 - 32 h.
The comparative evaluation of treatment performance of both YMBR and BMBR with
the low organic-feed wastewater (1,000 mg/L COD and 32 g/L NaCl) was examined. COD
removal of both processes were above 90% at HRT of 5 h. Under the same operating
conditions, the YMBR could run under transmembrane pressure 10 times lower than the
BMBR with a significantly reduced membrane fouling rate. This may be due to low
production of adhesive extracellular polymers (ECP) and the secondary filtration layer formed
from large free yeast cells. ECP production of bacterial sludge was increased considerably at
high salt contents and high sludge retention time (SRT). For the bacterial sludge, the increase
salinity led to increase in ECP value, whereas the ECP content of the yeast sludge was
relatively very small.
iv
Table of Contents
Chapter Title Page
Title page i
Acknowledgements ii
Abstract iii
Table of Contents iv
List of Figures vii
List of Tables x
Abbreviations xii
1 Introduction
1.1 Background 1
1.1.1 Environmental Concerns 1
1.1.2 Effects of High Salinity on Biological Treatment Processes 2
1.1.3 Salt-Tolerant or Halophilic Microorganisms 3
1.1.4 Membrane Bioreactor Process 3
1.2 Objectives of the Study 4
1.3 Scope of the Study 4
2 Literature Review 6
2.1 Introduction 6
2.1.1 The Seafood Processing Industry 6
2.1.2 Pickled Vegetable Processing 9
2.1.3 Other Saline Wastewaters 13
2.2 Effects of High Salinity on Biological Waste Treatment Process 13
2.2.1 Aerobic Treatment 14
2.2.2 Anaerobic Treatment 17
2.2.3 Nutrient Removal 18
2.3 Application of Halophilic Bacteria for Saline Wastewater Treatment 19
2.4 Yeasts 22
2.4.1 General 22
2.4.2 Applications of Yeasts for Wastewater Treatment 24
2.5 Theoretical Modeling Consideration 33
2.5.1 Growth without Inhibition 33
2.5.2 Growth with Inhibition 35
2.6 Respirometric Method 38
2.6.1 Respirometer 38
2.6.2 Experimental Procedure 38
2.6.3 Determination of Kinetic Constants 40
2.7 Membrane Bioreactor (MBR) 41
2.7.1 Advantage of the MBR Process 42
2.7.2 Main Design Parameters 42
2.7.3 Membrane Fouling 45
3 Methodology
3.1 Biokinetic Study 49
3.1.1 Seed Sludge 51
3.1.2 Acclimation 52
3.1.3 Biokinetic Experiments 52
3.2 Parametric Study 54
v
3.2.1 pH values 54
3.2.2 Sludge Retention Time (SRT) 55
3.3 Biomembrane Study 56
3.3.1 High COD loading 56
3.3.2 Low COD loading 59
3.4 Sludge Characterization Study 60
3.5 Analytical Methods 60
4 Results and Discussion
4.1 Biokinetic Study 62
4.1.1 Enrichment and Acclimation of Yeast and Mixed Bacterial Sludge 62
4.1.2 Evaluation and Comparison of Biokinetic Coefficients 69
4.2 Parametric Study 73
4.2.1 DO and pH 74
4.2.2 Nitrogen Variation in Mixed Yeast and Bacterial Cultures 78
4.2.3 Effect of SRT on COD and Nitrogen Removal 80
4.3 Biomembrane Study 81
4.3.1 High COD loading 82
4.3.2 Low COD loading 87
4.4 Sludge Characterization Study 93
4.4.1 Culture Study 94
4.4.2 YMBR and BMBR 96
4.4.3 Microscopic Observations of Mixed Yeast Sludge 97
4.4.4 Nutrient Uptake 98
5 Conclusions and Recommendations
5.1 Conclusions 100
5.2 Recommendations 101
Appendix A: Pictures of Experiments A-1
Appendix B: Experimental Data of Acclimation B-1
Appendix C: Experimental Data of Biokinetic Study C-1
Appendix D: Experimental Data of Parametric Study D-1
Appendix E: Experimental Data of Biomembrane Study E-1
vi
List of Figures
Figure Title Page
2.1 Flow diagram of steamed canned shrimp processing 7
2.2 Flow diagram of Dried and Salted fish processing 9
2.3 Flow diagram of kim chi pickles processing 11
2.4 Variation of COD removal rate with salt contents (Kargi and Uygur, 1996) 15
2.5 Diagram of a nitrogen treatment system 18
2.6 Schematic diagram of percolation reactor (Kargi and Uygur, 1996) 20
2.7 Variation of COD removal rate (R) as function of salt content (Kargi and Dincer,
2000) 20
2.8 Schematic diagram of the biofilter and trickling filter treatment system
(Yang et al., 2000) 21
2.9 Diagram of yeast cell (Salle, 1961) 22
2.10 Budding is a common reproductive process in yeasts 23
2.11 True mycelium (formed by fission) and pseudomycelium (formed by budding) 23
2.12 Specific growth rate of Candida ingens vs DO and VFA concentration
(Anciaux et al., 1989) 25
2.13 Traditional carbon and nitrogen removal system can be altered with anaerobic
and yeast treatment system (Ortiz et al. 1997) 27
2.14 Schematic diagram of the Yeast Cycle System (YCS) 28
2.15 Comparison between Yeast Cycle System (YCS) and complete mixed activated
sludge (AS) (Nishihara ESRC Ltd., 2001) 29
2.16 SCP from confectionery effluent (Gray, 1989) 31
2.17 The Symba process (Gray, 1989) 31
2.18 Growth curve of microorganisms in a culture 33
2.19 The effects of a limiting substrate on the specific growth rate (Monod model) 34
2.20 Curves of inhibition growth models (n =1: Ghose and Tyagi; n= 0.5: Bazua and
Wilke model) 35
2.21 Curves of substrate inhibition growth models 36
2.22 pH and DO models 37
2.23 Schematic diagram of respirometer 38
2.24 Recorder chart with a typical respirogram (Cech et al., 1984) 39
2.25 OUR response in respirometer (Ekama, et al., 1986) 40
2.26 Diagram of membrane bioreactor processes 42
2.27 Diagram of fouling mechanisms (adsorption and deposition) 45
2.28 Schematic illustration of membrane biofouling process (Ridgway and Flemming,
1996). 46
2.29 Schematic diagram of biofloc or biofilm 47
3.1 Flowchart of different phases of experimental study 49
3.2 Flowchart of biokinetic experiments 51
3.3 Schematic diagram of enrichment procedure 51
3.4 Respirometer set-up 53
3.5 Membrane reactor systems in the high COD loading 58
3.6 Schematic diagram of biomembrane reactor 59
4.1 Appearance of yeast cells predominantly grown in glucose-feed wastewater 62
4.2 Acclimation of yeast sludge cultured with glucose at high salt contents 63
vii
4.3 Acclimation of microbial mixed culture with glucose-feed wastewater as
function of salt 64
4.4 Typical COD and COD removal profile of mixed yeast batch in glucose-feed
wastewater at 32 g salt/L 65
4.5 Variation in COD removal rate versus salt contents in acclimatized yeast and
bacterial mixed cultures 66
4.6 Acclimation of yeast and bacterial sludges to fish-protein-feed wastewater
containing 32 g/L salt 67
4.7 Predominance of wild yeast strains in the cultures fed with fish-protein wastewater
(at 32 g/L salt) 68
4.8 OUR curves of mixed yeast and bacterial sludges feed with 50 mg/L COD and
32 g/L salt (glucose-feed wastewater) 70
4.9 OUR curves of mixed yeast and bacterial sludges feed with 100 mg/L COD and
32 g/L salt (protein-feed wastewater) 70
4.10 Variation in specific growth rate of yeast sludge as function of COD at different
salt contents for glucose-feed wastewater 71
4.11 Variation in specific growth rate of bacterial culture as function of COD
concentration at different salt contents for glucose-feed wastewater 71
4.12 Inhibition effect of salt contents on mixed yeast and bacterial cultures on glucose-
feed wastewater 73
4.13 Inhibition effect of salt contents on mixed yeast and bacterial cultures on protein-
feed wastewater 73
4.14 DO and COD changes of yeast batch fed with glucose and protein wastewater at
32 g salt/L 74
4.15 DO and COD changes of mixed bacterial batch fed with glucose and protein
wastewater at salt content of 32 g/L 75
4.16 pH changes of yeast culturefed with glucose and protein wastewater at
32 g salt/L 75
4.17 pH changes of mixed bacterial batch fed with glucose and protein wastewaters at
32 g salt/L 76
4.18 Variation in OUR as funtion of initial pHs for mixed yeast fed with protein
wastewater at 32 g salt/L 77
4.19 Variation in OUR as funtion of initial pHs for mixed bacterial fed with glucose
wastewater at 32 g salt/L 77
4.20 Variation in nitrogen components as funtion of time in the mixed yeast at 32 g
salt/L NaCl (Nitrite and nitrate concentration of both feed wastewaters were not
dectected) 79
4.21 Variation in nitrogen components vs. time in the mixed bacterial culture at
32 g salt/L NaCl 79
4.22 Variation in MLSS as funtion of SRT 80
4.23 Variation in COD, nitrogen removal and MLSS in funtion of SRT in mixed yeast
culture at VLR of 5 kg COD/m3.d (32 g salt/L) 81
4.24 Variation in flux as function of membrane transmembrane pressure (Viscosity of
water at 26oC = 8.70 x 10
-4 kg/m.sec) 82
4.25 Variation in COD, biomass and transmembrane pressure in the YMBR as
function of volumetric loading 83
4.26 Variation in COD, biomass and transmembrane pressure in the BMBR as
function of volumetric loading 84
4.27 Variation in COD removal in function of volumetric loading rate 86
4.28 Variation in COD removal rate in function of F/M ratio (initial COD = 5,000 mg/L) 86
4.29 Variation in COD, biomass and transmembrane pressure in the YMBR as
function of volumetric loading 88
viii
4.30 Variation in COD, biomass and transmembrane pressure in the BMBR as
function of volumetric loading 89
4.31 Variation in COD removal as function of HRTs in YMBR and BMBR 89
4.32 Variation in specific growth rate of yeast and bacteria at 32 g salt/L in function
of COD 90
4.33 Possible mechanisms for flux enhancement by yeast cells 93
4.34 Variation in ECP and CST in function of salt content 95
4.35 Variation in SVI, SS, ECP and viscosity with salt content in mixed bacterial
cultures 96
4.36 ECP contents of mixed yeast and bacterial sludges in YMBR and BMBR 97
ix
List of Tables
Table Title Page
1.1 Comparison of pollutant loads from seafood processing and other industries in the
Saigon-Dong Nai river catchment area (DOSTE-HCMC and CEFINEA, 1998) 2
2.1 Characteristics of herring brine waste (Balslev-Olesen et al., 1990) 7
2.2 Characteristics of wastewater from the dried salted fish plant (Dan, 2000) 8
2.3 Composition of brines used for canning vegetables (Joslyn and Timmons, 1967) 9
2.4 Raw waste loads and quality of wastewater from some pickling industries 10
2.5 Characteristics of the waste brine from four different kim chi factories located in
Suwon city and Kyunggi province, Korea (Park and Choi, 1999) 10
2.6 Wastewater characteristics of from various fishery product and vegetable
pickling industries 12
2.7 Characteristics of oil field brine (Dalmacija et al., 1996) 13
2.8 Characteristics of leachates (Pirbazari, 1996) 13
2.9 Adverse effects of high salinity in activated sludge process 16
2.10 Adverse effects of high salinity in anaerobic treatment processes 18
2.11 Summary of adverse effects of high salinity in nutrient removal processes 19
2.12 Effects of using halophilic bateria for high salinity wastewater treatment 22
2.13 Basic composition of Candida utilis yeast biomass (Defrance, 1993) 24
2.14 A comparison between yeast and anaerobic treatment process (Defrance, 1993) 27
2.15 Operating conditions of YCS (Nishihara ESRC Ltd., 2001) 28
2.16 Quality of treated water and efficiency of the YCS for seafood processing
wastewater treatment (Nishihara ESRC Ltd., 2001) 29
2.17 Summary of studies on yeast treatment of high salinity wastewater 30
2.18 Kinetic models for inhibition growth (Han and Levenspiel, 1988) 35
2.19 Comparison between biological performances of MBR process and conventional
AS process 44
3.1 Composition of glucose-feed wastewater (Defrance, 1993) 50
3.2 Composition of protein-feed wastewater 50
3.3 Operating conditions for high salinity acclimation 52
3.4 Operating conditions for the respirometric experiments 53
3.5 Operating conditions for the pH effect experiments 54
3.6 Operating conditions of the experiments on SRT effect 55
3.7 Difference between the high COD loading and low COD loading 56
3.8 Experimental operating conditions of YMBR and BMBR systems 57
3.9 Composition of the low COD wastewater 59
3.10 Effects of different HRTs and SRTs on yeast and bacterial membrane reactors 60
3.11 Operating conditions for the sludge characterization study 60
3.12 Parameters and their analytical method 61
4.1 Performance of mixed yeast and bacterial batches adapted to glucose-feed
wastewater with high salt 64
4.2 Performance of mixed yeast and bacterial sludges adapted to protein-feed
wastewater with high salt contents (Initial COD cof 5,000 mg/L). 69
4.3 Biokinetic coefficients of the yeast and bacterial sludges at different salt contents
for glucose and protein-feed wastewaters 72
4.4 Variation of parameters during various SRTs (Initial COD of 5000 mg/L) 81
x
4.5 Operating parameters of the YMBR, BMBR, some yeast treatments, MBR
processes treating different wastewaters and conventional AS system 85
4.6 Operating parameters and performance of YMBR and BMBR in high COD
loading phase 91
4.7 Values of different parameters during YMBR and BMBR filtration cycle 92
4.8 Yeast and bacterial sludges characterization 94
4.9 Composition of mixed bacterial and mixed yeast sludge 98
xi
List of Abbreviations
AF Anaerobic Filter
AS Activated Sludge
BOD Biochemical Oxygen Demand
BMBR Bacterial Membrane Bioreactor
COD Chemical Oxygen Demand
CST Capillary Suction Time
DO Dissolved Oxygen
DOSTE Department of Science, Technology and Environment
ECP Extracellular Polymers
EPS Extracellular Polymer Substances
ESRC Environmental Sanitation Research Center
F/M Food/Microorganism ratio
HCMC Hochiminh City
HRT Hydraulic Retention Time
J Permeate flux
MBR Membrane Bioreactor
MF Microfiltration
MLSS Mixed Liquor Suspended Solids
MLVSS Mixed Liquor Volatile Suspended Solids
N Nitrogen
NH3-N Ammonia Nitrogen
NO2-N Nitrite Nitrogen
NO3-N Nitrate Nitrogen
OUR Oxygen Uptake Rate
P Phosphorus
SBR Sequencing Batch Reactor
SCP Single-Cell-protein Production
SRT Sludge Retention Time
SS Suspended Solids
SSL Spent Sulphite Liquor
SVI Sludge Volume Index
TDS Total Dissolve Solid
TOC Total Organic Carbon
TKN Total Kjedahl Nitrogen
TS Total Solids
TVS Total Volatile Solids
U Substrate Utilization Rate
UASB Upflow Anaerobic Sludge Blanket
UNEP United Nations Environment Programme
UF Ultrafiltration
VFA Volatile Fatty Acid
VLR Volumetric Loading Rate
VOC Volatile Organic Carbon
VSS Volatile Suspended Solids
Y Yield coefficient
YCS Yeast Cycle System
YR Yeast Reactor
YMBR Yeast Membrane Bioreactor
P Transmembrane Pressure
1
Chapter 1
Introduction
1.1 Background
High salinity wastewater containing high inorganic salt content is mostly generated
from industries such as seafood processing, vegetable canning, pickling and cheese
processing. Among these, the seafood processing industry is an industrial sector that produces
large volumes of saline wastewater with high organic and nutrient concentration. Therefore it
causes heavy pollution to receiving waters. At present, the seafood processing industry plays
an important role in South East Asia’s economy. Under stringent environmental regulations,
this industry is now facing both high treatment costs and problems in the operation of
conventional wastewater treatment plant. These operational problems are linked to high
organic loading, high salt content and very large seasonal variation leading to change in waste
characteristics.
1.1.1 Environmental Concerns
In seafood processing, the main environmental concern is the use of large amounts of
fresh water for processing, including washing raw material and products, for cleaning of
machines, containers or flushing the working floor, for de-icing, thawing and salt soaking. In
general, 90-95% of water consumed is converted into highly polluting wastewater. Frozen
seafood processing consumes particularly large volumes of water, ranging from 70 to 120
m3/ton of product, the equivalent of 32-60 m
3/ton of raw fish (DOSTE-HCMC and
CEFINEA, 1998). The wastewater generated by fish processing factories has high loads of
organic and nutrients. This waste is commonly discharged directly into coastal areas. Another
important aspect of this industrial waste is its high salinity (Na+, Cl
-, SO4
2-), caused both by
the raw materials and seawater used in various processes. Here, using pre-filtered seawater for
processing leads to high salinity in the wastewater, which reduces the biodegradation rate in
effluent treatment units (Mendez et al., 1992).
Because factories process a broad range of products with large seasonal variation,
pollution characteristics vary significantly both from plant to plant, and even within the same
plant. In Ho Chi Minh City (HCMC), the seafood processing sector is one of the major
industrial contributors to the heavy pollution to receiving waters. The average BOD5 generally
ranges from 1,200 to 1,800 mg/L (COD of 1,600 - 2,300 mg/L) (DOSTE, 1994). In addition,
the wastewater contains high levels of suspended solids (150-200 mg/L), and is rich in
nutrients with total nitrogen ranging from 70 to 110 mg/L. The pollutant loads from the
seafood processing industry and other industries in the Saigon-Dong Nai river catchment area
is shown in Table 1.1. These data indicate that seafood processing is a sector that causes
considerable pollution to the environment in this part of Viet Nam.
2
Table 1.1 Comparison of pollutant loads from seafood processing and other industries in the
Saigon-Dong Nai river catchment area (DOSTE-HCMC and CEFINEA, 1998)
Industry
Flow rate
m3/day Pollution load, kg/day
SS BOD5 TKN
Seafood processing 18,900 4,200 28,400 1,700
Pulp and paper 49,200 54,900 104,800 340
Cassava 47,100 30,600 590,000 NA
Textiles & dyeing 32,500 5,600 17,300 NA
Beverages 15,600 4,400 19,000 630
Latex processing 11,600 2,500 86,600 2,800Meat processing and slaughterhouse 6,400 4,000 13,300 1,020
Sugar (sugar cane) 5,520 6,900 32,000 72
Vegetable canning 3,700 520 2,700 70SS – suspended solids
NA – None available
1.1.2 Effects of High Salinity on Biological Treatment Processes
Past studies on saline wastewater treatment reveal that salinity decreases BOD5 removal
efficiencies, increases effluent turbidity due to sludge settling in the secondary sedimentation
unit, solid losses, and changes in the mixed liquor floc protozoan population in an activated
sludge system (Dalmacija et al., 1996; Woolard and Irvine, 1995; Kargi and Dincer, 1998).
Kargi and Uygur (1996) reported many adverse effects of salt on aerobic attach growth such
as trickling filter and rotating biological contactors. The efficiency of COD removal decreased
significantly with increases in salt contents over 20g/L.
The anaerobic digester were much more sensitive to chlorides than activated sludge
processes (Burnett, 1974). Biogas production and COD removal of anaerobic treatment
processes such as anaerobic filter, UASB and batch reactor were inhibited significantly at salt
content above 30g NaCl/L (Baere et al., 1984; Feijoo et al., 1995). In addition, high salt
content also depressed the treatment ability of nitrifying and denitrifying bacteria, even
though pre-acclimation had been done (Dahl et al., 1997; Panswad and Anan, 1999).
The adverse effects of high salinity on conventional biological processes can be
attributed to high osmotic stress or inhibition of the reaction pathways in the organic
degradation process. In addition, high salt content induces cell lysis, which increases effluent
solids. The population of protozoa for proper flocculation is also significantly reduced at high
salt contents. Here, although salt acclimation can be expected from conventional processes,
the extent of adaptation is limited, and thus conventional processes can not be used to treat
wastewaters containing more than 3% salt (Woolard and Irvine, 1995).
Currently many saline wastewater treatment plants are able to overcome the technical
problems associated with high salinity by diluting the saline waste stream with fresh water.
Nevertheless, this practice is unsustainable, due to continuous pressure on the industries to
reduce fresh water consumption.
3
1.1.3 Salt-Tolerant or Halophilic Microorganisms
In order to improve organic and nitrogen removal efficiency, application of salt-tolerant
microorganisms in biological treatment of saline wastewater has been investigated
experimentally by several researchers (Nishihara ESRC Ltd., 2001; Woolard and Irvine,
1995; Hinteregger and Streichsbier, 1997; Park and Choi, 1999; Kargi and Dincer, 2000). Salt
tolerant microorganisms are those which can tolerate high salt content during their growth.
This utilization of halophilic microorganisms (e.g. Halobacter halobium) along with activated
sludge culture resulted in better treatment performances at salt contents above 2% (Kargi and
Dincer, 2000). Woolard and Irvine (1995) studied the treatment of hypersaline wastewater by
a moderate halophilic bacterial mixture isolated from soil of a saltern and fed in sequencing
batch reactor. They found that over 99% phenol removal was possible from 15% saline
wastewater.
In investigating the application of yeast in the food processing wastewater treatment,
researchers have investigated this potential in the treatment and reuse of wastes containing
solids and high concentrations of salt, fat and antibiotics. Park and Choi (1999) studied the
possibility of culturing an osmotolerant yeast, Pichia guilliermondii A9, using waste brine
from Kim Chi factory. The growth of Pichia guilliermondii A9 in waste brine was not
inhibited by NaCl concentrations of up to 60 g/L. In the Yeast Cycle System, wild yeasts were
utilized for treatment of wastewater, and the recovered excess sludge could be reused
(Nishihara ESRC Ltd., 2001). This yeast system is used as a pre-treatment to reduce the
organic pollutants, followed by a conventional activated sludge process. Primary treatment
plays an important role in removal of high organic and nutrient loadings. Moreover, excess
yeast sludge that had high protein, vitamins, lipid content could be used as animal feedstuff,
mushroom growing or as fertilizer.
1.1.4 Membrane Bioreactor Process
Application of the membrane bioreactor (MBR) concept in high salinity wastewater
treatment offers the possibility of overcoming low biodegradation rate and poor sludge
settling in the secondary sedimentation tank. MBR process can be operated at high MLSS and
thus organic removal can be improved. This results in sludge wastage and plant size reduction
(Visvanathan et al., 2000). Moreover, the selection of microorganisms present in the
membrane bioreactor is no more dependent on their ability to form biological flocs and
settling characteristics.
However, the membrane fouling problems lead to rapid flux reduction in MBR. This
secondary effect results in increase in energy consumption, and more frequent chemical
cleaning is required. These major problems hinder the widespread application of MBR in
effluent treatment processes. The membrane fouling might be the result of (a) the biofilm
growth or attachment of bacterial flocs on the upper surface of the membrane, and (b) the
deposition of macromolecules at the pore entrances or within the internal pore structure of the
membrane. The macromolecules can be protein from wastewater, extracellular polymers
(ECP) or long chain organic by-products generated during the biodegradation process.
4
1.2 Objectives of the Study
The overall objectives of this research were:
(1) To evaluate variation of biokinetic coefficients of salt-tolerant yeast mixture and
bacterial mixture with high salt contents.
(2) To find out suitable operating parameters for membrane bioreactor systems using salt-
tolerant yeast and bacterial mixture to treat saline seafood processing wastewater.
(3) To investigate membrane fouling of both microorganisms in terms of sludge
characteristics and ECP production.
1.3 Scope of the Study
To accomplish the above objectives, four studies were carried out:
(1) Biokinetic study. Biokinetic coefficients of mixed yeast and mixed bacterial treatment at
high salt contents were evaluated using respirometric experiments. The three salt
contents examined were 20, 32 and 45 g/L NaCl. Two feed wastewaters were used,
namely glucose-feed wastewater and fish-protein-feed wastewater. In the fish-protein-
feed wastewater, commercial tuna fish protein extract was mixed to obtain wastewater
composition similar to tuna fish processing wastewater. Whereas, glucose-feed
wastewater was composed of glucose as carbon source and inorganic ammonia as the
nitrogen source.
(2) Parametric study (optimization of operating conditions). The fish-protein wastewater
with salt content of 32 g/L was used in this study.
a. Optimum pH for mixed yeast and bacterial treatment at a salt content of 32 g/L was
evaluated in terms of oxygen uptake rate (OUR) by respirometric experiments.
Based on theresponse of maximum OUR at different pH values, the optimum pH
range was determined.
b. The variation of COD, DO and nitrogen (organic nitrogen, ammonia, nitrite and
nitrate) with aeration time in acclimatized mixed yeast and bacterial cultures was
monitored. Based on these COD, nitrogen profile data, suitable hydraulic retention
time (HRT), organic removal rate and nutrient uptake or nitrogen removal were
evaluated.
c. Five sludge retention times (5, 7, 10, 20 and 45 days) were investigated for mixed
yeast treatment. Every the sludge retention time (SRT) experiment was conducted
in two-liter batch reactors with fill-and-draw operation. Based on the COD and
nitrogen removal, optimum SRT was suggested.
(3) Biomembrane study. This study consisted of two phases:
a. High COD loading: Two parallel experimental set-ups were carried out, namely (1)
Yeast pretreatment followed by Bacterial Membrane Bioreactor (BMBR), and (2)
the Yeast Membrane Bioreactor (YMBR). Fish-protein-extract wastewater with
feed COD of 5,000 mg/L and salt of 32 g/L was used. The experimental
investigations were carried out by step-wise increase in volumetric loading at SRT
of 50 d.
b. Low COD loading: The fish-protein wastewater with 1,000 mg/L COD and 32 g/L
NaCl was used in this phase. Two experimental set-ups were conducted (1) YMBR
and (2) BMBR. Treatment performance of both reactors was investigated at
different HRTs and SRTs (5 and 10 days).
5
The process efficiency was evaluated in terms of organic removal and membrane
filtration flux for various volumetric loading rates vis-à-vis HRTs.
(4) Sludge characterization study: Variation of sludge characteristics with different salt
contents (0.5, 15, 32 and 45 g/L) was investigated by using two-liter batch reactors with
the fill-and-draw operation. Sludge properties were evaluated in terms of extracellular
polymer content, CST, SVI, viscosity and nutrient contents.
6
Chapter 2
2
Literature Review
2.1 Introduction
High salinity wastewaters are usually generated from industries such as seafood
processing, vegetable canning, pickling, tanning and chemical manufacturing. Seafood
processing factories located in arid zones use treated seawater or reused or recycled water in
processing steps such as defrosting, butchering and washing raw materials. Thus, the effluent
from these industries contains high salinity, which is approximately the same as that of
seawater. In addition, the adoption of waste minimization techniques within these industries
has led to reductions in waste volume, while waste concentration has been increased
(Woolard and Irvine, 1995). These wastewaters are often difficult to treat with conventional
treatment processes such as activated sludge, trickling filter and anaerobic processes. High
salinity can cause osmotic stress or inhibit the reaction pathways in the organic degradation
process. This results in a significant decrease in biological treatment efficiency or
biodegradation kinetics. In addition, high salt content induces cell lysis which causes
increased effluent solids. The populations of protozoa and filamentous organisms required for
proper flocculation are also significantly reduced at elevated salt content (Burnett, 1974).
Therefore, conventional treatment process can hardly meet the effluent standards for high
salinity wastewater.
2.1.1 The Seafood Processing Industry
The seafood processing sector contributes serious organic pollution loads and high
salinity to receiving waters. This feature leads to difficulty in biological treatment processes
(Mendez et al., 1992). The fish processing industry with cooking and brine filling operations
normally produces high strength organic matter, high level of oil and grease, and high salt
content. Typical water consumption ranges from 18–74 m3/ton of fish processed (Battistoni
and Fava, 1994).
Process for canning shrimp is shown in Fig 2.1. In this process, receiving, peeling and
washing discharge large quantities of wastewater containing 90% of total COD. High salinity
wastewater is generated from precooking or brine treatment. In the precooking operation,
shrimp is boiled in brine solution for 3-5 minutes, or it is steamed. These operations curl the
meat, extract moisture and develop the pink or red color of the finished product. The salt
content of precooking wastewater is in range of 2 - 3% (UNEP, 1999).
Sardine and herring are classified as small, oily fish. These fishes contain a considerable
amount of oil or fat located between the skin and the flesh. The hot brine separates the oil
from the sardines. The oil rising to the surface of the brine forms a thick oil film on the top
which is then skimmed off. After cooking, the cans are taken out the brine and the remaining
brine in the cans is drained off. After cooling in a drying chamber, the cans are sealed, washed
down and packed.
The sardine and herring processing industry regenerates two kinds of wastewater: (1) sardine
and herring brine and (2) wastewater from cleaning and rinsing operations. The flow rate of
wastewater from these processes is as large as 6 times that of sardine or herring brine.
However, herring brine is a very concentrated wastewater consisting of a mixture of acetic
7
acid, sugar, fish protein and fish oil, as well as a number of spices and salt. The characteristics
of the herring brine is shown in Table 2.1.
water, debris
ice, water
water, debris
water
water
water, salt
water, salt
hot water
water, salt
meat particles
debris, shells
RECIEVING
THAWING
COOKING
SHELLING
WASHING
INSPECTION
shells
to wastewater treatment plant
SALTING
CANNING
RETORTING
COOLING
Product
water
water
water, salt
water
water
water, salt
water, salt
steam
water
Figure 2.1 Flow diagram of steamed canned shrimp processing
Table 2.1 Characteristics of herring brine waste (Balslev-Olesen et al., 1990)
Parameter Units Average value
COD mg/L 90,000
BOD5 mg/L 78,000
Oil/fat mg/L 4,000
Ntotal mg/L 3,000
Ptotal mg/L 1,000
SS mg/L 10,000
VSS mg/L 7,000
Chloride g/L 65
TDS g/L 110
pH 3.8
The canning process for mollusks such as mussel, oysters, clams or scallops also
generates large quantities of wastewater with salt content above 2%. The mollusks are shelled
and washed with 3 to 6% salt solution. Then they are drained and steamed or cooked for 10 to
15 minutes at 100 C. After inspection and grading, the cooked mollusks are packed in cans
8
with 1 to 2% brine. Mendez et al. (1992) reported that wastewater from processing mollusks
contained very high organic, nitrogen and salt content (18.5 g COD/L, 4.0 g N/L and above
2% salt).
The typical processing of dried salted fish is schematized in Figure 2.2. Slime, blood
and other contaminating substances of raw fish are washed off using a 3% solution of clean
salt in water. This reduces bacterial loads on the fish during subsequent salting. Large fish like
mackerel are split open at the ventral side from the head down. All visceral matter and blood
are removed. The fish is then cut into large pieces (1.5-2 cm thick, 10 cm wide and 20 cm
long). Fishes have an odor of ammonia, the dressed fish, or fish fillets, are soaked in mild
brine (10%) and crushed iced for six to 10 hours. This may be followed by salting. After
washing in clean brine solution, the eviscerated fish is salted in 21% brine for about 15 hours.
Salted fish is placed on bamboo trays and sun dried for two to three days in full sunshine,
depending on the size of the fish. Salted fish can also be dried in ovens. Fishes are then
packed and stored.
The characteristics of wastewater from the dried salted fish plant are shown in Table
2.2. This wastewater contains very high salt contents, ranging from 17g to 46g NaCl/L. A
large volume of wastewater is produced from soaking and washing operations. The volume
ranges from 10 to 12 m3/ton of preprocessed fish, and 20-30 m
3/ton of iced or fresh fish. The
preprocessed fish, namely, is eviscerated or beheaded and cleaning at fishing boats or villages
in coastal zone before it is transported to sea-food processing plant.
Table 2.2 Characteristics of wastewater from the dried salted fish plant (Dan, 2000)
Concentration Parameter Unit
Washing + soaking tank Combine Wastewater(*)
COD mg/L 5,250 873
SS mg/L 371 119
TDS g NaCl/L 46 17
Cl- g/L 27 10
SO42-
mg/L 1,240 164
TKN mg N/L 747 128
Total P mg P/L 5 5 (*)
Except brine waste from soaking tank
9
wastewater
wastewater, blood
wastewater
wastewater, salt
wastewater, salt
meat particles
eviscera
RECEIVING
THAWING
EVISCERATION
FILLETING
WASHING
SOAKING
SALTING
meat particles
WWTP(*)
for combined wastewater
GRADING
COOLING
Product
water
water, ice
Energy
water
water
water, ice
PACKING
DRYING
10% salt
21% salt
organics
Offal recovery (animal feed)
wastewater, salt
organics
WWTP for high salinity
wastewater
eviscera, skin
meat particles
energy
water, ice
Energy
Energy
Energy
ice, debris
(*)WWTP – Waste water treatment plant
Figure 2.2 Flow diagram of Dried and Salted fish processing
2.1.2 Pickled Vegetable Processing
Salt is used widely in vegetable canning processing to enhance flavor, to preserve, or for
conditioning. Therefore this industry in general produce wastewater containing high salt
content. The composition of the brines commonly used in canning vegetables is presented in
Table 2.3.
Table 2.3 Composition of brines used for canning vegetables (Joslyn and Timmons, 1967)
Product Brine, g/L
Asparagus 21.5 – 24.0 salt
Green bean 19.2 – 27.5 salt
Cabbage 15.6 – 25.2 salt
Beets 24.0 salt, 18 – 24 sugar may be added
Peas 21.5 salt and 36 – 48 sugar
Brine waste from fermenting pickles contains high salt content (3 to 20%) and
extremely high organic concentrations. This is due to extraction from pickled vegetable tissue
during fermentation process. In addition, large amounts of saline wastewater are also
10
generated from washing or rinsing pickles after fermentation and rinsing equipment, soaking
or fermenting tanks. Table 2.4 presents waste loads in some pickling industries. EPA (1975)
reported that the BOD:N:P ratio of pickles and sauerkraut wastewaters were 100:1:0.2 and
100:4:0.5, respectively. These ratio show that nutrient concentrations in the pickling
wastewater are low for microorganism growth (BOD:N:P for bacteria growth is 100:5:1).
Table 2.4 Raw waste loads and quality of wastewater from some pickling industries
Raw waste loads Concentration Category Flow
m3/ton(*)BOD5
kg/ton
SS
kg/ton
BOD5
mg/L
SS
mg/L pH
Pickles: (EPA, 1975)
- Fresh packed 7.76 8.61 1.72
- Process packed 8.70 16.7 2.97
- Salting stations 0.96 7.21 0.38
Total 17.4 32.5 5.07 1,500-
5,800
(3,280)
135-825
(400)
4.3 – 6.3
(5.3)
Sauerkraut: (EPA, 1975)
- Canning 3.19 3.18 0.55
- Trimming 0.39 1.13 0.17
Total 3.58 4.31 0.72 1,400 –
6,300
60 - 630 4.3 – 6.3
Sauerkraut:
- (Woodroof, 1975) 19 6.8 1.36
- (NCA, 1971) 15 6.4 0.45 (*) Ton of raw material
Kim chi pickle which is pickled celery cabbage is a well-known food in Korea, Japan
and Vietnam. Figure 2.3 illustrates kim chi processing. Park and Choi (1999) reported that the
volume of waste brine produced from a kim chi factory is approximately 0.53 - 0.67 m3/ton of
product. Typically, the composition of waste brine contains sugars and other nutrients
extracted from the vegetables during fermentation, as well as a high salt content
(approximately 10%). Table 2.5 describes characteristics of the waste brine from four
different Kim chi factories located in Suwon city and Kyunggi province in Korea.
Table 2.5 Characteristics of the waste brine from four different kim chi factories located in
Suwon city and Kyunggi province, Korea (Park and Choi, 1999)
Parameter Factory A Factory B Factory C Factory D
pH 5.36 4.91 5.48 5.80
NaCl, g/L 116 95 84 70
BOD5 , mg/L 1,100 1,200 1,060 1,040
COD, mg/L 1,300 1,790 1,550 1,250
TKN, mg/L 25 28 20 25
11
Fresh celery
CUTTING
FERMENTATION
water, debrisRINSING
BOILING
CANNING
WASHING
water, salt
water
Kim Chee Juice to WWTP
Animal feedsrecovery
water water, salt
Cabbage
discarded leaves
FERMENTATION
RINSING
CAN FILLING
WASHING
water
water
Outer leavesand bases
Squares
Liquidand bases
SEALING SEALING
wateronion, chili
brinebrine
ginger citric acid
water, acid
water, salt
debris
water, salt
water, spices
spoiled leaves/bases
Kim Chee Nappato WWTP
Figure 2.3 Flow diagram of kim chi pickles processing
The characteristics of high salinity wastewater generated from seafood processing and
vegetable pickling industries is shown in Table 2.6. In seafood processing, the main
environmental issue concerns the use of large amounts of fresh water for processing, and its
emission as wastewater. The volume of wastewater discharged depending on the type of
products or raw materials range from 10 to 120 m3/ton of product. In comparison with the
vegetable pickling process, a ten-fold increase in organic loading (COD or BOD5) is
discharged by the seafood processing industry. However, the salt content of wastewater from
vegetable pickling is normally very high, possibly as much as 200 g/L.
12
Tab
le 2
.6
Was
tew
ater
ch
arac
teri
stic
s of
fro
m v
ario
us
fish
ery
pro
duct
an
d v
eget
able
pic
kli
ng i
nd
ust
ries
Type o
f pro
duct
U
nit
Canned
sard
ine
Canned
shrim
p(2
)Canned
muss
el/oyst
er
Tuna
Fis
h m
eal
Kim
chi
pic
kle
s(3
) Cucu
mber
pic
kle
s(1
)Sauerk
raut(1
)
Wate
r &
wast
ew
ate
r volu
me
m3/t
on
9
60
20-1
20
22
97
0.6
17
4
BO
D5 load
kg/t
on
9
120
60
15
194
0.7
33
4
SS load
kg/t
on
5-6
54
- 11
- -
5
0.7
Oil
& G
rease
kg/t
on
27
42
- 6
- -
- -
Opera
tion u
nits
genera
ting s
alin
e
wast
ew
ate
r
O
ff-load,
sauce
filli
ng/c
an
wash
ing
Brine f
illin
g,
cookin
g,
sealin
g, ca
n
wash
ing
cookin
g,
wash
ing
cookin
g,
sauce
filli
ng/s
ealin
g/
can
wash
ing
Off
-load,
centr
ifugin
g,
stora
ge
Ferm
enting
Ferm
enting,
pic
kle
wash
ing
Ferm
enting,
pic
kle
wash
ing
Salin
e w
ast
ew
ate
r volu
me
% o
f to
tal
volu
me
39
(seaw
ate
r)
2 –
2.5
(2)
3 (3
) 3 (3
) 95
(seaw
ate
r)
100
56
89
TD
S o
f sa
line
wast
ew
ate
r
g N
aCl/L
30-3
5
20 –
30 (2
) 21
(3)
23 (3
)
100
30 –
200
30 –
200
Sourc
es
U
NEP, 1999
UN
EP, 1999
UN
EP, 1999
UN
EP, 1999
UN
EP, 1999
Choi &
Park
,1999
Mid
dle
bro
ok
1979
Mid
dle
bro
ok
1979
(1)
ton
of
raw
mat
eria
l
(2)
So
der
qu
ist,
19
71
(3)
Est
imat
ed f
or
was
te b
rin
e o
nly
13
2.1.3 Other Saline Wastewaters
Wastewater generated from oil field exploitation contains high salt content and is refered to as
oil-field brine. Its characteristics are presented in Table 2.7.
Table 2.7 Characteristics of oil field brine (Dalmacija et al., 1996)
Value Parameter
Max Min Average
COD, mg/L 1,200 200 400
pH 7.6 7.3 7.5
TS, g/L 34.6 29.3 32.3
Phenol, mg/L 0.14 0.01 0.05
Oil, mg/L 315 139 237
Cl-, g/L 17.9 17.4 17.6
SO4-, mg/L 17.7 9.2 11.9
In coastal areas, when subsurface water rises, infiltration of saline water into sewers can
result in high concentrations of chloride and sulfate in wastewater. Therefore, large variation
of salinity in domestic wastewater occurs normally in this areas. This can cause salt shocks or
adverse effects on conventional biological treatment methods.
Hypersaline wastes are produced in significant quantities in chemical industries such as
oil and gas production. These wastes contain organic compounds and high concentrations of
salt (>3.5%). High salinity is also found in landfill leachates. Pirbazari (1996) reported that
the leachates from domestic waste landfill (Los Angeles) and hazardous waste landfill for
chemical and petroleum waste (Niagara) had high strength organic matters and high total
dissolved solids (TDS). The characteristics of two leachates are described in Table 2.8.
Table 2.8 Characteristics of leachates (Pirbazari, 1996)
Leachate Parameters Unit
Domestic waste landfill Hazardous waste landfill
COD mg/L 3,050 – 3,450 9,000 – 10,500
BOD5 mg/L 1,505 – 1,710 6,950 – 7,500
TOC mg/L 905 – 965 3,040 – 3,500
SS mg/L 460 – 565 862 – 946
TDS mg/L 5,800 – 6,250 22,600 – 25,900
TKN mg/L 75 - 84 160 – 180
Oil and grease mg/L 60 – 80
pH - 4.3 – 6.0
2.2 Effects of High Salinity on Biological Waste Treatment Process
In wastewater treatment, there are conflicting reports on the influence of salt on the
biological processes. Some reports have indicated adverse effects of high salinity, or shocks
of NaCl on organic removal efficiency and sludge settleability (Burnett, 1974). Others have
reported that constant application of NaCl to biological treatment systems does not upset the
organic removal efficiency, and results in good flocculation of the biomass. This shows that
acclimation of the biomass and level of salt are important factors that may explain these
different observations (Hamoda and Al-Attar, 1995).
14
2.2.1 Aerobic Treatment
Previous studies reported that operation of activated sludge process at salt contents
higher than 20 g/L is characterized by poor flocculation, high effluent solids, and a severe
decrease in substrate utilization rate (Burnett, 1974).
Microscopic observations of the mixed liquor flocs (Burnett, 1974) showed that
alterations in saline wastewater caused alterations in the mixed liquor floc ecology. There was
a rapid die-off for rotifers, stalked protozoa and motile ciliate protozoa coincided with a
decrease in BOD removal and disruption in clarifier performance. After a few days, motile
ciliated protozoa were again observed, but rotifers and stalked ciliata were absent.
Tokuz and Eckenfelder (1979) estimated the effects of inorganic salts (NaCl and
Na2SO4) on continuous flow activated sludge with low F/M. The results indicated that the
relative high concentration of NaCl (up to 35 g/L) had only a slight effect on the performance
of the activated sludge process and the effluent SS did not increase. This was probably due to
a decrease in the F/M ratio. A further increase of NaCl over 35 g/L caused sudden increases in
effluent SS. The effect of sodium sulfate on the system was not significant. They also
observed that the protozoa population decreased gradually and disappeared at salt contents
above 35 g/L. The disappearance of the protozoa coincided with the sudden increase in
effluent turbidity or SS. Likewise the effects of high sodium chloride concentrations in an
activated sludge process was studied by Hamoda and Al-Attar (1995). The results showed that
the organic removal efficiency, and the treated effluent quality of activated sludge process did
not deteriorate as constant application of NaCl up to 30 g/L. COD removal efficiency ranged
from 93 to 99%. However, in order to obtain equivalent substrate removal, three-fold-lower
F/M ratios were applied in conventional AS at salt content of 30 g/L compared to those
applied in AS at salt-free wastewater (at the same SRT). Thus the substrate removal rate
decreases at high salinity. The MLVSS in the activated sludge reactor increased at salt content
up to 30 g/L. This result differs from the research conducted by Burnett (1974). An
explanation for this may be the long acclimation of microorganisms to the saline wastewater
might result in the growth of halophilic microorganisms in the system. Burnett (1974) also
noted that, although the substrate utilization rate decreased, the biomass yield obtained was
increased at higher NaCl concentrations. This may be due to a change in the efficiency of
microbial metabolism and the selection of salt tolerant species in the system. The salt tolerant
species may be halophilic micro-organisms such as Zooglea ramugera or Halobacteriaceae
which are aerobic heterotrophs.
Dalmacija et al. (1996) reported that the nature of pollutants and the high salinity (about
29 g/L) of oil-field brine has an unfavorable effect on the activated sludge process. High
hydraulic loadings (above 2.5 m3/m
3.day) increased the wash-out of the activated sludge from
the reactor. The addition of PAC improved the sludge volume index and increased the rate of
biodegradation. This is due to the ability of biofilm formation on the activated carbon surface.
Kargi and Uygur (1996) investigated the effects of high salinity on the Rotating Biodisc
Contactor (RBC). The results indicated that the rate and efficiency of COD removal decreased
significantly with increases in salt content above 10g/L. COD removal efficiency with salt
free wastewater was 95%. Due to the adverse effect of salt on microorganisms, the COD
removal was down to 60% at 5% salt content. The increase in salt content causes a linear
reduction in COD removal rate as shown in Fig. 2.4.
15
0.0 1.0 2.0 3.0 4.0 5.0
Salt concentration, %
2500
3000
3500
4000
4500
5000
CO
D r
em
ova
l ra
te, m
g/m
2. h
Figure 2.4 Variation of COD removal rate with salt contents (Kargi and Uygur, 1996)
A review of the literature (Table 2.9) confirms the presence of adverse effects of the high
salinity on the conventional activated sludge systems. Major problems encountered in the
biological treatment of saline wastewater were summarized by Kargi and Dincer (2000). They
are:
Limited extent of adaptation: Conventional cultures cannot be effectively used to treat
saline wastewaters with salt contents above 3%.
Sensitivity to changes in ionic strength: Shifts in salt content from 0.5 to 2% usually
cause disruptions in system performance. Rapid change in salt contents causes more
adverse effects than gradual change. Equalization to constant salt content is necessary
before biological treatment.
Reduced degradation kinetics: Biological degradation rates decrease with increasing salt
content. Therefore, saline wastewaters should be treated at lower F/M ratios.
High effluent SS: Salt content in wastewater reduces the population of protozoa,
resulting in low settlability. Salt content in wastewater increases the buoyancy forces,
causing low sedimentation efficiencies.
16
Tab
le 2
.9
Adver
se e
ffec
ts o
f hig
h s
alin
ity
in
act
ivat
ed s
lud
ge
pro
cess
Auth
ors
Experim
ent
Resu
lts
Ludza
ck a
nd N
ora
n (
1965)
Incr
easi
ng influent
from
100 m
g C
l/L
20,0
00 m
g C
l/L (
33 g
NaCL/L
) over
2 t
o 3
weeks
-Solid
loss
es
dis
rupting c
larifier
-10%
loss
in B
OD
5 r
em
oval
-In
hib
itin
g n
itrifica
tion
Burn
ett
(1974)
Changin
g T
DS u
p t
o 3
5,5
g N
aCL/L
-
Decr
easi
ng B
OD
5 r
em
oval fr
om
97%
to 2
5%
for
6 d
ays
aft
er
-Rapid
die
-off
of ro
tife
rs &
sta
lked/m
obile
cili
ata
pro
tozo
a
-Turb
id in e
fflu
ent
Tokuz
& E
ckenfe
lder
(1979)
Opera
ting c
ontinuous
flow
act
ivate
d s
ludge w
ith low
F/M
ra
tio a
t 3
5 g
NaCl/L
-Slig
ht
eff
ect
on B
OD
rem
oval
-Eff
luent
SS d
id n
ot
incr
ease
d d
ue t
o low
F/M
If
salt c
onte
nt
> 3
5g/L
-
Decr
easi
ng t
he p
opula
tion o
f pro
tozo
a a
nd t
hen d
isappeare
d
-In
creasi
ng e
fflu
ent
SS
Ham
oda &
Al-Att
ar
(1995)
Incr
easi
ng s
alt c
onte
nt
to 1
0 g
/L a
nd 3
0 g
NaCL/L
-
Decr
ease
in s
ubst
rate
utiliz
ation r
ate
-
But
incr
easi
ng b
iom
ass
yie
ld d
ue t
o s
ele
ctin
g s
alt t
ole
rant
speci
es
(halo
phili
c bact
eria s
uch
as
Zoogle
a r
am
ugera
, H
alo
bact
eriace
ace
etc
.)
Dalm
aci
ja e
t al. (
1996)
Oil-
field
brine w
ith s
alt c
onte
nt
of
29 g
/L
-In
creasi
ng w
ash
-out
of
act
ivate
d s
ludge a
s hydra
ulic
loadin
gs
> 2
.5m
3/m
3.d
ay
Karg
i &
Uygur
(1996)
Incr
easi
ng in influent
salt c
onte
nts
over
1%
for
RBC
-D
ecr
easi
ng C
OD
rem
oval ra
te &
eff
icie
ncy
. -
CO
D r
em
oval w
as
dow
n t
o 6
0%
at
5%
salt c
onte
nt.
-
Incr
easi
ng s
alt c
onte
nt
cause
d lin
ear
reduct
ion in C
OD
rem
oval eff
icie
ncy
.
17
2.2.2 Anaerobic Treatment
Anaerobic process has become one of the most interesting treatment for highly organic
polluted wastewaters. However, the presence of salts may cause inhibition and toxicity
problems in the methanogenic activity. High salt levels can dehydrate anaerobic bacterial cells
because of osmotic pressure.
Anaerobic digester was much more sensitive to chlorides than activated sludge (Burnett,
1974). Baere et al. (1984) examined influence of high NaCl on methanogenic activity on
anaerobic filter (AF) process. The AF reactor was filled with ether-based polyurethane foam
with a specific surface area of 600 m2/m
3. The results showed that initial inhibition occurred
at 30g NaCL/L. A shock treatment with 35g/L had a sharp decrease in gas production, which
dropped by 65%, and TOC removal efficiency, which decreased from 98% to 70%. The pH
dropped significantly after each shock treatment, from about 6.8 in the influent to a pH of 5.4.
When Methanosarcina, a halophilic anaerobic bacteria strain, was predominant (>99% of the
methanogenic biomass) in the reactor, TOC removal was improved. The methanogenic
activity of these bacteria was inhibited at 60 g NaCl/L. TOC removal was less than 20% and
the gas production dropped below 15% at 50 g/L.
An anaerobic and aerobic system consisting of an aerobic contactor followed by
activated sludge was tested for the biological treatment of high salinity wastewater (Belkin et
al. 1993). This wastewater generated from several chemical factories. The mean salt and COD
concentration were 32 g/L and 4900 mg COD/L respectively. Low COD removal efficiency
(55%) of the whole system was obtained at F/M ratio of 0.56 g COD/gMLSS.d for anaerobic
process and 0.28 for aerobic process. The COD efficiency increased to 74% at very low F/M
ratios (0.02 and 0.04 for the anaerobic and aerobic process, respectively).
Feijoo et al. (1995) examined the continuous exposure of high salinity in pilot scale
UASB and AF reactors. The results shows that the methanogenic activity of both anaerobic
processes was reduced by 50% at sodium concentrations above 20g/L. For unadapted sludge,
the anaerobic reactors could be shocked in the concentrations ranging from 6 to 13 g/L. In
addition, sodium inhibition in anaerobic digestion process was conducted with batch assays
(Feijoo et al., 1995). At low concentrations, sodium is essential for methanogens. The
optimum concentrations were reported to be about 0.23 – 0.35 g Na/L. The effect of NaCl on
the methanogenic activity depends on the type of sludge. When the sodium concentration was
increased by 4 to 10g/L (10 to 25 g NaCl/L), methanogenic activity reduced to 50%. After 40
days of digestion, the relative methanogenic activity of the sludge increased from 0% to 45%.
The sludge pregrown in the presence of high salt content showed a higher tolerance to
sodium, probably due to the adaptation of methanogenic bacteria to sodium. However, the
treatment efficiency after recovery was still low (45%).
18
Table 2.10 Adverse effects of high salinity in anaerobic treatment processes
Authors Experiment Results
Baere et al. (1984) AF with surface area of 600m2/m3
at 30 g NaCl/L - Decrease in gas production (dropped 65%) - TOC removal was decreased from 98% to 70% - Decrease in pH from 6.8 to 5.4
at salt content of 60 g/L - Gas production dropped below 15% - TOC removal < 20%
Feijoo et al. (1995) UASB and AF - Reducing 50% methanogenic activity at salt content > 33 g NaCl /L
- Shocked at concentrations ranging from 10-21 g NaCl/L for unadapted sludge
Belkin et al. (1993) Anaerobic and aerobic system at 32 g salt/L
- Low COD removal for whole system (50%) - COD removal (70%) could be improved at very
low F/M ratio (0.02 for anaerobic and 0.04 for aerobic process
Feijoo et al. (1995) Anaerobic batch digestion - Decreasing 50% of methane activity as increasing TDS by 10-25 g NaCl/L.
2.2.3 Nutrient Removal
In the nitrogen removal processes (Fig. 2.5), the oxidation of ammonia to nitrite and
then nitrite to nitrate (nitrification process) takes place under aerobic conditions (autotrophic
bacteria) and reduction of nitrate to nitrogen gas (denitrification process) occurs under anoxic
condition (hetetrophic bacteria). Dahl et al. (1997) found inhibition of the nitrifiers in the case
of a rapid increase of chloride concentration. The decrease in nitrification activity resulting
from increasing salt content from 16 g NaCl/L to 32 g/L, was approximately 30%.
Figure 2.5 Diagram of a nitrogen treatment system
Panswad and Anan (1999) investigated the effects of various salinity levels on ammonia
and nitrate uptake rates of the biological nutrient removal systems
(Anaerobic/anoxic/aerobic). In the steady state, the specific ammonia and nitrate uptake rates
decreased with increase in chloride concentrations. The total nitrogen removal dropped from
85% to 70% at high salt contents (20 and 30 g NaCl/L). Concurrently, COD removal of the
system also was dropped from 90% at 5 g NaCl/L to 71% at 30 g/L. This indicated that the
nitrifying and denitrifying bacteria are very sensitive to sudden high salt content even with a
high degree of pre-acclimation. Similarly in conventional activated sludge process,
acclimation was clearly proven to be an important factor in improving the nitrification and
denitrification performance of the system. The phosphorous removal of this system decreased
from 38 to 10% with gradually increase in salt content from 0 to 30 g NaCl/L. This indicates
that poly-P bacteria have intense sensitivity to high salinity condition.
19
Dincer and Kargi (1999) reported that the salt content reduced the rate and the
efficiency of nitrification and denitrification at salt contents above 2% and 1% respectively.
Nitrobacter was more adversely affected by high salinity than Nitrosomonas resulting in
accumulation of nitrite in the effluent at salt contents above 2%. The denitrification rate
seemed to be more sensitive to salt inhibition than nitrification is. A summary of adverse
effects of high salinity in nutrient removal is shown in Table 2.11.
Table 2.11 Summary of adverse effects of high salinity in nutrient removal processes
Authors Experiment Results
Dahl et al. (1997) Synthetic wastewater, combined biological nitrification and denitrification lab-scale experiment
- Nitrification and denitrification rates were reduced with increase in salt content (32 g/L)
Dincer and Kargi (1999)
Activated sludge for nitrification and downflow packed column for denitrification
- Salt concencentrations > 3% resulted in significant reductions in performance of both nitrification & denitrification.
Panswad and Anan (1999)
Lab-scale anaerobic/anoxic/aerobic with synthetic wastewater
- Shocked at 70 g NaCl/L. - Specific nitrate and ammonia decreased as
increasing Chloride concentration - Nitrifying and denitrifying bacteria were very
sensitive to sudden salt content. - More intense sensitivity of P-bacteria to high
salinity
2.3 Application of Halophilic Bacteria for Saline Wastewater Treatment
Removal of salt content from wastewaters by reverse osmosis, electrodialysis before
biological treatment, is normally expensive. However, because of salt inhibition of bacteria
growth, application of conventional treatment processes does not obtain acceptable efficiency.
In recent years, several studies have shown that, in utilization of salt-tolerant microorganisms
in biological treatment such as halophilic bacteria, yeasts could be a reasonable approach for
treatment of high salinity wastewater.
Non-halophilic bacteria grow well in media which contain less than 1% salt content.
True halophilic bacteria require salt for survival. These bacteria can be divided into two
groups, namely moderate and extreme halophiles. The moderate halophiles are
microorganisms which grow best in medium containing 3-15% NaCl (0.5-2.5M). While
extreme halophiles exhibit optimum growth in media containing 15 – 30% NaCl (Woolard
and Irvine, 1995). To tolerate the osmotic forces present in saline environments, halophilic
microorganisms accumulate compatible solutes to equalize the ionic strength of the cytoplasm
with external environment. Moderate halophiles accumulate a mixture of inorganic cations
(K+, Na
+) and organic compounds (amino acids, glycerol) for osmosis regulation.
Kargi and Dincer (1996) examined the treatment ability of Zooglea ramigera, a
moderate halophilic bacteria strain, at different salt contents using a fed-batch reactor. This is
different from sequencing batch reactor (SBR). The fed batch operation involves slow
addition of highly concentrated or wastewater into an aeration reactor until the tank is full.
With slow feeding, concentrated/toxic wastewater gets diluted inside the reactor, resulting in
less inhibition and higher BOD removal rates. COD removal efficiency for salt-free
wastewater was about 85%. This was not effected at salt content of 0.5%. However, the
efficiency dropped quite significantly with increasing salt contents above 1%, and attained
nearly 60% at 5% salt content.
20
In order to estimate the removal efficiency of salt tolerant microorganisms, Kargi and
Uygur (1996) used different types of microbial flora, namely Zooglea ramigera and
Halobacter halobium. The experiments were conducted using an aerated percolation reactor
with 1% salt content (Fig. 2.6). The percolator column was filled with crushed ceramic
particles ( = 4 mm) on which microorganisms were immobilized on the medium surface
(fixed biofilm).
Ceramic particles
air diffusor
Feed tank
air pump
Effluent tank
Percolate reactor
Figure 2.6 Schematic diagram of percolation reactor (Kargi and Uygur, 1996)
The highest COD removal efficiency obtained (90%) corresponded to the mixed culture
of activated sludge and Halobacter halobium. Kargi and Dincer (2000) conducted a further
study with Halobacter halobium. This species was cultured along with activated sludge in the
fed-batch reactor. The organic removal rate was significantly improved (Fig. 2.7). COD
removal of 85% was obtained within 9 hours at high salt contents (3-5%).
0.0 1.0 2.0 3.0 4.0 5.0
Salt concentration, %
100
200
300
400
CO
D rem
oval ra
te, g/m
3 h
Halobacter halobium+activated sludge
Only activated sludge
Figure 2.7 Variation of COD removal rate (R) as function of salt content (Kargi and Dincer,
2000)
Woolard and Irvine (1995) investigated the treatment of phenolic wastewater with
extremely high salinity. A moderate halophilic bacteria was seeded in a sequencing batch
reactor. Over 99% phenol removal was achieved from 15% saline wastewater. Tellez et al.
(1995) evaluated biokinetic coefficients in biodegradation of oil field produced wastewater. It
is generated during recovery of natural gas and crude oil from onshore and offshore
21
operations. A commercial bacterium sp. (Petrobac-S) was used in this study. This is a
hydrocarbon degrader specially formulated for degrading crude or refined hydrocarbons in
moderated saline environments. The result indicated that when TDS was increased from 50 to
100 g/L, the maximum specific growth rate reduced from 0.137 to 0.047 h-1
. There was a
slight increase in the half-velocity-constant (Ks) at higher salt contents. Ks is substrate
concentration at one-half the maximum growth rate. The affinity level for subtrate can be
evaluated in terms of KS .
The phenol removal capacity of a new moderate halophilic bacterium, Halomonas sp.
was investigated by Hinteregger and Streichsbier (1997). This bacteria consumed phenol as a
source of carbon at NaCl concentrations between 10 and 140 g/L. Under optimum conditions,
the degradation of 0.1g phenol/L in the aerated reactor was completed after 13 hours at salt
contents in the range of 30 and 50 g NaCl/L.
Dincer and Kargi (1999) reported that biological treatment of pickling industry
wastewater usually resulted in low COD removal efficiencies because of plasmolysis of cells
caused by high salt content (3-5%). Utilization of halophilic microorganisms (e.g. Halobacter
halobium) along with the activated sludge culture usually resulted in a better treatment
performance. COD removal of 97% was obtained at HRT of 30 hrs and sludge age of 10 days.
An aerobic, submerged biofilter, coupled with a trickling filter was investigated to treat
emulsified diesel fuel wastewater with high salinity (2% salt) (Yang et al., 2000). Figure 2.8
illustrates the schematic diagram of this system. The biofilter was randomly packed with
plastic media particles. The salt-tolerant-bacteria were isolated from the sediments on an
estuary. This system could give high removal efficiency (TOC removal > 90%) at volumetric
loading of 1.5 kg TOC/m3.d. The biodegradation of captured VOCs in the trickling filter was
effective (68% removal). The adsorption of VOCs was accomplished by countercurrent flow
of the gas and liquid phases through media bed. Based on the biodegradability tests at high
salt contents of 3.4% and 4.0%, the authors postulated that the bacterial mixture could
undergo high salinity up to 4.0%.
Air blower
Feed tank
Biofilter Trickling fliter
DO meter
pH meter
Figure 2.8 Schematic diagram of the biofilter and trickling filter treatment system (Yang et
al., 2000)
22
Table 2.12 Effects of using halophilic bateria for high salinity wastewater treatment
Authors Experiment Results
Kargi and Dincer (1996)
Feed batch reactor with Zooglea ramigena
- Not affected at salt content of 0.5% ( 5 g/L)
- The efficiency dropped fast at increasing salt contents above 1%.
Kargi & Uygur (1996)
Percolator with Zooglea ramigena, Halobacter halobium
- Zooglea ramigera culture obtained COD efficiency of about 77% at 1% salt.
- Halobacter alone obtained lowest efficiency
- Mixed culture of activated sludge and Halobacter obtained the highest efficiency at 1% salt and COD removal of 70- 80% at 4 – 5% salt
Woolard and Irvine (1995)
SBR with moderate halophilic bacteria
- 99% phenol removal was obtained at 15% salt
Tellez et al. (1995)
Biokinetic experiments for oil field produced wastewater
- Maximum specific growth rate reduced from 0.14 to 0.05 h-1 when TDS was increased from 50 to 100 g/L
Hinteregger and Streichsbier (1997)
Using moderate halophilic bacterium, Halomonas sp. to treat phenolic wastewater
- 0.1 g/L phenol was completely degraded after 13 h at 30 g/L salt
Dincer and Kargi (1999)
Using Halobacter halobium to treat pickling wastewater
- 97% COD removal was obtained at HRT of 30 hrs and salt content of 3-5%
Yang (2000) Aerobic-submerged biofilter coupled with trickling filter cultured with salt-tolerant-bacteria
- TOC of above 90% was obtained at VLR of 1.5 kg TOC/m3.d at salt content of 3.4 %.
2.4 Yeasts
2.4.1 General
Yeasts are eucaryotic, heterotrophic, unicellular microorganisms with a variety of
shapes ranging from spherical to egg-shaped (common shape) and ellipsoidal, and from
cylindrical to considerably elongated and even filamentous (mycelium). Yeast have no
flagella or other organs of locomotion. In general, yeast cells are larger than bacteria, ranging
from 1 to 5 m in width and from 5 to 30 m or more in length. Yeasts have a complex
internal structure as shown in Figure 2.9. The vegetative budding yeast cell, in the log growth
phase, contains a very large vacuole and has rigid walls.
Yeasts multiply as single cells, which divide by budding or direct division (fission
which is similar to that by bacteria reproduce), or sporulation (sexual reproduction takes place
by means of ascospores) (Fig 2.10). In some unicellular varieties, large numbers of cells
attach themselves after budding, to form a pseudomycelium. In other cases, true mycelia are
formed by fission (Fig. 2.11). Plasma membrane
Centrosome Centrochromatin
Cell wall Cytoplasm
Mitochondrium
Nuclear membrane
Vacuole
Figure 2.9 Diagram of yeast cell (Salle, 1961)
23
Figure 2.10 Budding is a common reproductive process in yeasts
a. Pseudomycelium b. True mycelium
Figure 2.11 True mycelium (formed by fission) and pseudomycelium (formed by budding)
The dissimilation of organics may occur anaerobically (fermentation) or aerobically
(oxidation). The most typical yeast process applied in food or beverage industries is
anaerobic, also known as alcoholic fermentation. The end products of a fermentation can be
alcohols, acids, esters, glycerol and aldehydes. Prior to fermentation, polymeric substances
(carbohydrates, lipids, proteins) are hydrolyzed by enzymes (hydrolases). A typical reaction
of sugar fermentation by yeasts is shown in the following reaction:
yeasts, nutrients
Carbohydrate C2H5OH + CO2 + new yeast cells
Under aerobic process (assimilation), complete oxidation of organics yields carbon
dioxide and water. Abundant supply of oxygen enhances considerable yeast growth. When
yeast are supplied with both sugar and oxygen, the colonies grow up to 20 times faster
through cell division than without oxygen. However, incomplete oxidation may generate acids
and other intermediary products. There are differences in the compounds which can be
assimilated by various species of yeasts. Some can degrade pentoses, polysaccharides
(starch), sugars, alcohols, organic acids (lactic, acetic, citric) and other organic substrates.
yeasts, nutrients
Organics + O2 CO2 + H2O + new yeast cells+ end products
Fission
Budding
Fission
Sporulation
24
Yeasts may utilize the nitrogen required in their metabolism for the synthesis of protein
from organic (amino acids, urea, vitamins, peptone, aliphatic amines, etc.) and inorganic
sources (ammonia, nitrite and nitrate). Most species can utilize the ammonium ion. Other
nutrients required for yeast growth include phosphorous, sulfur (organic sulfur and sulphate),
minerals (potassium, magnesium, sodium and calcium). Trace amounts of boron, copper, zinc,
manganese, iron, iodine, molybdenum are required to obtain optimum yields in synthetic
culture media. Basic components of Candida utilis are showed in Table 2.13.
Table 2.13 Basic composition of Candida utilis yeast biomass (Defrance, 1993)
Element C O N H P S Mg Ash
Value (%) 43.7 32 10.2 6.7 2.4 0.6 0.2 7.4
Based on these components, the chemical composition of Candida utilis yeast strain is
formulated as follows:
C3.64H6.7N0.73P0.07S0.02.
The nutrient demands can also be found from this formula. The C:N:P ratio of Candida
utilis biomass is 100:20:5, corresponding to BOD5:N:P ratio of 100:7.5:2. Therefore, nutrient
demands of yeasts are higher than that of bacteria whose BOD5:N:P ratio is 100:5:1
(Defrance, 1996).
Yeasts can grow in temperatures ranging from 0 to 470C. The optimum temperature for
most yeasts is 20 to 30oC. It is noted here that osmophilic yeasts are cable of growing in high
osmotic pressure habitats such as high concentrations of salt or sugar which restrict the
availability of moisture. On the other hand, yeasts can grow in a wide pH range (from 2.2 to
8.0). In general, yeasts grow well on media with acid reactions (3.8-4.0), whereas optimum
pH values for bacteria growth range from 7.5 to 8.5.
Fungi or yeasts may be found wherever nonliving organic matter exists. Unpolluted
stream water generally has relatively large numbers of species. Therefore, because of the
relation between fungi and yeasts densities and organic loading, it is suggested that fungi and
yeasts may be useful indicators of pollution. A survey of yeast populations along the St
Lawrence River that received domestic wastewater from Quebec Province (Simard 1971;
Simard and Blackwood, 1971). The results indicated that the blooms of pink yeasts
(Rhodotorula spp.) and black yeasts (Candida, Crytococcus, Torulopsis and Pullularia spp.)
occurred after bacteria had utilized the easily degradable components of the raw sewage.
Their density can be used as indicator of pollution.
2.4.2 Applications of Yeasts for Wastewater Treatment
a. Domestic and Industrial Wastewater Treatment
The use of yeasts in biological treatment of domestic and industrial waste has been
studied since 1970. Thanh and Simard (1971) studied the biological treatment of domestic
wastewater with different yeast strains. All the tests were carried out with shaken 500mL-
flasks at 26-28oC for 3 days. The initial pH was adjusted to 5.0. The result indicated that the
yeast strains which gave high ammonia-nitrogen and COD removal efficiency were
Rhodotorula marina (85% NH3-N and 67% COD removals) and Candida krusei (91% NH3-N
and 72% COD removals). Especially, yeast strain Rhodotorula glutinis and Trichothecium
roseum could completely remove phosphorous compounds in domestic wastewater (Simard
25
and Thanh, 1973). However, COD reduction was not as high as had initially been expected.
The authors analyzed the cause to be the result of rapid uptake of phosphorous and nitrogen
compounds before the organics could be assimilated. Deocadiz (1977) studied yeast treatment
of mixture of domestic and paper mill white wastewater. Two yeast strains Candida utilis and
Rhodotorula glutinis were cultured in shaking flasks. Approximately 80% of COD, 50% of N
and 62% of P were removed after 24h. Rhodotorula yeast strain also gave the highest removal
efficiency for the biological treatment of potato chips wastewater. The COD, N and P
removals were 80%, 96% and 57%, respectively (Simard et al., 1973). The yeast sludge
contained high protein content (53%).
Henry and Thomson (1979) observed that Candida ingens yeast spontaneously grew
and formed a thick film on the supernatant of anaerobic piggery waste digesters. Based on this
observation, the authors investigated treatment ability of C. ingens for the effluent from these
digesters. The yeast was cultured in the stirring batch reactor. The results indicates that C.
ingens could utilize almost all the VFA up to a concentration of 0.09 mol/L after 24h growth
period. C.ingens grew well at pH ranged from 4.8 to 5.0 and at temperature of 29-32oC.
Miskiewicz et al. (1982) developed further yeast treatment of fresh piggery wastes by
adding carbon source (beet molasses or sucrose). Four yeast strains, Candida tropicalis,
Candida tropicalis, Candida robusta and Candida utilis were cultured in a batch aerated
reactor. The study shows that the use of raw piggery waste without carbon supplement leads
to low biomass yield and low treatment efficiency, even though the nutrients (N, P) are high.
It is found that molasses are the most appropriate carbon source. The culture of C. utilis on
molasses-enriched piggery waste (5570 mg COD/L) could obtain high treatment efficiencies.
76% TKN, 60% COD, 84% total P were removed at HRT of 7 hours and F/M ratio of 1.73 g
COD/g MLSS.d. The maximum specific growth rate of C. utilis was 0.19 h-1
.
Anciaux et al. (1989) investigated the influence of DO, substrate concentration, type of
VFA on the growth of C. ingens in the aerated and stirred batch reactor. The result shows that
higher the DO concentration, the shorter the lag phase and max increased with the DO
concentration according to the trend of the Monod model curve. Thus DO becomes a rate-
limiting factor at a very low concentration (Figure 2.12a). The effect of substrate
concentration is shown in Figure 2.12b. This figure shows inhibition of growth by the
substrate concentration. At VFA concentration of 0.25 g /L (approximately 270 mg COD/L),
the best growth rate ( ) and yields obtained were 7.5 d-1
and 0.6 g DS/ g acid consumed,
respectively.
0 20 40 60 8010 30 50 70
% DO (% saturation)
5.20
5.60
6.00
6.40
6.80
7.20
Sp
ecif
ic G
ro
wth
Ra
te
da
y-1
0 0.1 0.2 0.3 0.4 0.50.05 0.15 0.25 0.35 0.45
VFA concentration (g/L)
2.00
4.00
6.00
8.00
Sp
ecif
ic G
ro
wth
Ra
te
da
y-1
a. Specific growth rate vs. DO b. Specific growth rate vs. VFA concentration
Figure 2.12 Specific growth rate of Candida ingens vs DO and VFA concentration (Anciaux
et al., 1989)
26
Katayama-Hirayama et al. (1994) cultured the yeast strain of Rhodotorula glutinis with
phenolic wastewater. The cultures were propagated in shaking flasks and incubated at 30oC.
Phenol and monochlorophenols were completely degraded after 2 days. COD removal
obtained ranged from 79% to 83%. When compared with cell yields on the glucose (0.66 g/g)
and acetate (0.39 g/g) media, phenol is an excellent carbon source (y = 0.61 g/g). None of the
studies reviewed above evaluated the settling ability of yeasts.
Hu (1989) used ten different yeast strains in cultures to treat vermicelli wastewater
which contains high concentration of starch, lactic acid and protein with BOD ranging from
24,000 to 44,000 mg/L. Based on the ability of starch degradation, protein hydrolysis and
lactic acid tolerance, these yeast strains were screened from 391 colonies isolated from soil
samples. Most could grow well within pH range of 3.0-5.0, with pH 4.0 being the optimum.
The results shows that the two strains could reduce soluble COD by 92% at HRT of 7 days,
F/M ratio of 0.48 g COD/g MLSS.d and VLR of 1.03 kg COD/m3.d. The long HRT in this
process is due to the poor settling ability of yeasts. The yeasts could not be flocculated or
settled as in a conventional activated sludge process, and were easily washed out with the
effluent. Therefore, the HRT and SRT were kept constant. The author postulated that the
fungi contamination prevented the formation of yeast flocs.
Similarly, Chigusa et al. (1996) used nine different strains of yeasts capable of
decomposing the oil to treat wastewater from oil manufacturing plants. A pilot scale yeast
treatment system had been run for one year. The results showed that 10,000 mg/L of hexane
extracts in the raw wastewater were reduced by the yeast mixture to about 100 mg/L.
Also, Elmaleh et al.(1996) investigated the yeast treatment of highly concentrated acidic
wastewater from the food processing industry. The strain Candida utilis was cultured in
continuously completed mixing reactors. This system did not have a separate settling tank; the
SRT and HRT of the system are identical. The carbon source of feed wastewater was a
mixture of acetic acid, propionic and butyric acid. The pH was maintained at 3.5 to prevent
any bacterial contamination. The TOC removal obtained was 97% at high loading rates (30 kg
TOC/m3.day). The growth yield and maximum specific growth rate of yeasts were similar to
those for conventional activated sludge ( max = 0.5 h-; Y = 0.85-1.05 kg SS/kg TOC for acetic
acid). In this study, the authors only evaluated biokinetic constants, but did not focus on the
settling ability of yeasts.
Olive mill wastewater normally contains high concentration of fats, sugars, phenols,
volatile fatty acids which contribute to a very high COD concentration (100-200 g/L). Scioli
and Vollaro (1997) reported that Yarrowia lipolytica cultured in the 3.5L-aerated fermenter
was capable of reducing the COD level of olive oil processing wastewater by 80% in 24 hrs.
The effluent had a pleasant smell due to the presence of methanol and ethanol, while fats and
sugars were completely assimilated. The authors postulated that using membrane to filter
effluent before discharging into the sewage system might be a feasible approach for pollution
reduction in olive-oil-producing countries. Useful biomass (40% protein) and valuable lipase
enzyme could also be obtained in this process.
Silage is produced by the controlled fermentation of a crop with high moisture content,
such as grass or maize. This silage can be used as an animal feedstock. Its effluent is
extremely polluting, having very high BOD (30 – 80 g/L) and low pH (3.0-4.5). Arnold et al.
(2000) investigated the ability of selected yeast strains (C. utilis and Galactomyces
geotrichum) to purify silage effluent on the shaker-flask scale. High removal efficiencies of
COD (74-95%), VFA (85-99%) and phosphate (82-99%) were obtained after 24 hrs. Some
27
ammonia was also removed. pH rose during treatment to 8.5-9.0 from initial values of 3.7-5.8.
This was presumably due to removal of lactic acid and VFAs. The dramatic decrease in P
(resulting in extreme P removal) may be attributed to the shortage of phosphorus.
In general, carbon and nitrogen removal from high organic-strength wastewater can be
conducted with different processes, namely, anaerobic and aerobic processes, nitrification and
denitrification. Ortiz et al (1997) proposed an effective and economic alternative process in
which it is possible to achieve both carbon and nitrogen removal in two stages: anaerobic
bacterial treatment and yeast treatment process. The fermentative bacteria transforms the
organic nitrogen and the carbonaceous substrates into ammonia and volatile fatty acids (VFA)
which are degradable substrates for the yeast growth (Fig. 2.13). A comparison between yeast
and anaerobic treatment process is presented in Table 2.14.
Anaerobic
process Nitrification
process Dinitrification
process
a. Traditional coupling for carbon and nitrogen removals
Acidogenesis
process
Yeast
process
b. Coupling for Anaerobic acidogenesis and Yeast treatment
Figure 2.13 Traditional carbon and nitrogen removal system can be altered with anaerobic and
yeast treatment system (Ortiz et al. 1997)
Table 2.14 A comparison between yeast and anaerobic treatment process (Defrance, 1993)
Yeast process Anaerobic process
Cannot degrade complex organic compounds Degrades cellulose
High nutrient requirement (BOD5:N:P) Low nutrient requirement
Need for oxygen and agitation Slight agitation
Exothermal reaction need for cooling Producing methane as valuable bio-fuel
Sensitive to variation of temperature
Can consume VFAs produced from
acidogenesis
Dependent on two phases: liquefaction and
gasification
High organic loadings Low organic loadings (< 3 kg COD/m3.d)
Short HRT High HRT (minimum 10 days)
Valuable biomass Poor sludge production
Generally, dairy industry effluents contain large quantity of milk constituents such as
casein, lactose, fat and high inorganic salts. Marwaha et al.(1999) investigated the effect of
nitrogen supplements (urea and yeast extract) on the treatment ability of two yeast strains
Candida parapsilosis and Candida haemulonii isolated from the dairy effluents. All tests
were conducted with shaker-flasks and incubated at 30oC for 24 hrs. The pH of the medium
was adjusted to 5.5. The result indicated that maximum BOD (90%) and COD (82%)
28
removals could be obtained when 0.6% yeast extract was supplemented. The relation between
biomass growth and organic removal was not determined in this study.
b. High salinity wasterwater treatment
Choi and Park (1999) studied the treatment ability of an osmotolerant yeast, Pichia
guilliermondii A9, for waste brine from a kim chi factory using a shaker-flask scale. The
growth of Pichia guilliermondii A9 in waste brine was not inhibited by NaCl concentrations
up to 100 g/L. However, it was affected at concentrations above 120 g/L. Approximately 90%
of BOD was removed from the waste brine after 24 hrs. The maximum cell yield was 0.69 g
of dry cells per liter, containing 40% of protein. Cell growth was highest at pH 4, and
declined slightly when pH increased to 8.
Nishihara ESRC Ltd. (2001) studied the effect of the Yeast Cycle System on marine
products processing wastewater. In this system, yeasts are used for wastewater, where the
excess yeast from the treatment process is recovered and reused. The system consists of
pretreatment by yeast and secondary treatment by activated sludge process (Fig. 2.14). Table
2.15 and Figure 2.15 give a comparison with conventional complete mixing activated sludge
in terms of the operating conditions. Marine products processing wastewater has BOD5 and
SS concentrations ranging from 3,550-8,850 mg O2/L and 680-940 mg/L respectively. The
chloride concentration was 5,160 mg/L. Some yeast strains, Candida edax, Candida
valdivana and Candida emobii, were predominantly grown during enrichment with this raw
wastewater. The predominance of yeast strains with the enrichment culture technique is based
on free competition among different organisms in real wastewater. It was found that the yeast
treatment process can obtain high efficiency at a higher volumetric loading (5 – 6 times), F/M
ratio (2 – 3 times) when compared with the AS process.
Yeast reactor AS reactorSettling tank I Settling tank II
yeast sludge return AS sludge return
excess yeast sludge excess AS
Figure 2.14 Schematic diagram of the Yeast Cycle System (YCS)
Table 2.15 Operating conditions of YCS (Nishihara ESRC Ltd., 2001)
YCS AS(*) Parameter Unit
Range Mean Range
Influent BOD5 mg/L 3,550 –8,650 6,100 110-400
Salt content g/L NaCl 5-8 6 0.05-0.45
BOD5 volumetric loading kg/m3.day 4.5 – 10.4 7.5 0.8 – 1.9
Yeast concentration mg/L 8,000 – 10,000 9,000 2,500 – 4,000
BOD5 sludge loading (F/M) kgBOD5
/kgVSS.day
0.56 – 1.04 0.9 0.2 – 0.6
Water temperature C 23 – 30 26 23 – 30
pH 4.3 – 5.2 4.8 6.5 – 8.5
DO mg/L 0.51 – 0.95 0.7 2
SVI ml/g 60 – 72 66 100 – 120
Y kgVSS/kgBOD5 - 0.16 0.4 – 0.8
(*) Complete mixed activated sludge (Metcalf and Eddy, 1991)
29
L MLSS/L F/M Y SVI0
1
2
3
4
5
6
No. tim
es h
igher th
an v
alu
e
of activate
d s
ludge p
rocess
Activated sludge
Yeast Cycle System
Note:L = BOD5 volumetric loading (kg/m3.day) MLSS/L = Concentration of microorganism (mg/L) F/M = BOD5 sludge loading (kgBOD5/kgSS.day) Y = Sludge yield (kg VSS/kg BOD5)SVI = Sludge Volume Index (ml/g)
Figure 2.15 Comparison between Yeast Cycle System (YCS) and complete mixed activated
sludge (AS) (Nishihara ESRC Ltd., 2001)
Large flocs formed in the yeast treatment system were able to settle quickly. Thus the
MLSS could be maintained at high concentration of about 10,000 mg/L. The yeast sludge has
low SVI of 50-60 mL/gram, which was equal to half the SVI value for activated sludge. This
makes reducing the size of the sedimentation tank, and thickening of yeast sludge or chemical
conditioning for dewatering are not necessary. The efficiency of this system is presented in
Table 2.15 and Figure 2.23.
Table 2.16 Quality of treated water and efficiency of the YCS for seafood processing
wastewater treatment (Nishihara ESRC Ltd., 2001)
Parameter Influent After pretreatment
by yeast E%
After
activated sludge E, %
BOD5 , mg/L 5,450 150 97% 4 99%
SS, mg/L 798 113 86% 15 87%
T-N, mg/L 153 72 53% 10 86%
T-P, mg/L 33 18 46% 15 17 %
Cl-, mg/L 5,160 5,080 5.080
BOD5 and SS removal by yeast were more than 95% and 87% respectively. The
nitrogen (52%) and phosphorous removal (46%) during the yeast treatment were equal to
those found in the components of excess yeast sludge. The company postulated that the
structure of yeast flocs facilitated oxygen diffusion. Therefore energy could be saved through
the reduction of the supplied air flow. Moreover, the excess yeast with high protein, vitamins,
and lipid content could be used for animal feedstuff, mushroom growing or fertilizer.
30
Table 2.17 Summary of studies on yeast treatment of high salinity wastewater
Authors Experiment Results
Nishihara ESRC Ltd. (2001)
Yeast treatment system for marine products process wastewater
- BOD5 and SS removal were more than 95, 97% respectively
- High BOD5 volumetric loading (7.5 kg/m3.day)- High BOD5 sludge loading (0.9 kg BOD5/kgVSS.day) - Low excess yeast, 0.16 kg VSS/kg BOD5
Choi and Park(1999)
Yeast treatment for Kim Chi waste brine
- Pichia guilliermondii can tolerate NaCl up to 100 g/L - 90% BOD removal obtained for 24h
c. Waste recycle
Single-cell protein production (SCP)
Linkage between biomass for food production and waste and wastewater treatment has
been widely developed. A number of organisms are utilized for biomass and protein
production. These include: (1) protein-rich algae, fish, duckweed and water hyacinth in
oxidation/stabilization ponds; (2) bulrush, cattails and other plants in constructed wetlands;
(3) worms from composting waste and sludge and (4) yeasts and fungi cultured from
carbohydrate-rich wastewater. Therefore, utilization of food-processing wastewater as
substrates for biomass production or single-cell protein production (SCP). SCP results in
purification of effluent. This application can also obtain savings from decrease in disposal and
treatment costs. Single-cell protein production (SCP) is defined as microbial biomass
produced by some biological process and it can be used as food or food additives. Industries
that produce large volumes of carbohydrate-rich-containing wastewaters free of toxic
materials are most promising substrates for SCP. Such industries include milk, cheese-
processing, confectionery manufacturing, and food canning. Effluents from these industries
which contain high concentrations of COD and nutrients are costly to treat. Thus their
utilization for SCP is an attractive alternative.
Yeasts such as Saccharomyces cerevisiae, Candida utilis and most fungi, are quite
acceptable to animals and man. Whereas algal and bacterial biomass are less pleasant and
contain undesirable levels of certain cellular materials (such as high nucleic acid content,
toxic or carcinogenic substances absorbed from the growth substrate). In addition, due to
abundance of valuable nutritious substances such as proteins and vitamins, yeasts are the most
feasible and acceptable microorganisms in the production of SCP. In their simplest
processing, yeasts can be cultured in a suitable substrate, normally carbohydrates, such as
molasses, whey or starch, and under suitable conditions. The yeast biomass is harvested from
the fermentation or assimilation followed by separation (settling, filtration, centrifugation,
membrane), washed and dried to produce a free-flowing powder, rich in protein (Gray, 1989).
For example, 50 thousand tons of yeasts per year are produced in North America from
sulphide liquor (paper mill waste) containing high pentoses and hexoses.
A novel SCP process developed by George Bassett Co. in Sheffield is shown in Fig.
2.16. In this process, Candida utilis is cultured with confectionery wastewater and then
harvested by centrifuging and drying. The SCP is packed and sold as a high-protein additive
for animal feed. The present output of this plant is about 140 tons/year. It is able to remove
65% COD of the wastewater. Therefore, the wastewater after the SCP process contains
remaining COD concentration, which is low enough to be discharged directly to the sewer
(Gray, 1989).
31
Sterilizer Centrifuger Spray dryer
Crude effluent
Equalization tank
Inocula reactor Main reactor
Package
Figure 2.16 SCP from confectionery effluent (Gray, 1989)
The Symba process developed by Swedish Sugar Company is based on the symbiotic
culture of yeasts Endomycopsis fibuliger and Candida utilis with potato processing waste and
wastewater (Figure 2.17). The basic substrate for SCP production in wastewater is starch,
which is not easily assimilated by Candida utilis. However, the starch can be hydrolyzed to
low molecular weight sugars (glucose, maltose) by the enzyme amylase. This enzyme is
produced in large quantity by E. fibuliger. These hydrolyzed products (sugars) are then easily
degradable substrates for the growth of C. utilis which has high nutritional value. In the
process, the wastewater is strained in order to remove any large particles and then sterilized
by heating to destroy any microbial contamination. The sterilized substrate is introduced to a
preliminary reactor, in which E. fibuliger is grown to provide its population in the main
reactor, as its growth rate is much lower than that of C. utilis. These reactors are maintained in
aerobic condition. A large quantity of heat is generated from yeast assimilation activity and
removed by cooling towers or heat exchangers. Excess biomass is purified by sieving and
then concentrated by centrifuge and dried by spray drier. The Symba process can achieve
BOD removal of 90% (BOD of wastewater is reduced from 15,000 mg/L to 1,500 mg/L),
both N and P removals of about 50%. The yeast contains about 45% protein with less nucleic
acids but rich vitamins, especially vitamin B (Gray, 1989).
Storage tank
Sterilizer
Endomycopis
reactor
Cooling tower
Air blower
Storage tank Separater
Sewer
Separater
Sewer Packing
Symbiosis
reactor
Spray drier
Figure 2.17 The Symba process (Gray, 1989)
32
Simard and Cameron (1974) evaluated the growth of Candida utilis on dilutions of
spent sulphite liquor (SSL) with addition of different nitrogen sources. These were urea,
ammonia sulphate and ammonium hydroxide. The result shows that the dilution of SSL (2
water:1 SSL) increased significantly the dried biomass. Urea, ammonia sulphate gave high
conversion efficiency (70%).
Candida utilis yeast can be used to purify and uptake effectively ammonia present in the
anaerobic digester supernatant supplemented with molasses as a source of carbohydrate (Irgen
and Clark, 1976). The result shows that 100 g of molasses could yield 41 g of dried Candida
utilis yeast cells and 20 g of protein.
Likewise, Barker et al. (1982) cultured Hansenula anomala, Candida krusei and
Geotrichum candidum with whisky distillery spent waste. The influent COD of this waste
ranges from 15 g/L to 58 g/L. These yeast strains were isolated from whisky distillery
effluent. The results indicated that the yeasts could give high yield of protein biomass and
COD removal of 55% .
Rashad et al.( 1990) found that mango peel waste from drink processing industries can
be used for SCP production in which Pichia pinus is cultivated under optimum conditions (pH
= 4.8-5.0; temperature = 30oC). The maximum yield obtained after 3 days of growth was 6.2
g/ biomass/L of wastewater and dried yeast biomass contained high crude protein (62%) and
low nucleic acid (12%).
Liquid waste (deproteinized leaf juices) is generated from vegetable protein production.
Chanda and Chakrabarti (1996) reported that depended on type of vegetable, BOD5 and COD
of the waste can range from 12.9 to 19.0 g/L and 20.2 to 28.5 g/L respectively. This liquid
waste can be a good substrate for the cultivation of S. cerevisiae, T. utilis, C. lipolytica. The
yeast biomass obtained was rich in protein and vitamins. BOD of wastewater reduced
significantly (74 – 97%) by the growth of these yeasts. The shrimp shell waste could also be
converted into proteins by using the yeast Scharomyces cerevisiae KIV-1116 (Ferrer et al.,
1996).
Bio-fuel
Fuel-alcohol production by yeast/fungi fermentation of the industrial or agricultural
wastes has received considerable attention in Brazil and India in early 1990. Development of
such technologies could reduce the oil imports and a partial solution to disposal of wastes.
Nigam (1999) reported that the culture of Saccharomyces cerevisiae on pineapple cannery
waste had high ethanol productivity (0.98g ethanol /g yeast.h) and sugar uptake rate (2.3 g
sugar/g of yeast. h).
Yu et al. (1987) used the yeast Candida shehatae to ferment the spent sulphite pulp and
paper mill. The ethanol yield gained was 0.46 g ethanol/g initial sugar at HRT of 18 h. This
corresponds to the sugar removal of 95.5%. In general, the municipal primary wastewater
solids contains about 10% cellulose and 26% lignin. These cellulose components can be
effectively converted to ethanol by fungal cellulase and yeast Sacharomyces cerevisiae
(Cheung and Anderson, 1997).
Lark et al. (1997) studied on the reuse of recycled paper sludge. At least 72% of
cellulose in the sludge was converted into ethanol by the yeast Kluyveromyces marxianus.
The paper sludge volume was reduced to 30 –35% of the original volume after 72h of
33
fermentation. The ethanol production by yeasts from cassava grate waste was also studied
(Agu et al.,1998) where 60% of cellulose and lignin materials was hydrolyzed and converted
to ethanol.
The cheese waste normally contains very high content of lactose which can be suitable
substrate for ethanol production. Its concentration can be up to 50 g lastose/L. Ghaly and El-
Taweel (1997) reused this waste for continuous ethanol production by yeast Candida
pseudotropicalis. The results shows that high ethanol concentration (58 g/L) could be
achieved at HRT of 42 hours.
2.5 Theoretical Modeling Consideration
2.5.1 Growth without Inhibition
a. Specific growth rate ( )
Jackson and Edwards (1975) estimated specific growth rates of microorganisms in a
culture by the following expressions:
Xdt
dX(2-1)
)( ott
oeXX (2-2)
oo XttX ln)(ln (2-3)
o
o
tt
XX lnln(2-4)
Where X = Biomass concentration at the time t (mg/L)
Xo = Biomass concentration at the time to (mg/L)
= Specific growth rate (1/h).
Time
Bio
mass
conce
ntr
ation Ln(Xt)
Ln(Xo)
to t
Figure 2.18 Growth curve of microorganisms in a culture
Generally, Monod's model is used to estimate different biokinetic reactions between
microorganisms and the substrate in a continuous culture (Metcalf and Eddy, 1991).
According to this model, specific growth can be related to substrate by the following
relations:
S
mKS
S(2-5)
Where
= Specific growth rate of microorganism (d-1
)
34
m = Maximum specific growth rate (d-1
)
S = Substrate concentration (mg/L)
KS = Half-velocity constant or Monod constant (mg/L).
Substrate concentration, mg/L
Specific
gro
wth
rate
()
tim
e-1
max
max
2
Ks
Figure 2.19 The effects of a limiting substrate on the specific growth rate (Monod model)
b. Substrate Utilization rate (U)
Substrate utilization rate(U) vis-a-vis COD removal rate, can be expressed as:
X
SSU o
.
)((2-6)
Where
U = Substrate utilization rate, (mg substrate removed/ mg MLSS.day)
So = Initial substrate concentration (mg/L)
X = Biomass concentration (mg/L)
S = Final substrate concentration (mg/L)
= Hydraulic retention time (day).
c. Growth Yield Coefficient (Y)
The relation between new cell production and soluble substrate consumption can be
represented as follows:
dt
dSY
dt
dX* (2-7)
Where
Y = true growth yield coefficient (g SS/ g substrate removed.day)
X = Biomass concentration (mg/L)
S = Substrate concentration (mg/L)
For a given microorganism and essential nutrient/substrate under the same environmental
conditions, the weight of microbial cells produced per weight of nutrient/substrate consumed
is constant. This relationship is expressed as:
Y = Weight of organisms produced/Weight of substrate utilized
35
2.5.2 Growth with Inhibition
Han and Levenspiel (1988) proposed generalization of the Monod expression which
takes into account inhibition effects caused by high concentration of substrate, product or
toxics, ammonia, ion strength (salts) and other inhibitory substances.
IS
n
Im
K
IKS
S
K
I
1
1(2-8)
Where
I = concentration of inhibitor
KI = the critical inhibitor concentration above which reaction stops
n = constants
Some kinetic models for growth with inhibitory substances are shown in Table 2-18.
Table 2.18 Kinetic models for inhibition growth (Han and Levenspiel, 1988)
Equation Model name
S
obs
SI
mKS
S
KS
S
K
I1 (2-9) Ghose and Tyagi
Sobs
SIm
KS
S
KS
S
K
I5.0
1 (2-10) Bazua and Wilke
Sobs
S
n
Im
KS
S
KS
S
K
I1
(2-11)Han and Levenspiel
Where
m = Maximum specific growth rate at inhibitor concentration of zero (I = 0).
obs = Observed maximum specific growth rate at certain inhibitor concentration (I).
Inhibitor concentration (I), mass/vol
Obse
rved s
peci
fic
gro
wth
rate
(obs
), t
ime
-1
KI
max
n = 0.3
n = 0.5n = 1.0
Figure 2.20 Curves of inhibition growth models (n =1: Ghose and Tyagi; n= 0.5: Bazua and
Wilke model)
Eq.2-11 is a generalized form of Eq.2-9 and Equ.2-10, in which the constant n is 1.0 and
0.5 respectively.
.
36
a. Substrate Inhibition
Some authors have suggested models for the growth inhibition of Candida utilis on
acetic acid as substrate. These are:
cSm
S
S
SK
S1 (2-12) Defrance (1993)
iS
m
K
SSK
S
1)((2-13) Haldane (Ortiz et al. 1997)
iS
m
K
SSK
S2
(2-14) Andrew (Ortiz et al. 1997)
Where Sc = The critical substrate concentration above which reaction stops
Ki = Inhibition constant
Defrance’s model is also a modification of Han & Levenspiel’s model (Eq. 2-11) in which KI
=Sc and n = 1.
Substrate concentration (S) , mass/vol
Observ
ed s
pecific
gro
wth
rate
(obs)
,tim
e-1
max
Monod
Andrew
Haldane
Sc
Defrance
Figure 2.21 Curves of substrate inhibition growth models
b. Salt/ions Strength Inhibition
Webb’s model presents ion-strength inhibition for microorganism growth:
)17.1exp(
1max
IS K
KS
S
(2-15)
Where = Ion strength
Dincer and Kargi (1999) proposed expression to estimate salt inhibition for nitrification
and denitrification.
sTN
TNON
oN
CK
KR
NNR (2-16)
Where RN = Rate of nitrification and denitrification, kg/m3.h
KTN = Salt inhibition constant for nitrification and denitrification, g/L
37
Cs = Salt content, g/L
RON = Nitrification and denitrification rates for salt-free wastewater.
c. Other Inhibition Factors
The effect of pH, ammonia on the aerobic growth of Candida utilis at constant
temperature (30oC) are illustrated by the following models:
Effect of
ammonia
3
23
33
3
NHiNH
m
K
NHNHK
NH
(2-17) (Ortiz et al., 1997)
Effect of pH
1
21
1
K
H
H
Km
(2-18) (Jackson and Edards, 1975)
Where K1 , K2 = pH constants
Henze et al. (1997) described effects of pH, temperature and DO on aerobic
heterotrophic micro-organisms by the following kinetic models:
Effect of pH:
IK
KpHpH
pH
pHoptmm )()( (2-19)
Effect of temperature: )]20(exp[*)20()( CCtCCt ooom
om (2-20)
Effect of DO:
22
2
OO
O
Sm
SK
S
SK
S(2-21)
Where
I =110
)( pHoptpH
KpH = pH constant
= Constant
SO2 = DO in the mixed liquor
KS,O2 = Saturated constant for oxygen
2 4 6 8 10
pH
Obs
erv
ed s
pecific
gro
wth
rate
tim
e-1
pHopt
)
0 2 4 6 8
DO concentration, mg/L
Observ
ed s
peci
fic g
row
th rate
(D
O), tim
e-1
(DOopt)
Figure 2.22 pH and DO models
38
2.6 Respirometric Method
The respirometry measurement technique is used to measure the biochemical oxygen
uptake rate (OUR) under well-defined experimental conditions. The respirometers are based
on measuring the rate at which biomass takes up dissolved oxygen from the liquid phase
(Vanrolleghem et al., 1999). Assessment of wastewater components is often referred to as
wastewater characterization. The procedures for characterization involve a combination of
physic-chemical and biodegradation tests. Using this method, the biodegradable components
in the wastewater can be quantified (Vanrolleghem et al., 1999).
2.6.1 Respirometer
In principle, the respirometer consists of an oxygen electrode, DO meter, recorder,
respirometric reactor and water jacket vessel to maintain a constant temperature. It is placed
on a magnetic mixer in order to obtain a complete mixing of the reactor volume. A ceiling of
the respirometric cell is oblique, so that the air bubbles can easily escape from the cell. The
expansion funnel is used for adding the substrate solution and for escaping air bubbles during
periods of aeration. A cross-sectional area of the funnel stalk is small enough to minimize
oxygen absorption during the measurement. (Fig.2.23).
Figure 2.23 Schematic diagram of respirometer
1. Respirometric cell 2. Water jacket 3. DO probe
4. Air diffuser 5. Magnetic bar 6. Expansion funnel
7. DO meter 8. Recorder
2.6.2 Experimental Procedure
An expected concentration of endogenous activated sludge is transferred into the respirometry
and aerated to increase the dissolved oxygen concentration to 6-8 mg/L. When these
concentrations are reached, the aeration is stopped. A slow decrease in oxygen concentration
is due to heterotrophic endogenous respiration. A typical respirogram is shown in Fig. 2.24,
and can be interpreted as follows (Cech et al., 1984):
39
Time, minute
DO
, m
g/L
AB
C
ED
OURx,e
OURx,t
OC
Adding substrate
Endogenous phase
Figure 2.24 Recorder chart with a typical respirogram (Cech et al., 1984)
During the endogenous phase of respiration, heterotrophic microorganisms utilize
oxygen at a constant rate over a relatively long period of time, as demonstrated by the line A-
B-C. At time B, a small volume of concentrated substrate solution is injected into the cell by
means of a hypodermic syringe. Addition of a limited amount of substrate to the respirometry
reactor causes a temporary increasing respiration rate, as shown by the line B-D. This line is a
maximum-value tangent to the curve B-E. It represents the constant total respiration rate at the
substrate concentration S added. When the substrate concentration decreases with time, the
respiration rate also decreases. When the substrate has been removed (at point E) the
respiration rate returns to a value (line D-E), which is equal to, or perhaps slightly different
from, the original endogenous rate.
When the measurement of one concentration is finished, a new dose of substrate can be
injected into the cell and a next respirogram is recorded (Cech et al.,1984). In order to
evaluate a respirogram, the endogenous respiration rate (OURx,e), the total respiration rate
(OURt) and net oxygen consumption (OC) are calculated. The line section CE is equal to net
oxygen consumption (Fig. 2.25). If the OC value is higher than 4 mg/L O2, the determination
of OC is conducted using Ekama et al. (1986) method (Fig. 2.25). The high OC value occurs
when a high substrate concentration is introduced. This method is normally used for
determination of COD fraction (i.e. biodegradable COD/total COD).
In this test a preselected volume of wastewater of known total COD is mixed with a
preselected volume of mixed liquor of known MLVSS concentration in a batch reactor. After
mixing, the OUR is measured approximately every 5 to 10 minutes until OUR attains to a
constant value that is approximate or equal to OUR in the endogenous phase (Ekama et al.
1986). The respirogram is obtained by plotting the curve of OUR versus time (Fig. 2.25).
40
Time, minute
OU
R, m
g/L
.h
f
A
B
C
D
T
g
e
Figure 2.25 OUR response in respirometer (Ekama, et al., 1986)
Where
Area A: This area gives the concentration of Readily Biodegradable COD
(RBCOD) oxidized by the biomass. This is useful for assessing the
amount of volatile fatty acids (VFA) that needs to be added in a
biological phosphorus removal plant.
Area B: This area represents the amount of less readily biodegradable material
being oxidized.
Area C: This area shows the amount of oxygen being used to convert ammonia
into oxidized nitrate (nitrification).
Area D: The area under the whole curve shows the total oxygen demand of the
liquor. This is the total amount of oxygen which must be supplied to the
sludge to achieve full treatment.
OUR at line e: The respiration rate at the end of the curve, when at least 95% of the
organic waste has been treated, is the endogenous respiration rate. This
rate is proportional to the activity of the biomass.
OUR at line f: This rate is termed the Average Viability, and it is the average
respiration rate for the period where nitrification and the breakdown of
less readily biodegradable substrates are occurring.
OUR at line g: This is the maximum respiration rate observed at the start-up of the
respiration cycles. At this point all oxidative reactions take place,
including the oxidation of carbon and nitrogen compounds and the
uptake of phosphates.
Time T: The time for the sample to reach an endogenous respiration rate. This is
a direct method to determine the minimum HRT required to achieve at
least 95% treatment efficiency.
2.6.3 Determination of Kinetic Constants
Specific OUR of substrate oxidation at a substrate concentration S (OURx,ox) is given
by:
41
eXtXoxX OUROUROUR ,,, (2-22)
Where:
OURx,t = Total respiration rate (mg O2/mg VSS.h)
OURx,e = Endogenous respiration rate (mg O2/mg VSS .h)
Further specific substrate removal rate at a substrate concentration S (RX) is given by:
SOC
OURR oxX
X/
,(2-23)
Where
RX = Substrate removal rate (mg COD removed/mg VSS.h)
OC = Net oxygen consumption (mg O2/L)
S = Substrate concentration (mg COD/L)
OC is then equal to the area between the OUR curve and the second plateau level where
the OUR decreases rapidly and levels off (OC = Area A+area B) (Figure 2.25)
Biomass yield coefficient (Y) is expressed as:
S
OC
fY 1
1(2-24)
and the specific growth rate ( ) as:
XRY . (2-25)
Where
= Specific growth rate (h-1
)
f = COD/VSS ratio of the sludge (mg COD/mg VSS)
Y = Yield coefficient (mg VSS/mg COD removed)
2.7 Membrane Bioreactor (MBR)
MBR is the combination of two basic processes: (1) biological degradation and (2)
membrane seperation into a single process where suspended solids and microorganisms
responsible for biodegradation are seperated from the treated water by a membrane filtration
unit (Manem and Sanderson, 1996).
MBR systems have two principal configurations: (1) The submerged MBR (or
integrated MBR) in which outer-skinned membrane is submerged inside the bioreator and
permeate is extracted by suction or by pressuring the bioreactor; (2) The external circuit MBR
(or recirculated MBR) in which the mixed liquor is recirculated at high pressure through a
membrane module placed outside the bioreactor (Fig. 2.26). The permeate is extracted by
high cross-flow velocity through the membrane and the concentrated mixed liquor at the feed
side is returned to the bioreactor. Excess sludge is withdrawn in order to maintain constant
sludge age and the membrane is regularly cleaned by air or water backwash and chemical
cleaning (Visvanathan et al. 2000).
42
Effluent
Membrane filtration
Bioreactor
Air
Inffluent
Recirculation
Inffluent
Air diffuser
Bioreactor
Air
Inffluent
Membrane module
a. Recirculated MBR b. Surbmerge MBR process
Figure 2.26 Diagram of membrane bioreactor processes
2.7.1 Advantage of the MBR Process
The main advantages of MBR process can be listed as follows:
High quality of treated water: Biological treatment using the MBR process can obtain
extreme high removal efficiency of SS, COD, BOD and pathogen concentration.
Therefore the treated water can be discharged directly into the surface water, or reused
for cooling, toilet flushing and lawn watering.
Flexible in the operation: SRT is independent on HRT and can be controlled completely.
Long SRT can be maintained to allow the development of slow-growing microorganisms,
such as nitrifying bacteria.
Compact plant size:Due to the high biomass concentration that can be maintained in
MBR, a high volumetric loading rate can be applied which results in the reduced size of
bioreactor. In addition, secondary settling tank, sludge thickener or post treatment for
further BOD and SS removal are not necessary in the MBR process, and thus the plant
becomes more compact.
Independence of settling ability: the selection of microorganisms present in the
membrane bioreactor is no longer dependent on either their ability to form biological
flocs or the settling characteristics (Manem and Sanderson, 1996).
Low sludge production: Maintaining a low F/M ratio results in minimum sludge waste.
High degradation rate: High tangential velocities limit floc size and lead to an increase in
mass transfer rates of microorganisms.
2.7.2 Main Design Parameters
In order to have an optimal MBR process from the economic point of view, many
parameters should be considered. These can involve membrane selection, membrane
43
performance (permeate flux, transmembrane pressure, viscosity) and biological performance
(MLSS, SRT, HRT, F/M ratio) and economic considerations (energy consumption, sludge
treatment and disposal cost). These parameters can influence each other, and are mutually
dependent. For example, high MLSS can be maintained by controlling long SRT and thereby
increasing the volumetric loading (reducing HRT) and reducing the investment costs.
However, high MLSS requires high maintenance energy consumption due to increases in
viscosity that results in flux decline, high oxygen demand for aerobic organic degradation and
cell growth (Manem and Sanderson, 1996; Visvanathan et al. 2000).
Table 2.19 presents a comparison between biological performances of MBR and a
conventional AS process. This table shows that the mixed-liquor volatile suspended solid
concentrations (MLVSS) in aerobic MBR process are much higher than those in conventional
AS processes, reaching concentrations of 30 g/L. Based on particle size distribution tests,
Nazim et al. (1999) indicated the AS sludge contained large flocs, while the MBR sludge
contained small flocs. The high mass loads applied to aerobic MBR can be explained by high
biomass concentrations and high specific substrate removal rate (Manem and Sanderson,
1996). HRTs of 2-4 hrs and F/M ratio of 0.1 g COD/g MLSS.d are normally applied to
domestic wastewater treatment (Table 2.19). A combination of carbonaceous BOD removal
and nitrification can achieved high efficiency in MBRs. Muller et al. (1995) reported using
MBR process for domestic wastewater treatment could obtain 90% carbon removal and 100%
nitrification efficiency at COD loading rates of 0.9-2.0 kg/m3.d. (HRT of 2.0 h - 7.5 h). Thus
slow-growing nitrifying bacteria are retained by the membrane in the reactor. The aerobic
MBRs are capable of high-strength industrial wastewater treatment. High organic loading
rates (1.7- 8.6 kg COD/m3.d) can also applied.
44
Tab
le 2
.19 C
om
par
ison b
etw
een b
iolo
gic
al p
erfo
rman
ces
of
MB
R p
roce
ss a
nd c
onven
tional
AS
pro
cess
Par
amet
ers
Do
mes
tic
was
tew
ater
Do
mes
tic
was
tew
ater
Co
mbin
ed
do
mes
tic
was
tew
ater
+
ind
ust
rial
was
tew
ater
Veg
etab
le
cannin
g
Oil
y
was
tew
ater
Ice
crea
m
Fer
men
tati
on
was
tew
ater
conv
enti
onal
AS
Ty
pe
of
mem
bra
ne
MF
0.1
m
MF
0.1
m
MF
0.1
mM
F
MF
T
ub
ula
r U
F
UF
C
om
ple
ted
mix
ing
VL
R,
kg B
OD
/m3.d
1.3
5
5.4
-
12.8
k
g C
OD
/m3.d
1.5
1.2
0.4
7
- 8.6
6.2
0.8
-1.7
0.8
-1.9
HR
T
4 h
-
24 h
-
13 h
5.8
d
2.7
d
6-8
h
F/M
rat
io, g/g
.d
0.1
0.1
0.2
0.5
0.6
-0.8
-
0.4
0.2
-0.6
Slu
dg
e co
nce
ntr
atio
n,
g/L
14
11
2.5
-3.0
11
15-2
5
16
10-1
5
2.5
-4.0
SR
T
50-1
00
50
25
5-1
0
CO
D r
emov
al, %
>
95%
96
96
99
97
97
92-9
8
99
DO
-
0.5
-1.5
>
3
6.3
-
- 2.0
-3.5
2
.0
Y
0.2
0.2
-
So
urc
e Y
amam
oto
et a
l. (
1989)
Bu
isso
n e
t
al. (1
99
8)
Tro
uv
e et
al.
(19
94
)
Kra
uth
and
Sta
ab
(19
93
)
Fu
chs
and
Sch
olz
(20
00
)
Sco
tt a
nd
Sm
ith
, 1
97
7
Lu
et
al.
(19
99
)
Met
calf
an
d
Ed
dy
(19
91
)
45
2.7.3 Membrane Fouling
Membrane fouling or flux decline which leads to high-energy consumption and a large
cleaning chemical requirement is a major problem hindering the widespread application of
biomembrane reactor process. Membrane clogging in the MBR process might be the result of
(a) the biofilm growth, or adsorption or deposition of foulants on the top surface of the
membrane (external fouling) and (b) at the pore entrances or within the internal pore structure
of the membrane (internal fouling). Fig. 2.27 schematises the fouling mechanisms.
Adsorption is used here to mean an interaction between foulants and membrane. The
adsorption arises from physical forces which involve electrostatic forces (surface electric
charges), Van Der Waals forces (attractive forces in close proximity), solvation forces
(hydrogen bonds) and steric forces (attachment of polymers on the surfaces).
ConcentrateFluid flow
Permeate
macromoleculeparticle
membrane pore
Figure 2.27 Diagram of fouling mechanisms (adsorption and deposition)
The phenomenon of fouling is very complex and depends on physical chemical
parameters such as concentration, pH, ion strength and surface properties of particles and
membrane such as electrical charges, hydrophilicity or hydrophobicity. For example, a more
hydrophilic membrane can decrease the ability of adsorption and fouling rate.
Macromolecules (proteins, EPS) which are normally hydrophobic adsorb easily on
hydrophobic membranes. The adsorption layer is also more difficult to wash out from a
hydrophobic surface than from a hydrophilic one. Surface charges can also have an effect on
fouling. If there is electrostatic repulsion between the membrane surface and bioflocs or
macromolecules, fouling is decreased. Stability of macromolecules is influenced by the pH of
mix-liquor. Macromolecules are more compact at their isoelectric point, pI, where
intramolecular electrostatics repulsion is minimum. Thus rapid flux decline occurs at this pH
value due to an increase in the amount of deposited macromolecules.
a. Foulants
Three main types of foulant can be differentiated (Mulder, 1996):
Organic precipitates (biological substances, macromolecules, etc.): Macromolecules can
be protein molecules in wastewater, extracellular polymers (ECP) or long chain organic
by-products generated from biodegradation process.
Inorganic precipitates (metal hydroxides, calcium salts, etc.): Changes in environmental
conditions (pH, solute or ion strength) due to microorganism actions in MBR can form
precipitates. Gelatinous precipitates (such as hydrated complex of calcium phosphate and
citrate) can seriously foul membranes (Howell and Nystrom, 1993).
46
Particulates (cells, debris, microbial flocs, etc.): Particulates in the mixed liquor build-up
the solid cake on the surface of membrane, which results in a decline in flux.
b. Biofouling
Biofouling can be defined as adsorption/adhesion and growth of microorganisms which
forms biofilm on the membrane surfaces. Adhesion can be due to bonding interactions
between the membrane surface and adhesive structures such as flangella, fimbria, or
macromolecules (proteins, extracellular polymers) on the cell surface. Once attached, cells
may grow and multiply by using substrates and nutrients from the bulk solution (Fig. 2.28).
Harry et al.(1996) postulated flux decline can be significantly attributed to extracellular
polymers (EPS) rather than to the colloidal nature of bacterial cells.
cell # 1
primary adhesionEPS
surface EPS charges
cell # 1
GrowthGrowth
secondary adhesion
membrane
Figure 2.28 Schematic illustration of membrane biofouling process (Ridgway and Flemming,
1996).
Extracellular polymers (EPS)
The production of EPS is a general property of micro-organisms in natural
environments, and occurs in bacteria, algae, yeast and fungi (Flemming and Wingender,
2001). EPS are major component of activated sludge (AS) matrix and biofilms. They play an
important role in bioflocculation, settling and dewatering of AS process and in biofilms
development in attached-growth process. They consist largely of proteins, polysaccharides
and humid substances. EPS act as a bridge between cell surface and therefore initiate floc or
biofilm formation (Bura et al. 1998). However, the presence of high EPS concentration may
result in poor settling (bulking phenomenon) or dewatering condition (i.e. increasing sludge
volume index SVI, or capillary suction time CST) in conventional AS process. In addition,
high EPS concentration can increase the specific hydraulic resistance (R) of the filtration cake
in MBR process (Manem and Sanderson, 1996). Fig. 2.29 presents the schematic diagram of
the matrix of biofloc or biofilm.
47
PO43-
Divalent cation
COO-
bacteria cell
EPS
Figure 2.29 Schematic diagram of biofloc or biofilm
Various experimental studies have demonstrated the important role of macromolecules
in fouling and flux decline. Hodgson et al.(1993) investigated the role of the bacterial EPS in
cake resistance of MF system (0.2 m membranes at 100 kPa in batch filtration cells). The
gram-negative marine bacterium SW8 was used in this study. The role of the EPS in resistance
was confirmed by changes in flux through treated and untreated bacterial cakes. The treated
bacterial suspensions here means bacterial cells whose EPS were removed by proteolytic
enzyme and chelating agent (EDTA). Whereas the untreated cake consists of particular cells
with the void spaces filled with EPS proteins and polysaccharide The authors found that the
untreated cake had more higher resistance than the treated cakes. This confirmed that the
major cause of resistance was not the bacterial particles themselves, but the EPS associated
with those bacteria.
Likewise, Nagaoka et al. (1996) carried out a study on the influence of bacterial EPS on the
membrane separation AS process. Loop-type hollow fiber membrane modules with pore size
of 0.1 m were used. The experiments were intermittently operated with a cycle of 10-
minutes-run and 5-minutes-off. Feed wastewater was a mixture of acetic acid (as carbon
source) and necessary nutrients. The results indicate that EPS which was accumulated in the
aerations and also on the membrane caused an increase of viscosity of the mixed liquor and an
increase in the filtration resistance. There was a linear relationship between the filtration
resistance and viscosity of the mixed liquor, which is caused by rapid attachment of the
suspended EPS.
Mukai et al.(2000) estimated flux decline of ultrafiltration membrane in different
cultural growth phases, different EPS and metabolic product concentrations in AS process.
The authors reported that flux decline was effected by protein to sugar ratio of EPS and
metabolic products. Lower permeate flux occurred with higher retention of protein and greater
amounts of retained protein during filtration.
The other influences
Autolysis occurring at high SRT or starving conditions (low F/M ratio) can lead to
increase in concentrations of cell debris and soluble microbial by-products (such as humid
substances and proteins). Absorption of these substances on the membrane surface can boost
biofouling (mainly internal fouling). In addition, high hydraulic stresses may also enhance
cell and floc breakage that releases metabolites and cell debris.
48
The effects of MLSS, soluble COD and viscosity on membrane fouling was estimated
by Sato and Ishii’s model as follows (Manem and Sanderson, 1996): 326.0368.1926.0 )(*)(*)(**7.842 CODMLSSPR (2-26)
Where
R = Filtration resistance, m-1
P = Transmembrane pressure, Pa
= Viscosity, Pa.s
MLSS = mixed liquor suspended solid, mg/L
COD = Soluble chemical oxygen demand, mg/L
Rt = Total resistance for filtration, m-1
It can be seen that soluble COD concentration can contribute significantly to increase in
filtration resistance (Eq. 2.26).
49
Chapter 3
3
Methodology
This research comprises four main studies: (1) biokinetic study; (2) parametric study
(optimization of operating conditions); (3) biomembrane study, and (4) sludge
characterization study. Flowchart of the different phases of the experimental studies is shown
in Fig. 3.1.
(1) Biokineticexperiments
Aclimatized yeast andbacterial sludge
(2) Parametric study(Optimization of
operating conditions)
Glucose
Protein extractSRT
pH
(3) MBR study
High influent COD
5,000 mg/L
VLR
Low influent COD1,000 mg/L
SRT
(4) Sludgecharacteristics
(4) Sludgecharacteristics
DO
Figure 3.1 Flowchart of different phases of experimental study
3.1 Biokinetic Study
The objective of this study was to evaluate the biokinetic coefficients of mixed yeast
and mixed bacterial treatment of high salinity wastewater by means of respirometric
techniques. Two feed synthetic wastewaters were used, namely: glucose-feed wastewater and
protein-feed wastewater. In the protein-feed wastewater, commercial tuna fish protein extract
(T.C. Union Agrotech Co., Thailand) was mixed with tap water to obtain wastewater whose
composition was similar to that of tuna fish processing wastewater. The composition of two
feed wastewaters is presented in Tables 3.1 and 3.2. The flowchart of biokinetic study is
shown in Fig. 3.2.
50
Table 3.1 Composition of glucose-feed wastewater (Defrance, 1993)
Component Concentration (*)
(mg/L)
Glucose 4,673
(**)
(NH4)2SO4 1,870
Yeast extract 94
KH2PO4 235
MgSO4.7H2O 467
ZnSO4.7H2O 0.5
CaCl2 1.0
MnCl2 1.0
FeCl2 1.0
(NH4)6Mo24.4H2O 0.2
CuSO4 0.2
CoCl2 0.2 (*) Composition of synthetic wastewater used for bacterial system was similar to that for yeast culture, but the
concentration of each component was five times lower than that of the yeast culture. (**) Corresponding to COD concentration of 5,000 mg/L and BOD20 concentration of 4,440 mg/L
(Biodegradability of glucose is 0.95 g BOD20/g glucose (Henze et al., 1997).
Table 3.2 Composition of protein-feed wastewater
Parameters Concentration (mg/L)
COD 5,000
BOD5 3,850
Organic-nitrogen 690
Ammonia-nitrogen 26
Total phosphorous 45
Seed bacterial sludge
Enrichment
Acclimation ofhigh salinity
- Glucose as substrate- up to 45 g/L NaCl
Glucose Protein extract
Biokineticexperiments
Seed yeast sludge
Acclimation ofprotein extract
51
Figure 3.2 Flowchart of biokinetic experiments
3.1.1 Seed Sludge
a. Yeast sludge and enrichment
The term mixed yeast sludge implies the mixture of all wild yeasts which exist in the
raw wastewater and then quantitatively propagate under proper enrichment conditions. The
procedure of enrichment for yeasts was carried out according to the Standard Methods for the
examination of water and wastewater (APHA et al., 1995). The yeast strains were selected by
the enrichment culture technique based on free competition among different organisms in
wastewater (Nishihara Ltd., 2001). Figure 3.3 shows the procedure of the enrichment process.
The osmotolerant yeast sludge was enriched from the bottom sediments of an
equalization tank of a fish sauce factory located in Rayong province, Thailand. This tank
received wastewater containing high salt and organic content (30.2 g/L NaCl and 800 mg/L
COD). The enrichment was conducted using two-liter container and the fill-and-draw
operation. In the first batch, the raw sediment was added to two containers containing two
liters of feed wastewater with 20 g/L and 32 g/L NaCl. MLSS concentration of mixed liquors
obtained were around 1,000 mg/L. The feed wastewater (glucose as substrate) was mixed
using a diffused aeration system, and pH was adjusted to 3.5 in order to optimize yeast growth
and limit the bacterial contamination (Pelczar and Reid, 1972; Elmaleh, et al., 1996). After
eight hours of aeration, the biomass suspension was allowed to settle for 12 hours. Yeast cells,
normally, settle to the bottom; acid-tolerant bacteria and filamentous fungi remain in
suspension. The bacteria and fungi are removed by decanting supernatant. 1.5 liters of
supernatant was decanted and a fresh medium was added for next batch. When yeast biomass
(MLSS) exceeded 3,000 mg/L, the enrichment process was stopped.
Filling
Aeration(32 h)
Completion
Seed yeast sludge(sediments)
MLSS>3000 mg/L
Settling(10 h)
Drawing
no
yes
Feed-glucose-wastewater
Figure 3.3 Schematic diagram of enrichment procedure
52
b. Bacterial sludge
The bacterial seed sludge was obtained from the activated sludge process of the same
fish sauce processing wastewater treatment plant. This plant treats combined wastewater with
salt content of approximate to 1g/L NaCl and mean COD concentration of 540 mg/L.
3.1.2 Acclimation
Acclimation was carried out to obtain mixed bacterial and yeast sludges that can
tolerate salt contents (32 and 45 g /L NaCl). The two-litre-batch-reactors with fill-and-draw
operation were used in the acclimatization stage. Table 3.3 shows the operating conditions. In
order to obtain operating conditions similar to saline seafood processing wastewater treatment
using yeast treatment followed by mixed bacterial system, the initial COD concentrations of
5,000 mg/L for yeast and 1000 mg/L for mixed bacterial culture were selected.
Table 3.3 Operating conditions for high salinity acclimation
Operating conditions Yeast sludge Bacterial sludge
Initial COD (mg/L) 5,000 1,000
pH 3.5 7.5
Temperature ( oC) 25 – 32 25 – 32
MLSS (mg/L) 5,000 4,000
HRT (h) 36 24
After aeration (24 h for bacterial sludge and 32 h for yeast sludge), the biomass suspension
was allowed to settle for 12 hours, and the supernatant was sampled and centrifuged at 4,000
rpm for 15 min. COD of the supernatant was analyzed. When the COD removal efficiency
was less than 80%, the experiment was repeated at the same operating conditions. When the
COD removal exceeded 80%, the NaCl concentration was increased by 3 g/L. Acclimation to
high salt contents (32 and 45 g NaCl/L) was assumed to be completed when 80% COD
removal was attained.
3.1.3 Biokinetic Experiments
The kinetic coefficients of the acclimated yeast and bacterial cultures at different salt
contents with two substrates (glucose and protein extract) were assessed using a closed 0.9
liter batch respirometer, equipped with a recorder, DO meter and water jacket vessel to
maintain a constant temperature as shown in Figure 3.4. Table 3.4 presents the operating
conditions of the respirometric experiments for yeast and bacterial culture. Initially, glucose
was used as a substrate complemented with necessary nutrients. The So/Xo ratio (initial
substrate concentration/biomass concentration) governs the quality of the batch respirometric
tests (Cech et al., 1984). In order to obtain So/Xo ratio ranging from 0.01 to 0.20 (Mathieu and
Etienne, 2000), the Xo and So values should be ranged from 1,500 to 2,000 mg VSS/L and 20-
500 mg COD/L, respectively.
53
8
6
9
5
4
7
3
2
1
1. Respirometric cell 2. Water jacket 3. Air diffuser
4. DO probe 5. Magnetic bar 6. Magnetic mixer
7. Expansion funnel 8. DO meter 9. Recorder
Figure 3.4 Respirometer set-up
Table 3.4 Operating conditions for the respirometric experiments
Operating conditions Yeast sludge Bacterial culture
Initial pH 3.5 7.5
Temperature (oC) 30 0.5 30 0.5
Xo (mg MLSS/L) 1,500 1,500 (1)
Substrate (mg COD/L) 20 – 500 20 – 200
So/Xo ratio 0.01 – 0.35 0.01 – 0.15 (2)
Suppressing nitrification None Adding 70 mg N-ammonia/L (3)
Sources: (1)
Cech et al., (1984) (2) Chudoba. et al., (1992) (3) Liebeskind, (1999)
The experimental procedure for OUR determination are summarized below:
1) Obtaining endogenous sludge: The respirometer was filled with fresh sludge without
substrate and aerated at least for 2 hours.
2) Suppressing nitrification: NH4Cl was used with the concentration of 70 mg N-
ammonia/L. Liebeskind (1999) postulated that if ammonia was present in wastewater,
organic oxidation and nitrification simultaneously occurred. At high enough ammonia
concentration (70 mg/L N), OUR of nitrification is constant during organic oxidation.
When this ammonia dose is added to the endogenous sludge, nitrification OUR will be
determined. Thus, OUR of organic oxidation will be the difference between total OUR
and the sum of endogenous OUR and nitrification OUR.
3) Recording endogenous OUR: After suppressing the nitrification process, the mixture
was aerated at least half an hour before measuring endogenous OUR.
54
4) Adding substrate: An accurate amount of substrate was added to the respirometer and
total OUR was recorded by respirogram. New reaeration was necessary when the
dissolved oxygen concentration dropped below 2 mg/L.
The results of the respirometric experiments provided OUR data that are used for
calculating specific growth rates ( ) using Equations 2.22 to 2.25 in Chapter II. By using
OUR values and specific growth rates ( ) with respect to corresponding substrate
concentrations (S), maximum oxygen utilization rate (OURmax), observed specific growth rate
( obs) and half velocity constant (Ks) at the selected salt content were determined by
regression analysis based on Monod kinetics.
By using observed maximum specific growth rates ( obs) at the corresponding salt
contents (I), the critical salt content above which reaction stops (KI) and maximum specific
growth rate of free-salt solution (d-1
) were evaluated by linear regression analysis based on the
Ghose and Tyagi model (2-9). The regression analysis was done by Grapher sofware.
3.2 Parametric Study
3.2.1 pH values
The effects of pH values on bacterial and yeast treatment systems were also evaluated in
terms of OUR by respirometer. The protein-feed wastewater with 32 g salt/L was used. The
experiments were done at constant temperature of 30 0.5oC. The operating conditions are
presented in Table 3.5.
Table 3.5 Operating conditions for the pH effect experiments
Operating Condition Units Values
COD mg/L 50
MLSS mg/L 2,000
Temperature oC 30 0.5
Salt g/L NaCl 32
pH values:
+ For mixed yeast sludge 2.5 – 9.0
+ For mixed bacterial sludge 4.0 – 10.5
The main procedure of the experiment can be described as follows:
Obtaining endogenous sludge: The respirometer was filled with fresh sludge without
substrate and then aerated for at least 2 hours.
Suppressing nitrification (for mixed bacterial sludge): NH4Cl was used with a
concentration of 70 mg N-ammonia/L
Adjusting pH value: H2SO4 solution 0.1N or NaOH 0.1N solution was added into the
mixture until the desired pH value.
Recording endogenous OUR: After 10 minutes of aeration, the endogenous OUR was
recorded.
Adding substrate: Stock solution of fish extract (25,000 mg/L) had been prepared and
adjusted to the desire pH value before adding into the mixture. An accurate amount of
55
substrate (300 mg/L COD) was injected into the respirometer and then total OUR was
recorded. New reaeration was necessary when the DO dropped below 2 mg/L.
3.2.2 Sludge Retention Time (SRT)
Each of the SRT variation experiments was conducted in a two-liter batch reactor using
the fill-and-draw operation. Here, the MLSS variation was monitored for a minimum period
of 3 weeks. The steady state condition was reached when the MLSS values remain constant
for at least 5 days. The experimental operating conditions are presented in Table 3.6. To avoid
the effects of protein precipitation (at low pH) on the nitrogen uptake ability of yeast sludge,
removal of protein precipitates prior to feeding is necessary. Raw wastewater was acidified to
pH 3.5 and left to settle at least for 12 hours at 40C.
Table 3.6 Operating conditions of the experiments on SRT effect
Operating conditions Unit Value
COD mg/L 5,000
pH 3.5-4.0
MLSS mg/L 7,000
HRT h 24
Salt g/L 0.5, 15, 32 and 42
SRT d 5, 7, 10, 20 and 45
Sludge retention time is defined as the average time for which a unit of biomass remains
in the system. For a completely mixed process with a sludge return arrangement, SRT can be
expressed mathematically as follows:
eww
rc
XQQXQ
XV
)( (3-1)
Where
c = Sludge retention time, d
Qw = Sludge waste rate, L/d
Q = Influent flow rate, L/d
Vr = Volume of aeration tank, L
Xe = Volatile suspended solids in effluent
X = Mixed liquor volatile suspended solids in the aeration tank, mg/L
To facilitate operation of lab-scale experiment, it is assumed that solids lost in the effluent can
be neglected (Xe = 0). Equation 3-1 can be rewritten as:
w
rc
Q
V
c
rw
VQ (3-2)
Equation 3-2 provides the calculation for sludge volume to be wasted daily from the
reactors. After 24 hours of aeration, waste sludge was withdrawn. Due to poor settling ability
of mixed yeast sludge, settling step in a typical batch process was replaced with
56
centrifugation. The remaining of suspended biomass was centrifuged at 3,000 rpm for 15 min.
The centrifuged sludge was returned to the reactor for the next batch.
3.3 Biomembrane Study
The biomembrane study was conducted with the fish-protein-feed wastewater at 32 g/L
NaCl. This study consisted of two phases, namely: (1) high COD loading, and (2) low COD
loading. The difference between the two phases is shown in Table 3.7.
Table 3.7 Difference between the high COD loading and low COD loading
Parameter High COD loading Low COD loading
Influent COD (mg/L) 5,000 1,000
Experimental
set-ups Feed wastewater
(5,000 mg/L COD)
BMBRYeast reactor
YMBR
Feed wastewater (1,000 mg/L COD)
BMBR
YMBR
SRT 50 d for both YMBR and BMBR;
15 d for yeast reactor
50 d and 10 d were
investigated for both YMBR
and BMBR
3.3.1 High COD loading
The fish-protein-feed wastewater used in this phase was similar to that used for
biokinetic study (Table 3.2). Both yeast and bacterial sludges were excess sludges obtained
from the biokinetic study. Two parallel experimental set-ups were conducted, namely: (1)
yeast pretreatment followed by the Bacterial Membrane Reactor (BMBR) as schematized in
Figure 3.5, and (2) the Yeast Membrane Reactor (YMBR).
Both YMBR and BMBR tanks were made of transparent acrylic tube of 10 cm in
diameter with working volumes of 8 L and 3.6 L respectively. In order to provide enough
effluent volume to BMBR at different HRTs, the working volume of the yeast reactor (YR)
was made adjustable by changing outlet points installed along the height of the column. The
total volume of the yeast reactor was 21 L. These reactors were continuously aerated through
the stone diffusers placed at the bottom, and equipped to monitor pH and DO. For the YMBR
system, through an external pH dosing pump, the reactor pH values were maintained at the
required level. In each reactor, a polyethylene 0.1 m hollow fiber membrane module with a
surface area of 0.42 m2 was fixed on the upper end.
Speed-controlled roller pumps were used to withdraw the permeate from these
membrane modules. Both bioreactors were operated with periodic air backwashing (20
minutes filtration and 2 minutes of air backwashing at a pressure of 400 kPa arrangement. The
alternative operations of filtration and air injection was controlled by an intermittent
controller and solenoid valves. The transmembrane pressure was measured using a mercury
manometer.
For the BMBR system the feed wastewater was pretreated with the yeast reactor with
continuous aeration through diffusers with a mean HRT of 32 h. The pH was maintained at
3.5. The treated effluent from this reactor was continuously withdrawn and sent to the
sedimentation unit where the yeast sludge is separated then returned to the yeast reactor. The
hydraulic retention time of the settling tank was 5 h, which corresponds to the shortest HRT
of BMBR (4.5 h). From the yeast reactor, excess sludge was periodically withdrawn to
57
maintain a mean sludge retention time (SRT) of 15 days and mean biomass concentration of
4,500 mg/L MLSS. The settled effluent was stored in an intermittent storage “effluent tank”
and fed into the BMBR reactor system. In the BMBR system, the pH was maintained at above
7.0, whereas in the YMR, the feed water did not undergo this pre-treatment, but was fed
directly into the YMBR tank where the pH was maintained at 3.5.
The investigations were carried out by step-wise increase in volumetric loading. The
different loading steps used is summarized in Table 3.8. Each of volumetric loading rate
(VLR) variation was maintained for at least 7 days. The mean influent COD concentration fed
into the BMBR was the effluent COD from the yeast reactor.
Table 3.8 Experimental operating conditions of YMBR and BMBR systems
YMBR BMBR
Stage Time
days
VLRkg
COD/m3.d
Mean influent
COD
mg/L
Mean
HRT
h
Time
days
VLR kg
COD/m3.d
Mean
influent
COD(*)
mg/L
Mean
HRT
h
I 1-22 5.0 5,000 24.0 1–11 2.1 1,200 13.7
II 22-29 3.4 5,000 36.0 11-21 3.4 1,280 9.1
III 30-40 6.6 5,000 18.2 22-31 7.9 1,450 4.5
IV 41-51 9.9 5,000 12.2 32-41 5.3 1,080 4.9
V 52-62 16.3 5,000 7.1 42-51 1.7 1,170 16.1
VI 63-78 23.0 5,000 5.1 52-75 2.4 1,140 11.7
VII 77-90 3.6 1.450 3.6
(*) Effluent from the yeast reactor
58
Fee
d t
ank
Bac
teri
al e
xce
ss s
ludge
BM
R e
fflu
ent
tank
Eff
luen
t ta
nk 1
Yea
st e
xce
ss s
ludge
Overflow
YE
AS
T R
EA
CT
OR
BA
CT
ER
IAL
ME
MB
RA
NE
RE
AC
TO
R S
YS
TE
M
Com
pre
ssed
air
Sulp
huri
c ac
id s
ol.
Yea
st e
xce
ss s
ludge
Com
pre
ssed
air
Tim
er
Air
bac
kw
ash
Yea
st e
fflu
ent
tank
Lev
el w
ater
tank
Overflow
to f
eed t
ank
Sulp
huri
c ac
id s
ol.
YE
AS
T M
EM
BR
AN
E R
EA
CT
OR
Mem
bra
ne
reac
tor
Outl
et p
oin
t
(pH
3.5
-3.8
)
(pH
3.5
)
Yea
st e
xce
ss s
ludge
Com
pre
ssed
air
Tim
er
Air
bac
kw
ash
Lev
el w
ater
tank
Hg-m
anom
eter
Hg-m
anom
eter
NaO
H s
ol.
(to p
H 6
.8-7
.5)
Bac
teri
al
Fig
ure
3.5
Mem
bra
ne
reac
tor
syst
ems
in t
he
hig
h C
OD
load
ing
59
3.3.2 Low COD loading
This phase was a sequence of the high COD loading. Therefore, yeast and bacterial
sludges were acclimated. Both YMBR and BMBR had the same working volume and influent
wastewater with COD of 1,000 mg/L. The experimental set-up is schematized in Fig. 3.6.
Both YMBR and BMBR tanks were made of transparent acrylic tube of 15 cm diameter with
working volume of 10 L. Likewise, in the high COD loading, the reactor pH values were
maintained to 3.5 in the YMBR set-up. In each reactor a 0.1 m hollow fiber membrane
module with a surface area of 0.42 m2 was fixed on the upper end. The air-backwash
operation of both set-ups in this phase was similar to that of the high COD loading.
The composition of the low COD wastewater is presented in Table 3.9. The experiments
were conducted using a step-wise increase in the flux rate (i.e. increase in the volumetric
loading) at sludge retention times of 10 and 50 d. During the transition period (between SRT
of 10d and 50 d), no sludge was wasted from the reactors except for sampling. The operating
conditions for this phase is shown in Table 3.10.
Table 3.9 Composition of the low COD wastewater
Parameters Unit Concentration
COD mg/L 1,000
BOD5 mg/L 780
Organic-nitrogen mg/L 120
Ammonia-nitrogen mg/L 25
Total phosphorous mg/L 12
Salt content g/L NaCl 32
Excess sludge
Compressed air
Timer
Air backwash
Effluent tank
Level water
tank
(H SO solution only for YMR)
Feed tank
2 4
Hg manometer
Membrane
Bioreactor
Figure 3.6 Schematic diagram of membrane bioreactor
60
Table 3.10 Effects of different HRTs and SRTs on yeast and bacterial membrane reactors
YMBR BMBR
SRT
days
VLR
kg COD/m3.d
Mean HRT
h
SRT
days
VLR kg COD/m3.d
Mean HRT
h
10 2.66 8.8 10 2.97 8.1
10 2.95 7.7 10 3.57 6.3
10 3.66 6.1 10 4.30 5.2
10 4.59 5.0 Transition 3.63 7.1
Transition 3.58 7.2 Transition 4.08 6.0
Transition 4.28 6.1 50 5.56 4.7
50 4.93 5.3 50 6.35 4.0
50 6.55 4.0
3.4 Sludge Characterization Study
To investigate variation of sludge characteristics with salt contents and simultaneously
obtain a comparison between membrane bioreactor and batch systems, yeast and bacterial
batch reactors were operated at different salt contents (0.5, 15, 32 and 45 g/L). Two-liter-
batch reactors with fill-and-draw operation were used. The initial mixed yeast and bacterial
sludges were withdrawn from the YMBR and BMBR run at SRT of 10 d and 32 g salt/L.
These sludges were later acclimatized at salt contents by gradual increases or decreases. Table
3.11 presents the experimental operating conditions. The sludge was sampled when each
batch reached the steady state condition (i.e. COD removal was above 80% with a stable
MLSS value). The sludge was examined for ECP content, dewatering property (CST) and
sludge settleability (SVI). In addition, N and P contents of both yeast an bacterial sludge were
evaluated.
Table 3.11 Operating conditions for the sludge characterization study
Operating conditions Value
Initial COD (mg/L) 1,000
pH:
- Mixed yeast sludge
- Mixed bacterial sludge
3.5-4.0
7.0-7.5
Initial MLSS (mg/L) 7,000
SRT (d) 10
HRT (h) 8
Salt (g/L) 0.5, 15, 32 and 42
3.5 Analytical Methods
All analyses were conducted according to Standard Methods (APHA et al., 1995). COD
was analyzed by the potassium dichromate close reflux method with correction for chloride
interference.
The extraction of extracellular polymers (ECP) is based on the thermal extraction and
ethanol precipitation method (Brown and Lester, 1980). The sludge was separated by
centrifugation (2,000 g for 15 minutes), then washed and resuspended in distilled water. A
portion was taken for measurement of suspended solids and the remaining part was heated at
61
800C for 1h. The extracted polymers were collected by removing the sludge by centrifugation
(9,500 g for 15 minutes). The extracellular polymers in supernatant were precipitated by
adding two volumes of solvent mixture (1:1 acetone and ethyl alcohol) to one volume of
supernatant. It was then left overnight at 4oC. The ECP content was measured by means of
suspended solid analysis. Viscosity of the mixed liquor was measured using a rotating torque
cylinder. Table 3.12 listed parameters and their analytical methods used in this study.
Table 3.12 Parameters and their analytical method
Parameter Analytical method Analytical equipment
Interference of salt
Treatment Source
pH pH meter pH meter None APHA et al., 1995
DO DO meter DO meter None APHA et al., 1995
COD Dichromate Reflux Titration Yes Adding HgSO4
according to the 10:1 ratio of HgSO4:Cl.
APHA et al., 1995
Ammonia Distillation UV-vis Spectro.
None APHA et al., 1995
Nitrite Colorimetric UV-vis Spectro.
None APHA et al., 1995
Nitrate Cadmium reduction UV-vis Spectro.
None APHA et al., 1995
TKN Macro-Kjeldahl Titration Yes Adding conc. H2SO4 APHA et al., 1995
Phosphate Ascorbic acid UV-vis Spectro.
None APHA et al., 1995
TS Dried at 103-105oC Oven None APHA et al., 1995
VS Ignited at 550oC Furnace None APHA et al., 1995
CST Capillary time CST apparatus
None APHA et al., 1995
SVI Settled sludge volume after 30 minutes
1000mL cylinder
None APHA et al., 1995
Viscosity Rotating torque cylinder
None
EPS Thermal and centrifugation method
None Brown and Lester, 1980
SS Dryed at 103-105oC Filter/Oven None APHA et al., 1995
62
Chapter 4
4
Results and Discussions
4.1 Biokinetic Study
Respirometric experiments were used in this study to evaluate biokinetic coefficients in
yeast and mixed bacterial treatment of saline wastewaters containing 20, 32 and 45 g/L NaCl.
Two carbon sources were investigated, namely (1) Glucose-feed wastewater (glucose as
carbon source) and (2) Fish-protein-feed wastewater (fish protein extract as carbon and
nitrogen sources).
4.1.1 Enrichment and Acclimation of Yeast and Mixed Bacterial Sludge
Prior to the biokinetic study, enrichment and acclimation were carried out to obtain a
mixed yeast sludge and mixed bacterial sludge able to tolerate high salt contents (32 and 45 g
NaCl/L). In order to propagate all wild yeasts present in the raw sediments taken from a fish
sauce factory, the enrichment technique was applied prior to the acclimation. pH was adjusted
to 3.5 in order to limit bacterial contamination. The enrichment was completed when the yeast
concentration reached to 3000 mg/L. The enrichment and acclimation were conducted with
two-litre batch reactors using the fill-and-draw operationes.
a. Yeast Enrichment with Glucose-feed Wastewater
During enrichment and acclimation of the yeast culture, it was found that the color of
the sludge changed from black to brown and finally to white. A change in color is a typical
indication of the change in the proportion of different species in any microbiological culture.
Further, microscopic observations revealed that the yeast sludge contained predominantly
spherical yeast cells with few egg-shaped cells and few hyphal filaments. However round
cells budding multilaterally, bipolarly and unipolarly could be easily recognized in the culture
(Fig. 4.1).
Cultural characteristics were estimated by differences of yeast colonies in terms of
shape, texture and margin. To evaluate cultural characteristics, isolation was done for three
yeast mixtures. Isolation was carried out on the yeast-glucose-peptone agar (APHA et al.,
1995). The results of isolation for yeast mixtures at different salt contents show that there
were two predominant colony types. The majority of colonies were round shape, smooth
surface, opaque in color and round edge. The second colony had irregular shape, rough
surface and curled edges (Appendix A).
Figure 4.1 Appearance of yeast cells predominantly grown in glucose-feed wastewater
10 m10 m x 1500
20 m
x 500
63
b. Acclimation of Mixed Yeast and bacterial sludges to High Salt
The acclimation of mixed yeast and bacterial sludges to high salt was conducted with
the glucose-feed and fish-protein wastewaters. In order to obtain feed wastewater composition
similar to saline seafood processing wastewater using yeast treatment followed by mixed
bacterial system, the initial COD of 5,000 mg/L for yeast and 1000 mg/L for mixed bacterial
culture were used. When COD removal reached more than 80% (after 24 hours of aeration),
the NaCl concentration was increased by 3 g/L.
Biomass
The time required for yeast acclimation was about 16 days for an initial F/M ratio of
1.12 g COD/g MLSS.day compared to 26 days for bacterial culture for an initial F/M ratio of
0.5 g COD/g MLSS.day, at 45 g salt/L. Acclimation was assumed to be complete when COD
removal exceeded 80 %. Fig. 4.2 shows variation in COD removal with time at 45 g salt /L
for mixed yeast culture. Similar trend curves were obtained for the mixed bacterial culture
(Fig. 4.3). The asymptotic nature of the curves indicates that COD removal efficiency was
stable after a certain time, which marks the completion of acclimation.
4000
8000
12000
16000
MLS
S, m
g /L
MLSS
30
40
50
60
Salt c
oncentration (
g/L N
aCl)
50
60
70
80
90
CO
D rem
oval
(%
)
0 10 20 30 40 50
Time (days)
COD%
Salt concentration
Figure 4.2 Acclimation of yeast sludge cultured with glucose at high salt contents
During the acclimation, the biomass increased 4.5 times that of the initial biomass
concentration for yeast, as compared to 1.7 times for bacterial culture. Yeast biomass
concentration increased 10,700 mg/L after 40 days, whereas bacterial sludge concentration
increased from 2,040 to 3,400 mg/L. Differences in biomass concentration can be attributed to
the application of higher volumetric loading for yeast sludge (3.3 kg COD/m3.d for yeast and
1.0 kg COD/m3.d for bacterial sludge). Therefore, a large quantity of substrate consumed was
converted to yeast biomass. Even at a higher volumetric loading, the time required for yeast
acclimation was about 60 % that of the bacterial sludge showing a far better adaptability of
yeast at high salt. A better acclimation influences the start-up time of a wastewater treatment
64
plant, and also indicates the tolerance of the culture to occasional salt variation in the glucose-
feed wastewater.
2000
2400
2800
3200
3600
MLS
S, m
g /L
MLSS
10
20
30
40
50
Salt
concentrat
ion (
g/L N
aCl)
50
60
70
80
90
CO
D rem
oval (
%)
0 10 20 30 40
Time (days)
COD%
Salt concentration
Figure 4.3 Acclimation of microbial mixed culture with glucose-feed wastewater as
function of salt
Organic removals
In order to estimate organic removal rates, the COD profiles of acclimatized yeast and
bacteria batches were examined. The COD profile is defined as the quantity of COD varies
with aeration time in a culture batch. The typical COD profiles of a mixed yeast batch at 32 g
salt /L is shown in Fig 4.4. All COD profile data of yeast and bacterial batches are presented
in Appendix B. Table 4.1 summarizes operating conditions and COD removal of yeast and
bacterial sludges acclimatized to glucose-feed wastewater at high salt contents. Optimum
HRT referred to aeration time at which COD removal exceeded 90%.
Table 4.1 Performance of mixed yeast and bacterial batches adapted to glucose-feed
wastewater with high salt
Yeast batch Bacterial batch
Salt content Salt content Parameters Unit
20 g/L 32 g/L 45 g/L 20 g/L 32 g/L 45 g/L
Mean MLSS mg/L 8700 9500 9750 3050 3650 3200
Optimum HRT h 5 9 13 2.5 8 17
F/M g/g.d 2.77 1.39 0.96 3.27 0.84 0.44
Effluent COD mg/L 220 255 290 20 30 70
%COD % 95.6 94.9 94.2 98 97 93
COD removal rate g /g.d 2.65 1.32 0.90 3.20 0.81 0.41
65
20
40
60
80
100
CO
D rem
oval (%
)
0
1000
2000
3000
4000
5000
CO
D (
mg/L
)
0 2 4 6 8 10 12 14
Time (days)
COD%
COD
HRTopt
Figure 4.4 Typical COD and COD removal profile of mixed yeast batch in glucose-feed
wastewater at 32 g salt/L
The COD removal efficiency for acclimated yeast and bacterial cultures at 25, 32 and
45 g salt/L was studied with respect to substrate utilization rate U as shown in Figure 4.5.
Since the F/M ratio for the two cultures differs significantly, the effect of the variation has to
be taken into account for determining COD removal efficiency. COD removal efficiency is
therefore expressed by Equation (4-1) (Metcalf and Eddy, 1991).
24*.
*/
XHRT
EMFU
opt(4-1)
Where
U = Substrate utilization rate (g COD removed/g MLSS.day)
F/M = Food:microorganism ratio (g COD applied/g MLSS.day)
24*.
/ inf
XHRT
CODMF
opt
CODinf = Initial COD concentration (mg/L)
HRTopt = Optimum hydraulic retention time (h)
X = Biomass concentration (mg/L MLSS)
E = COD removal efficiency (%)
It could be observed that, while increasing the salt content, U is significantly decreased
for both cultures due to salt inhibition. However, the rate of decrease is found to be much
higher for bacterial culture than for yeast, indicating that the bacterial culture is much more
sensitive to changes in high salt content. When salt content increased from 20 to 45 g/L, U
decreased from 3.26 to 0.40 g COD/g MLSS.d for the bacterial culture, while U decreased
from 2.65 to 0.88 g COD/g MLSS.d for the yeast culture.
66
0.0
1.0
2.0
3.0
4.0
CO
D rem
oval
rate
(g C
OD
/g M
LSS.d
)
20 30 40 50
NaCl (g/L)
Mixed bacterial sludge
Mixed yeast sludge
XeY 048508906 ..
XeY 10404825 ..
R2= 0.978
R2 = 0.989
Figure 4.5 Variation in COD removal rate versus salt contents in acclimatized yeast and
bacterial mixed cultures
It can be concluded that wastewater containing high salt is better for yeast culture, it
may be better to opt for yeast culture while at low salt bacterial culture is preferred. The
intersection point was found to be at 25 g salt/L, which indicates that below this value
bacterial culture may be a better solution and vice versa.
As indicated in Section 4.1.1.b, the yeast culture is subjected to a higher F/M ratio than
the mixed bacterial culture in order to obtain a comparable substrate removal efficiency. This
can be considered a specific advantage of the mixed yeast culture over bacterial culture.
Normally, for aerobic treatment systems, higher F/M ratio vis-à-vis higher organic loading
puts greater stress on the system. This generally results in lower efficiency of substrate
removal and oxygen utilization. Therefore, in practice, aerobic systems are not subject to
volumetric loading exceeding 1.2 kg COD/m3.d. The normal range is 0.3 to 0.8 kg COD/m
3.d,
which avoids low efficiency and higher O2 requirement (because of low utilization
efficiency). Thus, by allowing a higher F/M ratio on the yeast culture without sacrificing
efficiency, the problem of higher organic loading on an aerobic treatment system is to some
degree addressed.
However, higher organic loading allows downsizing the treatment reactors, which
improves the overall economy of the system. In this study, the mixed bacterial culture was
loaded at 1.0 kg COD/m3.d (corresponding to a F/M ratio of 0.50 g COD/g MLSS.d) as
compared to 5.0 kg COD/m3.d (F/M ratio of 1.12 g COD/g MLSS.d) for yeast culture. This
range of loadings (for yeast) is generally acceptable for anaerobic treatment systems having a
low efficiencies normally in the range of 60 - 70% at best. Moreover, it has also been found
(Feijoo et al., 1995) that anaerobic microorganisms are highly sensitive to salt and total
inhibition could be noted for some treatment systems at a salt exceeding 20 g/L, whereas a
COD removal efficiency of more than 80 % (for 45 g/L) and 90 % (for 20 g/L) can be easily
obtained for yeast culture at HRT of 24 h. Thus, the yeast system is more useful than an
anaerobic system at high salt, and can be considered a better substitute for anaerobic systems
in terms of COD removal efficiency.
67
Protein-feed wastewater
Prior to the estimation of biokinetic constants for protein-feed wastewater, acclimation
of yeast and bacteria sludges to this new substrate was necessary. Both yeast and bacterial
sludges were the ones that had been adapted previously to glucose-feed wastewater at high
salt contents (20, 32 and 45 g/L). This acclimation lasted two weeks. The COD removal of
yeast batches at high salt contents was higher than 80% after 12 days of acclimation (Fig.
4.6). Whereas, COD removal of bacterial batches was higher than 90% after a few days. Thus,
the mixed bacterial sludge was able to acclimatize faster to the substrate with high protein
content. Unlike composition of glucose feed wastewater, the protein-feed wastewater contains
a large complex of organics such as proteins, colloidals and polysaccharides. These substance
can be slowly degraded by most yeast strains grown predominantly in the previous glucose
culture (Defrance, 1993). During acclimation to protein-feed wastewater, yeast strains could
be inhibited due to limitation of glucose, while other strains which are able to degrade the
complex organics grow rapidly and become predominant after acclimation. In general,
complex organics must be hydrolyzed by enzymes (hydrolases) before yeasts can degrade
them. The type of hydrolases produced by yeasts are dependent on the species (Pelczar and
Reid, 1972). This reveals that utilization of mixed yeast culture, based on a symbiotic process,
for treatment of wastewater having complex composition is more efficient than pure yeast
culture.
2000
4000
6000
8000
10000
MLS
S, m
g /L
Yeast sludge
Bacterial sludge
50
60
70
80
90
CO
D rem
oval (
%)
0 2 4 6 8 10 12 14 16 18
Time (days)
Figure 4.6 Acclimation of yeast and bacterial sludges to fish-protein-feed wastewater
containing 32 g/L salt
Sludge characteristics
The change in color of the sludge and microscopic observations strengthen the fact that
acclimation of yeast and bacterial culture is completed. The yeast sludge that was fed with
glucose wastewater was milky white and became dark brown to black often acclimation to
protein wastewater. Microscopic observation detailed the changes by predominant yeast
68
strains when substrate was changed. The yeast sludge fed with glucose mainly contained
spherical yeast cells with multilateral or bipolar budding, whereas mycelia (hypha filament)
and large size egg-shaped cells with monopolar budding were predominant in the yeast
mixture fed with protein wastewater (Fig. 4.7).
a. Predominance of egg-shaped cells with monopolar budding in suspension
b. Predominance of mycelial yeast (hypha filament) in settled sludge
Figure 4.7 Predominance of wild yeast strains in the cultures fed with fish-protein
wastewater (at 32 g/L salt)
Similarly, the bacterial sludge changed in color and settling properties. Its color changed
from light brown or orange to brown or dark brown when glucose-feed wastewater was
altered with protein-feed wastewater. Its flocs were larger and easily trapped fine particles
during settling. Therefore, supernatants of bacterial batches fed with protein wastewater were
clearer than those from glucose wastewater after 30 minutes of settling. SS of supernatant and
SVI in the batch fed with glucose wastewater (at 32g/L salt) were 275 mg/L and 16 mL/g,
respectively, while SS and SVI in the protein-feed wastewater were 230 mg/L and 81 mL/g.
The result revealed that there is a large difference in SVI but not in supernatant SS. Thus, the
bacterial sludge fed with protein wastewater may have poor thickening or dewatering ability;
details are discussed in Section 4.4. These can be explained by proteins in the wastewater
enhancing formation of extracellular polymers (ECP) which have a great influence on
bacterial floc structure, settling and dewatering ability (Dignac et al. 1998).
20 m
x 500
20 m x 800
69
Organic removals
All COD profile data of yeast and bacterial cultures fed with protein wastewater were
presented in Appendix B. Table 4.2 summarizes operating conditions and COD removal
efficiency of yeast and bacteria acclimatized to protein-feed wastewater at high salt contents.
Here, optimum HRT refers to aeration time at which COD removal efficiency obtained was
higher 80%.
Table 4.2 Performance of mixed yeast and bacterial sludges adapted to protein-feed
wastewater with high salt contents (Initial COD cof 5,000 mg/L).
Yeast Bacteria
Salt content Salt content Parameters Unit
20 g/L 32 g/L 45 g/L 20 g/L 32 g/L 45 g/L
Mean MLSS mg/L 6050 5810 6430 4300 3720 4120
HRT h 31 34 37 9 21 28
F/M g/g.d 0.64 0.61 0.51 0.64 0.31 0.21
CODeff mg/L 790 830 950 40 50 90
%COD % 84.2 83.4 81 96 95 91
COD removal rate (U): g /g.d
+ Protein ww 0.54 0.51 0.41 0.61 0.29 0.19
+ Glucose ww 2.65 1.32 0.90 3.20 0.81 0.41
The COD removal rate of both acclimatized yeast and bacterial sludges were
significantly reduced when glucose-feed wastewater was added protein-feed wastewater as
shown in Table 4.2. However, the COD removal rates of mixed yeast sludge at higher salt (32
and 45 g/L NaCl) were still higher than those of bacterial sludge when protein wastewater was
used. The difference in U of yeast sludge with salt increases was relatively minor. In addition,
in order to obtain equivalent COD removal efficiency, lower F/M ratios were required at
higher salt contents for both yeast and bacterial sludges. However, F/M reduction for the
mixed yeast sludge was small. This enhanced the advantages of the yeast system for seafood
processing wastewater having high organic strength and high salinity.
4.1.2 Evaluation and Comparison of Biokinetic Coefficients
Specific growth rates ( ) were obtained through OUR measurement by respirometric
method. The values of different initial CODs (20 to 500 mg/L) at 20, 32 and 45 g salt/L
were determined. Observed maximum specific growth rate ( obs) and the half-velocity
constant (KS) were determined from regression analysis.
Figures 4.8 and 4.9 show typical OUR curves of yeast and bacterial sludges fed with
glucose and protein wastewater at 32 g salt/L. OUR curves for a specific COD concentration
can represent maximum oxygen uptake rate of microorganisms for a given substrate, knowing
its biodegradability. For example, Fig. 4.8 indicates that for glucose-feed wastewater, OURs
of the mixed yeast sludge was higher than that of the bacterial sludge. Thus, yeast culture was
able to degrade glucose more efficiently at 50 mg/L COD and high salinity (32 g/L salt),
while the degradation ability of the mixed bacterial culture fed with protein-feed wastewater
was better at COD lower than 100 mg/L. Discussion is presented in Section 4.3.2.a (Fig.
4.32). The fraction of readily biodegradable matter is represented by the area A in these
figures. By comparison between the area A of the two substrates, it is noted that glucose-feed
wastewater mainly contains readily biodegradable matter, whereas the fraction of readily
biodegradable matter was lower for protein-feed wastewater.
70
A
A
0
10
20
30
40
50
OU
R (m
g O
2/m
g V
SS.h
)
0 5 10 15 20 25 30
Time (minutes)
bacterial sludge
yeast sludge
Figure 4.8 OUR curves of mixed yeast and bacterial sludges feed with 50 mg/L COD and
32 g/L salt (glucose-feed wastewater)
A
A
B
B
0
5
10
15
20
25
30
OU
R (m
g O
2/m
g V
SS.h
)
0 20 40 60 80 100 120
Time (days)
bacterial sludge
yeast sludge
Figure 4.9 OUR curves of mixed yeast and bacterial sludges feed with 100 mg/L COD and
32 g/L salt (protein-feed wastewater)
A given microorganism will survive in the system if it is able to reproduce at a faster
rate than the rate at which it was removed from the system. The mechanisms to remove the
microorganism can involve predation or wash-out in the effluent. Therefore the growth rate is
important in the biological treatment process. It is normally used to estimate the effects of
toxic substances, inhibitors or overloads on performance of the process. Figures 4.10 and 4.11
show the variation in specific growth rate with COD concentration (glucose-feed wastewater)
at different salt contents, for yeast and bacterial cultures, respectively. It can be seen that the
specific growth rate progresses in according to the Monod’s model to reach a maximum
value, then this value decreases as the salt content increases. Table 4.3 summarizes values of
obs, KS, and Y of yeast and bacterial sludges at different salt contents for glucose and protein-
feed wastewater.
71
0 100 200 300 400 500
COD (mg/L)
0.0
1.0
2.0
3.0
4.0
5.0
Specific
Gro
wth
Rate
day-1
20 g NaCl/L
32 g NaCl/L
45 g NaCl/L
20 g/L NaCl:
S
S
15860.5
32 g/L NaCl:
S
S
11874.4
45 g/L NaCl:
S
S
12970.2
R2 = 0.971 R2 = 0.975 R2 = 0.967
Figure 4.10 Variation in specific growth rate of yeast sludge as function of COD
concentration at different salt contents for glucose-feed wastewater
0 40 80 120 160 200
COD (mg/L)
0.0
2.0
4.0
6.0
8.0
Specific
Gro
wth
Rate
day-1
20 g NaCl/L
32 g NaCl/L
45 g NaCl/L
20 g/L NaCl:
S
S
4595.9
32 g/L NaCl:
S
S
5580.2
45 g/L NaCl:
S
S
5315.1
R2 = 0.965 R2 = 0.974 R2 = 0.950
Figure 4.11 Variation in specific growth rate of bacterial culture as function of COD
concentration at different salt contents for glucose-feed wastewater
For both substrates, the specific growth rate of mixed yeast culture is higher than the
bacterial one at high salt contents, while the opposite is observed for lower salt contents. This
observation is in line with the nature of the COD removal rate (Fig. 4.5). It was also found
that irrespective of the salt content, the yield constant Y of mixed yeast batch fed with glucose
is lower than that of the bacterial culture, whereas there was no considerable difference
between the Y constants of yeast and bacterial systems fed with protein at any salt content.
This may be due to change in the predominant species or changes in the carbon assimilation
metabolism as substrates change. The yield constants for both yeasts and bacteria grown on
72
protein-feed wastewater were slightly lower than for glucose-feed wastewater. The yield
constant estimated by the respirometric method is considered as the maximum yield
coefficent (Ymax) under certain environmental conditions such as temperature and type of
wastewater. In a biological treatment system, the observed yield constant Yobs may vary from
0 and upto Ymax. The observed yield constant Yobs values depend on design of the system such
as F/M ratio and sludge age (Henze et al., 1997). Therefore, unlike specific growth rate, Y
constant should not be used to evaluate the effects of inhibitors or to compare performance of
two systems with different operating conditions.
Table 4.3 Biokinetic coefficients of the yeast and bacterial sludges at different salt contents
for glucose and protein-feed wastewaters
obs Y KS
day-1
g VSS/g COD mg COD/L
Sg /L
NaCl Substrate
Yeasts Bacteria Yeasts Bacteria Yeasts Bacteria Glucose feed wastewater
20 5.60 9.95 0.46 0.57 158 45
32 4.74 2.80 0.48 0.58 118 55
45 2.70 1.15 0.41 0.53 130 53 Protein feed wastewater
15 4.69 5.65 0.43 0.40 201 54
32 3.66 1.95 0.44 0.47 329 93
45 2.46 1.11 0.40 0.40 396 96 Candida ingens in culture
using VFAs (Anciaux et
al., 1989)
6.8-7.2 0.56 N/A
C.utilis in culture using
acetic acid
Jackson and Edward, 1975
7.2-9.6 0.38 N/A
<1.0Domestic wastewater
(Henze et al., 1997)N/A 4-8 N/A 0.35-0.50 N/A 5-30
Ks values of bacterial cultures (both substrates) at high salt contents were found to be
higher than Ks of normal activated sludge. This indicates that heterotrophic aerobic
microorganisms living at high salinity show lower affinity for the substrate than those at low
salinity, probably because of reduced functioning and multiplication. Whereas Ks values of
yeasts were found to be 3-6 times higher than those of bacteria. Thus maximum growth rate of
yeast culture is only obtained at high organic substrate concentration.
Figures 4.12 and 4.13 compare maximum specific growth rate of mixed yeast and
bacterial cultures with glucose and protein-feed wastewaters at different salt contents. At
higher salt contents, the bacterial growth is severely inhibited, while the growth rate of yeast
mixture is sustained. The inhibition effect of high salt contents on yeasts and bacteria based
on the Ghose and Tyagi model is also shown. Salt inhibition constants (KI) have been
calculated from the linear relationship. It was 70 g/L (80g/L) for yeast and 46 g/L (51g/L) for
the bacterial culture with glucose (protein). This indicates that the inhibitory salt content is
much lower for bacterial culture compared to the yeast which is in line with the previous
observations. However, actual critical salt limits may be higher than the values derived, found
from the Ghose and Tyagi model. This can be recognized from studies concerning
osmotolerant microorganisms by Choi and Park (1999). They reported that growth of Pichia
guilliermondii on Kimchi brine waste is still sustained at up to 120 g NaCl/L. Similarly,
Dincher and Kargi (2000) also reported that Halobacter sp. in activated sludge culture could
73
continue removing COD even at 50 g salt/L. Therefore, the limits obtained in this study may
only be used to compare the relative performance.
701*1.8
I
R2 = 0.870
461*9.15
I
R2 = 0.959
0 10 20 30 40 50 60 70 80
Salt (g/L NaCl)
0.0
4.0
8.0
12.0
16.0
Observ
ed S
pecific
Gro
wth
Rate
day
-1)
Bacterial sludge
Yeast sludge
Figure 4.12 Inhibition effect of salt contents on mixed yeast and bacterial cultures on
glucose-feed wastewater
0 10 20 30 40 50 60 70 80
Salt (g/L NaCl)
0.0
4.0
8.0
12.0
16.0
Observ
ed S
pecific
Gro
wth
Rate
day-1
)
Bacterial sludge
Yeast sludge
491*65.7
I
R2 = 0.928
R2 = 0.985
801*86.5
I
Figure 4.13 Inhibition effect of salt contents on mixed yeast and bacterial cultures on
protein-feed wastewater
4.2 Parametric Study
This study focused on several operating parameters such as pH, DO, SRT and nitrogen
content for the acclimatized mixed yeast and bacterial cultures. In fact, the effect of pH on
bacterial and yeast systems were evaluated in terms of OUR by respirometric assays. The
optimum pH values for bacterial and yeast growth under high salinity conditions were
evaluated. Variation in DO and nutrient components in the yeast and bacterial batches was
monitored for the two feed wastewaters (glucose and fish-protein) with 32 g/L salt. The SRT
variation experiments were conducted in mixed yeast cultures using the fill-and-draw
operation. Based on nutrient and COD removals, a suitable SRT value for yeast treatment was
determined.
74
4.2.1 DO and pH
a. DO
Figures 4.14 and 4.15 compare DO profiles of mixed yeast and bacterial cultures fed
with glucose and protein wastewaters at 32 g salt/L. These figures also show that there is a
link between DO of mixed liquor and remaining COD (COD profile). DO of the mixed-liquor
in the yeast batch fed with glucose was low (0.7 mg/L) in the first 5 hours. DO then sharply
increased to saturated value (6.3 mg/L), whereas low DO level of 0.7 mg/L in the batch fed
with protein was steady for longer duration (28 h of aeration). DO then was also sharply
raised to 5.6 mg/L. DO values were lower for both substrates during the first hours of
aeration. This can be attributed to the presence of high initial concentration of organics (5,000
mg/L COD) after filling. The oxygen uptake rate of yeasts exceeded oxygen diffusion rate
from aeration at the first hours. Based on COD profiles, DO increased to saturated value when
most of the organic matters in the system was completely degraded.
0.0
2.0
4.0
6.0
DO
(m
g/L)
0
1000
2000
3000
4000
5000
CO
D (
mg/L
)
0 10 20 30 40
Time (h)
DO-protein-yeast
DO-glucose-yeast
COD-protein-yeast
COD-glucose-yeast
Figure 4.14 DO and COD changes of yeast batch fed with glucose and protein wastewater at
32 g salt/L
Similarly, DO in mixed bacterial cultures for both glucose and protein wastewater was
low during the first hour of aeration. It then increased sharply to the saturated concentration of
6.3 mg/L. This suggests that COD loading (influent COD of 1000 mg/L) applied to bacterial
batches was not as high as for the yeast batch. Oxygen consumption rate of bacterial sludge
did not exceed oxygen supply rate through aeration.
75
0.0
2.0
4.0
6.0
DO
(m
g/L)
0
200
400
600
800
1000
CO
D (
mg/L
)
0 5 10 15 20 25 30
Time (h)
COD-protein-bacteria
DO-protein-bacteria
COD-glucose-bacteria
DO-glucose-bacteria
Figure 4.15 DO and COD changes of mixed bacterial batch fed with glucose and protein
wastewater at salt content of 32 g/L
b. pH
A comparison of pH profiles of mixed yeast and bacterial batches fed with glucose and
protein wastewater (at 32 g/L salt) is presented in Figures 4.16 and 4.17. In the yeast batch fed
with protein, pH increased from 3.5 to above 4.0 after 2 hours of aeration. In order to maintain
pH at 3.5, 0.1N H2SO4 was added after every two or three hours. Inversely, pH of yeast fed
with glucose decreased pH to 2.6 after 8 h of aeration. Then it slightly increased to 2.9. Based
on the COD profile of the yeast batch fed with glucose-feed wastewater (Fig. 4.14), pH
started to raise when carbon supply was limited (about 250 mg/L COD).
Adjust pH to 3.5
2.0
3.0
4.0
5.0
pH
0 10 20 30 40
Time (h)
yeast-protein
yeast-glucose
Figure 4.16 pH changes of yeast culturefed with glucose and protein wastewater at 32 g
salt/L
Thus, there is a significant difference in pH variation between glucose and protein-feed
wastewaters. This may be attributed to difference in substrates involving carbon and nitrogen
sources. The main components of glucose wastewater are the glucose as carbon source and
inorganic nitrogen (ammonium sulphate) as nitrogen source, while organic acids or complex
organics such as lipids and polysaccharides may be predominant carbon sources in protein-
76
feed wastewater. The amount of organic nitrogen (such as protein, amino acids) is very high
in the protein-feed wastewater (690 mg /L organic-N; 26 mg/L ammonia-N).
The increase in pH for the protein wastewater has been previously observed by Lu
(1983) who used yeast mixture to treat vermicelli wastewater. pH increased from 4.0 to 8.5.
Arnold et al. (2000) examined silage wastewater treatment using Candida utilis and
filamentous yeast Galactomycetes geotrichum. The initial pH was 3.65, rising to 8.8 after
treatment. The authors postulated that the increase in pH was due to lactic acids removal,
VFAs or consumption of H+ during oxidation of organic N into ammonia. Lu (1983) also
suggested that the degradation of protein and release of ammonia caused increase of pH.
By contrasts, the decrease in pH for the yeast fed with glucose wastewater can be due to
the production of organic acids or the use of basic compounds such as ammonia by the cells
or the absorption by the medium of CO2 produced by yeasts (Thanh and Simard, 1973).
6.0
6.5
7.0
7.5
8.0
8.5
9.0
pH
0 5 10 15 20 25 30
Time (h)
Bacteria-glucose
Bacteria-protein
Figure 4.17 pH changes of mixed bacterial batch fed with glucose and protein wastewaters at
32 g salt /L
Unlike yeast growth, variation in pH in the mixed bacterial batch was not much
dependent on type of substrates. Both wastewaters had same initial pH of 7.5 and after
treatment, pH of both batches were increased. The difference in pH variation between yeast
and bacteria culture on glucose substrate may be explained by the difference in pathways of
nitrogen assimilation under respiratory conditions with no contribution by carbon metabolism
(Vicente, et al. 1998). Yeasts liberate H+ ion during ammonia uptake. In general, bacteria
produce alkalinity during ammonification and consume alkalinity in nitrification. However at
high salinity, nitrification is inhibited. Nitrite and nitrate concentrations of treated wastewater
in bacterial batches were lower than 2.5 mg/L N at 32 and 45 g/L salt. Fig. 4.17 shows that pH
of the mixed bacteria fed with protein wastewater (8.5) was higher than that fed with glucose
(7.8). This might be due to the increase of alkalinity during conversion of protein to ammonia.
c. Evaluation of optimum pH
Optimum pH of mixed yeast and bacterial cultures were evaluated in terms of OUR
using respirometric experiments. These were conducted with the protein-feed wastewater at
50 mg/L COD and 32 g salt /L. The bacterial and yeast sludges used in these experiments
were from the biomembrane reactors operated with the protein-feed wastewater at 32 g salt/L.
77
Figures 4.18 and 4.19 show suitable range of pH values for yeast and bacteria growth in high
salinity.
2.0 4.0 6.0 8.0 10.0
pH
0
2
4
6
8
10
12
14
16
18
OU
Rm
ax (m
gO
2/gVSS.h
)
Total OUR
Endogenous OUR
Figure 4.18 Variation in OUR as funtion of initial pHs for mixed yeast fed with protein
wastewater at 32 g salt/L
4.0 5.0 6.0 7.0 8.0 9.0 10.0 11.0
pH
0
5
10
15
20
25
OU
Rm
ax (m
gO
2/gVSS.h
)
Total OUR
Oxidation OUR
Figure 4.19 Variation in OUR as function of initial pHs for mixed bacterial fed with glucose
wastewater at 32 g salt/L
OUR obtained in the mixed yeast culture was highest at pH values of 5.0 - 5.5 and
declined slightly as pH increased to 8.0 or decreased to 3.0. The respiration rate of yeasts was
inhibited at pH 2.5 and above 9.0, whereas bacterial culture attained the highest OUR at pH
range of 7.5-9.0 where OUR declined slightly when pH was increased to 9.7, or decreased to
6.3. The respiration rate of bacteria was inhibited at pH below 5.3 or pH above 10.0. Thus, the
osmotolerant yeasts were able to tolerate a wider pH range than bacterial culture. Choi and
Park (1999) obtained similar results for Pichia guilliermondii, an osmotolerant yeast used to
treat kim chi waste brine with 80 g/L salt. Cell growth was not affected at pH ranging 4.0 to
8.0. By scanning electron micrographs, they showed that yeast cells only shrunk in size, but
did not rupture at the high osmotic shock pressure or high ion strength. This appears to be a
typical advantage of using yeasts to treat industrial effluents having large pH fluctuation.
However, at neutral pH (6.5 - 8.5), bacteria do multiply in significant numbers. Bacterial
growth should be inhibited by pH control in the yeast treatment process for the following
reasons:
78
a. If the yeast biomass is used as animal feed, large numbers of bacteria or pathogens will
reduce the quality of the yeast biomass product;
b. Excessive bacterial growth leads to operating problems, especially membrane clogging.
Previous studies have reported that bacterial contamination could be inhibited in pH
range of 3.5 - 3.8 (Peczar and Reid, 1972; Elmaleh, et al., 1966). The results of this study
shows that there was not a significant difference between OUR at 3.5 and OUR at the
optimum pH (5.0 - 5.5). The difference was 9%. Therefore, maintenance of pH 3.5 in the
reactor cannot reduce considerably COD removal rate.
4.2.2 Nitrogen Variation in Mixed Yeast and Bacterial Cultures
Figure 4.20 shows the variation in nitrogen components in the mixed yeast cultures fed
with glucose and protein wastewater. In the yeast fed with glucose wastewater, nitrogen
source was an inorganic salt (NH4)2SO4 with initial concentration of 365 mg N/L
(corresponding to COD:N = 100:7.2). Ammonia-nitrogen removal of 65% was obtained after
eight hours. Total nitrite and nitrate-N concentration of treated wastewater was very low
(around 2.0 mg N/L). This indicated that nitrification did not occur in the yeast reactor.
Likewise, total nitrite and nitrate concentration was not important in the yeast batch fed with
protein wastewater. The initial total nitrogen of the protein-feed wastewater was 790 mg N/L
mostly organic-N (745 mg/L organic-N). The total-N concentration was reduced to 446 mg/L
after 32 h, whereas ammonia-N increased from 45 mg/L to 420 mg/L. In comparison to the
yeast culture with glucose-feed wastewater, ammonia content and ammonia removal was
dependent to the availability of nitrogen sources (e.g. proteins, amino acids or ammonium
salt) in the feed wastewater and the BOD:N ratio. Due to the high BOD:N ratio (100:18) of
the protein-feed wastewater, the total nitrogen concentration of treated wastewater was still
high and mainly in the form of ammonia. If reuse of yeast biomass for single-cell-protein
production vis-à-vis further removal of nitrogen is considered, the combination with
carbohydrate-rich wastes (lack of nitrogen) such as molasses or pulp and paper wastewater
would be possible.
Moreover, the increase in ammonia may be related to the increase in the accumulated
acid volume used for adjusting pH to 3.5 as shown in Fig. 4.21. The link between H+ ion
consumption and ammonia release was discussed in Section 4.2. The amount of acid
consumed per gram ammonia-N released was about 26 meq H+.
79
0 10 20 30 40
Time (hours)
0
200
400
600
800
N (
mg /
L N
)
6
8
10
12
14
16
18
Acc
um
ula
ted 0
.1N
H2SO
4 v
olu
me (
mL)
Total N (protein ww)
NH4-N (protein ww)
Accumulated acid volume
NH4-N (glucose ww)
Figure 4.20 Variation in nitrogen components as funtion of time in the mixed yeast at 32 g
salt/L NaCl (Nitrite and nitrate concentration of both feed wastewaters were not
dectected)
The total-N removal obtained in the batch fed with protein wastewater after 32 h was
45%. Because nitrification did not occur in the yeast treatment, all nitrogen removed was
uptaken in the yeast sludge. Based on biomass produced in the batches and the amount of
nitrogen removed, the nitrogen content of the yeast sludge was estimated to be about 7.5% of
dried solids for glucose wastewater, and 13.8% for protein wastewater. The high nitrogen
content of the yeast sludge fed with protein wastewater can be attributed to precipitation of
protein at low pH (3.5). This can be confirmed through the sudden decrease in total nitrogen
concentration during the first hour of aeration (Fig. 4.20). The nitrogen uptake ability of
yeasts will be further discussed in Section 4.4.
0 10 20 30
Time (hours)
0
40
80
120
160
200
Nitro
gen (
mg /
L N
)
Total-N (protein ww)
Ammonia (protein ww)
NO2+NO3 (protein ww)
Ammonia (glucose ww)
NO2+NO3 (glucose ww) < 1.5 mg/L N
Figure 4.21 Variation in nitrogen components vs. time in the mixed bacterial culture at 32 g
salt/L NaCl
In the bacterial culture fed with protein wastewater, 24% of total nitrogen was removed
at HRT of 21 hrs. Fig. 4.21 shows that the oxidation of organic-N (ammonification) results in
an increase from 49 to 137 mg N/L after 21 hrs, while in the culture fed with glucose
wastewater, the initial ammonia concentration of 53 mg N/L was reduced to 14 mg/L (about
72% N removal) after 9 hrs. Thus the variation of ammonia during the bacterial cultures was
similar to that of mixed yeast cultures. Nitrite+nitrate-N concentration was lower than 2.5
80
mg/L and 1.2 mg/L for the cultures fed with protein and glucose, respectively. Whereas nitrite
and nitrate-N (fed with protein) at 15 g salt/L was 12.3 mg/L after 21 hours (Appendix A).
Thus nitrification was inhibited at 32 g salt/L. Lower reduction at higher salt (32g/L) is
consistent with previous studies (Dincer and Kargi, 1999; Panswad and Anan, 1999). In fact,
Dincer and Kargi (1999) reported that nitrification efficiency dropped quite sharply at salt
content above 3%. Depending on nitrogen balance, nitrogen uptake in the mixed bacterial
sludge for both feed wastewaters was reached to 4.5 % of dry solids. This result will be
confirmed by nutrient analysis of sludge in the sludge characterization study.
4.2.3 Effect of SRT on COD and Nitrogen Removal
The optimum SRT was evaluated on the basis of COD and nitrogen removal. Five
mixed yeast batch experiments corresponding to SRT of 5, 7, 10, 20 and 45 days were
conducted at the same organic loading of 5 kg COD/m3.d (HRT of 24 h) for 25 days. The seed
sludge used was taken from the yeast membrane bioreactor (YMBR) operated with SRT of 50
days. As the degradation of protein and the release of ammonia caused an increase in pH
during the culture, pH was adjusted to 3.5 - 4.0 using 0.1 N H2SO4 during aeration. Figure
4.22 shows that all the runs reached steady state after 15 days.
The mixed yeast culture run at higher SRT reached higher biomass concentration. At the
VLR of 5.0 kg COD/m3.d, when SRT increased from 5 d to 45 d, the MLSS increased from
2,400 to 10,300 mg/L. Thus a long SRT implies low F/M ratio. This resulted in high organic
removal efficiency at long SRT. The results are shown in Table 4.4 and Fig 4.23. COD
removal increased from 43 to 85% when SRT increased from 5 d to 45 d. Furthermore, the
sludge production (yield constant) was also minimized at long SRT (Table 4.4). Therefore,
SRT has a mutual relationship with the net specific growth rate of the mixed yeast sludge. The
net specific growth rate (’) was the difference between the specific growth rate ( ) and the
endogenous decay rate (kd), which includes endogenous respiration, death and subsequent
lysis. At long SRT, the system is operated in the endogenous phase. Thus kd will have a
significant effect on the net amount of biomass produced. This means that large fraction of the
substrate removal is oxidized for energy required for cell maintenance rather than for
synthesis of new cells.
0 5 10 15 20 25
Time (days)
2000
4000
6000
8000
10000
12000
MLSS (
mg/L)
45 d
20 d
10 d
7 d
5 d
Figure 4.22 Variation in MLSS as funtion of SRT
81
Table 4.4 Variation of parameters during various SRTs (Initial COD of 5000 mg/L)
SRT
(d)
CODeff
(mg/L)
TKNeff
(mg/L)
COD
removal
(%)
N
removal(*)
(%)
Mean
MLSS
(mg/L)
Biomass
production
rate(mg SS
produced/d)
Yield(g SS/g COD
removed)
5 2850 524 43 7.3 3245 657 0.305
7 1950 515 61 8.9 5351 765 0.251
10 1050 509 79 9.8 8150 810 0.205
20 900 535 82 5.2 9455 475 0.116
45 950 550 85 2.7 10335 228 0.053(*) Influent TKN concentration was 565 mg/L N.
The data are average values of at least three steady state batches.
Figure 4.23 indicates that if the highest COD removal was obtained at SRT of 45 d,
maximum nitrogen removal was achieved at SRT of 10 d. Uptake of nitrogen by the biomass
is a major nitrogen reduction mechanism in yeast culture as discussed in Section 4.2.2. Hence,
the nitrogen removal efficiency will depend on the biomass production rate. Table 4.4 shows
that the highest biomass production was obtained at SRT of 10 d. It can be suggested that
selection of optimum SRT should be based on the purpose of yeast application. SRT of 10 d is
optimum for single-cell-protein production, while longer SRT is suitable for enhancing
treatment efficiency.
10 20 30 40 50987654
SRT (days)
0
20
40
60
80
100
Rem
oval eff
icie
ncy
)
2000
4000
6000
8000
10000
12000
14000
MLSS (
mg/L
)
COD removal
Nitrogen removal
MLSS
Figure 4.23 Variation in COD, nitrogen removal and MLSS in funtion of SRT in mixed
yeast culture at VLR of 5 kg COD/m3.d (32 g salt/L)
4.3 Biomembrane Study
The objective of this study was to examine the potential for development of membrane
bioreactor systems using wild salt-tolerant yeast mixture and bacteria mixture to treat high
salinity wastewater (32 g/L NaCl). Based on COD of the feed-wastewater, this study was
divided into two phases, namely (1) high COD loading with 5000 mg COD/L and (2) low
COD loading with 1000 mg COD/L. The process efficiency was investigated in terms of
organic removal and membrane filtration flux for various volumetric loading rates, F/M ratio
and SRT values.
82
Two membrane modules having 0.1 m pore size and 0.42 m2 area were used for
YMBR and BMBR. When the pressure reached a value of 70 kPa, the membrane was
removed from the reactor, and chemical cleaning was conducted. During the chemical
cleaning, the external sludge cake layer was initially washed with water and the attached
biomass (on the membrane surface) was collected and analysed. After removal of the sludge
cake, the membrane was washed with tap water and backwashed with 2.5% sodium hydroxide
for 15 minutes followed by 1% nitric acid for 5 minutes before reuse.
After every chemical cleaning, the initial membrane resistance was measured, to verify
the cleaning efficiency. The initial Rm were determined by filtering the tap water through a
new or chemically cleaned membrane. Here, linear flux variation with applied pressure was
obtained. This variation for two fresh membrane modules is presented in Fig. 4.24. The
obtained initial Rm of two cleaned membrane modules were quasi equivalent (7.11x1011
and
7.18x1011
m-1
).
0 5 10 15 20 25 30 35
J (L/m2.h)
0
1
2
3
4
5
P (
kPa)
Y = 0.174* X - 0.1472
0 5 10 15 20 25 30 35
J (L/m2.h)
0
1
2
3
4
5
P (kPa)
Y = 0.172 * X - 0.1372
a. First membrane b. Second membrane
Figure 4.24 Variation in flux as function of membrane transmembrane pressure (Viscosity of
water at 26oC = 8.70 x 10
-4 kg/m.sec)
4.3.1 High COD loading
In this phase, fish-protein wastewater with 5,000 mg COD/L and 32 g salt/L was used.
Two experimental set-ups were investigated: (1) Yeast pretreatment followed by BMBR and
(2) YMBR. These were run at different HRTs at SRT of 50 d (as presented in Fig. 3.5).
a. Organic removal
The performance of YMBM and BMBR systems for various volumetric loading are
shown in Figures 4.25 and 4.26. Here it can be note that after 11 days of acclimation (stage I),
yeast biomass increased from 3,700 to 14,500 mg/L at a volumetric loading of 5.0 kg
COD/m3.d (average HRT of 24 h). COD and BOD removal obtained were above 76% and
85%, respectively.
In contrast, the BMBR system which involved the yeast reactor (YR) and BMBR
reached the steady state after 11 days. The YR reached the mean biomass of 6,500 mg/L and
COD removal of 76% at SRT of 15 days and average HRT of 36 h. The bacterial biomass in
BMBR increased from 4,000 to 11,000 mg/L at a volumetric loading of 2.1 kg COD/m3.d
(HRT of 13.7 h). COD and BOD removal obtained was above 85% and 97%, respectively.
These results demonstrate the rapid adaptability of the mixed yeast and bacterial cultures to
degrade the high salinity-organic wastewater.
83
20
40
60
80
Tra
nsm
em
bra
ne (
kPa)
10
20
30
40
Mean H
RT (
h)
Mean HRT
Transmembrane Pressure
0
4000
8000
12000
16000
20000
MLSS (
mg/L)
2000
4000
6000
CO
D (
mg/L)
0 10 20 30 40 50 60 70 80
Time (days)
Effluent COD
Influent COD
MLSS
Stage I Stage II Stage III Stage IV Stage V Stage VI
Figure 4.25 Variation in COD, biomass and transmembrane pressure in the YMBR as
function of volumetric loading
In the YMBR, as the loading rate was progressively increased through different stages
(3.4 - 16.3 kg/m3.d), the COD removal efficiency decreased from 85 to 60%, with the COD in
the effluent increasing from 870 to 2,300 mg/L. For the BMBR process, when the VLR was
increased from 2.1 to 7.9 kg COD/m3.d (F/M of 0.08 - 0.41), COD removal efficiency
decreased from 91 to 76 % (Fig. 4.7). Effluent BOD5 of the BMBR ranged from 45 - 60 mg/L
at low F/M ratio. Thus, the low BOD5:COD ratio (0.12 - 0.17) indicates that BMBR effluent
also contains a high proportion of non-degradable organic compounds due to the presence of
these products from yeast pretreatment.
YMBR could attain a COD removal efficiency higher than 60% at VLR ranging from 5
- 15 kg COD/m3.d as shown in Figure 4.27. These VLRs are generally within the acceptable
range for anaerobic treatment systems, which normally have low efficiency in the range of 60
- 70%. Moreover, it has also been found (Feijoo et al. 1995) that anaerobic microorganisms
are highly sensitive to salt content and total inhibition could be noted for some treatment
systems at a salt content above 20 g/L. Thus the yeast system is more useful than an anaerobic
system at a high salt content, and can be considered a better substitute for anaerobic systems
in terms of COD removal efficiency.
In addition, YMBR attained a lower COD removal rate at F/M ratios lower than 0.34
g/g.d (the corresponding average VLRs less than 7 kg/m3.d), compared to BMBR as shown in
Fig.4.28. However, YMBR achieved higher specific COD removal rate at F/M ratios higher
than 0.34 g/g.d. Thus, it can be concluded that the YMBR is subjected to a higher F/M ratio
and higher VLR than the BMBR to obtain a comparable COD removal efficiency. This can be
considered as an advantage for the yeast sludge compared to bacterial sludge.
84
20
40
60
80
Tra
nsm
em
bra
ne P
ressure
(kPa)
5
10
15
20
Mean H
RT (
h)
Mean HRT
Pressure
0
10000
20000
30000
MLSS (
mg/L)
0
400
800
1200
1600
CO
D (
mg/L
)
0 10 20 30 40 50 60 70 80 90
Time (days)
Effluent COD
Influent COD
MLSS
Stage I Stage II
Stage III
Stage IV Stage V Stage VI
Stage VII
Figure 4.26 Variation in COD, biomass and transmembrane pressure in the BMBR as
function of volumetric loading
The mean biomass concentration of the yeast reactor was 6,500 mg/L after steady state.
77% of BOD5 removal was achieved at VLR of 2.6 - 3.1 kg BOD5/m3.d and F/M ratio of 0.39
- 0.47 d-1
, while the Yeast Cycle System (YCS) for seafood processing wastewater treatment
(with 8 g salt/L) reached 97% BOD removal at higher VLR (4.5-10.4 kg BOD5/m3.d) and
higher F/M ratio (0.6 - 1.0 d-1
) (Nishihara ESRC Ltd., 2001). This may be due to yeast ability
to grow in relative low salinity environment with higher specific degradation rate. Similar
results were found in the biokenetic study (Section 4.1). In comparison to the yeast reactor,
the YMBR that was run at higher biomass concentration (11,000 mg SS/L) enables the
enhancement of volumetric loading to 5.0 kg BOD5/m3.d with higher BOD removal efficiency
(81%).
Table 4.5 compares operating conditions between YMBR, BMBR and few yeast
treatments, MBR process treating different wastewaters. Krauth and Staab (1993) found that
using BMBR for treatment of vegetable canning wastewater could achieve COD removal
efficiency exceeding 99% at high F/M ratio (0.5 g/g.d) and high VLR (5.4 kg/m3.d). BMBR
could also efficiently treat oily wastewater at mean F/M ratio of 0.7 g/g.d and VLR between
8.6 and 12.9 kg/m3.d (Scholz and Fuchs, 2000). Whereas the BMBR in this experiment can
only be subjected to lower VLR (3.4 - 5.0 kg/m3.d) and lower F/M ratio (0.1 - 0.3 g/g.d) to
obtain a comparable COD efficiency. It is important to note that all the above mentioned
BMBR systems were operated with salt contents lower than 1.0 g/L NaCl. However, this
current work was carried out at 32 g/L, the high salinity could be the major cause of the drop
in specific organic removal rate of bacterial sludge. Thus high salinity also reduces the
specific organic removal rate of bacterial sludge. However, compared to conventional
activated sludge systems for low salinity wastewater treatment (maximum VLR of 1.2
kg/m3.d), the BMBR could be operated efficiently at higher VLRs (3.4 - 6.0 kg/m
3.d).
85
Tab
le 4
.5 O
per
atin
g
par
amet
ers
of
the
YM
BR
, B
MB
R,
som
e y
east
tr
eatm
ents
, M
BR
pro
cess
es
trea
ting
dif
fere
nt
was
tew
ater
s an
d
conv
enti
onal
AS
sy
stem
Par
amet
er
Yea
st
Bac
teri
al m
emb
ran
e b
iore
acto
r p
roce
ss
Co
nv
enti
on
al A
S
Was
tew
ater
P
rote
in
extr
act
S
eafo
od
pro
cess
ing
V
erm
icel
li
Pro
tein
extr
act
Veg
etab
le
can
nin
g
Oil
y
was
tew
ater
Fer
men
tati
on
was
tew
ater
G
luco
se
Dom
esti
c
was
tew
ater
Pro
cess
M
BR
wit
h M
F
YC
S
(Co
nti
nu
ou
s
com
ple
ted
mix
ing
)
shak
ing
cult
ure
M
BR
wit
h M
F
MB
R w
ith
MF
MB
R w
ith
MF
M
BR
wit
h U
F
Co
nti
nu
ou
s
com
ple
ted
mix
ing
Com
ple
ted
mix
ing
Mic
roo
rgan
ism
s Y
east
mix
ture
Yea
st
mix
ture
Mix
ture
of
10
yea
st
stra
ins
AS
AS
A
S
AS
A
S
AS
VL
R, kgC
OD
/m3.d
4
.9-6
.5
4.5
– 1
0.4
(*)
1.0
3
3.4
-6.0
5
.4
8.6
-12
.9
1.7
3
- 0
.8-1
.9
F/M
rat
io, g
/g.d
0
.40
.6 –
1.0
(*
) 0
.5
0.1
-0.3
0
.5
0.6
-0.8
-
0.3
-0.6
(0.8
-2.1
) 0
.2-0
.6
ML
SS
, g
/L
15
8 –
10
2
.6
16
-20
1
1
15
-25
1
0
2
.5-4
.0
CO
D r
emoval
, %
8
6-9
1
97
9
2
91
-97
9
9
97
9
2
99
9
9
pH
3
.5-3
.8
4.3
–5
.2
3.0
-4.0
7
.5-8
.0
>6
.5
7.0
-7.8
6
.8-7
.2
6
.5-8
.5
DO
, m
g/L
1
.50
.5-0
.9
2
.16
.3
- 2
.0-3
.5
> 2
.0
2.0
Sal
t, g
/L
32
8
<1
3
2<
1
<1
<
1
30
(sal
t fr
ee)
<1
Ref
eren
ce
Th
is s
tud
y
Nis
hih
ara
ES
RC
,
Ltd
.(2
00
1)
Hu (
1989)
Th
is s
tud
y
Kra
uth
and
Sta
ab (
19
93
)
Sch
olz
and
Fuch
s (2
000
)
Lu e
t al
. (1
999
) H
amod
a an
d
Al-
Att
ar
(199
5)
Met
calf
and
Ed
dy
(19
91
)
Note
:(*
) :
kg B
OD
5/m
3.d
M
BR
:
Mem
bra
ne
Bio
reac
tor
M
F
: M
icro
filt
rati
on
U
F
: U
ltra
filt
rati
on
Y
CS
:
Yea
st C
ycl
e S
yst
em
A
S
: A
ctiv
ated
slu
dge
86
This can be explained by high sludge concentration and high substrate removal rate.
Similar results have been reported by Manem and Sanderson (1996), who found that VLR for
dairy wastewater was six times greater than for conventional activated sludge process,
although the biomass concentration was only twice as high. Moreover, the effluent suspended
solids of both membrane reactors were less than 5 mg/L and was almost constant throughout
all experiments.
0 2 4 6 8 10 12 14 16 18 20 22 24
Volumetric Loading (kg COD/m3.d)
20
30
40
50
60
70
80
90
100
CO
D rem
oval (%
)YMBR
BMBR
YMBR: Y = -0.044*X2-1.045*X+88.0
R2 = 0.929
BMBR: Y = -0.572*X2 + 3.568*X + 82.7
R2 = 0.836
Figure 4.27 Variation in COD removal in function of volumetric loading rate
YMBR: Y = -0.337*X2+ 1.156*X– 0.1041*X
R2 = 0.951
BMBR: Y = -1.285*X2+ 1.331*X - 0.0323
R2 = 0.938
0.0 0.5 1.0 1.5 2.0 2.5
F/M ratio (d-1)
0.0
0.2
0.4
0.6
0.8
1.0
CO
D rem
oval ra
te (
g C
OD
/ g
MLSS.d
)
BMBR
YMBR
Figure 4.28 Variation in COD removal rate in function of F/M ratio (initial COD = 5,000 mg/L)
b. Transmembrane pressure and membrane clogging
Variation in transmembrane pressure in YMBR and BMBR at different operating stages
is shown in Figures 4.25 and 4.26. The pressure ( P) in YMBR was almost constant
throughout the various stages of VLR for a total duration of 72 days. It then increased rapidly
after the 76th
day (63 kPa), indicating rapid membrane clogging. This may be due to the
increase in high filtration flux (89.6 L/d.m2), corresponding to HRT of 5 h in the last stage.
However, when YMBR was run at short HRT (5h), uncompleted biodegradation by yeast
results in high soluble COD and accumulation of fine particles (from the influent) retained in
the reactor which may cause a rapid fouling rate. The high soluble COD and fine particles in
the reactor could increase the filtration resistance (Manem and Sanderson, 1996). Whereas P
in BMBR sharply increased from 2 to 60 kPa after 12d, 6 d and 2 d at hydraulic retention time
87
(HRT) of 14h, 9 h and 4h, with average biomass concentrations of 6.1, 15 and 20 g MLSS/L
in stage I, II and III, respectively.
Values of different parameters during YMBR and BMBR filtration cycle in both phases
are presented in Table 4.7. As noted, increasing biomass concentration promotes the
membrane clogging, and difference between bacterial and yeast sludge results in different
filtration performances. In fact, characteristics of yeast mixture in the YMBR could prolong
the filtration cycle period. These characteristics are responsible for reducing membrane
clogging rate and can result in large yeast cells, low operating pH, poor adhesion capacity,
inhibiting biofilm formation, low net negative surface charge, low viscosity and low
production of the adhesive extracellular polymers (ECP) that play an important role in floc or
biofilm formation. These characteristics will be discussed in the next section (low COD
loading) and in the sludge characteristics study (Section 4.4).
4.3.2 Low COD loading
Fish-protein wastewater with 1,000 mg COD/L and 32 g salt/L was used in this phase.
Two experiments were conducted: (1) YMBR and (2) BMBR. The objective of this phase was
to obtain a comparative evaluation of treatment performance of YMBR and BMBR at
different HRTs and SRTs of 10 and 50 days (as presented in Fig. 3.6).
a. COD removals
Figures 4.29 and 4.30 present the overall performance of YMBR and BMBR process for
various HRTs in this phase. The mean influent COD concentration was maintained at 1,000
mg/L, and the VLR was increased from 2.7 - 6.5 kg COD/m3.d with a corresponding decrease
in HRT from 9 to 5 hours. The MLSS concentration for SRT of 10 days ranged from 4,500 -
5,100 mg/L and 5200 - 5500 mg/L for yeast and bacterial membrane reactors respectively. By
contrast, when the SRT was increased to 50 days, the MLSS concentration increased to within
the range of 13,600 - 15,200 mg/L in the YMBR and 15,200 - 16,300 mg/L in the BMBR. In
all experiments, the DO was maintained above 2.0 mg/L with salt content of 32 g/L. The
effect of changing VLR (vis-à-vis HRT) on COD removal at SRT of 10 days and 50 days is
shown in Fig. 4.31.
It was observed that at SRT of 10 days, the COD removal efficiency of the YMBR
remained low (about 76%) at lower HRTs (5h), but increased to 94% with the increase in
HRT (> 8 h). Whereas for the BMBR, the COD removal efficiency remained constant within
the range of 92 - 97% when HRT ranged from 5 to 8 h. Thus the COD removal efficiency of
the YMBR is lower than that of the BMBR at short HRTs, but converged at longer HRTs. In
general, the removal efficiency of a biological system increases with HRT (until a certain
limit), this was also observed for both systems.
88
20
40
60
Tra
nsm
em
bra
ne P
ressure
(kPa)
3
4
5
6
7
8
9
10
Mean H
RT (
h)
Mean HRT
Tran.. pressure
4000
8000
12000
16000
20000
MLSS (
mg/L)
0
200
400
600
800
1000
1200
CO
D (
mg/L
)
0 10 20 30 40 50 60 70 80 90
Time (days)
Effluent COD
Influent COD
MLSS
SRT = 10 d
SRT = 50 d
Figure 4.29 Variation in COD, biomass and transmembrane pressure in the YMBR as
function of volumetric loading
However, due to lower MLSS vis-à-vis the higher F/M ratio (1.02 g/g.d) in the YMBR
or possibly lower specific growth rate of yeast, the efficiency was low at lower HRT. Indeed,
the difference in the specific growth rates of yeast and bacteria at 32 g salt/L can be found in
the biokinetic study. Figure 4.32 reveals that although yeasts had higher maximum specific
growth rate ( max) at 32 g salt/L. Its specific growth rate ( ) at low substrate concentrations
(less than 180 mg/L) was lower than that of bacteria. Thus, the yeast growth was more
inhibited at low COD in the membrane reactor.
Higher HRT vis-à-vis low F/M ratio (0.5 d-1
) enabled better conversion of organic
matters with higher yeast mass available. The MLSS concentration in BMBR was relatively
high at SRT of 10 d. This might be due to higher specific growth rate of bacteria in a
substrate-limiting condition (COD < 200 mg/L) as shown in Fig. 4.9. Therefore, the COD
removal efficiency of BMBR remained high through out HRTs. A peak efficiency of 97%
was obtained for BMBR at a HRT of 7 - 8 hours, which represents the best range of operating
conditions.
89
20
40
60
80
100
Tra
nsm
em
bra
ne P
ressure
(KPa)
3
4
5
6
7
8
9
Mean H
RT (
h)
Mean HRT
Trans.Pressure
4000
8000
12000
16000
20000
MLSS (
mg/L)
0
400
800
1200
CO
D (
mg/L)
0 20 40 60 80
Time (days)
Effluent COD
Influent COD
MLSS
SRT = 10 d
SRT = 50 d
Figure 4.30 Variation in COD, biomass and transmembrane pressure in the BMBR as
function of volumetric loading
60
70
80
90
100
CO
D r
em
oval (%
)
YMBR
BMBR
4.0 5.0 6.0 7.0 8.0 9.0 10.0
HRT (h)
50
60
70
80
90
100
CO
D rem
oval (%
)
YMBR
BMBR
SRT of 50 d
SRT of 10 d
YMBR: Y = -1.491 * X2 + 26.0218 * X - 18.294
R2 = 0.885
BMBR: Y = -1.149 * X2 + 16.744*X + 36.636
R2 = 0.767
YMBR: Y = -0.683 * X2 + 10. 824 * X + 53.67
R2 = 0.885
BMBR: Y = -0.502 * X2 + 7.815*X + 67.01
R2 = 0.701
Figure 4.31 Variation in COD removal as function of HRTs in YMBR and BMBR
90
0 50 100 150 200 250 300 350 400 450 500
COD (mg/L)
0.00
0.50
1.00
1.50
2.00
2.50
Speci
fic
Gro
wth
Rate
day-1
)
Bacterial sludge
Mixed yeast sludge
Figure 4.32 Variation in specific growth rate of yeast and bacteria at 32 g salt/L in function
of COD
At SRT of 50 d, there was no significant difference between the COD removal
efficiency of YMBR (86-91%), and BMBR (91–93%) probably due to the lower F/M ratio
(0.35 - 0.40) in both reactors. As mentioned earlier, yeast growth is limited at low substrate
concentration. Thus, this result shows that maintenance of high MLSS (long SRT) in the
MBR can significantly enhance treatment efficiency for substrate limited growth. Moreover,
there was no appreciable change in the COD removal efficiency in the transition phase (Table
4.6). In this phase, low F/M ratios were maintained within the range of 0.35 - 0.55 by
controlling high HRT. High efficiency (95%) could be obtained at these low F/M ratios for
both YMBR and BMBR.
However, a conventional activated sludge system can be operated at a maximum VLR
of 1.2 kg/m3.d and F/M of 0.6 d
-1, and degradation rates reduced considerably with an
increase in salinity. Therefore, this high salinity wastewater should be treated at lower F/M
ratios (Kargi and Dincer, 2000). Three-fold-lower F/M ratios were applied in conventional
activated sludge at 30 g salt/L compared to those applied in salt-free wastewater (at the same
SRT) in order to obtain equivalent substrate removal.
In comparison, a comparable performance is obtained from the membrane bioreactors at
a very high salinity and VLR (3.0 - 5.0 kg/m3.d). Similar results have been reported by
Manem and Sanderson (1996) who found that a six-fold higher VLR could be applied for
dairy wastewater compared to conventional activated sludge system without deteriorating
performance. The results show that both YMBR and BMBR can be effectively used to treat
high salinity wastewater such as pickling and seafood processing wastewater to conform to
effluent standards of COD lower than 120 mg/L and BOD lower 20 mg/L.
91
Tab
le 4
.6 O
per
atin
g p
aram
eter
s an
d p
erfo
rman
ce o
f Y
MB
R a
nd B
MB
R i
n h
igh
CO
D l
oad
ing p
has
e
YM
BR
B
MB
R
SR
T
(day
s)
VL
R
(kg
CO
D/m
3.d
)
Mea
n
HR
T
(h)
ML
SS
(mg
/L)
F/M
(g/g
.d)
CO
D
rem
ov
al
(%)
Eff
luen
t
CO
D
(mg/L
)
Eff
luen
t
BO
D
(mg
/L)
SR
T
(day
s)
VL
R
(kg
CO
D/m
3.d
)
Mea
n
HR
T
(h)
ML
SS
(mg
/L)
F/M
(g/g
.d)
CO
D
rem
ov
al
(%)
Eff
luen
t
CO
D
(mg
/L)
Eff
luen
t
BO
D
(mg
/L)
10
2.6
6
8.8
505
0
0.4
9
94
50
10
10
2.9
7
8.1
48
00
0.3
8
97
25
< 5
10
2.9
5
7.7
503
0
0.5
6
94
50
15
10
3.5
7
6.3
46
00
0.7
8
97
25
< 5
10
3.6
6
6.1
444
0
0.8
3
84
15
0
90
10
4.3
0
5.2
57
30
0.7
6
92
70
30
10
4.5
9
5.0
453
0
1.0
2
76
23
0
15
0
10
-50
(*)
3.6
3
7.1
64
30
0.5
7
97
30
5
10
-50
(*)
3.5
8
7.2
530
0
0.5
8
96
45
15
10
-50
(*)
4.0
8
6.0
106
00
0.3
9
95
55
15
10-5
0(*
)4.2
8
6.1
115
00
0.3
7
95
50
15
50
5.5
6
4.7
131
00
0.4
3
93
70
30
50
4.9
3
5.3
145
00
0.3
5
91
90
40
50
6.3
5
4.0
163
00
0.3
9
91
10
0
55
50
6.5
5
4.0
150
00
0.4
4
86
15
0
90
(*)
Tra
nsi
tion
ph
ase
92
b. Membrane clogging
Membrane systems are often subjected to clogging, and this poses serious problems for
operation and maintenance. In order to investigate membrane clogging, experiments were
carried out in a continuous operational mode. Figures 4.29 and 4.30 show variations in
transmembrane pressure with time for YMBR and BMBR. The trend of pressure variation in
the YMBR and BMBR was similar to that in the high COD loading. Table 4.7 shows the
mean flux, transmembrane pressure and accumulated permeate volume during the filtration
cycle of the BMBR and the YMBR in both phases. In the high COD loading, it was observed
that the transmembrane pressure ( P) of the YMBR remained almost constant for
approximately 72 days before rising sharply. Whereas, the P increased sharply after 32 days
for the low COD loading in which the mean flux was three times higher than that in the high
COD loading. Total operating time (before P reaching to 70 kPa) for a SRT of 10 days is
higher than 50 days. For YMBR, there was no significant difference in the volume of the
permeate collected, although mean MLSS concentration at SRT of 50 days was 2.5 times that
at SRT of 10 days. This indicates that the clogging rate of the membrane is not entirely
dependent on the concentration of MLSS.
Table 4.7 Values of different parameters during YMBR and BMBR filtration cycle
SRT
(days)
Reactor Influent
COD
Mean
MLSS
(mg/L)
Mean flux
(L/m2.d)
Accumulated
permeate
(L/cycle)
Filtration
cycle
(d)
High COD loading:
50 YMBR 5,000 11000 34 2550 76
BMBR 1,200 20000 60 63 1.0
Low COD loading:
10 YMBR 1,000 4,500 95 3,100 38
BMBR 1,000 4,600 91 320 3.5
50 YMBR 1,000 11,000 98 2,910 32
BMBR 1,000 10,600 95 120 1.5
Membrane clogging in the BMBR was much more severe than in the YMBR. The
average filtration time for the BMBR, at SRT of 10 days was 3.5 days. This decreased to 1.5
days for SRT of 50 days, requiring frequent membrane washing. The rapid development of
biofilm for BMBR was in relation to the MLSS concentration in the reactor. At higher MLSS
concentrations (10,600 mg/L) for 50-day SRT, the rate of clogging was much higher than in
the YMBR. The difference in the performance between both YMBR and the BMBR is
probably due to the mechanism by which the biofilm develops on the membrane surface.
For bacterial system, biofilm develops by two mechanisms; first, by attachment of cells
on the membrane surface, which promotes further growth of bacteria; secondly by capturing
more cells (from the mixed liquor) in the already developed matrix formed by the first
process. The formation of biofilm on the membrane surface could be related to the production
of adhesive extracellular polymers (ECP) by bacteria. ECP is partly soluble in water and goes
into a colloidal suspension after production, which results in ECP accumulation in the mixed
liquor. During filtration, the macromolecular ECP compounds accumulate on the membrane
surface, and play the key role of binding the cells on the membrane surface and entrapping
larger organic particles in the slimy matrix. The availability of organic substance in the ECP
matrix promotes growth of new bacteria. This is further aided by the entrapment of more cells
93
from the mixed liquor, which seriously impairs membrane performance and module life.
Thus, at higher concentration of microorganisms in the mixed liquor, production of ECP
increases and biofilm develops at a much higher rate, which leads to rapid development of
transmembrane pressure.
The mechanism of biofilm development in the YMBR is different to that in the BMBR.
In the YMBR the yeasts attached physically to the membrane surface during filtration instead
of getting trapped in a matrix. Though ECP is produced in the YMBR, the quantity is much
less than that of the BMBR where a dense gel matrix is formed. The yeast cells are attached
together by physical interwinding of mycelia or pseudomycelia (Nishihara ESRC Ltd., 2001).
In addition larger yeast cells (2.5 - 3.9 m) in the YMBR also forms a secondary layer on the
membrane surface that acts like a second barrier to fouling particles and aggregates (Guell et
al. 1998). However, with increase in thickness of the cake, it gets heavier and starts sloughing
off by uplifting air flow or air-backwash, as soon as the cake is unable to sustain its own
weight. Thus the cake thickness cannot enlarge indefinitely and thus reduce the problem of
frequent membrane clogging. Mechanisms of flux enhancement by yeast sludge are shown in
Fig. 4.33.
Hydrogel(high resistance)
Porous cake(secondary filter layer)
Yeast cells
biofilm(ECP matrix)
Air bubble(from aeration)
membrane
Filtration Backwash
Partial detachment
Complete detachment
Figure 4.33 Possible mechanisms for flux enhancement by yeast cells
The membrane clogging depends upon the MLSS concentration of bacteria in the
BMBR, but to a much lesser extent for YMBR, primarily due to the difference in the
mechanisms of clogging. Due to this difference between BMBR and YMBR, a prolonged
filtration cycle (about 10 times higher compared to BMBR) could be obtained for the YMBR
without much problem with membrane clogging.
4.4 Sludge Characterization Study
In the membrane bioreactor system, fouling problem can be linked to the sludge
characteristics as discussed in Section 4.3.2.b. In order to investigate the difference in fouling
phenomenon in both YMBR and BMBR, sludge characterization was carried out. Series of
yeast and bacterial cultures were run at various salt contents. The sludge was examined for
ECP content, dewatering property (CST), viscosity as well as sludge settleability (SVI). These
results were also compared with sludge obtained from YMBR and BMBR systems operated at
32 g salt/L. Furthermore, in both these bioreactors, prior to chemical cleaning, the
characteristics of the sludge cake formed on the membrane surface was also analyzed.
94
4.4.1 Culture Study
Results of ECP, CST and viscosity of the mixed yeast and bacterial sludges at different
salt contents are presented in Table 4.8 and Figure 4.34. ECP content in the mixed yeast
sludge was very low at all salt contents compared to that in bacterial sludge. While the ECP
concentration of bacterial sludge rises considerably with salt content. Similar phenomena
have been reported for on flocculation of marine bacteria (Watanabe et al., 1998). They
suggested that the ECPs did not originate from autolysis, but seem to be excreted by living
marine bacterial cells. Unlike ECP production of bacteria living in salt-free water, which only
occurs during the stationary phase, marine bacteria can produce large amount of ECP in the
exponential growth phase (Flemming and Wingender, 2001). This leads to significant
accumulation of ECP in a high salinity environment. In addition, due to alteration of genetic
structure under high osmotic stress by NaCl concentration, bacteria can synthesize more
specific proteins that may contribute to increase in the ECP production (Vijaranakul et al.,
1997). Eikelboom (2000) found that Zoogloea bulking problems were due to high production
of ECP in the bacterial flocs. This phenomenon could explain the acceleration of biofouling
problem in BMBR.
In contrast, the production of ECP in yeasts is much lower and its concentration remained
low. Reeslev et al. (1996) reported that Aureobasidium pullulans, a mycelial yeast, can only
synthesized ECP when growth was nitrogen-limited while no ECP was produced when the
culture was carbon-substrate-limited. The feed wastewater was rich nitrogen (COD:N =
100:15), and therefore growth of yeast is carbon-substrate-limited. This may result in less
ECP production, saving it from frequent clogging.
Table 4.8 Yeast and bacterial sludges characterization
Sludge Salt
(g/L)
MLSS
(mg/L)
SS (*)
(mg/L)
SVI
(mL/g)
ECP
(mg/g)
CST
(s/g) (**)Viscosity
centipoises
Bacteria
0.5 3320 80 72 28.4 1.48 5.04
15 3860 152 63 29.0 1.22 5.14
32 3200 232 81 34.6 8.41 5.60
45 3840 220 108 48 14.9 7.14
Bacterial MBR 32 6320 N/A 159 57.8 9.8 7.26
Yeast
0.5 3260 1373 N/A < 1 1.19 5.01
15 4020 1767 N/A l<1 1.12 5.28
32 4480 1440 N/A 5 0.94 6.00
45 4580 1530 N/A 7.5 0.94 6.34
Yeast MBR 32 4300 N/A N/A 11.4 1.44 4.86(*) Suspended solids concentration of supernatant after 2 h of settling (**) s/g = second/g MLSS
95
0 10 20 30 40 50
Salt (g/L)
0
10
20
30
40
50
ECP (
g /
mg)
Mixed yeast sludge
Mixed bacterial sludge
0 10 20 30 40 50
Salt (g/L)
0
4
8
12
16
CST (
s/g M
LSS)
a) ECP vs. salt b) CST vs. salt
Figure 4.34 Variation in ECP and CST in function of salt content
Sludge dewaterability is measured by CST, which is used as a relative indicator to
characterize the performance of sludge dewatering. While SVI is used as a field measure of
sludge settleability. CST of bacterial sludge increased from 1.5 to 15 s/g with the increase of
salt content from 15 to 45 g/L. Similarly, SVI of the bacterial sludge also increased with an
increase in salt content, although the SVI values remained lower than 150 (a value
corresponding to bulky sludge) even for a salt content of 45 g/L. However, the bacterial
sludge at high salinity (above 30 g/L) was not compacted and remained fluffy compared to the
sludge at salt contents of 15 g/L or lower. Similarly, it was that increasing viscosity trend with
salt content was comparable to CST and ECP. In general, viscosity is dependent on MLSS.
MLSSs of cultures at different salt contents were maintained relatively stable in this study.
Thus, increase in viscosity with salt increase may be due to the ECP content of the sludge.
Microscopic examination did not show any filamentous bacteria or fungi in the bacterial
sludge even at high salt contents, and relationships could be established between the increase
in the SVI, CST and ECP with salt content for the bacterial sludge. The linear correlation
between SVI and ECP was found in some previous studies (Goodwin and Foster, 1985;
Urbain et al., 1993). The increase in CST and SVI as function of salt content in bacterial
sludge may be attributed to excess sodium which can cause a deterioration in the floc
structure, and an increase in ECP. This results in weak flocs, poor settling (increase in SVI)
and dewatering (increase in CST). Likewise, Urbain et al. (1993) reported that ECP content
was correlated linearly with SVI, high polysaccharides and proteins resulting in a worsening
of sludge settleablity. In addition, the increase of ECP can be explained by the structure of the
three-dimensional ECP matrix kept together by divalent cations such as Ca2+
, Mg2+
. An
exchange of divalent cations (Ca2+
, Mg2+
) with Na+ will take place when the ratio of Na
+ to
divalent cation (Na+:M
2+) exceeds two. In general, the Na
+:M
2+ ratio is very high in saline
environment. In such conditions, large amount of polymers produced resulted in weaken the
floc strength (Bruus et al., 1992). These ECP molecules extend out from cell surfaces and
form a dense gel that retains water in gel pores (Liao et al., 2001). Liss et al. (1996) also
reported that these complexes and hydrated ECPs within the floc matrix have a large capacity
to retain water, which decreases dewaterability. Furthermore, suspended solids in the
supernatant of bacterial batches appear to increase when the CST or SVI (vis-à-vis ECP)
increases (Fig. 4.35). This may be due to disruption of weakened floc structure at high salinity
(by shearing forces from aeration) which forms pin-point flocs and thus results in high
effluent suspended solids.
96
0 10 20 30 40 50
Salt (g/L)
60
80
100
120
140
SVI (mL/g)
10
20
30
40
50
ECP (mg/g)
SVI
ECP
0 10 20 30 40 50
Salt (g/L)
5.0
6.0
7.0
8.0
Viscosity (centipois)
0
50
100
150
200
250
SS of supernatant (mg/L)
SS of supernatant
Viscosity
Figure 4.35 Variation in SVI, SS, ECP and viscosity with salt content in mixed bacterial
cultures
For mixed yeast sludge, CST remained relatively constant with salt increases. This can
be understood in terms of the negligeable amount of ECP produced and indicates that the
yeast sludge has better dewatering or thickening ability than bacterial sludge at high salinity,
which is a specific advantage of the yeast system. The viscosity of yeast mix liquor, was
slightly higher than that of bacteria. This can be attributed to the difference between MLSS of
mixed yeast (4500 mg/L) and bacterial sludges (3500 mg/L). Thus, unlike bacterial sludge,
viscosity of yeast sludge is mainly effected by biomass concentration, but not by ECP.
Table 4.8 shows that SS in the supernatant of the mixed yeast sludge was high even
after 2 hours of settling. SS removal efficiency was only around 65%. The SS in supernatant
were mainly dispersed as small yeast cells, which had not been captured by mycelia or
pseudomycelial yeast matrix as shown by microscopic examination. Settled solids after two
hours of settling were found to be compacted, and mainly consisted of yeasts interwind with
each other and large yeast cells. Unlike activated sludge flocs, yeast flocs could not trap all
fine yeast cells or fine particles, which led to high SS in the supernatant even after prolonged
settling time. Conflicting results were obtained by different studies concerning the floc
characteristic of yeast sludge. Nishihara ESRC Ltd. (2001) obtained large yeast flocs that
settled quickly while other authors (Hu,1989; Arnold et al., 2000) obtained poor settling
sludge when dealing with different yeast strains. This suggests that the predominance of
mycelial yeasts may depend on the free competition among different yeast strains in
wastewater.
4.4.2 YMBR and BMBR
Large differences of ECP obtained from both YMBR and BMBR and the batch reactors
(run at 32 g salt/L) are shown in Figure 4.36. This may be due to of washing-out of
macromolecules (ECPs and proteins) with fine particles in supernatant during the decantation
stage, while most of these substances were retained by membrane module with spore size of
0.1 m. However, in comparison between YMBR and BMBR, the ECP concentration of
mixed bacterial sludge was higher than that of yeast sludge at both SRTs. This observation
was similar to that obtained in batch studies.
97
YMBR BMBR0
100
200
300
400
500
600
mg E
CP/g
dried b
iom
ass
SRT of 10 d
SRT of 50 d
Sludge cake
(attached to the membrane surface)
1139
147
58
232
490
Figure 4.36 ECP contents of mixed yeast and bacterial sludges in YMBR and BMBR
Meanwhile, visual observation of the clogged membrane revealed that large amount of
extremely viscous and gelatinous sludge cakes was attached to the BMBR (Appendix A).
However, the YMBR has a very thin layer of sludge cake, which could be easily washed out
with tap water. This difference in the nature of the attached sludge contributes to the
improvement of the membrane performance in the YMBR process.
In both the YMBR and BMBR, the ECP content seems to vary with the SRT, as shown
in Fig. 4.36. It was found that ECP concentration increased with increase in SRT for both
BMBR and YMBR. This can be explained by the variation in ECP production rate at different
growth phases. Low SRT corresponds to the stationery phase, while long SRT (50 d)
corresponds to the decay phase in which cell lysis takes place. Similarly, Pavoni et al. (1972)
reported that during the early decay phase, the rate of ECP production was maximum.
Sheintuch et al. (1986) reported that ECP content is a function of SRT in continuous
bioreactors and it increased linearly with SRT.
4.4.3 Microscopic Observations of Mixed Yeast Sludge
From microscopic observations, it was noted that there were changes in predominant
yeast strains and presence of other microorganisms (protozoa, rotifers) when operating
conditions was changed. The operating conditions consisted of influent COD concentration,
organic loading, SRT and the type of substrate (glucose and protein extract as protein source)
and salt content. Photographs of mixed yeasts and bacteria flocs are shown in Appendix A.
Most yeast cells in the cultures were larger than those in the YMBR, even though both
reactors was run at the same operating conditions (Protein-feed wastewater, SRT of 50 days,
COD of 5,000 mg/L). Mycelia (hypha filament) and large size egg-shaped cells with
monopolar budding were predominant in the cultures. While round or smaller egg-shaped
cells, which may be different yeast strains from the cultures, were predominant in the YMBR
(mean size of 2.4 m). The mean size of the mother cells was 3.9 m. This might be due to
membrane with pore sizes of 0.1 m retaining most fine size cells among which minority of
acid-tolerant bacteria was involved. Fine cells suspended in supernatant were washed out by a
decanting step after more than 2 hours of settling in batch operation.
98
For substrate as glucose, most yeast cells are round with multilateral budding and mean
size of 5.5 m. In the YMBR, predominant yeast strains were also changed as organic loading
rate varied. The white or light brown color of yeast mixture gradually changed to dark brown
when the high COD loading (5,000 mg COD/L) changed to low COD loading (1,000 mg
COD/L). The majority of yeast colonies were black, orange and yellow growing on the yeast-
glucose-peptone agar when growing in the low COD loading (influent COD of 100mg/L).
Most yeast cells were egg-shaped and mycelia. It could be easily observed that free-living
ciliates (protozoa) grow well in this yeast mixture. These ciliates move freely and rapidly in
the mixed liquor. This change may be due to the increase in DO concentration. DO of all runs
in the low COD loading is above 4.0 mg/L, whereas DO in that with COD of 5,000 mg/L was
less than 2.0 mg/L.
4.4.4 Nutrient Uptake
Nutrients of mixed yeast and bacterial sludges are presented in Table 4.9. The volatile
solids of both yeast and bacterial sludges were not significantly different, ranging from 89 to
95% of dried solids. The nitrogen content of the mixed yeasts sludge fed with glucose
wastewater was found to be 7.1% on average compared to 3.2 % for bacterial sludge.
Similarly, the phosphorus uptake ability of the yeasts was approximately twice as high as that
of bacterial sludge. For fish-protein wastewater, nitrogen content of the mixed yeast cultures
was more than 15%. In general, nitrogen content of the yeast and fungi biomasses was in the
range of 7-12% (Westhuizen and Pretorius, 1998; Defrance, 1993). Thus the amount of
nitrogen uptaken for protein-feed wastewater is too high (above 15%) compared to normal
values or values of yeast sludge fed with glucose wastewater (7.1%). This may be attributed
to the precipitation of a part of influent protein at low pH (3.5). The nitrogen content of yeast
sludge in the YMBR was not excessive (7.64%) and was in the normal range. Thus, it is
postulated that in the YMBR process with a low F/M ratio, yeast can produce enzyme to
hydrolyze protein aggregates and can assimilate them.
Table 4.9 Composition of mixed bacterial and mixed yeast sludge
Volatile solids
(%)
Nitrogen content
(% of dries solids)
Phosphorous content
(% of dried solids) Microorganisms
Glucose Protein Glucose Protein Glucose Protein
Yeast sludge:
+ 20 g /L: 94.1 90.3 6.79 15.2 2.23 3.61
+ 32 g /L: 92.3 93.4 7.28 17.1 1.79 3.43
+ 45 g /L: 95.0 90.5 7.19 15.7 1.70 3.53
YMBR sludge 91.5 7.64 3.52
(low-COD loading)
Bacterial sludge
+ 20 g /L: 89.5 91.6 2.70 5.61 0.61 1.56
+ 32 g /L: 92.0 92.9 3.05 5.02 0.86 2.37
+ 45 g /L: 94.2 90.2 3.74 5.32 0.90 2.23
BMBR sludge 89.7 5.21 1.51
(High COD loading)
The nitrogen and phosphorous contents of the bacterial biomass fed with protein-feed
wastewater were approximately twice higher than that fed with glucose wastewater as shown
in Table 4.9. By exception of nutrient uptake into real bacterial cells, the remaining of these
99
nutrients may entered into composition of floc structure such as ECP and phosphate bonds
which can enhance ECP production.
Even when different wastewaters were used, the mixed yeast sludge from the cultures or
the YMBR contained higher nutrients than the bacterial sludge. The average crude protein
content obtained was 45% (corresponding to 7.2% N). This value was similar to the protein
content obtained from the well-known “Symba” process which is a single-cell-protein
production (SCP) for human food consumption by using potato processing waste to culture
yeasts Endomycopsis fibuliger and Candida utilis. In general, algal and bacterial biomasses
are less pleasant to taste because they contain undesirable levels of certain cellular materials
such as high nucleic acid content, toxic or carcinogenic substances absorbed from the growth
substrate. By contrast, yeasts and most fungi are quite acceptable to animals and man due to
the abundance of valuable nutritious substances such as proteins and vitamins. Thus, it is
suggested that a combination of yeast treatment and SCP production can be a cost-effective
approach for seafood processing industries which, at present, face difficulties in treatment
efficiency and high costs.
100
Chapter 5
5
Conclusions and Recommendations
This study investigated biological processes in using wild salt-tolerant yeast and
bacteria for treatment of saline seafood processing wastewater. Basic studies on biokinetic
coefficients and optimum operating parameters of yeast and bacterial treatment were
conducted. The effects of high salt contents (20, 32 and 45 g/L NaCl) on the biokinetic
coefficients were evaluated using respirometric method. Then the optimum operating
parameters for the yeast and bacterial treatments were found from the parametric study using
the acclimatized mixed yeast and bacterial cultures.
The main part of this study focused on the membrane bioreactor. The potential for
developing membrane bioreactor systems using salt-tolerant yeast and bacteria to treat saline
seafood processing wastewater was examined. A comparative evaluation of treatment
performance of both systems was done. The last section focused on sludge characteristics
concerning membrane clogging. The relationship between sludge properties and membrane
flux decline was investigated. The conclusions drawn from these results are presented below.
5.1 Conclusions
From the biokinetic study, it can be concluded that the yeast is more efficient for
treating wastewater containing high organic load and high salt content. It can be reasoned that
they would be more suitable for varying salt loads because they require lower acclimation
time than does bacterial culture. This was attributed to their better osmotolerant properties.
The salt inhibition was found to be much higher for yeast culture. The maximum specific
growth rate for yeast is higher than for bacterial culture at high salt contents. However, yeast
growth was more inhibited at low CODs.
The results of the parametric study indicate that the osmotolerant yeasts were able to
tolerate wider pH range than bacterial culture. Total OUR of yeast sludge was highest for pHs
5.0 –5.5. The respiration rate of yeasts was inhibited at pH 2.5 or pH above 9.0. The OUR of
yeast at pH 3.5 was slightly lower than that at pH 5.0 - 5.5. However, pH of YMBR was
maintained at pH 3.5 in order to limit bacterial contamination. The results of SRT study show
that the highest nitrogen removal by uptake into yeast biomass was obtained at SRT of 10 d,
whereas the maximum COD removal efficiency was obtained at SRT above 45 d.
In the high COD loading of membrane bioreactor study, the COD removal rate for
BMBR was lower than the YMBR at high VLRs at high salt contents (32 g/L). Thus, the
mixed yeast system could be subjected to higher F/M ratio. Even though DO concentration of
yeast mixed liquor was lower than 1.0 mg/L at the high F/M ratio, the treatment efficiency of
the yeast system does not decline. This may be due to the structure of yeast flocs facilitating
oxygen diffusion. It is suggested that the yeast system represent a better substitute for an
anaerobic system in terms of COD removal rate at high salinity.
The low COD loading phase revealed that both YMBR and BMBR give high COD
removal efficiency (>90%) at high salt content, low F/M ratio and high SRT. Both reactors
generated good effluent quality (COD < 120 mg/L, BOD < 20 mg/L and SS < 5 mg/L).
101
Yeast sludge in fact achieve significantly better reduction in the membrane clogging
rate than bacterial sludge. BMBR is highly prone to membrane clogging, whereas YMBR can
be operated at a relatively low pressure for prolonged filtration cycle. The filtrate cycle of
YMBR was approximately 10 times higher than BMBR. Thus using yeasts in the
biomembrane reactor can enhance membrane performance and has the potential to improve
the economics of treatment system due to reducing operating and maintenance costs.
Several factors such as mechanisms of biofilm formation, concentration of ECP and size
of cells contributed to the better filtration cycle of the YMBR. Reduction of problems
associated with membrane clogging supports the use of YMBR in practice. Variation of ECP
content that is responsible for biofilm as well as floc formation was found to vary with salt
content for bacterial sludge, whereas for yeast sludge, ECP concentration remained practically
constant even at high salt contents. Along with increases in ECP, CST of bacterial sludge was
increased at high salinity, while CST of yeast sludge remained practically constant, indicating
that dewatering would be easy for high salinity wastewater. Thus, using yeasts in membrane
bioreactor can enhance membrane performance and reduce the operational problems
associated with sludge dewatering and disposal.
Nitrogen and phosphorous contents of the yeast sludge were approximately twice higher
than those of bacterial sludge. This suggests that yeasts have high nutritional values. In
addition, yeasts and most fungi, are quite acceptable to animals and man due to the abundance
of other valuable nutritious substances (vitamins). Thus, a combination of yeast treatment and
SCP production can be a cost-effective approach for seafood processing industries.
5.2 Recommendations
Based on the extensive experimental data obtained, several recommendations for future
studies can be outlined:
High Salinity Wastewater
1. Due to flux-enhancing ability of yeast sludge, operation modes for yeast membrane
bioreactor can be examined to shorten backwash time. This may results in reduction of
operation costs and increase in total permeate flux.
2. This study has not evaluated in depth the yeast sludge properties at different salt
contents which may be related to membrane flux. These properties can consist of
specific filtration resistance, hydrophobicity, surface charge, bound water, cell size and
composition of ECP. In order to understand thoroughly the fouling inhibition
mechanism of yeast sludge, a detailed study of the sludge properties at various salt
content should be undertaken.
3. ECP production of mixed yeast sludge or activated sludge under varying nutrient
compositions or cation concentrations of influent wastewater may be effected. A
balance of nutrient contents or cations (divalent cation: monovalent cation) may lead to
significant decrease in the membrane clogging rate. A study on the effects of nutrient
and cation contents on ECP production would perhaps be useful.
4. Since seafood factories process a large range of products with important seasonal
variations, changing in pollution characteristics vary significantly from plant to plant,
and even within the same plant. Large variations in salt content can therefore be
expected. Thus, a study on the effects of salt shock loading on yeast sludge may be
102
necessary. The shock salt loading experiment for bacterial sludge may be conducted in
parallel to obtain a comparative evaluation.
5. In order to confirm the effectiveness of yeast treatment, pilot-scale yeast membrane
bioreactors for long-term treatment of real saline seafood processing wastewater should
be developed.
Biomembrane process and Membrane clogging
1. The wastewater pH of the environment influences surface charges, protein deposition
and deflocculation. Low pH can cause an increase in surface charges, protein aggregates
and deflocculation, which may enhance membrane filtration water flux. However, low
pH values will also inhibit bacterial growth. Therefore, optimum pH values to control
membrane clogging in the bacterial system should be considered.
2. A low pH environment may result in predominance of acid-tolerant microorganisms
such as fungi, yeasts and acidogenic bacteria. Thus, under suitable operating conditions
(such as pH, DO, organic loading and SRT) there may be a symbiotic relationship
between these acid-tolerant microorganisms. For example, acidogenic bacteria enable to
hydrolyze and convert quickly organic complexes (such as protein, lipids, and
polysaccharides) to lower molecular-weight intermediate compounds (such as VFAs,
amino acids and short-chain carbohydrates) which may be suitable substrates for yeast
and fungal growth in a low pH environment. Therefore, using a mixture of acid-tolerant
microorganisms in the biomembrane process for treatment of certain wastewaters can be
investigated.
3. In this study, it was observed that ECP production increases with salt content. Thus,
combination of activated carbon (AC) adsorption and the bacterial membrane process
should be examined. Activated carbon addition to the BMR process can enhance
permeate flux by forming porous cakes and can remove refractory organic matters by
AC adsorption. Therefore, this method may be applied successfully for TOC removal as
a pre-treatment of raw water in the RO process.
Using Yeasts and Fungi for treatment of toxic wastewaters
Based on tolerance ability of yeasts or fungi in extremely strict conditions, further
investigation are proposed:
1. A comparative study of bacterial and yeast sludge response to acute and chronic heavy
metal stress using respirometric method is recommended. The results from such a study
might provide information on quantitative toxicity evaluation of the heavy metals for
both yeast and bacterial treatment systems. Resting and growing cells (biomass) can
demonstrate completely different mechanisms of resistance to acute and chronic heavy
metal stress. Thus, mechanisms of heavy metal biosorption by yeasts and bacterial
sludge may also be investigated.
2. Using yeast or fungi biomembrane process with low F/M ratio for treatment of high
strength hazardous organic wastewater may be a feasible biological approach. Toxic
wastes or chemicals suggested include pesticides, herbicides, phenolic derivatives,
aromatic compounds, cyanides and tannery wastewater.
3. Identification of the various yeast or bacteria present at high salt conditions.
103
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A-1
Fig. A-1 Yeast colonies cultured with glucose-feed wastewater containing high salt concentrations
Round, smooth colony
Irregular, rough colony
Round, smooth colony
Irregular, rough colony
Fig. A-2 Two predominant yeast colonies cultured with glucose-feed wastewater
Fig. A-3 Predominant yeasts in the batch culture with glucose-feed wastewater
containing 32 g/L salt (x 1500)
10 m
10 m
A-2
Fig. A-4 Predominant yeasts in the batch culture with protein-feed wastewater
containing 32 g/L salt (x 500)
Fig. A-5 Predominant yeasts in the batch culture with protein-feed wastewater
containing 45 g/L salt (Growth of mycelia yeasts occurred) ( x 500)
20 m
20 m
A-3
Fig. A-6 Mycelia yeasts (cultured with fish-protein wastewater) ( x 800)
Fig. A-7 Yeast flocs formed by interwinding of mycelia yeasts
(settled sludge in theYMBR) ( x 500)
20 m
20 m
A-4
Fig. A-8 Fine yeast cells suspended in supernatant of YMBR
Fig. A-9 Free-living ciliates (protozoa) grew well in the yeast mixture of YMBR
(in COD loading) (x 250)
Protozoa
20 m
10 m
A-5
Fig. A-10 Bacterial flocs in the batch culture with protein-feed wastewater at 15 g/L salt
(rounded and compacted sludge flocs) (x250)
Fig. A-11 Bacterial flocs in the batch culture with protein-feed wastewater at 32 g/L salt
(Open and weak flocs) ( x 250)
Fig. A-12 Bacterial flocs in the BMBR
(Fine and weak flocs) (x 250)
10 m
10 m
10 m
A-6
Fig. A-13 Respirometer system (Respirometer, recorder, DO meter and thermostat)
Fig. A-14 Respirometer
A-7
Fig. A-15 Respirogram (low COD dose)
Fig. A-16 Respirogram (high COD dose)
A-8
Fig
. A
-17 Ye
ast
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-18
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-20
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. A
-21
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-23
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B-1
Table B-1 Acclimation of mixed bacterial sludge to 20 g salt/L with the glucose-feed
wastewater
Time Salt Cl HRT CODin CODeff Ecod MLSS pHo pHt SVI
day g/L NaCl g/L Cl h mg/L mg/L % mg/L mL/g
2 20.5 24 952 175 81.6 2,116 6.79 6.79 88
4 19.0 23 931 143 84.6 2,120 7.17 7.47 83
5 18.9 22 940 143 84.8 2,132 6.76 6.35 81
7 19.6 24 930 154 83.4 2,274 6.99 6.11 71
9 20.0 24 930 136 85.4 2,488 7.61 7.01 70
11 20.9 29 1,010 32 96.8 2,640 7.80 7.10 80
13 20.6 28 870 51 94.1 2,516 7.75 6.85 87
15 21.3 25 915 35 96.2 2,668 7.84 6.40 47
17 21.3 12.9 25 915 60 93.4 2,672 7.69 6.33 45
19 21.3 12.9 26 915 55 94.0 2,888 7.32 6.22 43
28 20.0 12.2 24 1,090 52 95.2 2,750 7.80 6.26 36
40 20.5 12.4 22 1,070 102 90.5 2,724 7.47 6.28 34
42 20.5 12.4 15 1,070 127 88.1 2,760 7.75 6.52 35
44 20.5 12.0 14 1,070 24 97.8 2,837 6.97 6.16 34
46 20.5 12.4 17 1,070 71 93.4 2,993 7.64 6.53 32
Table B-2 Acclimation of mixed bacterial sludge to 32 g salt/L with the glucose-feed
wastewater
Time Salt Cl HRT CODin CODeff Ecod MLSS pHo pHt SVI
day g/L NaCl g/L Cl h mg/L mg/L % mg/L mL/g
0 20.5 24 952 181 81.0 2,072 6.79 6.88 91
2 19.8 23 931 145 84.4 1,948 7.11 7.65 79
4 20.9 22 940 134 85.7 2,152 6.89 6.34 81
5 24.9 24 930 185 80.1 2,324 6.86 6.06 62
7 25.0 24 930 121 87.0 2,292 7.44 6.64 71
9 26 24 950 145 84.7 2,208 7.51 7.33 76
11 26.6 24 950 95 90.0 2,190 7.5 6.88 78
13 29 26 950 100 89.5 2,176 7.85 7.67 82
15 31.8 26 870 120 86.2 2,328 7.73 7.61 82
17 32.6 25 870 31 96.4 2,576 7.84 6.75 74
19 34 25 915 12 98.7 2,852 7.85 7.19 49
28 33.5 20.3 25 915 88 90.4 3,260 7.28 7.93 40
30 33.9 20.6 26 915 61 93.3 3,000 7.31 7.93 43
38 32.6 19.8 24 1,090 169 84.5 2,300 7.42 7.24 24
40 33.6 20.4 24 1,090 118 89.2 2,164 7.43 7.41 25
42 32.2 19.6 15 1,090 125 88.5 2,253 7.44 7.72 23
44 32.2 19.6 14 1,090 23 97.9 2,337 7.22 7.23 29
46 32 20.0 17 1,090 29 97.3 2,501 7.46 7.99 16
B-2
Table B-3 Acclimation of mixed bacterial sludge to 45 g salt/L with the glucose-feed
wastewater
Time Salt Cl HRT CODin CODeff Ecod MLSS pHo pHt SVI
day g/L NaCl g/L Cl h mg/L mg/L % mg/L mL/g
0 20.5 24.0 952 415 56 2,040 6.8 6.81 86.5
2 19.3 23.0 931 420 55 1,888 6.93 7.5 79.4
4 21.4 22.0 940 431 54 2,100 6.78 6.37 71.8
5 24.8 24.0 930 445 52 1,738 6.89 6.47 81.7
7 24.9 24.0 930 395 58 2,408 7.51 7.71 62.3
9 25.5 24.0 1,010 375 63 2,344 7.33 7.51 64.0
11 27.4 24.0 1,010 310 65 2,143 7.2 6.68 70.0
13 30.5 24.0 1,010 325 70 2,048 7.7 6.64 75.0
15 33.5 25.0 870 215 67 2,200 7.85 7.72 70.5
17 33.5 25.0 870 201 77 2,412 7.87 6.89 71.3
19 37.4 26.0 875 215 75 2,792 7.85 7.31 58.7
21 40.0 25.0 915 197 78 2,748 7.85 7.41 62.2
23 42.0 26.0 1,031 220 79 2,848 7.14 7.86 54.1
25 42.2 25.0 1,031 231 81 2,936 7.85 7.2 48.7
26 46.2 24.0 1,031 205 80 3,252 7.26 7.75 44.9
28 45.3 27.5 24.0 1,031 157 85 3,270 7.38 7.93 35.9
30 43.5 26.4 24.0 1,031 120 88 3,310 7.42 7.92 33.1
38 44.3 26.9 24.5 1,090 135 88 3,100 7.48 7.79 21.0
40 47.0 28.6 24.0 1,090 148 86 3,400 7.54 7.98 18.0
42 44.4 27.0 24.0 1,090 167 89 3,230 7.56 7.98 18.0
Table B-4 Acclimation of mixed yeast at 20 g salt/L with the glucose-feed wastewater
Time Salt Cl HRT CODin CODeff Ecod MLSS pHo pHt
day g/L NaCl g/L Cl h mg/L mg/L % mg/L
0 20.5 24 5,100 325 93.6 3,700 5.61 3.05
2 21.5 24 5,100 305 94.0 5,420 5.87 2.86
4 20.8 26 5,100 370 92.7 6,732 5.97 2.92
6 21.1 26 4,980 272 94.5 7,916 5.54 2.91
8 20.6 25 7,400 363 95.1 6,676 5.64 2.54
10 20.1 15.9 27 7,400 1,230 83.4 6,416 5.72 3.25
21 20.1 15.9 26 7,400 1,270 82.8 5,028 5.66 4.22
23 20.9 14.5 27 5,350 447 91.6 6,596 3.50 2.93
33 20.2 23.1 22 7,490 352 95.3 7,756 3.57 2.37
35 20.0 14.5 15 7,490 352 95.3 9,340 5.00 2.40
37 20.0 14.6 21 7,490 331 95.6 10,367 4.79 2.51
39 20.6 15.5 15 7,490 444 94.1 12,216 4.07 2.32
B-3
Table B-5 Acclimation of mixed yeast at 32 g salt/L with the glucose-feed wastewater
Time Salt Cl HRT CODin CODeff Ecod MLSS pHo pHt
day g/L NaCl g/L Cl h mg/L mg/L % mg/L
0 32.8 24 4,980 920 81.5 4,679 4.09 2.78
2 33.0 24 4,980 1,010 79.7 4,710 5.52 2.83
4 32.3 25 4,980 415 91.7 5,704 5.32 2.83
5 32.3 24 4,980 428 91.4 5,828 5.51 3.13
7 31.3 24 4,980 351 93.0 3,804 5.49 3.07
9 32.6 24 5,100 368 92.8 4,242 5.84 2.86
11 33.5 27 4,980 413 91.7 6,596 5.92 2.98
13 32.2 27 4,980 409 91.8 8,040 5.56 2.84
15 32.6 25 7,400 355 95.2 7,496 5.64 2.50
17 33.8 26 7,400 382 94.8 10,168 5.46 3.34
28 31.9 23.0 24 7,400 478 93.5 9,312 5.78 2.78
30 31.7 23.5 24 7,400 423 94.3 10,152 5.08 2.52
38 32.0 22.5 24 7,500 369 95.1 13,944 4.46 2.64
40 31.7 21.7 23 7,400 449 93.9 14,324 5.14 2.65
42 32.0 22.5 24 7,400 396 94.6 14,713 4.56 2.59
44 33.0 22.5 21 7,400 278 96.2 14,557 5.36 2.58
46 31.8 21.1 24 7,400 411 94.4 15,550 4.57 2.53
Table B-6 Acllimation of mixed yeast to 45 g salt/L with the glucose-feed wastewater
Time Salt Cl HRT F/M CODin CODeff Ecod MLSS pHo pHt
day g/L NaCl g/L Cl h d mg/L mg/L % mg/L
0 32.0 23 1.12 5,100 1,720 66.3 3,768 5.54 3.12
2 32.0 24 1.15 5,100 1,560 69.4 3,984 5.88 2.92
4 32.0 24 0.82 4,980 1,150 76.9 6,100 5.97 2.96
5 35.0 26 0.63 4,980 1,095 78.0 7,920 5.66 2.92
7 38.0 25 0.58 4,980 950 80.9 8,652 5.84 2.77
9 40.8 24 0.56 5,050 765 84.9 8,950 5.68 2.87
11 43.3 23 0.55 5,100 680 86.7 9,240 5.66 2.83
13 45.3 24 0.51 5,100 719 85.9 10,060 5.43 2.73
15 44.1 31.7 24 0.50 5,100 539 89.4 10,144 5.79 2.84
17 43.1 31.6 25 0.46 5,050 471 90.7 10,984 5.26 2.53
30 44.9 29.1 27 0.38 5,100 369 92.8 13,480 4.99 2.54
40 45.0 30.7 24 0.36 5,200 345 93.4 14,456 5.00 2.4
42 44.1 30.1 26 0.32 5,200 362 93.0 16,047 5.23 2.53
44 45.0 31.7 24 0.33 5,200 411 92.1 15,733 5.37 2.49
46 45.0 32.2 24 0.32 5,100 342 93.3 15,967 4.88 2.54
B-4
Table B-7 Acclimation of mixed yeasts to protein-feed wastewater at 20 g/L salt
Time CODin CODeff COD% MLSS F/M
days mg/L mg/L mg/L d-1
0 4930 5450
2 5120 1587 69 5400 0.63
4 5210 1667 68 5870 0.59
6 5020 1456 71 6780 0.49
8 4970 1243 75 7920 0.42
10 4760 809 83 9250 0.34
12 5010 752 85 9350 0.36
14 5150 567 89 10850 0.32
16 5090 458 91 10750 0.32
Table B-8 Acclimation of mixed yeasts to protein-feed wastewater at 32 g/L salt
Time CODin CODeff COD% MLSS F/M
days mg/L mg/L mg/L d-1
0 5010 5700
2 4850 1795 63 5400 0.60
4 5120 1640 68 5990 0.57
6 4950 1730 65 5930 0.56
8 4760 1190 75 6950 0.46
10 5230 1200 77 7050 0.49
12 5100 870 83 8345 0.41
14 5010 701 86 9050 0.37
16 5120 770 85 9710 0.35
Table B-9 Acclimation of mixed yeasts to protein-feed wastewater at 45 g/L salt
Time CODin CODeff COD% MLSS F/M
days mg/L mg/L mg/L d-1
0 4790 5500
2 5100 2091 59 5200 0.65
4 5020 1640 61 5350 0.63
6 4970 1730 61 5320 0.62
8 5120 1587 69 6240 0.55
10 5070 1200 72 8705 0.39
12 5025 870 83 8956 0.37
14 4790 701 84 9310 0.34
16 4990 770 83 9420 0.35
B-5
Table B-10 Acclimation of mixed bacterial sludge to protein-feed wastewater at 20 g/L salt
Time CODin CODeff COD% MLSS F/M
days mg/L mg/L mg/L d-1
0 3120
2 1020 92 91 3650 0.28
4 990 79 92 4270 0.23
6 1070 86 92 4750 0.23
8 1120 56 95 5120 0.22
10 1010 51 95 5250 0.19
12 980 29 97 5430 0.18
14 990 50 95 5910 0.17
16 1130 45 96 5670 0.20
Table B-11 Acclimation of mixed bacterial sludge to protein-feed wastewater at 32 g/L salt
Time CODin CODeff COD% MLSS F/M
days mg/L mg/L mg/L d-1
0 3560
2 1100 165 85 3750 0.29
4 980 78 92 4340 0.23
6 995 90 91 4950 0.20
8 1020 51 95 5120 0.20
10 1050 73 93 5010 0.21
12 950 38 96 5700 0.17
14 970 78 92 5600 0.17
16 1105 55 95 6310 0.18
Table B-12 Acclimation of mixed bacterial sludge to protein-feed wastewater at 45 g/L salt
Time CODin CODeff COD% MLSS F/M
days mg/L mg/L mg/L d-1
0 3950
2 1010 222 78 3650 0.28
4 1200 228 81 3750 0.32
6 1105 221 80 4210 0.26
8 1030 175 83 4760 0.22
10 970 175 82 4790 0.20
12 990 109 89 5700 0.17
14 1025 113 89 5970 0.17
16 990 89 91 6105 0.16
B-6
Table B-13 COD and DO profile data of the mixed yeast batch culture with glucose-feed
wastewater at 20g/L salt
Time COD DO COD% U MLSS
h mg/L mg/L g/g.d mg/L
0.0 5020 8250
2.5 1250 1.2 75 4.15
5.0 220 6.4 95.6 2.65
7.5 230 6.3 95.4 1.76
9.0 205 6.4 95.9 1.48
10.5 220 6.4 95.6 1.26 9130
Average 8700
Table B-14 COD and DO profile data of the mixed yeast batch culture with glucose-feed
wastewater at 32 g/L salt
Time COD DO COD% U MLSS
h mg/L mg/L g/g.d mg/L
0.0 4950 9350
2.5 2750 0.7 45 2.25
5.0 1505 0.9 69.9 1.75
8 540 4.2 89.2 1.49
9.0 255 6.3 94.9 1.32
10.5 210 6.2 95.8 1.14 9645
13 245 6.3 95.1 0.91
Average 9500
Table B-15 COD and DO profile data of the mixed yeast batch culture with glucose-feed
wastewater at 45 g/L salt
Time COD DO COD% U MLSS
h mg/L mg/L g/g.d mg/L
0.0 5050 9360
2.5 4200 0.7 16 0.80
5.0 2740 0.9 45 1.12
7.5 2000 4.2 60 0.99
9.0 1560 6.3 69 0.95
10.5 790 6.2 84 1.00 10130
13 290 6.3 94 0.90
Average 9750
B-7
Table B-16 COD and DO profile data of the mixed bacterial batch culture with glucose-feed
wastewater at 20 g/L salt
Time COD DO COD% U MLSS
h mg/L mg/L g/g.d mg/L
0 1036 2752
0.25 821 1.97 18 5.84
1.25 146 6.3 85 5.57
2.5 20 6.3 98 3.20
3.5 26 6.5 97 2.27
6.5 45 6.4 96 1.20 3351
Average 3050
Table B-17 COD and DO profile data of the mixed bacterial batch culture with glucose-feed
wastewater at 32 g/L salt
Time COD DO COD% U MLSS
h mg/L mg/L g/g.d mg/L
0 1020 3245
0.25 760 2.1 24 6.44
1.25 320 4.95 68 3.65
2.25 235 5.7 77 2.28
3.5 160 6.2 84 1.61
6.5 45 6.3 96 0.99
8 30 6.4 97 0.81 4053
11.5 35 6.3 97 0.56
14.5 25 6.4 98 0.45
Average 3650
Table B-18 COD and DO profile data of the mixed bacterial batch culture with glucose-feed
wastewater at 45 g/L salt
Time COD DO COD% U MLSS
h mg/L mg/L g/g.d mg/L
0.25 945 0.9 6 1.65
1.25 836 1.2 16 0.98
2.25 769 2.3 23 0.77
3.5 578 3.4 42 0.90
6.5 405 4.7 60 0.69
8.5 259 5.4 74 0.65 3450
11.5 218 5.9 78 0.51
14.5 113 6.1 89 0.46
17 70 6.2 93 0.41
Average 3200
B-8
Table B-19 COD and DO profile data of the mixed yeast batch culture with fish-protein-feed
wastewater at 20 g/L salt
Time COD DO COD% U MLSS
h mg/L mg/L g/g.d mg/L
0.0 5010 5432
2.5 0.6
5.5 3180 0.7 36.5 1.32
7.5 2800 0.7 44.1 1.17
10.5 2490 0.6 50.3 0.95
16.0 2191 0.7 56.3 0.70
18.5 1650 0.6 67.1 0.72
20.5 1500 0.6 70.1 0.68
26.0 0.6
29.0 1010 2.2 79.8 0.55
31.0 790 4.5 84.2 0.54
34.0 680 4.9 86.4 0.51 6670
37.0 490 5.1 90.2 0.48
42.0 540 5.6 89.2 0.42
Average 6050
Table B-20 COD and DO profile data of the mixed yeast batch culture with protein-feed
wastewater at 32 g/L salt
Time COD DO COD% U MLSS
h mg/L mg/L g/g.d mg/L
0.0 5010 5325
5.5 3234 0.7 35.4 1.33
7.5 2707 0.7 46.0 1.27
10.5 2392 0.6 52.3 1.03
16.0 2191 0.7 56.3 0.73
18.5 2031 0.6 59.5 0.67
20.5 1913 0.6 61.8 0.62
26.0 0.6
29.0 968 2.2 80.7 0.58
32.0 772 4.0 84.6 0.55
34.0 830 4.2 83.4 0.51 6302
37.0 584 4.3 88.3 0.49
42.0 541 5.6 89.2 0.44
Average 5810
B-9
Table B-21 COD and DO profile data of the mixed yeast batch culture with fish-protein-feed
wastewater at 45 g/L salt
Time COD DO COD% U MLSS
h mg/L mg/L g/g.d mg/L
0.0 5010 5915
2.5 0.3
5.5 4030 0.2 19.6 0.67
7.5 3830 0.4 23.6 0.59
10.5 3690 0.4 26.3 0.47
16.0 3230 0.5 35.5 0.42
18.5 2700 0.4 46.1 0.47
20.5 2590 0.5 48.3 0.44
26.0 0.5
29.0 1500 0.8 70.1 0.45
32.0 1330 1.1 73.5 0.43
34.0 1240 2.5 75.2 0.41 6946
37.0 950 3.9 81.0 0.41
42.0 920 4.1 81.6 0.36
Average 6430
Table B-22 COD and DO profile data of the mixed bacterial batch culture with fish-protein-
feed wastewater at 20 g/L salt
Time COD DO COD% U MLSS
h mg/L mg/L g/g.d mg/L
0 1020
0.25 805 1.9 21 4.80 3730
1.25 619 2.1 39 1.79
2.25 416 4.2 59 1.50
3.5 300 4.5 71 1.15
6.5 143 5.9 86 0.75
9.0 40 6.5 96 0.61 4870
11.5 25 6.4 98 0.48
14.5 35 6.5 97 0.38
Average 4300
B-10
Table B-23 COD and DO profile data of the mixed yeast batch culture with fish-protein-feed
wastewater at 32 g/L salt
Time COD DO COD% U MLSS
h mg/L mg/L g/g.d mg/L
0 976 3270
0.25 870 1.6 11 2.80
1.25 645 1.8 34 1.75
2.25 568 2.5 42 1.20
3.5 509 4.0 48 0.88
6.5 383 5.7 61 0.60
9 320 6.0 67 0.48
11.5 200 6.2 80 0.45
14.5 110 6.3 89 0.39
18.5 65 6.2 93 0.33
21.0 50 6.3 95 0.29 4172
26.0 45 6.2 95 0.24
28.0 20 6.2 98 0.23
Average 3721
Table B-24 COD and DO profile data of the mixed yeast batch culture with fish-protein-feed
wastewater at 45 g/L salt
Time COD DO COD% U MLSS
h mg/L mg/L g/g.d mg/L
0 985 3650
0.25 935 0.7 5 1.17
1.25 941 0.9 4 0.21
2.25 888 1.5 10 0.25
3.5 760 1.5 23 0.37
6.5 691 1.6 30 0.26
9.0 534 2.4 46 0.29
11.5 473 3.8 52 0.26
14.5 337 5.5 66 0.26
18.5 298 5.7 70 0.22
21.0 206 6 79 0.22
26.0 127 5.7 87 0.19
28.0 90 6.2 91 0.19 4589
30.0 75 6.3 92 0.21
4120
Salt 20
Yeast
Time CODin CODeff COD%y MLSSy F/M
0 4930 5450
2 5120 1587 69 5400 0.63
4 5210 1667 68 5870 0.59
6 5020 1456 71 6780 0.49
8 4970 1243 75 7920 0.42
10 4760 809 83 9250 0.34
12 5010 752 85 9350 0.36
14 5150 567 89 10850 0.32
16 5090 458 91 10750 0.32
Bacteria
Time CODin CODeff COD% MLSS F/M
0 3120
2 1020 92 91 3650 0.28
4 990 79 92 4270 0.23
6 1070 86 92 4750 0.23
8 1120 56 95 5120 0.22
10 1010 51 95 5250 0.19
12 980 29 97 5430 0.18
14 990 50 95 5910 0.17
16 1130 45 96 5670 0.20
Salt 32
Yeast
Time CODin CODeff COD%y MLSSy F/M
0 5010 5700
2 4850 1795 63 5400 0.60
4 5120 1640 68 5990 0.57
6 4950 1730 65 5930 0.56
8 4760 1190 75 6950 0.46
10 5230 1200 77 7050 0.49
12 5100 870 83 8345 0.41
14 5010 701 86 9050 0.37
16 5120 770 85 9710 0.35
Bacteria
Time CODin CODeff COD% MLSS F/M
0 3560
2 1100 165 85 3750 0.29
4 980 78 92 4340 0.23
6 995 90 91 4950 0.20
8 1020 51 95 5120 0.20
10 1050 73 93 5010 0.21
12 950 38 96 5700 0.17
14 970 78 92 5600 0.17
16 1105 55 95 6310 0.18
Salt 45
Yeast
Time CODin CODeff COD%y MLSSy F/M
0 4790 5500
2 5100 2091 59 5200 0.65
4 5020 1640 61 5350 0.63
6 4970 1730 61 5320 0.62
8 5120 1587 69 6240 0.55
10 5070 1200 72 8705 0.39
12 5025 870 83 8956 0.37
14 4790 701 84 9310 0.34
16 4990 770 83 9420 0.35
Bacteria
Time CODin CODeff COD% MLSS F/M
0 3950
2 1010 222 78 3650 0.28
4 1200 228 81 3750 0.32
6 1105 221 80 4210 0.26
8 1030 175 83 4760 0.22
10 970 175 82 4790 0.20
12 990 109 89 5700 0.17
14 1025 113 89 5970 0.17
16 990 89 91 6105 0.16
CO
DB
f
0
200
400
600
800
1000
1200
05
10
15
20
25
30
35
CO
DB
f
S2
0
Tim
e5000
CO
Dy-p
rD
Oy-p
r%
CO
DU
MLS
S
0.0
5010
54
32
2.5
0.6
5.5
3180
0.7
36
.51.3
2
7.5
2800
0.7
44
.11.1
7
9.0
10.5
2490
0.6
50
.30.9
5
12.5
16.0
2191
0.7
56
.30.7
0
18.5
1650
0.6
67
.10.7
2
20.5
1500
0.6
70
.10.6
8
26.0
0.6
29.0
1010
2.2
79
.80.5
5
31.0
790
4.5
84
.20.5
4
34.0
680
4.9
86
.40.5
16670
37.0
490
5.1
90
.20.4
8
42.0
540
5.6
89
.20.4
2
6050
S3
2
Tim
e5000
CO
Dy-p
rD
Oy-p
rC
OD
%U
MLS
S
0.0
5010
53
25
2.5
0.6
5.5
3234
0.7
35
.41.3
3
7.5
2707
0.7
46
.01.2
7
9.0
10.5
2392
0.6
52
.31.0
3
12.5
16.0
2191
0.7
56
.30.7
3
18.5
2031
0.6
59
.50.6
7
20.5
1913
0.6
61
.80.6
2
26.0
0.6
29.0
968
2.2
80
.70.5
8
32.0
772
4.0
84
.60.5
5
34.0
830
4.2
83
.40.5
16302
37.0
584
4.3
88
.30.4
9
42.0
541
5.6
89
.20.4
4
5810
S4
5
Tim
e5000
CO
Dy-p
rD
Oy-p
rC
OD
%U
MLS
S
0.0
5010
59
15
2.5
0.3
5.5
4030
0.2
19
.60.6
7
7.5
3830
0.4
23
.60.5
9
9.0
10.5
3690
0.4
26
.30.4
7
12.5
16.0
3230
0.5
35
.50.4
2
18.5
2700
0.4
46
.10.4
7
20.5
2590
0.5
48
.30.4
4
26.0
0.5
29.0
1500
0.8
70
.10.4
5
32.0
1330
1.1
73
.50.4
3
34.0
1240
2.5
75
.20.4
16946
37.0
950
3.9
81
.00.4
1
42.0
920
4.1
81
.60.3
6
6430
S2
0
Tim
e5000
CO
Dy-p
rD
Oy-p
r%
CO
DU
MLS
S
0.0
5010
54
32
2.5
0.6
5.5
3180
0.7
36
.51.3
2
7.5
2800
0.7
44
.11.1
7
9.0
10.5
2490
0.6
50
.30.9
5
12.5
16.0
2191
0.7
56
.30.7
0
18.5
1650
0.6
67
.10.7
2
20.5
1500
0.6
70
.10.6
8
26.0
0.6
29.0
1010
2.2
79
.80.5
5
31.0
790
4.5
84
.20.5
4
34.0
680
4.9
86
.40.5
16670
37.0
490
5.1
90
.20.4
8
42.0
540
5.6
89
.20.4
2
6050
S3
2
Tim
e5000
CO
Dy-p
rD
Oy-p
rC
OD
%U
MLS
S
0.0
5010
53
25
2.5
0.6
5.5
3234
0.7
35
.41.3
3
7.5
2707
0.7
46
.01.2
7
9.0
10.5
2392
0.6
52
.31.0
3
12.5
16.0
2191
0.7
56
.30.7
3
18.5
2031
0.6
59
.50.6
7
20.5
1913
0.6
61
.80.6
2
26.0
0.6
29.0
968
2.2
80
.70.5
8
32.0
772
4.0
84
.60.5
5
34.0
830
4.2
83
.40.5
16302
37.0
584
4.3
88
.30.4
9
42.0
541
5.6
89
.20.4
4
5810
S4
5
Tim
e5000
CO
Dy-p
rD
Oy-p
rC
OD
%U
MLS
S
0.0
5010
59
15
2.5
0.3
5.5
4030
0.2
19
.60.6
7
7.5
3830
0.4
23
.60.5
9
9.0
10.5
3690
0.4
26
.30.4
7
12.5
16.0
3230
0.5
35
.50.4
2
18.5
2700
0.4
46
.10.4
7
20.5
2590
0.5
48
.30.4
4
26.0
0.5
29.0
1500
0.8
70
.10.4
5
32.0
1330
1.1
73
.50.4
3
34.0
1240
2.5
75
.20.4
16946
37.0
950
3.9
81
.00.4
1
42.0
920
4.1
81
.60.3
6
6430
S20
S45
Tim
eC
OD
Bf
DO
Bf
CO
D%
UM
LS
ST
ime
CO
DB
fD
OB
fC
OD
%U
ML
SS
01
02
00
98
53650
0.2
58
05
1.9
21
4.8
03730
0.2
59
35
0.7
51
.17
1.2
56
19
2.1
39
1.7
91
.25
94
10
.94
0.2
1
2.2
54
16
4.2
59
1.5
02
.25
88
81
.510
0.2
5
3.5
30
04
.57
11
.15
3.5
76
01
.523
0.3
7
6.5
14
35
.98
60
.75
6.5
69
11
.630
0.2
6
94
06
.59
60
.61
4870
95
34
2.4
46
0.2
9
11
.52
56
.49
80
.48
11
.54
73
3.8
52
0.2
6
14
.53
56
.59
70
.38
14
.53
37
5.5
66
0.2
6
4300
4300
18
.52
98
5.7
70
0.2
2
21
.02
06
679
0.2
2
26
.01
27
5.7
87
0.1
9
28
.09
06
.291
0.1
94589
30
.07
56
.392
0.2
1
Tim
eC
OD
Bf
DO
Bf
CO
D%
UM
LS
S4120
09
76
3270
0.2
58
70
1.6
11
2.8
0
1.2
56
45
1.8
34
1.7
5
2.2
55
68
2.5
42
1.2
0
3.5
50
94
.04
80
.88
6.5
38
35
.76
10
.60
93
20
6.0
67
0.4
8
11
.52
00
6.2
80
0.4
5
14
.51
10
6.3
89
0.3
9
18
.56
56
.29
30
.33
21
.05
06
.39
50
.29
4172
26
.04
56
.29
50
.24
28
.02
06
.29
80
.23
3721
20
g/L
Tim
eC
OD
B2
DO
B2
CO
D%
UM
LS
S2
0 g
/L3
2 g
/L4
5 g
/L
01036
27
52
30
50
36
50
32
00
0.2
5821
1.9
71
85.8
42
.58
17
1.2
5146
6.3
85
5.5
73
.27
0.8
40
.44
2.5
20
6.3
98
3.2
02
03
07
0
3.5
26
6.5
97
2.2
79
89
79
3
6.5
45
6.4
96
1.2
03
.20
.81
0.4
1
8.5
15
6.4
99
0.9
43351
30
50
32
g/L
Tim
eC
OD
B2
DO
B2
CO
D%
UM
LS
S
01020
32
45
0.2
5760
2.1
24
6.4
4
1.2
5320
4.9
56
83.6
5
2.2
5235
5.7
77
2.2
8
3.5
160
6.2
84
1.6
1
6.5
45
6.3
96
0.9
9
830
6.4
97
0.8
14053
11.5
35
6.3
97
0.5
6
14.5
25
6.4
98
0.4
5
36
50
45
g/L
Tim
eC
OD
B2
DO
B2
CO
D%
UM
LS
S
01000
29
50
0.2
5945
0.9
61.6
5
1.2
5836
1.2
16
0.9
8
2.2
5769
2.3
23
0.7
7
3.5
578
3.4
42
0.9
0
6.5
405
4.7
60
0.6
9
8.5
259
5.4
74
0.6
53450
11.5
218
5.9
78
0.5
1
14.5
113
6.1
89
0.4
6
17
70
6.2
93
0.4
1
32
00
C-1
Tab
le C
-1
Bio
kin
etic
exper
imen
tal
dat
a of
mix
ed y
east
slu
dge
wit
h g
luco
se-f
eed w
aste
wat
er a
t 2
0 g
/L s
alt
Sm
g/L
CO
D
Rx
,t
mg O
2/g
VS
S.h
Rx
,e
mg O
2/g
VS
S. h
OC
m
g/L
R
x,o
x
mg O
2/g
VS
S.d
OC
/S
Rx
mg
CO
D/g
VS
S.d
YC
OD
g C
OD
/g C
OD
Yv
ss
g V
S/g
CO
D
day
-1
10
10.4
2
.74
3.3
3
7.6
1
0.3
36
21.8
0.6
64
0.4
68
0.2
4
15
16.1
3
.09
5.2
7
13
0.3
55
37.3
0.6
45
0.4
54
0.4
1
20
24.3
2
.73
7.1
7
22
0.3
62
61.8
0.6
38
0.4
49
0.6
8
30
38.5
3
.88
10
.2
35
0.3
42
99.1
0.6
58
0.4
63
1.0
9
50
57.0
3
.99
16
.9
53
0.3
41
152
0.6
59
0.4
64
1.6
7
100
60.7
5
.16
35
.6
56
0.3
60
159
0.6
40
0.4
51
1.7
5
200
98.0
4
.47
94
268
2.9
5
300
131.4
3
.55
128
366
4.0
3
500
136.1
4
.43
132
377
4.1
5
A
ver
age
Y
0.4
57
Tab
le C
-2
Bio
kin
etic
exper
imen
tal
dat
a of
mix
ed y
east
slu
dge
wit
h g
luco
se-f
eed w
aste
wat
er a
t 3
2 g
/L s
alt
Sm
g/L
CO
D
Rx
,t
mg O
2/g
VS
S.h
Rx
,e
mg O
2/g
VS
S. h
OC
m
g/L
R
x,o
x
mg O
2/g
VS
S.d
OC
/S
Rx
mg
CO
D/g
VS
S.d
YC
OD
g C
OD
/g C
OD
Yv
ss
g V
S/g
CO
D
day
-1
10
8.1
2.8
1
3.1
0
5.2
0.3
13
16
0.6
87
0.4
84
0.1
9
15
12.8
2.8
9
4.8
3
9.9
0.3
25
31
0.6
75
0.4
75
0.3
6
20
23.1
2.7
2
6.1
2
20
0.3
09
64
0.6
91
0.4
87
0.7
4
30
25.1
2.7
9
9.5
0
22
0.3
20
70
0.6
80
0.4
79
0.8
1
50
42.9
5.1
2
15
.69
38
0.3
17
119
0.6
83
0.4
81
1.3
7
200
95
5.4
3
63
.56
90
0.3
21
282
0.6
79
0.4
78
3.2
5
300
103
5.2
9
98
308
3.5
5
500
102
3.6
0
98
309
3.5
6
A
ver
age
Y
0.4
80
C-2
Tab
le C
-3
Bio
kin
etic
exper
imen
tal
dat
a of
mix
ed y
east
slu
dge
wit
h g
luco
se-f
eed w
aste
wat
er a
t 4
5 g
/L s
alt
Sm
g/L
CO
D
Rx
,t
mg O
2/g
VS
S.h
Rx
,e
mg O
2/g
VS
S. h
OC
m
g/L
R
x,o
x
mg O
2/g
VS
S.d
OC
/S
Rx
mg
CO
D/g
VS
S.d
YC
OD
g C
OD
/g C
OD
Yv
ss
g V
S/g
CO
D
day
-1
10
9.2
2.8
2
4.2
3
6.4
0.4
27
15.2
0.5
73
0.4
04
0.1
5
30
28.1
2.6
3
12
.62
25
0.4
25
61.0
0.5
75
0.4
05
0.6
0
50
28.4
2.0
5
20
.00
26
0.4
04
63.0
0.5
96
0.4
20
0.6
2
100
57.5
2.7
3
41
.48
55
0.4
19
131.1
0.5
81
0.4
09
1.2
9
200
64.2
2.6
2
82
.17
62
0.4
15
147.4
0.5
85
0.4
12
1.4
5
300
90.1
2.5
7
88
209.4
2.0
6
500
91.8
2.1
1
90
214.5
2.1
1
A
ver
age
Y
0.4
11
0100
200
300
400
500
CO
D c
oncentration S
(m
g/L C
OD
)
0.0
1.0
2.0
3.0
4.0
5.0
Specific Growth Rate day-1
20 g
NaCl/
L
32 g
NaCl/
L
45 g
NaCl/
L
Fig
. C
-1
V
aria
tion o
f sp
ecif
ic g
row
th r
ate
of
yea
st s
ludg
e ver
sus
CO
D c
on
cen
trat
ion
at d
iffe
rent
salt
conte
nts
for
glu
cose
-fee
dw
aste
wat
er
20 g
/L N
aCl:
S
S
158
60
.5
R2=
0.9
71
32 g
/L N
aCl:
S
S
118
74
.4 R2=
0.9
82
45 g
/L N
aCl:
S
S
129
70
.2 R2=
0.9
67
C-3
Tab
le C
-4
Bio
kin
etic
exper
imen
tal
dat
a of
mix
ed b
acte
rial
slu
dge
wit
h g
luco
se-f
eed w
aste
wat
er a
t 20 g
/L s
alt
Sm
g/L
CO
D
Rx
,t
mg O
2/g
VS
S.h
Rx
,e
mg O
2/g
VS
S. h
OC
m
g/L
R
x,o
x
mg O
2/g
VS
S.d
OC
/S
Rx
mg
CO
D/g
VS
S.d
YC
OD
g C
OD
/g C
OD
Yv
ss
g V
S/g
CO
D
day
-1
5
17
2.6
4
1.0
4
14
0.2
10
75
0.7
90
0.5
56
1.0
3
7
22
2.4
9
1.4
2
19
0.2
05
101
0.7
95
0.5
60
1.3
8
10
24
2.4
8
1.7
9
21
0.1
81
112
0.8
19
0.5
77
1.5
4
30
48
2.8
8
5.0
5
46
0.1
70
240
0.8
30
0.5
85
3.2
8
50
84
3.1
5
8.9
1
81
0.1
80
427
0.8
20
0.5
77
5.8
5
100
109
3.7
9
18
.91
105
0.1
91
555
0.8
09
0.5
70
7.6
0
200
109
3.6
9
105
554
7.5
8
A
ver
age
Y
0.5
70
Tab
le C
-5
Bio
kin
etic
exper
imen
tal
dat
a of
mix
ed b
acte
rial
slu
dge
wit
h g
luco
se-f
eed w
aste
wat
er a
t 32 g
/L s
alt
Sm
g/L
CO
D
Rx
,t
mg O
2/g
VS
S.h
Rx
,e
mg O
2/g
VS
S. h
OC
m
g/L
R
x,o
x
mg O
2/g
VS
S.d
OC
/S
Rx
mg
CO
D/g
VS
S.d
YC
OD
g C
OD
/g C
OD
Yv
ss
g V
S/g
CO
D
day
-1
5
7.2
4
.79
0.9
2
2.4
2
0.1
85
13.7
0.8
15
0.5
74
0.1
9
10
10.9
4
.14
1.7
9
6.7
4
0.1
81
38.1
0.8
19
0.5
77
0.5
3
15
11.6
3
.82
2.6
1
7.7
6
0.1
76
43.9
0.8
24
0.5
80
0.6
1
20
12.9
3
.72
3.5
0
9.1
6
0.1
77
51.8
0.8
23
0.5
80
0.7
2
30
14.7
3
.64
5.1
1
11
.07
0.1
72
62.5
0.8
28
0.5
83
0.8
7
50
22.0
3
.90
8.3
7
18
.07
0.1
69
102.1
0.8
31
0.5
85
1.4
2
100
30.5
4
.20
0.0
0
26
.34
148.8
1.0
00
0.7
04
2.0
7
200
31.1
4
.62
26
.47
149.5
2.0
8
A
ver
age
Y
0.5
83
C-4
Tab
le C
-6
Bio
kin
etic
exper
imen
tal
dat
a of
mix
ed b
acte
rial
slu
dge
wit
h g
luco
se-f
eed w
aste
wat
er a
t 45 g
/L s
alt
Sm
g/L
CO
D
Rx
,t
mg O
2/g
VS
S.h
Rx
,e
mg O
2/g
VS
S. h
OC
m
g/L
R
x,o
x
mg O
2/g
VS
S.d
OC
/S
Rx
mg
CO
D/g
VS
S.d
YC
OD
g C
OD
/g C
OD
Yv
ss
g V
S/g
CO
D
day
-1
10
6.8
3
.85
2.5
8
2.9
1
0.2
61
11.8
0.7
39
0.5
20
0.1
5
20
10.9
3
.75
5.0
9
7.1
8
0.2
57
29.1
0.7
43
0.5
23
0.3
7
30
12.2
4
.61
6.8
9
7.5
7
0.2
32
30.6
0.7
68
0.5
41
0.3
9
50
13.6
4
.05
11
.93
9.5
1
0.2
41
38.5
0.7
59
0.5
35
0.4
9
100
20.5
4
.00
24
.26
16
.50
0.2
45
66.8
0.7
55
0.5
32
0.8
5
200
20.8
4
.10
16
.69
67.6
0.8
6
A
ver
age
Y
0.5
31
040
80
120
160
200
CO
D c
oncentration (
mg/L C
OD
)
0.0
2.0
4.0
6.0
8.0
Specific Growth Rate day-1
20 g
NaCl/
L
32 g
NaCl/
L
45 g
NaCl/
L
Fig
. C
-2
V
aria
tion
of
spec
ific
gro
wth
ra
te
of
mix
ed
bac
teri
al
sludg
e v
ersu
s C
OD
con
cen
trat
ion a
t d
iffe
ren
t sa
lt c
on
ten
ts f
or
glu
cose
-fee
d w
aste
wat
er
20 g
/L N
aCl:
S
S
44
95
.9
R2=
0.9
65
32 g
/L N
aCl:
S
S
52
80
.2
R2=
0.9
69
45 g
/L N
aCl:
S
S
53
14
.1
R2=
0.9
47
C-5
Tab
le C
-7
Bio
kin
etic
exper
imen
tal
dat
a of
mix
ed y
east
slu
dge
wit
h f
ish-p
rote
in-f
eed
was
tew
ater
at
20 g
/L s
alt
Sm
g/L
CO
D
Rx
,t
mg O
2/g
VS
S.h
Rx
,e
mg O
2/g
VS
S. h
OC
m
g/L
R
x,o
x
mg O
2/g
VS
S.d
OC
/S
Rx
mg
CO
D/g
VS
S.d
YC
OD
g C
OD
/g C
OD
Yv
ss
g V
S/g
CO
D
day
-1
20
18
1.3
0
6.8
16
0.4
35
41.8
0.5
65
0.3
98
0.4
3
50
44
1.2
6
16
.0
43
0.4
10
110.4
0.5
90
0.4
15
1.1
4
100
49
1.4
6
28
.9
47
0.3
71
121.1
0.6
29
0.4
43
1.2
5
300
118
1.3
0
80
.7
116
0.3
45
298.3
0.6
55
0.4
61
3.0
8
500
123
1.4
6
121
311.4
3.2
1
Tab
le C
-8
Bio
kin
etic
exper
imen
tal
dat
a of
mix
ed y
east
slu
dge
wit
h f
ish-p
rote
in-f
eed
was
tew
ater
at
32 g
/L s
alt
Sm
g/L
CO
D
Rx
,t
mg O
2/g
VS
S.h
Rx
,e
mg O
2/g
VS
S. h
OC
m
g/L
R
x,o
x
mg O
2/g
VS
S.d
OC
/S
Rx
mg
CO
D/g
VS
S.d
YC
OD
g C
OD
/g C
OD
Yv
ss
g V
S/g
CO
D
day
-1
45
14
3.8
1
14
.9
10
0.4
25
26.3
0.5
75
0.4
05
0.2
8
100
37
4.2
1
30
.5
32
0.3
91
87.1
0.6
09
0.4
29
0.9
2
200
47
3.1
0
54
.8
44
0.3
51
117.7
0.6
49
0.4
57
1.2
5
300
75
4.2
0
74
.6
71
0.3
19
189.6
0.6
81
0.4
80
2.0
1
500
77
4.1
0
73
196.9
2.0
9
C-6
Tab
le C
-9
Bio
kin
etic
exper
imen
tal
dat
a of
mix
ed y
east
slu
dge
wit
h f
ish-p
rote
in-f
eed
was
tew
ater
at
45 g
/L s
alt
Sm
g/L
CO
D
Rx
,t
mg O
2/g
VS
S.h
Rx
,e
mg O
2/g
VS
S. h
OC
m
g/L
R
x,o
x
mg O
2/g
VS
S.d
OC
/S
Rx
mg
CO
D/g
VS
S.d
YC
OD
g C
OD
/g C
OD
Yv
ss
g V
S/g
CO
D
day
-1
20
11
3.5
1
7.4
7
0.4
72
16.6
0.5
28
0.3
72
0.1
6
50
13
3.5
1
17
.6
9
0.4
52
21.8
0.5
48
0.3
86
0.2
1
100
41
3.2
9
31
.6
38
0.4
05
88.2
0.5
95
0.4
19
0.8
5
300
73
3.2
9
91
.5
70
0.3
91
163.0
0.6
09
0.4
29
1.5
7
500
70
70
163.0
1.5
7
A
ver
age
Y
0.4
03
0100
200
300
400
500
CO
D c
oncentration S
(m
g/L
CO
D)
0.0
1.0
2.0
3.0
4.0
Specific Growth Rate day-1
20 g
NaCl/
L
32 g
NaCl/
L
45 g
NaCl/
L
Fig
. C
-3 V
aria
tion o
f sp
ecif
ic g
row
th r
ate
of
yea
st s
ludg
e ver
sus
CO
D c
on
cen
trat
ion
at d
iffe
ren
t sa
lt c
on
ten
ts f
or
fish
-pro
tein
-fee
d w
aste
wat
er
15
g/L
NaC
l:
S
S
201
69
.4 R2=
0.9
64
32
g/L
NaC
l:
S
S
322
62
.3 R
2=
0.9
42
45
g/L
NaC
l:
S
S
228
46
.2 R
2=
0.9
43
C-7
Tab
le C
-10
B
iok
inet
ic e
xp
erim
enta
l d
ata
of
mix
ed b
acte
rial
slu
dg
e w
ith
fis
h-p
rote
in-f
eed
was
tew
ater
at
20
g/L
sal
t
Sm
g/L
CO
D
Rx
,t
mg O
2/g
VS
S.h
Rx
,e
mg O
2/g
VS
S. h
OC
m
g/L
R
x,o
x
mg O
2/g
VS
S.d
OC
/S
Rx
mg
CO
D/g
VS
S.d
YC
OD
g C
OD
/g C
OD
Yv
ss
g V
S/g
CO
D
day
-1
20
100
3.7
3
7.2
96
0.4
61
223.2
0.5
39
0.3
80
2.1
5
50
148
3.4
3
17
.4
145
0.4
45
337.4
0.5
55
0.3
91
3.2
5
100
196
4.1
6
32
.1
192
0.4
11
446.3
0.5
89
0.4
15
4.3
0
150
221
4.2
6
47
.0
216
0.4
02
503.4
0.5
98
0.4
21
4.8
5
300
220
220
510.7
4.9
2
A
ver
age
Y
0.4
02
Tab
le C
-11
B
iok
inet
ic e
xp
erim
enta
l d
ata
of
mix
ed b
acte
rial
slu
dg
e w
ith
fis
h-p
rote
in-f
eed
was
tew
ater
at
32
g/L
sal
t
Sm
g/L
CO
D
Rx
,t
mg O
2/g
VS
S.h
Rx
,e
mg O
2/g
VS
S. h
OC
m
g/L
R
x,o
x
mg O
2/g
VS
S.d
OC
/S
Rx
mg
CO
D/g
VS
S.d
YC
OD
g C
OD
/g C
OD
Yv
ss
g V
S/g
CO
D
day
-1
20
14.1
3
.81
5.5
10
0.3
52
31.0
0.6
48
0.4
56
0.3
5
36
18.9
2
.82
9.7
16
0.3
45
48.6
0.6
55
0.4
61
0.5
5
100
34.7
3
.66
25
.0
31
0.3
21
93.7
0.6
79
0.4
78
1.0
6
200
37.8
2
.70
47
.0
35
0.3
01
106.1
0.6
99
0.4
92
1.2
0
300
50.2
4.3
46
138.9
1.5
7
A
ver
age
Y
0.4
71
C-8
Tab
le C
-12
B
iok
inet
ic e
xp
erim
enta
l d
ata
of
mix
ed b
acte
rial
slu
dg
e w
ith
fis
h-p
rote
in-f
eed
was
tew
ater
at
45
g/L
sal
t
Sm
g/L
CO
D
Rx
,t
mg O
2/g
VS
S.h
Rx
,e
mg O
2/g
VS
S. h
OC
m
g/L
R
x,o
x
mg O
2/g
VS
S.d
OC
/S
Rx
mg
CO
D/g
VS
S.d
YC
OD
g C
OD
/g C
OD
Yv
ss
g V
S/g
CO
D
day
-1
20
14.8
4
.06
7.0
11
0.4
49
24.9
0.5
51
0.3
88
0.2
4
50
23.7
3
.65
17
.1
20
0.4
38
46.7
0.5
62
0.3
96
0.4
5
100
33.3
2
.94
32
.4
30
0.4
15
70.6
0.5
85
0.4
12
0.6
8
150
43.2
3
.93
48
.9
39
0.4
18
91.3
0.5
82
0.4
10
0.8
8
300
41.8
3.9
38
88.2
0.8
5
A
ver
age
Y
0.4
01
0100
200
300
CO
D c
oncentration S
(m
g/L C
OD
)
0.0
1.0
2.0
3.0
4.0
5.0
Specific Growth Rate day-1
20 g
NaCl/
L
32 g
NaCl/
L
45 g
NaCl/
L
Fig
. C
-4
Var
iati
on
of
spec
ific
gro
wth
ra
te
of
mix
ed
bac
teri
al
sludge
ver
sus
CO
D
con
cen
trat
ion a
t d
iffe
ren
t sa
lt c
on
ten
ts f
or
fish
-pro
tein
-fee
d w
aste
wat
er
15 g
/L N
aCl:
S
S
33
65
.5
R2=
0.9
82
32 g
/L N
aCl:
S
S
93
95
.1
R2=
0.9
73
45 g
/L N
aCl:
S
S
64
11
.1
R2=
0.9
44
D-1
Table D-1 Profile data of pH and nitrogen components of mixed yeast culture with glucose-
feed wastewater at 32 g salt/L (mean MLSS = 9500 mg/L)
Time pH COD NH 3-N NO 3+NO 2- N Total-N N removal
h mg/L mg/L mg/L mg/L %
0 4950 365 365
0.25 3.50
1.25 2.88 1200 317 N/A 317 13.2
2.50 2.72 2750 267 2.13 269 26.3
3.50 2.64 N/A 237 2.00 239 34.5
6.50 2.60 950 172 2.01 174 52.3
9.00 2.65 255 139 2.05 141 61.4
11.50 2.71 210 127 2.00 129 64.7
14.50 2.80 245 117 2.20 119 67.3
17.00 2.85 N/A 112 N/A 112 69.3
Table D-2 Profile data of pH and nitrogen components of mixed bacterial culture with
glucose-feed wastewater at 32 g salt/L (mean MLSS = 3650 mg/L)
Time pH COD NH 3-N NO 3+NO 2- N Total-N N removal
h mg/L mg/L mg/L mg/L %
0 1020 53 0 53
0.25 7.50 760
1.25 7.76 320
2.25 8.10 235 43 0.70 44 17.5
3.50 8.26 160 32 1.74 34 36.3
6.50 8.28 45 20 1.72 21 59.8
8.00 8.31 30 17 1.85 19 64.8
11.50 8.28 35 14 1.90 16 70.0
14.50 8.25 25 14 2.50 17 68.9
D-2
Table D-3 Profile data of pH and nitrogen components of mixed yeast culture with fish-
protein-feed wastewater at 32 g salt/L (mean MLSS = 5810 mg/L)
Time COD pH Adjusted
pHOrganic-N NH 3-N
NO 3+NO 2-
NTotal N N removal
h mg/L mg/L mg/L mg/L mg/L %
0.0 5010 745 45 ND 790
2.3 3.57 3.50 560
5.5 3234 4.26 3.47 63 ND
7.5 2707 3.78 3.45 458 97 554 37.8
10.5 2392 4.06 3.52 127 ND
16.0 2191 4.79 3.51 393 161 554 37.8
18.5 2031 4.00 3.46 209 ND
20.5 1913 4.14 3.50 290 234 524 41.3
26.0 4.37 3.56
29.0 968 4.09 3.50 109 329 1.20 438 50.9
32.0 772 4.21 3.52 31 396 1.50 427 52.1
34.0 830 4.63 3.50
38.0 584 4.47 3.41 424 2.10
42.0 541 3.99 3.46 16 430 2.30 446 50.0
ND- None detected
Table D-4 Profile data of pH and nitrogen components of mixed bacterial culture with fish-
protein-feed wastewater at 32 g salt/L (mean MLSS = 3720 mg/L)
Time COD pH Organic-N NH 3-NNO 3+NO 2-
NTotal N N removal
h mg/L mg/L mg/L mg/L mg/L %
0 976 150 49 0 199
0.25 870 7.50
1.25 645 7.93 120 52 0.5 172 13.6
2.25 568 8.20 67 1.09
3.5 509 8.40 75 1.51
6.5 383 8.55 63 97 1.5 161 19.1
8.5 320 8.54 107 2.5
11.5 200 8.53 110 2.40
14.5 110 8.54 126 3.10
18.5 65 143 2.6
21.0 30 8.52 2.1 142 4.9 149 25.1
26.0 45 8.53 137 4.3
28.7 20 5.1 142 4.9 152 23.6
D-3
Table D-5 H2SO4 amount consumed to maintain pH 3.5 of the mixed yeast culture with
fish-protein-wastewater at 32 g salt/L
Time pH Adjusted pH 1N H 2SO4 volume
consumed
Accumulated acid
volume
h mL mL
0.0 0 6.6
2.3 3.57 3.50
5.5 4.26 3.52 0.9 7.5
7.5 3.78 3.47 0.7 8.2
10.5 4.06 3.54 0.9 9.2
16.0 4.79 3.50 1.7 10.8
18.5 4.00 3.46 0.7 11.5
20.5 4.14 3.47 0.7 12.3
26.0 4.37 3.49 0.9 13.2
28.7 4.09 3.55 0.5 13.7
31.7 4.21 3.51 0.6 14.3
34.8 4.63 3.50 0.8 15.2
38.3 4.47 3.41 0.9 16.1
41.8 3.99 3.46 0.5 16.6
Table D-6 Results of optimum pH experiment for mixed yeast sludge cultured with fish-
protein-wastewater at 32 g/L salt (COD dose = 50 mg/L; MLVSS = 1190 mg/L)
pH OUR total OUR endo OUR ox
mgO2/gVSS.h mgO 2/gVSS.h mgO 2/gVSS.h
2.5 4.2 2.3 2.0
3.0 13.9 4.2 9.7
4.0 14.9 4.2 10.7
5.1 16.1 3.8 12.4
5.5 16.3 3.8 12.5
6.6 14.7 2.9 11.8
7.5 14.9 2.7 12.1
7.9 13.4 3.0 10.4
8.7 11.7 2.5 9.2
9.1 8.1 2.2 5.8
D-4
Table D-7 Optimum pH experiment for mixed bacterial sludge cultured with fish-protein-
wastewater at 32 g/L salt (COD dose = 50 mg/L; MLVSS = 1380 mg/L)
pH OUR total OUR endo OUR ox
mgO2/gVSS.h mgO 2/gVSS.h mgO 2/gVSS.h
4.5 2.8 1.10 1.7
5.3 7.8 2.31 5.5
6.3 14.8 4.14 10.7
7.6 21.3 5.45 15.8
8.1 21.0 5.23 15.3
8.9 21.0 5.88 15.6
9.7 17.7 4.8 12.9
10.1 12.0 3.4 8.6
10.8 7.1 2.2 4.9
10.9 7.5 2.3 5.2
Table D-8 Variation of MLSS during SRT experiments
(Unit: mg/L)
Time SRT
day 5d 7d 10d 20d 45d
0 9000 9000 9000 9000 9000
3 6200 7950 8400 9100 9150
5 4140 7570 8300 8800 10130
7 3640 6800 8000 9300 9600
9 3920 6500 8500 8910 9860
11 2490 6200 8700 9140 9550
13 2770 6550 7640 9400 10270
15 2820 5770 8330 8900 10710
17 2950 5820 7770 9440 10230
19 3230 5110 8150 9740 10320
21 3450 5360 7950 9300 10510
25 3380 5210 8250 9520 10130
D-5
Table D-9 Experimental data of SRT variation runs for mixed yeast batch with fish-protein-
feed wastewater at 32 g/L salt (Inititial COD = 5000 mg/L, HRT = 24 h)
SRT = 5 d SRT = 7 d SRT = 10 d
CODeff TKN effMean
MLSSCODeff TKN eff
Mean
MLSSCODeff TKN eff
Mean
MLSSParameter
mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L mg/L
Batch 1 3120 467 3230 2075 532 5115 1054 495 8150
Batch 2 2655 534 3450 1811 502 5355 1145 497 7950
Batch 3 2776 572 3380 1964 512 5210 952 534 8250
Average 2850 524 3340 1950 515 5330 1050 509 8117
(continuous)
SRT = 20 d SRT = 45 d
CODeff TKN effMean
MLSSCODeff TKN eff
Mean
MLSSParameter
mg/L mg/L mg/L mg/L mg/L mg/L
Bach 1 854 517 9740 937 521 10320
Bach 2 1023 572 9295 925 576 10520
Bach 3 845 515 9525 988 553 10130
Average 907 535 9520 950 550 10320
Table D-10 Result of SRT experiment
SRT 5 7 10 20 45
Volume of wasted sludge, mL/d 400 286 200 100 44
F/M, g COD/g MLSS.d 1.54 0.93 0.62 0.53 0.49
COD removal, % 43 61 77 82 85
N removal, % 7.24 8.90 9.83 5.22 2.67
Amount of SS produced, mg/d 649 765 809 475 228
Amount COD removed, mg/d 2150 3050 3850 4100 4250
Observed yield coefficient 0.302 0.251 0.210 0.116 0.054
E-1
Table E-1 Experimental data for determination of initial membrane resistance of two
membrane modules (A = 0.42 m2; pore size = 0.1 m, temperature = 31.7
oC )
a. Module 1 b. Module 2
Pressure Pressure Flowrate
(L/h)
Flux
(L/m2.h) (mmHg) (kPa)
Flowrate
(L/h)
Flux
(L/m2.h)(mmHg) (kPa)
12.8 30.5 37 4.93 2.6 6.1 6 0.80
10.4 24.6 31 4.13 4.2 10.1 12 1.60
8.9 21.1 27 3.60 7.1 17.0 22 2.93
7.3 17.4 22 2.93 10.0 23.9 31 4.13
5.9 14.1 18 2.40 12.7 30.2 37 4.93
4.3 10.2 12 1.60
2.7 6.3 6 0.80
1.2 2.8 2.5 0.33
Table E-2 Experimental data of YMBR in high COD loading
Time HRT F/M L MLSS CODinf CODeff COD% Time Pressure
(day) (h) g/g.d (kg/m3.d) (mg/L) (mg/L) (mg/L) (%) (day) (kPa)
0 23.7 3650 4830 1691 0 0.3
5 24.2 0.60 4.8 8100 4780 1162 80 4 0.5
11 24.0 0.32 4.8 14200 4870 998 81 8 0.9
13 14100 4950 1105
16 12900 4750 1050
17 23.8 0.39 5.3 13500 5200 1456 83 13 0.0
19 24.4 0.34 5.0 12000 5100 969 85 19 0.8
22 23.5 0.23 3.4 13600 5120 870 86 25 0.9
25 32.3 0.23 3.4 11700 4780 884 86 30 0.8
29 31.7 0.25 3.3 11000 4900 784 84 33 1.0
31 32.2 0.53 6.3 10500 4780 1147 87 37 0.9
35 18.3 0.66 6.5 9890 4850 1261 76 39 0.9
38 17.9 0.69 7.0 10120 5290 1323 74 43 1.2
40 18.2 0.68 6.4 9530 4940 1136 75 46 1.2
43 18.4 0.93 9.7 10440 4882 1528 77 48 1.6
45 12.0 0.95 10.1 10610 5128 1436 71 51 2.0
47 12.2 0.88 9.9 11250 5100 1683 69 53 2.0
51 12.4 0.87 9.8 11200 4970 1541 68 54 2.3
54 12.2 1.43 14.7 10260 4872 1705 69 59 3.3
56 8.0 1.49 16.9 11290 5000 1909 62 62 4.3
59 7.1 1.59 17.1 10710 5120 2340 54 64 4.3
62 7.2 1.47 16.8 11450 4972 1756 65 67 15.0
65 7.1 2.22 23.3 10530 4870 3120 39 70 45.0
73 5.0 2.04 21.8 10730 4670 2870 37 73 54.0
75 5.1 2.15 23.7 11050 5120 3215 41 75 62.0
78 5.2 2.15 10050 5230 3430 76 65.0
79 65.0
E-2
Table E-3 Experimental data of BMBR in high COD loading
Time HRT F/M L MLSS CODinf CODeff COD% Time Pressure
(day) (h) g/g.d (kg/m3.d) (mg/L) (mg/L) (mg/L) (%) (day) (kPa)
0 12.4 4005 1115 321 71.2 1 3
2 12.7 5200 1097 307 72.0 3 5
4 12.7 6200 1176 312 73.5 5 6
5 13.8 2.11 0.35 5970 1216 216 82.2 6 8
6 13.9 2.11 0.32 6700 1224 160 86.9 7 10
9 15.1 2.00 0.23 8600 1258 198 84.3 8 14
11 15.3 1.92 0.18 10950 1224 132 89.2 9 32
11 9.7 2.97 0.24 12500 1200 228 81.0 10 43
12 9.2 3.17 0.27 11540 1215 172 85.8 11 61
14 8.3 3.64 0.22 16500 1260 142 88.7 12 1
16 8.2 3.78 0.22 17450 1290 232 82.0 16 4
19 10.0 3.44 0.20 17400 1435 202 85.9 17 5
21 4.8 6.98 0.35 19800 1395 272 80.5 18 8
21 4.0 8.52 0.42 20100 1420 450 68.3 19 60
24 5.0 6.98 0.35 19850 1455 520 64.3 21 13
26 4.0 9.24 0.50 18500 1540 550 64.3 22 18
31 5.0 4.90 0.25 19400 1020 126 87.6 23 23
31 4.2 5.80 0.32 18200 1015 158 84.4 23 47
32 4.5 5.46 0.27 20050 1023 99 90.3 24 73
34 5.8 5.21 0.23 22400 1260 158 87.5 23 1
39 15.5 1.87 0.09 20150 1210 120 90.1 25 9
40 16.1 1.52 0.07 22500 1020 95 90.7 26 18
41 16.5 1.92 0.10 18950 1320 158 88.0 28 39
41 16.3 1.65 0.08 21750 1120 97 91.3 29 57
44 12.3 2.61 0.13 19970 1340 154 88.5 30 65
46 12.5 1.99 0.10 19750 1035 123 88.1 32 17
49 12.1 2.01 0.11 18769 1012 117 88.4 33 36
51 13.2 2.45 0.12 19700 1347 145 89.2 35 74
51 10.5 2.81 0.13 21720 1230 110 91.1 37 17
54 10.3 2.37 0.11 21434 1015 134 86.8 38 27
60 10.7 2.20 0.11 20352 980 85 91.3 41 2.0
63 3.2 9.38 0.48 19500 1250 320 74.4 43 2.7
66 3.7 9.41 0.45 21000 1450 420 71.0 46 5.0
70 4.1 8.78 0.43 20500 1500 570 62.0 48 7.0
72 3.2 10.50 0.53 19700 1400 630 55.0 49 15
75 3.6 9.00 0.42 21200 1350 790 41.5 50 35
51 65.0
E-3
Table E-4 Experimental data of YMBR in low COD loading
Time HRT F/M L MLSS CODinf CODeff COD% Time Pressure
(day) (h) g/g.d (kg/m3.d) (mg/L) (mg/L) (mg/L) (%) (day) (kPa)
1 9.7 2.60 0.51 5120 1050 263 75 1 0.67
3 9.0 2.59 0.43 6040 970 233 76 3 0.93
5 8.3 2.66 0.44 6100 920 288 69 5 1.20
7 8.1 2.98 0.49 6100 1005 185 82 7 0.93
9 8.2 3.22 0.56 5700 1100 150 86 9 0.80
11 9.1 2.24 0.42 5400 850 45 94.7 9 0.80
12 9.2 2.48 0.52 4750 950 24 97.5 12 0.93
13 8.7 2.39 0.52 4600 866 25 97.1 13 0.93
15 8.7 2.61 0.47 5500 945 80 91.5 15 0.93
17 9.0 2.69 0.54 4970 1009 62 93.9 17 2.00
18 7.5 3.17 0.57 5520 990 40 96.0 17 2.00
19 7.8 2.99 0.58 5200 972 35 96.4 19 2.00
21 7.8 2.74 0.55 4950 892 45 95.0 21 1.86
22 7.5 3.23 0.67 4790 1010 24 97.6 22 2.13
23 7.5 2.86 0.58 4950 893 95 89.4 23 2.13
24 8.1 2.71 0.56 4800 913 80 91.2 24 6.00
25 6.1 3.44 0.72 4750 874 110 87.4 24 6.00
27 6.5 3.58 0.87 4133 970 170 82.5 27 6.00
29 5.9 3.46 0.80 4300 850 140 83.5 29 6.00
31 5.8 4.18 0.92 4560 1010 170 83.2 31 9.0
33 4.9 4.46 1.00 4450 910 240 73.6 31 9.0
35 5.1 4.00 0.86 4670 850 180 78.8 35 25.0
37 5.2 4.57 0.98 4670 990 270 73 36 45.0
38 5.1 4.64 1.03 4520 985 240 75.6 38 62.0
41 4.8 5.28 1.21 4350 1056 240 77.3 45 65.0
46 7.5 3.42 0.70 4900 1070 51 95.2 45 0.52
50 7.0 3.60 0.74 4850 1050 20 98.1 49 0.67
52 7.0 4.01 0.78 5120 1170 45 96.2 50 0.67
54 6.9 3.51 0.63 5600 1009 65 93.6 56 0.67
56 6.8 3.35 0.54 6200 950 45 95.3 56 0.67
59 6.3 4.03 0.39 10400 1059 70 93.4 60 1.20
62 6.2 4.53 0.42 10900 1170 45 96.2 61 2.30
64 6.2 3.91 0.34 11500 1009 20 98.0 62 2.30
66 5.8 4.63 0.35 13200 1120 65 94.2 62 2.30
67 5.3 5.00 0.38 13500 1120 75 93.3 66 5.00
70 5.4 5.00 0.35 14200 1125 65 94.2 66 5.00
71 5.0 5.09 0.36 14300 1060 120 88.7 70 12.0
72 5.3 4.51 0.32 14000 995 110 88.9 71 25.0
73 5.3 5.07 0.35 14700 1120 95 91.5 72 32
75 5.3 4.94 0.33 15050 1090 95 91.3 73 63
77 4.0 6.75 0.47 14500 1125 148 86.8 75 65
80 4.2 6.05 0.40 15150 1059 190 82.1
82 4.0 6.36 0.43 14750 1060 135 87.3 76 1.5
84 4.0 7.00 0.46 15200 1170 190 83.8 76 1.5
86 4.0 6.27 0.40 15500 1050 150 85.7 79 1.7
89 3.9 6.89 0.46 15100 1120 101 91.0 80 2.3
81 2.5
E-4
Table E-5 Experimental data of BMBR in low COD loading
Time HRT F/M L MLSS CODinf CODeff COD% Time Pressure
(day) (h) g/g.d (kg/m3.d) (mg/L) (mg/L) (mg/L) (%) (day) (kPa)
1 7.8 3.23 0.29 5500 1050 50 95.2 1 20.0
3 8.1 2.87 0.28 4750 970 20 97.9 2 33.3
6 8.4 2.63 0.28 4500 920 45 95.1 3 79.8
9 7.9 3.34 0.46 5200 1100 25 97.0 4 82.5
12 8.2 2.78 0.51 4650 950 15 98.0 5 85.1
15 5.9 3.84 0.82 4700 945 45 95.2 7 20.0
17 6.5 3.73 0.76 4920 1009 45 95.5 8 33.3
19 6.0 3.89 0.86 4530 972 25 97.4 9 59.9
21 6.5 3.29 0.77 4300 892 20 97.8 10 79.8
25 6.8 3.08 0.68 4550 874 20 97.7 12 86.5
27 4.9 4.75 0.96 4950 970 110 88.7 13 86.5
29 5.1 4.00 0.78 5120 850 45 94.7 15 27.9
31 5.5 4.41 0.82 5400 1010 90 91.1 16 46.6
33 4.9 4.46 0.69 6500 910 75 91.8 17 85.0
35 5.2 3.92 0.62 6300 850 45 94.7 19 23.9
37 5.6 4.24 0.70 6100 990 70 92.9 20 59.9
41 6.8 3.73 0.61 6100 1056 20 98.1 21 86.0
42 6.8 4.13 0.70 5900 1170 45 96.2 22 86.5
44 7.1 3.21 0.59 5450 950 15 98.4 23 27.9
46 7.3 3.52 0.53 6700 1070 20 98.1 24 53.2
48 7.4 3.58 0.44 8100 1103 45 95.9 25 86.5
50 5.7 4.42 0.47 9500 1050 110 89.5 27 33.3
52 5.9 4.76 0.45 10600 1170 65 94.4 29 85.0
54 6.3 3.84 0.37 10500 1009 25 97.5 30 85.1
56 5.9 3.29 0.28 11700 810 20 97.5 31 86.5
59 4.4 5.78 0.48 12050 1059 45 95.8 34 58.5
62 4.7 5.97 0.46 12950 1170 95 91.9 35 82.0
64 5.2 4.66 0.35 13200 1009 65 93.6 36 85.1
66 4.6 5.84 0.41 14300 1120 76 93.2 37 86.5
67 3.9 6.89 0.46 15070 1120 155 86.2 42 20.0
70 4.1 6.59 0.45 14700 1125 120 89.3 43 32.0
71 4.3 5.92 0.38 15700 1060 45 95.8 44 78.0
72 3.8 6.28 0.38 16500 995 120 87.9 45 81.0
73 4.1 6.56 0.40 16300 1120 105 90.6 46 86.0
75 4.3 6.08 0.35 17400 1090 110 89.9 47 86.5
77 4.5 6.13 0.35 17100 1150 76 93.4 48 86.0
80 4.0 6.35 0.38 16900 1059 76 92.8 49 25.0
82 4.2 6.06 0.35 17200 1060 110 89.6 50 42.0
51 65.0
52 86.5
53 86
F-4
0 2 4 6 8 10 12 14 16 18 20 22 24
Volumetric Loading (kg COD/m3.d)
20
30
40
50
60
70
80
90
100
CO
D rem
oval (%
)
YMBR
BMBR
Table F-1 Regression analysis of Fig.
YMBR BMBR
Equation Y = 88.00264 - 1.04453 * X - 0.04398 * X2
Degree = 2
Number of data points used = 22
Average X = 10.45
Average Y = 70.3889
Degree: 0
Residual sum of squares = 4949.26
Coef of determination, R-squared = 0
Degree: 1
Residual sum of squares = 403.182
Coef of determination, R-squared = 0.918537
Degree: 2
Residual sum of squares = 350.384
Coef of determination, R-squared = 0.929205
Equation Y = 82.7394 + 3.5684* X - 0.5722 * X2
Degree = 2
Number of data points used = 34
Average X = 4.71309
Average Y = 81.5303
Coefficients:
Degree 0 = 82.7394
Degree 1 = 3.5683
Degree 2 = -0.57221
Orthogonal Polynomial Factors:
X Shift = 6.010248447204969
X Scale = 0.4454589472227987
Degree: 1
Residual sum of squares = 1091.09
Coef of determination, R-squared = 0.732687
Degree: 2
Residual sum of squares = 668.19
Coef of determination, R-squared = 0.836296
F-4
0.0 0.5 1.0 1.5 2.0 2.5
F/M ratio (d-1)
0.0
0.2
0.4
0.6
0.8
1.0
CO
D rem
oval ra
te (
g C
OD
/ g
MLSS.d
)
BMBR
YMBR
Table F-2 Regression analysis of Fig.
YMBR BMBR
Equation Y = -0.10406 + 1.1556 * X - 0.33699 * X2
Degree = 2
Number of data points used = 22
Average X = 0.951382
Average Y = 0.559181
Coefficients:
Degree 0 = -0.104058111
Degree 1 = 1.155606227
Degree 2 = -0.3369879819
Degree: 1
Residual sum of squares = 0.345839
Coef of determination, R-squared = 0.76943
Degree: 2
Residual sum of squares = 0.0467869
Coef of determination, R-squared = 0.968807
Equation Y = -0.03234 + 1.3306 * X - 1.2847 * X2
Degree = 2
Number of data points used = 34
Average X = 0.264056
Average Y = 0.202548
Coefficients:
Degree 0 = -0.03233949848
Degree 1 = 1.330610919
Degree 2 = -1.284738775
Degree: 0
Residual sum of squares = 0.273731
Coef of determination, R-squared = 0
Degree: 1
Residual sum of squares = 0.0350683
Coef of determination, R-squared = 0.871888
Degree: 2
Residual sum of squares = 0.0169625
Coef of determination, R-squared = 0.9380
F-4
60
70
80
90
100
CO
D r
em
oval (%
)
YMBR
BMBR
4.0 5.0 6.0 7.0 8.0 9.0 10.0
HRT (h)
50
60
70
80
90
100
CO
D rem
oval (%
)
YMBR
BMBR
SRT of 50 d
SRT of 10 d
Table F-3 Regression analysis of Fig.
YMBR BMBR
SRT 10 days
Equation Y = -18.2938 + 26.02184 * X - 1.4905 * X2
Degree = 2
Number of data points used = 19
Average X = 7.13158
Average Y = 88.2956
Coefficients:
Degree 0 = -18.29383197
Degree 1 = 26.02183739
Degree 2 = -1.490548568
Degree: 1
Residual sum of squares = 260.884
Coef of determination, R-squared = 0.799519
Degree: 2
Residual sum of squares = 149.022
Coef of determination, R-squared = 0.885482
Y = 91.219 - 9.232 * X + 2.896 * X2 - 0.2059 * X3
Degree = 3
Number of data points used = 14
Average X = 6.24286
Average Y = 94.8309
Coefficients:
Degree 0 = 91.21936071
Degree 1 = -9.231993703
Degree 2 = 2.895593329
Degree 3 = -0.2059391806
Degree: 2
Residual sum of squares = 36.227
Coef of determination, R-squared = 0.652589
Degree: 3
Residual sum of squares = 35.8103
Coef of determination, R-squared = 0.656585
F-4
(Continuous)
YMBR BMBR
SRT 50 days
Equation Y = 53.6689 + 10.8237 * X - 0.6825 * X2
Degree = 2
Number of data points used = 21
Average X = 5.49667
Average Y = 91.6494
Coefficients:
Degree 0 = 53.66891328
Degree 1 = 10.82366557
Degree 2 = -0.6825361179
Degree: 0
Residual sum of squares = 419.6
Coef of determination, R-squared = -2.22045E-016
Degree: 1
Residual sum of squares = 123.575
Coef of determination, R-squared = 0.705495
Degree: 2
Residual sum of squares = 110.933
Coef of determination, R-squared = 0.735623
Equation Y = 67.0097 + 7.8151 * X - 0.50207 * X2
Degree = 2
Number of data points used = 22
Average X = 5.24091
Average Y = 93.4416
Coefficients:
Degree 0 = 67.00973409
Degree 1 = 7.815103512
Degree 2 = -0.5020654975
Degree: 1
Residual sum of squares = 118.846
Coef of determination, R-squared = 0.576309
Degree: 2
Residual sum of squares = 113.208
Coef of determination, R-squared = 0.59641
