I
Sustainable aquaculture effluent treatment systems utilizing biodegradable plastics through microbial remediation
A Thesis
Presented to
The Faculty of the College of Arts and Sciences
Florida Gulf Coast University
In Partial Fulfillment
of the Requirement for the Degree of
Master of Science
By
Cristina Rose Lopardo
2018
II
APPROVAL SHEET
This thesis is submitted in partial fulfillment of
the requirements for the degree of
Master of Science
Cristina Rose Lopardo
Approved: 2018
Hidetoshi Urakawa, Ph.D. Committee Chair / Advisor
William J. Mitsch, Ph.D.
Li Zhang, Ph.D.
Jong-Yeop Kim, Ph.D.
The final copy of this thesis has been examined by the signatories, and we find that both the content and the form meet acceptable presentation standards of scholarly work in the above mentioned discipline. III
Acknowledgements
I would like to thank my mentor Dr. Hidetoshi Urakawa for his support and guidance throughout my years of study at Florida Gulf Coast University. You helped me to further my knowledge in the field of microbiology and to reach my academic goals through being a member of your laboratory and as a mentee. Thank you very much for helping me to contribute to the scientific field through publication of my research, something I did not think possible before working with you. I thank Dr. William J. Mitsch as a member of my committee and co-collaborator for his insight and expertise in the field of wetland ecology which helped to bring a different insight into my studies which I otherwise would not have. I thank Dr. Li Zhang as a member of my committee and co-collaborator for her guidance throughout our time working together and for your feedback of my thesis. I thank Dr. Jong-Yeop Kim for being a member of my committee and for providing feedback and constructive criticism of my thesis. Special thanks to the Everglades
Wetland Research Park team under the guidance of Dr. William Mitsch and Dr. Li Zhang for their help and support in field and laboratory work. Also special thanks to Megan Feeney with laboratory work and Matthew Gamel and Nicolas Culligan for help with my aquarium setup and care of my pinfish, in addition to the rest of our lab team whom I had the privilege to know and work with over the past 2 ½ years who provided feedback and emotional support. I thank Haruka
Urakawa for her guidance and care, your assistance contributed to my successes both professionally and personally at FGCU. You are much appreciated for all that you do.
A special thanks to my parents, without your support I would literally not be where I am today. Your guidance and teachings have shaped me to be the person I became and helped me to reach my goals in life. I thank Stefanie, you are not only my sister but my best friend. You always help to remind me to not be so serious all the time and to live life with childish enthusiasm. Thanks IV to all my family and friends for your emotional support and guidance throughout my program and to always keep me grounded, you are too many to list but are always in my heart. Last but certainly not least, I would like to thank God. Without Him my dreams would not have happened and am grateful for all the opportunities I have been given and people I have met along the way.
I thank the following scholarship foundations for their monetary support allowing me to complete my masters program at FGCU; the Marco Island Shell Club, the Blair Foundation, and the Dorothy M. Rygh Fellowship Foundation. I also thank Joe Veradino at JoeFish Aquatics for providing aquariums and Dr. Timothy Hovanec at Dr Tim’s Aquatics for providing polyhydroxyalkanoate plastic beads. I thank our funding agencies for allowing us to complete our research, FGCU Office of Research and Graduate Studies internal grant program and the Florida
Sea Grant college program with support from the National Oceanic and Atmospheric
Administration (NOAA), Office of Sea Grant, U.S. Department of Commerce, Grant (PD-15-2).
V
Abstract
Demand of marine aquaculture has risen worldwide with human population growth and need for reliance on a steady protein supply. Direct discharges from on-land marine aquaculture systems are exceptionally high in nitrogen and phosphorous, which have negative impacts to downstream waters leading to eutrophication. Bioremediation methods using biodegradable plastic applications were explored through this thesis; a closed marine recirculating aquarium system (RAS) and using two different biodegradable plastic media in a bioreactor system and a marine aquaculture effluent from pinfish (Lagodon rhomboides) and wetlands using a vertical-flow (VFTW) and floating treatment (FTW) construction. The objectives explored were to; apply ecological engineering techniques for the treatment of nutrient rich marine aquaculture effluent, determine the nitrogen and phosphorous removal rates from these systems, and to determine the impact of biodegradable plastics on microbial community composition. In the marine RAS study, the TN removal efficiency was 92.1-98.5% with the use of biodegradable plastic, polycaprolactone (PCL) and polyhydroxyalkanoate (PHA), and an ASN medium and 61.5-62% TN removal efficiency with pinfish effluent. TN removal efficiency was shown to be greater with biodegradable plastic than a nonbiodegradable plastic control for both conditions. Alternatively, use of a biodegradable plastic
(PCL) as an external carbon input to a treatment wetland showed minimal increases in TN retention in a FTW system with greater efficiency effects shown between wetland construction, the VFTW system showed an increased TN, 87-91%, and TP, 74-81%, nutrient retention compared to the
FTW system. The use of biodegradable plastic for treatment of high nutrient marine aquaculture effluent was shown to be effective and this technology can be used on a larger scale application for sustainable wastewater treatment practices.
VI
Table of Contents
Acknowledgements ...... III Abstract ...... V Table of Contents ...... VI List of Tables ...... VII List of Figures ...... VIII Chapter 1: Introduction - Plastic waste in marine environments ...... 2 Objectives of Thesis ...... 8 Figures...... 13 References ...... 14 Chapter 2: Performance and microbial diversity of a closed recirculating aquaculture effluent treatment system using polycaprolactone and polyhydroxyalkanoate as carbon source and biofilm carrier ...... 21 Abstract ...... 22 Introduction ...... 23 Materials and Methods ...... 24 Results & Discussion ...... 30 Conclusion ...... 37 Figures & Tables ...... 39 References ...... 49 Chapter 3: Comparison of nutrient retention efficiency between vertical-flow and floating treatment wetland mesocosms with and without biodegradable plastic ...... 57 Abstract ...... 58 Introduction ...... 59 Materials and Methods ...... 61 Results & Discussion ...... 66 Conclusion ...... 74 Figures & Tables ...... 75 References ...... 85 VII
List of Tables
Chapter 2
Table 2.1 Operating conditions of aquaria during nutrient removal experiment ...... 39 Table 2.2 Bioreactor performance based on the removal efficiency of nutrients ...... 40 Table 2.3 Summary high-throughput sequencing and diversity indices ...... 41 Table 2.4 Summary percent abundance at the genus level ...... 42
Chapter 3
Table 3.1 Nutrient flux of two constructed wetland systems...... 75 Table 3.2 Removal efficiency of various constructed wetlands for aquaculture effluent treatment …………………………………………………………………………………………………….………76 Table 3.3 Summary of DNA sequencing and diversity indices ...... 77 Table 3.4 Functional groups at the genus level...... 78
VIII
List of Figures
Chapter 1
Figure 1.1 Outline of the SAFE system ...... 13
Chapter 2
Figure 2.1 Experimental setting of aquaria ...... 44 Figure 2.2 Succession of oxygen, TN and TP in the experiment with ASN medium ...... 45 Figure 2.3 Succession of oxygen, TN and TP in the experiment with aquaculture wastewater ....46 Figure 2.4 Biofilm growth on plastic surface and percent decrease of plastic weight ...... 47 Figure 2.5 Class-level relative abundance of microbial communities ...... 48
Chapter 3
Figure 3.1 Design of two wetland systems ...... 80 Figure 3.2 Mesocosms used in this study ...... 81 Figure 3.3 Change in inflow and outflow nutrient concentrations ...... 82 Figure 3.4 Change in stem height of Spartina patens ...... 83 Figure 3.5 Relative bacterial and archaeal abundance at the phylum level ...... 84
1
Sustainable aquaculture effluent treatment systems utilizing
biodegradable plastics through microbial remediation
2
CHAPTER 1
Plastic waste in marine environments
3
1. INTRODUCTION
1.1. Plastic uses and sources of environmental pollution
The world’s oceans are subject to major anthropogenic disturbance from both land based and oceanic activities such as overfishing, natural resource extraction, wastewater discharges, and debris accumulation. Oceanic debris are comprised of approximately 80% land and 20% sea origin, with a distribution of 30% occurring as free floating and beach washup and 70% collecting at the ocean floor (Jambeck et al., 2015; Frias et al., 2016). One of the most common anthropogenic debris present in oceans today is plastic, which has been recognized as an environmental pollutant since the 1970s globally, mainly due to global plastic production (250-280 million tons yr-1 in
2011) and improper plastic disposal (Erikson et al., 2013; Van Cauwenberghe et al., 2013; Vianello et al., 2013; Fischer et al., 2015; Jambeck et al., 2015; Frias et al., 2016). Marine plastic debris occur from poor waste management, fishing, industry, shipping, and tourism losses (Jambeck et al., 2015; Steer et al., 2017). Plastics are considered a persistent pollutant due to their nonbiodegradable nature, plastics are a synthetic byproduct of petroleum which is difficult to degrade environmentally (Erikson et al., 2013; Wessel et al., 2016). The pathway of plastic degradation occurs from a combination of ultraviolet radiation from the sun, hydrolysis, and mechanical breakdown, which can take hundreds of years to break down (Erikson et al., 2013;
Vianello et al., 2013; McCormick et al., 2014; Melli et al., 2017).
A new problem of concern regarding plastic debris is the emergence of microplastics in the environment. Microplastics are defined as plastic particles ranging in size from ˂ 1 millimeter to
˂ 5 millimeters (Fischer et al., 2015; Leslie et al., 2017). Microplastics occur in marine environments in two distinct forms, primary microplastics as a wide range of products from cosmetics to microplastic fibers, from polyester and polyamide synthetic fabrics (Van 4
Cauwenberghe et al., 2013; Fischer et al., 2015; Steer et al., 2017) and secondary microplastics from fragmentation of larger plastic particles (Fischer et al., 2015). Primary microplastics enter the environment through wastewater treatment plant discharges, where the plastic is unable to be removed from the water column before discharging due to their small size and buoyancy
(McCormick et al., 2014; Jambeck et al., 2015; Nel et al., 2017).
There is now concern of microplastics entering the food chain, it has been observed that microplastics consumed by zooplankton can be transferred to higher trophic levels with unknown implications for ocean species and even human health (Steer et al., 2017; Tsang et al., 2017; Chiba et al., 2018). After microplastic ingestion it is either stored in the tissues of the organism or released back into the environment through expulsion (Van Cauwenberghe et al., 2013). Due to the varying sizes of microplastics they pose a threat to a wide array of organisms (i.e. cetaceans, sea turtles, sea birds, fishes, and zooplankton) and consumption has been shown to negatively impact feeding behavior, survival, and fecundity (Wessel et al., 2016; Brach et al., 2018). In addition to negative impacts from the physical presence of microplastics, there are also negative chemical impacts which can occur from the absorption and concentration of persistent organic pollutants on plastic surfaces and the release of hazardous chemicals composing these plastics (Woodall et al., 2015;
Wessel et al., 2016; Tsang et al., 2017).
Microplastics have been found in every environment globally from the Mediterranean Sea,
Pacific Ocean, remote islands, and even Antarctic communities and deep-seas (Van Cauwenberghe et al., 2013; McCormick et al., 2014; Chiba et al., 2017; Nel et al., 2017). However, microplastics have distinct biofilm communities from the surrounding water column and may encourage novel microbial communities, especially those able to degrade plastics such as Pseudomonas spp.
(McCormick et al., 2014). In addition to unique microbial community composition, microplastics 5 have been shown to be a vector for alien species (Angiolillo et al., 2015). Several policy measures have now been put in place globally to help reduce plastic waste inputs such as the Marine Strategy
Framework Directive for European Union member states to have “good environmental status” by
2020 (Vianello et al., 2013) and the United Nations Sustainable Development Goal 14.1 to “take action to reduce marine pollution by 2025” (Chiba et al., 2017). These initiatives aim to reduce plastic inputs through more responsible waste management initiatives, however, these policies are unable to control plastics in deep oceanic environments.
1.2. Role of biodegradable plastics: a step toward sustainability
An alternative to traditional plastics and to reduce anthropogenic plastic debris effects are biodegradable plastics. Biodegradable plastics are a step toward “greening” of industrial chemistry with the promise of reducing the carbon footprint and environmental impacts associated with traditional fossil fuel produced plastics (Gross & Kalra, 2002; Emadian et al., 2016; Thakur et al.,
2018). Biodegradable plastics are classified into three types of polymers; chemically synthesized polymers (petroleum based), bacterial, and plant-based polymers (synthesized from biomass), which can be degraded by heterotrophic microorganisms into simpler carbon compounds (Ishigaki et al., 2004; Numata et al., 2009; Song et al., 2009; Tokiwa et al., 2009). Biodegradable plastics are broken down environmentally through both physical and chemical means and are used as a substrate binding surface for microbial biofilm formation (Boley et al., 2000; Tokiwa et al., 2009).
However, different plastics have different degradation rates depending on environmental conditions (i.e. pH, temperature, salinity), and degradation optimization based on location (i.e. soil, water column, compost) in addition to their molecular composition (Lenz and Marchessault,
2005; Emadian et al., 2016). 6
Due to global policies toward reducing plastic wastes and carbon emissions, the production of biodegradable plastics was estimated to increase from 1.7 million tons worldwide in 2014 to 6.2 million tons by 2018 (Emadian et al., 2016; Spierling et al., 2018; Thakur et al., 2018).
Biodegradable plastics such as poly(3-hydroxybutryate) (PHB) belong to a family of microbially produced plastics produced through fermentation using Wautersia eutropha, additionally PHB can be enzymatically produced intracellularly by a variety of organisms under carbon limiting conditions, Bacillus megaterium and Alcaligenes eutrophus, allowing for a wide range of applications and genetic engineering such as implanting the reductase and synthase genes of A. eutrophus within a transgenic plant producing PHB (Gross and Kalra, 2002; Lenz and
Marchessault, 2005). Due to production cost and recycling ability, the dominant biodegradable plastics used in single-use plastic packaging are either petroleum-based or have the same chemical composition as current petroleum derived plastics such as bio-polyethylene (Lenz and
Marchessault, 2005; Spierling et al., 2018). Until the market for biodegradable plastics increases this is unlikely to change due to the small quantity of this type of plastic waste. Common sources of biodegradable plastic waste are from wastewater treatment plant discharges containing non- water-soluble plastics. A proposed solution is to convert biodegradable plastic waste to compost, chemical intermediates for industrial use, and energy from aerobic and anaerobic microbial processing (Gross and Kalra, 2002). In addition to single-use packaging material, there are currently a wide variety of applications using biodegradable plastics such as medical applications, wastewater treatment systems, disposable packaging, building materials, agricultural and aquacultural applications (Gross and Kalra, 2002; Lenz and Marchessault, 2005; Tan et al., 2013;
Emadian et al., 2016; Thakur et al., 2018). 7
1.3. Biodegradable plastic uses and applications in wastewater treatment: aquaculture
effluents
Demand of marine aquaculture has risen worldwide with human population growth and need for reliance on a steady protein supply as the world’s marine fish resources are being depleted.
The global fish production from intensive aquaculture practices was predicted to be around 100 million tons in 2008 with 20% of that being cultured fishes, a number that is expected to increase substantially in the next decade (De Lange et al., 2013; Turcios and Papenbrock, 2014). In these systems, management of water is key, over time oxygen can be quickly depleted as well as nitrogen accumulation in the form of toxic nitrite and ammonia as fish excrete waste and uneaten food accumulates in the system, leading to the need for frequent water changes to keep fish stocks healthy (van Rijn, 1996; van Rijn et al., 2006; Turcios and Papenbrock, 2014). Direct discharges from on-land marine aquaculture systems are exceptionally high in nitrogen, phosphorous, and salts which have negative environmental impacts to downstream or adjacent waters leading to eutrophication and salinization (Konnerup et al., 2011; Liang et al., 2017; Zhang et al., 2016).
Several approaches to water treatment in these systems are to utilize bioremediation through algae and microorganisms within aquaculture ponds and to utilize an external biofiltration system (Avnimelech et al., 2006). Alternatively, the development of closed recirculating systems with the use of a biofilter allows for removal of nitrogen and dissolved organic matter within the water column improving water quality and decreasing large energy inputs (Gutierrez-Wing and
Malone, 2006). In cases of large scale aquaculture operations, the use of constructed wetlands for wastewater treatment have also been explored in recirculating systems, use of aquaponics, and in flow-through systems for treatment of high nutrient effluents (Turcios and Papenbrock, 2014).
2. Objectives of the investigation 8
2.1. Sustainable aquaculture for environments (SAFE).
Biodegradable plastic applications for use as an external carbon source in low C/N ratio wastewater, as found in aquarium and aquaculture effluents, were explored through two different studies in this thesis; a closed recirculating marine aquarium system using two different biodegradable plastic media in a bioreactor system and a study exploring the use of a biodegradable plastic as an external carbon source in a marine aquaculture effluent treatment wetland. This project was completed as part of the sustainable aquaculture for environments (SAFE) multidisciplinary study aimed at the development of novel recirculating aquaculture systems composed of multiple components (e.g. constructed wetland, edible seaweed culture, and bioreactor biopolymer-based denitrification) for the treatment of nutrient rich aquaculture effluent
(Fig. 1.1). Components B1 and B4 were completed through this thesis work and will be explored further in the following sections.
2.2. Biodegradable plastic bioreactor system for marine aquaculture effluent treatment.
Biodegradable plastics, produced from petroleum byproducts and synthesized from biomass, have been used successfully as a bioreactor medium for effective wastewater treatment of a wide variety of sources. Recently, municipal wastewater treatment systems have employed the use of biofilm reactors using biodegradable plastics to increase denitrification efficiency in wastewater treatment to increase the C/N ratio (Tan et al., 2013). Additionally, within aquarium systems bioreactor filters can be effective with zero discharge using various biodegradable plastics for marine and freshwater treatment (Boley et al., 2000; Gutierrez-Wing et al., 2012). A closed recirculating system is seen as a cost effective and environmentally sound approach as this can dramatically reduce eutrophication from high nutrient effluent discharges to adjacent water bodies
(Avnimelech et al., 2006). Closed recirculating aquaculture and aquarium systems utilizing 9 bioreactor filtration have been studied to determine the efficiency rate of nutrient removal (Shen and Wang, 2011; Gutierrez-Wing et al., 2014; Wu et al., 2014) and microbial community composition (Tal et al., 2003; Qui et al., 2017).
In this study two biodegradable plastics, polycaprolactone (PCL) and polyhydroxyalkanoate (PHA), were used as bioreactor medium for a closed recirculating marine aquarium system for water treatment to determine the degradation rate and nutrient cycling efficiency within this type of treatment system to be discussed further throughout this thesis. PHA belongs to the group of PHBs which degrade naturally across a wide variety of environments (e.g. soil, freshwater, marine, activated sludge) due to bacterial production of PHA depolymerases
(Numata et al., 2009; Gutierrez-Wing et al., 2012; Emadian et al., 2016). Alternatively, PCL is a petroleum-based biodegradable plastic which can be broken down through biofilm growth and physical degradation in a bioreactor system, and additionally in fresh and saltwater, soil, and compost environments (Ishigaki et al., 2004; Emadian et al., 2016). These carbon compounds in aquarium systems allow for uptake of carbon as a proton source initiating breakdown of nitrate through either denitrification or nitrate assimilation (Arbiv and van Rijn, 1995).
The overall objectives explored were to determine the nutrient removal efficiency within a closed recirculating marine aquarium filtration system and to explore microbial community composition. Specific objectives were:
● Develop a cost-effective marine aquarium effluent treatment system utilizing a
biopolymer-based denitrification system using PHA and PCL as a carbon source.
● Compare the nutrient cycling efficiency associated with each biodegradable plastic used,
PHA or PCL, compared with a traditional nonbiodegradable plastic bioreactor fill,
polypropylene. 10
● Determine if there are differences in nutrient removal efficiency based on flowrate.
● Determine microbial community composition and unique microorganisms associated with
each biodegradable plastic.
2.3. Biodegradable plastic-based constructed wetlands for treatment of aquaculture effluent.
The utilization of a wetland treatment system for saltwater aquaculture effluent using a biodegradable plastic, polycaprolactone (PCL), as an external carbon source to increase nutrient cycling efficiency was explored. In this study PCL was preferred over polyhydroxyalkanoate
(PHA) due to its cost efficiency. Constructed treatment wetlands have been effective to treat effluents of urban runoff, agricultural areas, and aquaculture through their nutrient cycling and uptake ability and ability to treat high salinity waters (Brown et al., 1999; Liang et al., 2017).
Newly constructed wetlands have a low concentration of organic carbon, the addition of an external carbon source in the form of organic matter such as periphyton algae, addition of glucose, and fructose in the wetland influent have been shown to be effective for increasing carbon and nutrient cycling efficiency (Wu et al., 2014).
In this study, the use of a biodegradable plastic in a constructed wetland system to treat marine aquaculture effluent was explored to increase the organic carbon content in newly constructed wetland mesocosms and to determine if this would lead to an increase in nutrient cycling efficiency. Two different wetland mesocosm systems were explored, a vertical-flow constructed wetland with embedded biodegradable plastic and a floating treatment wetland with a bioreactor filtration system utilizing biodegradable plastic as a medium. A vertical-flow constructed wetland is a form of a subsurface flow wetland commonly used for aquaculture effluent treatment effective for both solids and nutrient removal, characterized by having a sand/gravel substrate that water percolates through (Konnerup et al., 2011; De Lange et al., 2013; 11
Mitsch and Gosselink, 2015). Additionally, a floating treatment wetland is commonly used for stormwater, agricultural, and aquaculture treatment through aquaponics systems where aquatic or terrestrial plants are grown hydroponically on a floating mat system directly in open water
(Hubbard et al., 2004; Vymazal, 2007; Olguín et al., 2017; Pavlineri et al., 2017; Urakawa et al.,
2017).
The application of a biodegradable plastic as an external carbon source in constructed wetlands to increase heterotrophic denitrification was done on a microcosm scale, 0.04 m2 – 0.071 m2, with the use of PCL/starch (Shen et al., 2015) and use of poly-3-hydroxybutryate-co-3- hydroxyvalerate/polyacetic acid (PHBV/PLA) (Yang et al., 2018). However, the use of this technology is relatively new and no large-scale system studies, using 1 m2 mesocosms, have been conducted, to our knowledge at this time, and was the motivation for exploring the novel use of
PCL in this way.
The overall objectives explored were to determine the nutrient removal efficiency within a vertical-flow treatment wetland and a floating treatment wetland mesocosm system and their microbial community composition. Specific objectives were:
● Apply ecological engineering techniques for the design and maintenance of a constructed
wetland for treatment of nutrient rich marine aquaculture effluent.
● Determine the nitrogen and phosphorous removal rates for both a vertical-flow treatment
wetland and a floating treatment wetland mesocosm system.
● Improve removal efficiency with the presence of an external carbon source in the form of
PCL plastic beads incorporated in wetland construction.
● Understand the microbial community composition within each constructed wetland type
looking at soil, root, water column, and PCL biofilm. 12
● Determine the impact of PCL on microbial community composition and unique
microorganisms observed in the presence of PCL.
Biodegradable plastics are a green alternative to traditional petroleum based non- biodegradable plastics for reduction of plastic pollution globally. Additionally, biodegradable plastic waste can be utilized for the creation of industrial uses and a form of clean energy as previously stated. With the development of more biodegradable plastic-based technologies, their use will become more widespread and help to continue the shift from petroleum-based to biodegradable alternatives. Green energy initiatives such as those discussed will bring about change in global energy uses based on fossil fuels and petrochemical production with biodegradable plastics as a step in this direction.
13
Figure 1.1 Outline of the Sustainable Aquaculture for Environments (SAFE) system. Components (A) and the combination of multiple (interchangeable) revolver-type multiple components (B) that will treat downstream, recirculating nutrient rich water in 4 ways, biopolymer-based denitrification (B1), edible seaweed culture and the following market research (B2), energy production using algal aquaculture and methane production (B3) and constructed wetland (B4). Purple boxes indicate each component. Blue boxes indicate components B1 and B4 completed through thesis work.
14
References
Angiolillo, M., di Lorenzo, B., Farcomeni, A., Bo, M., Bavestrello, G., Santangelo, G., Cau, A.,
Mastascusa, V., Cau, A., Sacco, F., Canese, S., 2015. Distribution and assessment of
marine debris in the deep Tyrrhenian Sea (NW Mediterranean Sea, Italy). Marine Pollution
Bulletin. 92, 149-159.
Arbiv, R., van Rijn, J., 1995. Performance of a treatment system for inorganic nitrogen removal in
intensive aquaculture systems. Aquacultural Engineering. 14, 189-203.
Avnimelech, Y., 2006. Bio-filters: The need for a new comprehensive approach. Aquacultural
Engineering. 34, 172-178.
Boley, A., Muller, W.-R., & Haider, G., 2000. Biodegradable polymers as solid substrate and
biofilm carrier for denitrification in recirculated aquaculture systems. Aquacultural
Engineering. 22, 75-85.
Brach, L., Deixonne, P., Bernard, M-F., Durand, E., Desjean, M-C., Perez, E., van Sebille, E., ter
Halle, A., 2018. Anticyclonic eddies increase accumulation of microplastic in the North
Atlantic subtropical gyre. Marine Pollution Bulletin. 126, 191-196.
Brown, J.J., Glenn, E.P., Fitzsimmons, K.M., Smith, S.E., 1999. Halophytes for the treatment of
saline aquaculture effluent. Aquaculture. 175, 225-268.
Chiba, S., Saito, H., Fletcher, R., Yogi, T., Kayo, M., Miyagi, S., Moritaka, O., Fujikura, K., 2018.
Human footprint in the abyss: 30 year records of deep-sea plastic debris. Marine Policy.
96, 204-212. 15
De Lange, H. J., Paulissen, M.P.C.P., Slim, P.A., 2013. 'Halophyte filters': The potential of
constructed wetlands for application in saline aquaculture. International Journal of
Phytoremediation. 15, 352-364.
Emadian, S.M., Onay, T.T., Demirel, B., 2017. Biodegradation of bioplastics in natural
environments. Waste management. 59, 526-536.
Erikson, M., Maximenko., N., Thiel., M., Cummins., A., Lattin, G., Wilson, S., Hafner, J., Zellers,
A., Rifman, S., 2013. Plastic pollution in the South Pacific subtropical gyre. Marine
Pollution Bulletin. 68, 71-76.
Fischer, V., Elsner, N.O., Brenke, N., Schwabe, E., Brandt, E., 2015. Plastic pollution of the Kuril-
Kamchatka Trench area (NW pacific). Deep-Sea Research II. 111, 399-405.
Frias, J.P.G.L., Gago, J., Otero, V., Sobral, P., 2016. Microplastics in coastal sediments from
Southeastern Portuguese shelf waters. Marine Environmental Research. 114, 24-30.
Gutierrez-Wing, M.T., Malone, R.F., 2006. Biological filters in aquaculture: trends and research
directions for freshwater and marine applications. Aquacultural Engineering. 34, 163-171.
Guiterrez-Wing, M.T., Malone, R.F., Rusch, K.A., 2012. Evaluation of polyhydroxybutryate as a
carbon source for recirculating aquaculture water denitrification. Aquacultural
Engineering. 51, 36-43.
Guiterrez-Wing, M.T., Malone, R.F., Rusch, K.A., 2014. Development of a model for PHA-based
denitrification in a packed bed reactor. Aquacultural Engineering. 60, 41-47.
Gross, R.A., Kalra, B., 2002. Biodegradable polymers for the environment. Science. 297, 803-807. 16
Hubbard, R., Gascho, G., Newton, G., 2004. Use of floating vegetation to remove nutrients from
swine lagoon wastewater. Transactions of the ASAE. 47, 1963-1972.
Ishigaki, T., Sugano, W., Nakanishi, A., Tateda, M., Ike, M., Fujita, M., 2004. The degradability
of biodegradable plastics in aerobic and anaerobic waste landfill model reactors.
Chemosphere. 54, 225-233
Jambeck, J.R., Geyer, R., Wilcox., C., Siegler., T.R., Perryman., M., Andrady, A., Narayan., R.,
Law., K.L., 2015. Plastic waste inputs from land into the ocean. Marine Pollution. 347,
768-770.
Konnerup, D., Trang, N.T.D., Brix, H. 2011. Treatment of fishpond water by recirculating
horizontal and vertical flow constructed wetlands in the tropics. Aquaculture. 313, 57-64.
Lenz, R.W., Marchessault, R.H., 2005. Bacterial polyesters: biosynthesis, biodegradable plastics
and biotechnology. Biomacromolecules. 6, 1-8.
Leslie, H.A., Brandsma, S.H., van Velzen, M.J.M., Vethaak, A.D., 2017. Microplastics en route:
Field measurements in the Dutch river delta and Amsterdam canals, wastewater treatment
plants, North Sea sediments and biota. Environmental International. 101, 133-142.
Liang, Y., Zhu, H., Bañuelos, G., Yan, B., Zhou, Q., Yu, X., Cheng, X., 2017. Constructed
wetlands for saline wastewater treatment: A review. Ecological Engineering. 98, 275-285.
McCormick, A., Hoellein, T.J., Mason, S.A., Schluep, J., Kelly, J.J., 2014. Microplastic is an
abundant and distinct microbial habitat in an urban river. Environmental Science &
Technology. 48, 11863-11871. 17
Melli, V., Angiolillo, M., Ronchi, F., Canese, S., Giovanardi, O., Querin, S., Fortibuoni, T., 2017.
The first assessment of marine debris in a site of community importance in the north-
western Adriatic Sea (Mediterranean Sea). Marine Pollution Bulletin. 114, 821-830.
Mitsch, W.J., Gosselink, J.G., 2015. Wetlands, 5th Edition: John Wiley & Sons, Inc., Hoboken,
NJ.
Nel, H.A, Hean, J.W., Noundou, X.S., Froneman, P.W., 2017. Do microplastic loads reflect the
population demographics along the southern African coastline?. Marine Pollution Bulletin.
115, 115-119.
Numata, K., Abe, H., Iwata, T., 2009. Biodegradability of Poly(hydroxyalkanoate) materials.
Materials. 2, 1104-1126.
Olguín, E.J., Sánchez-Galván, G., Melo, F.J., Hernández, V.J., González-Portela, R.E., 2017.
Long-term assessment at field scale of floating treatment wetlands for improvement of
water quality and provision of ecosystem services in a eutrophic urban pond. Science of
the Total Environment. 584, 561-571.
Pavlineri, N., Skoulikidis, N.Th., Tsihrintzis, V.A., 2017. Constructed floating wetlands: A review
of research, design, operation and management aspects, and data meta-analysis. Chemical
Engineering Journal. 308, 1120-1132.
Qui, T., Xu, Y., Gao, M., Han, M., Wang, X., 2017. Bacterial community dynamics in a
biodenitrification reactor packed with polylactic acid/poly (3-hydroxybutryate-co-3-
hydroxyvalerate) blend as the carbon source and biofilm carrier. Journal of Bioscience and
Bioengineering. 123, 606-612. 18
Shen, Z., Wang, J., 2011. Biological denitrification using cross linked starch/PCL blends as solid
carbon source and biofilm carrier. Bioresource Technology. 102, 8835-8838.
Shen, Z., Zhou, Y., Liu, J., Xiao, Y., Cao, R., Wu, F., 2015. Enhanced removal of nitrate using
starch/PCL blends as solid carbon source in a constructed wetland. Bioresource
Technology. 175, 239-244.
Song, J. H., Murphy, R. J., Narayan, R., Davies, G. B., 2009. Biodegradable and compostable
alternatives to conventional plastics. Philosophical Transactions of the Royal Society of
London Series, B Biological Sciences. 364, 2127-2139.
Spierling, S., Röttger, C., Venkatachalam, V., Mudersbach, M., Herrmann, C., Endres, H.-J., 2018.
Bio-based plastics – A building block for the circular economy?. Procedia CIRP. 69, 573-
578.
Steer, M., Cole, M., Thompson, R.C., Lindeque, P.K., 2017. Microplastic ingestion in fish larvae
in western English Channel. Environmental Pollution. 226, 250-259.
Tal, Y., Watts, J.E., Schreier, S.B., Sowers, K.R., Schreier, H.J., 2003. Characterization of the
microbial community and nitrogen transformation processes associated with moving bed
bioreactors in a closed recirculated mariculture system. Aquaculture. 215, 187-202.
Tan, C., Ma, F., Qui, S., Zeng, H., Zhou, Y., 2013. Study of biodegradable polyurethane foam as
carriers for low C/N ratio wastewater. Applied Mechanics and Materials. 284-287, 352-
356. 19
Thakur, S., Chaudhary, J., Sharma, B., Verma, A., Tamulevicius, S., Thakur, V.K., 2018.
Sustainability of bioplastics: Opportunities and challenges. Current Opinion in Green and
Sustainable Chemistry. 13, 68-75.
Tokiwa, Y., Calabia, B.P., Ugwu, C.U., Aiba, S., 2009. Biodegradability of plastics. International
Journal of Molecular Science. 10, 3722-3742.
Tsang, Y.Y., Mak, C.W., Liebich, C., Lam, S.W., Sze, E.T-P., Chan, K.M., 2017. Microplastic
pollution in the marine waters and sediments of Hong Kong. Marine Pollution Bulletin.
115, 20-28.
Turcios, A.E., Papenbrock, J., 2014. Sustainable treatment of aquaculture effluents—what can we
learn from the past for the future? Sustainability. 6, 836-856.
Urakawa, H., Dettmar, D.L., Thomas, S., 2017. The uniqueness and biogeochemical cycling of
plant root microbial communities in a floating treatment wetland. Ecological Engineering.
108, 573-580.
Van Cauwenberghe, L., Vanreusel, A., Mees, J., Janssen, C.R., 2013. Microplastic pollution in
deep-sea sediments. Environmental Pollution. 182, 495-499. van Rijn, J., 1996. The potential for integrated biological treatment systems in recirculating fish
culture—A review. Aquaculture. 139, 181-201. van Rijn, J., Tal, Y., Schreier, H.J., 2006. Denitrification in recirculating systems: Theory and
applications. Aquacultural Engineering. 34, 364-376.
Vianello, A., Boldrin., A., Guerriero, P., Moschino, V., Rella., R., Sturaro, A., Da Ros, L., 2013.
Microplasic particles in sediments of Lagoon of Venice, Italy: First observations on 20
occurrence, spatial patterns and identification. Estuarine, Coastal and Shelf Science. 130,
54-61.
Vymazal, J., 2007. Removal of nutrients in various types of constructed wetlands. Science of the
Total Environment. 380, 48-65.
Wessel, C.C., Lockridge, G.R., Battiste, D., Cebrian, J., 2016. Abundance and characteristics of
microplastics in beach sediments: Insights into microplastic accumulation in northern Gulf
of Mexico estuaries. Marine Pollution Bulletin. 109, 178-183.
Woodall, L.C., Gwinnett, C., Packer, M., Thompson, R.C., Robinson L.F., Paterson, G.L.J., 2015.
Using a forensic science approach to minimize environmental contamination and to
identify microfibers in marine sediments. Marine Pollution Bulletin. 95, 40-46.
Wu, S., Kuschk, P., Brix, H., Vymazal, J., Dong, R., 2014. Development of constructed wetlands
in performance intensifications for wastewater treatment: a nitrogen and organic matter
targeted review. Water Resources. 57, 40-55.
Yang, Z., Yang, L., Wei, C., Wu, W., Zhao, X., Lu, T., 2018. Enhanced nitrogen removal using
solid carbon source in constructed wetland with limited aeration. Bioresource Technology.
248, 98-103.
Zhang, S., Ban, Y., Xu, Z., Cheng, J., Li, M., 2016. Comparative evaluation of influencing factors
on aquaculture wastewater treatment by various constructed wetlands. Ecological
Engineering. 93, 221-225.
21
CHAPTER 2
Performance and microbial diversity of a closed recirculating aquaculture effluent treatment system using polycaprolactone and polyhydroxyalkanoate as carbon source and biofilm carrier
The original work found in the following chapter has been submitted for review in the journal
Aquaculture International
Cristina R. Lopardo & Hidetoshi Urakawa
22
ABSTRACT
Nitrogen removal is essential for a successful management approach in a marine recirculating aquaculture system (RAS) through microbial remediation. In this study two biodegradable polymers, polycaprolactone (PCL) and polyhydroxyalkanoate (PHA), were used as a carbon source and biofilm carrier for marine RAS wastewater treatment. Results showed that the use of an artificial saltwater nitrate (ASN) medium with seeded polymers had a high nitrogen removal efficiency regardless of flow rate, with a nitrate-nitrite removal efficiency of 95.4-98.9%. The TN removal efficiency was greater with PHA, 98.3-98.5%, than PCL, 92.1-96.8%. TP removal efficiency was greater under a high flow rate condition with PCL and PHA, 24.5-59.1%. Use of marine aquaculture effluent showed an increased nutrient removal efficiency with PCL having a nitrate-nitrite removal efficiency of 89%, TN removal efficiency of 62%, and a TP removal efficiency of 58.5%. Microbial community analysis demonstrated the complexity of microbial consortia within marine RAS. Two major classes identified in our study were Alphaproteobacteria
(2-68%) and Flavobacteriia (0.5-58%). We identified Crocinitomix, Oceanicola,
Meridianimaribacter, and Ruegeria as potential PCL degraders, while Winogradskyella,
Muricauda, Marinobacter, Alteromonas as potential PHA degraders. The microbial communities responded to varying water chemistry differences and developed unique consortia based on biodegradable polymer types. Biodegradable plastics can be utilized in marine RAS to treat wastewater through microbial processes.
23
1. INTRODUCTION
A closed recirculating aquaculture system (RAS) is seen as an environmentally sound approach for culturing fish and is widely used worldwide because this approach can produce high yields of fish stock through controlled feeding and rearing conditions and limits harm to natural fisheries and water usage (van Rijn, 1996). However, drawbacks to this approach are the accumulation of nutrient and fish culture waste in the system, which are too high to directly discharge into adjacent water bodies. Thus, effluent needs to be treated before being discharged in many cases (van Rijn, 1996).
An effective approach to solve this issue is through a microbially mediated treatment for aquaculture effluents. Waste reduction could be accomplished by incorporation of denitrification and sludge digestion (Shen et al., 2013a, b; Gutierrez-Wing et al., 2014; Wang and Chu, 2016;
Zhang et al., 2016). In RAS, glucose, methanol, and starch are mainly used as carbon sources to stimulate nitrogen removal (Boley et al., 2000; Gutierrez-Wing et al., 2012; Zhu et al., 2015).
Understanding microbial diversity and composition in a marine aquaculture system is crucial to understand nitrogen reduction efficiency and predominant groups of organisms contributing to nitrification, denitrification, or nitrate assimilation processes (Tal et al., 2003; Gao et al., 2012).
Several studies aimed to increase the nitrogen removal efficiency within RAS bioreactor systems by adding external carbon sources (Boley et al., 2000; Gutierrez-Wing et al., 2012, 2014; Luo et al., 2018), however, application to marine RAS is limited.
Recently application of biodegradable plastics as an electron donor in denitrification have been shown to be an effective approach for wastewater treatment of landfill leachate (Ishigaki et al., 2004) and municipal wastewater treatment systems (Tan et al., 2013). Biodegradable plastics are often used as a filtration medium through breakdown of plastic by both physical and chemical 24 hydrolysis and are used as a structural basis for biofilm formation (Tokiwa et al., 2009). Despite a large potential to use biodegradable plastic in the denitrification process of RAS, previous microbial community analyses were mainly focused on the RAS equipped with non-biodegradable plastic material as biofilm carrier (Tal et al., 2003; Michaud et al., 2009; Gao et al., 2012; Kumar et al., 2013).
Here we investigated the nutrient removal performance and microbial community of marine RAS enhanced by the use of polycaprolactone (PCL) and polyhydroxyalkanoate (PHA) as biodegradable plastics. To determine the factors that influence the nutrient removal efficiency and plastic degradation patterns, two flow rate conditions were tested in an artificial saltwater nitrate
(ASN) medium first, then marine aquaculture effluent from pinfish (Lagodon rhomboides) were used in the second set of experiments. Overall, we confirmed the great nitrogen removal efficiency and the presence of potential polymer degraders in the complexity of microbial consortia within marine RAS.
2. MATERIALS AND METHODS
2.1 Acclimation of aquarium sludge under batch experiments
An initial pilot scale study was conducted over a three-month period to determine the nutrient removal ability of microbial communities using two biodegradable plastics, polycaprolactone
(PCL), a petroleum-based polymer (IC3D, TechTack Moldable plastic) and polyhydroxyalkanoate
(PHA), a bacterial based polymer (NP-active Pearls, Dr. Tim’s Aquatics), as a carbon source.
Aquarium sludge used as a seed was isolated from the biofilter of a 60 L saltwater aquarium, which housed a few Goldtail Demoiselle (Chrysiptera parasema) and Banggai Cardinalfish (Pterapogon kauderni) and maintained at 25°C and 35 ppt for one year. A batch PCL degradation experiment 25 was conducted in 50 mL glass test tubes containing an equal volume of aquarium sludge and artificial saltwater nitrate (ASN) medium amended with 10 mg-N / L of KNO3 and 2 mg-P / L of
KH2PO4. PCL beads were added to be 5% total volume. The test tubes were prepared in duplicate and incubated at 30°C for three months. Ammonium, nitrate, and phosphate concentrations were colorimetrically determined and monitored during the incubation period. A similar experiment was conducted in triplicate for PHA using the same ASN medium also with 5% PHA beads by volume.
2.2. Bed expansion: seeded beads incubation for scale up
Once initial small-scale experiments were conducted, PCL and PHA beads were harvested
(0.8 g) and then used for biomass expansion for further large-scale bioreactor experiments. The plastic beads were incubated in two 1 L glass autoclaved medium bottles with lids at 30°C to ensure no cross contamination can occur, over a two-month period, doubling bead biomass every week with unseeded sterile bioplastic beads until a volume of 2 L was reached for both PCL and
PHA. These samples were monitored for parameters such as ammonia, nitrate, nitrite, oxygen and pH. This monitoring proved that both PCL and PHA beads during bed expansion were maintained at aerobic conditions similar to the original pilot scale experiment.
2.3 Bioreactor experiment using an ASN medium and seeded biodegradable plastic beads
Experimentation using the ASN medium (35 ppt, 10 mg-N / L KNO3 and 2 mg-P / L
KH2PO4) and cultured biodegradable plastic beads was conducted in the temperature-controlled laboratory at Florida Gulf Coast University (26°46’43”N, 81°77’34”W) and consisted of five 76
L (20 gallon) aquariums filled with 38 L of medium with flow through systems, differing only in the type of biodegradable plastic used and flowrate. Temperature remained stable with a mean of
25°C for Phase I and 26.1°C for Phase II, with an average salinity of 32.6 and 32.7 ppt, respectively 26
(Table 2.1). Experimental setup consisted of five AQUAMAXX bioreactors (1 L volume) with
500 mL media connected to a pump (Cobalt Aquatics model MJ-1200) with flow nozzles to control flowrate. A submerged filtration pump (SUNSUN JP-032F, Sun Microsystems) was used to collect microbial biomass formed during the experimental period (Fig. 2.1). Flowrate was determined by using a stopwatch and a 20 L bucket as outlined in Greiner and Timmons (1998). Two aquarium tanks tested a low flow rate condition (0.3 L / min) and two tested a high flow rate condition (1 L
/ min) with one containing PCL and the other containing PHA beads as reactor fill under each condition using 0.47 kg and 0.38 kg equal parts cultured and new unseeded beads respectively.
One control aquarium tank under high flowrate was prepared with 45 g of polypropylene plastic
(PP) (1” Bio Barrels, Pentair) as bioreactor medium, which is a standard plastic substrate used in aquaculture bioreactors. Phase I of experimental period occurred day 0 to 23 upon feeding tanks with ASN medium, followed by Phase II, day 44 to 53, starting with a complete ASN medium change in each tank.
Water samples were collected with sterilized 250 mL polypropylene sampling bottles
(ThermoScientific Nalgene) every 2 days during Phase I and every 3 days during Phase II and frozen at -20°C before analysis. Total nitrogen (TN), nitrate plus nitrite, total phosphorous (TP), and phosphate were analyzed using a Seal Analytical autoanalyzer using the persulfate digestion method for TN and nitrate plus nitrite and EPA guideline 365.1 for TP and phosphate (Luebbe,
2005; USEPA, 1993c). The removal efficiency of water quality parameters was calculated according to Luo et al. (2018), the amended equation for removal efficiency (%) was calculated from the difference between the starting concentration (determined by the average of the first two measurements) and mean water column concentration of last two measurements. Physical 27 parameters such as water temperature, pH, salinity, and dissolved oxygen (DO) were measured using a YSI Pro Plus meter.
2.4 Bioreactor experiment using aquaculture wastewater
Experimentation using aquaculture wastewater with uncultured biodegradable plastic bead fill was conducted in the temperature-controlled laboratory at the Vester Marine and
Environmental Sciences Research Field Station (26°19’50”N, 81°50’15”W), Florida Gulf Coast
University. Temperature remained relatively stable with a mean of 23.8°C for Phase I and 25.6°C for Phase II with an average salinity of 31.9 and 45.1 ppt, respectively (Table 2.1). Bioreactor flow through systems were setup identical to the ASN medium cultured plastics experiment and the same plastic media (Fig. 2.1). Experimental setup consisted of three 57 L aquaria, filled with
38 L saltwater aquaculture effluent with a flow rate of 1 L / min. The setup consisted of one tank with each bioreactor fill at a volume of 600 mL; PP (0.63 kg), PCL (0.62 kg), and PHA (0.60 kg).
Pinfish (Lagodon rhomboides) housed in an outdoor 1,900 L (500-gallon) tank at the Vester Field
Station provided the aquaculture effluent used in this experiment.
Water samples were collected in the same manner as the ASN medium experiment and frozen at -20°C before analysis, during Phase I, day 0 to 9, collection occurred every 3 days. A complete water change and replacement with new aquaculture effluent occurred on day 45 initiating Phase II, day 45 to 54, collection of water samples occurred daily. In the aquaculture wastewater reactor experiments, nutrient concentrations were analyzed for TN, nitrate-nitrite, TP, and phosphate, and determined using the same methods as described above. Physical parameters such as water temperature, pH, salinity, and dissolved oxygen (DO) were measured using a YSI
Pro Plus meter in the same manner. 28
2.5. Microscopy
Microbial biomass was collected from the filter sponge in the submerged filtration pump
(SUNSUN) upon completion of Phase II of testing for each experimental condition using sterile collection cups to be used for microscopy, cell counting, and DNA analysis. A 10 mL volume from each filter sponge biomass sample was collected in a sterile 15 mL plastic centrifuge tube and fixed with formalin (2% final concentration [vol / vol]). 4’,6-diamidino-2-phenylidole (DAPI) was used to stain cells and fixed samples (0.8 mL) were passed through black 0.22-µm polycarbonate isopore membrane filters (GTBP, MilliporeSigma) using standard hand vacuum pump operation.
An anti-bleaching agent was used as mounting medium (AF1; Citifluor). Cells were observed under 600x magnification using an Olympus BX51 epifluorescence microscope system. More than
10 random fields in each filter were viewed to determine cell numbers.
2.6. Biodegradation of plastics
Plastic beads were weighed at the end of experiments to determine the amount of plastic material loss over the course of experimentation. Wet weights of plastics were used to determine microbial growth directly on beads, once wet weights were obtained plastics were washed clean of their biomass with deionized water and physical separation before drying at 50°C for 4 days in a drying oven for determination of dry weight. Plastic percent degradation was calculated, based on the difference of initial and final dry weight of plastic beads as outlined by Tal et al. (2003).
2.7. High-throughput sequencing of 16S rRNA genes
DNA samples were extracted from filter sponge samples using the Qiagen DNeasy
Powersoil DNA kit according to the manufacturer’s instructions. Extracted DNA was sequenced using the Illumina MiSeq System at RTL Genomics. The PCR amplification of archaeal and 29 bacterial 16S rRNA genes were performed using the primer set, 515yF
(5’GTGYCAGCMGCCGCGGTAA) and 926pfR (5’CCGYCAATTYMTTTRAGTTT) (Parada et al., 2016) tagged with the Illumina i5 forward
(TCGTCGGCAGCGTCAGATGTGTATAAGAGACAG) and i7 reverse
(GTCTCGTGGGCTCGGAGATGTGTATAAGAGACAG) sequencing primer. Each PCR reaction (25 µL) contained Qiagen HotStar Taq master mix, with equal amount of forward and reverse primers (5 µM each), and 1 µL of DNA template (1 to 20 ng). The PCR cycle conditions were as follows; an initial denaturation at 95°C for 5 min, followed by 35 cycles of 94°C for 30 sec, annealing at 54°C for 40 sec, and extension at 72°C for 1 min, with a final extension of 10 min at 72°C. The first stage PCR product was transferred to a second PCR based on qualitatively determined concentrations with primers based on the Illumina Nextera PCR primers forward
(AATGATACGGCGACCACCGAGATCTACAC-[i5 index]-TCGTCGGCAGCGTC) and reverse (CAAGCAGAAGACGGCATACGAGAT-[i7 index]-GTCTCGTGGGCTCGG). The second stage amplification was run the same as the first, except for 10 cycles instead of 35 cycles.
Amplicons were visualized with eGels (Life Technologies), products were pooled equimolar with each size selected quantified using the Quibit 2.0 fluorometer (Life Technologies). Amplicons were then loaded on an Illumina MiSeq (Illumina) 2 x 300 flow cell at 10 pM (RTL Genomics).
For sequence data analysis, FASTQ files were merged in the PEAR Illumina paired-end read merger (Zhang et al., 2013). The USEARCH algorithm was used to complete prefix dereplication (Edgar et al., 2011). Clustering at a 3% divergence level was conducted using
USEARCH (Edgar et al., 2011). Operational taxonomic unit (OTU) selection was performed using the UPARSE-OTU algorithm (Edgar, 2013). Chimera identification was completed using
UCHIME (Edgar, 2010) and detected chimera sequences were removed from further analysis. 30
Representative OTU was used to determine taxonomic information through a basic local alignment search tool (BLAST) at the National Center for Biotechnology Information (NCBI), and MG-
RAST (Meyer et al., 2008). The high-throughput sequences determined in this study were deposited in GenBank under BioProject number PRJNA496040.
2.8. Data analysis
Statistical difference among flowrate conditions and between uses of different reactor media from water quality results was determined using a one-way analysis of variance (ANOVA) with the JMP data analysis software (SAS Institute) (Lehman, 2005). When applicable a Tukey-
Kramer method was employed when one-way ANOVA showed significant differences. Statistical tests were completed using two-tailed and unpaired data analyses with a significant difference shown when p < 0.05. Sequence data statistics were performed using PAST statistical software and MG-RAST (Meyer et al., 2008). Diversity index calculations; Shannon index (Chao and Shen,
2003), Menhinick’s richness (Menhinick, 1964), and Pielou’s evenness (Jost, 2010) were implemented using Microsoft Excel.
3. RESULTS AND DISCUSSION
3.1 Reactor performance
3.1.1 Performance of ASN medium reactor system
The effect of flow rate and carbon source, PCL or PHA biodegradable polymers, were explored in this study (Fig. 2.2). The nitrate-nitrite removal efficiency for Phase I was found to be greater with use of PCL under a high flow and PHA systems, with a removal efficiency of 87.8-
90.3%, PCL low flow and control systems were significantly lower (p ˂ 0.001) (Table 2.2). During
Phase II the nitrate-nitrite removal efficiency increased for all systems ranging 74.7-98.9%, control 31 was significantly lower (p ˂ 0.001). This is consistent with an increase in microbial biomass over time and utilization of biodegradable plastics by heterotrophic bacteria. Nitrate removal efficiency within a biodegradable plastic denitrification reactor system was found to range 67.1 to 98% (Wu et al., 2012, 2013; Chu and Wang, 2016) and 90% within a marine RAS system (Zhu et al., 2015) consistent with our findings. During Phase I the TN removal efficiency was greater with PCL high flow and PHA systems with an efficiency ranging 96.4-97%, PCL low flow and control systems were significantly less efficient (p ˂ 0.001) (Table 2.2). The TN removal efficiency increased in
Phase II for all systems except for a slight decrease in the PCL high flow system, with control efficiency being significantly lower (p ˂ 0.001) (Fig. 2.2). The use of biodegradable polymers enhanced TN removal efficiency with a range of 92.1-98.5%, consistent with similar studies (Chu and Wang, 2011; Yang and Yang, 2011; Chu and Wang, 2013; Zhang et al., 2016). These findings showed that biodegradable plastics enhanced denitrification in a marine RAS with cultured beads regardless of polymer differences.
The phosphate and TP removal efficiencies were variable compared to nitrogen (Fig. 2.2).
In Phase I the phosphate removal efficiency was significantly greater in high flow systems with biodegradable plastic ranging 72.3% to 76.9% (p ˂ 0.001, one-way ANOVA), compared to all other systems (Table 2.2). During Phase II the phosphate removal rate increased in PP and PCL low flow systems, with decreases in efficiency for all other systems, PCL high flow and PHA low flow systems had a significantly lower removal efficiency (p ˂ 0.05, f = 8.42). Overall phosphate removal efficiency for all systems was 0-47.7% and lower than similar findings of phosphate removal ranging 65 to 96% (Martins et al., 2010). Phosphorous is released by bacterial biomass under anaerobic conditions which occurred with the PHA high and low flowrate conditions (van
Rijn, 2006). In Phase I the TP removal efficiency followed the same trend as phosphate, with a 32 significantly greater removal efficiency in the PCL and PHA high flow rate systems (p ˂ 0.001)
(Table 2.2). The TP removal rate increased for PCL and PHA low flow rate systems and control in Phase II, however, PCL and PHA high flow rate systems had a significantly higher efficiency
(p = 0.01), with an overall efficiency ranging 24.5-59.1% (Table 2.2).
3.1.2 Performance of aquaculture wastewater reactor system
The effect of biodegradable plastics to increase the C/N ratio and nutrient removal efficiency with marine aquaculture effluent was explored in this study (Fig. 2.3). During Phase I nitrate-nitrite removal efficiency was greater with the use of biodegradable plastics, 77.9-97%, with no significant differences (Table 2.2). During Phase II, a decrease in nitrate-nitrite removal efficiency occurred, however, the PCL and PHA systems had a significantly greater removal rate than the control (p ˂ 0.001) ranging 37.8-89% which was similar to studies demonstrating a nitrate removal efficiency of 67.1-98% (Chu and Wang, 2013; Zhu et al., 2015; Chu and Wang,2016).
The TN removal efficiency in Phase I was similar for all systems ranging 61.7-71.2%, with a decrease in efficiency in Phase II, 0-62% with a significantly greater efficiency using PCL and
PHA (p ˂ 0.001) (Table 2). This efficiency was lower than that exhibited in the ASN medium experiment as well as similar studies (Chu and Wang, 2011; Yang and Yang, 2011; Chu and Wang
2013; Zhang et al., 2016). The use of PCL and PHA increased nitrogen removal efficiency compared to our control supporting an increase in oxygen consumption and heterotrophic activity with an increased C/N ratio (Fig. 2.3).
In Phase I the phosphate removal efficiency ranged 21.4-26.3% for all systems with no significant differences (Table 2.2). In Phase II, the phosphate removal efficiency increased with the use of PCL and PHA ranging 46.8-70.7% and was significantly lower in the control (p ˂ 0.05).
Some heterotrophic denitrifiers were shown to assimilate phosphate from the breakdown of PHA, 33 supporting our finding in this condition of an increased phosphate removal efficiency with the use of PCL and PHA (van Rijn, 2006). The TP removal efficiency for Phase I ranged 39.5-48% for all systems with no significant differences (Fig. 2.3). Additionally, during Phase II the TP removal efficiency increased with use of PCL and PHA ranging 40.8-58.5%, the control efficiency was significantly lower (p ˂ 0.05) (Table 2.2).
3.2. Bacterial abundance and plastic degradation
The ASN medium experiment produced the highest microbial biomass with PHA beads in both high and low flow conditions. These values ranged from 1.8 (± 1.1) x 106 to 2.8 (± 2.3) x 106 cells / mL (mean ± standard deviation) for low and high flowrate respectively. PCL under the ASN medium experimental conditions produced a microbial biomass of 7.4 (± 4.6) x 105 to 1.1 (± 0.7) x 106 cells / mL. Additionally, the use of PHA reactor produced a higher bacterial biomass than the use of PCL in the aquaculture wastewater experiment with 5.2 (± 2.5) x 106 and 6.6 (± 3.1) x
106 cells / mL for PCL and PHA, respectively. Our experimental controls using PP produced a biomass of 1.3 (± 0.6) x 106 to 2.2 (± 1.0) x 106 cells / mL for both studies, with the lowest bacterial numbers found in the aquaculture wastewater experiment. These counts were within the range of similar studies with cell counts ranging from 106 to 109 per milliliter (Khan et al., 2002; Horiba et al., 2005; Michaud et al., 2014).
Bioplastic degradation was found to correlate with the amount of biofilm growth directly on the plastic beads within the bioreactor chamber and not with the bacterial biomass from the filter sponge, supporting heterotrophic degradation and not physical breakdown of the plastic types. The highest degradation occurred for PHA with ASN medium, ranging 27.6-30.3% (Fig.
2.4). The degradation of PCL from both experiments and PHA with aquaculture wastewater experiment performed similarly 6.6-16.5% (Fig. 2.4). The lowest degradation was observed with 34
PP 0.46-1.52% (Fig. 2.4). A study by Gutierrez-Wing et al. (2012) found that salinity and water temperature did not affect degradation of PHB (polyhydroxy butyrate), however, molecular weight and biofilm biomass did have an affect which is consistent with our finding that use of PHA with an ASN medium resulted in an anoxic condition (Fig. 2.5) but had a greater biofilm biomass formation than PCL, and therefore a greater degradation rate. Additionally, a degradation study of
PCL under varying conditions found a different percent degradation depending on the type of water used, with a degradation of 29.8% using natural seawater and 3.1% using a sterilized artificial seawater after a 52-week period (Lu et al., 2018). These studies demonstrate how environmental factors can alter degradation rate of biodegradable plastics depending on media used.
3.3 Microbial community analysis
A total of 17,763 sequences were analyzed and resulted in 1,473 operational taxonomic units (OTUs) with the highest distribution found in bioreactor samples having PP medium for both experimental conditions (Table 2.3). To examine the microbial diversity three indices were applied, the Menhinick richness, Pielou evenness, and Shannon index. The result of the Menhinick richness, showed PP had a greater richness when compared to all other conditions which supported the OTU distribution result. Community richness decreases as a more selective microbial consortium develops due to water chemistry factors and heterotrophic bacterial selection for polymer selective degraders, conditions found within a RAS system (Zhu et al., 2015). The
Shannon index, looking at both evenness and richness, showed no significant differences when comparing plastics and operating conditions (i.e. ASN medium and aquaculture wastewater) (p ˃
0.05) (Table 2.3). The taxonomic analysis identified 25 classes from all samples: 9-17 collected from the ASN medium samples and 13-23 in the aquaculture wastewater samples. The dominant 35 classes present in our study were Alphaproteobacteria (2-68%), Gammaproteobacteria (2-30%),
Deltaproteobacteria (0.08-8%), Planctomycetia (0.2-24%), Flavobacteriia (0.5-57%), and
Sphingobacteriia (0.2-13%) (Fig. 2.5).
Alphaproteobacteria was found to contain the most diverse genera (Table 2.4). Members of the family Rhodobacteraceae dominated in our study (i.e. Roseobacter, Ruegeria, Rhodovulum,
Oceanicola, Jannaschia, and Stappia) from both experimental systems, ranging 1-24%.
Rhodobacteraceae are known as primary surface colonizers in marine biofilms and comprise physiologically and morphologically diverse members (Cho and Giovannoni, 2004; Michaud et al., 2009; Blancheton et al., 2013). The heterotrophic genera Ruegeria (0-20%) and Roseobacter
(0-4%) were identified in both experimental systems and are major marine lineages able to colonize in marine biofilms (Buchan et al., 2005; Michaud et al., 2009; Schreier et al., 2010;
Blancheton et al., 2013; Higuchi-Takeuchi et al., 2016), suggesting their importance to the complexity of the aquatic microbial community. The genus Rhodovulum (1%) was identified in both experimental systems in the PCL high flow and PHA samples. The Rhodovulum species
Rhodovulum sulfidophilum can synthesize the polymeric storage molecule polyhydroxybutryate
(PHB) (Cai et al., 2012; Higuchi-Takeuchi et al., 2016). Additionally, the genus Oceanicola (24%) identified in the PCL aquaculture wastewater sample contain the species Oceanicola granulosus and O. batsensis known to synthesize PHB (Cho and Giovannoni, 2004), these findings may suggest that these clades are able to degrade these storage molecules for intracellular use in a carbon limited environment such as in a marine RAS.
Heterotrophic denitrification is the major nitrate removal pathway within RAS where a biodegradable polymer is utilized as a substrate and carbon source to initiate breakdown of nitrate to gaseous nitrogen (van Rijn et al., 2006; Shen et al., 2013a, b; Zhu et al., 2015). The dominant 36 genera of denitrifying bacteria found in marine and freshwater RAS can be quite extensive and differ depending on fish-rearing conditions and fish species (Wang and Chu, 2016). Marinobacter presented in all samples in the ASN medium experiment (1-9%), whereas, Nitratireductor was specifically identified with use of PCL. The genera Marinobacter and Nitratireductor are common marine denitrifiers belonging to the classes Alphaproteobacteria and Gammaproteobacteria, respectively, and are shown to play a role in bioreactor systems (Labbè et al., 2004; Madigan et al., 2012; Blancheton et al., 2013; Shi et al., 2013). This suggests denitrification in each system was dependent on different genera based on the type of biodegradable polymer and effluent conditions. The heterotrophic genera Fluviicola and Crocinitomix within Flavobacteriia were identified in the ASN medium experiment and can utilize complex carbon sources (Rodrigues-
Sanchez et al., 2016), suggesting the ability to utilize PCL and PHA (Table 2.4). Crocinitomix was a major genus identified in this study with an abundance of 2-40% with use of PCL medium, a greater abundance under low flow rate, and also present in minor amount in the low flow PHA system.
Two genera known to contain polymer degrading species were identified in this study,
Clostridium and Rhodococcus. The finding of Clostridium in the ASN medium experiment with
PHA media (7%) and PCL low flow rate (1%), this genus contains organisms able to degrade biopolymers under anaerobic conditions and it was proposed that the availability of degraded byproducts can enhance denitrifiers (Tokiwa et al., 2012; Chu and Wang, 2016; Qiu et al., 2016).
Additionally, several studies found Clostridium as the dominant genus present utilizing PHBV
(poly-3-hydroxybutryate-co-hydroxyvelate) and an artificial wastewater medium under anaerobic conditions (Chu and Wang, 2016; Qui et al. 2016). This was supported by our taxon distribution showing Clostridia was found with the sulfate-reducing order Desulfovibrionales and genus 37
Desulforhopalus in the PHA high and low flow rate samples from the ASN medium experimental condition, which exhibited the highest biomass production and subsequent oxygen depletion within the system (Fig. 2.4&2.5). Additionally, the genus Rhodococcus was identified in the ASN medium study in PCL systems, 1%. According to Dash et al. (2013), Rhodococcus ruber was found to degrade plastic and Rhodococcus sp. has been shown to degrade naphthalene when isolated from marine environments. Additionally, Rhodococcus ruber was found to degrade polystyrene in a flask study (Mor and Sivan, 2008). This suggests that Rhodococcus sp. may be able to utilize other biodegradable polymers, however, not enough evidence was present to demonstrate use of PCL as a carbon source. The genera Alteromonas, Muricauda and
Winogradskyella were identified in greater than a 10% abundance with the use of PHA and the genus Meridianimaribacter showed 27% abundance with PCL suggesting the use of these polymers as a carbon source, but no studies were found to support this finding (Table 2.4).
4. CONCLUSIONS
Biodegradable plastics can be utilized in marine RAS to treat wastewater through microbial processes. It was found that the nutrient cycling efficiency was greater with an ASN medium and seeded polymers, allowing for major heterotrophic denitrifiers to develop compared to a system which utilized aquaculture wastewater with the microbial community naturally present in the system. Heterotrophic bacteria (e.g. Nitratireductor, Marinobacter, Fluviicola and Crocinitomix) were identified in this study along with known genera capable of polymer degradation (e.g.
Clostridium and Rhodococcus) showing the complexity of consortia present within aquarium sludge. Aquaculture environments can develop complex consortia dependent on physiochemical factors and our results demonstrated these differences based on wastewater source and varying 38 systems management (Schreier et al., 2010; Blancheton et al., 2013; Michaud et al., 2014). Overall, biodegradable plastic showed great potential for nutrient removal in a marine RAS under varying flow rate and water conditions.
39
Table 2.1 Operating conditions of aquaria during nutrient removal experiments* Phase Influent type Salinity (ppt) Temperature (°C) pH Operation time (d) I ASN medium 32.7 ± 1.77 25.0 ± 1.07 7.2 ± 0.71 0 - 23 II 32.6 ± 1.34 26.1 ± 0.62 7.8 ± 0.31 44 - 53 I Aquaculture wastewater 31.9 ± 0.51 23.8 ± 0.50 7.1 ± 0.56 0 - 9 II 45.1 ± 2.82 25.9 ± 0.60 7.8 ± 0.10 45 - 54 *Average salinity, temperature, and pH over experimental period for Phases I and II. Operation time denotes Phase I and Phase II in days.
40
Table 2.2 Bioreactor performance based on the removal efficiency of nutrients
Influent type Removal efficiency (%) - - 3- Phase I NO3 + NO2 TN PO4 TP PP 4.0 19.0 34.6 2.3 PCLhf 90.3 96.4 72.3 62.0 PCLlf 13.4 9.2 0.0 0.0 PHAhf 90.2 97.0 76.9 60.7 PHAlf 87.8 83.9 27.9 0.0 ASN medium Phase II PP 74.7 78.2 47.7 55.3 PCLhf 95.4 92.1 0.0 29.2 PCLlf 98.1 96.8 42.6 55.3 PHAhf 98.9 98.3 41.2 59.1 PHAlf 98.8 98.5 0.0 24.5
- - 3- Phase I NO3 + NO2 TN PO4 TP PP 49.3 61.7 26.3 47.8 PCL 97.0 67.2 21.4 39.5 PHA 77.9 71.2 25.0 48.0 Aquaculture wastewater Phase II PP 0.0 0.0 0.0 0.0 PCL 89.0 62.0 70.7 58.5 PHA 37.8 61.5 46.8 40.8
*Removal efficiency calculated as described previously for Phase I and Phase II. Influent type denotes experimental condition and sample name showing media type used and flowrate where applicable.
41
Table 2.3 Summary high-throughput sequencing and diversity indices*
Mean sequence Pielou Menhinick Shannon Influent type Sequences length OTU evenness richness index PP 1,817 411 ± 3 203 0.55 0.82 2.44 PCLhf 1,867 411 ± 4 131 0.59 0.72 2.51 ASN medium PCLlf 2,154 410 ± 2 140 0.53 0.79 2.31 PHAhf 1,949 410 ± 10 81 0.67 0.49 2.61 PHAlf 1,195 410 ± 16 122 0.70 0.72 2.99 PP 3,694 412 ± 8 532 0.62 1.58 3.13 Aquaculture wastewater PCL 3,451 411 ± 9 157 0.66 0.69 2.81 PHA 1,636 406 ± 8 107 0.65 0.51 2.56 *Diversity indices were calculated after normalization to 10,000 reads.
42
Table 2.4 Summary percent abundance at the genus level*
Class ASN medium Aquaculture wastewater Genus/Order PP PCLhf PCLlf PHAhf PHAlf PP PCL PHA Acidobacteriia Acidobacteriales 1 0 0 0 0 0 0 1 Actinobacteria Gordonia 0 4 5 0 0 0 0 0 Mycobacterium 12 0 0 0 0 2 0 1 Rhodococcus 0 1 1 0 0 0 0 0 Microbacterium 0 0 0 0 0 0 2 0 Cytophagia Cytophaga 0 0 0 0 1 0 0 0 Ekhidna 1 0 0 0 0 0 0 0 Marinoscillum 1 0 0 0 0 0 0 1 Cytophagales 1 0 1 0 1 1 0 0 Flavobacteriia Crocinitomix 0 11 40 0 2 0 0 0 Fluviicola 0 3 0 0 0 0 0 0 Formosa 0 2 1 0 0 0 0 0 Meridianimaribacter 0 27 0 0 0 0 0 1 Muricauda 0 1 0 4 11 0 0 0 Olleya 0 0 0 0 0 0 0 4 Winogradskyella 0 1 0 15 0 1 0 23 Flavobacteriales 0 0 16 0 0 1 0 0 Sphingobacteriia Lewinella 11 1 3 2 7 0 0 0 Sphingobacterium 1 0 3 0 0 2 0 1 Sphingobacteriales 0 0 0 4 2 3 0 0 Dehalococcoidia Dehalococcoides 1 0 0 0 0 0 0 0 Cyanobacteria Xenococcus 0 0 0 0 0 1 0 0 Clostridia Clostridium 0 0 1 7 7 0 0 0 Planctomycetia Pirellula 1 0 0 0 0 0 0 1 Planctomyces 18 4 1 0 1 4 3 3 Rhodopirellula 3 1 1 0 0 0 0 0 Planctomycetales 2 3 1 0 0 3 4 1 Alphaproteobacteria Kordiimonas 0 0 0 0 0 0 0 1 Terasakiella 0 0 0 1 1 0 0 0 Mesorhizobium 0 0 0 1 0 0 4 0 Nitratireductor 1 0 0 0 0 0 4 0 Anderseniella 0 0 0 0 0 0 1 0 Bauldia 0 0 0 0 0 0 1 0 Methyloceanibacter 0 0 0 0 0 0 1 0 Rhizobiales 0 0 0 0 0 0 2 0 Hyphomonas 0 0 0 0 2 0 1 0 Jannaschia 0 0 0 0 0 0 2 0 Oceanicola 0 0 0 0 0 0 24 0 Rhodovulum 0 1 0 0 0 0 0 1 Roseobacter 0 1 0 1 2 1 4 0 43
Ruegeria 0 20 1 2 1 1 18 1 Stappia 0 1 0 0 0 0 0 1 Rhodobacterales 0 0 0 0 0 0 2 0 Thalassospira 0 0 0 3 0 0 0 0 Rhodospirillales 2 0 0 0 0 2 0 0 Sphingopyxis 0 0 0 0 0 0 0 1 Deltaproteobacteria Bacteriovorax 0 0 0 0 0 0 0 1 Desulforhopalus 0 0 0 1 1 0 0 0 Desulfovibrionales 0 0 0 5 5 0 0 0 Kofleria 0 0 0 0 0 0 0 1 Nannocystis 0 0 0 0 0 1 0 0 Myxococcales 1 0 0 0 1 0 0 0 Gammaproteobacteria Alteromonas 0 0 0 17 6 0 3 0 Marinobacter 1 3 0 9 8 2 0 0 Alteromonadales 0 0 1 0 0 0 0 0 Granulosicoccus 0 2 1 2 3 0 0 0 Coxiella 0 0 0 0 0 0 1 0 Legionella 1 0 1 0 0 0 1 0 Neptuniibacter 0 0 0 1 1 0 0 0 Oceanobacter 0 0 0 0 2 0 0 0 Oceanospirillum 0 0 0 1 0 0 0 0 Haemophilus 0 0 0 0 0 0 0 1 Verrucomicrobiae Verrucomicrobium 0 0 0 0 1 0 0 0 *Shown are percent bacterial relative abundance greater than 1% at least in one sample after normalization to 10,000 reads per sample. Cyanobacteria are shown at the phylum level as class level is unidentified. Order level identification was made when genus level was unidentifiable.
44
Figure 2.1 Experimental setting of aquaria. (a) The reactors used and (b) schematic overview including the tubing system. 45
Figure 2.2 Succession of oxygen, TN and TP in the experiment with ASN medium. Phase I and Phase II are shown over experimental period.
46
Figure 2.3 Succession of oxygen, TN and TP in the experiment with aquaculture wastewater. Phase I and Phase II are shown over experimental period. 47
Figure 2.4 Biofilm growth on plastic surface and percent decrease of plastic weight. Biomass growth represents biofilm growth on the plastic media used in each bioreactor calculated for PP, PCL, and PHA over experimental period for all conditions and correlation to decrease plastic weight. 48
Figure 2.5 Class-level relative abundance of microbial communities. Bacterial and archaeal abundance of sludge samples denoted by experiment and conditions used. The relative abundance shown is based on the normalized data per 10,000 reads. Shown are abundance greater than 30 reads, with those below grouped together as category Others. Inset shows the oxygen level on the ending date of Phase II experiments.
49
References
Blancheton, J.P., Attramadal, K.J.K., Michaud, L., d'Orbcastel, E.R., Vadstein, O., 2013. Insight
into bacterial population in aquaculture systems and its implication. Aquacultural
Engineering. 53, 30-39.
Boley, A., Muller, W.-R., Haider, G., 2000. Biodegradable polymers as solid substrate and biofilm
carrier for denitrification in recirculated aquaculture systems. Aquacultural Engineering.
22, 75-85.
Buchan, A., González, J. M., Moran, M. A., 2005. Overview of the marine Roseobacter lineage.
Applied and Environmental Microbiology. 71, 5665-5677.
Cai, J., Wei, Y., Zhao, Y., Pan, G., Wang, G., 2012. Production of polyhydroxybutyrate by the
marine photosynthetic bacterium Rhodovulum sulfidophilum P5. Chinese Journal of
Oceanology and Limnology. 30, 620-626.
Chao, A., Shen, T.-J., 2003. Nonparametric estimation of Shannon’s index of diversity when there
are unseen species in sample. Environmental and Ecological Statistics. 10, 429-443.
Cho, J.-C., Giovannoni, S.J., 2004. Oceanicola granulosus gen. nov., sp. nov. and Oceanicola
batsensis sp. nov., poly-β-hydroxybutyrate-producing marine bacteria in the order
‘Rhodobacterales’. International Journal of Systematic and Evolutionary Microbiology.
54, 1129-1136.
Chu, L., Wang, J., 2011. Nitrogen removal using biodegradable polymers as carbon source and
biofilm carriers in a moving bed biofilm reactor. Chemical Engineering Journal. 170, 220-
225. 50
Chu, L., Wang, J., 2013. Denitrification performance and biofilm characteristics using
biodegradable polymers PCL as carriers and carbon source. Chemosphere. 91, 1310-1316.
Chu, L., Wang, J., 2016. Denitrification of groundwater using PHBV blends in packed bed reactors
and the microbial diversity. Chemosphere. 155, 463-470.
Dash, H.R., Mangwani, N., Chakraborty, J., Kumari, S., Das, S., 2013. Marine bacteria: potential
candidates for enhanced bioremediation. Applied Microbiology and Biotechnology. 97,
561-571. doi:10.1007/s00253-012-4584-0
Edgar, R.C., 2010. Search and clustering orders of magnitude faster than BLAST. Bioinformatics.
26, 2460-2461.
Edgar, R.C., 2013. UPARSE: highly accurate OTU sequences from microbial amplicon reads.
Nature Methods. 10, 996-998.
Edgar, R.C., Haas, B.J., Clemente, J.C., Quince, C., Knight, R., 2011. UCHIME improves
sensitivity and speed of chimera detection. Bioinformatics. 27, 2194-2200.
Gao, X.-Y., Xu, Y., Liu, Y., Liu, Z.-P., 2012. Bacterial diversity, community structure and function
associated with biofilm development in a biological aerated filter in a recirculating marine
aquaculture system. Marine Biodiversity. 42, 1-11.
Greiner, A.D., Timmons, M.B., 1998. Evaluation of the nitrification rates of microbead and
trickling filters in an intensive recirculating tilapia production facility. Aquacultural
Engineering. 18, 189-200. 51
Guiterrez-Wing, M.T., Malone, R.F., Rusch, K.A., 2012. Evaluation of polyhydroxybutryate as a
carbon source for recirculating aquaculture water denitrification. Aquacultural
Engineering. 51, 36-43.
Guiterrez-Wing, M.T., Malone, R.F., Rusch, K.A., 2014. Development of a model for PHA-based
denitrification in a packed bed reactor. Aquacultural Engineering. 60, 41-47.
Higuchi-Takeuchi, M., Morisaki, K., Toyooka, K., Numata, K., 2016. Synthesis of high-
molecular-weight polyhydroxyalkanoates by marine photosynthetic purple bacteria. PLOS
ONE. 11, e0160981. doi:10.1371/journal.pone.0160981
Horiba, Y., Khan, S., Hiraishi, A., 2005. Characterization of the microbial community and
culturable denitrifying bacteria in a solid-phase denitrification process using poly(ε-
caprolactone) as the carbon and energy source. Microbes and Environments. 20, 25-33.
Ishigaki, T., Sugano, W., Nakanishi, A., Tateda, M., Ike, M., Fujita, M., 2004. The degradability
of biodegradable plastics in aerobic and anaerobic waste landfill model reactors.
Chemosphere. 54, 225-233.
Jost, L., 2010. The relation between evenness and diversity. Diversity. 2, 207-232.
Khan, S.T., Horiba, Y., Yamamoto, M., Hiraishi, A., 2002. Members of the family
Comamonadaceae as primary poly (3-hydroxybutyrate-co-3-hydroxyvalerate)-degrading
denitrifiers in activated sludge as revealed by a polyphasic approach. Applied and
Environmental Microbiology. 68, 3206-3214.
Kumar, R.V.J., Sukumaran, V., Achuthan, C., Joseph, V., Philip, R., Singh B.I.S., 2013. Molecular
characterization of the nitrifying bacterial consortia employed for the activation of 52
bioreactors used in brackish and marine aquaculture systems. International
Biodeterioration & Biodegradation. 78, 74-81.
Labbé, N., Parent, S., Villemur, R., 2004. Nitratireductor aquibiodomus gen. nov., sp. nov., a
novel α-proteobacterium from the marine denitrification system of the Montreal Biodome
(Canada). International Journal of Systematic and Evolutionary Microbiology. 54, 269-
273.
Lehman, A., 2005. JMP for basic univariate and multivariate statistics: a step-by-step guide: SAS
Institute.
Luebbe, B., 2005. Nitrate and nitrite in water and seawater and total nitrogen in persulfate digests,
Autoanalyzer method no. G-172-96-RIO, Rev. 10.
Lu, B., Wang, G.-X., Huang, D., Ren, Z.-L., Wang, X.-W., Wang, P.-L., Zhen, Z.-C., Zhang, W.,
Ji, J.-H., 2018. Comparison of PCL degradation in different aquatic environments: Effects
of bacteria and inorganic salts. Polymer Degradation and Stability. 150, 133-139.
Luo, G., Liu, Z., Gao, J., Hou, Z., Tan, H., 2018. Nitrate removal efficiency and bacterial
community of polycaprolactone-packed bioreactors treating water from a recirculating
aquaculture system. Aquaculture International. 26, 773-784.
Madigan, M.T., Martinko, J.M., Stahl, D. A., Clark. D.P., 2012. Brock biology of microorganisms:
Pearson, California.
Martins, C.I.M., Eding, E.H., Verdegem, M.C.J., Heinsbroek, L.T.N., Schneider, O., Blancheton,
J.P., d’Orbcastel, R.E., Verreth, J.A.J., 2010. New developments in recirculating 53
aquaculture systems in Europe: A perspective on environmental sustainability.
Aquacultural Engineering. 43, 83-93.
Menhinick, E.F., 1964. A comparison of some species-individuals diversity indices applied to
samples of field insects. Ecological Society of America. 45, 859-861.
Meyer, F., Paarmann, D., D'Souza, M., Olson, R., Glass, E., Kubal, M., Paczian, T., Rodrigues,
A., Stevens., R., Wilke, A., Wilkening, J., Edwards, R., 2008. The metagenomics RAST
server – a public resource for the automatic phylogenetic and functional analysis of
metagenomes. BMC Bioinformatics. 9, 386.
Michaud, L., Lo Giudice, A., Interdonato, F., Triplet, S., Ying, L., Blancheton, J.-P., 2014. C/N
ratio-induced structural shift of bacterial communities inside lab-scale aquaculture
biofilters. Aquacultural Engineering. 58, 77-87.
Michaud, L., Lo Giudice, A., Troussellier, M., Smedile, F., Bruni, V., Blancheton, J.-P., 2009.
Phylogenetic characterization of the heterotrophic bacterial communities inhabiting a
marine recirculating aquaculture system. Journal of Applied Microbiology. 107, 1935-
1946.
Mor, R.,Sivan, A., 2008. Biofilm formation and partial biodegradation of polystyrene by the
actinomycete Rhodococcus ruber. Biodegradation. 19, 851-858.
Parada, A.E., Needham, D.M., Fuhrman, J.A., 2016. Every base matters: assessing small subunit
rRNA primers for marine microbiomes with mock communities, time series and global
field samples. Environmental Microbiology. 18, 1403-1414. 54
Qiu, T., Liu, L., Gao, M., Zhang, L., Tursun, H., Wang, X., 2016. Effects of solid-phase
denitrification on the nitrate removal and bacterial community structure in recirculating
aquaculture system. Biodegradation. 27, 165-178.
Rodriguez-Sanchez, A., Purswani, J., Lotti, T., Maza-Marquez, P., van Loosdrecht, M.C.M.,
Vahala, R., Gonzalez-Martinez, A., 2016. Distribution and microbial community structure
analysis of a single-stage partial nitritation/anammox granular sludge bioreactor operating
at low temperature. Environmental Technology. 37, 2281-2291.
Schreier, H. J., Mirzoyan, N., Saito, K., 2010. Microbial diversity of biological filters in
recirculating aquaculture systems. Current Opinion in Biotechnology. 21, 318-325.
Shen, Z., Zhou, Y., Hu, J., Wang, J., 2013. Denitrification performance and microbial diversity in
a packed-bed bioreactor using biodegradable polymer as carbon source and biofilm
support. Journal of Hazardous Materials. 250, 431-438.
Shen, Z., Zhou, Y., Wang, J., 2013. Comparison of denitrification performance and microbial
diversity using starch/polylactic acid blends and ethanol as electron donor for nitrate
removal. Bioresource Technology. 131, 33-39.
Shi, Z., Zhang, Y., Zhou, J., Chen, M., Wang, X., 2013. Biological removal of nitrate and
ammonium under aerobic atmosphere by Paracoccus versutus LYM. Bioresource
Technology. 148, 144-148.
Tal, Y., Watts, J.E., Schreier, S.B., Sowers, K.R., Schreier, H.J., 2003. Characterization of the
microbial community and nitrogen transformation processes associated with moving bed
bioreactors in a closed recirculated mariculture system. Aquaculture. 215, 187-202. 55
Tan, C., Ma, F., Qui, S., Zeng, H., Zhou, Y., 2013. Study of biodegradable polyurethane foam as
carriers for low C/N ratio wastewater. Applied Mechanics and Materials. 284-287, 352-
356.
Tokiwa, Y., Calabia, B.P., Ugwu, C.U., Aiba, S., 2009. Biodegradability of plastics. International
Journal of Molecular Science. 10, 3722-3742.
United States Environmental Protection Agency [USEPA], 1993c. Method 365.1: Determination
of phosphorous by semi-automated colorimetry (Revision 2.0), Cincinnati, Ohio. van Rijn, J., 1996. The potential for integrated biological treatment systems in recirculating fish
culture—A review. Aquaculture. 139, 181-201. van Rijn, J., Tal, Y., Schreier, H.J., 2006. Denitrification in recirculating systems: Theory and
applications. Aquacultural Engineering. 34, 364-376.
Wang, J., Chu, L., 2016. Biological nitrate removal from water and wastewater by solid-phase
denitrification process. Biotechnology Advances. 34, 1103-1112.
Wu, W., Yang, F., Yang, L., 2012. Biological denitrification with a novel biodegradable polymer
as carbon source and biofilm carrier. Bioresource Technology. 118, 136-140.
Wu, W., Yang, L., Wang, J., 2013. Denitrification performance and microbial diversity in a
packed-bed bioreactor using PCL as carbon source and biofilm carrier. Applied
Microbiology and Biotechnology. 97, 2725-2733.
Yang, S., Yang, F., 2011. Nitrogen removal via short-cut simultaneous nitrification and
denitrification in an intermittently aerated moving bed membrane bioreactor. Journal of
Hazardous Materials. 195, 318-323. 56
Zhang, J., Kobert, K., Flouri, T., Stamatakis, A., 2013. PEAR: a fast and accurate Illumina Paired-
End reAd mergeR. Bioinformatics. 30, 614-620.
Zhang, Q., Ji, F., Xu, X., 2016. Effects of physicochemical properties of poly-ε-caprolactone on
nitrate removal efficiency during solid-phase denitrification. Chemical Engineering
Journal. 283, 604-613.
Zhu, S.-M., Deng, Y.-L., Ruan, Y.-J., Guo, X.-S., Shi, M.-M., Shen, J.-Z., 2015. Biological
denitrification using poly (butylene succinate) as carbon source and biofilm carrier for
recirculating aquaculture system effluent treatment. Bioresource Technology. 192, 603-
610.
57
CHAPTER 3
Comparison of nutrient retention efficiency between vertical-flow and floating treatment wetland mesocosms with and without biodegradable plastic
The original work found in the following chapter has been submitted for review in the journal
Ecological Engineering
Cristina R. Lopardo, Li Zhang, William J. Mitsch, Hidetoshi Urakawa
58
ABSTRACT
Treatment wetlands are ecological systems that are engineered to improve polluted water quality through macrophyte, soil, and microbial remediation and are used commonly for urban and agricultural runoff treatment. However, constructed wetlands used for marine aquaculture effluent treatments are understudied when compared to their freshwater counterpart. We compared the nutrient retention and the microbial communities of two types of constructed wetland mesocosms, a vertical-flow treatment wetland (VFTW) and floating treatment wetland (FTW) in subtropical south Florida. To enhance nutrient retention efficiency, we implemented biodegradable plastic
(polycaprolactone), as an external carbon source and monitored the performance of VFTW and
FTW for the treatment of marine aquaculture effluent. Polycaprolactone surface were covered by various cyanobacterial genera including Oscillatoria, Leptolyngbya, Brasilonema, and Trichormus and some plastic-degrading bacteria such as Pseudomonas. The presence of a biodegradable plastic in FTW improved the overall performance of nitrogen removal (nitrite plus nitrate) by 14% through denitrification. The pattern of nutrient removal between two treatment wetland mesocosms were significantly different (p ˂ 0.01), with over 87-91% retention of total nitrogen in
VFTW and no retention in FTW, the latter due to poor retention of nitrite plus nitrate and production of organic nitrogen from the system not present in inflow waters. Total phosphorus was retained in both mesocosm types, with higher retention (74-81%) in the VFTW than in the FTW
(17-40%). The nutrient retention in VFTW was higher overall compared with FTW mesocosms regardless of biodegradable plastic presence.
59
1. INTRODUCTION
Wetland construction or wetland restoration has been effective in water quality enhancement through nutrient reductions from agricultural and urban runoff (Fink and Mitsch,
2004; Nahlik and Mitsch, 2006; Mitsch et al., 2012, 2015; Griffiths and Mitsch, 2017). Treatment wetlands are ecological systems that are engineered to treat polluted water through macrophyte, soil, and microbial remediation and have some varieties (Vymazal, 2007). Vertical-flow treatment wetlands (VFTWs) are fed inflows intermittently or continuously with a relatively short hydraulic residence time (Stottmeister et al., 2003; De Lange et al., 2013) and effective for solids removal from the water column and nutrient cycling by means of phytoremediation and microbial processes
(e.g. denitrification and nitrification) (Fuchs et al., 2011; De Lange et al., 2013). A floating treatment wetland (FTW) is a relatively new phytoremediation technique to reduce the impact of excess nutrient loading within the waterbody itself. FTWs consist of aquatic or terrestrial plants grown hydroponically on a floating mat directly in the open water of the system (Hubbard et al.,
2004; Vymazal, 2007; Olguín et al., 2017; Pavlineri et al., 2017). The plant roots are exposed directly to the water column instead of buried in a sand or gravel substrate allowing for nutrients to be absorbed hydroponically, reducing the nutrient load internally (Zhou and Wang, 2010; White and Cousins, 2013). Plant roots have a greater surface area exposure in the water column, which allow for greater bacterial colonization and unique rhizosphere microbial functions (Zhao et al.,
2012; White and Cousins, 2013; Urakawa et al., 2017). These two wetland designs (VFTW and
FTW) have been proven to be effective for agriculture and storm water treatments (Faulwetter et al., 2011; Zhang et al., 2013b; Liu et al., 2016; Fu et al., 2017; Urakawa et al., 2017).
Since treatment wetlands have been specifically designed for wastewater treatment removing high nutrients and suspended solids (Turcios and Papenbrock, 2014; Mitsch and 60
Gosselink, 2015), it is possible to apply treatment wetlands to remediate aquaculture wastewater as a cost-effective approach (Brown et al., 1999; Lin et al., 2010; Liang et al., 2017). One of the most common treatment wetlands for aquaculture effluent is characterized as subsurface flow construction, which has a sand or gravel substrate, where water flows either vertically (vertical- flow) or horizontally (horizontal-flow), and treated water is either reused in a closed system or discharged in an open system (Konnerup et al., 2011; Mitsch and Gosselink, 2015). Enhancement of aquaculture wastewater treatment capacity could be possible through the addition of various external carbon sources such as methanol, glucose, starch, and cellulose (Wu et al., 2014). Several studies aimed to explore different denitrification activity with external carbon in constructed wetlands, use of periphyton as a producer of organic carbon (Sirivedhin and Gray, 2006), and addition of different sugars (e.g. glucose and fructose) to wetland influents (Lin et al., 2002; Lu et al., 2009).
Using biodegradable plastic as an external carbon source in treatment wetlands is a new approach and two benchtop scale wetland microcosms were previously designed with the use of a cornstarch/polycaprolactone blend (Shen et al., 2015) and poly-3-hydroxybutryate-co-3- hydroxyvalerate/polyacetic acid (PHBV/PLA) (Yang et al., 2018). However, no application has been made in a medium scale outdoor treatment wetland and understanding the microbial community composition in a treatment wetland with biodegradable plastic is the next question to improve the performance of nutrient removal.
In this study, we evaluated nutrient retention efficiency between vertical-flow and floating treatment wetland mesocosms with and without biodegradable plastic (polycaprolactone) for treatment of marine aquaculture effluent to enhance nutrient cycling (e.g. denitrification).
Likewise, determining how microbial community composition could change with the addition of 61 a biodegradable plastic and how microbial community composition differs between vertical-flow and floating treatment wetland mesocosms was concurrently studied for a better understanding of microbial community composition of these two treatment wetland systems.
2. MATERIALS AND METHODS
2.1 Vertical-flow treatment wetland construction
In May 2016, the experimental units were constructed at the Everglades Wetland Research
Park of Florida Gulf Coast University (26°06.452´N, 81°46.334´W) in two rows of four wetland mesocosms (1.33 m x 0.47 m x 0.61 m polyethylene tubs) with one row modeling vertical-flow treatment wetlands (VFTW) and one row modeling floating treatment wetlands (FTW) in a batch system (Fig. 3.1). Vertical-flow constructed mesocosms were filled with a 10 cm layer of gravel followed by an approximate 30 cm of sand fill according to methods outlined in Ahn et al. (2001) and Ahn and Mitsch (2002) (Fig. 3.1). Cordgrass (Spartina patens) was collected from a nearby
23-ha restored brackish marsh (5 ppt) at the Naples Botanical Garden (26°06.181´N,
81°46.534´W) (Zhang et al., 2017) and planted in August 2016. Salinity was gradually increased from 0 to 5 ppt to acclimate plants for 10 months prior to starting the experiment. Two mesocosms were incorporated with a 1 cm layer of polycaprolactone beads (3.5 mm in diameter, IC3D,
TechTack Moldable plastic) buried at a depth of 9 cm (1.82 kg) within the upper substrate layer prior to effluent feeding to mesocosms to allow for settling of the plastics, which are buoyant in water, as upper substrate layer was no longer densely packed after burial. Two other mesocosms were used as a control without biodegradable plastic incorporation. A nutrient removal experiment was conducted during June 2017. 62
2.2 Floating treatment wetland construction
Four floating mat treatment mesocosms were filled with lake water pumped up from an adjacent lake. Each mat had 18 plantings (9 cm diameter) spaced 25 cm apart from the center of each hole (Fig. 3.1). Seven-cm cordgrass (S. patens) plants were placed in aerator pots seated within the floating mats. Artificial saltwater (Instant Ocean) was used to adjust salinity to be 5 ppt.
A recirculating bioreactor system was equipped in all four floating treatment wetlands: two mesocosms had bioreactors with polycaprolactone (PCL) plastic beads as a reactor medium with two mesocosms having empty bioreactors used as control. The recirculating bioreactor setup consisted of 250 mL biodegradable beads (472 g) in AQUAMAXX bioreactors (1 L volume) connected with a filter pump (Cobalt MJ-1200) and a flow nozzle controlled to a flowrate at 1 L min-1.
2.3 Upstream tank setup
A 560-liter upstream tank (dimensions 1.00 x 0.8 x 0.7 m3) housed 10 Pinfish (Lagodon rhomboides) used to generate the brackish aquaculture wastewater (Fig. 3.2). A filtration system consisted of a canister filter with ultraviolet sterilizer lamp (55.9 cm h x 35.6 cm d, Red Sea brand) and a 3.8 L bioreactor (NextReef, MR1 XL) with polypropylene plastic fill (2.54 cm Bio Barrels,
Pentair) and filtration pump (Maxi-jet Pro Powerhead, Pentair). The brackish aquaculture wastewater was fed manually once a week for one month prior to starting the experiment and then every six days after start for two months to both systems.
2.4 Water sampling and chemical analysis
Water samples were collected in 250 mL autoclaved polypropylene sampling bottles
(ThermoScientific Nalgene) from the outflow pipe of mesocosms (Fig. 3.1) and stored at -20°C 63 until analysis. The hydraulic loading rate of the vertical-flow systems were set to be 3.03 L day-1
(48.4 cm day-1), manually fed to the system from the upstream tank, which allowed for a complete flow-through of three days to the outflow pipe. The hydraulic loading rate (HLR) was determined according to Mitsch and Gosselink (2015) using the following equation, q = 100Q / A, where q =
(HLR), (cm day-1), Q = inflow rate, m3 day-1, and A = wetland surface area, (m2). Water quality parameters such as water temperature, pH, salinity, and dissolved oxygen (DO) were measured in the FTW mesocosms using a YSI Pro Plus meter. Turbidity was determined using a Trilogy fluorometer with a turbidity module (Turner Design). Ammonia concentration was colorimetrically conducted using a Spectronic Genesys 20 spectrophotometer (Thermo Scientific) using a standard sodium salicylate method. Nutrients in water samples were colorimetrically determined using a SmartChem Autoanalyzer to measure nitrate-nitrite nitrogen and Total
Kjeldahl nitrogen (TKN) according to EPA guidelines 353.1 and 351.2 respectively (USEPA,
1993b, a). Total nitrogen was determined from the combined TKN and nitrate-nitrite concentrations. Total phosphorus (TP) concentration was determined according to the EPA guideline 365.1 (USEPA, 1993c).
2.5 Plant tissue samples
Aboveground plant tissue samples (9 cm2) were randomly collected from all mesocosms at the start and end of the experiment. Changes in plant stem height were measured for an estimate of daily growth rate over each system period.
2.6 Microscopy
Water samples were collected from the outflow pipes of each mesocosm and fixed with formalin (2% final concentration [vol/vol]). Cells were stained with 4’, 6-diamidino-2- 64 phenylindole (DAPI), then part of the fixed water samples (0.8 mL) were filtered onto black 0.22-
µm polycarbonate isopore membrane filters (GTBP, MilliporeSigma) with a standard hand vacuum pump operation. An anti-bleaching agent was used as the mounting medium (AF1;
Citifluor). Cells were observed under 600x magnification using an Olympus BX51 epifluorescence microscope system. For each filter, more than 10 random fields were viewed to determine cell numbers.
2.7 Sample collection for microbial analysis
Biodegradable plastics were collected in clean 50 mL plastic centrifuge tubes from those embedded in the VFTW mesocosms and bioreactors on the FTW mesocosms. Root samples were collected using sterilized scissors and stored in 50 mL centrifuge tubes, consisting of a mixture of
0 - 15 cm depth segments from two distinct locations within each mesocosm. Soil samples were collected from two distinct locations in each vertical-flow mesocosm at a depth of 5 cm. The collected soil samples were vortexed for homogenization after initial collection. Water samples
(250 mL) collected from all FTW mesocosms were filtered using 0.2 µm cellulose nitrate membrane filters (47 mm diameter, Fischer Scientific Nalgene Analytical Test Filter) for further
DNA extraction. All samples were stored at -20°C for DNA extraction.
2.8 High throughput sequencing
DNA samples were extracted from biofilm on PCL beads, root, soil, and water filter using the MagAttract PowerSoil DNA KF kit (Qiagen) according to the manufacturer’s instructions.
Extracted DNA was eluted into 100 µL EB solution. Archaeal and bacterial 16S rRNA genes were amplified using the primer set, 515yF (5’GTGYCAGCMGCCGCGGTAA) and 926pfR
(5’CCGYCAATTYMTTTRAGTTT) (Parada et al., 2016) tagged with the Illumina i5 forward 65
(TCGTCGGCAGCGTCAGATGTGTATAAGAGACAG) and i7 reverse
(GTCTCGTGGGCTCGGAGATGTGTATAAGAGACAG) sequencing primer. Each PCR reaction contained 25 µL reactions with Qiagen HotStar Taq master mix, equal amount of forward and reverse primers (5 µM each), and 1 µL of DNA template (1 to 20 ng). Thermal cycling consisted of an initial denaturation at 95°C for 5 min, followed by 35 cycles of 94°C for 30 sec, annealing at 54°C for 40 sec, and extension at 72°C for 1 min, with a final extension of 10 min at
72°C. PCR product from the first stage was then transferred to a second PCR based on qualitatively determined concentrations with primers for the second PCR based on the Illumina Nextera PCR primers forward (AATGATACGGCGACCACCGAGATCTACAC-[i5 index]-
TCGTCGGCAGCGTC) and reverse (CAAGCAGAAGACGGCATACGAGAT-[i7 index]-
GTCTCGTGGGCTCGG). The second stage amplification was run the same as the first except for
10 cycles instead of 35 cycles. Amplicons were visualized with eGels (Life Technologies), products were pooled equimolar with each size selected quantified using the Quibit 2.0 fluorometer
(Life Technologies). Amplicons were then loaded on an Illumina MiSeq (Illumina) 2 x 300 flow cell at 10 pM (RTL Genomics).
For analysis, FASTQ formatted files were merged using the PEAR Illumina paired-end read merger (Zhang et al., 2013a). Prefix dereplication was completed using the algorithm of
USEARCH (Edgar et al., 2011). Clustering at a 3% divergence level was conducted using the
USEARCH (Edgar et al., 2011). Operational taxonomic unit (OTU) selection was performed using
UPARSE-OTU algorithm (Edgar, 2013). Chimera checking was completed using UCHIME
(Edgar, 2010) and detected chimera sequences were removed. Representative OTUs were used to determine taxonomic information through a basic local alignment search tool (BLAST) at National
Center for Biotechnology Information (NCBI), and MG-RAST (Meyer et al., 2008). The high- 66 throughput sequence datasets were deposited in GenBank under BioProject number
PRJNA496041.
2.9 Data analysis
The significant differences were determined when p < 0.05. Tukey-Kramer method was employed in conjunction with a one-way analysis of variance (ANOVA) using JMP data analysis software (SAS Institute) according to Lehman (2005) for testing statistical differences among multiple mesocosm settings. Student’s t-test was also implemented to determine if two sets of data were significantly different from each other. All statistics (one-way ANOVA, Tukey-Kramer, and
Student’s t-test) were completed using two-tailed and unpaired data analyses. Data were presented by mean ± standard deviation unless otherwise noted. General statistics of high-throughput sequence data were performed using MG-RAST (Meyer et al., 2008). Diversity index calculations
(Shannon index, Menhinick’s richness and Pielou’s evenness indices) were implemented using
Microsoft Excel (Menhinick, 1964; Chao and Shen, 2003; Jost, 2010).
3. RESULTS AND DISCUSSION
3.1 Physical parameters
Average rainfall over the experimental period was 7.6 ± 0.9 mm day-1 measured using the real-time hydrologic, water quality monitoring, and meteorological field station at the Everglades
Wetland Research Park (Zhang et al., 2017), which was less than the calculated HLR 48.4 m day-
1, with an average rain gauge depth of 1.2 ± 1.3 cm. Physical parameters were measured for FTW mesocosms with an average water temperature of 29.9°C, salinity of 5.1 ppt, pH of 7.95, and DO of 3.2 mg L-1. Turbidity (NTU, nephelometric turbidity unit) was high in the VFTW mesocosms 67
(23.5 ± 3.0 NTU and 47.0 ± 7.3 NTU with and without biodegradable plastic, respectively) and low in the FTW mesocosms (1.71 ± 0.1 NTU and 2.26 ± 0.2 NTU) due to the impact of soil.
3.2 Effect of biodegradable plastics for nutrient removal
Total nitrogen retention was significantly different (p ˂ 0.001, n = 8, one-way ANOVA) and higher in the vertical-flow system (Table 3.1). TN retention performance in our vertical-flow system (86.9-90%) was consistent with other saline aquaculture effluent treatment wetland systems with a mean removal efficiency of 98% TN (Brown et al., 1999), and 98.2% TDIN (Webb et al., 2012) (Table 3.2). On the contrary, TN retention was not observed in FTW mesocosms over the experimental period as outflow concentrations exceeded the inflow concentration of effluent
(Table 3.1), leading to a negative retention rate, consistent with a previous aquaculture wastewater treatment study showing outflow concentration of TKN exceeded feed water TKN (Lin et al.,
2010). However, the presence of PCL significantly increased TN retention in the FTW mesocosms
(p ˂ 0.0001, Tukey-Kramer pairwise). This finding was not identified in the VFTW system (Table
3.1). A comparison study in China found that floating treatment systems had a lower TN removal efficiency compared with vertical-flow systems (Zhang et al., 2015). Newly constructed or newly restored wetlands are found to have a low C:N ratio, therefore the addition of an external carbon can enhance denitrification in these wetland systems (Bachand and Horne, 1999), as evidenced in the FTW system in our study. Overall, these lines of evidence indicate the strict carbon limitation in the FTW than the other treatment wetland systems (Zhang et al., 2015).
Inflow TN was composed of over 98% inorganic nitrogen, mainly in the form of nitrate and nitrite. Aquaculture effluent contains low organic nitrogen and phosphorus in the water column with 7-32% of nitrogen found in suspended solids (Turcios and Papenbrock, 2014). In the VFTW system nitrate plus nitrite ratio decreased to be 17.6% in the control and 16.2% with 68 embedded PCL, along with a decrease of the TN:TP ratio from 8.9 to 4.7 and 4.5 respectively, indicating microbial nitrogen removal processes (i.e. denitrification or DNRA [dissimilatory nitrate reduction to ammonium]) (Fig. 3.3a). Our ammonia measurements showed the depletion of ammonia from 9.2% in inflow to 7.1% in the control and 5.7% with embedded PCL, with an increase in organic nitrogen to 75.3% and 78.2%, respectively. Based on these findings we concluded that denitrification, not DNRA, was the major process in the removal of nitrate plus nitrite pool in the VFTW. In the FTW system nitrate plus nitrite ratio decreased from 94.2% to be 40.5% in the FTW control and 10.7% in FTW with PCL condition, with an increased removal efficiency with presence of PCL (Fig. 3.3a). Additionally, we found a decrease of ammonia in outflow water from 5.8% to 3.2% in control and 4.3% with PCL conditions, with an organic nitrogen increase to 56.3% in control and 85.0% in FTWs with embedded PCL, a greater proportion of organic nitrogen with presence of PCL indicated PCL stimulated the denitrification activity. As previously discussed, TN retention was not observed in FTW mesocosms due to the production of organic nitrogen from the system not present in inflow waters and possible nitrification activity due to the presence of oxygen in the water column, however, the inorganic nitrogen concentrations decreased. In contrast with the VFTW system, the TN:TP ratio in outflow water increased from 13.2 to 27 in control and 17 with PCL medium (Fig. 3.3b).
TP retention was significantly higher in the VFTW system than for the FTW system (p ˂
0.0032, t-test), however, no significance was found with the presence of PCL in both mesocosm systems (Table 3.1). The TP retention in the VFTW mesocosms had a mean retention of 74 -
81.1%, which was lower than similar studies, 99% (Brown et al., 1999) and 88% (Lymbery et al.,
2006) (Table 3.2). Zhang et al. (2015) found TP removal efficiency ranged from 26-70% and was more variable than nitrogen in constructed wetland systems. The mean TP retention in the FTW 69 system was 17.4-39.5%, consistent with findings by Lin et al. (2010) with 2-18% removal and
Pavlineri et al. (2017) of 18.2% removal efficiency, with an increase over time (Table 3.2).
3.3 Growth of Spartina in vertical-flow and floating treatment systems
Change in plant height ranged from 1.7 to 17.3 mm day-1 with the highest growth rate occurring in the FTW mesocosms with PCL medium (Fig. 3.4). Even though there seemed to be an increased growth rate associated with use of PCL medium in the FTW system, the range of measurements overlapped when looking at mean growth rate. Due to a low number of replicates in this study (n = 2), no statistical comparison was made.
3.4 Bacterial abundance
Total bacterial abundance of wetland water columns generally range from 105 to 106 cells mL-1 (Urakawa and Bernhard, 2017), which was similar with our findings. No significant differences were found in the outflow bacterial abundance with the following distribution; VFTW control (3.7 x 106 ± 1.8 x 106 cells mL-1), VFTW with embedded PCL (3.5 x 106 ± 1.7 x 106 cells mL-1), FTW control (1.0 x 106 ± 6.3 x 105 cells mL-1), and FTW with PCL medium (3.2 x 106 ±
2.2 x 106 cells mL-1).
3.5 High-throughput sequencing of 16S rRNA gene
3.5.1 Taxonomic overview of dominant phyla
A total of 86,547 sequences were analyzed and resulted in 2346 operational taxonomic units (OTUs) (Table 3.3). Shannon index indicated significant differences between sample means
(p = 0.02, one-way ANOVA) with the lowest diversity in water samples and the highest diversity in root samples. There were significant differences found between VFTW root and FTW water samples (p = 0.03, Tukey-pairwise). The highest diversity found in soil samples was consistent 70 with previous reports (Urakawa and Bernhard, 2017). The taxonomic analysis identified 29 phyla from all samples: 12-21 collected from PCL plastic biofilm, 15-24 in root samples, 17-20 in soil samples, and 8-11 in water samples. No statistical difference was found in PCL biofilm samples
(p = 0.06, t-test) and soil samples (p = 0.5, t-test). Root samples were significantly different (p =
0.03, n = 2, one-way ANOVA) between VFTW control and FTW with PCL (p = 0.03, Tukey- pairwise). Water samples in floating treatment system conditions having PCL bioreactor medium were significantly lower than the control samples (p = 0.0002, t-test).
The three predominant phylum present in all samples were Proteobacteria (2-44%),
Cyanobacteria (0.04-51%) and Bacteroidetes (0.02-30%) (Fig. 5). These results were consistent with previous studies of wetland microbial communities (Bai et al., 2014; Liu et al., 2016; Urakawa and Bernhard, 2017). Members of Proteobacteria are important in wetlands because of their strong involvement in biogeochemical cycling (Liu et al., 2016) and they dominated in a majority of samples except for water column samples from FTW. The two most abundant phyla in the water column samples were Firmicutes (59-90%) and Actinobacteria (6-17%) (Fig. 3.5). Unexpectedly, the most dominant member of Firmicutes was identified as Bacillus (57-88%), this trend agreed between four samples assuring good reproducibility of the method used. We attributed this finding to the presence of soil within plant pots (Fig. 3.2). Bacillus is recognized as a representative degrader of biodegradable plastics. For example, Bacillus pumilus, isolated from a freshwater pond and river were shown to degrade poly (e-caprolactone) hydrolytically (Tezuka et al., 2004).
However, presence of Bacillus was found regardless of PCL medium indicating that Bacillus was not directly enriched by the biodegradable plastics (Table 3.4).
Soil microbial communities in VFTW were dominated by Proteobacteria (40-85%),
Cyanobacteria (5-40%), Bacteroidetes (2-10%), and Planctomycetes (2-7%) (Fig. 3.5). The soil 71 was covered with approximately 5-10 cm of water layer (Fig. 3.2). Cyanobacteria is a typical phylum found in freshwater sediment and water column communities (Paerl, 2014; Urakawa and
Bernhard, 2017; Paerl, 2018). Thus, the observed microbial community might resemble a typical freshwater sediment community rather than a typical soil community (Zhang et al., 2013b). A steep oxic-anoxic gradient contributes to maintain high microbial diversity and functionally diverse organisms (Urakawa et al., 2017). Our results supported this finding by having the highest diversity found in soil samples of our wetland.
3.5.2 Comparison of rhizosphere communities in soil and water
In wetland plants, the rhizosphere acts as an interface between the surface of roots and the surrounding soil, which transports oxygen and other minerals to the roots which results in unique microbial communities distinct from surrounding soil and water column in a case of floating macrophytes (Mitsch and Gosselink, 2015; Urakawa et al., 2017).
The nitrogen cycle plays an important role in wetland plant metabolisms through the transformation of nitrogen species (i.e. ammonia and nitrate). Mesorhizobium and Rhizobium are essential diazotrophs and plant growth-promoting rhizosphere bacteria found in wetland systems
(Zhang et al., 2013b; Urakawa et al., 2017). Mesorhizobium was identified in root, soil, water, and
PCL biofilm samples while Rhizobium was identified only in vertical-flow root samples (Table
3.4). Nitrogen-fixing bacteria were more abundant in the VFTW than FTW mesocosms. The only nitrifying bacterium identified was Nitrospira, in root, soil, and VFTW PCL biofilm samples.
Methanogenesis is an important process in wetlands through which methane is naturally produced by methanogens and methane oxidation occurs from methanotrophic bacteria to convert methane to carbon dioxide (Mitsch and Gosselink, 2015). Archaea are important methanogens in 72 wetland sediments contributing to methane production (Madigan et al., 2012; Urakawa and
Bernhard, 2017), three genera of methanogenic archaea found were Methanobacterium,
Methanoregula, and Methanosarcina. Six methanotrophic genera were also found, Methylocystis,
Methylobacter, Methylococcus, Methylosoma, Methylocella and Hyphomicrobium. These methanogens and methanotrophs were more abundant in the VFTW than in the FTW mesocosms
(Table 3.4). Hyphomicrobium belonging to Alphaproteobacteria and Methylibium belonging to
Betaproteobacteria were the two most abundant facultative methylotrophic genera and widely distributed in our constructed wetland systems, which supported a previous wetland study (Zhang et al., 2013b). Methylocella was the most widespread methanotroph found in this study.
Coexistence of methanogens and methane oxidizers suggests the existence of the methane cycle and the skewed relative abundance of these microorganisms indicated more imperative role of this process in the vertical-flow system than in the floating wetland system.
Sulfate-reducing bacteria (SRB) were the predominant sulfur cycling microorganisms found in root samples and PCL biofilm. Although SRB were found in both systems, the vertical- flow system contained a greater diversity of organisms (i.e. Desulfobulbus, Desulfatitalea,
Desulfobacterium, Desulfonema, Desulfocapsa, Desulfopila, Desulfomicrobium, and
Desulfovibrio) than were found in the floating treatment system (i.e. Desulfovibrio and
Desulfobulbus) (Table 3.4). The floating treatment system contained very minor amount of SRB in contrast to Urakawa et al (2017) which found a very rich SRB community in floating treatment rhizosphere. SRB communities in rhizosphere and soil in a Phragmites australis planted wetland
(Zhang et al., 2013b) and wetland soils (Faulwetter et al., 2009; Wang et al., 2012) were very diverse and consistent with our findings.
3.5.3 Denitrification in vertical-flow and floating treatment systems 73
Denitrification is the main nitrogen removal process in treatment wetland systems as discussed previously and paired with nitrification, a process in which nitrate is produced from ammonium, can fully remove nitrogen microbially from wastewater systems (Faulwetter et al.,
2009). Predominant denitrifiers found in our study were Bacillus in water column samples,
Nitratireductor, a marine denitrifier (Labbè et al., 2004) represented in all samples in minor amount, and Pseudomonas (0.2%) in soil samples with embedded biodegradable plastics (Table
3.4). Pseudomonas has been found to degrade plastic particles in an urban river environment
(McCormick et al., 2014), soil environments (Emadian et al., 2017), and the deep-sea (Sekiguchi et al., 2011). The presence of Pseudomonas only in soil samples with PCL may indicate the possibility of PCL use as a substratum or degradability, as indicated by similar findings of
Pseudomonas on plastic pot biofilm from a floating treatment wetland (Urakawa et al., 2017).
These findings support our observation of increased denitrification activity in the VFTW and FTW construction with the presence of PCL.
3.5.4 PCL degradation in a vertical-flow and floating treatment constructed wetland
The most abundant genera found in PCL biofilm samples collected from VFTW sediment were identified as Oscillatoria (7%) and Leptolyngbya (6%) and from FTW bioreactors were
Brasilonema (8%) and Trichormus (9%) belonging to the phylum Cyanobacteria (Table 3.4).
Additionally, Leptolyngbya was identified in VFTW sediment with embedded PCL. The localization of Cyanobacteria in VFTW plastics was attributed to a partial exposure of plastics to the surface (Fig. 3.2). We identified many Cyanobacteria within our study in presence of PCL plastic, consistent with previous marine plastic debris research (Bryant et al., 2016; Debroas et al.,
2017; Quero and Luna, 2017). Cyanobacteria were identified as the key species in the microbial network which is formed on the surface of plastics (Debroas et al., 2017). However, none of these 74 studies confirmed if Cyanobacteria are actively involved in the biodegradation of the plastics
(Debroas et al., 2017; Quero and Luna, 2017). Bryant et al. (2016) and Debroas et al. (2017) identified Leptolyngbya on the surface of plastics collected from the surface water of the North
Atlantic. It should be noted that Cyanobacteria are able to synthesize polyhydroxybutryate, an intracellular storage compound and bioplastic, under photoautotrophic or chemoheterotrophic conditions (Balaji et al., 2013; Singh et al., 2017). Additionally, several genera can synthesize polyhydroxyalkanoate (PHA) and contain PHA biosynthesis genes (e.g. Oscillatoria limosa,
Anabaena cylindrica, Synechoccocus spp.), these findings can lead to the speculation they are also able to degrade these bioplastic storage compounds for intracellular use.
4. CONCLUSIONS
Wetlands play a vital role in water purification and nutrient cycling which can be utilized to treat agricultural runoff and aquaculture discharges in a sustainable fashion. The nutrient retention rate for TN and TP was greater in the vertical-flow than in the floating wetland systems.
The use of a biodegradable plastic increased nutrient retention in the FTW mesocosms. The presence of PCL altered microbial community composition through microbial selection such as abundant cyanobacterial genera and other bioplastic-degrading microorganisms. Further long-term studies are needed to have a greater understanding of microbial plastic degradation and associated nutrient cycling in constructed wetland systems.
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Table 3.1 Nutrient flux of two constructed wetland systems.
TN (mg m-2 day-1) TP (mg m-2 day-1) Removal Removal Wetland PCL Inflow Outflow Reduction (%) Inflow Outflow Reduction (%) efficiency efficiency VFTW (-) 8.63 ± 0.26 77.41 ± 0.45a 90.0 1.82 ± 0.09 7.82 ± 0.11a 81.1 86.04 ± 0.19 9.64 ± 0.02 (+) 11.29 ± 0.53 74.75 ± 0.72b 86.9 2.51 ± 0.19 7.13 ± 0.63a 74.0
FTW (-) 27.48 ± 0.59 (-5.09) ± 0.72c 0 1.01 ± 0.06 0.66 ± 0.1b 39.5 22.39 ± 0.13 1.67 ± 0.04 (+) 23.95 ± 0.65 (-1.56) ± 0.78d 0 1.38 ± 0.05 0.29 ± 0.09b 17.4 *Data are mean ± standard error (n = 8) for all experimental conditions with percent retention quantified from ((inflow concentration – outflow concentration) (inflow concentration)) x 100 (Olguín et al., 2017). The letter next to the monthly retention denotes statistical significance from completing, one-way ANOVA for TN, and significance shown for TP retention, t-test. 76
Table 3.2 Removal efficiency of various constructed wetlands for aquaculture effluent treatment.
Removal efficiency Temperature Study Scale Construction type Plant species used Salinity (ppt) Location (%)* (°C) Suaeda eseroa, Salicornia TN:98, TIN:94, Tucson, AZ Brown et al., 1999 Mesocosm Subsurface flow bigelovii, 10 22.6 - 37.4 TP:99 USA Altriplex barclayana Lymbery et al., Horizontal Mesocosm Juncus kraussii TN:69, TP:88 6.6 - 24.8 N/A Australia 2006 subsurface flow Canna indica, Typha latifolia, Li et al., 2007 Pilot-scale Vertical - flow TN:54.6 TP:80.1 0 23.6 - 24.0 China Acorus calamus, Arave sisalana Vertical - Canna indica, N/A Zhang et al., 2010 Pilot-scale downflow vertical Typha latifolia, TN:48, TP:17 N/A China Freshwater - upflow hybrid Acorus calamus
Salicornia TDIN:98.2, North Wales, Webb et al., 2012 Pilot-scale Subsurface flow 22 23.1 europaea DIP:36 - 89 UK N/A Free water surface Phragmites Litopenaeus Lin et al., 2003 Pilot-scale – subsurface flow TIN:68.2, PO4-P:5.4 23.5 Taiwan australis vannamei hybrid culture Eichhornia crassipes, Pistia stratiotes, Typha Floating angustifolia, macrophyte – Lin et al., 2010 Pilot-scale Phragmites TN:0 - 18, TP:2 - 18 0 N/A Taiwan subsurface flow communis, Canna hybrid generalis, Cyperus alternifalius Chrysopogon zizaniodes, Typha De Stefani et al., Floating treatment N/A in-stream latifolia, TN:13 - 29, TP:65 10.0 - 14.0 Italy 2011 wetland Freshwater Sparganium erectum Floating treatment Ipomenea N/A Li and Li, 2009 Pilot-scale TN:30.6, TP:18.2 24.4 China wetland aquatica Freshwater TN: 86.9 - 90.7 This study Mesocosm Vertical - flow Spartina patens 7.4 28.8 FL, USA TP: 74 - 81.1 Floating treatment TN: 0 Mesocosm Spartina patens 5.1 29.9 FL, USA wetland TP: 17.4 - 39.5 *Removal efficiencies are denoted TN (total nitrogen), TIN (total inorganic nitrogen), TDIN (total dissolved inorganic nitrogen), TP (total phosphorous), DIP (dissolved inorganic phosphorous), and phosphate. Temperature is denoted as air temperature for subsurface flow and water temperature for floating treatment systems. N/A shows data not measured.
77
Table 3.3 Summary of DNA sequencing and diversity indices.
Mean Menhinick PCL Pielou Shannon Samples Sequences OTU sequence richness presence evenness index length index
PCL Biofilm
VFTW1 21,519 776 411 ± 19 0.42 2.53 2.32
VFTW2 13,069 504 412 ± 6 0.41 2.04 2.18
FTW1 13,118 211 411 ± 2 0.62 1.09 2.91
FTW2 13,496 278 412 ± 8 0.53 1.35 2.62
Root VFTW1 (-) 22,640 622 412 ± 8 0.39 2.5 2.14
VFTW2 (-) 23,988 871 411 ± 13 0.48 2.98 2.73
VFTW1 (+) 23,930 764 412 ± 6 0.48 2.53 2.14
VFTW2 (+) 11,796 525 411 ± 15 0.48 1.93 2.43
FTW1 (-) 27,026 472 411 ± 6 0.47 1.86 2.46
FTW2 (-) 19,676 316 411 ± 6 0.51 1.33 2.48
FTW1 (+) 15,924 194 411 ± 4 0.48 0.89 2.14
FTW2 (+) 17,078 355 411 ± 3 0.48 1.56 2.43
Soil VFTW1 (-) 18,851 646 412 ± 8 0.6 2.26 3.23
VFTW2 (-) 17,544 529 412 ± 8 0.6 2.15 3.24
VFTW1 (+) 20,381 208 411 ± 14 0.29 1.04 1.37
VFTW2 (+) 22,693 827 412 ± 8 0.55 2.57 3.07
Water FTW1 (-) 20,580 143 412 ± 5 0.29 0.75 1.24
FTW2 (-) 21,014 135 412 ± 4 0.22 0.76 0.95
FTW1 (+) 25,287 113 412 ± 4 0.17 0.64 0.69
FTW2 (+) 26,658 128 412 ± 3 0.35 0.71 1.5 Diversity indices were calculated after normalization to 10,000 reads per sample. Samples 1 and 2 denote replicates and (-) and (+) denotes presence of PCL in construction when applicable.
78
Table 3.4 Functional groups at the genus level.
PCL biofilm Roots Soil Water VFTW VFTW FTW FTW VFTW VFTW FTW FTW Nitrogen - fixing bacteria VFTW FTW (+) (-) (+) (-) (+) (-) (+) (-) Alphaproteobacteria Nitrospirillium 77 1 15 9 0 0 20 2 0 0 Rhizobium 1 0 33 4 0 0 4 3 0 0 Bradyrhizobium 0 0 11 5 4 0 0 0 1 0 Mesorhizobium 83 15 7 9 20 6 12 13 2 4 Azospirilium 9 0 5 0 0 0 8 0 0 0 Betaproteobacteria Azohydromonas 18 2 8 1 4 11 15 51 0 0 Azonexus 2 0 13 4 0 0 1 1 0 0 Derxia 39 0 10 14 0 0 7 115 0 0 Cyanobacteria Anabaena 12 0 0 0 0 0 6 0 0 0 Nostoc 79 115 43 19 2 1 1063 92 0 2 Calothrix 84 6 9 30 0 0 51 30 0 0 Cylindrospermum 36 0 79 36 0 0 526 56 0 0 Nitrifying bacteria Nitrospira Nitrospira 1 0 1 0 265 88 0 8 0 0 Gammaproteobacteria Nitrosococcus 5 1 23 59 114 63 14 21 0 0 Denitrifying bacteria Bacilli Bacillus 3 0 7 6 0 0 26 24 7589 7245 Alphaproteobacteria Nitratereductor 3 1 24 16 4 1 51 7 0 0 Gammaproteobacteria Pseudomonas 27 2 4 0 1 0 0 19 0 0 Sulfate-reducing bacteria Deltaproteobacteria Desulfobulbus 25 1 29 140 0 2 1 22 0 0 Desulfatitalea 302 0 0 76 0 0 0 0 0 0 Desulfobacterium 0 0 0 19 0 0 0 0 0 0 Desulfonema 28 0 5 11 0 0 0 0 0 0 Desulfocapsa 24 0 59 69 0 0 4 3 0 0 Desulfopila 1 0 11 9 0 0 2 0 0 0 Desulfomicrobium 11 0 5 12 0 0 15 6 0 0 Desulfovibrio 40 24 137 138 6 7 44 12 0 0 Sulfur-oxidizing bacteria Chlorobia Chlorobium 9 0 13 8 0 0 4 2 0 0 Betaproteobacteria Thiobacillus 74 4 15 11 73 99 17 40 1 1 Thiobacter 2 0 5 1 0 0 0 1 0 0 Gammaproteobacteria Thiothrix 0 2 0 0 1 32 0 0 0 0 Methanogenic archaea Methanobacteria Methanobacterium 19 0 43 13 0 0 51 11 0 0 Methanomicrobia Methanoregula 0 0 3 0 0 0 0 0 0 0 Methanosarcina 1 0 2 0 0 0 2 0 0 0 79
Methanotrophic bacteria Alphaproteobacteria Methylocystis 0 0 7 0 0 0 0 0 0 0 Methylobacter 0 0 0 0 0 2 0 0 0 0
Methylocella 0 3 0 1 26 41 0 0 0 2 Gammaproteobacteria Methylococcus 0 0 4 0 0 0 0 0 0 0 Methylosoma 0 0 0 0 9 3 0 0 0 0 Methylotrophic bacteria Methylobacterium 0 0 1 1 5 8 0 0 0 0 Alphaproteobacteria Methylobacillus 0 0 5 5 0 0 0 2 0 0
Hyphomicrobium 4 3 16 6 89 80 14 8 8 9
Betaproteobacteria Methylophilus 0 0 0 0 15 4 0 0 0 0 Methylibium 200 5 13 53 7 4 171 550 0 0 Shown are average bacterial and archaeal relative abundance of sample distribution (n = 2) after normalization to 10,000 reads per sample. Cyanobacteria are shown at the phylum level as class level was unidentified.
80
Figure 3.1 Design of two wetland systems. a) Overview of experimental setup b) schematic of vertical- flow treatment wetland (VFTW) mesocosm and c) schematic for floating treatment wetland (FTW) mesocosm. b) and c) schematic denoting length (1.33 m) x width (0.47 m) x height (0.61 m) with different design depth and width of mesocosms with PCL, polycaprolactone, identical design was used for controls without PCL aspect.
81
Figure 3.2 Mesocosms used in this study. a) Overview of VFTW mesocosms in setup, b) Spartina patens location within mesocosm relative to edge of tub, c) VFTW mesocosm containing PCL plastic beads which had the ability to float when flooded, d) overview of FTW mesocosm with bioreactor setup (rear row), e) FTW mesocosm containing PCL plastic as reactor medium, f) configuration of floating mat with 18 planting holes, g) view of aerator pot and plant root, and h) overview setup of upstream aquaculture tank with double filtration system that housed Pinfish (Lagodon rhomboides).
82
Figure 3.3 Change in inflow and outflow nutrient concentrations. a) Percent composition of nitrogen in inflow and outflow and b) TN:TP ratio change over time, with (-) denoting control and (+) presence of PCL in construction. A solid horizontal line indicates mg-based Redfield ratio between N and P (8.9).
83
Figure 3.4 Change in stem height of Spartina patens. Measured in mm day-1 over the month testing period with (-) denoting control and (+) presence of PCL in construction. Data are shown as mean ± range (n = 2).
84
Figure 3.5 Relative bacterial and archaeal abundance at the phylum level. Percent relative abundance distribution after normalization to 10,000 reads per sample. Proteobacteria are shown at the class level. Sample naming uses 1 and 2 showing replication and (-) and (+) denoting presence of PCL.
85
References
Ahn, C., Mitsch, W. J., Wolfe, W. E., 2001. Effects of recycled FGD liner material on water quality
and macrophytes of constructed wetlands: a mesocosm experiment. Water Resources. 35,
633-642.
Ahn, C., Mitsch, W.J., 2002. Scaling considerations of mesocosm wetlands in simulating large
created freshwater marshes. Ecological Engineering. 18, 327-342.
Bachand, P. A. M., Horne, A. J., 1999. Denitrification in constructed free-water surface wetlands:
II. Effects of vegetation and temperature. Ecological Engineering. 14, 17-32.
Bai, Y., Liang, J., Liu, R., Hu, C., Qu, J., 2014. Metagenomic analysis reveals microbial diversity
and function in the rhizosphere soil of a constructed wetland. Environmental Technology.
35, 2521-2527.
Balaji, S., Gopi, K., Muthuvelan, B., 2013. A review on production of poly β hydroxybutyrates
from cyanobacteria for the production of bio plastics. Algal Research. 2, 278-285.
Brown, J. J., Glenn, E.P., Fitzsimmons, K.M., Smith, S.E., 1999. Halophytes for the treatment of
saline aquaculture effluent. Aquaculture. 175, 225-268.
Bryant, J. A., Clemente, T. M., Viviani, D. A., Fong, A. A., Thomas, K. A., Kemp, P., Karl, D.
M., White, A. E., DeLong, E. F., 2016. Diversity and activity of communities inhabiting
plastic debris in the North Pacific gyre. mSystems. 1, e00024-16.
doi:10.1128/mSystems.00024-16
Chao, A., Shen, T.-J., 2003. Nonparametric estimation of Shannon’s index of diversity when there
are unseen species in sample. Environmental and Ecological Statistics. 10, 429-443. 86
De Lange, H. J., Paulissen, M.P.C.P., Slim, P.A., 2013. 'Halophyte filters': The potential of
constructed wetlands for application in saline aquaculture. International Journal of
Phytoremediation. 15, 352-364.
Debroas, D., Mone, A., Ter Halle, A., 2017. Plastics in the North Atlantic garbage patch: A boat-
microbe for hitchhikers and plastic degraders. Science of the Total Environment. 599-600,
1222-1232.
De Stefani, G., Tocchetto, D., Salvato, M., Borin, M., 2011. Performance of a floating treatment
wetland for in-stream water amelioration in NE Italy. Hydrobiologia. 674, 157-167.
Edgar, R. C., 2010. Search and clustering orders of magnitude faster than BLAST. Bioinformatics.
26, 2460-2461.
Edgar, R. C., 2013. UPARSE: highly accurate OTU sequences from microbial amplicon reads.
Nature Methods. 10, 996-998.
Edgar, R. C., Haas, B. J., Clemente, J. C., Quince, C., Knight, R., 2011. UCHIME improves
sensitivity and speed of chimera detection. Bioinformatics. 27, 2194-2200.
Emadian, S. M., Onay, T. T., Demirel, B., 2017. Biodegradation of bioplastics in natural
environments. Waste Management. 59, 526-536.
Faulwetter, J., Burr, M. D., Cunningham, A. B., Stewart, F., Camper, A. K., Stein, O. R., 2011.
Floating treatment wetlands for domestic wastewater treatment. Water Science and
Technology. 64, 2089-2095. 87
Faulwetter, J. L., Gagnon, V., Sundberg, C., Chazarenc, F., Burr, M. D., Brisson, J., Camper, A.
K., Stein, O. R., 2009. Microbial processes influencing performance of treatment wetlands:
A review. Ecological Engineering. 35, 987-1004.
Fink, D.F. and Mitsch, W. J., 2004. Seasonal and storm event nutrient removal by a created wetland
in an agricultural watershed. Ecological Engineering. 23, 313-325.
Fu, G., Huangshen, L., Guo, Z., Zhou, Q., Wu, Z., 2017. Effect of plant-based carbon sources on
denitrifying microorganisms in a vertical flow constructed wetland. Bioresource
Technology. 224, 214-221.
Fuchs, V. J., Mihelcic, J. R., Gierke, J. S., 2011. Life cycle assessment of vertical and horizontal
flow constructed wetlands for wastewater treatment considering nitrogen and carbon
greenhouse gas emissions. Water Resources. 45, 2073-2081.
Griffiths, L.N., Mitsch, W.J., 2017. Removal of nutrients from urban stormwater runoff by storm-
pulsed and seasonally pulsed created wetlands in the subtropics. Ecological Engineering.
108, 414-424.
Hubbard, R., Gascho, G., Newton, G., 2004. Use of floating vegetation to remove nutrients from
swine lagoon wastewater. Transactions of the ASAE. 47, 1963-1972.
Jost, L., 2010. The relation between evenness and diversity. Diversity. 2, 207-232.
Konnerup, D., Trang, N. T. D., Brix, H., 2011. Treatment of fishpond water by recirculating
horizontal and vertical flow constructed wetlands in the tropics. Aquaculture. 313, 57-64.
Labbé, N., Parent, S., Villemur, R., 2004. Nitratireductor aquibiodomus gen. nov., sp. nov., a
novel α-proteobacterium from the marine denitrification system of the Montreal Biodome 88
(Canada). International Journal of Systematic and Evolutionary Microbiology. 54, 269-
273.
Lehman, A., 2005. JMP for basic univariate and multivariate statistics: a step-by-step guide: SAS
Institute.
Li, W., and Li, Z., 2009. In situ nutrient removal from aquaculture wastewater by aquatic vegetable
ipomoea aquatica on floating beds. Water Science and Technology. 59, 1937-1943.
Li, G., Wu, Z., Cheng, S., Liang, W., He, F., Fu, G., Zhong, F., 2007. Application of constructed
wetlands on wastewater treatment for aquaculture ponds. Wuhan University Journal of
Natural Sciences. 12, 1131-1135.
Liang, Y., Zhu, H., Bañuelos, G., Yan, B., Zhou, Q., Yu, X., Cheng, X., 2017. Constructed
wetlands for saline wastewater treatment: A review. Ecological Engineering. 98, 275-285.
Lin, Y. F., Jing, S. R., Lee, D. Y., Chang, Y. F., Sui, H. Y., 2010. Constructed wetlands for water
pollution management of aquaculture farms conducting earthen pond culture. Water
Environment Research. 82, 759-768.
Lin, Y. F., Jing, S. R., Wang, T. W., Lee, D. Y., 2002. Effects of macrophytes and external carbon
sources on nitrate removal from groundwater in constructed wetlands. Environmental
Pollution. 119, 413-420.
Lin, Y. F., Jing, S. R., Lee, D. Y., 2003. The potential use of constructed wetlands in a recirculating
aquaculture system for shrimp culture. Environmental Pollution. 123, 107-113. 89
Liu, J., Yi, N.-K., Wang, S., Lu, L.-J., Huang, X.-F., 2016. Impact of plant species on spatial
distribution of metabolic potential and functional diversity of microbial communities in a
constructed wetland treating aquaculture wastewater. Ecological Engineering. 94, 564-573.
Lu, S., Hu, H., Sun, Y., Yang, J., 2009. Effect of carbon source on the denitrification in constructed
wetlands. Journal of Environmental Sciences. 21, 1036-1043.
Lymbery, A. J., Doupe´, R.G., Bennett, T., Starcevich, M., R., 2006. Efficacy of a subsurface-flow
wetland using the estuarine sedge Juncus krausii to treat effluent from inland saline
aquaculture. Aquacultural Engineering. 34, 1-7.
Madigan, M. T., Martinko, J. M., Stahl, D. A., Clark. D. P., 2012. Brock Biology of
Microorganisms: Pearson, California.
McCormick, A., Hoellein, T. J., Mason, S. A., Schluep, J., Kelley, J. J., 2014. Microplastic is an
abundant and distinct microbial habitat in an urban river. Environmental Science &
Technology. 48, 11863-11871.
Menhinick, E.F., 1964. A comparison of some species-individuals diversity indices applied to
samples of field insects. Ecological Society of America. 45, 859-861.
Meyer, F., Paarmann, D., D’Souza, M., Olson, R., Glass, E.M., Kubal, M., Paczian, T., Rodriguez,
A., Stevens, R., Wilke, A., Wilkening, J., Edwards, R.A., 2008. The metagenomics RAST
server - a public resource for the automatic phylogenetic and functional analysis of
metagenomes. BMC Bioinformatics. 9, 386.
Mitsch, W.J, Zhang, L., Stefanik, K. C., Nahlik, A. M., Anderson, C. J., Bernal, B., Hernandez,
M., Song, K., 2012. Creating wetlands: Primary succession, water quality changes, and
self-design over 15 years. BioScience. 62, 237-250. 90
Mitsch, W.J Zhang, L, Marois, D., Song, K., 2015. Protecting the Florida Everglades wetlands
with wetlands: Can stormwater phosphorus be reduced to oligotrophic conditions?
Ecological Engineering. 80, 8-19.
Mitsch, W. J., Gosselink, J. G. 2015. Wetlands, 5th Edition: John Wiley & Sons, Inc., Hoboken,
NJ.
Nahlik, A. M., Mitsch, W. J., 2006. Tropical treatment wetlands dominated by free-floating
macrophytes for water quality improvement in Costa Rica. Ecological Engineering. 28,
246-257.
Olguín, E. J., Sánchez-Galván, G., Melo, F. J., Hernández, V. J., González-Portela, R. E., 2017.
Long-term assessment at field scale of floating treatment wetlands for improvement of
water quality and provision of ecosystem services in a eutrophic urban pond. Science of
the Total Environment. 584, 561-571.
Paerl, H. W., 2014. Mitigating harmful cyanobacterial blooms in a human-and climatically-
impacted world. Life. 4, 988-1012.
Paerl, H. W., 2018. Mitigating toxic planktonic cyanobacterial blooms in aquatic ecosystems
facing increasing anthropogenic and climatic pressures. Toxins. 10, 76.
doi:10.3390/toxins10020076
Parada, A. E., Needham, D. M., Fuhrman, J. A., 2016. Every base matters: assessing small subunit
rRNA primers for marine microbiomes with mock communities, time series and global
field samples. Environmental Microbiology. 18, 1403-1414. 91
Pavlineri, N., Skoulikidis, N. T., Tsihrintzis, V. A., 2017. Constructed floating wetlands: a review
of research, design, operation and management aspects, and data meta-analysis. Chemical
Engineering Journal. 308, 1120-1132.
Quero, G. M., Luna, G. M., 2017. Surfing and dining on the “plastisphere”: Microbial life on
plastic marine debris. Advances in Oceanography and Limnology. 8, 199-207.
Shen, Z., Zhou, Y., Liu, J., Xiao, Y., Cao, R., Wu, F., 2015. Enhanced removal of nitrate using
starch/PCL blends as solid carbon source in a constructed wetland. Bioresource
Technology. 175, 239-244.
Singh, A. K., Sharma, L., Mallick, N., Mala, J., 2017. Progress and challenges in producing
polyhydroxyalkanoate biopolymers from cyanobacteria. Journal of Applied Phycology. 29,
1213-1232.
Sirivedhin, T., Gray, K. A., 2006. Factors affecting denitrification rates in experimental wetlands:
field and laboratory studies. Ecological Engineering. 26, 167-181.
Sekiguchi, T., Sato, T., Enoki, M., Kanehiro, H., Uematsu, K., Kato, C., 2011. Isolation and
characterization of biodegradable plastic degrading bacteria from deep-sea environments.
JAMSTEC Report of Research and Development. 11, 33-41.
Stottmeister, U., Wiessner, A., Kuschk, P., Kappelmeyer, U., Kastner, M., Bederski, O., Muller,
R. A., Moormann, H., 2003. Effects of plants and microorganisms in constructed wetlands
for wastewater treatment. Biotechnology Advances. 22, 93-117.
Tezuka, Y., Ishii, N., Kasuya, K.-i., Mitomo, H., 2004. Degradation of poly (ethylene succinate)
by mesophilic bacteria. Polymer Degradation and Stability. 84, 115-121. 92
Turcios, A. E., Papenbrock, J., 2014. Sustainable treatment of aquaculture effluents—what can we
learn from the past for the future? Sustainability. 6, 836-856.
USEPA., 1993a. Method 351.2: Determination of total kjeldahl nitrogen by semi-automated
colorimetry (Revision 2.0), Cincinnati, Ohio.
USEPA., 1993b. Method 353.2: Determination of nitrate-nitrite nitrogen by automated colorimetry
(Revision 2.0), Cincinnati, Ohio.
USEPA., 1993c. Method 365.1: Determination of phosphorous by semi-automated colorimetry
(Revision 2.0), Cincinnati, Ohio.
Urakawa, H., Bernhard, A. E., 2017. Wetland management using microbial indicators. Ecological
Engineering. 108, 456-476.
Urakawa, H., Dettmar, D. L., Thomas, S., 2017. The uniqueness and biogeochemical cycling of
plant root microbial communities in a floating treatment wetland. Ecological Engineering.
108, 573-580.
Vymazal, J., 2007. Removal of nutrients in various types of constructed wetlands. Science of the
Total Environment. 380, 48-65.
Wang, Y., Sheng, H.-F., He, Y., Wu, J.-Y., Jiang, Y.-X., Tam, N. F.-Y., Zhou, H.-W., 2012.
Comparison of the levels of bacterial diversity in freshwater, intertidal wetland, and marine
sediments by using millions of illumina tags. Applied and Environmental Microbiology.
78, 8264-8271. 93
Webb, J. M., Quinta, R., Papadimitriou, S., Norman, L., Rigby, M., Thomas, D. N., Le Vay, L.,
2012. Halophyte filter beds for treatment of saline wastewater from aquaculture. Water
Resources. 46, 5102-5114.
White, S. A., Cousins, M. M., 2013. Floating treatment wetland aided remediation of nitrogen and
phosphorus from simulated stormwater runoff. Ecological Engineering. 61, 207-215.
Wu, S., Kuschk, P., Brix, H., Vymazal, J., Dong, R., 2014. Development of constructed wetlands
in performance intensifications for wastewater treatment: a nitrogen and organic matter
targeted review. Water Resources. 57, 40-55.
Yang, Z., Yang, L., Wei, C., Wu, W., Zhao, X., Lu, T., 2018. Enhanced nitrogen removal using
solid carbon source in constructed wetland with limited aeration. Bioresource Technology,
248, 98-103.
Zhang, D. Q., Jinadasa, K. B., Gersberg, R. M., Liu, Y., Tan, S. K., Ng, W. J., 2015. Application
of constructed wetlands for wastewater treatment in tropical and subtropical regions (2000-
2013). Journal of Environmental Science. 30, 30-46.
Zhang, J., Kobert, K., Flouri, T., Stamatakis, A., 2013. PEAR: a fast and accurate illumina paired-
end read merger. Bioinformatics. 30, 614-620.
Zhang, L., Thomas, S., Mitsch, W. J., 2017. Design of real-time and long-term hydrologic and
water quality wetland monitoring stations in South Florida, USA. Ecological Engineering.
108, 446-455.
Zhang, S.Y., Zhou, Q.H., Xu, D., He, F., Cheng, S.P., Liang, W., Du, C., Wu, Z.B., 2010. Vertical-
flow constructed wetlands applied in a recirculating aquaculture system for Channel catfish 94
culture: effects on water quality and zooplankton. Polish Journal of Environmental Studies.
9, 1063e1070.
Zhang, W., Wu, X., Liu, G., Chen, T., Zhang, G., Dong, Z., Yang, X., Hu, P., 2013.
Pyrosequencing reveals bacterial diversity in the rhizosphere of three Phragmites australis
ecotypes. Geomicrobiology Journal. 30, 593-599.
Zhao, F., Xi, S., Yang, X., Yang, W., Li, J., Gu, B., He, Z., 2012. Purifying eutrophic river waters
with integrated floating island systems. Ecological Engineering. 40, 53-60.
Zhou, X., Wang, G., 2010. Nutrient concentration variations during Oenanthe javanica growth
and decay in the ecological floating bed system. Journal of Environmental Science. 22,
1710-1717.