Louisiana State University LSU Digital Commons
LSU Master's Theses Graduate School
2014 Assimilation, a Biological Nitrogen Removal Strategy For Freshwater Ornamental Fish Hatcheries Fatemehsadat Fahandezhsadi Louisiana State University and Agricultural and Mechanical College, [email protected]
Follow this and additional works at: https://digitalcommons.lsu.edu/gradschool_theses Part of the Civil and Environmental Engineering Commons
Recommended Citation Fahandezhsadi, Fatemehsadat, "Assimilation, a Biological Nitrogen Removal Strategy For Freshwater Ornamental Fish Hatcheries" (2014). LSU Master's Theses. 3909. https://digitalcommons.lsu.edu/gradschool_theses/3909
This Thesis is brought to you for free and open access by the Graduate School at LSU Digital Commons. It has been accepted for inclusion in LSU Master's Theses by an authorized graduate school editor of LSU Digital Commons. For more information, please contact [email protected]. ASSIMILATION, A BIOLOGICAL NITROGEN REMOVAL STRATEGY FOR FRESHWATER ORNAMENTAL FISH HATCHERIES
A Thesis
Submitted to Graduate Faculty of the Louisianan State University and Agricultural and Mechanical College in partial fulfillment of the requirements for the degree of Master of Science in Civil Engineering
in
The Department of Civil & Environmental Engineering
by Fatemehsadat Fahandezhsadi B.E Chemical Engineering University of Tehran, 2008 May 2015
i
ACKNOWLEDGMENTS
My gratitude goes to my adviser and major professor, Dr. Malone, for his great help, support and advice during my master program at Louisiana State University. I also thank him for his patience by listening to me and encourage me spiritually and academically when the research didn’t go in the right direction. Thanks to Dr. Gutierrez-Wing and Dr. Hall for serving on my committee. Thanks to Dr. Hall for helping me through the various problems I had with lab works and supplements. I thank Dr. Gutierrez-Wing for her guidance and extensive amount of time she spend to discus and help me to accomplish this research. Thanks Dr. Blouin and Fan Wing for helping me with statistical analysis. I would also like to thank Mrs. Sandy Malone for bringing me into her family and her spiritual supports when I was far from my family.
Thanks to my friends and lab workers Asmita Phadke, Leslie Pipkin, Yasmin Mohammad,
Jonathan Barnett, Davis Lofton and Marlon Greensword for their great help with my lab works. I thank Daniel Alt who helped me a lot to design and build the setup at first when I didn’t know enough experience of building experimental set-up. I also thank Matthew Louque for helping me to build the final set-up. My special thanks to my husband, Nima Chitsazan, for always being there and encouraging me to finish the research when I was feeling down. I also thanks my parents and my brothers for their financial and spiritual support. Thanks God who makes all things happen and helps me in every step of my life. This research was supported by Southern Regional Aquaculture
Center (SRAC).
ii
TABLE OF CONTENTS
ACKNOWLEDGMENTS ...... ii
ABSTRACT ...... v
CHAPTER 1. INTRODUCTION ...... 1 1.1. General introduction ...... 1 1.2. Research objectives ...... 3 1.3. Organization of the thesis ...... 3
CHAPTER 2. BACKGROUND ...... 4 2.1. Ammonia in aquaculture systems ...... 4 2.2. Ammonia removal pathways ...... 6 2.2.1. Assimilation by photoautotrophic algae ...... 7 2.2.2. Conversion by chemoautotrophic bacteria...... 7 2.2.3. Assimilation by heterotrophic bacteria ...... 10 2.3. Polyhydroxyalkanoates ...... 11 2.3.1. Application of polyhydroxyalkanoates (PHAs) for nitrogen removal ...... 12 2.4. Discussion ...... 13
CHAPTER 3. ASSIMILATION STRATEGY FOR AMMONIA-N REMOVAL FOR LOW PH FRESHWATER ORNAMENTAL FISH BREEDING SYSTEMS ...... 16 3.1. Introduction ...... 16 3.2. Background ...... 17 3.2.1. Estimation of volumetric TAN removal rate ...... 19 3.3. Materials and methods ...... 21 3.3.1. Experiment 1: Batch study of pH effects ...... 21 3.3.2. Experiment 2: Batch Study of PHB Consumption rate ...... 24 3.4. Results and discussion ...... 26 3.4.1. Experiment 1 (pH effects) ...... 26 3.4.2. Experiment 2 (PHB consumption rate) ...... 33 3.5. Conclusions ...... 36
CHAPTER 4. A CONTNIUOUS LAB SCALE EVALUATION ...... 38 4.1. Introduction ...... 38 4.2. Background ...... 39 4.3. Materials and methods ...... 42 4.3.1. Experiment 1: Continuous study of assimilation strategy ...... 42 4.3.2. Experiment 2: Air pulsing frequency under continuous loading regime ...... 46 4.4. Results and discussion ...... 47 4.4.1. Continuous study ...... 47 4.4.2. Air pulsing frequency ...... 50 4.5. Conclusions ...... 52
CHAPTER 5. SUMMARY AND OUTLOOK ...... 53
iii
5.1. Experimental findings ...... 53 5.2. Recommendations for further research ...... 54
REFERENCES ...... 56
APPENDIX A: OBSERVED TAN CONCENTRATIONS FROM BATCH SYSTEM @ PH 8 (CHAPTER 3, EXPERIMENT 1) ...... 66
APPENDIX B: OBSERVED TAN CONCENTRATIONS FROM BATCH SYSTEM @ PH 6.5 (CHAPTER 3, EXPERIMENT 1) ...... 68
APPENDIX C: STATISTICAL ANALYSIS (CHAPTER 3, EXPERIMENT 1) ...... 70
APPENDIX D: OBSERVED TAN CONCENTRATIONS FROM CONTINUOUS SYSTEM WITH TAN LOADING RATE (CHAPTER 4, EXPERIMENT1) ...... 71 D.1 Data observed from continuous TAN loading rate @ 750 mg-N/day ...... 71 D.2 Data observed from continuous TAN loading rate @ 650 mg-N/day ...... 72 D.3 Data observed from continuous TAN loading rate @ 550 mg-N/day ...... 74 D.4 Data observed from continuous TAN loading rate @ 450 mg-N/day ...... 75 D.5 Data observed from continuous TAN loading rate @ 350 mg-N/day ...... 77 D.6 Data observed from continuous TAN loading rate @ 250 mg-N/day ...... 78
APPENDIX E: OBSERVED DISSOLVED OXYGEN (DO) DATA UNDER CONTINUOUS TAN LOADING @ 350 MG/DAY (CHAPTER 4, EXPERIMENT 2) ...... 80
VITA ...... 83
iv
ABSTRACT
Freshwater ornamental fish production is a major component of aquaculture in the southeastern United States. Closed recirculating aquaculture systems (RASs) allow freshwater ornamental fish hatcheries to mimic native water quality conditions for sensitive species. Total ammonia-N (TAN) removal is the main concern for closed RASs. Biological nitrification is the common method to control TAN. However, the nitrification process might be impaired in acidic water condition with pH<7. Another issue of using the nitrification process for TAN removal is that the nitrification is not sufficient to compensate for the amount of ammonia released in larval production systems because of slow growth rate of nitrifying bacteria. To address these issues, this study suggests an aerobic assimilation strategy in which fast growing heterotrophic bacteria directly convert TAN into microbial biomass. The heterotrophic bacteria needs an external carbon source. This study employed polyhydroxybutyrate (PHB) which is a solid insoluble bioplastic, as a carbon source in the aerobic assimilation process to avoid releasing carbon into the system. In order to understand the applicability of the aerobic assimilation process instead of nitrification, this study compared the TAN removal rate by aerobic assimilation biofilters operating @ pH 8 and 6.5 in a batch system. . The relationship between VTR and TAN appears to be well represented
3 by Michaelis-Menton kinetics which gave the mean values of VTRmax and K1/2 of 600 g-N/m -day and 0.1 g-N/m3, respectively. The study also verified the performance of the assimilation strategy under different continuous nitrogen loading regimes. This investigation can be used to establish the biofilter air pulsing frequency, which is the optimized air pulsing frequency that avoids clogging in the system. The results showed that there was no apparent release of carbon in the water tank. Unlike nitrification, assimilation strategy for TAN removal was not significantly different in two pH of 8 and 6.5 @ 28 ° C. The average consumption rate of the PHB was found
v as 21.93± 0.43 grams of PHB consumed per gram of TAN. The results from the continuous system, at a continuous feeding rate of 350 mg-N/d, suggested that the optimum air pulsing frequency through the PHB bed for releasing the excess biosolids was about 6-hour. In general, the study found that the aerobic assimilation using polyhydroxybutyrate (PHB) as a carbon source is an applicable method for ammonia-N removal particularly for freshwater ornamental fish production systems. The drawback against this approach is that the PHB beads are expensive. However, our investigations showed that the average cost of the PHB beads have been decreased dramatically during the last few years. This might indicate that the assimilation using PHB beads can be a reasonable choice for ammonia-N removal particularly for freshwater ornamental fish production systems.
vi
CHAPTER 1. INTRODUCTION
1.1 General introduction Recirculating aquaculture system (RAS) technology for fish culture has been well established for more than three decades (Nazar et al. 2013; Timmons and Center 2002; Masser et al. 1999; Timmons and Losordo 1994; Malone and Beecher 2000; Gutierrez-Wing and Malone
2006). RAS offers many advantages for successful commercial fish production systems like reducing the water demand and treating effluent wastes such as ammonia and organic matter instead of discharging them. RAS can be designed environmentally friendly to use up to 90% less water in comparison to other aquaculture systems. Closed RASs can also provide an optimum environmental condition for fish culture (Malone and Gudipati 2005; Andrews 1990; Timmons and Losordo 1994; Masser et al. 1999; Halachmi 2006).
Freshwater ornamental fish production is a major sector of aquaculture in the southeastern
United States (Chapman et al. 1997; Livengood and Chapman 2007; Yanong 1996). The ornamental fish hobby has grown in the U.S. over the last few decades (Miller-Morgan 2010).
About 95% of all U.S. freshwater ornamental fish production occurs in the state of Florida (Hill and Yanong 2010; Yanong 1996). Overall the ornamental fish trade in the United States was estimated at about $1.5 billion per year, with 80% for freshwater species and 20% for marine species (Andrews 1990; Larkin and Degner 2001).
Biological filtration is the most common method to remove ammonia-N from RAS.
Nitrification followed by denitrification is a conventional biological process for ammonia-N removal in aquaculture wastewater treatment systems (Crab et al. 2007; Malone and Pfeiffer 2006).
1
Nitrification is an aerobic oxidation of ammonia-N to nitrite-N and then nitrate-N via chemoautotrophic bacteria (Hagopian and Riley 1998; Sastry et al. 1999; Zhu and Chen 1999;
Gutierrez-Wing and Malone 2006). Nitrate-N reduction occurs through the anaerobic denitrification process with nitrogen gas as the final product (Van Rijn et al. 2006; Whitson et al.
1993).
Several factors such as dissolved oxygen (DO) concentrations, substrate, organic matter, temperature, pH, alkalinity and salinity, limit the nitrification process specially for freshwater larval production system (Crab et al. 2007). Acidic water with low pH (pH < 7) regimes can impair the process (Malone and Pfeiffer 2006; Wortman and Wheaton 1991). A biofilter’s ability to nitrify is also inherently limited by low TAN and nitrite-N concentrations (0.1 mg-N/l) (Malone and
Beecher 2000; Chen et al. 2006).
To address these issues, an aerobic assimilation strategy can be used. In this strategy, heterotrophic bacteria quickly assimilate the available ammonia-N compounds into their microbial biomass (Schneider et al. 2007; Ebeling et al. 2006). Heterotrophic bacterial growth is encouraged and the bacteria accumulate nitrogen. The nitrogen is then removed from the system by removing the bacterial biomass (Avnimelech 1999; McIntosh 2001).
The heterotrophic bacteria require an external carbon source as an electron donor, which is often provided by dosing water soluble carbon sources such as methanol, ethanol, glucose or other carbohydrates (Akunna et al. 1993; Sauthier et al. 1998; Zhu and Chen 2001). However, the approach requires constant supervision to prevent excessive carbon release. Employing a biodegradable solid non-water soluble polymer as a carbon source can improve the water quality by reducing the risk of the carbon release (Gutierrez-Wing et al. 2007).
2
1.2 Research objectives The main goal of this study was to replace the conventional nitrification/denitrification method of TAN removal from U.S. freshwater ornamental fish culture with the aerobic assimilation strategy using polyhydroxybutyrate (PHB) bioplastic as a solid, non-water-soluble organic carbon source. The objectives of this effort were to:
1) Measure the ability of a PHB based assimilation biofilter to assimilate ammonia-N under an
aerobic regime.
2) Quantify the impact of reducing pH on kinetic constants.
3) Measure kinetics formulations that can be used to size PHB assimilation biofilters.
4) Define the consumption rate of PHB per gram of nitrogen removed from the system.
5) Determine if the PHB bead is releasing carbon.
6) Verify the assimilation biofilter performance under different continuous nitrogen loading
regime.
7) Establish a biofilter air pulsing frequency through the PHB bed.
1.3 Organization of the thesis Chapter 2 describes the background information about different ammonia-N removal strategy from recirculating aquaculture systems focusing on the new aerobic assimilation method.
Chapter 3 investigates the effect of low pH on volumetric TAN removal rate (VTR) through the aerobic assimilation process using polyhydroxybutyrate (PHB) as an organic carbon source.
Chapter 4 includes the results from experiments at different continuous nitrogen loading rates.
Summary of the results, conclusions and recommendation for further studies were provided in
Chapter 5.
3
CHAPTER 2. BACKGROUND
2.1 Ammonia in aquaculture systems Ammonia is the major nitrogenous waste product produced by fish excretion and protein digestion of fish feed (Walsh and Wright 1995; Wright 1995). In general, an average of 30 grams of total ammonia-N (TAN) is produced per 1 kg of 35% protein fish feed (Malone and Beecher
2000; Malone et al. 1990). Ammonia is highly soluble in water and exists in water in two forms:
+ un-ionized (free) ammonia (NH3) and ionized ammonium (NH4 ). The following equation shows
+ the transformation of the ionized NH4 to un-ionized NH3 in water:
+ NH4 + OH ↔ NH3 + H2O (1)
The proportion of ionized ammonium to free ammonia existing in the water depends on salinity, pH and temperature of the water. At a pH higher than 7, Equation 1 is driven to the right and NH3 become the predominant form (Metcalf and Eddy 2003; Emerson et al. 1975; Trussell
1972). The un-ionized ammonia is extremely toxic to fish (Colt 2006; Meade 1985; Ip et al. 2001;
Masser et al. 1999; Körner et al. 2001). The sum of the two forms are called total ammonia- nitrogen (TAN) (Thurston et al. 1981). Although biologists more precisely use free ammonia as a guide for ammonia concentration, engineers tend to use TAN as a limiting factor in recirculating aquaculture systems because they do not control pH (Colt 2006).
Both TAN and nitrite-N are toxic to fish and must be controlled in closed recirculating aquaculture systems (Masser et al. 1999). Larvae are particularly susceptible to TAN and nitrite-
N concentrations. Hatching and larval production systems typically require extremely high water quality condition. TAN and nitrite-N levels should be held below 0.1 ppm for larval production systems (Young-Lai et al. 1991; Malone and Pfeiffer 2006). High TAN and nitrite concentrations can adversely affect hatching rate by posing stress to the larvae and consequently damaging the
4 fish eggs (Yanong 1996; Barimo and Walsh 2005) . Larval production systems require extremely high water quality condition for a few weeks as each egg is first hatched then reared as larvae
(Reinbold and Pescitelli 1982; Young-Lai et al. 1991). Previous attempts at larval rearing of Bala
Shark in Florida have failed due to decline in water quality. The sudden surge of organic and total ammonia-N induced the rapid decay of un-hatched eggs and empty egg membranes which complicates the use of traditional nitrifying biofilters as the bacteria are slow to respond to the sudden change in loading (Rusten, Eikebrokk et al. 2006).
Malone and Pfeiffer (2006) classified freshwater ornamental fish larval production systems as an “Acidic-Oligotrophic” area, which demands high water quality condition with a low pH regime that is inhibitory to the biofiltration process (Table 1).
Table 1: Biofiltration classification based on pH and trophic levels; after Malone and Pfeiffer (2006).
TAN/nitrite Class Application pH performance range (g-N//m3)
Ultra-oligotrophic Larval 7-8 0.0-0.1
Acidic-oligitrophic Ornamental 6-7 0.1-0.3
Oligotrophic Broodstock 7-8 0.1-0.3
Mesotrophic Fingerling 7-8 0.3-0.5
Eutrophic Growout 7-8 0.5-1.0
5
2.2 Ammonia removal pathways Biological filtration is the most cost effective way to control TAN and nitrite-N concentrations below the toxic levels in a closed recirculating aquaculture system (Hochheimer and Wheaton 1998; Yang et al. 2001; Abeysinghe et al. 1996; Crab et al. 2007). The process is called biofiltration because it employs bacteria that filter waste products. RAS systems use fixed film reactors with media like sand grains or plastic beads to support attached bacterial biofilm.
Typically, the bacteria convert ammonia-N and nitrite-N to nitrate-N which is the less toxic compound. (Masser et al. 1999; Malone and Beecher 2000).
Three nitrogen conversion pathways for ammonia-N removal in recirculating aquaculture systems are: ammonia-N conversion (1) to algae biomass using photoautotrophic algae, (2) to nitrate-N by autotrophic bacteria (nitrification process), and (3) to microbial biomass cells via heterotrophic bacterial growth (assimilation process) (Figure 1) (Ebeling et al. 2006; Crab et al.
2007).
Figure 1: Nitrogen conversion pathways for biological ammonia-N removal in recirculating aquaculture systems via different groups of bacteria after Ebeling et al (2006).
6
2.2.1 Assimilation by photoautotrophic algae Photoautotrophic algae based systems have been conventionally used to control ammonia-
N accumulation in recirculating aquaculture systems via adding some nutrient to stimulate algae
growth (Brune et al. 2004; Crab et al. 2007; Aslan and Kapdan 2006). Ebeling et al (2006)
describes the saltwater algae biosynthesis using ammonia as a nitrogen source:
+ - 2- 16 NH4 + 92 CO2 + 92 H2O +14 HCO3 + HPO4 → C106H263O110N16P + 106O2 (2)
For every gram of ammonia-N consumed, 15.14 grams of oxygen and 15.85 grams of algae
biomass are produced. Equation 2 predicts a carbon to nitrogen ratio (C/N) of 5.67 for the algae
based systems. Besides the need for sunlight, the variation of dissolved oxygen, pH, and ammonia-
N concentration are the disadvantages of this system because it is hard to take the huge amount of
algae biomass produced out of the water (Burford et al. 2003; Stumm and Morgan 2012).
2.2.2 Conversion by chemoautotrophic bacteria Nitrification followed by denitrification is currently the most common biological method
for ammonia-N removal (Reeves 1972; Prosser 1989; Metcalf and Eddy 2003; Sharma and Ahlert
1977; Sawyer et al. 1971). Nitrification is a two-step process in which ammonia-N is oxidized to
nitrite-N by autotrophic “Ammonia Oxidizing Bacteria” (AOB), and then nitrite-N is oxidized to
nitrate-N as the final product via “Nitrite Oxidizing Bacteria” (NOB) (Metcalf and Eddy 2003;
Henze 2002; Eding et al. 2006). Both steps are aerobic and are mediated by different groups of
chemoautotrophic bacteria, known as nitrifiers. In the first step, oxidation of ammonia-N to nitrite-
N is carried out primarily by bacteria genera as Nitrosomonas. This step is also carried out by other
autotrophic bacteria genera (prefix with Nitoso-) such as Nitrosococcus, Nitrospira, Nitrosolobus,
and Nitrosorobrio obtaining energy from oxidation of ammonia to nitrite:
7
+ - + 2NH4 + 3O2 → 2NO2 + 4H + 2H2O (3)
The second step is accomplished simultaneously by NOB, commonly known as
Nitrobacter, which oxidize nitrite-N to nitrate-N. Besides Nitrobacter, other autotrophic bacteria
genera such as Nitrococcus, Nitrospira, Nitrospina, and Nitroeystis (prefix with Nitro-) can also
oxidize nitrite-N to nitrate-N (Metcalf and Eddy 2003):
- - 2NO2 + O2 → 2NO3 (4)
The following formula describes the total oxidation reaction:
+ - + NH4 + 2O2 → NO3 + 2H + H2O (5)
According to Equation 5, for every gram of ammonia-N oxidized to nitrate-N, 4.57 grams
of dissolved oxygen are consumed. Equation 6 represents the complete nitrification process,
considering C5H7NO2 as the chemical formula for nitrifier bacterial cells:
+ - + NH4 + 1.863O2 + 0.098CO2 → 0.0196 C5H7NO2 + 0.98NO3 + 0.0941H2O + 1.98H (6)
Nitrification is an acid producing process (Losordo and Westers 1994). The amount of
alkalinity consumed for nitrogen oxidation can be estimated by the following equation:
+ - - NH4 + 2 HCO3 + 2O2 → NO3 + 2CO2 + 3H2O (7)
To convert 1 gram of NH4-N to NO3-N, 7.14 grams of alkalinity as a calcium carbonate
(CaCO3) is required (Masser et al. 1999; Metcalf and Eddy 2003). Chemoautotrophic nitrifying
bacteria obtain their energy from oxidation of inorganic compounds and use CO2 as a primary
carbon source (Sawyer et al. 1971; Ebeling et al. 2006; Hagopian and Riley 1998).
8
The rate of nitrification is influenced by environmental conditions such as dissolved oxygen (DO) concentration, pH and alkalinity (Eding et al. 2006; Satoh et al. 2000; Chen et al.
2006). Nitrification rate can dramatically drop with low DO concentration (< 2 mg/l) and stop at
DO concentration below 0.5 mg/l (Boyd 1998; Timmons and Center 2002; Sharma and Ahlert
1977; Stenstrom and Poduska 1980). The optimum temperature range for nitrifier bacteria is between 30-35 °C (86-95 °F), and nitrification rates dramatically drop at temperatures below 10
°C (Malone and Pfeiffer 2006; Zhu and Chen 2002; Saidu 2009). The optimum pH value for
Nitrosomonas and Nitrobacter is between 7 to 8.8, and pH levels below 7 can impair nitrification process (Wortman and Wheaton 1991). The alkalinity level recommended for maximum nitrification rates was more than 200 mg/l and nitrification rates can be reduced at alkalinity level below 40 mg/l as CaCO3 (Biesterfeld et al. 2003; Chen et al. 2006).
TAN and nitrite-N (substrate) concentrations in the water are the important factors affecting the nitrification process (Chen et al. 2006; Wheaton et al. 1994). Monod-type equation serves as a model representing the TAN concentration as the limiting factor for nitrifier bacterial growth (Chen et al. 2006):
(푆−푆0) 푅 = 푅푚푎푥 ∗ , (8) 퐾푆 + (푆−푆0) where R is the substrate conversion rate (g/m3.day)
3 Rmax is the maximum substrate conversion rate (g/m .day)
S is the limiting substrate concentration (g/m3), and
3 Ks is the half saturation constant ((g/m ).
9
The minimum TAN concentration (S0) for nitrification biofilters was estimated to be 0.07 ± 0.05 mg/l at temperature 27 ° C (Zhu and Chen 1999; Ester et al. 1994).
2.2.3 Assimilation by heterotrophic bacteria Another type of nitrogen management system relies on heterotrophic bacterial growth in a biofilter. At a high carbon to nitrogen ratio, heterotrophic bacteria quickly assimilate ammonia-
N into bacterial biomass. The typical organic carbonaceous substrates consumed are glucose, sucrose, or any other form of carbohydrates (Hagopian and Riley 1998; Michaud et al. 2006). The following equation predicts the stoichiometry of an assimilation via aerobic heterotrophic bacterial
+ growth using ammonia (NH4 ) as a nitrogen source, O2 as an electron acceptor and sucrose as an electron donor (Ebeling et al. 2006):
+ - NH4 + HCO3 + 1.18 C6H12O6 + 2.06 O2→ C5H7O2N + 3.07 CO2 + 6.06 H2O (9)
Table 2 summarizes the stoichiometry of aerobic assimilation via heterotrophic bacterial metabolism. For every gram of ammonia-N converted to heterotrophic bacterial biomass, 15.17 g of carbohydrates (sucrose), 4.71 grams of oxygen, and 3.57 grams of alkalinity are consumed. The oxygen demand of the aerobic assimilation process is higher but the alkalinity requirement is about half of the nitrification (dissimilatory) process. The heterotrophic assimilation process produces
40 times more biomass cells than nitrification (8.07 versus 0.2 g) (Chen et al. 2006). For every gram of biomass (VSSH) removed, 0.125 gram of nitrogen was removed from the system. Soluble carbon dosing demands a management system to control the rates while reducing the risk of releasing excess carbon (Zhu and Chen 2001; Kaiser and Schmitz 1988; Lin et al. 2002). This problem can be avoided by using sophisticated process control which requires higher technology
(Lee et al. 2000).
10
Table 2: Stoichiometry for heterotrophic bacteria metabolism of 1.0 g NH4+-N with sucrose as supplemental organic carbon after Ebeling et al (2006). Parameter Consumed (g) + NH4 ‒N 1.0
C6H12O6 15.17 Alkalinity 3.57 Oxygen 4.71
Soluble carbon dosing demands a management system to control the rates while reducing the risk of releasing excess carbon (Zhu and Chen 2001; Kaiser and Schmitz 1988; Lin et al. 2002).
This problem can be avoided by using sophisticated process control which requires higher technology (Lee et al. 2000).
2.3 Polyhydroxyalkanoates The use of a solid non water-soluble bioplastic as a carbon source and support media for bacterial growth can reduce the risk of carbon release. Polyhydroxyalkanoates (PHAs) are a family of biodegradable, environmental-friendly, thermoplastic linear polyesters produced in nature by many bacteria as their energy and carbon reserve compounds (Williams and Peoples 1996; Peoples and Sinskey 1989; Hankermeyer and Tjeerdema 1999). Bacteria naturally accumulate PHAs as a response to lack of some nutrients such as nitrogen, phosphorous, oxygen or magnesium in the presence of excess carbon source (Shang et al. 2003; Dawes 1990; Steinbüchel et al. 1992; Yu
2001; Du and Yu 2002; Ojumu et al. 2004). The first PHA was discovered in 1926 by Maurice
Lemoigne as a constituent of the bacterium Bacillus megaterium (Salit 2014; Verlinden et al. 2007;
Lemoigne 1926). More than 100 different types of PHAs can be produced by different types of bacteria. The production is also affected by using different nutrients, carbon substrate, and specific enzymes (Kawaguchi et al. 1992; Ashby et al. 2001).
11
The general structure formula of PHAs is shown in Figure 2. Most of PHAs are composed of R (-)-3-hydroxyalkanoic acid monomers with different numbers of carbon atoms from C3 to
C14. Different saturated or unsaturated bonds attached to straight or branched chain with the functional aliphatic or aromatic groups (Ojumu et al. 2004; Kawaguchi et al. 1992; De Smet et al.
1983). In Figure 2, R group is methyl (Reddy et al. 2003).The wide variety of PHAs can be designed and produced wi th different physical and chemical properties based on using different types and sizes of the functional group (Ojumu et al. 2004; Ashby et al. 2001). Poly (3- hydroxybutyrate) (PHB) is the simplest and most commonly occurring form of PHA family, which is synthesized in the bacterium Alcaligenes eutrophus. (Ojumu et al. 2004; Ackermann and Babel
1998; Choi and Lee 1997).
Figure 2: General structure of polyhydroxyalkanoates. R=alkyl groups C1-C13, X=1-4, n=100- 30000 (Philip et al. 2007).
2.3.1 Application of polyhydroxyalkanoates (PHAs) for nitrogen removal
Insoluble PHAs can be completely degraded in aerobic conditions releasing CO2 and water.
Besides CO2 and water, methane is also produced when the PHA degradation occurs in anaerobic conditions (Ojumu et al. 2004; Brandi et al. 1995; Philip et al. 2007). The degradation rate of PHAs in water depends on their physical properties, chemical structure (especially the functional group), microbial population in the environment, PHA surface area and environmental conditions such as temperature and pH (Domb and Kumar 2011; Ojumu et al. 2004; Philip et al. 2007).
12
Assimilation strategy is a sort of paradigm change to replace the existing biological filtration process (nitrification/denitrification) with aerobic PHA degradation (Gutierrez-Wing et al. 2007; Boley and Mller 2005; Boley et al. 2000). In contrast to the traditional method, passive assimilation using insoluble PHA biopolymer requires less supervision while avoiding the risk of carbon release as well as providing more water quality stabilization (Gutierrez-Wing et al. 2007;
Joye and Hollibaugh 1995).
This study represents the aerobic assimilation strategy for ammonia-N removal from US freshwater ornamental aquaculture systems using polyhydroxybutyrate (PHB) bioplastic as a solid, non-water-soluble carbon source instead of the conventional ammonia-N removal method
(nitrification/denitrification).
2.4 Discussion As reported in the literature, nitrifying bacteria are sensitive microrganisms which are susceptible to environmental conditions such as temperature, dissolved oxygen and substrate concentrations, pH, alkalinity and salinity (Zhu and Chen 2002; Malone and Pfeiffer 2006; Sharma and Ahlert 1977; Stenstrom and Poduska 1980; Eding et al. 2006; Wheaton et al. 1994). The nitrification rate dramatically declines at a low pH below 7 (Chen et al. 2005). Acidic-oligotrophic conditions inhibit the nitrification process. The problem becomes more complicated by the sudden surge of total ammonia-N (TAN) and organic materials when the eggs are introduced to the system.
Water quality is reduced by the sudden organic and ammonia load on the traditional filter due to rapid decay of un-hatched eggs and empty egg membranes. The problem is often complicated further by poor biofilter acclimation due to fact that the slow growing nitrifiers do not have enough time to adjust.
13
Nitrification process can be replaced by assimilation strategy. At high C/N ratios, heterotrophic bacteria assimilate ammonia-N and convert it to biomass. Since heterotrophic bacteria grow significantly faster than autotrophs, they don’t need too much time to adjust (Ebeling et al. 2006; Sedlak 1991; Avnimelech 1999; Hagopian and Riley 1998). As opposed to the nitrification/denitrification process demanding an aerobic followed by an anaerobic bioreactor, assimilation process occurs in only one aerobic bioreactor.
The use of insoluble carbon substrate shows advantages over traditional assimilation filters
(mainly anaerobic) (Lee et al. 2000; Boley and Mller 2005; Boley et al. 2000). The dosing soluble carbon source requires management and a control system, which can increase the cost of the process (Gutierrez-Wing et al. 2007). The use of solid insoluble bioplastic provides a system with a lower risk of carbon release, eliminating the requirement of having sophisticated control system of dosing soluble carbon source (Hiraishi and Khan 2003; Anderson and Dawes 1990;
Hankermeyer and Tjeerdema 1999).
The cost of the PHB media compared with soluble carbon source like methanol is an obstacle to commercialize the process (Gutierrez-Wing et al. 2007; Boley et al. 2000; Castilho et al. 2009). The bioplastic market price reported in 2010 was $2.11 per kg of PHB, in comparison to $0.39 per kg of methanol (Chanprateep 2010; Boley et al. 2000) .
PHB production has been recently improved to find different extraction techniques, reducing the current cost of the bioplastic (Castilho et al. 2009; Nath et al. 2008; Quillaguamán et al. 2008). Table 3 shows the PHB cost has dropped at 79.1% from 2000 to 2010. The cost reduction will make the assimilation solid based process more economically feasible.
14
Table 3: Cost estimates of different substrates for nitrate-N removal in years 2000 and 2010 after (Boley et al. 2000; Chanprateep 2010) Average Average cost Average cost Price (per Price (per consumption of of Substrate kg in kg in of substrate (kg denitrification denitrification - - 2000) 2010) substrate/kg ($/kg NO3 -N ($/kg NO3 -N - NO3 -N) in 2000) in 2010) Methanol $1.02 $0.39 3.03 3.09 1.18
Ethanol $1.21 $0.43 2.0 2.42 0.86
PHB $10.1 $2.11 2.9 29.29 6.12
15
CHAPTER 3. ASSIMILATION STRATEGY FOR AMMONIA-N REMOVAL FOR LOW PH FRESHWATER ORNAMENTAL FISH BREEDING SYSTEMS
3.1 Introduction Recirculating aquaculture systems (RAS) for fish culture have been well established for more than three decades (Timmons and Losordo 1994; Masser et al. 1999; Timmons and Center
2002; Nazar et al. 2013; Malone and Beecher 2000). Closed recirculating aquaculture systems offer the advantage of providing a controlled environmental condition for fish production, as well as conserving heat and water by reusing the water after physical and biological treatment (Andrews
1990; Lazur and Britt 1997). A well designed RAS can provide a sustainable environment by using
90-99% less water than other aquaculture systems (Nazar et al. 2013; Timmons and Losordo 1994;
Malone and Gudipati 2005).
Larval production systems (acidic oligotrophic RAS) can impair nitrification process because of the low pH regime (Malone and Pfeiffer 2006). The problem becomes more serious by the sudden surge of total ammonia-N (TAN), hatching enzymes, and organic material, which occurs when the eggs are introduced to the system. The rapid decay of un-hatched eggs and empty egg membranes put a sudden organic and ammonia load on the traditional biofilter generally causing a decline in water quality. The problem associated with the ammonia shock loading is often complexed by poor biofilter acclimation and the slow growing nitrifiers that do not have time to adjust (Rusten et al. 2006). Nitrification process is also inherently limited by low TAN and nitrite-N concentration (<0.1 mg/l) (Chen et al. 2006; Malone and Beecher 2000).
Assimilatory uptake of ammonia-N via fast growing heterotrophic bacteria can address the limitations associated with the nitrification process. The assimilation process is typically done by
16 dosing water soluble organic carbon sources like methanol, ethanol, and glucose, which requires a control system and supervision to avoid carbon release into the breeding tank (Hagopian and
Riley 1998). The use of solid insoluble bioplastics for an aerobic assimilation process can reduce the risk of carbon release as well as preventing the toxic sulfide production. The objectives of this chapter are to:
1) Compare the ammonia-N removal rate via aerobic assimilation biofilters operating @ pH 8 and
6.5.
2) Establish the carbon costs of the proposed aerobic assimilation filtration strategy.
3.2 Background Ammonia is the major nitrogenous waste product by fish as a result of protein metabolism, which exists in water in two forms: un-ionized (free) ammonia (NH3) and ionized ammonium
(NH4+) (Walsh and Wright 1995). The sum of the two forms is known as total ammonium nitrogen
(TAN). Control TAN and nitrite-N is an important aspect in recirculating aquaculture systems.
Biological filtration is the most common method to remove ammonia-N from RAS.
Biofilters employs suspended growth or attached growth (fixed film) process (Metcalf and Eddy
2003). Recirculation aquaculture systems have been mainly focused on using aerobic, fixed film bioreactors. The bacteria attach to the surface of media like rock, sand and plastic beads to grow
(Lekang and Kleppe 2000; Sandu et al. 2002; Singh et al. 1999; Malone and Pfeiffer 2006; Chen et al. 2006; Fitch et al. 1998; Akhbari et al. 2011). These fixed film processes display a stability of treatment that is difficult to maintain with suspended growth operating in a low substrate regime
(Malone and Beecher 2000).
17
Biological nitrification/denitrification via autotrophic bacteria has been traditionally used to control ammonia-N concentration in recirculating aquaculture systems (Gutierrez-Wing and
Malone 2006; Hagopian and Riley 1998; Sastry et al. 1999; Sharma and Ahlert 1977; Wheaton et al. 1994). Nitrification is a two-step biological process of oxidizing ammonia (NH4-N) to nitrite
(NO2-N) and then nitrite to nitrate (NO3-N), accomplished by chemoautotrophic aerobic bacteria called Nitrosomonas and Nitrobacter, respectively. Denitrification is a biological reduction of nitrate (NO3-N) to nitrogen gas (N2), which is mostly accomplished by facultative anaerobic heterotrophic bacteria (Metcalf and Eddy 2003; Van Rijn et al. 2006; Ebeling et al. 2006).
Nitrification process can be limited by many factors. Strictly aerobic nitrifying bacteria are susceptible to environmental conditions such as temperature, dissolved oxygen (DO), salinity, pH, alkalinity and the presence of organic carbon source (Zhu and Chen 2002; Malone and Pfeiffer
2006; Sharma and Ahlert 1977; Stenstrom and Poduska 1980; Eding et al. 2006; Wheaton et al.
1994). Moreover, slow-growing nitrifying bacteria demand longer acclimation time to adjust to reach the optimum nitrification rate (Rusten et al. 2006; Boller and Gujer 1986; Metcalf and Eddy
2003). Aerobic assimilation strategy can address the limitations associated with the nitrification process. In this process, fast-growing heterotrophic bacteria (5-20 times faster than nitrifiers) can convert the available ammonia-N into bacterial biomass (Metcalf and Eddy 2003; Henze 2002).
At high organic carbon to nitrogen (C/N) ratios, heterotrophic bacteria dominate the biofilter inhibiting nitrifiers growth by occupying the outer level of the biofilm and preventing the diffusion of nutrient and DO to the nitrifiers, which are placed in the inner layer of the biofilm (Ohashi et al. 1995; Nogueira et al. 2002; Satoh et al. 2000; Michaud et al. 2006). It is reported that nitrification rate is decreased by 70% at C/N ratios from 1.0 to 2.0 as compared to C/N=0 (Zhu and Chen 2001; Michaud et al. 2006).
18
An organic carbon source is usually provided by dosing soluble carbon sources such as methanol, ethanol and glucose (Avnimelech 1999; Kaiser and Schmitz 1988; Zhu and Chen 2001).
Although the soluble organic carbon compounds are readily available for bacteria, the dosing requires careful controls to prevent carbon release into the fish tank (Boley et al. 2000; Kaiser and
Schmitz 1988; Lee et al. 2000; Hagopian and Riley 1998). The problem compounds further for anaerobic denitrification in which the toxic sulfide can be produced as a result of excess carbon supply (Whitson et al. 1993). In the absence of oxygen, if the nitrate-N level fall too low because of excess carbon, the bacteria start to use sulfate as the next electron acceptor producing extremely toxic sulfide (Lee et al. 2000; Gutierrez-Wing et al. 2007).
A solid, non-water soluble carbon sources, polyhydroxybutyrate (PHB), has been used to remove nitrate-N through anaerobic denitrification process to prevent the carbon release
(Gutierrez-Wing et al. 2007). PHAs are insoluble biodegradable polymers that are produced by bacteria in which they naturally accumulate PHAs as a carbon and energy storage as a response to the lack of some nutrients such as nitrogen, phosphorous, oxygen or magnesium in the presence of excess carbon source (Shang et al. 2003; Dawes 1990; Steinbüchel et al. 1992; Yu 2001; Du and Yu 2002; Ojumu et al. 2004). The main goal of this study is to investigate the ability of aerobic assimilation process using polyhydroxybutyrate (PHB) bioplastic as a solid, non-water-soluble carbon source for ammonia-N removal from US freshwater ornamental aquaculture systems instead of the conventional ammonia-N removal method (nitrification/denitrification).
3.2.1 Estimation of volumetric TAN removal rate Biofilters are designed based on their ability to remove ammonia-N in terms of volumetric
TAN removal rate (VTR) with the unit of mass per unit volume of biofilter media per day (g/m3 media.day). Monod equation supporting by Michaelis-Menton enzyme kinetics has been used to
19 describe the kinetics of ammonia removal processes (Malone and Pfeiffer 2006; Malone and
Beecher 2000; Zhu and Chen 2002; Saidu 2009)
(푇퐴푁−푇퐴푁0) 푉푇푅 = 푉푇푅푚푎푥 ∗ (10) 퐾1/2 + (푇퐴푁−푇퐴푁0)
3 Where VTR is the volumetric TAN conversion rate (g-N/m media-day)
3 VTRmax is the maximum specific rate of TAN utilization (g-N/m media-day)
3 K1/2 is the half saturation constant (g-N/m )
TAN is the TAN concentration in the water (g-N/m3)
3 TAN0 is the minimum TAN concentration for nitrification biofilter (0.07 ± 0.05 g/m @27 ° C)
(Zhu and Chen 1999; Ester et al. 1994).
Equation 10 shows the strong dependency of the substrate removal rate on the TAN concentration. The higher TAN concentration yields higher TAN removal rate (Zhu and Chen
1999, 2001; Losordo and Hobbs 2000). At high TAN concentrations (TAN>>KA), Equation 14 can be simplified to the zero order kinetic reaction. In this case, the TAN removal rate no longer depends on TAN concentration:
푉푇푅 = 푉푇푅푚푎푥 (11)
At low TAN concentrations, the Equation 10 predicts a first kinetic reaction:
푉푇푅 ∗푇퐴푁 푉푇푅 = 푚푎푥 (12) 퐾퐴
Malone et al (2005) represented a new term, “τ”, for the linear regression of the plot between TAN and VTR with a zero intercept:
푉푇푅 = τ *TAN (13)
20
The term τ can be used as a parameter showing the biofilter performance in low TAN concentration regimes. The main goal of this study was to investigate the use of up-flow packed- bed biofilter for ammonia-N removal through aerobic assimilation process using biodegradable
PHAs.
3.3 Materials and methods Two sets of experiments were performed to investigate the effect of pH on the ammonia-N assimilation rates and one experiment was run to determine the bioplatsic consumption rate during the aerobic assimilation process. All the experiments were run in batch laboratory scale recirculating systems using polyhydroxybutyrate (PHB) as a carbon source.
3.3.1 Experiment 1: Batch study of pH effects Experiment 1 was performed to measure the kinetic formulations for sizing PHB assimilation biofilters and quantify the influence of low pH regime on the assimilation rate
(biofilter performance) in a batch system. Temperature and flow rate were kept constant during the experimental run.
The experimental system consisted of two identical independent expandable fixed-film up- flow bench scale biofilters filled with 1 liter PHB media (Figure 3). Each filter was constructed of
4” ID acrylic PVC pipe loaded with 1 L volume of the PHB beads. The unexpanded biofilter height was 5”. Water was recirculating from the 50 L water tank to the biofilter using 1/50 HP centrifugal pump (Little giant 3X-MDX). Bacteria convert the available ammonia-N into the biomass in the biofilter using PHB as an organic carbon source. Air was through the PHB bed for 1 minute by using air pump to control the biomass accumulation and avoid bed clogging issues. The air injection was set to expand the bed about 50%. The air impulse was set at 4 hr interval using
21 automatic controller connected to the air pump. The water with heavy biomass flowed from the biofilter to the PolyGeyserR floating bead filter to remove the excess biosolids from the recirculating system(Malone and Gudipati 2005).
Figure 3: Schematic diagram of aerobic fixed-film bioreactor to control ammonia-N concentration and PolyGeyserR biofilter for solid removal.
The PolyGeyserR was constructed using acrylic PVC pipe with 4” ID and was 25” in height.
Air was injected to the PolyGeyserR air chamber which backwashed the PolyGeyserR every 20 minutes. As the water flow through the floating bead, the solids were captured by the plastic bead, settled down and removed daily from the sludge valve located at the bottom of the PolyGeyserR.
Clear water flowed from the top of the PolyGeyserR to the water tank.
22
The experiment was performed under controlled conditions. The dissolved oxygen was maintained above 3 mg/l after the PHB bed using aquarium air pumps to blow air to the bottom of the water tank. Temperature was set at 28 ±1 °C using a heater in each water tank. The feed water for the experiments was tap water which was aerated for 2 days. The first set of experiments were run using aged tap water @ pH 8. Then, the desired amount of 85% phosphoric acid was added to the water tank to lower the pH to 6.5. The same experimental procedure was then performed for pH @6.5. Each set of experiment was performed in triplicate using both identical set-up. During the acclimation time (3 days), the PHB bead filters were not pulsed with air to accelerate the bacterial growth rate and to assure the population of bacteria is high.
A mixture of ammonia chloride (NH4Cl) and dipotassium phosphate (K2HPO4) were dissolved and added to the water tank to stimulate the bacterial growth in the system. The other minerals necessary for bacterial growth support was assumed to be supplied in tap water. After acclimation, the desired amount of ammonia chloride and dipotassium phosphate were dissolved and added to the water tank to obtain 2.5 mg/l NH4Cl (as N) and 7.5 mg/l K2HPO4 concentrations.
Samples were taken every hour from the middle of the water tank and analyzed for ammonia-N, nitrite-N, nitrate-N and chemical oxygen demand (COD). Samples were collected until the ammonia-N concentration was dropped to zero. Nitrate-N concentrations were measured for the first and the last samples to see if there is any accumulation of nitrate-N in the system.
PHBs were taken out of the reactor after each set of batch experiments to sanitize the acrylic
PVC pipe, screens and all the pipes. After taking out all the beads from the biofilter, the reactor cylinder, water tank and all the pipes and screens were washed with diluted bleach (1250 ppm) to minimize the problems associated with the existence of bacteria outside of the bioreactor. The chlorine was flushed out of the system with tap water. The reactor was rinsed with deionized water 23 before loading the beads. The bioreactors were reloaded with the same acclimated beads after cleaning to accelerate the acclimation time.
Dissolved oxygen (DO) of the water tank was checked by DO meter (YSI model 85 D) to insure that the DO concentration maintained above 2 mg/l in the biofilter. pH in the water tank was recorded by pH meter (YSI pH 100 meter). Ammonia-N, nitrite-N and nitrite/nitrate were measured using a SmartChem 170 Discrete Analyzer (Unity Scientific Inc.) based on EPA methods 350.1, 354.1, and 353.2, respectively. COD was measured using TNT 821 vial (HACH
Inc.) which is automatically read by spectrophotometer DR 6000 (HACH Inc.) based on EPA method 5220 D.
3.3.2 Experiment 2: Batch Study of PHB Consumption rate This experiment was performed to define the consumption rate of PHB per gram of nitrogen removed through aerobic assimilation process in a batch system. Temperature and flow rate were kept constant during the experimental run.
The aerobic assimilation process is a three-step process of ammonia-N removal by growing, releasing and then removing the biosolids from the system. In the first step, heterotrophic bacteria directly convert the available ammonia-N into their biomass cells generating biosolids.
The “poor man” biofilter was taken out of the water and manually shaken after each experimental run to release the excess biosolids which grew on the surface of the PHB media. In the third step, the PHB beads were rinsed with DI water to remove the biosolid.
The poor man biofilter was constructed of a small plastic container with some holes at the bottom filled with 1 liter PHB media. The biofilter was placed in the bottom of a 20 gallon aquarium tank (Figure 4, a & b). Air was injected into an airlift placed above the media using an
24
aquarium air pump. The three basic operations of recirculating aquaculture systems (circulation,
aeration, and degasification) were performed by an airlift that draws the water through the bottom
of the media.
(a)
Sampling
(b) Figure 4: “poor man” biofilter: (a) Experimental unit consisting of a “Poor man” biofilter using air-lift to draw the water through the bottomn of the media. (b) Schematic of a “poor man” biofilter incorporating airlift for water recirculation, aeration and degasification.
25
A mixture of ammonia chloride (NH4Cl) and dipotassium phosphate (K2HPO4) was dissolved and added to the water tank to stimulate heterotrophic bacterial growth (Zweifel et al.
1993; Wortman and Wheaton 1991; Suthersan 1996). Temperature was set at 28 ± 0.5°C.
After an acclimation period (3 days), NH4Cl (as N) and K2HPO4 were added to a fresh batch system at 2.5 and 7.5 mg/l concentrations to estimate the average consumption rates of PHBs per gram of ammonia-N removed from the system (Zhu and Chen 2001). Samples were analyzed for ammonia from 2.5 mg/l until it dropped to zero. The first and last samples were analyzed for nitrite/nitrate to see if there was any nitrite/nitrate accumulation in the system.
Samples were taken from the aquarium tank every 2-hour for ammonia-N determination until the ammonia-N was removed from the water. Samples for ammonia analyzing were collected in 50 ml glass vials and immediately analyzed.
Samples were analyzed for ammonia and nitrite/nitrate by a SmartChem 170 Discrete
Analyzer (Unity Scientific Inc.) based on EPA methods 350.1 and 353.2. The dry new PHBs were initially weighted. Once the added ammonia-nitrogen was removed from the water, the PHB media were taken out, rinsed with deionized water (DI water), air-dried overnight and weighed. The consumption rates were obtained by comparing the final weight of bioplastics to the amount of ammonia-nitrogen utilized by the system.
3.4 Results and discussion
3.4.1 Experiment 1 (pH effects) The experiment was run to investigate the influence of pH on assimilation rate and the biofilter kinetics constants. Ammonia removal was observed 3 days after the system was hooked up indicating the bacteria were acclimated. The acclimation time is significantly lower than the
26 nitrification process (2-3 weeks). This faster acclimation is due to higher reproduction rates of heterotrophic bacteria than autotrophic nitrifiers (Masser et al. 1999). Temperature was maintained constant at 28 ±1 °C using a water heater. The dissolved oxygen values were maintained above 3 mg/l after PHB biofilter during the experimental run. Each batch of ammonia-N added to the system was completely disappeared in several hours. The experiment was run in triplicate. The average chemical oxygen demand was 8.674 ± 2.60 and 8.449 ± 2.26 for experimental run @ pH
8 and 6.5, respectively showing there was no apparent carbon release into the water tank (see
Appendix A & B). Figure 5 shows the picture taken from the new PHB media and the media used in assimilation biofilters. The left picture shows the extensive bacterial growth on the surface of the PHBs which needs to be mobilized and removed to prevent the PHB bed clogging. The PHB beads maintained their rigidity below the soft outer layer of the bacteria.
Figure 5: The PHB media supplied by Metabolix, Inc. (USA) before and after 4 days using in assimilation biofilters showing the aggressive surface attacks of the beads by bacteria.
Figures 6 & 7 show the observed ammonia-N concentrations versus time. Rapid decrease of total ammonia-N concentrations with time for both experiments using aged tap water @ pH 8 and pH 6.5 were observed.
27
3
) 3 2.5
2
1.5 First Run Second Run 1 Third Run
0.5 TAN CONCENTRATION (G/M CONCENTRATION TAN
0 0 2 4 6 8 10 TIME (HR)
Figure 6: Rapid decrease of TAN concentration through assimilation process using aged tap water @ pH 8.
3
2.5
2 First Run
1.5 Second Run
1 Third Run
TAN CONCENTRATION (G/M 3) CONCENTRATION TAN 0.5
0 0 2 4 6 8 10
TIME (HR)
Figure 7: Rapid change in TAN concentration through assimilation process using low pH water @ 6.5.
28
The figures show that the experiments were replicated very well. The observed data shows there is no apparent evidence of nitrite-N and nitrate-N production, which indicates that nitrification process was inhibited by assimilation process via fast growing heterotrophic bacteria
(Appendix A & B).
Volumetric TAN removal rates (VTR) represent the biofilter performances calculated from observe data based on the differences between TAN concentrations. The unit of VTR is mass per unit volume of biofilter media per day (g/m3 media-day). Ammonia-N assimilation rates were plotted in Figures 8 & 9 for experimental run @ pH 8 and 6.5, respectively. The figures show the apparent correlation between the TAN concentrations and the VTRs. Higher assimilation rates can be achieved at the higher ammonia concentrations. The best Michaelis-Menton Monod type graph of reaction kinetics (Equation 10) was fitted to the observed data resulting VTRmax and K1/2 values of 640 g/m3-day and 0.08 g/m3 for pH 8 and 560 g/m3-day and 0.12 g/m3 for pH 6.5.
900
800
700 First Run
600 Second Run
DAY)
- 3 500 Third Run 400 Michaelis- Menton
VTR(G/M 300
200
100
0 0 0.5 1 1.5 2 2.5 3 3 TAN CONCENTRATION (G/M ) Figure 8: Volumetric TAN removal rates (VTR) for aerobic assimilation biofilters using aged tap water @ pH 8 as compared to the Michaelis-Menton kinetics.
29
900
800
700
600 First Run
500 Second Run
DAY) 400 -
3 Third Run 300 Michaelis-Menton 200
VTR (G/M VTR 100
0 0 0.5 1 1.5 2 2.5 3
TAN CONCENTRATION (G/M3)
Figure 9: Volumetric TAN removal rates for aerobic assimilation biofilters using tap water @ pH of 6.5 as compared to the Michaelis-Menton kinetics.
The plotted observed data for two different pH @ 8 and 6.5 showed no significant differences between the assimilation rates of the pH regimes. This was verified by a log transformed linear model using SAS software version 9.4. The analysis of covariance (ANCOVA) was run on the linearized data proved the volumetric TAN conversion rates were not statistically significantly different for pH 8 and 6.5 regimes (p value of 0.0603 > 0.05). Malone and Beecher
(2000) classified the recirculating aquaculture applications into three main categories of oligotrophic, mesotrophic and eutrophic. The classification was based on TAN and nitrite-N levels.
Oligotrophic zone (broodstock category) demands extremely high water quality with TAN and nitrite-N levels below 0.3 mg/l. The Ammonia and nitrite-N levels can be a little higher for the second zone, fingerling category (<0.5 mg/l). The eutrophic term was used for grow out category in which the TAN and nitrite-N levels can be increased to 1.0 mg/l. The results were separately plotted for each category in order to define the kinetics constants for each application area.
30
Figure 10 shows the linear curve fitted to the observed data obtained for oligotrophic zone at two different pH levels of 8 and 6.5.The term “τ” is the slope of the linear regression under the assumption of zero intercept.
Oligotrophic Zone
900
800
700 y = 2004.6x R² = 0.7671
600
day) - 3 500 pH 6.5
400 pH 8 Linear (pH 6.5)
VTR (g/m VTR 300 y = 1640.2x Linear (pH 6.5) R² = 0.7124 200 Linear (pH 8)
100
0 0 0.05 0.1 0.15 0.2 0.25 0.3 TAN Concentration (g/m3)
Figure 10: Direct linear regression fitted to the observed data for oligotrophic water quality range @pH 8 and 6.5.
The τ values for pH 8 and 6.5 are 1758 and 1452 compared to the value for nitrification process of 750(Hearn 2009). Since the log transformation of linearized data using ANCOVA test showed no significant differences between the VTR for the two pH regime, the mean value of
1822 can be used as a general τ value for oligotrophic zone for aerobic assimilation process
(Appendix C).
However the linear line was not properly fitted to the data from mesotrophic RAS zone,
1466 and 1129 can be used as conservative values for τ at pH 8 and 6.5, respectively (Figure 11).
31
Mesotrophic Zone 900 800 700 y = 1466x 600
day) R² = 0.6313
- 3 500 pH 6.5 400 pH 8 y = 1129.2x 300 R² = 0.4874 Linear (pH 6.5) VTR (g/m VTR 200 Linear (pH 8) 100 0 0 0.1 0.2 0.3 0.4 0.5 TAN Concentration (g/m3)
Figure 11: Linear regression line fitted to the observed data for mesotrophic water quality range @pH 8 and 6.5.
The average τ value for mesotrophic zone was 1297. The observed data for eutrophic RAS applications (0.5-1.0 g-TAN/m3) were followed the zero order kinetic equation which leads to
VTRmax 0f 645 ± 80 and 550 ± 80 for pH 8 and 6.5, respectively (Figure 12).
Eutrophic Zone 900
800
700 )
600
day - 3 500
400 pH 6.5
300 pH 8 VTR (g/m VTR 200
100
0 0 0.5 1 1.5 2 2.5 3 TAN Concentration (g/m3)
Figure 12: Observed data for eutrophic water quality area @pH 8 and 6.5 follows zero order kinetics reaction.
32
The mean value of 594 ± 93 can be used for two different pH of 8 and 6.5 regime and eutrophic RASs. Although the τ values are higher than nitrification rate for oligotrophic and mesotrophic RAS applications, the conversion rate is not superior in eutrophic zone.
Table 4 summarize the kinetics equations that can be used to design a biofilter for ammonia-N removal through aerobic assimilation process based on the related RAS categories.
Table 4: Volumetric TAN conversion rates for different zones and different pH regimes
Class pH Kinetics Equation
Oligotrophic 8 & 6.5 푉푇푅 = 1822 ∗ (푇퐴푁)
Mesotrophic 8 & 6.5 푉푇푅 = 1297 ∗ (푇퐴푁)
Eutrophic 8 & 6.5 푉푇푅 = 594 ± 93
3.4.2 Experiment 2 (PHB consumption rate) The half reaction for ammonia removal using polyhydroxybutyrate (PHB) as the electron donor (organic carbon source) was calculated based on the half reaction for ammonia-N as a nitrogen source and oxygen as an electron acceptor.
Bacterial cell synthesis (Metcalf and Eddy 2003):
- + + - 0.2 CO2 + 0.05 HCO3 + 0.05 NH4 + H + e = 0.05 C5H7O2N + 0.45 H2O (10)
Electron acceptor (Metcalf and Eddy 2003):
+ - 0.25 O2 + H + e = 0.5 H2O (11)
Electron donor (PHB) (Gutierrez-Wing et al. 2007):
33
+ - 0.0556 C4H6O2 + 0.3333 H2O = 0.2222 CO2 + H + e (12)
Total reaction:
+ NH4 + HCO3- +1.6 C4H6O2 + 2.14 O2 = C5H7O2N + 2.35 CO2 +3.77 H2O (13)
Equation 13 describes the total reaction of ammonia-N assimilation, using PHB as an organic carbon source. By considering the molecular weight of ammonia-N and PHB, the consumption rate of PHB can be predicted as about 9.84 g of PHBs for every g of ammonia- nitrogen converted into bacterial biomass.
Table 5 provides the average and standard deviation of PHB consumption rates based on the differences between the initial and final dried bioplastic weights comparing to the amount of ammonia-N removed from the system.
The experimental results show the average consumption of 21.93 ± 0.43 grams of PHB consumed per gram of ammonia-N removed during assimilation process.
Table 5: PHA consumption rate per gram of ammonia-N removed in the batch experiment at 28 ± 0.5 ° C.
Initial Pan pan +dry Dry residue plastic NH4-N g PHB /g weigh Weight residue(g) (g) consumed consumed NH4-N (g) (g) (g) (mg) 607.42 141.4 466.02 149.450 8.050 375.15 21.46
607.18 141.6 465.58 150.159 8.559 383.47 22.32
607.2 141.2 466 149.460 8.260 375.15 22.02
mean 21.93
St. Dev 0.43
34
The experimental consumption rate value was higher than the theoretical value. Aerobic assimilation is a three-step process. The three steps are: (1) biosolid generation, (2) mobilization and (3) removal. Failure of the third step might cause an internal recycling of nitrogen within the system that might be the cause of the difference between the experimental and the theoretical values of the PHBs consumption rates. The decay rate of the biosolid (Kd) in a recirculating aquaculture system was estimated at 36% in the first 24 hours (Chen et al. 1997). Particularly in this experiment, the sludge was removed daily after each experimental run allowing the biosolids to remain in the system during the experiments which increased the sludge residence time and released some of the nitrogen to the water.
The PHB consumption rate was significantly higher (under aerobic condition in comparison to anaerobic condition. The average consumption rate of the PHB was found to be
2.92 g of bioplastic consumed for every gram of nitrate-N removed through anaerobic denitrification process (Gutierrez-Wing et al. 2007).
The experimental results showed the consumption rate of the PHB was 10 times higher for ammonia-N removal through aerobic assimilation process.
Although the consumption rate of PHBs in aerobic condition is higher than anaerobic process, the application of aerobic assimilation technology using insoluble bioplastic offers many advantages for oligotrophic systems. The VTR obtained for oligotrophic zone was superior to the data from nitrification process. The available TAN can be completely removed through the aerobic assimilation process. The experimental results proved that low pH regime cannot limit the performance of assimilation biofilters. The presence of oxygen avoids the risk of toxic sulfide production in the system.
35
In general, an average of 30 grams of total ammonia-N (TAN) is produced per 1 kg of 35% protein fish feed added to the system (Malone and Beecher 2000; Malone et al. 1990). The average cost estimated of PHB to remove ammonia-N produced by 1 kg of fish feed with 35% protein was calculated $6.66 in 2000 and dropped to $1.39 in 2010. Although the average cost of removing the same amount of ammonia-N using soluble carbon source like methanol or ethanol was lower than the one using insoluble PHB, the 79.1% reduction of the PHB price can make the process economically feasible (Chanprateep 2010).
The aerobic assimilation strategy seems most promising for oligotrophic systems, particularly for larval systems. Although the current price of the PHB ($2.11/kg of PHB) might preclude the use for eutrophic broad systems, nitrogen limitation environment (oligotrophic condition) can meet the economic requirement by reducing the PHB consumption. In a freshwater larval production application, the PHA based system can be used for a few weeks as each batch of eggs is first hatched, then reared as a larvae. Several studies have been focused on using different substrates to decrease the PHA production cost (Castilho et al. 2009; Nath et al. 2008;
Quillaguamán et al. 2008; Lenz and Marchessault 2005; Lee and Chang 1995; Lee et al. 1999;
Tsuge 2002). Other than the low consumption rate of bioplastic for larval production systems, the use of PHB can create a superior tank environment to regular nitrification filters which is inhibited by the TAN and nitrite-N concentrations below 0.1 mg-N/L. The low nitrogen level environment limits the nutrient for bacterial growth creating an almost free disease environment for larvae.
3.5 Conclusions The obtain results indicates that ammonia-N can be removed directly from recirculating systems through aerobic assimilation strategy. The log transformation of linearized model showed no significant differences between VTRs under two pH regimes. The relationship between VTR
36 and TAN appears to be well represented by Michaelis-Menton kinetics which gives the mean
3 3 values of VTRmax and K1/2 of 600 g-N/m -day and 0.1 g-N/m , respectively. The observed value for COD during the experiment was 15.14 ± 5.43 showing no apparent release and accumulation of carbon source in the water tank. The consumption rate of the PHB using ammonia-N as a nitrogen source under aerobic condition found 21.93± 0.43 which is almost 10 times higher than the consumption rate under anaerobic denitrification process.
PHB production has been recently improved to find different extraction techniques, reducing the current cost of the bioplastic (Castilho et al. 2009; Nath et al. 2008; Quillaguamán et al. 2008).
The price per kg of PHB was $10.1 in 2000 and decreased to $2.11 per kg of PHB (Boley et al.
2000; Chanprateep 2010). The cost reduction can make the aerobic assimilation solid based process more economically feasible for commercial scale system.
37
CHAPTER 4. A CONTNIUOUS LAB SCALE EVALUATION
4.1 Introduction The use of closed recirculating aquaculture systems (RAS) for fish culture has been developed over the last three decades because of the limitation of weather, water and land supplies and the increasing environmental restrictions in countries (Timmons and Losordo 1994; Masser et al. 1999; Timmons and Center 2002; Nazar et al. 2013; Malone and Beecher 2000). Closed RAS technology can provide a sustainable, environmental friendly approach that reduces diseases and losses by controlling the water quality such as dissolved oxygen (DO), temperature, and inorganic nitrogen (Martins et al. 2010; Gutierrez-Wing and Malone 2006).
Closed recirculating aquaculture systems allow freshwater ornamental fish hatcheries to mimic native water quality conditions for sensitive species (i.e. low alkalinity, low hardness, neutral to acidic pH, and temperature 26-28°C). In a flow through format, the use of water softener, reverse osmosis systems, heaters, acid injectors and similar control systems are costly; therefore, the reuse of water is necessary.
Many attempts at freshwater ornamental fish reproduction in the U.S. have been inhibited by poor survival of larval fish. Acidic oligotrophic water quality is particularly hard to achieve in a commercial setting (Malone and Pfeiffer 2006). The nitrification processes have been impaired due to the susceptibility of nitrifying bacteria to low pH (<7) and TAN concentration (<0.1 ppm)
(Chen et al. 2006; Malone and Pfeiffer 2006; Malone and Beecher 2000).
Anaerobic assimilation strategy using an insoluble biodegradable polymer polyhydroxybutyrate (PHB), a simple form of polyhydroxyalkanoates (PHAs), is proposed to meet the water quality criteria defined for the acidic oligotrophic category. In this approach,
38 heterotrophic bacteria convert available ammonia-N into biomass through the aerobic assimilation processes, bacterial biomass is mobilized by air injection beneath the bed, and nitrogen rich solids are removed from the system.
This study will address the two following objectives:
1) Quantity the aerobic assimilation strategy performance under different continuous nitrogen
loading regime.
2) Establish a biofilter air pulsing frequency that prevents clogging of the PHB bed.
4.2 Background Ammonia is the major nitrogenous waste product of fish excretion. In general, an average of 30 grams of total ammonia-N (TAN) is produced per 1 kg of 35% protein fish feed added to the system (Malone and Beecher 2000; Malone et al. 1990). Ammonia exists in water in two forms of
+ un-ionized ammonia (NH3) and the ammonium ion (NH4 ). (Walsh and Wright 1995; Wright 1995;
Malone and Burden 1988). The sum of the two forms are called total ammonia-nitrogen (TAN)
(Thurston et al. 1981). Total ammonia-N (TAN) removal is the main biological filtration objective for recirculating aquaculture systems (Walsh and Wright 1995; Masser et al. 1999; Colt 2006;
Russo and Thurston 1991).
Ammonia has been traditionally converted to less toxic nitrogenous compounds through biological nitrification process (Van Rijn et al. 2006; Lucchetti and Gray 1988). Nitrification involves a two-step process mediated by two different genera of bacteria which drives their energy from the oxidation of ammonia-N and nitrite-N. Nitrosomonas are responsible for the oxidation of ammonia-N to nitrite-N in the first step. Nitrobacter are the other group of bacteria that subsequently oxidize nitrite-N less toxic nitrate-N. Both nitrifying bacteria are strictly aerobic. The
39 nitrate-N can accumulates unless denitrification occurs. Denitrification is a biological reduction of nitrate (NO3-N) to nitrogen gas (N2), which is mostly accomplished by facultative anaerobic heterotrophic bacteria in some systems (Metcalf and Eddy 2003; Van Rijn et al. 2006; Ebeling et al. 2006).
The biofilters ability to remove ammonia-N is defined in terms of volumetric TAN removal rate (VTR). Equation 14 Monod relationship supporting by Michaelis-Menton kinetics reaction can be used as a model for ammonia-N removal (Malone and Pfeiffer 2006; Malone and Beecher
2000; Zhu and Chen 2002; Saidu 2009)
(푇퐴푁−푇퐴푁0) 푉푇푅 = 푉푇푅푚푎푥 ∗ (14) 퐾퐴 + (푇퐴푁−푇퐴푁0)
3 Where VTR = volumetric TAN conversion rate (g-N/m media-day)
3 VTRmax = maximum specific rate of TAN utilization (g-N/m media-day)
3 KA= the half saturation constant (g-N/m )
TAN = TAN concentration in the water (g-N/m3)
TAN0 = minimum TAN concentration for nitrification biofilter
Malone et al (2005) represented a new term, “τ”, for the slope of the linear regression for the plot between TAN concentration and VTR under the assumption of zero intercept. Equation 14 can be simplified to:
푉푇푅 = τ *TAN (15)
Although the aerobic nitrification process has been commonly used for ammonia-N removal from RAS, the approach is limited by the pH and low substrate encountered in acidic oligotrophic ornamental fish systems. (Reeves 1972; Prosser 1989; Metcalf and Eddy 2003;
40
Sharma and Ahlert 1977; Sawyer et al. 1971; van Rijn and Rivera 1990). Acidic water can dramatically declines the biofilter nitrification capacity (Szwerinski et al. 1986; Nogueira et al.
2002; Chen et al. 2005). Zhu and Chen (1999) reported that the TAN and nitrite-N levels should be higher than 0.07 ± 0.05 g-N/m3 @27 ° C to prevent nitrification inhibition. These issues can be addressed using aerobic assimilation strategy for ammonia-N removal via heterotrophic bacterial growth, and remove the sludge from the system. Heterotrophic bacteria. Ammonia-N can be directly assimilated from the RAS into bacterial biomass by the addition of carbonaceous compounds. The ammonia-N can be removed only if the biosolids were taken out of the system.
Heterotrophic bacteria uses organic carbon sources as an electron donor which are usually provided by dosing soluble organic carbon sources such as methanol, ethanol, glucose or other carbohydrates (Avnimelech 1999; Zhu and Chen 1999; Losordo and Hobbs 2000). Although the soluble carbon sources are readily available for bacteria, the dosing based system requires expensive control systems to avoid the overdosing and prevent the releasing of soluble carbon source to the fish tank (Boley et al. 2000; Kaiser and Schmitz 1988; Lee et al. 2000).
A family of solid insoluble biodegradable polymer, polyhydroxyalkanoates (PHAs), can be used for nitrogen removal through assimilation process. Bacteria accumulate PHAs as a carbon and energy storage (Dawes 1990; Steinbüchel et al. 1992; Yu 2001; Du and Yu 2002; Shang et al.
2003). Polyhydroxybutyrate (PHB) is the simplest form of PHA that has been previously used as an organic carbon source for nitrate-N assimilation through anaerobic denitrification process
(Boley and Mller 2005; Boley et al. 2000; Lee et al. 2000; Gutierrez-Wing et al. 2007).
Ammonia-N can be completely removed from aquaculture system through a three-step aerobics assimilation process using insoluble organic carbon sources (PHAs). The three steps are
41
(1) The bacteria grow on the surface of the PHAs converting ammonia-N into the biomass, (2) The bead is mobilized in order to detach the excessive biosolids that trapped between the PHAs, and
(3) The excessive biomass is removed using a solid removal system due to avoids the internal nitrogen cycling in the biofilter column. The main goal of this study was to affirm the ammonia-
N removal through aerobic assimilation process using biodegradable PHAs.
4.3 Materials and methods In this study, two sets of experiments were run, which are: (1) Validate the aerobic assimilation biofiter performance using biodegradable PHA under 7 different continuous nitrogen loading regime, and (2) Determine the optimum backwash frequency based on the oxygen uptake rates of the bacteria.
4.3.1 Experiment 1: Continuous study of assimilation strategy Two identical systems were constructed to accelerate the experimental run. Each setup was comprised of a lab-scale expandable fixed-film bioreactor, a water tank, a PolyGeyserR bead filter and a peristaltic pump connected to a nutrients tank (Figure 13). The up-flow biofilter was constructed of 4-inch ID and 5-foot height acrylic PVC pipe. The biofilter was loaded with 1 liter volume of the PHB beads with the unexpanded bed height of 5 inches. PHBs used in this study were supplied by Metabolix Inc. (USA) with the average size of 1-1.5 mm. The PolyGeyserR was made of 4-inch ID clear PVC pipe and 25-inch height filled with 5-inch deep plastic floating media having an average size of 2-3 mm. The up-flow biofilter was connected to a 50 L volume water tank filled with 2 days aged tap water @ pH 8. A 1/50 HP centrifugal pump (Little giant 3X-MDX) was used to circulate water from the reservoir tank to the biofilter and then to the PolyGeyserR bead filter.
42
Figure 13: Schematic diagram of experimental set-up continuous loading system using low rate peristaltic pump.
Ammonia-N was assimilated by bacteria as the water passed through the PHB media. The bacteria used ammonia-N as a nitrogen source to grow on the surface of the PHBs producing biomass. The bacterial growth in the reactor was controlled by automatically pulsing air into the bottom of the PHB bed every 4 hours for 1 minutes using an air pump connected to an automatic controller. The bacterial biomass released was captured by PolyGeyserR as the water flowed through the floating bead.
Air was injected to the PolyGeyserR air chamber using aquarium air pump. Once the air chamber was filed, the air was released through the floating bead to detach the excess biofilm. The air injection was set at the desired rate providing backwash frequency of every 20-min. The sludge
43 was settled into the sludge chamber and removed daily through the sludge valve located at the bottom of the PolyGeyserR (Malone and Gudipati 2005).
Temperature was set at 28 ±1 °C using a heater in the water tank to accelerate the ammonia-
N removal rate. The flow rate was set at 4 ± 0.5 lpm, which provides approximately 115 recirculating passes per tank volume per day. Dissolved oxygen was maintained above 3 mg/l after the PHB bed during the experimental run using aquarium air pump connected to an air stone to inject air into the bottom of the water tank. A nutrient solution containing 5 mg/l ammonia chloride
(NH4Cl) and 15 mg/l dipotassium phosphate (K2HPO4) was added to the water tank to encourage the heterotrophic bacterial growth. The other nutrients are assumed to be available in the tap water.
The air pump was turned off to stop the pulsing and accelerate the bacterial growth. After acclimation, the desired amount of ammonia chloride and dipotassium phosphate were dissolved in a 5 gallon bucket and continuously pumped to the water tank by using a peristaltic pump (CHEM
- TECH model CTPD2HSA – PAP1) set at 4 mL/min (Figure 14). For each data point, the TAN loading was maintained for at least 24 hours allowing the system to reach to steady state condition giving more stable results. The experiment was run in triplicate for 7 different loading rates at 750,
650, 550, 450, 350, and 250 mg-N/day. The experiments were run from the high to low loading rates. Samples were collected every hour for the entire 8 hours. Each sample was taken from the water tank in a 50 ml glass vials.
Dissolved oxygen and pH were measured by DO meter (YSI model 85 D) and pH meter
(YSI pH 100 meter). Sample were analyzed for ammonia-N, nitrite-N, nitrate and chemical oxygen demand (COD). Nitrate-N concentrations were measured for the first and the last samples to see if there was any accumulation of nitrate-nitrogen in the system. Ammonia-N, nitrite-N and nitrite/nitrate were measured using a SmartChem 170 Discrete Analyzer (Unity Scientific Inc.)
44 based on EPA methods 350.1, 354.1, and 353.2. COD was measured using TNT 821 vial (HACH
Inc.), which was automatically read by spectrophotometer DR 6000 (HACH Inc.) based on EPA method 5220 D.
5gallon nutrient bucket
Peristaltic pump
Figure 14: Peristaltic pump connected to a 5 gallon plastic bucket containing nutrient substrate (NH4Cl and K2HPO4).
Between each experimental run, the reactor cylinder, water tank, screens and all the pipes were sanitized in order to minimize the bacterial growth outside of the bioreactor. PHBs were taken out of the reactor. The reactor cylinder, water tank, screens and the pipes were washed with diluted bleach (1250 ppm). The chlorine was flushed out of the system with tap water. The reactor was rinsed with deionized water before loading the beads. The bioreactor was reloaded with the same acclimated beads after cleaning.
45
4.3.2 Experiment 2: Air pulsing frequency under continuous loading regime A huge amount of biofilm was produced through the aerobic assimilation process due to the high reproduction rates of heterotrophic bacteria. Nitrogen cannot be removed from the system unless the bacterial biomass removed from the water. The excessive biofilm trapping between the
PHBs can cause a bed clogging. Once the biofilm reached a certain level, the media with the biofilm began to rise to the top of the column. If the air was not pulsed through the PHB bed, the biofilm cover the outer layer of the PHB grains preventing the diffusion of oxygen and nutrients into bacteria in the inner layer. Air was periodically inject through the PHB media from the bottom of the bioreactor. The air pulsing expanded the packed bed 30 to 50% which allowed the excessive biofilm to be detached from the PHBs and flow through the solid removal section (PolyGeyserR).
This experiment was run to define the optimum air pulsing frequency, which is an important parameter in the biofilter design.
The experimental set-up was the same as experiment 1. TAN loading was maintained at
350 mg-N/day for at least 24 hours allowing the system to reach to steady-state condition. Air was injected from the bottom of the bioreactor for the entire 1 minutes to agitate and knock off the bacteria growing on the surface of the PHBs.
Samples were collected from the bottom of the water tank (right before the PHBs bed) and right after the PHB bed in 300 ml BOD bottles. The BOD bottles were submerged in the water and immediately capped while they were submerged to avoid any air trapping in the bottle. Experiment ran in triplicate. Samples were taken at 2-hour intervals for 8 hours and were analyzed immediately using Winkler test to determine the dissolved oxygen (DO) concentration of water before and after the biofilters.
46
4.4 Results and discussion
4.4.1 Continuous study Experiment 1 was performed using a bench-scale laboratory set-up to test the assimilation biofilter performance under different continuous TAN loading regimes. The temperature and pH were maintained at 28 ±1 and 7.92 ± 0.1, respectively. The observed data didn’t show any significant increase of nitrite-N and nitrate-N concentrations. This might be due to high heterotrophic bacterial growth which inhibits the nitrification process by taking the available oxygen and nutrients. The mean value of COD for all dosing rates was 15.14 with the standard deviation of 5.43 showing no apparent release of organic compounds into the water tank (Appendix
D). The mean observed TAN concentrations were plotted in Figure 15. The VTR range was from
250 to 750 g-N/m3-day across the TAN concentration of 0.04 to 1.2 g/m3.
800
700
600 y = 602.4x R² = 0.4639
500
DAY)
- 3 400
300 VTR (G/M VTR 200
100
0 0 0.2 0.4 0.6 0.8 1 1.2 1.4 TAN CONCENTRATION (G/M3)
Figure 15: Triplicated experiments run of Volumetric TAN removal rates (VTR) for aerobic assimilation biofilters under different continuous TAN loading.
47
The relationship between VTR and TAN concentration in the continuous regimes did not follow Michaelis-Menton kinetics. Although the data were fitted the linear zero intercept curve, the R2=0.46 indicates that the behavior is clearly non-linear. This data set was not complied the assumption of the theory with τ. The non-linear behavior might be interpreted that the bacteria adjusted better to the loading having more time to get to the population in the continuous loading regime.
The bacteria was aggressively grow after they have been acclimated. Figure 16 demonstrates the huge amount of biofilm produced and attached to the surface of the beads after running 2 days experiment under continuous loading rate of 350 mg TAN /day.
Figure 16: Biofilm growth on the surface of the PHB bead after 2 days experimental run under continuously TAN dosing of 350 mg/day (photo wastaken by Leica MZ7 stereo microscope with the total magnification of 63x).
48
The photo was taken by a stereomicroscope, which show the PHB beads were completely coated by the tough biofilms in the water. Biofilm production through aerobic assimilation process was much higher than nitrification due to higher growth rate of the heterotrophic bacteria in comparison to nitrifiers.
The two SEM micro photos (Figure 17 & 18) were taken after 2 days of experimental run under continuously Tan loading 350 mg-N/day. The photos show a huge amount of microorganisms growing on the surface of the beads. Figure 17 particularly shows the high amount of bacterial population on the surface of the PHB.
Figure 17: SEM micro photo of the biofilm growing on the surface of the PHB beads after 2 days continuous experimental run showing the bacterial growth.
49
Figure 18: SEM micro photo sowing the microbial growth on the surface of PHB after 2 days experimental run.
4.4.2 Air pulsing frequency The air pulsing through the PHB bed and the biosolids removal are the important factors which affect the aerobic assimilation biofilter performance. The low air pulsing frequency increases the residence time of the biofilm in the PHB bed. Once the bacteria died and decayed within the bed, they released nitrogen into the water impairing the biofilter performance. In the other hand, the high pulsing frequency can aggressively abrade the biofilm thus the bacteria doesn’t have enough time to grow and thicken on the surface of the beads for the optimum biofiltering. The amount of oxygen consumed in filtration (OCF) used to determine the best time intervals for air pulsing (Malone and Beecher 2000). Assuming the majority of the bacterial population exist in the biofilter, Equation 15 can be used to calculate OCF (Manthe et al. 1988).
This equation was derived based on the mass balance of dissolved oxygen. 50
푂퐶퐹 = 푄 ∗ (퐶푖 − 퐶푂) (16)
In equation (16), OCF = the amount of oxygen consumed in filtration (mg-O/day)
Q = the flow rate through the biofilter (l/day)
Ci=the oxygen concentration of the influent biofilter (mg-O/l)
Ci=the oxygen concentration of the effluent biofilter (mg-O/l).
Oxygen consumption rates were calculated from the observed data using Equation 16. The
OCFs were plotted versus time in hour in Figure 19.
3 2 6 Y = -0.003X - 0.014X + 0.5308X + 3.3205 R² = 0.8398 First Run 5 second run y = -0.0051x3 + 0.0551x2 - 0.0097x + 2.6268 R² = 0.5254
4 third run
DAY) -
3 Poly. (First 3 Run) Poly. (second run) 2 OCF OCF (G/M y = -0.0026x3 + 0.0037x2 + 0.1696x + 2.7553 Poly. (third run) R² = 0.5945 1
0 0 2 4 6 8 10 12 TIME (HR) Figure 19: OCF with the optimum air pulsing frequency of 6.18 ± 0.98 hr. at 350 mg-N/day continuous loading rate.
The data used for the Figure 19 were provided in Appendix E. The OCFs represent the bacteria respiration which is increased from the start-up, reached at the maximum point and then started to decrease. It was assumed that the optimum growth rate of the bacteria occurred at the peak OCF. 51
The time corresponding to the maximum OCF points can be used as the optimum air pulsing frequency. Considering the coefficient of determination (R2), the best curve fitting to the observed data was a third-degree polynomials. The equations showed on the figure 19 were solved to obtain the maximum OCF values. The maximum OCFs were obtained at 6.28, 7.11 and 5.16 hours with the mean value of 6.18 ± 0.98 hour. This can be interpreted that the optimum time for air pulsing was about every 6-hour for the optimum effective biofilm carrying capacity.
4.5 Conclusions Aerobic assimilation strategy for ammonia-N removal was tested and verified under different continuous nitrogen loading regimes. The relationship between VTR and TAN concentration didn’t follow the Michaelis-Menton kinetics equation. The relationship behavior was not linear and didn’t fit to the zero intercept assumption for τ strategy. However, the data showed at low TAN concentration regime (oligotrophic area), the VTR for continuous systems was much higher in comparison to the batch systems. The mean COD was 15.14 ± 5.43 that showing no apparent release of organic compounds into the water tank (Appendix D). The optimum air pulsing frequency through the PHB bed was found to be 6-hour for the continuous
TAN loading of 350 mg-N/day.
52
CHAPTER 5. SUMMARY AND OUTLOOK
5.1 Experimental findings
An aerobic assimilation strategy using insoluble carbon source for ammonia-N removal from
recirculating aquaculture systems via heterotrophic bacteria was examined with batch and
continuous loading lab scale systems. The aerobic assimilation processes can be divided to
three steps, they are : (1) converting the ammonia-N into biomass, (2) releasing of the biofilm
by an air burst through the biofilter bed, and (3) removing the released biosolids using a solid
capture system.
The study verified that aerobic assimilation approach can be successfully applied for ammonia-
N removal from freshwater RAS for both batch and continuous systems. The acclimation time
of the aerobic assimilation biofilter was found to be relatively fast (2-3 days).
The statistical analysis results showed no significant difference between the VTRs obtained
at pH 8 and pH 6.5 (p>0.05).
The kinetics equations that can be used to design a biofilter for ammonia-N removal through
aerobic assimilation process based on the related RAS categories were summarized as: 1)
VTR=1822 * TAN (oligotrophic), 2) VTR=1297 * TAN (mesotrophic) and 3) VTR=594 ± 93
(eutrophic).
The consumption rate of PHB through aerobic assimilation process was approximately 22 g
PHB/ g of ammonia-N removed from water.
By analyzing the chemical oxygen demand (COD) data for all experiments, this study revealed
that there was no apparent release of carbon source into the water tank, when employing the
PHB beads.
The continuous study verified the use of assimilation strategy for TAN removal. .
53
The optimum time for air pulsing frequency were obtained at 6-hour for the optimum effective
biofilm carrying capacity.
Although the market prices of soluble carbon source such as methanol and ethanol are
relatively lower than the PHB, the use of bioplastic offers a self-control system which
decreases the overall cost by eliminating the needs for sophisticated control dosing systems.
Although the consumption rate of PHB is higher under aerobic than anaerobic conditions, the
aerobic process avoids producing the toxic sulfide, which is the main issue associated with
anaerobic denitrification processes.
The solid based system using PHB is a cost effective process for larval system due to the low
nitrogen generation. Although the cost of PHBs is found to be an issue of using the bioplastic
as a carbon source for a long term aquaculture systems, the use of PHB can be reasonable for
a few weeks as each batch of eggs is first hatched then reared as a larvae.
The PHB market price found to be $10.1 and $2.11 in 2000 and 2010, respectively. This
change in price shows that the cost of PHB cost has dropped significantly over the last decade.
The rapid price reduction of the PHB might make the solid-based assimilation more
economically feasible.
5.2 Recommendations for further research Future study can be done on the feasibility of the aerobic assimilation strategy for salt water
RAS. The study for defining the effect of pH on aerobic assimilation rates can be extended in order to determine the lowest pH level that assimilation can be run successfully. The air pulsing frequency may be optimized for different loading rates using an unexpanded biofilter. Different kind of PHB can be tested in order to determine the effect of the PHB characterization on the biofilter performance.
54
The OCFs were tested under continuous loading rate of 350 mg-N/m3.d. The further study is recommended to run the experiment under different TAN loading regimes to optimize the air pulsing frequency. The solid removal system should be optimized in order to minimize the sludge retention time (SRT). After optimizing the design criteria and the kinetics constants, the biofilter should be tested on live animals for a commercial scale system.
55
REFERENCES
Ackermann, J.-U., and W. Babel. 1998. Approaches to increase the economy of the PHB production. Polymer degradation and stability 59 no. 1: 183-186.
Akhbari, A., A. Zinatizadeh, P. Mohammadi, M. Irandoust, and Y. Mansouri. 2011. Process modeling and analysis of biological nutrients removal in an integrated RBC-AS system using response surface methodology. Chemical Engineering Journal 168 no. 1: 269-279.
Akunna, J.C., C. Bizeau, and R. Moletta. 1993. Nitrate and nitrite reductions with anaerobic sludge using various carbon sources: glucose, glycerol, acetic acid, lactic acid and methanol. Water Research 27 no. 8: 1303-1312.
Anderson, A.J., and E.A. Dawes. 1990. Occurrence, metabolism, metabolic role, and industrial uses of bacterial polyhydroxyalkanoates. Microbiological reviews 54 no. 4: 450-472.
Andrews, C. 1990. The ornamental fish trade and fish conservation. Journal of fish Biology 37 no. sA: 53-59.
Ashby, R.D., D.K. Solaiman, T.A. Foglia, and C.-K. Liu. 2001. Glucose/lipid mixed substrates as a means of controlling the properties of medium chain length poly (hydroxyalkanoates). Biomacromolecules 2 no. 1: 211-216.
Aslan, S., and I.K. Kapdan. 2006. Batch kinetics of nitrogen and phosphorus removal from synthetic wastewater by algae. Ecological Engineering 28 no. 1: 64-70.
Avnimelech, Y. 1999. Carbon/nitrogen ratio as a control element in aquaculture systems. Aquaculture 176 no. 3: 227-235.
Barimo, J.F., and P.J. Walsh. 2005. The effects of acute and chronic ammonia exposure during early life stages of the gulf toadfish,< i> Opsanus beta. Aquatic toxicology 75 no. 3: 225-237.
Biesterfeld, S., G. Farmer, P. Russell, and L. Figueroa. 2003. Effect of alkalinity type and concentration on nitrifying biofilm activity. Water environment research: 196-204.
Boley, A., and W. Mller. 2005. Denitrification with polycaprolactone as solid substrate in a laboratory-scale recirculated aquaculture system. Water Science & Technology 52 no. 10- 11: 495-502.
Boley, A., W.-R. Müller, and G. Haider. 2000. Biodegradable polymers as solid substrate and biofilm carrier for denitrification in recirculated aquaculture systems. Aquacultural engineering 22 no. 1: 75-85.
Boller, M., and W. Gujer. 1986. Nitrification in tertiary trickling filters followed by deep-bed filters. Water Research 20 no. 11: 1363-1373.
Boyd, C.E. 1998. Pond aquaculture water quality management: Springer.
56
Brandi, H., R. Bachofen, J. Mayer, and E. Wintermantel. 1995. Degradation and applications of polyhydroxyalkanoates. Canadian journal of Microbiology 41 no. 13: 143-153.
Brune, D., K. Kirk, and A. Eversole. 2004. Autotrophic intensification of pond aquaculture; shrimp production in a partitioned aquaculture system. In Proceedings of the Fifth International Conference on Recirculating Aquaculture, Roanoke, Virginia, 201-210.
Burford, M.A., P.J. Thompson, R.P. McIntosh, R.H. Bauman, and D.C. Pearson. 2003. Nutrient and microbial dynamics in high-intensity, zero-exchange shrimp ponds in Belize. Aquaculture 219 no. 1: 393-411.
Castilho, L.R., D.A. Mitchell, and D.M. Freire. 2009. Production of polyhydroxyalkanoates (PHAs) from waste materials and by-products by submerged and solid-state fermentation. Bioresource technology 100 no. 23: 5996-6009.
Chanprateep, S. 2010. Current trends in biodegradable polyhydroxyalkanoates. Journal of bioscience and bioengineering 110 no. 6: 621-632.
Chapman, F.A., S.A. Fitz‐Coy, E.M. Thunberg, and C.M. Adams. 1997. United States of America trade in ornamental fish. Journal of the World Aquaculture Society 28 no. 1: 1- 10.
Chen, S., D.E. Coffin, and R.F. Malone. 1997. Sludge production and management for recirculating aquacultural systems. Journal of the World Aquaculture Society 28 no. 4: 303-315.
Chen, S., J. Ling, and J.-P. Blancheton. 2006. Nitrification kinetics of biofilm as affected by water quality factors. Aquacultural Engineering 34 no. 3: 179-197.
Chen, X.-l., P. Zheng, R.-c. Jin, S.-x. Zhou, and G.-s. Ding. 2005. Effect of pH on nitrification. Journal of Zhejiang University(Agriculture and Life Sciences) 31 no. 6: 755-759.
Choi, J.-i., and S.Y. Lee. 1997. Process analysis and economic evaluation for poly (3- hydroxybutyrate) production by fermentation. Bioprocess Engineering 17 no. 6: 335-342.
Colt, J. 2006. Water quality requirements for reuse systems. Aquacultural Engineering 34 no. 3: 143-156.
Crab, R., Y. Avnimelech, T. Defoirdt, P. Bossier, and W. Verstraete. 2007. Nitrogen removal techniques in aquaculture for a sustainable production. Aquaculture 270 no. 1: 1-14.
Dawes, E.A. 1990. Novel biodegradable microbial polymers.
De Smet, M., G. Eggink, B. Witholt, J. Kingma, and H. Wynberg. 1983. Characterization of intracellular inclusions formed by Pseudomonas oleovorans during growth on octane. Journal of bacteriology 154 no. 2: 870-878.
57
Domb, A.J., and N. Kumar. 2011. Biodegradable polymers in clinical use and clinical development: John Wiley & Sons.
Du, G., and J. Yu. 2002. Green technology for conversion of food scraps to biodegradable thermoplastic polyhydroxyalkanoates. Environmental science & technology 36 no. 24: 5511-5516.
Ebeling, J.M., M.B. Timmons, and J. Bisogni. 2006. Engineering analysis of the stoichiometry of photoautotrophic, autotrophic, and heterotrophic removal of ammonia–nitrogen in aquaculture systems. Aquaculture 257 no. 1: 346-358.
Eding, E., A. Kamstra, J. Verreth, E. Huisman, and A. Klapwijk. 2006. Design and operation of nitrifying trickling filters in recirculating aquaculture: a review. Aquacultural engineering 34 no. 3: 234-260.
Emerson, K., R.C. Russo, R.E. Lund, and R.V. Thurston. 1975. Aqueous ammonia equilibrium calculations: effect of pH and temperature. Journal of the Fisheries Board of Canada 32 no. 12: 2379-2383.
Ester, C., J. Novak, G. Libey, and G. Boardman. 1994. Rotating biological contactor performance in recirculating aquaculture systems. Original not available for examination.
Fitch, M.W., N. Pearson, G. Richards, and J.G. Burken. 1998. Biological fixed-film systems. Water environment research: 495-518.
Gutierrez-Wing, M.T., and R.F. Malone. 2006. Biological filters in aquaculture: trends and research directions for freshwater and marine applications. Aquacultural Engineering 34 no. 3: 163-171.
Gutierrez-Wing, M.T., K.A. Rusch, and R.F. Malone. 2007. Polyhydroxyalkanoates as a Carbon Source for Denitrification of Waters. In World Environmental and Water Resources Congress 2007@ sRestoring Our Natural Habitat, 1-13. ASCE.
Hagopian, D.S., and J.G. Riley. 1998. A closer look at the bacteriology of nitrification. Aquacultural engineering 18 no. 4: 223-244.
Halachmi, I. 2006. Systems engineering for ornamental fish production in a recirculating aquaculture system. Aquaculture 259 no. 1: 300-314.
Hankermeyer, C.R., and R.S. Tjeerdema. 1999. Polyhydroxybutyrate: plastic made and degraded by microorganisms. In Reviews of environmental contamination and toxicology, 1-24. Springer.
Hearn, R. 2009. Gas transfer in air-lifts used to recirculate aquaculture systems, Faculty of the Louisiana State University and Agricultural and Mechanical College in partial fulfillment of the requirements for the degree of Master of Science in Civil Engineering in The
58
Department of Civil and Environmental Engineering by Ryan Hearn BS Environmental Engineering, Louisiana State University.
Henze, M. 2002. Wastewater treatment: biological and chemical processes: Springer.
Hill, J.E., and R.P. Yanong. 2010. Freshwater ornamental fish commonly cultured in Florida.
Hiraishi, A., and S. Khan. 2003. Application of polyhydroxyalkanoates for denitrification in water and wastewater treatment. Applied microbiology and biotechnology 61 no. 2: 103- 109.
Hochheimer, J.N., and F. Wheaton. 1998. Biological filters: trickling and RBC design. In Proceedings of the Second International Conference on Recirculating Aquaculture, Roanoke, VA, 16-19.
Ip, Y., S. Chew, and D. Randall. 2001. Ammonia toxicity, tolerance, and excretion. Fish physiology 20: 109-148.
Joye, S.B., and J.T. Hollibaugh. 1995. Influence of sulfide inhibition of nitrification on nitrogen regeneration in sediments. Science 270 no. 5236: 623-625.
Kaiser, H., and O. Schmitz. 1988. Water quality in a closed recirculating fish culture system influenced by addition of a carbon source in relation to feed uptake by fish. Aquaculture Research 19 no. 3: 265-273.
Kawaguchi, Y., N. Koyama, S. Nakamura, M. Hiramitsu, Y. Yoshida, and H. Kimura. 1992. Synthesis and degradation of polyhydroxyalkanoates in Alcaligenes eutrophus. FEMS Microbiology Letters 103 no. 2‐4: 103-108.
Körner, S., S.K. Das, S. Veenstra, and J.E. Vermaat. 2001. The effect of pH variation at the ammonium/ammonia equilibrium in wastewater and its toxicity to< i> Lemna gibba. Aquatic botany 71 no. 1: 71-78.
Larkin, S.L., and R.L. Degner. 2001. The US wholesale market for marine ornamentals. Aquarium Sciences and Conservation 3 no. 1-3: 13-24.
Lazur, A.M., and D.C. Britt. 1997. Pond recirculating production systems: Southern Regional Aquaculture Center.
Lee, P.G., R. Lea, E. Dohmann, W. Prebilsky, P. Turk, H. Ying, and J. Whitson. 2000. Denitrification in aquaculture systems: an example of a fuzzy logic control problem. Aquacultural Engineering 23 no. 1: 37-59.
Lee, S.Y., and H.N. Chang. 1995. Production of poly (hydroxyalkanoic acid). In Microbial and Eznymatic Bioproducts, 27-58. Springer.
59
Lee, S.Y., J.-i. Choi, and H.H. Wong. 1999. Recent advances in polyhydroxyalkanoate production by bacterial fermentation: mini-review. International journal of biological macromolecules 25 no. 1: 31-36.
Lekang, O.-I., and H. Kleppe. 2000. Efficiency of nitrification in trickling filters using different filter media. Aquacultural engineering 21 no. 3: 181-199.
Lemoigne, M. 1926. Produits de dehydration et de polymerisation de l’acide ß-oxobutyrique. Bull Soc Chim Biol 8: 770-782.
Lenz, R.W., and R.H. Marchessault. 2005. Bacterial polyesters: biosynthesis, biodegradable plastics and biotechnology. Biomacromolecules 6 no. 1: 1-8.
Lin, Y.-F., S.-R. Jing, T.-W. Wang, and D.-Y. Lee. 2002. Effects of macrophytes and external carbon sources on nitrate removal from groundwater in constructed wetlands. Environmental pollution 119 no. 3: 413-420.
Livengood, E., and F. Chapman. 2007. The ornamental fish trade: An introduction with perspectives for responsible aquarium fish ownership. University of Florida IFAS Extension.
Losordo, T., and H. Westers. 1994. System carrying capacity and flow estimation. Developments in aquaculture and fisheries science 27: 9-9.
Losordo, T.M., and A.O. Hobbs. 2000. Using computer spreadsheets for water flow and biofilter sizing in recirculating aquaculture production systems. Aquacultural engineering 23 no. 1: 95-102.
Lucchetti, G.L., and G.A. Gray. 1988. Water reuse systems: a review of principal components. The Progressive Fish-Culturist 50 no. 1: 1-6.
Malone, R., and S. Gudipati. 2005. Polygeyser airlift filtration of high density recirculating aquaculture systems. Fort Pierce: The 2nd International Sustainable Marine Fish Culture Conferenee.
Malone, R.F., and L.E. Beecher. 2000. Use of floating bead filters to recondition recirculating waters in warmwater aquaculture production systems. Aquacultural Engineering 22 no. 1: 57-73.
Malone, R.F., and D.G. Burden. 1988. Design of recirculating soft crawfish shedding systems: Louisiana Sea Grant College Program, Center for Wetland Resources, Louisiana State University.
Malone, R.F., and T.J. Pfeiffer. 2006. Rating fixed film nitrifying biofilters used in recirculating aquaculture systems. Aquacultural engineering 34 no. 3: 389-402.
60
Malone, R.F., K.A. Rusch, and D.G. Burden. 1990. Kemp's Ridley sea turtle waste characterization study: precursor to a recirculating holding system design. Journal of the World Aquaculture Society 21 no. 2: 137-144.
Manthe, D.P., R.F. Malone, and S. Kumar. 1988. Submerged rock filter evaluation using an oxygen consumption criterion for closed recirculating systems. Aquacultural Engineering 7 no. 2: 97-111.
Martins, C., E. Eding, M. Verdegem, L. Heinsbroek, O. Schneider, J.-P. Blancheton, E.R. d’Orbcastel, and J. Verreth. 2010. New developments in recirculating aquaculture systems in Europe: A perspective on environmental sustainability. Aquacultural Engineering 43 no. 3: 83-93.
Masser, M.P., J. Rakocy, and T.M. Losordo. 1999. Recirculating aquaculture tank production systems. Management of recirculating systems. SRAC Publication 452.
McIntosh, R. 2001. High rate bacterial systems for culturing shrimp. In Proceedings from the Aquacultural Engineering Society's 2001 Issues Forum. Shepherdstown, West Virginia, USA. Aquaculture Engineering Society, 117-129.
Meade, J.W. 1985. Allowable ammonia for fish culture. The Progressive Fish-Culturist 47 no. 3: 135-145.
Metcalf, I., and H. Eddy. 2003. Wastewater engineering; treatment and reuse.
Michaud, L., J.-P. Blancheton, V. Bruni, and R. Piedrahita. 2006. Effect of particulate organic carbon on heterotrophic bacterial populations and nitrification efficiency in biological filters. Aquacultural Engineering 34 no. 3: 224-233.
Miller-Morgan, T. 2010. A brief overview of the ornamental fish industry and hobby. Fundamentals of ornamental fish health. Blackwell, USA: 25-32.
Nath, A., M. Dixit, A. Bandiya, S. Chavda, and A. Desai. 2008. Enhanced PHB production and scale up studies using cheese whey in fed batch culture of< i> Methylobacterium sp. ZP24. Bioresource technology 99 no. 13: 5749-5755.
Nazar, A.A., R. Jayakumar, and G. Tamilmani. 2013. Recirculating aquaculture systems.
Nogueira, R., L.s.F. Melo, U. Purkhold, S. Wuertz, and M. Wagner. 2002. Nitrifying and heterotrophic population dynamics in biofilm reactors: effects of hydraulic retention time and the presence of organic carbon. Water Research 36 no. 2: 469-481.
Ohashi, A., D. Viraj de Silva, B. Mobarry, J.A. Manem, D.A. Stahl, and B.E. Rittmann. 1995. Influence of substrate C/N ratio on the structure of multi-species biofilms consisting of nitrifiers and heterotrophs. Water Science and Technology 32 no. 8: 75-84.
Ojumu, T., J. Yu, and B. Solomon. 2004. Production of polyhydroxyalkanoates, a bacterial biodegradable polymers. African Journal of Biotechnology 3 no. 1: 18-24.
61
Peoples, O.P., and A.J. Sinskey. 1989. Poly-beta-hydroxybutyrate (PHB) biosynthesis in Alcaligenes eutrophus H16. Identification and characterization of the PHB polymerase gene (phbC). Journal of Biological Chemistry 264 no. 26: 15298-15303.
Philip, S., T. Keshavarz, and I. Roy. 2007. Polyhydroxyalkanoates: biodegradable polymers with a range of applications. Journal of Chemical Technology and Biotechnology 82 no. 3: 233-247.
Prosser, I. 1989. Autotrophic nitrification in bacteria. Adv. Microb. Physiol 20: 125-192.
Quillaguamán, J., T. Doan-Van, H. Guzmán, D. Guzmán, J. Martín, A. Everest, and R. Hatti- Kaul. 2008. Poly (3-hydroxybutyrate) production by Halomonas boliviensis in fed-batch culture. Applied microbiology and biotechnology 78 no. 2: 227-232.
Reddy, C., R. Ghai, and V.C. Kalia. 2003. Polyhydroxyalkanoates: an overview. Bioresource technology 87 no. 2: 137-146.
Reeves, T.G. 1972. Nitrogen removal: a literature review. Journal (Water Pollution Control Federation): 1895-1908.
Reinbold, K.A., and S.M. Pescitelli. 1982. Effects of exposure to ammonia on sensitive life stages of aquatic organisms.
Russo, R.C., and R.V. Thurston. 1991. Toxicity of ammonia, nitrite, and nitrate to fishes. Aquaculture and water quality: 58-89.
Rusten, B., B. Eikebrokk, Y. Ulgenes, and E. Lygren. 2006. Design and operations of the Kaldnes moving bed biofilm reactors. Aquacultural engineering 34 no. 3: 322-331.
Saidu, M. 2009. Temperature impact on nitrification and bacterial growth kinetics in acclimating recirculating aquacultural systems. A Dissertation. Louisiana State University, Baton Rouge, LA, United States.
Salit, M.S. 2014. Biopolymer. In Tropical Natural Fibre Composites, 39-52. Springer.
Sandu, S.I., G.D. Boardman, B.J. Watten, and B.L. Brazil. 2002. Factors influencing the nitrification efficiency of fluidized bed filter with a plastic bead medium. Aquacultural Engineering 26 no. 1: 41-59.
Sastry, B.N., A.A. DeLosReyes Jr, K.A. Rusch, and R.F. Malone. 1999. Nitrification performance of a bubble-washed bead filter for combined solids removal and biological filtration in a recirculating aquaculture system. Aquacultural engineering 19 no. 2: 105- 117.
Satoh, H., S. Okabe, N. Norimatsu, and Y. Watanabe. 2000. Significance of substrate C/N ratio on structure and activity ofnitrifying biofilms determined by in situ hybridization and the use of microelectrodes. Water science and technology 41 no. 4-5: 317-321.
62
Sauthier, N., A. Grasmick, and J. Blancheton. 1998. Biological denitrification applied to a marine closed aquaculture system. Water Research 32 no. 6: 1932-1938.
Sawyer, C.N., H. Wild Jr, and T. McMahon. 1971. Nitrification and Denitrification Facilities. In Design Seminar for Wastewater Treatment Facilities, Kansas City, Mo.
Schneider, O., M. Chabrillon‐Popelka, H. Smidt, O. Haenen, V. Sereti, E.H. Eding, and J.A. Verreth. 2007. HRT and nutrients affect bacterial communities grown on recirculation aquaculture system effluents. FEMS microbiology ecology 60 no. 2: 207-219.
Sedlak, R.I. 1991. Phosphorus and nitrogen removal from municipal wastewater: principles and practice: CRC Press.
Shang, L., M. Jiang, and H.N. Chang. 2003. Poly (3-hydroxybutyrate) synthesis in fed-batch culture of Ralstonia eutropha with phosphate limitation under different glucose concentrations. Biotechnology letters 25 no. 17: 1415-1419.
Sharma, B., and R. Ahlert. 1977. Nitrification and nitrogen removal. Water Research 11 no. 10: 897-925.
Singh, S., J. Ebeling, and F. Wheaton. 1999. Water quality trials in four recirculating aquacultural system configurations. Aquacultural engineering 20 no. 2: 75-84.
Steinbüchel, A., E. Hustede, M. Liebergesell, U. Pieper, A. Timm, and H. Valentin. 1992. Molecular basis for biosynthesis and accumulation of polyhydroxyalkanoic acids in bacteria. FEMS Microbiology Letters 103 no. 2‐4: 217-230.
Stenstrom, M.K., and R.A. Poduska. 1980. The effect of dissolved oxygen concentration on nitrification. Water Research 14 no. 6: 643-649.
Stumm, W., and J.J. Morgan. 2012. Aquatic chemistry: chemical equilibria and rates in natural waters: John Wiley & Sons.
Suthersan, S.S. 1996. Remediation engineering: design concepts: CRC Press.
Szwerinski, H., E. Arvin, and P. Harremoës. 1986. pH-decrease in nitrifying biofilms. Water Research 20 no. 8: 971-976.
Thurston, R.V., R.C. Russo, and G. Vinogradov. 1981. Ammonia toxicity to fishes. Effect of pH on the toxicity of the unionized ammonia species. Environmental science & technology 15 no. 7: 837-840.
Timmons, M.B., and N.R.A. Center. 2002. Recirculating aquaculture systems.
Timmons, M.B., and T.M. Losordo. 1994. Aquaculture water reuse systems: engineering design and management: Elsevier Science Publishers.
63
Trussell, R. 1972. The Percent Un-Ionized Ammonia in Aqueous Ammonia Solutions at Different p H Levels and Temperatures. Journal of the Fisheries Board of Canada 29 no. 10: 1505-1507.
Tsuge, T. 2002. Metabolic improvements and use of inexpensive carbon sources in microbial production of polyhydroxyalkanoates. Journal of bioscience and bioengineering 94 no. 6: 579-584.
Van Rijn, J., and G. Rivera. 1990. Aerobic and anaerobic biofiltration in an aquaculture unit— nitrite accumulation as a result of nitrification and denitrification. Aquacultural engineering 9 no. 4: 217-234.
Van Rijn, J., Y. Tal, and H.J. Schreier. 2006. Denitrification in recirculating systems: theory and applications. Aquacultural engineering 34 no. 3: 364-376.
Verlinden, R.A.J., D.J. Hill, M.A. Kenward, C.D. Williams, and I. Radecka. 2007. Bacterial synthesis of biodegradable polyhydroxyalkanoates. Journal of Applied Microbiology 102 no. 6: 1437-1449.
Walsh, P.J., and P.A. Wright. 1995. Nitrogen metabolism and excretion: CRC Press.
Wheaton, F., J. Hochheimer, G. Kaiser, and M. Krones. 1994. Nitrification Principles: Aquaculture and Reuse Systems. Amsterdam: Elsevier.
Whitson, J., P. Turk, and P. Lee. 1993. Biological denitrification in a closed recirculating marine culture system. Techniques for Modern Aquaculture: 458-466.
Williams, S., and O. Peoples. 1996. Biodegradable plastics from plants. Chemtech 26 no. 9: 38- 44.
Wortman, B., and F. Wheaton. 1991. Temperature effects on biodrum nitrification. Aquacultural engineering 10 no. 3: 183-205.
Wright, P.A. 1995. Nitrogen excretion: three end products, many physiological roles. The Journal of Experimental Biology 198 no. 2: 273-281.
Yang, L., L.-S. Chou, and W.K. Shieh. 2001. Biofilter treatment of aquaculture water for reuse applications. Water research 35 no. 13: 3097-3108.
Yanong, R.P. 1996. Reproductive management of freshwater ornamental fish. In Seminars in Avian and Exotic Pet Medicine, vol. 5, 222-235. Elsevier.
Young-Lai, W., M. Charmantier-Daures, and G. Charmantier. 1991. Effect of ammonia on survival and osmoregulation in different life stages of the lobsterHomarus americanus. Marine biology 110 no. 2: 293-300.
Yu, J. 2001. Production of PHA from starchy wastewater via organic acids. Journal of Biotechnology 86 no. 2: 105-112.
64
Zhu, S., and S. Chen. 1999. An experimental study on nitrification biofilm performances using a series reactor system. Aquacultural engineering 20 no. 4: 245-259.
———. 2001. Effects of organic carbon on nitrification rate in fixed film biofilters. Aquacultural Engineering 25 no. 1: 1-11.
———. 2002. The impact of temperature on nitrification rate in fixed film biofilters. Aquacultural Engineering 26 no. 4: 221-237.
Zweifel, U.L., B. Norrman, and A. Hagstrom. 1993. Consumption of dissolved organic carbon by marine bacteria and demand for inorganic nutrients. Marine Ecology-Progress Series 101: 23-23.
65
APPENDIX A: OBSERVED TAN CONCENTRATIONS FROM BATCH SYSTEM @ PH 8 (CHAPTER 3, EXPERIMENT 1)
FIRST RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration concentration Concentration (mg/L) 0 (mg/L)2.56 (mg/L)0.019 0.011(mg/L) 8.7 8.12
1 2.18 0.056 3.58 8.15
2 1.832 0.081 10.2 7.98
3 1.49 0.12 6.44 7.89
4 1.12 0.073 14.2 7.81
5 0.78 0.065 . 8.77 7.68
6 0.41 0.012 6.08 7.64
7 0.1553 0.051 6.8 7.61
8 0.007 0.063 0.051 7.02 7.5
SECOND RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) 0 (mg/L)2.57 (mg/L)0.013 0 .023(mg/L) 7.83 8.05
1 2.2 0.048 8.72 8.01
2 1.85 0.062 9.96 7.95
3 1.38 0.098 12.4 7.88
4 0.99 0.12 11.23 7.75
5 0.62 0.078 9.42 7.71
6 0.33 0.063 6.08 7.63
7 0.051 0.077 6.52 7.61
8 0.01 0.022 0.082 10.2 7.57
66
THIRD RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 2.53 0.032 0.007 8.56 8.02
1 2.16 0.071 9.54 8.07
2 1.75 0.083 8.15 7.93
3 1.47 0.12 5.84 7.85
4 1.14 0.065 10.2 7.72
5 0.81 0.033 15.3 7.64
6 0.531 0.012 10.07 7.62
7 0.2178 0.014 8.73 7.59
8 0.015 0.01 6.02 7.56
9 0.0012 0.007 0.09 7.84 7.53
67
APPENDIX B: OBSERVED TAN CONCENTRATIONS FROM BATCH SYSTEM @ PH 6.5 (CHAPTER 3, EXPERIMENT 1)
FIRST RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 2.47 0.008 0 .012 5.23 6.51
1 2.12 0.014 7.78 6.53
2 1.74 0.032 9.23 6.47
3 1.43 0.041 7.92 6.39
4 1.15 0.067 9.34 6.32
5 0.85 0.078 11.1 6.27
6 0.51 0.092 8.36 6.25
7 0.25 0.061 7.38 6.22
8 0.042 0.031 10.2 6.17
9 0.001 0.022 0.084 6.51 6.18
SECOND RUN TIME Ammonia-N Nitrite-N Nitrate-N COD (HR) concentration Concentration Concentration (mg/L) pH (mg/L) (mg/L) (mg/L) 0 2.56 0.0054 0 .02 3.12 6.50
1 2.22 0.0078 5.14 6.52
2 1.83 0.012 7.43 6.45
3 1.54 0.018 7.98 6.38
4 1.23 0.0096 9.76 6.37
5 0.98 0.0076 8.23 6.31
6 0.63 0.032 9.79 6.29
68
7 0.39 0.021 10.97 6.28
8 0.18 0.009 10.14 6.25
9 0.068 0.012 7.62 6.28
10 0.003 0.0078 0.054 9.54 6.31
THIRD RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 2.53 0.021 0 .0032 6.2 6.50 1 2.21 0.043 7.49 6.51 2 1.92 0.057 7.58 6.48 3 1.67 0.049 8.12 6.44 4 1.42 0.064 9.23 6.36 5 1.12 0.069 8.85 6.31 6 0.87 0.078 10.24 6.27 7 0.54 0.094 8.79 6.22 8 0.31 0.11 9.53 6.19 9 0.11 0.067 7.22 6.21 10 0.042 0.0071 0.098 8.2 6.28
69
APPENDIX C: STATISTICAL ANALYSIS (CHAPTER 3, EXPERIMENT 1)
Source DF Type III SS Mean Square F Value Pr > F PH 1 0.31391093 0.31391093 3.69 0.0603 LNCONC 1 19.58974012 19.58974012 230.59 <.0001 LNCONC*PH 1 1.12977104 1.12977104 13.30 0.0006
Analysis of Covariance for LNVTR 7
6 R T 5 V N L
4
3
-6 -4 -2 0 LNCONC
PH 6.5 8
70
APPENDIX D: OBSERVED TAN CONCENTRATIONS FROM CONTINUOUS SYSTEM WITH TAN LOADING RATE (CHAPTER 4, EXPERIMENT1)
D.1 Data observed from continuous TAN loading rate @ 750 mg-N/day FIRST RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 1.22 0.055 0 .016 17 7.98
2 1.132 0.012
4 1.28 0.069
6 1.12 0.004
8 1.134 0.034 0.081 25 7.92
Mean 1.177
St. DEV 0.062
SECOND RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 1.21 0.013 0 .011 18 7.97 2 1.27 0.078 4 1.02 0.032 6 1.14 0.036 8 1.123 0.034 0.1 24 7.92 Mean 1.1526 St. DEV 0.094
71
THIRD RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 1.207 0.023 0 .0123 27 7.98
2 1.324 0.198
4 1.16 0.037
6 1.05 0.76
8 1.18 0.1 0.12 28 7.89
Mean 1.1842
St. DEV 0.098
D.2 Data observed from continuous TAN loading rate @ 650 mg-N/day FIRST RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 1.082 0.036 0 .037 12 7.97
2 1.17 0.047
4 1.067 0.053
6 1.232 0.024
8 1.046 0.058 0.094 14 7.91
Mean 1.119
St. DEV 0.078
72
SECOND RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 1.082 0.012 0 .028 14 7.96
2 1.168 0.068
4 0.997 0.095
6 1.068 0.12
8 1.081 0.086 0.11 10.1 7.89
Mean 1.079
St. DEV 0.060
THIRD RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 1.098 0.068 0 .071 18 7.99
2 1.172 0.053
4 1.183 0.065
6 0..995 0.042
8 0.084 0.084 0.104 13.1 7.89
Mean 1.103
St. DEV 0.077
73
D.3 Data observed from continuous TAN loading rate @ 550 mg-N/day FIRST RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 1.019 0.022 0 .032 17 7.98
2 1.002 0.061
4 0.984 0.058
6 0.952 0.046
8 1.0518 0.087 0.0927 12 7.92
Mean 1.001
St. DEV 0.037
SECOND RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 0.932 0.034 0 .084 8.6 8.01
2 1.028 0.058
4 1.11 0.062
6 1.043 0.038
8 0.978 0.013 0.092 15 7.96
Mean 1.018
St. DEV 0.067
74
THIRD RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 0.989 0.027 0 .062 16 7.99
2 1.078 0.034
4 1.076 0.076
6 1.12 0.12
8 0.092 0.089 0.098 13.1 7.92
Mean 1.038
St. DEV 0.0787
D.4 Data observed from continuous TAN loading rate @ 450 mg-N/day FIRST RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 0.874 0.076 0 .076 8.9 8.09
2 0.782 0.0451
4 0.831 0.065
6 0.798 0.034
8 0.801 0.023 0.12 16 7.93
Mean 0.817
St. DEV 0.036
75
SECOND RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 0.854 0.034 0 .101 7 7.98
2 0.796 0.058
4 0.762 0.062
6 0.784 0.038
8 0.802 0.013 0.065 12 7.91
Mean 0.799
St. DEV 0.034
THIRD RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 0.901 0.076 0 .078 9 7.96
2 0.812 0.035
4 0.789 0.031
6 0.795 0.029
8 0.79 0.048 0.101 15.1 7.89
Mean 0.817
St. DEV 0.047
76
D.5 Data observed from continuous TAN loading rate @ 350 mg-N/day FIRST RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 0.345 0.010 0 .021 5 7.96
2 0.212 0.031
4 0.31 0.027
6 0.289 0.013
8 0.338 0.098 0.065 13 7.91
Mean 0.299
St. DEV 0.053
SECOND RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 0.287 0.009 0 .33 12 7.98
2 0.211 0.0113
4 0.301 0.028
6 0.198 0.031
8 0.254 0.012 0.078 8 7.91
Mean 0.250
St. DEV 0.045
77
THIRD RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 0.212 0.024 0 .1 13 7.97
2 0.298 0.007
4 0.315 0.023
6 0.276 0.032
8 0.223 0.0137 0.023 15 7.92
Mean 0.264
St. DEV 0.045
D.6 Data observed from continuous TAN loading rate @ 250 mg-N/day FIRST RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 0.054 0.074 0 .078 18 7.98
2 0.021 0.032
4 0.048 0.011
6 0.032 0.009
8 0.065 0.001 0.12 16 7.91
Mean 0.044
St. DEV 0.017
78
SECOND RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 0.046 0.018 0 .021 14 7.97
2 0.038 0.023
4 0.052 0.015
6 0.04 0.007
8 0.029 0.001 0.063 21 7.9
Mean 0.041
St. DEV 0.008
THIRD RUN TIME Ammonia-N Nitrite-N Nitrate-N COD pH (HR) concentration Concentration Concentration (mg/L) (mg/L) (mg/L) (mg/L) 0 0.044 0.021 0 .101 18 7.98
2 0.038 0.008
4 0.026 0.019
6 0.039 0.07
8 0.035 0.006 0.078 22 7.93
Mean 0.036
St. DEV 0.006
79
APPENDIX E: OBSERVED DISSOLVED OXYGEN (DO) DATA UNDER CONTINUOUS TAN LOADING @ 350 MG/DAY (CHAPTER 4, EXPERIMENT 2)
FIRST RUN
TIME DOin (mg/l) DOout (mg/l) Do consumption rate 3 (HR) (g-O2/m -day) 0 3.73 3.25 2.76
1 3.81 3.31 2.88
2 4.18 3.56 3.57
3 4.12 3.67 2.59
4 4.1 3.54 3.22
5 4.18 3.62 3.22
6 4.15 3.54 3.51
7 4.22 3.49 4.2
8 4.26 3.79 2.70
9 4.15 3.77 2.18
10 4.15 3.81 1.95
11 4.13 3.8 1.9
80
SECOND RUN
TIME DOin (mg/l) DOout (mg/l) Do consumption rate 3 (HR) (g-O2/m -day) 0 3.75 3.35 2.30
1 3.81 3.24 3.28
2 3.89 3.38 2.93
3 3.92 3.51 2.36
4 4.1 3.54 3.22
5 4.21 3.7 2.93
6 4.25 3.59 3.8
7 4.32 3.67 3.74
8 4.28 3.66 3.57
9 4.19 3.64 3.17
10 4.1 3.65 2.59
11 4.12 3.67 2.59
81
THIRD RUN
TIME DOin (mg/l) DOout (mg/l) Do consumption rate 3 (HR) (g-O2/m -day) 0 3.89 3.28 3.51
1 3.92 3.3 3.57
2 4.05 3.27 4.5
3 4 3.24 4.38
4 4.08 3.22 4.95
5 4.12 3.21 5.24
6 4.18 3.19 5.70
7 4.15 3.17 5.64
8 4.07 3.18 5.12
9 4 3.21 4.55
10 3.88 3.25 3.63
11 3.92 3.24 3.92
82
VITA
Fatemehsadat Fahandezhsadi was born in Tehran, Iran, in September, 1984. She received her Bachelor degree in chemical engineering from University of Tehran in 2008. In fall 2012,
Fatemeh started her Masters in Civil and Environmental Engineering. She is currently a candidate for the degree of Master of Science.
83