Improved Onsite Wastewater Disinfection and Nutrient Removal for Safe Discharge and Reuse

DISSERTATION

Presented in Partial Fulfillment of the Requirements for the Degree Doctor of Philosophy in the Graduate School of The Ohio State University

Kun Liu, M.S.

Graduate Program in Environmental Science

The Ohio State University

2017

Dissertation Committee:

Dr. Karen Mancl, Adviser

Dr. Jiyoung Lee

Dr. Olli H. Tuovinen

Dr. Peter Ling

Copyrighted by

Kun Liu

2017

Abstract

For onsite wastewater reuse or discharge, fixed media biofilters provide high treatment efficiency with relatively low cost and maintenance. However, these biofilters are not designed to provide disinfection or remove nutrients. Advanced treatment processes are needed to reduce public health and environmental risks before reuse or discharge into surface water. The objectives of this study were to improve onsite wastewater disinfection, to enhance cold weather ammonia removal, and to develop an advanced natural nutrient removal system.

The existing disinfection systems used for onsite wastewater treatment are flow- through UV and chlorine systems. They both have disadvantages. Flow-through systems using UV lamps have continuous lamp operation with fluctuating flows. In these water- cooled systems the lamps overheat, mineral and humic substances foul the lamps and solids accumulate in the disinfection chambers. Flow-through chlorine tablet systems introduce the disinfectant as it dissolves in the flowing water. Variations in flow rate results in inconsistent tablet dissolution and chlorine dosing. Batch UV and chlorine disinfection proved to overcome the drawbacks of conventional systems. The batch UV system shortened daily operation time by two-thirds and used higher flow-through rates in a recirculating system to scour the lamp and eliminate solids settling. The batch design reduced lamp fouling and maintenance while sustaining high disinfection effectiveness. Batch chlorine system using a slow-release chlorine agent provided ii precise control of chlorine dose, allowed for longer contact time and reduced dosing frequency. The 9- and 6-months field tests for UV and chlorine batch systems showed their effectiveness and reliability. This study showed that batch UV and chlorine systems are both improved alternative onsite disinfection systems.

Cold temperature often reduces ammonia removal efficiency in wastewater treatment systems. Improving ammonia removal by warming biofilters in cold weather has not been tested yet. This study investigated two methods to warm biofilters: covering the biofilter surface with polyethylene films and covering the biofilter with insulation in a greenhouse. The results showed that both treatments significantly increased the biofilter temperature by 4 °C on average. However, ammonia removal was not improved correspondingly. Instead, both warming methods might pose negative impacts on biofilters. Anaerobic conditions were established within the covered filters which resulted in loss of ammonia removal capability. The ammonia removal performance of sand biofilter might be affected by many other factors other than just temperature.

Biofilter effluent nitrogen and phosphorus are plant nutrient. An advanced nutrient removal system was developed using a hydroponic floating bed systems with perennial ryegrass. Hydroponic floats and floating beds were designed and tested to solve the issues of consistent seed germination and algae control. The results showed that customized opaque floats and shaded rack with misters offered even grass seeds germination with no surface algae growth. The customized floats that filled the hydroponic beds also insured algae control in the treated effluent. This simple system

iii allowed the plants to grow optimally and develop a healthy root mat under the hydroponic floats, which significantly enhanced the nutrient removal efficiency. Such floating bed systems could alleviate the environmental risk of onsite wastewater discharge and produce an easily harvested, usable crop while reducing the nutrient load into surface water.

Acknowledgments

I wish to thank my advisor, Dr. Karen Mancl, for her intellectual support, encouragement, and enthusiasm throughout this research project.

I am thankful to Dr. Olli Tuovinen for his sight, advice, and editorial assistance. I gratefully acknowledge Dr. Jiyoung Lee and Dr. Peter Ling for their invaluable suggestions, precious comments and encouragement.

I am grateful to Dr. Eunyoung Park for discussing with me various aspects of my projects.

I would also like to thank Nimesha Gunarathna for her assistance and advice in my projects.

I also gratefully acknowledge assistance from Joe Russler and Christopher Gecik for their assistance in experiment preparation and setup.

Vita

July 2006 ...... Quzhou Jr. High

2010 ...... B.S. Civil Engineering, Chongqing University

2012 ...... M.S. Civil Engineering, Purdue University

2013 to present ...... Graduate Research Associate, Department of

Food, Agricultural and Biological

Engineering, The Ohio State University

Fields of Study

Major Field: Environmental Science

Table of Contents

Abstract ...... ii

Acknowledgments ...... v

Vita ...... vi

Fields of Study ...... vi

Table of Contents ...... vii

List of Tables...... xii

List of Figures ...... xv

Chapter 1 Review of Literature ...... 1

Introduction ...... 1

Onsite or decentralized wastewater treatment ...... 3

Wastewater Reuse and Impacts ...... 7

Potential Environmental Benefits ...... 7

Potential Adverse Environmental Effects ...... 9

Public Health Risks ...... 10

Onsite Wastewater Disinfection for Reuse ...... 13

vii

Challenges to Onsite Wastewater Treatment in Nutrient Removal ...... 16

Nutrient Removal of Wastewater using Hydroponic Systems ...... 20

Conclusion ...... 22

Chapter 2: Onsite batch UV disinfection for reuse ...... 23

Introduction ...... 23

Methods ...... 26

Experimental set up and procedure ...... 26

Sampling and Storage ...... 30

Analytical technique ...... 30

Results and Discussion ...... 31

UV lamp performance ...... 31

UV system performance measured with laboratory and field tests ...... 33

Field test under three different flow rate scenarios ...... 40

UV batch irradiator versus continuous UV flow-through reactor ...... 43

Conclusion ...... 46

Chapter 3: Onsite chlorine batch disinfection ...... 48

Introduction ...... 48

Methods ...... 51

Chlorine batch disinfection ...... 52

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Wastewater samples ...... 54

Analytical techniques ...... 54

Results and Discussion ...... 55

Mechanism of NaDCC chlorination ...... 55

Chlorine batch system in field test ...... 58

Chlorine batch reactor vs. UV batch irradiator ...... 60

Conclusion ...... 62

Chapter 4: Temperature management in sand biofilters for enhanced ammonia removal in cold weather ...... 63

Introduction ...... 63

Methods ...... 67

Sand biofilters ...... 68

Greenhouse ...... 68

Biofilter testing and sampling ...... 69

Data measurement ...... 70

Data analysis ...... 70

Results and Discussion ...... 71

Historical data of biofilter performance ...... 71

Biofilter temperature ...... 75

Nutrient removal in three biofilter treatments...... 82

Relationship between biofilter temperature and ammonia removal ...... 88

Conclusion ...... 88

Chapter 5: Development of hydroponic floating bed system for wastewater advanced nutrient removal ...... 90

Introduction ...... 90

Methods ...... 92

Type I float in a pilot greenhouse...... 93

Type II, III and IV floats in laboratory hydroponic bed ...... 94

Type IV and V floats in a high tunnel greenhouse ...... 98

Results and Discussion ...... 99

Pilot greenhouse hydroponic system with type I float ...... 99

Comparison of type II, III and IV hydroponic floats in laboratory test ...... 102

Type IV and V floats in a full scale hydroponic floating bed system ...... 105

Factors affecting the performance of hydroponic systems ...... 110

Conclusion ...... 111

Chapter 6: Wastewater nutrient removal in hydroponic/greenhouse system...... 113

Introduction ...... 113

Methods ...... 116

Pilot test in an uncontrolled greenhouse environment ...... 116

Full-scale test in a high tunnel greenhouse ...... 118

Data Collection and Measurement ...... 119

Statistical analysis ...... 119

Results and Discussion ...... 120

Nutrient removal in pilot greenhouse ...... 120

Nutrient removal in full scale greenhouse ...... 131

Factors affecting hydroponic systems nutrient removal performance ...... 134

Conclusion ...... 137

Chapter 7: Summary and suggestions for future research ...... 138

Summary of Problems ...... 138

Research summary ...... 139

Suggestions for Future Research ...... 140

References ...... 142

Appendix A: Experimental data of batch disinfection ...... 160

Appendix B: Experimental data for biofilter nutrient removal ...... 164

Appendix C: Experimental data for hydroponic nutrient removal ...... 167

List of Tables

Table 1. Comparison of ammonia aquatic life criteria in the 1999 Update, 2009 Draft and

2013 Ambient Water Quality Criteria (AWQC) ...... 19

Table 2. BOD5 and turbidity (± SD) of the wastewater used in the laboratory and field tests ...... 33

Table 3. E. coli viable counts (± SD) after UV disinfection in three laboratory tests ...... 34

Table 4. E. coli viable counts (± SD) after UV batch disinfection in 5 h field tests ...... 37

Table 5. E. coli viable counts (± SD) after UV batch disinfection in 8 h field tests ...... 37

Table 6. The UV doses received in the laboratory and field tests (average light intensity

2.0 × 103 μW/cm2) ...... 39

Table 7. E. coli viable counts (± SD) after UV exposure (27.1 mJ/cm2) with different circulation flow rates in the field batch disinfection ...... 43

Table 8. Free, total residual chlorine (mg/L) and pH (± SD) in 48 hours test of NaDCC reaction with peat biofilter treated wastewater under different NaDCC doses...... 56

Table 9. Total, free chlorine concentration and E. coli count of chlorination after 15 minutes and 24 hours contact time in the field test, average E. coli count in effluent was

26830 (± 1350) CFU/100 mL before chlorination ...... 59

Table 10. Annual, cold and warm months sand biofilter effluent monitoring ...... 72

Table 11. Average water quality parameters raw turkey processing wastewater ...... 82

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Table 12. Effluent ammonia concentration for all 12 full scale biofilters during test period

...... 87

Table 13. Nutrient content of Maxigro hydroponic solution used in laboratory hydroponic test, the nutrient solution was made as 2.5 g/L of Maxigro granule ...... 97

Table 14. Average pollutant concentrations of sand biofilter treated wastewater ...... 117

Table 15. Nutrient contents of rye grass shoot and root after growing in wastewater for 60 days ...... 131

Table 16. E. coli count in three replicates from three lab batch UV test ...... 160

Table 17. E. coli count in three replicates from 3.5 h field batch UV test ...... 161

Table 18. E. coli count in three replicates from 5 h field batch UV test ...... 162

Table 19. E. coli count in three replicates from 8 h field batch UV test ...... 163

Table 20. Biofilter effluent ammonia concentration in three temperature treatment from

02/17/2017 to 05/15/2017 ...... 164

Table 21. Biofilter effluent nitrate concentration in three temperature treatment from

02/17/2017 to 05/15/2017 ...... 165

Table 22. Biofilter effluent total phosphorus concentration in three temperature treatment from 02/17/2017 to 05/15/2017...... 166

Table 23. Pilot greenhouse hydroponic system ammonia concentration and removal ratio

...... 168

Table 24. Pilot greenhouse hydroponic system nitrate concentration and removal ratio 169

Table 25. Pilot greenhouse hydroponic system phosphate concentration and removal ratio

...... 170

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Table 26. Pilot greenhouse hydroponic system HSD, LSD and control group ammonia concentration ...... 171

Table 27. Pilot greenhouse hydroponic system HSD, LSD and control group nitrate concentration ...... 172

Table 28. Pilot greenhouse hydroponic system HSD, LSD and control group phosphate concentration ...... 173

Table 29. High tunnel greenhouse hydroponic system grass and control group ammonia concentration ...... 174

Table 30. High tunnel greenhouse hydroponic system grass and control group nitrate concentration ...... 175

Table 31. High tunnel greenhouse hydroponic system grass and control group phosphate concentration ...... 176

xiv

List of Figures

Figure 1. Internal structure of the Salcor Model 3G UV unit ...... 27

Figure 2. Assembled UV batch system in laboratory test...... 28

Figure 3. Schematic diagram of the field test using UV batch disinfection system.

Samples were collected from (A) septic tank effluent, (B) peat biofilter effluent, and (C)

UV batch disinfection effluent ...... 29

Figure 4. Photo of UV batch system in field test: the UV unit was placed underground and above the storage tank ...... 30

Figure 5. Comparison of UV intensity during the warming up period between 200 h old lamp, old lamp after cleaning and brand new lamp...... 32

Figure 6. Log reduction of E. coli counts (± SD) in laboratory test of UV batch unit in

100L recirculation tank, 136.8 mJ/cm2 dosage received in 20 min ...... 34

Figure 7. E. coli survival ratio (± SD) as a function of the UV batch reactor contact time in A, laboratory test, 136.8 mJ/cm2 dosage received in 20 min; B, the field test, 38.0 mJ/cm2 dosage received in 3.5 h ...... 36

Figure 8. Survival ratio and log removal (± SD) of E. coli after 2.5 h UV exposure under different circulation flow rates: A, mean survival ratio (N/N0) vs. flow rate; B, mean log10 reduction vs. flow rates ...... 41

Figure 9. Maintenance of UV disinfection unit: cleaning up lamp and checking wiring .. 44 xv

Figure 10. UV lamp after 5 months of operation at 3 L/s recirculation flow rate, no obvious fouling was formed along the lamps and covering sleeves ...... 45

Figure 11. Schematic diagram of the field test using chlorine batch disinfection system. 52

Figure 12. Configuration of NaDCC capsules dispenser ...... 53

Figure 13. CT value (± SD) of NaDCC in 5 different dosages, free chlorine concentrations were measured at 10 min, 24 h and 48 h after chlorination ...... 57

Figure 14. Time series observation of nutrient in full scale biofilter effluent from

07/11/13 to 04/28/16: A, ammonia concentration; B, nitrate concentration; C, total phosphorus concentration ...... 73

Figure 15. Distribution of biofilter temperature in 2017 at four ground depths: A, in control biofilter; B, in plastic covered biofilter; C, in greenhouse covered and Styrofoam board insulated biofilter ...... 76

Figure 16. Average temperature (± SD) of three tested biofilter from 02/15/2017 to

05/15/2017 ...... 77

Figure 17. Hourly variation of temperature on 02/16/2017 at various depths of A, greenhouse covered biofilter; B, plastic covered and C, no cover (control); D, Hourly variation of greenhouse indoor air and outdoor air temperature...... 79

Figure 18. Average and standard deviation of three tested biofilter subsurface temperature on 02/16/2017 ...... 81

Figure 19. Average nutrient concentration (± SD) of effluent from three test sand biofilters ...... 83

xvi

Figure 20. Nutrient concentrations in three tested biofilter effluent A, ammonia; B, nitrate;

C, total phosphorus ...... 84

Figure 21. Type I float and pilot-scale hydroponic system in uncontrolled pilot greenhouse (W×L×H = 4 m × 8 m × 2.5 m) ...... 94

Figure 22. Prototypes of hydroponic floats: A, type II; B, type III; C, type IV ...... 95

Figure 23. hydroponic benches integrated with sand biofilter and high tunnel greenhouse ...... 99

Figure 24. Failure of grass seeds germination and burlap was covered with algae ...... 100

Figure 25. comparison of low (left) and high (right) seed density grass production after

30 days of seed germination...... 101

Figure 26. Algae development in hydroponic boxes, effluent showed greenish color due to abundant algae suspended ...... 102

Figure 27. Seed germination result of A, type II; B, type III and C, type IV float. Under the same seeding density (150 g/m2), type II showed lowest germination ratio, type III showed highest germination ratio and uniformity among the three types of floats ...... 103

Figure 28. Germination rack and mister system under hydroponic bed provided optimal germination results of ryegrass seeds ...... 106

Figure 29. Mortality of grass from the center of the float after one week cultivation due to failure of root penetration through the floats into hydroponic bed ...... 107

Figure 30. Grass shoot (A) and root (B) development after two-week cultivation using type V float...... 108

xvii

Figure 31. Wastewater in hydroponic bed, algae growth was eliminated due to complete blockage of light by using opaque floats ...... 109

Figure 32. Nutrient concentration before and after 2 HRT hydroponic float system

- treatments in plant growing stage (after germination): A, nitrate nitrogen (NO3 -N); B,

+ 3- ammonium (NH4 -N); C, Phosphate (PO4 -P) ...... 121

Figure 33. Change of nutrient removal ratio over plant growing stage ...... 124

Figure 34. Comparison of nutrient level before and after treatment of HSD, LSD and no

- + 3- grass: A, nitrate nitrogen (NO3 -N); B, ammonium (NH4 -N); C, phosphate (PO4 -P) . 126

Figure 35. Comparison of wastewater influent and nutrient removal with treatment of

- + HSD, LSD and no grass: A, nitrate nitrogen (NO3 -N); B, ammonium (NH4 -N); C,

3- phosphate (PO4 -P) ...... 129

Figure 36. Comparison of nutrient removal in hydroponic floats with and without rye

- + 3- grass: A, nitrate nitrogen (NO3 -N); B, Ammonium (NH4 -N); C, phosphate (PO4 -P) 133

- + Figure 37. Relationship of the NO3 /NH4 ratio with nitrogen removal in the pilot greenhouse test ...... 136

- + Figure 38. Relationship of NO3 /NH4 ratio with nitrogen removal in high tunnel greenhouse test ...... 136

xviii

Chapter 1 Review of Literature

Introduction

A centralized approach to wastewater treatment has been widely applied in urban areas for decades as it has been accepted as an efficient solution to treat wastewater in populated areas (Opher and Friedler, 2016). However, for rural and some decentralized urban areas, construction of infrastructure to send wastewater to municipal treatment plants is expensive, sometimes making it impractical. Onsite wastewater treatment systems (OWTSs) in these cases have been widely used to treat wastewater in situ and return effluent immediately for discharge or reuse. Conventional OWTSs consist primarily of a septic tank and a soil absorption field, also known as a subsurface wastewater infiltration system (US EPA, 2002). Alternatively, sand or media biofilters are often used to provide advanced treatment of settled wastewater or septic tank effluent

(US EPA, 2002).

The reuse of treated onsite wastewater has benefits such as reducing the nutrient load discharging into the environment and saving cost for discharge conduit, but barriers need to be considered and overcome. One major barrier to wastewater reuse is the associated health risk (Angelakis and Snyder, 2015; Asano, 2007; Park et al., 2016).

Depending on the degree of treatment, a variety of chemical constituents and pathogenic organisms are still present in reclaimed wastewater (Hoerger, et al., 2014; Kostich et al., 1

2014; Pavione et al., 2013). Contact or consumption of reclaimed wastewater or wastewater-irrigated crops presents a risk of waterborne enteric diseases. Another related concern is residual unregulated chemicals in the reclaimed wastewater, such as endocrine disrupting compounds (EDCs), pharmaceuticals and personal care products (PPCPs)

(Comerton et al., 2009).

Disinfection of treated wastewater before disposal or reuse is a critical step for minimizing public health risks (Park et al., 2016). During the past 50 years, a wide variety of disinfection processes have been developed using physical, chemical, and biological agents (Asano, 2007). For onsite wastewater disinfection, the most commonly used disinfectants are calcium hypochlorite tablets and ultraviolet light (UV) (US EPA,

2002).

In addition to pathogens, treated wastewater still contains ammonia, nitrate and phosphorus. Ammonia is toxic to all vertebrates causing convulsions, coma and death

(Randall and Tsui, 2002). When wastewater is discharged into surface water, ammonia can pose hazard to aquatic life, nitrate and phosphorus can cause eutrophication in receiving water bodies. In biological wastewater treatment systems, low temperature slows down or inhibits the biological activity in activated sludge or biofilms especially for ammonia oxidizers (Eighmy and Bishop, 1989). This becomes a concern in ammonia removal in areas with cold season or snow melting periods.

The treated wastewater can provide most of the required N, P for plant growth, making it suitable to fertilize plants (Chen et al., 2016; Ham et al. 2007). Natural systems such as constructed wetlands and hydroponic systems can be used for removal of

2 nutrients in wastewater and produce some useful crops (Chang et al., 2004; Chen et al.,

2009; Chen et al., 2016; Ge et al., 2007; Zurita et al., 2011). Many researches have investigated the feasibility and performance of using hydroponic root mats for wastewater treatment as controlled nutrient removal systems (Bartucca et al., 2016; Chen et al., 2016; Li et al., 2010; Ren et al., 2016).

The goal of this chapter is to review recent research on enhancing performance of onsite wastewater treatment systems to remove pathogens and nutrients including ammonia, nitrate and phosphorus. Specifically, the objectives are to discuss the benefits and risks of wastewater reuse, to explore the limits of disinfection systems, to analyze challenges faced by onsite wastewater treatment in nutrient removal, and to provide an overview of treatment performance of wastewater hydroponic systems in nutrient removal.

Onsite or decentralized wastewater treatment

Approximately 25% of the US population is served by onsite systems, most commonly by conventional septic tanks with soil absorption fields (US EPA, 2002). In typical subsurface wastewater infiltration systems, septic tanks remove most sediments and floating material and partially digest retained organic matter in raw wastewater. The absorption drain field receives effluent from septic tank for further treatment in the soil through biological processes, which work well in removing pathogens, ammonia and phosphorus if it is installed in areas with appropriate conditions and maintained regularly

(US EPA, 2002). The characteristics and performance of OWTSs in removing conventional pollutants such as nitrogen, phosphorous, and microorganisms as well as

3 some trace organic chemicals has been documented by many researchers (Lowe et al.,

2008; Teerlink et al., 2012; Van Cuyk et al., 2001).

Lowe et al. (2008) conducted a field experiment for 2 years testing OWTSs with three different surface architectures (open, stone, and synthetic) and two hydraulic loading rates (4 and 8 cm/day). Over the 2 years, the cumulative mass removal of dissolved organic carbon, total nitrogen, and total phosphorus averaged 94, 42, and 99% respectively. No significant difference of treatment performance was observed under different surface architectures and hydraulic loading rates, and the increased vadose zone depth slightly improved treatment performance.

Teerlink and others (2012) evaluated the efficacy of trace organic chemicals attenuation under different hydraulic loading rate (1, 4, 8, 12, and 30 cm/day) by using a series of bench scale soil treatment units (STUs). After testing with synthetic wastewater and spiked with 17 trace organic chemicals, the results suggested that soil at varying depths were able to attenuate much of the trace organic chemicals in domestic wastewater.

However, the efficacy of attenuation is compound specific.

Forbis-Stokes et al. (2016) simulated the performance of conventional OWTSs as well as two other variations: aerobic treatment units (ATUs) with spray distribution, and mounded OWTSs. The results showed both the conventional and ATU systems failed due to effluent ponding and E. coli transport to the land surface when rainfall intensity was greater than 0.25 cm/h. Only mound systems did not fail under existing conditions as they did not allow Eschericia coli to reach the surface or ponding to occur. Other research conducted by Cooper et al. (2015) compared conventional drain field and advanced

4 systems with inclusion of sand filter pretreatment. The results showed that advanced systems that include sand filter pretreatment and soil-based treatment had better N removal than in conventional treatment systems. Many existing OWTSs located too close to surface and ground water tend to have system failure when experiencing increasing wastewater flows, causing leaching of nutrients and pathogen contamination. US EPA

(2002) reported failure rates of 10%-20% for OWTSs due to hydraulic overloading, poor siting and design, lack of regulation, compliance, and maintenance. Borchardt et al. (2003) found 8-11% incidence of human enteric viruses in private drinking water wells in

Wisconsin, adjacent to onsite adsorption trenches. It was estimated based on this observation that approximately 1.2 million US households may be exposed to contaminated water sources resulting from inadequate treatment of effluent from onsite systems.

In the U.S., only about one-third of the land area has suitable soil conditions for conventional subsurface soil absorption fields (US EPA, 2002). In other areas, sand (or other media) filters are often used as an alternative or advanced treatment method. As a biofilm based technology for wastewater treatment, sand or media biofilters treat wastewater through aerobic digestion of biofilm along with physical processes such as straining and sedimentation. Suspended solids are removed within the pores of the media, and dissolved pollutants are chemically adsorbed onto media surfaces (US EPA, 2002).

These systems have a simple structure, low cost in construction and maintenance, and relatively high efficiency of treatment. The design of the biofilter can be modified for various types of wastewater, making it versatile. For example, high-rate trickling filters

5 have proven to be effective in handling fluctuations of wastewater quality that occur with slaughterhouse wastes (Moodie and GreenWeld, 1978).

Fixed media biofilters may be used for a broad range of applications, from single- family residences to small communities. They are often constructed with single or multiple layers of sand, as well as other media such as peat, coconut chips. For small onsite wastewater treatment, effluent from the septic tank enters the top of the biofilter.

The wastewater then percolates through the biofilter where treatment occurs (Humphrey et al., 2016). The biofilter effluent is collected by an under drain and discharged directly to a surface water or put into reuse through reuse systems.

Although primarily used to treat domestic wastewater, fixed media biofilters can also be modified to treat wastewaters high in organic materials such as wastewater from restaurants or supermarkets, and other food processing facilities (US EPA, 2002). Liu et al. (1999) compared single and multilayer sand bioreactors for dairy wastewater treatment. The 2-layer sand biofilter achieved higher BOD5 removal (85%) than single layer filters (76%). Kang and others (2007) investigated the feasibility of using coarse/fine sand filtration for removing organic materials from turkey processing wastewater. The sand biofilter was modified into a three-layer filter, with 5 cm layer of pea gravel at the bottom to support layers of fine sand (46 cm) and coarse sand (15 cm) to a height of 66 cm. The wastewater contained 1270 ± 730 mg COD/L and was applied to each sand bioreactor at hydraulic loading rates of 264, 132 and 66 L/m2/day. The results showed an over 94% removal of TOC and BOD5 during 80 days of testing at loading

6 rates <132 L/m2/day, which proved the multilayered sand biofilter to be a feasible treatment for turkey processing wastewater.

Wastewater Reuse and Impacts

Rising demands for water to supply agriculture, industry and cities are leading to competition over the allocation of limited fresh water. Water conservation, reuse and recycling can greatly increase the benefits obtained from limited supplies of freshwater resources (Anderson, 2003). At the beginning of twentieth century, the State of California implemented the first wastewater reuse projects in the U.S., driven by water scarcity for agricultural irrigation. With the awareness of potential environmental and health risks, the first regulation on wastewater reuse in agriculture was established in 1918 (California,

1918). Treated wastewater now has been widely used for more than just irrigation, but also involves industrial, urban and potable reuse (GWI, 2010; US EPA, 2012). The main uses of treated wastewater include: irrigation (both agricultural and landscape), recharge of aquifers, seawater barriers, industrial applications, dual-distribution systems for toilet flushing, and other urban uses (Angelakis and Snyder, 2015).

Wastewater reuse can pose great environmental impact as it has the potential to lessen water demand on sensitive water bodies, reduce the discharge of pollutants to environment, extend existing water supplies and benefit receiving ecosystems in reuse

(Atwater et al., 1998). Yet improper reuse of reclaimed wastewater may result in adverse impact to environment such as pollution to soil and groundwater. This section will discuss both potential benefits and adverse effects of wastewater reuse.

Potential Environmental Benefits

Numerous cases have been reported regarding to adverse impacts of wastewater discharge to the environment (Hindell and Quinn, 2000; Kevekordes and Clayton, 2000).

The reuse of wastewater reduces discharge of effluent into sensitive water body and therefore decreases the load of organics, nutrients and microorganisms. In 1997, a $140 million recycling project at the San Jose/Santa Clara Water Pollution Control Plant was completed to provide 80,000 m3 per day of recycled wastewater instead of discharging into the South San Francisco Bay, protecting the natural habitat for the salt marsh area

(US EPA, 1998).

Wastewater reuse helps to conserve freshwater resource and appropriate allocate water to the environment, ensuring good environmental condition for stressed water supplies (Hamilton et al., 2005). As the limited water resource is becoming an important issue, wastewater reuse alleviates this pressure and prevents over-extraction of water resource (US EPA, 2012).

The reuse of wastewater for agricultural irrigation can significantly increase crop growth and yield, as demonstrated by many researchers (e.g., Anderson, 2003; Jang et al.,

2010; Singh et al., 2012). Ham et al. (2007) studied paddy rice field irrigation with reclaimed wastewater after stabilization pond and constructed wetlands treatment. The resulting average crop yield was over 50% greater than control field. Singh et al. (2010) found that application of sewage water increased the yield of Rabi crops as well as total

N, P, K and organic carbon content of soil compared to irrigation with well water. The reuse process turned excess nutrients in wastewater into fertilizer, which also reduced the use of chemical fertilizer (Jang et al., 2012).

Groundwater recharge using reclaimed wastewater in areas facing water scarcity and over-extraction of groundwater has been proved to be effective in stabilizing groundwater table. The groundwater recharge project in southern Italy established an aquifer circulation between shallow groundwater and deep groundwater, draining treated wastewater into a swallow hole linked both levels (De et al., 2007). In coastal areas like

Whittier Narrows in Los Angeles, California, high sanitary quality reclaimed wastewater has been reused for direct groundwater recharge to maintain the level of water table and prevent seawater intrusion (Anderson, 2003). Other cases have been found in application of wetland habitat restoration using wastewater as augmented flow. Velty et al. (2006) investigated using purified municipal wastewater in rewetting degraded peatland as prerequisite for complete restoration in Northeast Germany. The results proved that municipal wastewater could be an alternative solution for water supply required during rewetting of peat soils.

Potential Adverse Environmental Effects

In some coastal areas, the salinity levels of reclaimed water can be higher than traditional sources of water, which would change soil chemical and physical properties

(Hamilton et al., 2005). Soil salinity and sodicity are considered as constraints when irrigation with wastewater in these areas (Jalali et al., 2008; Surapaneni and Olsson,

2002). When irrigating with sodic water, sodium absorption ratio (SAR) is often introduced which describes the relation between soluble sodium and soluble divalent cations (Ca2+ and Mg2+) (Hamilton et al., 2005). High SAR in reclaimed wastewater is one of the major concerns in reuse. Sodium and other forms of salinity can significantly

9 affect soil property by changing its cation exchange capacity (CEC), thus affecting soil permeability, water retention capacity and nutrient uptake by plants (Halliwell et al.,

2001; Saidi, 2012).

Large scale of irrigation with reclaimed wastewater may result in the leaching of salt and nitrate, which is considered to be the most significant environmental impact to groundwater (Paruch, 2014). The study of Paruch (2014) using baker's yeast wastewater for irrigation revealed that at high water table region, the irrigation had an extremely significant impact on the chemical status of groundwater with high values of COD, N-

+ - NH4 , N-NO3 , Cl and Na. Some other emerging contaminants may be found in reused wastewater such as EDCs, pharmaceutically active compounds (PhAC) and heavy metals originate either from industrial or domestic sources. A recent study in the Czech Republic and in Switzerland (Macikova et al., 2014) investigated 36 endocrine disrupting compounds and their impact to aquatic life. The result suggested a positive relationship of these chemicals to fish glucocorticoid plasma levels. An evaluation of 39 micro- pollutants in wastewater was conducted in Vidy Bay, Switzerland (Hoerger et al., 2014).

It was found that the micro-pollutants from wastewater could extend into the deep lake and in the direction of a downstream drinking water intake, posing potential ecotoxicological risk.

Public Health Risks

There has always been a concern of potential health effects associated with the wastewater reuse. Depending on the characteristic of reclaimed wastewater, degree of treatment, and their reuse pathways, the related public health impacts can be classified

10 into two categories: biological risks due to presence of microbial pathogens and chemical risks from pharmaceuticals and personal care products, heavy metals and nanoparticles.

As a major concern, risks arising from microbiological contamination limit the wide-spread application of wastewater reuse (Hamilton et al., 2007). The occurrence of pathogens in reclaimed wastewater and their transmission to human by direct consumption is of most concern as enteric pathogens are the most common microbial pathogens in wastewater (Becerra-Castro et al., 2015; Okoh et al., 2007; Solomon et al.,

2002). The risks posed by pathogens transmitted through wastewater irrigation are complicated to estimate, which depend upon the survival of pathogens, the infective dose, the extent of exposure and the immunity of host (Shuval and Fattal, 2003).

For onsite irrigation, commonly used coliform bacteria may not indicate the presence of viruses and protozoa in wastewater due to different susceptibility to different disinfection methods (Blatchley et al., 2007; Park et al., 2016). For example, Park et al.

(2016) investigated the onsite chlorination effectiveness on bacteria, viruses and spore- formers. The viable Cl. perfringens and F-specific coliphages were frequently detected in the storage tank for irrigation. In addition, the F-specific coliphage removal was significantly reduced during the winter/spring season (Park et al., 2016). These results indicated the potential viral and protozoan risks in onsite wastewater irrigation through ingestion or contact of plants and soils being irrigated.

Among factors affecting biological risks of wastewater irrigation, the type of plants produced and the method employed for wastewater irrigation plays an important role in transmission of pathogens through food chain (Becerra-Castro et al., 2015). Lower

11 risk of irrigation on fruit trees or vegetables cultivated on vines has been found than on leaf vegetables with direct contact with soil and irrigation wastewater (Cirelli et al., 2012).

For salad crops such as lettuce, the variation of irrigation wastewater quality has greater impact on its health risk (Pavione et al., 2013).

Studies regarding to interaction between microorganisms and plants being irrigated with wastewater have been conducted in recent years. Some species of opportunistic pathogens such as Citrobacter freundii, Enterobacter cloacae, E. coli,

Enterobacter sakazakii and Klebsiella pneumoniae were found recovered in vegetable after wastewater irrigation (Ibenyassine et al., 2007). Moreover, some pathogens can internalize into plants which poses tremendous health risk if this happens to fresh produce that are usually consumed raw (Ge et al., 2012; Ge et al., 2013; Wachtel et al.

2002). Wachtel et al. (2002) found that E. coli O157:H7 can be internalized into lettuce through plant roots through irrigation water. Ge et al. (2013) investigated the pattern of

Salmonella internalization into green onions and environmental factors affecting the extent of this internalization. The results showed that Salmonella Typhimurium can be taken up through plant top part and transported to the lower part, where viable internalized Salmonella can survive and maintain their viability. Many other researchers have also proven that plants serves as host for these bacteria, irrigation with wastewater can extend their survival and even promote their propagation (Ibekwe et al., 2004; Tyler and Triplett, 2008).

Some emerging contaminants in reclaimed wastewater are receiving increased attention due to the vast use of organic chemicals and pharmaceuticals including

12 compounds such as EDCs and PPCPs. There are still many uncertainties about the impact of EDCs on human health. Based on epidemiological data on adverse effects on human health from EDCs, low level of exposure to EDCs has not found to be harmful (Belgiorno et al., 2007). A risk assessment of PPCPs was performed by Roccaro and Vagliasindi

(2014) in agriculture application, the preliminary conclusion was that the presence of the

PPCPs posed a low hazard to the human health. However, Kostich et al. (2014) estimated the potential risks of 56 pharmaceuticals in 50 wastewater treatment plants in US, and found six with greatest risks: valsartan, hydrochlorothiazide, metoprolol, atenolol, lisinopril, and enalaprilat, suggesting more detailed study of potential health impacts.

Although no established regulation has yet been set to regulate EDCs and PPCPs, the current increasing number of known EDCs and PPCPs present in reclaimed wastewater suggests precautionary principles in selecting advanced treatment techniques for better removal of these contaminants.

Onsite Wastewater Disinfection for Reuse

Conventional disinfection processes, namely chlorination and UV radiation, have been used in wastewater disinfectant for onsite reuse to reduce environmental and public health risks (US EPA, 2002). Chlorine has been used as a water and wastewater disinfectant due to its high inactivation effectiveness and relatively low cost. Park et al.

(2016) investigated an onsite wastewater treatment and reuse system which treats a single family home wastewater with peat biofilter and reuses treated wastewater for lawn irrigation after chlorine disinfection. The results of pathogen inactivation showed 5.4, 2.3 and 2.5 log reductions to E. coli, coliphages and Cl. perfringens respectively. Among the

13 three pathogen indicators tested after biofilter and chlorination, geometric mean (GM) of viable counts of E. coli was reduced from 2.5 × 105 CFU 100 mL-1 to 8.4 × 101 CFU 100 mL-1 after peat biofilter treatment, and was not detected after chlorination. However, viable Cl. perfringens and F-specific coliphages were not completely inactivated. The results suggested adjustments in the onsite wastewater treatment to accommodate public health concerns about viruses, spore-formers and antibiotic resistance genes.

Another commonly applied technology for disinfection is ultraviolet (UV) light irradiation. Leverenz et al. (2007) compared chlorine with UV disinfection system in onsite wastewater treatment. They collected operation and maintenance data for onsite wastewater disinfection systems and to compare the performance, reliability, maintenance requirements and costs of calcium hypochlorite tablet chlorination unit and

UV disinfection unit. The results showed that for hypochlorite tablet system, the overall performance was excellent in bacteria and virus removal, and the effluent quality did not affect the effectiveness of disinfection. The dissolution rate of the tablet and consistency of chlorine dose are main concern of reliability. The vacation period resulted in high residual chlorine may require dechlorination facilities. Weekly inspection was necessary to ensure the tablets are present and feeding properly, thus easy maintenance is the advantage of this system. For UV system, the performance was maximized with TSS less than 5 mg/L and turbidity less than 3 NTU. Two types of fouling were discovered because of the hardness minerals and humic substances. The hardness of the water, the influent water quality and quantity of pretreatment and period while lamp was left on

14 without flow were constraints of its reliability. Maintenance should be more frequent than every six months for cleaning the lamp and Teflon liner.

UV disinfection for onsite reuse is effective at inactivating most viruses, spores and cysts and leaves no residual that may be harmful to humans and wild life (US EPA,

1999). However, microorganisms can sometimes repair by themselves after UV irradiation through photoreactivation or in the absence of light (US EPA, 1999). Yoon et al. (2006) conducted a pilot study of microorganism repair after UV disinfection was performed for agricultural reuse. In low dose UV disinfection (6 mJ/cm2) microorganisms reactivated within 12 h by approximately 5% and 1% due to photoreactivation and dark repair, respectively. However, this increase was not significant at high UV dose (16 mJ/cm2). Kollu and Örmeci (2015) investigated the regrowth potential of E. coli and indigenous wastewater bacteria after UV disinfection in the absence of light and obtained opposite results. Higher percent regrowth of E. coli and indigenous wastewater bacteria were observed after UV disinfection at 40 mJ/cm2 than at 15 mJ/cm2. Addition of nutrients after disinfection did not boost dark repair of E. coli but increased its regrowth.

Also the regrowth of indigenous coliform bacteria in wastewater was 50 and 200 times higher than the regrowth of laboratory cultured E. coli at 15 mJ/cm2 and 40 mJ/cm2, respectively.

Another disadvantage of UV disinfection is its less efficiency in wastewater with high turbidity and total suspended solids (TSS) (US EPA, 1999). Azimi et al. (2012) investigated the causes of the tailing phenomenon and its relationship to the structure of wastewater bioflocs. The results showed that biofloc structure had a significant effect on

15 disinfectability with UV light. The steep initial slope of the UV dose response curve

(DRC) of un-sheared biofloc samples was not caused by the free-floating organisms, but due to the presence of UV-susceptible bioflocs. The compact cores contained in the bioflocs played a significant role in UV radiation resistance. Larger bioflocs contained a larger number and a larger volume of cores were more likely to be resistant to UV disinfection.

Recent year studies focused on improving UV performance in disinfection by integration with other disinfection processes to overcome the limitations of UV stand- alone disinfection. Advanced oxidation processes (AOPs) have been successfully examined for removing a wide range of contaminants (Ferro et al., 2016; Malato et al.,

2009). Ferro et al. (2016) evaluated the effect of UV/H2O2 disinfection on inactivation of total coliforms, E. coli and antibiotic resistant E. coli. The detection limit (5 CFU mL-1) was achieved after 90 min treatment and the UV/H2O2 disinfection showed high efficiency in antibiotic resistant E. coli inactivation. Lee and others (2016) modeled and tested UV/H2O2 processes combined with O3/H2O2 in attenuating trace organic contaminants. The results showed that the combined process of O3/H2O2 followed by

UV/H2O2 can achieve superior trace organic contaminants abatement.

Challenges to Onsite Wastewater Treatment in Nutrient Removal

Onsite and decentralized wastewater effluents from biofilters often contains excessive nutrients that may cause eutrophication and have significant impact on ecosystem health (Cao et al., 2011; Tylova-Munzarova et al., 2005). In the U.S., over 20% of surface waters impairments are related to nutrient discharge, including oxygen

16 depletion, algal blooms, and toxicity (US EPA, 2009). Some toxic algae may produce toxins that pose hazard towards human and wildlife (Lapointe et al., 2015). For onsite treatment systems, the wastewater nitrogen entering the septic tank contains mostly organic nitrogen (ON) and ammonium (US EPA, 2002). Organic nitrogen is partially mineralized to ammonium via ammonification in the septic tank and the effluent is piped to the drain field trenches or fixed media filters, where ammonium may be nitrified into nitrate via nitrification process (Humphrey et al., 2010).

Humphrey and others (2016) tested the total dissolved nitrogen (TDN) treatment efficiency of conventional OWTS and single-pass sand filters. The conventional OWTSs had 98% and 70% reduction of TDN concentrations and masses, respectively, from samples taken 35 m downgradient. The sand filter OWTS reduced TDN concentration and mass by 80% and 50% respectively. Denitrification was found to be the mass removal mechanism in conventional OWTS and thus resulting in better TDN removal.

Diaz-Elsayed et al. (2017) conducted a cost-benefit analysis on nutrient management of the advanced OWTSs, including aerobic treatment units (ATUs) and passive nitrogen reduction systems (PNRs). The results showed that the advanced

OWTSs greatly improved the TN removal from 61-65% to 97%-100% than conventional systems, but at the expense of approximately twice the cost and the remaining environmental impact categories. Thus the transition from conventional to advanced

OWTSs for improved nutrient management remains to be a challenge.

The failure of OWTSs may result in ponding of drain field or media filter which often leads to insufficient nitrification process within the treatment system and exceeding

17 ammonium limit in effluent (US EPA, 2002). In 2013, US EPA announced new final ammonia aquatic life criteria, reflecting the latest scientific information on freshwater mussel and snail sensitivity to ammonia (US EPA, 2013). The comparison between 1999,

2009 and 2013 ammonia aquatic life criteria is shown in Table 1.

In biological wastewater treatment systems, low temperature slows down or inhibits the biological activity in activated sludge or biofilms especially for ammonia oxidizers (Eighmy and Bishop, 1989). This becomes a concern in ammonia removal in areas with cold season or snow melting periods. Lee and others (1999) investigated the effect of low temperature on ammonia removal in constructed wetland. As temperature reduced from 23°C to 5°C, ammonia removal and nitrification were decreased by 20%.

For oxidation ditch process in wastewater treatment plants (WWTP), low temperature also can pose a negative impact on ammonia removal (Yang, et al., 2013). Zhang and others (2016) compared the ammonia removal in activated sludge system and biofilm system. The results showed that at low temperatures, biofilm system had lower ammonia removal ability but higher stability compared to activated sludge system.

Table 1. Comparison of ammonia aquatic life criteria in the 1999 Update, 2009 Draft and 2013 Ambient Water Quality Criteria (AWQC)

2009 Draft 1999 AWQC 2013 AWQC Update AWQC Update Update Criteria Criteria Criteriac

Criterion Duration pH 7.0,T = 20°C (mg/L)

Acute (1-hr average) 24a 19 17

Chronic (30-drolling 4.5b 0.91 1.9* average)

*Not to exceed 2.5 times CCC or 4.8 mg TAN/L (at pH 7, 20°C) as a 4-day average within the 30-days, more than once in three years on average.

Criteria frequency: Not to be exceeded more than once in three years on average. a Salmonids present b Based on renormalization of data to pH 7 and 20°C c Mussels present

Kauppinen and others (2014) studied the nutrient removal properties of three pilot-scale sand filters (SFs) in a cold temperature climate over a one-year period.

Remarkable differences were noted between the three SFs which were related to the construction and planning of the filters. The winter did not seem to have any significant impact on nutrient removal, showing resistance of SFs to temperature changes in terms of nutrient removal.

Laaksonen and others (2017) investigated an onsite sand filter for treatment of household wastewater in cold weather areas. Throughout the test period, BOD5 and COD removal remained constantly high between 92% and 98%. The reduction of total nitrogen and total phosphorus declined after initial period of 3 months and varied between 5-25% and 50-65%, respectively. The nitrification was efficient and nitrogen in the effluent was predominantly in the form of nitrate. It was concluded that the sand filter had high efficiency in organic load reduction but its nitrogen and phosphorus load reduction was insufficient.

Nutrient Removal of Wastewater using Hydroponic Systems

Hydroponic systems are systems that grow plants in a nutrient solution instead of soil. The plant is usually provided with structural support media. The high nutrient content of treated wastewater makes it possible in hydroponic systems to be reused and applied to the root system of the plant instead of nutrient solution. Various system setups and plants were tested for feasibility to be used as wastewater treatment and nutrient removal systems (Haddad et al., 2011, 2012; Xu et al., 2014; Vaillant et al., 2003).

Vaillant et al. (2003) applied nutrient film technique (NFT) to treat primary wastewater to discharge standards. The research adopted commercial greenhouse NFT system to irrigate Datura innoxia plants with wastewater. The results showed it was effective in reducing wastewater TSS, BOD5, COD, TP and TN after 48 h of operation.

Three major nitrogen removal mechanisms have been identified: microbial denitrification, plant nitrogen uptake and volatilization. The nitrification process was affected by the

BOD5 level because of the oxygen competition. It was concluded that NFT system using

D. innoxia made it possible in wastewater treatment to achieve the permitted levels for discharge.

Haddad et al. (2011) used hydroponic system for decentralized wastewater treatment and reuse. A 4-year test was conducted using various plants to treat wastewater in hydroponic barrels and channels filled with soilless-media. The total nitrogen removal and total phosphorous removal were 13-47% and 30% respectively. After modification to five consecutive treatments, the TN removal was improved to 62-65%. Plants had a better growth and yield in hydroponic barrels than in horizontal channels. After the system upgrade to a gradual multi-stage vertical flow hydroponic system, higher removal efficiency of solubale organics, suspended solids and nitrogen was observed (Haddad et al., 2012). It was concluded that gradual hydroponic systems could be successfully used as small decentralized wastewater treatment systems.

Xu and others (2014) investigated efficiency of a laboratory-scale plate/fabric/grass hydroponic system in nutrient removal of turkey processing wastewater after secondary treatment. Average removal of 53% nitrate-N and 68% o-phosphate-P was attained at 48 h at 2-day hydraulic retention time. Peak nutrient removal was achieved on day 22 of grass growth, and multiple harvesting of the biomass can be used to keep the plants at the stage of highest nutrient removal efficiency. Except for plant root uptake, microbes associated with the plant roots and the fabric media likely played another important role due to nutrient assimilation and denitrification. The plate/fabric/grass hydroponic system showed its effectiveness in rapid removal of nutrients from secondary treated wastewater.

Conclusion

Onsite wastewater treatment system offers a simple and effective solution for wastewater reuse and disposal in decentralized areas. However, disinfection and advanced nutrient removal method may be required if wastewater is for reuse or discharge, respectively. With the increase of reuse application of onsite wastewater, its benefit and potential impacts will always coexist, and onsite disinfection process is able to reduce its public health risk. When facing challenges in nutrient removal, advanced nutrient removal process should be considered. Wastewater hydroponic system has been proven to be efficient in nitrogen and phosphorus reduction as a biological nutrient removal method.

Chapter 2: Onsite batch UV disinfection for reuse

Introduction

Onsite or decentralized wastewater treatment systems eliminate the infrastructure for transporting the wastewater, focus on adequate treatment of wastewater and dispersal of treated effluent at where it’s generated. To minimize the possible environmental and public health risks, disinfection of effluent before discharge or reuse is required. The goal of onsite wastewater disinfection after secondary treatment is to minimize the exposure of public to bacterial, viral, and protozoan pathogens. In the United States, Escherichia coli or fecal coliform counts in treated wastewater effluents are regulated for National Pollutant Discharge Elimination System

(NPDES) permits (US EPA, 2012). In Ohio, the Ohio Environmental Protection Agency uses E. coli limits in their discharging household sewage treatment systems for the

NPDES general permits (Ohio EPA, 2016).

Same as wastewater treatment plants, onsite wastewater treatment systems are multistage processes to reduce the environmental and health risks of effluent discharge.

The septic tank provides primary treatment to separate the suspended solids and grease from the effluent. Secondary treatment removes dissolved organic matter and a significant amount of pathogens (US EPA, 2002). Secondary treatment systems such as subsurface wastewater infiltration system and sand or media biofilters are used in 23 decentralized systems. To prevent the spread of waterborne diseases and to further reduce public health problems, tertiary treatment would be required such as chlorination and or ultraviolet irradiation as means of disinfection (Mounaouer and Abdennaceur,

2016; Zyara et al., 2016).

Onsite wastewater treatment systems may have long intervals between maintenance, a lack of redundant systems, high variability in flow rates, and other site- specific factors (Leverenz et al., 2007). These characteristics require onsite disinfection systems to be simple, reliable and efficient in variant conditions. Chlorine and ultra-violet

(UV) light are most commonly used agents in onsite wastewater disinfection according to these requirements. The conventional onsite chlorinator often use NaOCl tablets feeder to make contact with flow-through wastewater, which is an uncontrolled, passive process. The dissolution rate of chlorine tablets into wastewater depends on the effluent pipe flow rate, which may lead to uneven dosing as flow rate varies greatly (Leverenz et al., 2007). UV disinfection technologies are of increasing interest recently due to its simplicity to use onsite and its viability of reduction of all waterborne pathogens (Hijnen et al., 2006).

UV-based disinfection systems transmit high radiation energy at <400 nm that causes damage to cellular nucleic acids and other biomolecules, causing mutations in mild doses and mortality in high doses. UV dose, expressed as mJ/cm2, is the amount of

UV energy penetrating the water, multiplied by the amount of time the water is exposed to this energy; i.e., I (intensity W/cm2) × T (exposure times). UV doses are usually referenced to transmittance in distilled water measured at 254 nm (= 100%). Other than

UV dose, the efficacy of UV irradiation in disinfection process is also related to the characteristics of wastewater flow such as flow rate, microorganism concentration and transmittance (Mounaouer and Abdennaceur, 2016).

Utilities using disinfection typically target the doses to yield four log cycles of microbial inactivation. In general, doses of UV-based disinfection systems can inactivate pathogenic microorganisms including viruses, vegetative bacterial cells, spores and cysts

(Oguma et al., 2002; Hijnen et al., 2006; Nasser et al., 2006). For example, for four log reduction in drinking water disinfection the doses are in the approximate range of 6 for E. coli, 10 for Cryptosporidium oocysts and 30-36 mJ/cm² for human enteric viruses

(http://www.trojanuv.com/). UV radiation does not leave residual effects and does not produce chlorinated by-products, which may pose a health hazard. These features make

UV radiation an ideal system for onsite wastewater disinfection.

The onsite UV disinfection system is relatively simple to install, operate and maintain in small flow treatment plants. Microorganisms may recover and regrow if UV disinfection is inadequate, and regrowth of partially damaged or injured microbes is also possible in chemical disinfection (Talon et al., 2005; Gilboa and Friedler, 2008; Friedler and Gilboa, 2010; Hallmich and Gehr, 2010; Guo et al., 2011, 2013; Zhou et al., 2017).

UV irradiation combined with chlorine disinfection yields additive and more potent efficacy than either technique alone in bacterial inactivation (Mounaouer and

Abdennaceur, 2016; Zyara et al., 2016). UV irradiation coupled with electrocoagulation has also been tested in efforts to improve pathogen and turbidity removal (Cotillas et al.,

2014) and to prevent bacterial photoreactivation and biofouling of the UV lamp (Haaken

25 et al., 2013). When combined with chemical oxidants such as hydrogen peroxide and persulfate, UV irradiation has powerful potential in not only removing pathogens, but also destroying other pollutants such as pharmaceuticals and personal care products

(Kwon et al., 2015; Afonso-Olivares et al., 2016).

In a typical small flow onsite UV disinfection system, a low-pressure UV irradiator is connected to the outlet of the secondary treatment. In a residential setting, the peak water usage, wastewater treatment and outflow may only last a few hours each day, mostly in the morning and in the evening on weekdays. In a continuous flow system, the UV lamp is on at all times regardless of fluctuations in water usage.

Continuous irradiation leads to lamp overheating and fouling because the UV lamp is immersed in non-flowing wastewater for several hours each day (Leverenz et al., 2007;

Babcock Jr. et al., 2004). An excessive idle time for UV lamps significantly reduces their effective life.

The objectives of this study were to evaluate batch UV disinfection as an alternative to conventional flow-through UV disinfection. The efficacy of UV- disinfection was tested by monitoring survival of E. coli. The performance and reliability of UV batch disinfection system was monitored in the field, and factors that may affect its efficiency were evaluated.

Methods

Experimental set up and procedure

A Salcor Model 3G UV Ultra-Violet Septic Disinfection Unit (254 nm UV-C,

38W, Salcor, Inc., Fallbrook, CA) was used to generate UV irradiation. The UV unit had

26 one inlet and one outlet and contained a 25 W low-pressure UV lamp, as shown in Figure

1. To transfer this UV flow-through unit into a UV batch system, a recirculation chamber was added in both lab and field tests, allowing wastewater to circulate through UV unit for a predetermined time. A sump pump was used for recirculation.

Figure 1. Internal structure of the Salcor Model 3G UV unit

The laboratory UV disinfection test utilized a cylindrical 1.3 m × 0.65 m recirculation tank installed with the UV disinfection unit on top (Figure 2). The tank was filled with 100 L diluted raw wastewater (1 part wastewater + 100 parts tap water) from a turkey processing plant (Kopp Turkey, Inc., Harrison, OH). Before dilution, the 100 parts tap water was initially stored in the tank for over two days until no residual chlorine could be detected. A pump with flow rate of 0.5 L/s was installed inside the recirculation tank. The circulation time for 100 L wastewater treatment was set at 20 min, with both pump and UV lamp on simultaneously.

Figure 2. Assembled UV batch system in laboratory test

The field test was treating onsite residential wastewater for a three-bedroom, single-family house located in London, OH (39°57'44.4"N, 83°26'00.6"W). The wastewater was first treated onsite in a septic tank and then passed through a peat filter

(Premier Tech Environment model STC-650, Quebec, QC). The peat filter effluent was accumulated in a 3,785 L (1,000 gal) storage tank, which also served as a chamber for

UV disinfection (Figure 3). UV disinfection was set to turn on daily while the peat filter

28 effluent in the UV chamber was recirculated (Figure 4). Normal daily operation was 3.5 h at 0.5 L/s recirculation before landscape irrigation. Recirculation time of 5 h and 8 h in the field were also examined. Disinfection efficiency was compared at 0.5, 2.0, and 3.0

L/s recirculation flow rates, with 2.5 h of recirculation time and sample collection every

30 min. The UV disinfected effluents were eventually discharged for landscape irrigation nearby.

Figure 3. Schematic diagram of the field test using UV batch disinfection system. Samples were collected from (A) septic tank effluent, (B) peat biofilter effluent, and (C) UV batch disinfection effluent

Figure 4. Photo of UV batch system in field test: the UV unit was placed underground and above the storage tank

Sampling and Storage

For both field and laboratory tests, before the UV disinfection, wastewater was sampled and tested for biochemical oxygen demand (BOD5) and turbidity. For laboratory tests, water samples were collected from the start of the recirculation and at 5 min intervals till the end of disinfection. In the field test, samples were taken every 30 min during the 3 h recirculation period. Samples from disinfection tests were stored at 4 °C and analyzed within 12 h of storage.

Analytical technique

E. coli was enumerated on solid media using EPA Method 1603 m-TEC agar (US

EPA 2006). Water samples were filtered using 0.45 µm pore size membrane filters (47 mm). For E. coli, the membrane filters were incubated on Difco modified m-TEC agar at

35 °C for two h and then at 44.5  0.2 °C for 18-20 h. E. coli colony forming units (CFU)

30 were counted based on magenta colored colonies. All samples were processed in duplicates. For samples with above 200 CFU, appropriate dilutions (102 ~ 103) in 50 mM phosphate-buffered saline were applied to obtain countable numbers with accuracy.

Measurement of BOD5 was conducted according to Standard Methods (Rice et al., 2012).

Turbidity was measured using a Model 2100Q Portable Turbidimeter (Hach Company,

Loveland, CO).

UV intensity was measured (General Tools UV254SD Data Logging UVA &

UVC meter, Secaucus, NJ) at the distance of 10 cm to the UV lamp tube when UV unit was operating with circulation water going through. The UV dose (mJ/cm2) was calculated using the intensity multiplied by the exposure time.

Analysis of variance (ANOVA) was conducted using JMP 11 software to evaluate the data. ANOVA was used to screen for significance of circulation time and flow rate.

Tests were assessed at a significance level of 0.05. Where possible, the standard deviation

(SD) of the mean was calculated for the replicates.

Results and Discussion

UV lamp performance

The UV intensity was measured after the light intensity stabilized following the switching of the lamp on, which was after the first 12 minutes. Both a new lamp and one used for 200 h were tested. The UV lamp intensity was tested under laboratory conditions

(22 C). For both lamps, the light intensity stabilized after 10 min, with mean values of

1640 W/cm2 for the used lamp and 2040 W/cm2 for the new one (Figure 5). After over

200 h, a 20% loss of intensity was observed for the used lamp. UV disinfection is

31 affected by suspended solids and fouling on the lamps (Lin et al., 1999; Hua and

Thompson, 2000; Wait et al., 2007). The used lamp was cleaned with a soft cloth and isopropyl alcohol to remove fouling deposits and other residues. After cleaning, the lamp intensity significantly (p = 0.012) improved with a mean value of 2630 W/cm2, exceeding the new lamp output. These test results showed that the UV lamps require routine maintenance and cleaning to keep them at best performance.

3500

3000

)

2 2500

2000 Old lamp 1500 Cleaned New lamp

UV intensity (µw/cmintensity UV 1000

500

0 0 2 4 6 8 10 12 Time (min)

Figure 5. Comparison of UV intensity during the warming up period between 200 h old lamp, old lamp after cleaning and brand new lamp

UV system performance measured with laboratory and field tests

The BOD5 and turbidity data for the wastewater samples in the laboratory and field tests are listed in Table 2. The slaughterhouse wastewater used in the laboratory tests had a about 5-fold higher BOD5 and turbidity as compared to the peat filter effluent in the field test.

Table 2. BOD5 and turbidity (± SD) of the wastewater used in the laboratory and field tests

Test BOD5 (mg/L) Turbidity (NTU) Laboratory 67.2 8.8 33.4 11.2 Field 12.8 2.2 6.4 1.1

The results of E. coli inactivation as a function of the contact time in the laboratory tests are listed in Table 3. For E. coli, the results showed a 3.7 log reduction after 20 min circulation through the UV batch unit. The logarithmic reduction fits a polynomial regression trend, indicating a slight tailing effect over the 20 min period

(Figure 6). The diluted raw wastewater used in the lab test contained relatively high TSS which might result in this tailing effect.

Table 3. E. coli viable counts (± SD) after UV disinfection in three laboratory tests

Contact E. coli (CFU/100 mL) Time (min) Test 1 Test 2 Test 3 0 6.0 × 103 ± 1.4 × 102 1.4 × 103 ± 2.9 × 102 9.4 × 103 ± 9.0 × 102 5 5.0 × 102 ± 26 2.0 × 102 ± 35 2.1 × 102 ± 41 10 30 ± 3.1 45 ± 14 38 ± 4.2 15 5.0 ± 0.7 15 ± 1.3 5.0 ± 2.0 20 1.0 ± 0.6 6.0 ± 1.5 3.0 ± 0.3

4.5

4 3.5

3 CFU/100 mL) CFU/100

2.5 10

2 (log 1.5

E. coli coli E. 1

log 0.5 0 0 5 10 15 20 Time (min)

Figure 6. Log reduction of E. coli counts (± SD) in laboratory test of UV batch unit in 100L recirculation tank, 136.8 mJ/cm2 dosage received in 20 min

The survival ratios of E. coli in response to UV contact time in the laboratory tests are shown in Figure 7A. The survival ratio was calculated by Nt/N0, where N0 is the 34 initial count of E. coli at beginning of disinfection and Nt is after indicated time of disinfection. Over 90% of E. coli was inactivated within the first 5 min (1-log removal).

Figure 7B shows the survival ratio of E. coli over 3.5 h in the field test. The average E. coli counts were reduced from 2.2 × 104 CFU/100 mL to 1.1×103 CFU/100 mL

(geometric mean) within 3.5 h of circulation through the UV unit, amounting to only

1.3 log reduction. Also large variance of E. coli inactivation was found among the field tests compared to laboratory tests, indicating variation of wastewater quality and other environmental factors might have impact on performance of UV inactivation. The log reduction of E. coli was improved after increasing daily circulation time to 5 h and 8 h, shown in Tables 4 and 5. The two tests of 5 h recirculation achieved 2.0 log reductions in average, and the 8 h recirculation test gained 3.1 log reductions and brought E. coli count in effluent to a safety level for reuse.

100 90

80 70 A 60 50

survival ratio (%) ratio survival 40 30

E. coli coli E. 20 10 0 0 5 10 15 20 25 Time (min)

100 90 80 B 70 60 50

40 survival ratio (%) ratiosurvival 30

E. coli coli E. 20 10 0 0 1 2 3 4 Time (hour)

Table 4. E. coli viable counts (± SD) after UV batch disinfection in 5 h field tests

Contact E. coli (CFU/100 mL) Time (h) Test 1 Test 2 0 3.7 × 104 ± 7.5 × 102 3.4 × 104 ± 5.6 × 102 1 1.5 × 104 ± 6.8 × 102 1.4 × 104 ± 1.6 × 102 3 1.5 × 103 ± 65 1.5 × 103 ±29 4 6.8 × 102 ± 33 8.0 × 102 ± 21 5 2.9 × 102 ± 50 5.0 × 102 ± 17

Table 5. E. coli viable counts (± SD) after UV batch disinfection in 8 h field tests

Contact E. coli (CFU/100 mL) Time (h) Test 1 Test 2 0 2.8 × 104 ± 8.5 × 102 2.7 × 104 ± 1.1 × 103 2 3.1 × 103 ± 3.1 × 102 2.9 × 103 ± 1.6 × 102 4 5.6 × 102 ± 87 6.7 × 102 ±39 6 1.3 × 102 ± 12 1.3 × 102 ± 12 8 22 ± 3 25 ± 5

The time course of log reduction of E. coli in 3.5 h field tests followed a linear trend. It did not show a tailing effect at the latter stages of the disinfection. This may be due to the low concentration of total suspended solids (TSS) in peat biofilter effluents.

The suspended solids in wastewater during biological treatment process can results in the formation of microbial bioflocs, consisting of microorganisms, extracellular polymeric substances (EPS), organic and inorganic colloidal particles (Sanin and

Vesilind, 1994; Urbain et al., 1993). Thus suspended solids in secondary treatment

37 effluents protect microbes against UV irradiation (Brahmi et al., 2010; Azimi et al.,

2012). This causes a tailing effect in the inactivation over time, which is further enhanced by the high TSS concentration and large size suspended solids. In onsite UV disinfection, Leverenz et al. (2007) found that the performance was maximized with

TSS less than 5 mg/L and turbidity less than 3 NTU. Biofilm based secondary treatment such as peat filters can efficiently remove suspended solids including bioflocs, thus attenuating the tailing effect and increasing its efficacy when using UV disinfection system.

The performance of UV disinfection can be quantified by using UV doses:

UV dose = IT (1)

where I = UV light intensity ( W/cm2), T = contact time (s)

The only contact of wastewater with the UV lamp is when it is circulated in the chamber. The average UV intensity for the storage tank is the lamp intensity multiplied by the volume ratio between UV chamber and storage tank. Thus the UV dose can be calculated as

UV dose = Vchamber/Vtank IT (2)

where and are the volumes of the UV chamber and storage tank, respectively

The UV chamber volume was 5.7 L and the batch volumes for the laboratory and field tests were 100 L and 3,785 L, respectively. The dose received in the laboratory tests for 20 min was 4.7 times higher if compared to the 3.5 h field test, which reached 180 mJ/cm2 and achieved a 3 log inactivation of E. coli (Table 6).

Table 6. The UV doses received in the laboratory and field tests (average light intensity 2.0 × 103 μW/cm2)

Time UV dose Test (min) (mJ/cm2) Laboratory 5 34.2 10 68.4

15 102.6

20 136.8

Field 60 10.8 120 21.7

180 32.5

210 38.0

Complete inactivation of E. coli and Salmonella, reported by de Nardi et al.

(2011), was attained at 11 and 32 mJ/cm2 UV doses, respectively. This dosage was insufficient in inactivating E. coli using a UV batch irradiator in the present study. In the field test conditions, the intensity of the UV light varied greatly with the temperature and the fouling on the lamp over time. The actual UV dosage received by wastewater may not be as much as the calculated irradiation listed in Table 6. Based on the results for the UV lamp intensity and chamber size, the maximum volume of the batch reactor is 570 L, if the daily running time is set at 2 h.

The laboratory tests showed that the dose of UV radiation was sufficient for disinfection of the 100 L batch in 20 min, whereas in the field scale 3.5 h of 39 disinfection of the 3,785 L tank was insufficient to inactivate E. coli to acceptable levels. When extended to 8 h disinfection, E. coli was inactivated to below 30

CFU/100 mL. In the field scale, the circulation time of the wastewater in the batch reactor should, therefore, be extended to account for the larger volume.

Field test under three different flow rate scenarios

From equation 1, the UV dose received in the UV batch irradiator depends on the lamp specification and the contact time. The circulation flow rate does not directly affect the performance of the UV unit. Because of the lack of a residual disinfectant, UV irradiation is ineffective against bacterial growth in biofilm formation in the effluent chamber. Moreover, low flow rates may accumulate suspended solids as sediments in the chamber.

0.45 A 0.4

0.35

0.3 0.25 0.2

Survival ratio Survival 0.15 0.1 0.05 0 0.5 2 3 Flow rate (l/s)

2.5 B

1.5

2.5 hr log reductionlog hr2.5 0.5

0 0.5 2 3 Flow rate (l/s)

Three different circulation flow rates (0.5, 2.0 and 3.0 L/s) were applied and compared in the UV batch irradiator. The viable counts of E. coli before and after a 2.5 h disinfection period are listed in Table 7. The initial E. coli counts in biofilter effluent ranged from 3.0 × 102 to 3.4 × 104 CFU/100 mL. The mean log reductions after 2.5 h of disinfection were 0.7, 0.8 and 1.4 for 0.5, 2.0, and 3.0 L/s flow rates, respectively.

Figure 8 shows the average survival ratios and log reductions at the three circulation flow rates. No significant difference was observed among these three treatments (p =

0.335), consistent with equation 1 in which circulation flow rate is not a parameter.

However, a somewhat increasing trend in E. coli inactivation was observed between 0.5

L/s and 3.0 L/s flow rates. This difference may reflect better mixing wastewater in the large circulation tank. The increased flow rate through the reactor may also allow fewer solids to accumulate in the chamber and thus slow down the fouling of the UV lamp.

Table 7. E. coli viable counts (± SD) after UV exposure (27.1 mJ/cm2) with different circulation flow rates in the field batch disinfection

E. coli counts in the effluent (CFU/100 mL) Flow rate (L/s) before UV radiation after 2.5 h UV radiation 0.5 1.4 × 103 ± 1.4 × 102 3.2 × 102 ± 12 1.2 × 104 ± 8.2 × 102 9.8 × 102 ± 2.8 × 102

3.2 × 103 ± 4.5 × 102 7.5 × 102 ± 54

3.9 × 103 ± 2.2 × 102 1.3 × 103 ± 2.5 × 102

2 3.0 × 102 ± 71 2.0 × 10 ± 4 3.8 × 102 ± 42 1.4 × 102 ± 15

6.5 × 102 ± 82 2.2 × 102 ± 23

2.7 × 103 ± 1.4 × 102 1.5 × 102 ± 12

3 1.5 × 104 ± 1.6 × 103 2.6 × 102 ± 26 1.4 × 104 ± 2.1 × 103 2.4 × 102 ± 24

3.4 × 104 ± 4.5 × 102 8.0 × 102 ± 82

1.4 × 104 ± 8.3 × 102 4.0 × 103 ± 3.4 × 102

UV batch irradiator versus continuous UV flow-through reactor

The UV batch irradiator was tested in the field for nine months. Major maintenance was conducted in the fifth month, including lamp cleaning and checking of wiring and the control box (Figure 9). The system operated during this entire period uninterrupted with over 600 h of lamp lighting period.

Figure 9. Maintenance of UV disinfection unit: cleaning up lamp and checking wiring

One major issue in using a traditional flow-through UV reactor is that the UV lamp stays on all the time, regardless of effluent flow at low or high levels. The extended use of the UV lamp significant reduces the energy efficiency of the system.

Continuous use also increases the temperature of the lamp, increasing the possibility of lamp burn out. In the UV batch irradiator, the lamp is only turned on during operation

(3.5 h daily in field test), which increases the reliability of UV irradiation and the maintenance interval. After 5 month of operation, only slight fouling was observed on the sleeve covering the UV lamp (Figure 10). The higher flow rate being applied in the

UV batch reactor also assisted in prevention of biofouling on lamps. The normal operation flow rate if setup as conventional flow-through system is between 0.1-0.3 L/s and for our batch UV system test, the minimum flow rate was at 0.5 L/s and up to 3.0

L/s. The higher flow rate generates more turbulence and shearing force inside the UV unit, making biofouling more difficult to attach.

Figure 10. UV lamp after 5 months of operation at 3 L/s recirculation flow rate, no obvious fouling was formed along the lamps and covering sleeves

The power usage of the UV batch irradiator was comparable to the UV flow- through system because of the recirculation pump. With power of the lamp at 40 W, the flow-through system has a power consumption of 0.96 kWh on a daily basis as the system is operating continuously with UV lamp on for 24 h/day. For the UV batch system, however, the power consumption is proportional to the preset daily running time. With circulation pump operating at 400 W, the power usage of UV batch system is identical to UV flow-through system at approximately 130 min. However, this goal was not able to be achieved due to the large volume of the tested recirculation tank, which required over 8 h daily to reduce E. coli to a safe level. This comparison suggested a recirculation tank that has smaller size and provides sufficient water

45 amount for daily irrigation is preferred to make this system more economic. Another possible modification is to increase the UV dose by connecting more UV unit in series and using lamps with higher intensity.

Conclusion

This study evaluated batch UV disinfection as an alternative to conventional flow-through UV disinfection systems. The results showed that batch UV system was consistent and overcame the short-comings of flow-through systems. Batch treatment extended the lamp life-time by only operating for a few hours before irrigation. The energy use for both systems was comparable with the UV batch system running for only a short time daily. The UV batch disinfection system showed its potential to replace conventional flow-through systems with proper recirculation tank sizing and maintenance.

The efficacy of UV-disinfection was tested by monitoring survival of E. coli.

For the UV batch system in the field test, E. coli counts were reduced by 1.4 logs after

2.5 h contact time, and 2.0 log after 5 h contact time. After 8 h of contact time, 3.1 log reduction was achieved. Three recirculation flow rates, 0.5, 2.0 and 3.0 L/s, inside the

UV chamber did not change UV dosage received. However, higher flow rate generated more turbulence in the UV chamber, allowing fewer solids to accumulate in the UV chamber and slowing down the fouling of the UV lamp.

The performance and reliability of UV batch disinfection system was monitored in the field. Using a peat biofilter as a secondary treatment lowered the level of

46 suspended solid flocs. The field test lasted for about nine months and the UV lamp life was over 600 h, the batch UV system proved to be reliable over many months.

Chapter 3: Onsite chlorine batch disinfection

Introduction

The goal of onsite wastewater disinfection after secondary treatment is to minimize public exposure to bacterial, viral, and protozoan pathogens. Chlorine has been used as a water and wastewater disinfectant due to its high effectiveness and relatively low cost. Chlorine tablets are typically used for chemical disinfection in residential wastewater treatment units (Babcock Jr. et al., 2011). However, these systems are prone to clogging, inadequate dissolution and retention time, and require tablet refills. Due to fluctuation in effluent flow, the amount of chlorine mixed into the effluent tends to vary greatly (Leverenz et al., 2007). This will lead to under-dose or over-dose of disinfectant, where under-dose may cause insufficient pathogen inactivation and over-dose may cause damage to plants if the wastewater is reused for irrigation and can be toxic to fish and other aquatic life if discharged. US EPA (2002) suggested that residual chlorine should be monitored when the disinfected wastewater is reused for example for irrigation because it can damage vegetation at >5 mg/L levels.

For household sewage treatment systems (HSTS), the general NPDES permit requires total residual chlorine in the effluent to be less than 0.038 mg/L (Ohio EPA, 2017).

Some commonly used chlorine disinfectants include chlorine gas (Cl2), chlorine dioxide

(ClO2), sodium hypochlorite (NaOCl) and calcium hypochlorite (Ca(OCl)2). However, 48 chlorine gas (Cl2) and chlorine dioxide (ClO2) are not considered appropriate for small facilities due to the hazards presented by storage, handling, and application of these chemicals (Leverenz et al., 2007).

Another chlorine disinfectant is sodium dichloroisocyanurate (NaDCC), which has been shown to be an effective antimicrobial agent and is a promising alternative to

NaOCl in household based water treatment (Clasen and Edmondson, 2006). Both sodium hypochlorite and NaDCC release free available chlorine (FAC) in the form of hypochlorous acid (HOCl) as shown below:

However, unlike NaOCl which releases all of its chlorine as FAC, NaDCC releases only approximately 50% of the chlorine as FAC, the remaining chlorinated isocyanurate serves as ―chlorine reservoir‖ for further FAC release when FAC is consumed in equilibrium (Bloomfield and Miles, 1979). This slow release mechanism is an advantage over NaOCl in onsite wastewater disinfection.

Many studies have shown that chlorination on wastewater can result in formation of toxic disinfection by-products (DBPs) such as the trihalomethanes

(THMs), haloacetic acids (HAAs), bromate, and chlorite (Bayo et al., 2009; Buth et al.,

2011; Watson et al., 2012). Formation of DBPs derives from reaction of residual chlorine with organic compounds in wastewater (Mounaouer and Abdennaceur, 2016).

While chlorine has high efficiency on microorganisms such as bacteria, some pathogens were found to be more resistant to chlorination, such as Cryptosporidium parvum and

Giardia lamblia) (Blatchley et al., 2012; Peiran et al., 2014).

With limitations of chlorination, one of the commonly used alternatives for wastewater disinfection is ultraviolet (UV) irradiation (Mounaouer and Abdennaceur,

2016). E. coli and other bacteria are very sensitive to UV light as well as other pathogens such as protozoa (de Nardi et al., 2011). The UV disinfection technology is relatively simple to install in small treatment plants and is easy to operate and maintain as compared to chemical disinfection.

In a typical onsite UV disinfection system a UV irradiator is connected to the outlet of secondary treatment as a flow-through system. In residential settings, the peak water usage, wastewater treatment and outflow may only last a few hours each day.

Typical flow patterns are weekday morning and evening water use. In a continuous flow system, the UV lamp is on at all times without considering fluctuations in water usage. Because these lamps are water cooled, this leads to lamp overheating and fouling because the illuminated lamps sit in non-flowing wastewater for several hours each day

(Leverenz et al., 2007). An excessive idle time for UV lamps significantly reduces their effective lamp life.

Using a batch disinfection system as an alternative may be effective in addressing some of the issues of flow-through systems such as under-dosing or over- dosing due to uneven flow, and reducing UV lamp fouling with reduced daily radiation period. Some laboratory scale research on chlorination in a batch system is conducted to allow for accurate dosage and contact time (Bohrerova and Linden, 2006; Buth et al.,

2011) for large wastewater treatment plants. However, few long-term field test results were found for onsite chlorination using batch system. For UV irradiators, many lab-

50 scale research studies have been undertaken using batch UV systems (Blume and Nies,

2004; Gibson et al., 2008; Naddeo et al., 2009), however, these have not been tested in the field.

The performance of onsite wastewater disinfection is mainly monitored by monitoring fecal indicator bacteria (FIB) (Sanders et al., 2013). Despite the debate in the appropriateness of the selected indicators and methods used to estimate the human health risk, total coliforms, fecal coliforms, Escherichia coli, fecal streptococci, or enterococci have been adopted by most states in the US to monitor potential pathogens in water (Sanders et al., 2013). In Ohio, E. coli was used as indicator in HSTS General

NPDES permit and the effluent standard was set at 410 CFU/100 mL (Ohio EPA, 2017).

The objectives of this study were to investigate the field performance and feasibility of batch chlorine disinfection for onsite wastewater reuse. Its efficiency and reliability was compared to UV batch disinfection system as well as conventional flow- through systems.

Methods

To test performances of both chlorine and UV batch systems in onsite wastewater disinfection, a field test was conducted treating onsite peat biofilter effluent for a three-bedroom, single-family house located in London, OH (39°57'44.4"N,

83°26'00.6"W). This onsite treatment system was installed and operated for a three- bedroom single-family house. The wastewater was first treated by a septic tank and then a peat biofilter (Premier Tech Environment model STC-650, Quebec, Canada).

The biofilter effluent accumulated in a 3785 L storage tank, which also served as the

51 tank for batch disinfection with either chlorine or UV. The batch of disinfected effluent was then discharged for irrigation. The diagrams of the complete field test setup for UV and chlorine batch system were demonstrated in Figure 3 (Chapter 2) and Figure 11, respectively. UV unit was first installed and tested for 9 months in this system. Then it was replaced by chlorine system, which operated for another 6 months.

Figure 11. Schematic diagram of the field test using chlorine batch disinfection system

Chlorine batch disinfection

For chlorine batch disinfection, granular sodium dichloroisocyanurate (NaDCC) was used as disinfectant. In the field tests, the disinfectant was designed to be dosed once per day before scheduled daily irrigation. To meet these requirements, a special chlorine dispenser was designed, shown in Figure 12. An outdoor animal feeder was attached to the bottom of a 19 L bucket which holds capsules of granular NaDCC. The capsulized NaDCC was easily dosed by the animal feeder with a built-in timer to control the time and duration of each dose. The chlorine dispenser was placed above the storage tank in the access riser and dropped capsules directly into the storage tank.

Figure 12. Configuration of NaDCC capsules dispenser

To better understand the characteristic of NaDCC in wastewater, a laboratory test on chlorination of effluent samples from the peat biofilter was conducted. Five dosages were tested by adding 10 mg, 20 mg, 30 mg, 40 mg and 50 mg of NaDCC into

1 L of wastewater. Free and total chlorine and pH were measured at 10 minutes, 24 hours and 48 hours contact time. Sodium bisulfite (NaHSO3) was added at each time to stop the chlorine oxidation reactions.

The field test of chlorine disinfection was conducted continuously for 26 weeks.

Chlorine capsules were dosed 1 hour prior to scheduled irrigation, the dropped capsules usually dissolved within 15 minutes. The dosing frequency was set as once per day with assumption that the residual free chlorine could inactivate pathogens in storage tank for

24 hours. The free chlorine and E. coli concentration was measured on samples after 15

53 minutes of dosing, and 24 hours later just before the next dosing on weekly basis. For the first 13 weeks of test, each dose dropped approximately 50 grams of NaDCC capsules and capsules were refilled every two weeks. Then for the remaining 13 weeks, the dosage was reduced to 25 grams of NaDCC capsules and refilled every 4 weeks.

Wastewater samples

The wastewater samples were collected from both chlorine and UV batch reactors in the field test. The onsite wastewater from the single-family house received primary and secondary treatment with a 2-compartment septic tank and peat biofilter, respectively. The biofilter effluent was tested for BOD5 and turbidity. Samples were collected from the batch chlorine and UV irradiator attached to the effluent storage tank, before and after disinfection. Samples for E. coli were collected every 15 minutes for chlorine disinfection and every 30 minutes during UV operation, and were stored at

4 °C and analyzed within 12 h.

Analytical techniques

E. coli were enumerated using EPA Method 1603 m-TEC agar (US EPA 2006).

Water samples were filtered using 0.45 µm pore size membrane filters (47 mm). For E. coli, the membrane filters were incubated on Difco modified m-TEC agar at 35 °C for two hours and then at 44.5 °C for 18-20 h. E. coli colony forming units (CFU) were counted based on magenta colored colonies. All samples were processed in three duplicates. Measurement of BOD5 was conducted according to Standard Methods

(1998). Turbidity was measured using a Model 2100Q Portable Turbidimeter (Hach

Company, Loveland, CO).

For chlorine batch reactor, the residual chlorine was also measured after dosing and before the next dosing 24 hours later. Free chlorine concentrations were verified with the N,Ndiethyl-p-phenylenediamine (DPD) colorimetric method (Rice et al., 2012).

The DPD is oxidized by free chlorine, resulting in a solution with color intensity proportional to the free chlorine concentration. The absorption of the solution was quantified using the Free Chlorine Pocket Colorimeter (Hach Company, Loveland, CO).

Results and Discussion

Mechanism of NaDCC chlorination

The release of free and total chlorine of NaDCC dissolving in peat biofilter effluent in the laboratory test is presented in Table 8. The total chlorine residual showed a trend of first order decay from 10 minutes of contact time to 48 hours. In average 52 (±

0.04) % of NaDCC dissolved was detected as total chlorine after 10 minutes of contact time. For all five NaDCC concentrations, free chlorine residual after 24 hours were not decreased compared to 10 minutes level, suggesting that the slow free chlorine release mechanism allowed for at least 24 hours of effectiveness in disinfection. When applying

30 mg, 40 mg and 50 mg of NaDCC in wastewater, the free chlorine after 24 hours were even increased from 10 minutes, which was helpful in chlorine batch systems in order to reduce the dosing frequency. Significant decay of free chlorine residual was observed for all levels of NaDCC treatment after 48 hours, indicating that the effectiveness of NaDCC disinfectant may not be able to stay for over two days.

Table 8. Free, total residual chlorine (mg/L) and pH (± SD) in 48 hours test of NaDCC reaction with peat biofilter treated wastewater under different NaDCC doses

Time 10mg/L NaDCC 20mg/L NaDCC (min) free Cl total Cl pH free Cl total Cl pH 0.5 ± 5.5 ± 8.29 ± 1.8 ± 10.4 ± 8.09 ± 10 0.08 0.29 0.057 0.29 0.86 0.028 0.5 ± 1.2 ± 8.51 ± 1.9 ± 4.1 ± 8.28 ± 1440 0.14 0.08 0.075 0.08 0.16 0.042 0.1 ± 0.6 ± 8.78 ± 0.7 ± 8.78 ± 2880 3 ± 0.22 0.08 0.22 0.067 0.14 0.042

Time 30 mg/L NaDCC 40 mg/L NaDCC (min) free Cl total Cl pH free Cl total Cl pH 2.4 ± 17.2 ± 7.81 ± 2.7 ± 18.0 ± 7.84 ± 10 0.14 0.28 0.028 0.08 0.16 0.041 3.5 ± 7.3 ± 8.19 ± 3.8 ± 7.1 ± 8.31 ± 1440 0.29 0.45 0.042 0.12 0.12 0.022 1.3 ± 5.3 ± 8.52 ± 1.4 ± 4.4 ± 8.64 ± 2880 0.08 0.22 0.028 0.05 0.08 0.025

Time 50 mg/L NaDCC (min) free Cl total Cl pH 3.6 ± 25.5 ± 7.75 ± 10 0.16 0.49 0.024 4.2 ± 9.7 ± 8.34 ± 1440 0.24 0.73 0.073 1.8 ± 5.4 ± 8.64 ± 2880 0.08 0.08 0.049

Figure 13 shows the CT value for up to 48 hours. These results suggested that one dose of chlorine per day before daily field irrigation might be a good strategy for batch chlorination using NaDCC as disinfectant. For conventional flow-through tablet chlorination, NaDCC is usually not approved for use in wastewater disinfection due to its slow release mechanism of chlorine in limited contact time for flow-through systems

(Weaver and Lesikar, 2017). However, when using batch chlorination with NaDCC, the contact time can be easily controlled and extended, which avoids under-dose and allows for a long lasting effect of pathogen inactivation within the storage tank. Thus the slow release mechanism of NaDCC has its advantage over hypochlorite in batch chlorine systems which can reduce the dosing frequency by maintain stable free chlorine level and effectiveness in pathogen inactivation. The reduced dosage and dosing frequency reduces the cost of chlorine agents and refill frequency in field operation.

7000

6000

5000

4000 10 min 3000 1 day

CT (mg min/L) (mg CT 2 days 2000

1000

0 10mg/L 20mg/L 30 mg/L 40 mg/L 50 mg/L NaDCC dosage (mg/L)

Figure 13. CT value (± SD) of NaDCC in 5 different dosages, free chlorine concentrations were measured at 10 min, 24 h and 48 h after chlorination

Chlorine batch system in field test

The average BOD and turbidity in wastewater effluent after peat biofilter were

12.8 (± 2.2) mg/L and 6.4 (± 1.1) NTU respectively. Average E. coli count in effluent was 26830 (± 1350) CFU/100 mL. After 15 minutes of contact with chlorine the E. coli level dropped to 1 CFU/100 mL, equaling 4.4 log removals, shown in Table 9. E. coli stayed inactivated for 24 hours until next dosing. From all samples collected after 15 minutes and 24 hours of contact throughout the testing period, E. coli were not detected except for two samples which were measured at week 20 and week 23. In both cases, the capsules were partially jammed at the dispenser outlet, which led to insufficient dosing before the day of test and sampling. In these two cases, at 15 minutes E. coli counts were 20 and 38 CFU/100 mL respectively, but after 24 hours E. coli were not detected in wastewater.

It is known that E. coli is very sensitive to free chlorine exposure. Even at initial free chlorine concentrations of 0.20 mg/L, >99.9% inactivation was achieved within

0.50 minutes for all initial E. coli concentrations (Helbling and VanBriesen, 2007). At

50 g/day dosing rate, the free chlorine residuals at 15 minutes and 24 hours were 1.03

(± 0.15) and 0.4 (± 0.06) mg/L, and 0.62 (± 0.22) and 0.4 (± 0.18) mg/L at 25 g/day dosing rate. 15 minutes of contact time resulted in free chlorine ranged from 0.3 mg/L to 1.2 mg/L, and after 24 hours contact time it ranged from 0.2 to 0.6 mg/L. The variance of free chlorine in batch chlorinator is much smaller than traditional flow- through tablet chlorinator. Leverenz et al. (2007) tested free chlorine dose in tablet system and it ranged from 0.14 mg/L to 390 mg/L. Thus by switching to batch

58 chlorinator, the primary reliability issue of tablet chlorinator could be solved, which is the changing dissolution rate and resulting under dose and over dose of chlorine.

Table 9. Total, free chlorine concentration and E. coli count of chlorination after 15 minutes and 24 hours contact time in the field test, average E. coli count in effluent was 26830 (± 1350) CFU/100 mL before chlorination

Total Chlorine Free Chlorine E. coli Test Date (mg/L) (mg/L) (CFU/100mL) 15 min 24 h 15 min 24 h 15 min 24 h 10/28/2015 0.7 0.6 0.3 0.2 <1 <1 11/05/2015 0.6 0.4 0.4 0.2 1 <1 11/17/2015 2.2 1.8 0.8 0.6 <1 <1 11/23/2015 2 1.8 0.8 0.6 1 <1 12/01/2015 1.8 1.6 0.6 0.4 <1 <1 12/16/2015 2.4 1.6 0.8 0.4 <1 <1 02/02/2015 1.8 1.2 0.8 0.4 <1 <1 02/10/2015 2.2 1.2 1 0.5 <1 <1 03/01/2015 2.8 1.2 1.2 0.4 <1 <1 03/15/2016 2.6 1.2 1 0.4 20 <1 03/29/2016 2.4 1 1.2 0.4 38 <1 04/18/2016 3.2 0.8 1 0.3 <1 <1

Another issue of chlorine tablet is the aging of tablet on top of the chlorinator stack which lead to decay of active chlorine in the tablet (Leverenz et al., 2007). By sealing granular chlorine into gelatin capsules, the efficacy of chlorine can be better preserved for longer period to ensure consistent disinfection effectiveness for the same 59 dosage. However, the outlet of the dispenser allowed moist to get into the chlorine container making some capsules sticky and partially blocked the outlet, resulting in under dose of chlorine. Thus regular maintenance is necessary to clean up sticky capsules at the outlet and replace desiccant in the container.

Chlorine batch reactor vs. UV batch irradiator

The performance of UV batch disinfection system was evaluated in Chapter 2.

From these test and monitoring, chlorine and UV batch systems both proved to be practical in onsite wastewater disinfection. The chlorine disinfection showed high effectiveness on E. coli inactivation. Although the UV batch system did not meet full inactivation of E. coli, this can be improved by either extending the circulation time or adding another UV unit in series. The chlorine residual and slow release characteristic of NaDCC allows free chlorine to be maintained at an effective level that inactivates pathogens for a long time. The UV disinfection only takes place when the system is circulating wastewater from storage tank through UV chamber and no effects of disinfection afterwards.

The chlorine batch system was operated well for more than 6 months, with biweekly or monthly refill of the capsule dispenser and checking for clogging or broken capsules. Frequent dispenser refills were required due to the high moisture content inside the storage tank. Even though the chlorine dispenser was about 2 m above the water surface, the bucket was well-sealed on top and desiccant was put in along with

NaDCC capsules, the moisture in the tank could still find its pathway through the dispenser outlet channel and get into the bucket. This dampened some of the capsules in

60 dispenser especially at the feeder outlet, which resulted in capsules sticking together and broken capsules. The metal parts in the dispenser were also exposed to a corrosive environment. Thus for current design of chlorine dispenser, at least a monthly refill frequency is recommended. The UV batch irradiator was tested in the field for 9 months. Major maintenance was conducted in the 5th month, including lamp cleaning and checking of wiring and the control box. The system operated during this entire period uninterrupted.

Both batch disinfection systems overcame some of the primary issues of traditional chlorine tablet and UV flow-through system. Chlorine batch dispenser could precisely control the dosage of chlorine added into wastewater on each batch, making the residual chlorine much more consistent. Using capsules to seal granular chlorine slowed down the aging of chlorine, which keeps each chlorine capsule at a high efficacy even after a long time. The UV batch irradiator significantly reduced the daily lamp hours by transforming wastewater from one time flow-through to batch recirculation, which also stopped lamp fouling due to overheating of lamp that submerged in wastewater at very low flow rate. The UV batch system extended lamp life, enhanced lamp performance and reduced the maintenance frequency compared to

UV flow-through systems.

Running a chlorine dispenser is simple. The cost for building the dispenser using an animal feeder was less than $50. If dosing at 15g/day, the daily cost for

NaDCC is about $0.18. For the UV batch system, the initial equipment cost is $500.

With circulation pump operating at 400 W and UV lamp at 40 W, the cost of running

61 the system for one hour is $0.05 cents. Thus for the same operation cost to chlorine dispenser, UV batch system can operate 3.75 hours per day. However with 3.75 hours of operation the pathogen reduction was still not identical to chlorine disinfection, showing the limitation of single unit UV batch system. Batch UV disinfection is better suited for small storage tanks in terms of cost when compared with chlorine.

Conclusion

The efficacy and reliability of disinfection of biofilter secondary effluents was evaluated by testing chlorine and UV batch reactors. Both systems were found to be effective and reliable in onsite wastewater disinfection based on E. coli inactivation.

The chlorine batch disinfection system showed a 4.4-log reduction down to non-detect within 15 minutes after chlorine addition, and this effect remained over 24 hours. With a simple dispenser and timer, NaDCC capsules can be accurately dosed based on treatment requirement, and the disinfection in wastewater can be easily monitored using residual free chlorine concentration as an indicator.

Compared to the chlorine batch system, the UV batch disinfection system was less effective, with only 1-log reduction of E. coli after 3 hours of operation for the same batch volume. However, it overcame some of the shortcomings of chlorine dispenser such as the need for frequent refill and maintenance. The UV batch system can be operated maintenance-free for at least 6 months. Thus, for a smaller size batch storage tank (<1000 L), a batch UV batch disinfection unit is well suited. For tanks larger than 1000 L, chlorine batch system is preferred by taking advantage of its higher effectiveness in less contact time.

Chapter 4: Temperature management in sand biofilters for enhanced ammonia removal in cold weather

Introduction

Meat and poultry processing wastewater contains relatively high levels of fat and suspended solids as well as high biochemical oxygen demand (BOD5) and chemical oxygen demand (COD) (Moodie and GreenWeld, 1978). Hence common biological treatment systems such as activated sludge, oxidation ditches, and sequencing batch reactors (SBRs) require corresponding pretreatment for meat and poultry processing wastewater (Kang et al., 2007). A sustainable and low-cost alternative of treating such wastewater is to use biofilm based technology such as sand filters (Kauppinen et al., 2014;

US EPA, 1978; Widrig et al., 1996).

Biofilm based technology for wastewater treatment has simple structure, low cost in construction and maintenance, and relatively high efficiency of treatment (Moodie and

GreenWeld, 1978). The design of the biofilter can be modified for various types of wastewater. High-rate trickling filters had been proved to be effective in handling fluctuations in wastewater quality that occur with slaughterhouse wastes (Moodie and

GreenWeld, 1978). BOD5 removal of over 75% was achieved (Hopwood, 1977). Liu et al.

(1999) compared single and multilayer sand bioreactors for dairy wastewater treatment.

The two-layer sand biofilter achieved higher BOD5 removal (85%) than single layer

63 filters (76%). Another key factor affecting sand biofilter performance is loading rate or dosing rate. Different hydraulic loading rates were tested by Kang et al. (2007). At loading rates below 132 L/m2/day, excellent performance was achieved with over 94% of

TOC and 98% of BOD5 removal.

+ Ammonium-nitrogen (NH4 -N) is a primary concern in discharged water, because it has several adverse effects when released into the environment, which include fertilization-driven oxygen depletion in aquatic ecosystems and toxicity to aquatic life (US EPA, 2013). The specific criteria for ammonia are pH and temperature dependent. Although the standard considers cold weather condition which permits higher ammonia level as chronic criteria, it still poses great pressure upon wastewater treatment facilities, especially for months of November and March in years with cold winter. In

Ohio, the standard is 12.6 mg/L 30-day average total ammonia-nitrogen at pH 7.0, T = 0-

10°C during the months of December to February and 2.2 mg/L during the months of

March to November (Ohio EPA, 2014). Thus, it is important to address the cold temperature ammonia removal issue in wastewater treatment facilities.

Nitrogen removal in wastewater treatment is achieved by nitrification coupled with denitrification process, involving two kinds of bacteria—nitrifiers and denitrifiers

(Choi et al., 2008). In the nitrification process, ammonia (NH3) is first oxidized

- aerobically to nitrite (NO2 ) by ammonia-oxidizing bacteria (AOB). Nitrite is then

- converted to nitrate (NO3 ) by nitrite oxidizing bacteria (NOB), which is reduced to nitrogen gas eventually by denitrifiers (Rodriguez-Caballero et al., 2012).

Biofilm systems allow sufficient biomass retention time for nitrification, thus they are frequently used for nitrogen removal (Okabe et al., 1996). Kang et al. (2007) observed high BOD5 removal efficiency in turkey-processing wastewater treatment using multi-layered sand biofilters, which were constructed with fine sand and topped with a layer of coarse sand and a pea gravel cap. However, the efficiency of nitrogen removal and the effect of variables remain unknown for this specific multi-layered sand biofilter.

Wastewater treatment facilities are facing potential ammonia oxidation process failure through biological treatment during winter temperatures (Gilbert et al., 2014;

Hendrickx et al., 2012; Hu et al., 2013). The incomplete ammonia removal in cold weather is because the oxidation rate of ammonia by AOB is significantly affected by temperature in the first stage of nitrification, or partial nitrification (PN) (Kim et al., 2008;

Zhang et al., 2016).

The dynamics of AOB community structure are important and necessary for improving stable ammonia degradation and its removal in biofilm treatment (Zhang et al.,

2009), which can be altered under in adverse conditions (Park et al., 2009). AOB community can alter in low temperature (Siripong and Rittmann, 2007), and temperature is believed to be the most significant factor affecting the AOB community structure among other environmental variables (Park et al., 2009). The low temperature decreases the biomass of the biofilm as well as changes the AOB community (Park et al., 2008).

Other researchers reported that although changes occurred in the AOB community, the nitrification process remained stable (Layton et al., 2005). Choi et al. (2008) tested a pilot-scale aerated submerged biofilm reactor during cold months when average

65 temperature dropped to approximately 6°C. The biofilm continued to have high ammonia removal efficiency, reducing from 25 mg/L ammonium-N to zero within 40 to 48 hours.

Rodriguez-Caballero et al. (2012) related this maintained nitrification activity at low temperature to a shift of the AOB community composition, from a Nitrosomonas oligotropha-dominated community to a mixed community including also Nitrosomonas ureae-like ammonia oxidizers, in response to the change of ammonia-N and organic loading. Some researchers suggested inhibition of the AOB activity under high carbon:nitrogen (Michaud et al., 2006). The relationship between AOB community and heterotrophic bacteria is complex and many other factors are likely to be involved (Racz et al., 2010).

A greenhouse is an ideal solution for providing a more controlled and moderate environment in cold weather. The ground and underground soil can be passively heated by absorbing solar radiation. Ghosal et al. (2004) reported that the temperature at various depths inside greenhouse was on average 7-9 °C and 3-6 °C higher than the bare soil surface for daily and monthly variations, respectively. The heat exchange pattern with ground in a greenhouse changes throughout the year. During the winter and early spring, heat is transferred mostly from the greenhouse to the ground (Nawalany et al., 2014), helping to maintain higher average ground temperature than outside.

Few studies related to thermal performance of the ground in a greenhouse have been published to date (Nawalany et al., 2014). Temperature fluctuations are most significant in the top layer of the ground. At the depth of around 10 m, ground temperature is close to the annual average temperature of air in the given area (Popiel et

66 al., 2001). The distribution of ground temperature was reported to be higher in the central zone of the greenhouse and lower alongside the side and end wall zones (Al-Kayssi et al.,

2002; Tong et al., 2009).

The influence of temperature on the nitrification process is important for the design and operation of the attached growth reactor. In activated sludge reactor, the van’t

Hoff-Arrhenius equation has been used to estimate temperature impact on nitrification rate (Antoniou et al., 1990). In biofilm reactors, however, the impact of temperature change is poorly understood because the process is also influenced by other variables

(Mendez et al., 1995). No previous study has applied or tested greenhouse covered sand biofilter system in ammonia removal, which may be a simple and economic solution for restoring nitrification and nutrient removal in cold weather. A greenhouse cover may provide a good working environment for maintenance and can be used for ornamental plants propagation.

The objectives of this study were (1) to determine the impact of a greenhouse to the sand biofilter temperature profile; (2) to evaluate and compare the efficiency of nutrient removal of sand biofilters in cold weather under different test scenarios; and (3) to understand the possible relationship between temperature and biofilter ammonia removal efficiency.

Methods

The experiment examined multi-layered sand biofilters treating turkey processing wastewater at Whitewater Processing, Inc., Harrison, Ohio. A greenhouse and Styrofoam insulation board covered sand biofilter was constructed to increase biofilter temperature

67 during these cold periods. As another treatment to sand biofilters, polyethylene films were used to cover the biofilter surface. These two treatments were compared with sand biofilters with no treatment as a control in terms of their nutrient removal performance in cold weather, especially for ammonia removal.

Sand biofilters

The small sand biofilter constructed for greenhouse cover test had a dimension of 6 m × 12 m. The biofilter had top layer at ground level, each with a 45 cm (18 in.) layer of fine sand overlaid with 15 cm (6 in.) of coarse sand. Another 15 cm (6 in.) layer of pea gravel was covered over coarse sand layer. The fine sand was supported by

5 cm (2 in.) of pea gravel to facilitate drainage. The drainage zone on the bottom of sand biofilter used perforated drainage pipe surrounded by washed gravel. The fine sand had an effective size of 0.3 mm and a uniformity coefficient of 4.0. The coarse sand had an effective size of 2.4 mm and a uniformity coefficient of 1.3. The pea gravel had an effective size of 3.8 mm and a uniformity coefficient of 1.7.

A full-scale turkey processing wastewater treatment plant was adjacent to the small test greenhouse/biofilter system. It consisted of 12 biofilters in the dimension of

25 m × 54 m. Two biofilters were selected for this test. Filter 1 was fully covered with polyethylene films two weeks prior to the beginning of the test. Filter 2 had no extra treatment and served as the control. The remaining 10 biofilters were also monitored during the test as a reference for the control biofilter.

Greenhouse

The small sand biofilter was covered with a high tunnel structure greenhouse

(W×L×H = 9 m × 18 m × 4.3 m). The greenhouse was covered with double- polyethylene inflated sheets and both end walls were constructed with pressure treated studs and plywood. One of the two end walls had one 2 m × 1 m entry door and two

0.76 m × 0.76 m ventilation fans, the other had two 0.76 m × 0.76 m ventilation windows. No active heating or cooling was applied during the experiment period.

During the experiment, the small sand biofilter in greenhouse were also fully covered with Styrofoam insulation board to reduce heat transfer from/to the atmosphere in the greenhouse.

Biofilter testing and sampling

The 12 full-scale sand biofilters effluent were monitored from 07/11/13 to

04/28/16. Grab samples of turkey processing wastewater were collected from the discharge pipe in a storage pond following the grease trap. Sand biofilter effluent samples were collected from the effluent discharge pipe. Samples were collected weekly and tested for CBOD, TSS, nitrate, phosphorus and ammonia in Q Laboratories,

Cincinnati, OH. These data were used as a reference to learn the performance of sand biofilters and seasonal trends and variations.

The biofilter temperature treatment test was conducted from 02/15/2017 to

05/15/2017. Before the experiment, the sand biofilters received low hydraulic loading rate of 100 L/m2/day. During the testing, the hydraulic loading rate increased to 200

L/m2/day, which was close to the upper limit of the proper working range of the biofilters (Kang et al., 2007). The purpose was to operate the biofilters at their full

69 capacity during the test and better differentiate the treatments. Biofilter effluent samples were taken on a weekly basis during the day shift from effluent pipes when it did not rain. On each sampling day, effluent samples were collected at the beginning, middle and end of the day shift as a good representation of daily average. All samples were stored in refrigerator at 4 °C and tested in the laboratory within 24 hours after sampling.

Data measurement

Samples were prepared for each measurement from three tested biofilters. BOD5 and pH were measured using standard methods (Rice et al., 2012). The nutrient analysis for ammonia-N, nitrate-N and total phosphorus (TP) were conducted using the Hach Kit, with Salicylate Method (Method 8155), Cadmium Reduction Method (Method 8039) and Molybdovanadate with Acid Persulfate Digestion Method (Method 10127) respectively. All these test methods were approved by American Public Health

Association (Rice et al., 2012). A Hach colorimeter was calibrated with standard nutrient solutions before each measurement.

During the experiment, temperature measurements were taken inside biofilters with digital thermo sensors. Thermocouples were installed at 0.05, 0.25, 0.5 and 0.75 m depths of the sand biofilters in each measurement line at the center of each biofilter.

Temperature was recorded every hour and results were stored by HOBOTM 4-Channel

External Data Loggers (Onset Computer Corp, Bourne, MA).

Data analysis

Analysis of variance (ANOVA) screened for the significance of greenhouse cover for ammonia-N, nitrate-N and TP removal by comparing data collected from both greenhouse covered and control sand biofilter. Graphs plotting the relationship between temperature and ammonia removal were examined to reveal any patterns. Regression models were built and validated if typical linear or nonlinear pattern could be observed.

Temperature measurements along the vertical and horizontal measurement planes are presented in a graphic form. Thermal distribution throughout the cold period was established for both greenhouse covered and control biofilters. The significance of greenhouse cover without active heating and cooling to mean ground temperature was verified by ANOVA using JMP 11.

Results and Discussion

Historical data of biofilter performance

The three-year weekly monitoring of the newly constructed sand biofilters provided a comprehensive overview of the treatment performance of turkey processing wastewater. Table 10 listed the CBOD, TSS and nutrients in the sand biofilter effluent.

CBOD and TSS both achieved 99% removal after sand filtration. Average total phosphorus, ammonia and nitrate were 5.6 ± 1.8 mg/L, 1.9 ± 1.6 mg/L and 45.7 ± 24.2 mg/L respectively. These data were also separated into cold months’ period (November through April) and warm months’ period (May through October). When switching from warm to cold period, ammonia concentration in biofilter effluent increased from 1.0 ±

1.0 mg/L to 2.6 ± 1.6 mg/L, while nitrate decreased from 57.6 ± 26.6 mg/L to 35.2 ±

15.6 mg/L. Both ammonia and nitrate showed significant differences between cold and

71 warm periods (P<0.05). However, no significant difference was observed for phosphorus in these two different time periods.

Table 10. Annual, cold and warm months sand biofilter effluent monitoring

Warm Parameter All Year Cold Months* Months** Average (± SD) 6.1 ± 3.4 6.0 ± 2.5 6.2 ± 4.9 CBOD Maximum 24 14 24 (mg/L) Minimum 3 3 3 Average (± SD) 3.7 ± 3.7 3.4 ±2.7 4.2 ±4.9 TSS (mg/L) Maximum 26 20 26 Minimum 1 1 1 Average (± SD) 5.6 ± 1.8 6.1 ± 1.7 5.0 ± 1.8 TP (mg/L) Maximum 11.8 9.6 11.8 Minimum 0.87 2.62 0.87 Average (± SD) 1.9 ± 1.6 2.6 ±1.6 1.0 ±1.0 Ammonia Maximum 13.2 13.2 4.3 (mg/L) Minimum 0.1 0.1 0.1 Average (± SD) 45.7 ± 24.2 35.2 ± 15.6 57.6 ±26.6 Nitrate Maximum 165 75.6 165 (mg/L) Minimum 4.02 4.02 22.2 *Cold months refer to November to April **Warm months refer to May to October

Ammonia 5 A

Effluent (mg/L) Effluent 2

0 6/17/13 11/14/13 4/13/14 9/10/14 2/7/15 7/7/15 12/4/15 5/2/16 Date (MM/DD/YY)

180 Nitrate 160 B 140

120 100 80

Effluent (mg/L) Effluent 60 40 20 0 6/17/13 11/14/13 4/13/14 9/10/14 2/7/15 7/7/15 12/4/15 5/2/16 Date (MM/DD/YY)

Continued

Figure 14. Time series observation of nutrient in full scale biofilter effluent from 07/11/13 to 04/28/16: A, ammonia concentration; B, nitrate concentration; C, total phosphorus concentration 73

Figure 14 continued

14 Phosphorus 12 C

6 Effluent (mg/L) Effluent 4

0 6/17/13 11/14/13 4/13/14 9/10/14 2/7/15 7/7/15 12/4/15 5/2/16 Date (MM/DD/YY)

Figure 14 showed the biofilter effluent throughout a three-year monitoring period. The ammonia concentration demonstrated some seasonal patterns. A sudden decline of the ammonia concentration was observed during the end of April to beginning of May, and it started to increase in November (Figure 14A). From Figure 14

B, the nitrate concentration was at peak level from May to July, implying some negative correlation with the ammonia concentration. These results suggest an enhanced nitrification process when in transition from cold to warm periods, which may also indicate a possible shifting of AOB community components as some other researcher discovered (Rodriguez-Caballero et al., 2012). No obvious patterns could be observed for phosphorus concentration.

Biofilter temperature

The biofilter temperatures for three treatments were monitored from February through April of the test. Figure 15A, 15B and 15C show the biofilter temperature at

0.05 m, 0.25 m, 0.5 m and 0.75 m depth. The cyclical temperature waves were dampened for all three biofilters as the depth increased. The biofilter with a plastic cover reduced the variation of subsurface temperature, especially at 0.05 m and 0.25 m depth. Figure 16 compared the average biofilter subsurface temperature for the entire test period. The overall average temperature of greenhouse covered biofilter, plastic covered biofilter and control biofilter were 18.2 °C , 18.2 °C and 14.1 °C respectively.

Student t test showed a significant increase of biofilter temperature with greenhouse cover and plastic cover treatment (p<0.05), while no significant difference was found between these two treatments. This result indicated that both treatments raised biofilter temperature due to greenhouse effect and reduced heat losses to the ambient environment.

35 0.75 m 30 A 0.5 m

0.25 m

) 25 C

° 0.05 m 20

Temperature( 10

0 02/15 02/23 03/03 03/11 03/20 03/28 04/05 04/14 04/22 04/30 05/09 Date (MM/DD)

35 B 0.75 m

30 0.5 m

) 0.25 m C 25 ° 0.05 m 20

15 Temperature(

5 02/15 02/23 03/03 03/11 03/20 03/28 04/05 04/14 04/22 04/30 05/09 Date (MM/DD)

continued

Figure 15 continued

24 0.75 m 22 C

0.5 m

)

C 20 0.25 m ° 0.05 m 18

16 Temperature(

12 02/15 02/23 03/03 03/12 03/20 03/28 04/06 04/14 04/22 05/01 05/09 Date (MM/DD)

° 15 greenhouse 10 plastic cover

control Temperature ( Temperature 5

0 0.75 m 0.5 m 0.25 m 0 m Depth (m)

Figure 16. Average temperature (± SD) of three tested biofilter from 02/15/2017 to 05/15/2017

On one of the coldest days in the experiment, the hourly variation and average of biofilter and air temperature on 02/16/2017 are shown in Figures 17 and 18. The temperature variation close to the biofilter surface level are more related to air and surface temperature change while at depth of 0.5 m and 0.75 m, the diurnal changes of temperature became close to constant, shown in Figures 17A-C. Ghosal et al. (2004) reported that the temperature is constant throughout the day at 0.3 m depth or deeper.

For the greenhouse treatment, it was evident to observe the effect of greenhouse and insulation cover on biofilters at the depth of 0.25 m. The cyclic V-shape dives of the temperature were due to the scheduled dosing of wastewater from distribution pipes at that depth, which had lower temperature than biofilter sand. In a typical cold day, the mean temperature differences of these three treatments were increased, especially at depths of 0.05 m and 0.25 m.

14.5

)

C 14 °

13.5

13 0.75 m Temperature ( Temperature 0.5 m 12.5 A 0.25 m 0.05 m 12 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 Time of day (hour)

continued

Figure 17 continued

) C

° 13

11 0.75 m

Temperature ( Temperature 0.5 m 10 0.25 m 9 B 0.05 m 8 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 Time of day (hour)

) 10

C ° 8

6 0.75 m

Temperature ( Temperature 4 0.5 m 2 C 0.25 m 0.05 m 0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 Time of day (hour)

continued 80

Figure 17 continued

) C

° 10

0 Temperature ( Temperature -5 outdoor air temp -10 D indoor air temp

-15 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 Time of day (hour)

20 18

C) 14

° 12 10 greenhouse 8 plastic cover control

Temperature ( Temperature 6 4 2 0 0.75 m 0.5 m 0.25 m 0 m Depth (m)

Figure 18. Average and standard deviation of three tested biofilter subsurface temperature on 02/16/2017

Nutrient removal in three biofilter treatments

The raw turkey processing wastewater contained high levels of BOD5 and TSS

(Table 11). After sand biofilter secondary treatment, average BOD5 was reduced to 8.1

± 8.0 mg/L and TSS was decreased to 10.3 ± 8.5 mg/L, achieving 99.3% and 98.3% removal respectively. No differences in effluent BOD5 and TSS were observed among the three different sand biofilter temperature treatments.

Table 11. Average water quality parameters raw turkey processing wastewater

Constituent Average

NH3 (mg/l) 9.14 Nitrate (mg/l) 1.67 Nitrite (mg/l) 0.34

BOD5 (mg/l) 1170 Fecal coliform / 100 mL 7.56 × 105 E. coli / mL 1.26 × 106 FOG (mg/l) 236.7 TSS (mg/l) 617.4 Total nitrogen (mg/l) 474 Total phosphorus (mg/l) 12.7

The mean total phosphorus (TP) concentration in raw wastewater was 12.73 ±

5.43 mg/L. Figures 19 and 20C show the effluent TP concentrations after three tests of biofilter treatment. The results showed no significant difference (p = 0.49, 0.50 and

0.99 pairwise) among the three treatments. The mean removal of TP was 14.7%, indicating ineffective removal of TP through sand biofilter secondary treatment. This is not unexpected since the removal of phosphorus usually requires tertiary treatments which are specialized in TP removal such as enhanced biological phosphorus removal

(EBPR) and methods involve the use of algae and cyanobacteria for phosphorus removal (Cai et al., 2013; Rawat et al., 2011; Suka ov et al., 2015).

Greenhouse Cover 20

Plastic Cover

No Cover (Control) 15

10 Concentration (mg/L) Concentration 5

0 Ammonia Nitrate Total Phosphorus Wastewater nutrient

Figure 19. Average nutrient concentration (± SD) of effluent from three test sand biofilters

40 A Greenhouse Cover 35 Plastic Cover 30 No Cover 25

15 Concontration (mg/L) Concontration 10

0 02/17 03/03 03/16 04/01 04/11 04/29 05/08 Date (MM/DD)

30 B Greenhouse Cover Plastic Cover 25 No Cover 20

Concontration (mg/L) Concontration 10

0 02/17 03/03 03/16 04/01 04/11 04/29 05/08 Date (MM/DD)

Continued

Figure 20. Nutrient concentrations in three tested biofilter effluent A, ammonia; B, nitrate; C, total phosphorus

Figure 20 continued

14 C

6 Greenhouse Cover

Concontration (mg/L) Concontration 4 Plastic Cover No Cover 2

0 02/17 03/03 03/16 04/01 04/11 04/29 05/08 Date (MM/DD)

As showed in Table 10 and Figure 19, most of the nitrogen in raw turkey wastewater was in organic form since ammonia, nitrate and nitrite accounted for a small fraction of the total nitrogen. The aerobic sand biofilters provided organic carbon and partial organic nitrogen removal as well as ammonia oxidation. The mean and time series concentration of ammonia and nitrate can be found in Figure 19 and 20A, 20B. A

Student t-test of the mean nitrate concentration only showed significant difference between plastic and no cover treatment among three treatments (p = 0.012). The same wastewater was applied to all three treatments at the same loading rate, thus the higher nitrate concentration in the effluent may indicate less limitation in the nitrification process.

Both greenhouse and plastic covered biofilter experienced failure of ammonia removal compared to the biofilters with no cover. The effluent ammonia level from the plastic covered biofilter was significantly higher than the other two treatments in the first three months of the test and then decreased dramatically after April 21. Whereas in the greenhouse covered biofilter, high effluent ammonia removal was maintained until the beginning of April. Starting in April, the effluent ammonia increased to as high as

38.2 mg/L, indicating failure of the partial nitrification, as shown in Figure 20A. Going into the experiment, the plastic covered greenhouse biofilter received no rest and was loaded at a full loading rate of 100 L/m2/day for over a year, whereas the no cover biofilter was rested for 30% of the time over the last year. The poor performance of ammonia removal in the plastic covered greenhouse in late winter/early spring may be due to the continuous operation of the biofilter. For all 12 biofilters in the treatment plant, large variation of ammonia removal between each biofilter was observed. Table

12 lists the ammonia concentrations for all 12 biofilters, measured by the plant operator during the test period. The concentrations averaged at 4.5 mg/L. This might be a better representation of the average performance of biofilters in the low temperature period.

Another possible reason for poor ammonia removal in the plastic covered biofilter and greenhouse covered biofilter is the plastic film and thermal foam that completely covered the filter surface. The covers might hinder oxygen transfer close to the surface where the high strength wastewater was applied. This might limit the AOB bioactivity and affect the partial nitrification process. As no microbial analysis was conducted

86 during this test, the difference of AOB and NOB communities in the three tested biofilters remains unknown.

Zhang and others (2016) found biofilm systems had limitation in treating high ammonia wastewater compared to activated sludge system. The biofilm depth resulted in an oxygen and substrate gradient, leading to higher biodiversity of AOB preferred to different conditions (Zhang et al., 2016). Another explanation for the sudden increase is episodic ammonia release due to change of season or weather. All 12 biofilters had experienced periodic ammonia increase of different time and duration, more frequently in winter. The biofilms in biofilters may get more ―stressed‖ in cold weather, causing lower nitrification activity.

Table 12. Effluent ammonia concentration for all 12 full scale biofilters during test period

Ammonia Ammonia Filter Filter (mg/L) (mg/L) 1 13.5 ± 3.7 7 2.6 ± 0.8 2 0.4 ± 0.2 8 22.1 ± 3.8 3 6.0 ± 0.3 9 4.6 ± 0.8 4 1.2 ± 0.6 10 0.8 ± 0.3 5 3.4 ± 0.6 11 25.0 ± 2.3 6 16.0 ± 4.1 12 0.8 ± 0.2

Relationship between biofilter temperature and ammonia removal

The results of biofilter effluent ammonia in Figure 19 and 20A show that both greenhouse cover and plastic cover treatments failed to provide stable ammonia removal. The reason for system failure in the ammonia removal in greenhouse covered biofilter is unknown. Although some researchers have found the direct relationship between temperature and AOB bioactivity (Kim et al., 2008; Zhang et al., 2016), the ammonia removal of biofilters may be affected by other factors as well in addition to the temperature effect. During the test period, a few temperature drops occurred in February and March, which did not show an impact in ammonia removal in all three biofilters.

Thus short term cold periods did not affect the ammonia removal efficiency of biofilters.

Some researchers have not found a direct relationship between nutrient removal and cold temperature, suggesting that nutrient removal is rather resistant to temperature changes

(Christopherson et al., 2005; Kauppinen et al., 2014; Williamson, 2010). Another explanation is that the temperature during the test period was not low enough to reach a critical point where the AOB activity becomes a limiting factor. Thus the impact of low temperature to ammonia removal remains unknown since the two temperature treatments may also change other factors affecting the efficiency of ammonia removal.

Conclusion

The three-year monitoring of sand biofilters proved that loss of ammonia removal efficiency occurred in cold months from December through March. This study introduced two approaches to warm sand biofilters: greenhouse cover and plastic cover. Both treatment on sand biofilters resulted in average biofilter temperature 4.1 °C higher than

88 biofilter with no cover. However, ammonia removal was not improved with either treatment. Instead, both biofilters experienced ammonia removal degradation during the test. This might be related to the treatment provided or episodic ammonia release from biofilter effluent. Although historical data showed lower ammonia removal in winter seasons, a direct relationship was not found between temperature and biofilter ammonia removal efficiency in the tests. Other factors such as the condition of the biofilter, the oxygen level within the filter, loading rate and maintenance frequency might also have to be considered.

Chapter 5: Development of hydroponic floating bed system for wastewater advanced nutrient removal

Introduction

Treated wastewater often contains excessive nutrients that may have significant impact on ecosystem (Cao et al., 2011; Tylova-Munzarova et al., 2005). Too much nitrogen and phosphorus in the water causes algae bloom, endangering aquatic life and may also produce toxins in water. In the United States, the complete numeric nutrient criteria are still under development, but some states have already established partial N and/or P criteria in lakes, rivers and estuaries (US EPA, 2017).

For wastewater treatment, many advanced treatment technologies provide enhanced nutrient removal. Constructed wetlands (CWs) are one type of wastewater renovation system to remove nutrients and other pollutants (Chen et al., 2009; Zurita et al., 2011). However, the high construction costs and large land requirement have limited their application in intensive wastewater treatment systems (Zurita et al., 2011).

Furthermore, plants are a source of nutrients in CWs, and their role can switch from nutrient uptake to consistent leaching (Hatt et al., 2009). Rapid and substantial decomposition and release of organic matter were found in wetland plants after a growing season (Chen et al., 2009). Thus whole plant removal or harvesting shoot and root

90 biomass is necessary which is often difficult and expensive in large scale CWs (Chen et al., 2009).

Wastewater used as a nutrient solution in hydroponic systems is one alternative in remediation of wastewater pollutants in a land-saving and cost-effective manner (Li et al.,

2010). In general, hydroponic vegetation systems can be categorized in two types: floating hydroponic root mats on the water surface and non-floating hydroponic root mat filters with roots touching the rooting-proof bottom of the water body (Chen et al., 2016).

Both systems have been used to treat poultry processing wastewater (Todd et al., 2003), domestic wastewater (Mietto et al. 2013; Saeed et al. 2014) and storm water runoff

(Chang et al. 2013; Wang et al. 2014; 2015). In addition, the produced crops can be further used as animal feedstock or be processed into biogas, biofertilizer and biomaterial, which bring potential economic returns (Li et al., 2007).

Perennial ryegrass (Lolium perenne) is a macrophyte commonly used in vertical flow CWs (Cao et al., 2011; Chang et al., 2004; Ren et al., 2016). This species has been proven to be highly productive and establishes easier and more quickly than most of other long-lived pasture grass varieties (Cao et al., 2011). It also has strong capability in absorption of nitrogen and phosphorus from wastewater and produces high-quality forage for livestock (Matheson and Sukias, 2010; Ren et al., 2016). One typical hydroponic system for ryegrass and other wheat grass is non-floating hydroponic root mat filters used as fodder systems for animal feeding (Chen et al., 2016). This system requires large quantity of grass seeds to form a thick layer of root mat and needs to be harvested in a short period such as 7-10 days. Thus this system is not suitable for hydroponic systems

91 focusing on wastewater nutrient removal. An ideal hydroponic system for this purpose should allow grass to grow a longer period of time and a fully developed root mat can be more efficient in nutrient uptake. Li et al. (2010) introduced a floating bed system employing rye grass, freshwater clams and biofilm carrier. This system was complicated and expensive, making it impractical in many applications. A simple and effective hydroponic floating system is still needed for small scale wastewater nutrient removal.

The objective of this study were (1) to develop a grass hydroponic floating bed system that was simple to build and operate, could support plants throughout the whole propagation period from seed germination to harvest, and was tolerant to various environmental conditions; (2) to improve different components and system designs in order to optimize the hydroponic floating bed system performance; and (3) to evaluate factors that may have influence on hydroponic float system performance and plant propagation.

Methods

The hydroponic system development process initiated with design of hydroponic floats. Different prototypes of floats were evaluated in different test environments, including laboratory, pilot plant scale greenhouse and full size high tunnel greenhouse.

During the development stage, the performance of the hydroponic floats along with the whole system was evaluated. The design of prototypes and hydroponic benches were also modified during this process to overcome any shortcomings discovered from the previous design. For all stages of the test, perennial ryegrass (Lolium perenne) was selected as the

92 plant of propagation. For all tests of hydroponic systems, a 2-day hydraulic retention time (HRT) was adopted as recommended by Xu et al. (2014).

Type I float in a pilot greenhouse

Type I floats was made of a perforated plastic plates supported by 6 cm wide

Styrofoam along the sides. On every 2 cm of the plastic plate, a 1.2 cm diameter hole was drilled, creating an effective area of 28 cm × 28 cm for each float. The test to with rye grass was conducted in a pilot-scale greenhouse built close to sand biofilter treatment system at Whitewater Processing, Inc., Harrison, Ohio (Figure 21). Three plastic storage boxes with dimension of 115 cm × 50 cm × 16 cm were used as container for hydroponic floats. Each storage box held two 40 cm × 40 cm type I floats. During the test, a piece of woven burlap was placed on the effective area of each float, and grass seeds were sown directly on the burlap. Before grass seeds germination, tap water filled the hydroponic bench and was replaced every two days until the sowed seeds start to grow shoots and establish a stable root system. Two different seed densities, 200 g/m2 as high seed density (HSD), and 100 g/m2 as low seed density (LSD), were compared.

After the two-week germination, treated turkey processing wastewater from the discharge pipe of a sand biofilter was applied. The performance of type I float and hydroponic system was monitored for 60 days.

Figure 21. Type I float and pilot-scale hydroponic system in uncontrolled pilot greenhouse (W×L×H = 4 m × 8 m × 2.5 m)

Type II, III and IV floats in laboratory hydroponic bed

A large hydroponic bed (W×L×H = 2.4 m × 1.2 m × 15 cm) was built in the laboratory to test different types of hydroponic float design for rye grass. Three prototype hydroponic floats were tested (Figure 22). The type II float was a modification from the type I float, which used the same Styrofoam in larger dimension

(W×L = 55 cm × 60 cm). The effective growing area was covered by plastic hardware cloth (W×L = 45 cm × 50 cm) instead of the perforated plastic in type I float. The type

III float was constructed using a perforated Lexan™ UV resistant polycarbonate sheet

(W×L = 60 cm × 115 cm) with a loop of PVC pipe underneath as support. Extra support was necessary when using this float due to the heavy weight of the polycarbonate sheet.

The type IV float used 4 cm thick Styrofoam insulation panels with 1.5 cm diameter

94 drilled holes across the whole board at even spacing of 5 cm. All three types of floats used woven burlap as support media for grass seeds.

A continued

Figure 22. Prototypes of hydroponic floats: A, type II; B, type III; C, type IV

Figure 22 continued

To compare the performances of all three types floats in the laboratory scale hydroponic bed, both germination and growing stages were tested under the same

96 conditions in the laboratory environment with artificial nutrient solution and lighting.

During the germination stage, type II floats were placed directly in the hydroponic bench filled with tap water, the flexible supporting material allowed seeds to make contact with water. For the type III and IV floats, a fine mist of tap water was sprayed onto the seeds multiple times throughout the day to promote germination. After 10 days of seed germination, a nutrient solution was added into the hydroponic bench. The solution was made with Maxigro hydroponic nutrient as presented in Table 13. The growing stage lasted for 30 days.

Table 13. Nutrient content of Maxigro hydroponic solution used in laboratory hydroponic test, the nutrient solution was made as 2.5 g/L of Maxigro granule

Parameter Concentration (%) Ammonium Nitrogen 1.5 Nitrate Nitrogen 8.5

P2O5 5.0

K2O 14.0 Ca 6.0 Mg 2.0 S 3.0 Fe 0.12 Mn 0.05

Type IV and V floats in a high tunnel greenhouse

The full scale field test in a high tunnel greenhouse adopted type IV floats from the laboratory test at the beginning stage. They were shortly replaced by type V floats after shortcomings of type IV float were found in the full scale hydroponic bench. As an upgrade from type IV float, type V floats filled the holes with coco chips (coconut husk) as media upon perforation, in an effort to help seed fixation and germination.

The hydroponic benches (W×L×H = 1.2 m × 6 m × 1.2 m) were placed on top of a small underground sand biofilter (W×L = 6 m × 12 m), both covered with a simple structure high tunnel greenhouse (W×L×H = 9 m × 18 m × 4.3 m), shown in Figure 23.

The greenhouse was covered with double-polyethylene inflated sheets and both end walls was constructed with pressure treated studs and plywood. No active heating or cooling of indoor air was available for this greenhouse. One of the two end walls had one 2 m × 1 m entry door and two 0.76 m × 0.76 m ventilation fans, the other had two

0.76 m × 0.76 m ventilation windows. The hydroponic bench could hold 10 type IV or

V floats (0.6 m × 1.2 m × 4 cm) crafted with polystyrene foam. Each float was covered with burlap and then seeded with ryegrass seeds. A germination shelf was built in the lower portion of the hydroponic bench and equipped with a mister system to provide moisture for seed germination. The programmed mister pump delivered treated wastewater effluent from a 200 L storage tank to the grass seeds. The dosing frequency was programmed at once per hour during daytime and once every four hours over the night time, with 5 minute duration each time. After seed germinated and grass roots started to penetrate through the holes, floats were transferred up into the hydroponic

98 bench for root forcing and development of shoots. Temperature and humidity was monitored by digital temperature and humidity gauge during the test.

Figure 23. hydroponic benches integrated with sand biofilter and high tunnel greenhouse

Results and Discussion

Pilot greenhouse hydroponic system with type I float

The ryegrass seeds germinated in about two weeks with type I float in the pilot greenhouse. During the 60-day growing period, the temperature in the pilot greenhouse ranged from -2 to 48°C . The seeds did not uniformly germinate for both low and high seed densities floats. Areas on the burlap with no seed germination was soon covered with a thick layer of algae, while in the germinated areas grass shoot elongated and extended leaving no room for algae development. The worst case happened in one of

99 the high seed density floats, where 90% of the burlap area was covered with algae instead of grass, shown in Figure 24.

Figure 24. Failure of grass seeds germination and burlap was covered with algae

The high and low seeding density on burlap did not show significant difference in terms of grass shoot and root mass development after 30 days of growth, as shown in

Figure 25. The low density group had stronger grass root and stem compared to high density group. Once fully developed, ryegrass showed high tolerance in temperature variation. After 30 days of growth, the average temperature dropped from 7°C to below

0 °C, which did not pose any negative impact on the grass.

100

Figure 25. comparison of low (left) and high (right) seed density grass production after 30 days of seed germination

Throughout the test algae developed not only on the float burlap, but along the inside of the hydroponic boxes. Some algae also suspended in the wastewater making the treated effluent in green color, shown in Figure 26. Algae has been widely used in phytoremediation as an efficient pollutant removal media (Hultberg et al. 2013), special immobilized culture are often required to prevent algae particles mixing into the effluent (Shaker et al., 2015). The existence of algae generated two major problems for the wastewater hydroponic float system. One was the competition at germination stage with grass seeds, which might lead to unsuccessful germination of grass. The competition between grass seeds and algae during the grass germination determined the outcome of grass sprouting and development in the growing stage. The second issue is the algae mixed with the effluent. When released, the effluent would contain algae cells, increasing suspended solids in the effluent. The algal nutrients would also be released

101 back to the environment and encourage algae growth in the receiving water body. Thus the control of algae in hydroponic systems is essential for both hydroponic plant propagation and nutrient removal.

Figure 26. Algae development in hydroponic boxes, effluent showed greenish color due to abundant algae suspended

Comparison of type II, III and IV hydroponic floats in laboratory test

The type II, III and IV floats compared in the lab showed their advantages and disadvantages as supporting media for ryegrass germination and growth. Photos taken after two weeks of seed germination are shown in Figure 27. The type II float allowed seeds to directly germinate in the hydroponic bench with no additional watering onto seeds, providing convenience for seed germination compared to the other two types of

102 floats. The seeds planted started to sprout on Day 9. Germination time was similar using spray to wet the seeds for the type III and IV floats. However, some seeds on the type II float floated on top of the water and were detached from the burlap. This led to an uneven seeding density on the burlap and also a decreased germination ratio. The type III float achieved the best germination ratio and uniformity among the three designs due to its flat and rigid surface.

A continued

103

Figure 27 continued

In the early stage of grass growing, all three types of floats provided good support for root and shoot development. However, the type II float was the least 104 supportive as the grass mat grew and became heavier. Eventually the center of the float started to submerge. The submerged areas on the burlap showed less shoot and root density, which indicated its negative impact to plant cultivation. The type III float provided the best support for plant growth throughout the whole period with the best shoot and root development at the end of test. The shortcoming of the type III float was the cost and labor to craft it, if proper tools were unavailable. Also, the transparency of the float in the early stage of plant propagation may be susceptible for more algal growth. On the type IV float, ryegrass also achieved good root and shoot establishment.

Although the type III float proved to perform best in terms of seed germination, plant coverage and root/shoot development, it was not selected as a test float for the following full scale greenhouse test mainly because of the difficulty to drill holes on the polycarbonate sheet, especially in mass production. Instead, the type IV float was used for the next full scale greenhouse experiment as a simpler and more economical choice.

Type IV and V floats in a full scale hydroponic floating bed system

As the mister germination system delivered better germination outcomes, this approach was adopted in the full-scale hydroponic floating best system in the high tunnel greenhouse, as shown in Figure 28. The germination of grass seeds took 3 weeks due to the low indoor temperature when the test was conducted. The floats were moved up to the hydroponic bench when roots penetrated the float’s holes and could be seen from the bottom of the float.

105

Figure 28. Germination rack and mister system under hydroponic bed provided optimal germination results of ryegrass seeds

The type IV float test in hydroponic bench was conducted for 3 weeks and suspended early due to the mortality of grass from the center of the float. After one week of cultivation in the hydroponic bench, grass on all four tested floats started to wilt from the center and spread to the sides, as shown in Figure 29. The initial investigation indicated that most of the grass roots did not extend through the float holes but instead went sideways along the top of the floats with covered burlap. This severely restricted the acquisition of moisture from the wastewater. The burlap along the side of floats contained more moisture due to capillarity action, allowing the grass along the side to continue to grow.

106

Figure 29. Mortality of grass from the center of the float after one week cultivation due to failure of root penetration through the floats into hydroponic bed

To encourage the root establishment through the perforated holes and touch the water surface in the hydroponic bench, the type V float holes were filled with coco chips as a growing media to create a pathway for the grass root to the water and to provide support for the grass roots. The grass shoot and root establishment after two weeks in the hydroponic bench we shown in Figure 30. The modification to the type V float achieved successful root development and shoot elongation with the coco chips.

The algal development seen in the pilot greenhouse hydroponic boxes was not observed in this high tunnel greenhouse hydroponic system due to lower relative humidity as well as the complete block of solar radiation into wastewater by floats and benches. The treated effluent of wastewater at 2 HRT was clear with no green color, as seen in Figure

31. From the results, the type V float turned out to be a suitable amendment to prevent grass mortality in the hydroponic benches. Compared to floats designed by other

107 researchers (Bartucca et al., 2016; Li et al., 2010), type V float and its corresponding hydroponic bed offered a simple, cost-effective yet reliable solution for continuous production of ryegrass and wastewater nutrient removal.

Continued

Figure 30. Grass shoot (A) and root (B) development after two-week cultivation using type V float

108

Figure 30 continued

Figure 31. Wastewater in hydroponic bed, algae growth was eliminated due to complete blockage of light by using opaque floats

109

Factors affecting the performance of hydroponic systems

Strengthening the root mat development under hydroponic float is one of the main design parameters, since a well-established root mat often indicates optimized growing condition for plants as well as strong nutrient uptake capability. Rye grass could form dense root mat with proper hydroponic float design, such as type III and V float in our experiment. Some other indigenous aquatic plants often seen in wetlands may also be ideal for this hydroponic float system, such as species from the genera

Canna, Carex, Cyperus, Juncus, Phragmites and Typha (Kadlec and Wallace, 2009).

The microalgae development in wastewater hydroponic systems had potential risk to the treated effluent quality as well as grass production. In both the pilot greenhouse test and laboratory hydroponic bed test microalgae were observed along the side of hydroponic bed as biofilms and in wastewater as suspended solids. However, in the high tunnel greenhouse hydroponic test, no algae were found in wastewater and effluent. This was mainly because of the effective insulation of wastewater in hydroponic bed from sunlight by using type V hydroponic float and completed covered the water surface in hydroponic bed. This turned out to be a simple and effective way in controlling algae growth within hydroponic beds.

Although ryegrass has high tolerance to temperature variation, extreme high and low temperature can still pose negative impact on its production in hydroponic systems.

During the pilot greenhouse test with type I floats, temperature reached as high as 45-

48 °C for a few days, which resulted in yellow grass shoot and wilt of the whole plant in a few days. In type IV and V test in high tunnel greenhouse, high temperature led to

110 dry out of grass seeds during germination with mister system, which significantly reduced the germination ratio. A few days below 0 °C were also experienced during this germination when frost or ice crystal was observed on seeds and burlap, resulting in some areas with no seed germination on burlap. These results suggest that active heating or cooling may be occasionally needed in greenhouse hydroponic systems especially in germination stage to survive those temperature extremes.

Conclusion

The development process from pilot plant greenhouse hydroponic system to full scale high tunnel greenhouse floating bed system achieved improvement in overall performance of plant cultivation. Five prototypes of hydroponic floats were tested.

Type I through IV floats all showed their limitation in either laboratory or field test.

The type V float was made with perforated Styrofoam insulation panels and coco chips, which overcame most of the drawbacks in the other float designs, and addressed issues in seed germination and algae control. Algae development in hydroponic system may contribute to nutrient removal, but can inhibit grass growth and contaminate treated effluent. Using the type V floats in the full scale hydroponic bench blocked most of the solar radiation into the wastewater, which limited algal growth. The full-scale greenhouse hydroponic system provided a better growing environment for plants compared to uncontrolled pilot system. High germination ratios of seeds were achieved by using mister germination rack. The separation of germination and growth of plants into stacked hydroponic bed saved space in the greenhouse and enhanced productivity.

The developed full scale hydroponic floating bed system might be considered as an

111 economical and reliable solution for wastewater pollutant remediation and plant production.

112

Chapter 6: Wastewater nutrient removal in hydroponic/greenhouse system

Introduction

Nutrient pollution is one of the most widespread and challenging environmental problems in the United States. Over 20% of surface water impairment is related to nutrient discharge, including oxygen depletion, algal blooms, and toxicity (US EPA,

2009). It has been suggested that maintenance of stream water total phosphorus concentration at < 30 µg/L is necessary to limit benthic algal below nuisance levels of

100 mg/m2 (Dodds et al., 1997). Federal limits on phosphorus concentrations in fresh water have not been set (US EPA, 2017), but a discharge limit of 1 mg/L or lower may be achieved using existing technologies (Health Research Inc., 2014). Contamination of nutrients to groundwater can also be harmful. Nitrate is susceptible to leach through

- soil, causing elevated groundwater NO3 -N concentrations beneath and down gradient from the septic systems (Humphrey et al., 2016). High nitrate concentration in drinking water (>10 mg/L) can be hazardous to infants by causing disease known as methemoglobinemia (Sadeq et al., 2008).

Hydroponic root mats (HRMs) provide direct uptake of nutrients from the water through plant roots and also prevent algal reproduction by forming a mat of roots and thus creating shade. Commercial horticulture production systems often use the hydroponic floating bed system consisting of bench scale HRMs operated inside a 113 greenhouse. They can be used in controlled climatic conditions and plants can be easily harvested and processed. This approach can also be transferred into phytoremediation processes growing macrophytes in wastewater solutions. Primary or secondary treated wastewater can be used as a nutrient source for cultivating plants for horticultural purposes, or other non-food use such as forage or for renewable bioenergy production

(Darwish et al., 2007). With various field test conditions, design parameters and characteristics of wastewater, nutrient removal efficiencies have been reported for both floating and non-floating systems. TN removal efficiencies range from 3-92%, ammonia removal efficiencies from -46 to 94 %, and TP removal efficiencies from -5 to 88% (Chen et al., 2016).

Three major nitrogen removal mechanisms have been identified: microbial denitrification, plant nitrogen uptake and volatilization (Vaillant et al., 2003; Lee et al.,

2009; Li et al., 2012). Among these mechanisms, microbial denitrification plays a major role (Ge et al., 2007; Stottmeister et al., 2003). However, plants can have great impact on the N removal by affecting nitrification and denitrification intensity in the root zone (Stottmeister et al., 2003). Some studies suggest that the root zone of plants can release organic compounds providing a suitable environment for the nitrifying and denitrifying microorganisms (Sundaravadivel and Vigneswaran, 2001).

Perennial ryegrass (Lolium perenne) is a cool-season perennial gramineous terrestrial evergreen lawn plant (Ren et al., 2016). Bartucca et al. (2016) tested a laboratory scale ryegrass hydroponic floating system for wastewater nitrate removal

- and found almost complete removal of all NO3 added from the hydroponic solutions

114 with an initial concentration of 50 100, and 150 mg/L. Ren and others (2016) studied

+ removal efficiencies of NH4 -N and TP by cultivating L. perenne in a 10 cm hydroponic ditch (HD) and two constructed wetlands (CWs) in a cold region. During the two year test, the removal efficiencies were maintained around 82.8 %, 92.5 % and

+ 68.6 % in NH4 -N removal, and 27.7 %, 40.3 % and 42.3 % in TP removal in the 10 cm HD and the two hybrid CWs (Ren et al., 2016). These showed that perennial ryegrass has the potential to be adopted as a cultivated plant in HRMs. However, field test using hydroponic floating bed systems in greenhouses is still lacking to treat wastewater. The removal of N, P with rye grass in commercial greenhouse setting has not been studied. It is also poorly understood that whether the hydroponic floating bed system would provide efficient nutrient removal in real production, with vast variation of wastewater quality and environmental conditions.

The objective of this study was to investigate the N, P nutrient removal from secondary treated wastewater in a hydroponic floating bed system in greenhouse environment. Two test scenarios, small scale hydroponic bed and floats in pilot greenhouse and large scale hydroponic floating system in full scale greenhouse were compared in terms of their nutrient removal and grass productivity. Factors that may have influence on nutrient removal, such as seeding density and algae control were analyzed.

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Methods

Pilot test in an uncontrolled greenhouse environment

A small pilot greenhouse (W×L×H= 4 m × 8 m × 2.5 m) was built for field hydroponic system test. The wastewater used was collected from the discharge pipe of a pilot sand biofilter system at Whitewater Processing, Inc., Harrison, Ohio. The nutrient level in the wastewater fluctuated with the production activities within the turkey processing facility during the test period (Table 14). Three hydroponic benches were made with transparent boxes with dimension of 113.5 cm × 50.5 cm × 16 cm and placed inside the pilot greenhouse. Each hydroponic bench fit two floats on the top, providing support grass production. The square floats were made of perforated plastic plates surrounded by 6 cm wide Styrofoam along the edge. The effective area for plant growth in each float was 28 cm × 28 cm. Holes were drilled on the plastic plate with a diameter of 1.2 cm every 2 cm, allowing grass roots to go through and into the wastewater. A piece of woven burlap served as a seed substrate covered the top of the perforated plate growing area.

To start the hydroponic system, perennial ryegrass (Lolium perenne) seeds were directly sown on the wet burlap substrate. Each hydroponic bench was filled with 3 L of wastewater with a hydraulic retention time (HRT) of 2 days, suggested by Xu et al.

(2014). The corresponding hydraulic loading rate was 6 cm/day, which was a typical loading rate when treating domestic wastewater (Chen et al., 2016). An air pump was connected to submerged air stones to maintain a high dissolved oxygen level in the root zone. Water samples were taken before and after the hydroponic system treatment.

116

Table 14. Average pollutant concentrations of sand biofilter treated wastewater

Parameter Concentration (mg L-1)

BOD5 4.11 Total suspended solids (TSS) <5.00 Total nitrogen (TN) 64.7 Total phosphorus (TP) 13.5 Oil and grease <2.81 Nitrate 58.2 Nitrite 1.15 Ammonia 3.87

Two trials were conducted in the small pilot greenhouse. In the first trial, seed density was 110 g/m2. The ammonia concentration in wastewater ranged from 10 to 20 mg/L. In the second trial, two different seed densities were compared, 200 g/m2 as the high seed density (HSD) group, and 100 g/m2 as the low seed density (LSD) group, along with a control group with no seed sowed on floats. This was to investigate the algae development on the floats and their impact to nutrient removal in the hydroponic system. The ammonia concentration in wastewater was lower than that in the first trial, ranging from 0.05 to 8 mg/L. In both trials, the experiment was separated into two stages: germination stage and growing stage. In the germination stage, tap water was filled into the hydroponic bench and was replaced every two days until the sowed seeds started to grow shoots and established a stable root system, which took about 10 days.

117

After seed germination, the biofilter treated wastewater filled each bench to replace tap water in the growing stage and continued for 60 days.

Full-scale test in a high tunnel greenhouse

Two hydroponic benches (W×L×H= 1.2 m × 6 m × 1.2 m) were set on top of a small, underground sand biofilter (6 m × 12 m), all covered with a simple structure high tunnel greenhouse (W×L×H= 9 m × 18 m × 4.3 m). The greenhouse was covered with double-polyethylene inflated sheets and both end walls was constructed with pressure treated studs and plywood. Each of the end walls had two 0.76 m × 0.76 m ventilation fans. A heat recovering system was also installed in the greenhouse to harvest extra internal heat in cold weather and transfer the heat into the underground biofilter. The hydroponic bed on top of the bench could hold 10 floats (0.6 m × 1.2 m × 4 cm) crafted with polystyrene foam. The float perforations were filled with coco chips (coconut husk) as a growing media, which helped seed germination root fixation. To start the seed germination, seeds were applied directly onto the floats with 100 g/m2 seed density, and then covered with burlap. In one of the two hydroponic benches, a wooden rack was built underneath the hydroponic bed for seed germination (Figure 23). Above the germination rack, a mister system was attached under the hydroponic bed to provide moisture for germination. Grass seeds germinated in 2 weeks on the germination rack of hydroponic bench. Floats were then transferred into the hydroponic bed for nutrient removal. The other hydroponic bed held control group floats with no seeds applied.

Same as the pilot greenhouse test, 6 cm/day of hydraulic loading rate was applied. The nutrient removal test lasted for 30 days until the grass was ready for harvest.

118

Data Collection and Measurement

Water samples were collected every two days during the growing stage and tested within 2 hours after sampling. The nutrient analyses were conducted with methods using Hach kit, approved by American Public Health Association (Rice et al.,

2012). Specifically, USEPA PhosVer 3 (Ascorbic Acid) Method (Method 8048),

Salicylate Method (Method 8155) and Cadmium Reduction Method (Method 8039)

3- were used for measurement of o-phosphate-phosphorus (PO4 -P), ammonium-nitrogen

+ - (NH4 -N) and nitrate-nitrogen (NO3 -N), respectively. These three nutrients are usually the most abundant N and P nutrients in wastewater (Xu et al., 2014). The Hach colorimeters were calibrated monthly with standard nutrient solutions.

At the end of each test, all plant tissue samples were collected by separating grass shoot and root. The tissues were then dried until weight was constantly under

60°C for 48 h. Plant biomass samples were grounded and analyzed by the Service

Testing and Research Laboratory, Ohio Agricultural Research and Development Center,

Ohio State University, Wooster, Ohio. Nitrate-N concentrations in shoot and root samples were determined by a potentiometric method as described by Baker and Smith

(1969). Other trace elements were tested using dry ashing and acid dissolution method

(Mester and Sturgeon, 2003).

Statistical analysis

Analysis of variance (ANOVA) was conducted using JMP 11 software to analyze the data. Two experimental factors (nutrient concentration, seeding density)

119 were tested for the mass balance data and identify if the factors had significant effects on the nitrogen and phosphorus removal (p< 0.05).

Results and Discussion

Nutrient removal in pilot greenhouse

Figure 32 shows the nutrient concentration changes before and after 2 HRT hydroponic float system treatments during the first trial in the pilot greenhouse.

Throughout the whole growing stage with high ammonia levels (above 8 mg/l), the

- + average reduction of nitrate nitrogen (NO3 -N), ammonium nitrogen (NH4 -N) and o-

- phosphate-P (PO4 -P) after hydroponic float system treatment were 23.0% ± 10.8%,

65.5% ± 23.5% and 48.4% ± 19.9%, respectively. The ANOVA test showed significant treatment effects with hydroponic system in o-phosphate-P and ammonium-N concentration (p<0.01), and less significant effect on nitrate-N concentration

(p=0.0183). These results showed a less effective removal of nitrate compared to ammonium and phosphate.

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40 35 A

25 (mg/L) 20 Influent 15

Nitrate Effluent 10

0 0 10 20 30 40 50 60 days after germination

continued

Figure 32. Nutrient concentration before and after 2 HRT hydroponic float system - treatments in plant growing stage (after germination): A, nitrate nitrogen (NO3 -N); B, + 3- ammonium (NH4 -N); C, Phosphate (PO4 -P)

121

Figure 32 continued

25 B

15 Influent

10 Effluent Ammonium (mg/L) Ammonium

0 0 10 20 30 40 50 60 days after germination

25 C

Influent 10

Effluent Phosphate (mg/L) Phosphate 5

0 0 10 20 30 40 50 60 days after germination

122

Figure 33 showed the change of nutrient removal ratio over the growing stage during the first trial. Nitrate-N maintained at a low removal ratio throughout the whole period. Ammonium-N removal was low at the beginning and increased to peak level after 10 days. This may indicate that the grass uptake capability and nitrification microorganisms had not been fully established at this early stage of treatment. The o- phosphate-P removal was relatively high (60%) for the first 20 days, then dropped down to below 40% for the next 20 days, and increased back to 55% in the following

20 days. In the uncontrolled environment, from day 20 to day 40 in the growing stage, the ambient temperature increased by 5 °C, which led to an 8.5 °C average temperature increase in the greenhouse, the grass growth was inhibited based on observation of their shoot elongation and root development. Also browned and wilted shoots emerged during this period. As a result, both ammonium-N and o-phosphate-P removal dropped after day 20. It is obvious that the nutrient removal efficiency of the hydroponic systems was closely related to the condition of grass in the hydroponic floats, and the pilot greenhouse failed to provide protection to grass under high temperature conditions.

123

1 0.9 0.8

0.7

0.6 ammonium 0.5 nitrate 0.4

Removal ratio Removal phosphate 0.3 0.2 0.1 0 0 10 20 30 40 50 60 Days after germination

Figure 33. Change of nutrient removal ratio over plant growing stage

During the second trial, a biofilm of algae emerged on the empty burlap of the control floats after day 10 in the hydroponic bed. The nutrient concentrations before and after HSD, LSD and no grass (control) treatments are shown in Figure 34. All three groups show high removal of o-phosphate-P (69%) and ammonium-N (83%), but low in nitrate-N removal (26%), which was similar to nutrient removal in the first trial. The

ANOVA results showed that hydroponic treatment had significant effects on both o- phosphate-P (p<0.01) and ammonium-N (p<0.05) concentration. However, no significant effects on nitrate-N were observed in all three groups (p-value for HSD,

LSD and control were 0.2, 0.18 and 0.08 respectively). This was similar to the results in the first trial. Figure 34A shows a sharp decline of nitrate-N concentration in influent wastewater from day 10 to day 20 after seed germination and remained at lower level 124 for the remaining time of the test. The ammonium-N in influent remained very low for most time of the test. However, a sudden increase of ammonium-N concentration was observed in the influent at day 35, reaching peak level in about 10 days and dropped down towards the end of the test, shown in Figure 34B. This might be due to the variation of the production activities in the turkey processing facility as well as nitrification process variation in biofilter treatment, which was discussed with details in

Chapter 4. The nitrate-N removal ratio in this test was lower than those reported by some other researchers (Bartucca et al., 2016; Jie, 2014; Zheng et al., 2007), which ranged between 60 to 72%.

125

60 A

50 Influent 40 HSD LSD 30

Nitrate ( ( mg/L) Nitrate No grass 20

0 0 10 20 30 40 50 60 Days after germination

3.5

3 B

2.5 Influent 2 HSD

1.5 LSD No grass

Ammonium ( mg/L) ( Ammonium 1

0.5

0 0 10 20 30 40 50 60 Days after germination

continued Figure 34. Comparison of nutrient level before and after treatment of HSD, LSD and no - + 3- grass: A, nitrate nitrogen (NO3 -N); B, ammonium (NH4 -N); C, phosphate (PO4 -P)

126

Figure 34 continued

6 C

5 Influent 4 HSD LSD 3 No grass

Phosphate ( mg/L) ( Phosphate 2

0 0 10 20 30 40 50 60 Days after germination

The influent nutrient concentrations and removal ratio from HSD, LSD and control were compared in Figure 35. For nitrate-N, ammonium-N and o-phosphate-P, no significant difference in removal ratio was observed among HSD, LSD and control.

This showed that high seeding density may be helpful in faster germination, but does not lead to higher nutrient removal in growing stage. With higher seed density, the overall germination ratio will decrease, leaving behind more un-germinated seeds.

Some seeds started to rot and cultivated fungi on the burlap, which inhibited grass growth. In one of the HSD floats, the top left corner was completely infected by fungi on the burlap, with no germinated grass left. Thus in terms of grass production economics and nutrient removal, low seeding density seemed to be a better choice. The

127 nutrient removal in control group using empty burlap floats was mostly contributed by biofilms formed on burlap by algae and other microorganisms. This was not surprising since algae is known to have large requirement for N and P and can propagate rapidly with sufficient nutrient source, oxygen and light (Cai et al., 2013; Suka ov et al., 2015).

Hultberg et al. (2013) suggested the use of micro or macro algae for removal of excessive nutrients from various water sources, followed by a deliberate strategy to use the biomass produced, which would provide substantial benefits in terms of sustainability and food issues. When compared with HSD and LSD treatment, more algae were observed inside the control hydroponic bed without rye grass’s competition.

The competition between grass and algae on available nutrients in HSD and LSD hydroponic beds might pose challenges on grass production.

128

1 300 0.9 A

250 0.8 0.7 200 0.6 0.5 150 0.4 100 0.3

0.2 (mg/l) nitrate Influent Nitrate removal ratio removal Nitrate 50 0.1 0 0 0 2 5 7 9 12 14 17 20 22 24 27 29 31 34 37 39 41 44 50 56 59 Days

inflow HSD LSD No grass

1 14 0.9 B 0.8 12 0.7 10 0.6 8 0.5 0.4 6 0.3 4 0.2

2 (mg/l) ammonium Influent 0.1

Ammonia removal ratio removalAmmonia 0 0 0 2 5 7 9 12 14 17 20 22 24 27 29 31 34 37 39 41 44 50 56 59 Days

Inflow HSD LSD No grass

continued Figure 35. Comparison of wastewater influent and nutrient removal with treatment of - + HSD, LSD and no grass: A, nitrate nitrogen (NO3 -N); B, ammonium (NH4 -N); C, 3- phosphate (PO4 -P) 129

Figure 35 continued

1 10 0.9 C 9

0.8 8

0.7 7

0.6 6 0.5 5 0.4 4 0.3 3

0.2 2 Influent phophate (mg/l) (mg/l) phophate Influent 0.1 1

Phophaste removal ratio removal Phophaste 0 0 0 2 5 7 9 12 14 17 20 22 24 27 29 31 34 37 39 41 44 50 56 59 Days

Inflow HSD LSd No grass

Table 15 lists the nutrient contents in grass shoot and root after 60 day HSD and

LSD treatment. The nitrogen and phosphorus content in HSD and LSD grass was similar. Higher nitrogen content and phosphorus content were measured in grass shoot and root respectively. More calcium, magnesium and zinc were found in LSD root than

HSD root, indicating more accumulation of these nutrients by LSD roots in this test.

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Table 15. Nutrient contents of rye grass shoot and root after growing in wastewater for 60 days

N P K Ca Mg Na Zn Sample (µg/g) (µg/g) (µg/g) (µg/g) (µg/g) (µg/g) (µg/g) HSD shoot 48255 4960 39881 5416 2970 1653 74 HSD root 32738 5038 23692 11165 1516 6567 380 LSD shoot 49877 5273 41519 5043 3092 1855 228 LSD root 31732 6761 24016 23591 2059 6481 1817

Nutrient removal in full scale greenhouse

The results of nutrient removal in 30-day test in the full scale greenhouse are shown in Figure 36. The mean o-phosphate-P removal was 61.2% ± 26.5%, and mean nitrate-N and ammonium-N removal were 46.6% ± 15.0% and 63.3% ± 20.4%, respectively. In the first 8 days after germination, removal of all three nutrients was low and then significantly increased since day 8. This trend showed a healthy development of grass shoots above floats and root mat extension beneath floats after germination.

After day 20, the grass root mat was fully established, touching the bottom of the hydroponic bed. The nutrient removal also reached peak level during this time. The growing root mat increased direct uptake of nutrients from the wastewater and provided higher specific surface area for attachment and growth of other microorganisms

(Reinsel, 2014). In the control group, empty floats with bare burlap showed significantly less nutrient removal after day 8, with mean o-phosphate-P, nitrate-N and ammonium-N removal 21.4% ± 5.0%, 16.9% ± 8.1% and 23.2% ± 9.9%, respectively

(Figure 36). No algae development was found in both ryegrass and control hydroponic

131 beds. These results indicated that when algal growth was limited, the rye grass was responsible for the majority of the nutrient removal.

Compared to the nutrient removal in pilot plant hydroponic test, a higher nitrate-

N removal was achieved, the o-phosphate-P removal was identical, and the ammonium-

N removal ratio was lowered in the high tunnel greenhouse hydroponic beds. With the same hydraulic loading rate, the high tunnel hydroponic beds carried a much higher ammonium load with high influent ammonium, averaging at 31.8 (± 9.9) mg/L, whereas the concentration in pilot plant influent was only 0.7 (± 0.6) mg/L. This difference was mostly due to the seasonal variation of ammonia level in sand biofilter treated effluents and the different loading rates applied in pilot and full scale sand biofilters. Thus although the removal ratio was lower in high tunnel greenhouse beds, the ammonium-N concentration reduced was 21.5 (± 10.4) mg/L, much higher than pilot greenhouse systems (0.6 ± 0.6 mg/L). The average shoot length was 21.2 (± 11.3) cm in the high tunnel floats, where the pilot greenhouse floats only had 12.3 (± 7.8) cm.

Better grass production was observed in the high tunnel hydroponic beds compared to the pilot greenhouse, with higher shoot and root mat densities on floats. The optimized hydroponic beds and floats in high tunnel greenhouse prevented algal reproduction due to shading which enhanced both grass production and nutrient load reduction.

132

1 20 0.9 18 0.8 A 16 0.7 14 0.6 12 0.5 10 0.4 8 0.3 6

0.2 4 (mg/l) nitrate Influent Nitrate removal ratio removal Nitrate 0.1 2 0 0 0 2 4 6 8 10 12 14 16 18 20 22 24 26 28 30 32 Days

inflow Grass No grass

0.9 50

0.8 B 0.7 40 0.6 30 0.5 0.4 20 0.3

0.2 10 Ammonium removal ratio removal Ammonium 0.1 (mg/l) ammonium Influent 0 0 0 2 4 6 8 10 12 14 16 18 20 22 24 26 28 30 32 Days

Inflow Grass no grass

continued Figure 36. Comparison of nutrient removal in hydroponic floats with and without rye - + 3- grass: A, nitrate nitrogen (NO3 -N); B, Ammonium (NH4 -N); C, phosphate (PO4 -P)

133

Figure 36 continued

1 20

0.9 18

0.8 C 16 0.7 14 0.6 12 0.5 10 0.4 8 0.3 6

0.2 4

Influent phosphate (mg/l) phosphate Influent Phosphate removal ratio removal Phosphate 0.1 2 0 0 0 2 4 6 8 10 12 14 16 18 20 22 24 26 28 30 32 Days

Inflow Grass No grass

Factors affecting hydroponic systems nutrient removal performance

In the hydroponic tests, algae were observed to grow along with grass mostly in open areas in hydroponic bench. The existence of algae might be beneficial in terms of nutrient removal in hydroponic systems, but could compete with rye grass and inhibit their growth. Also algae could contaminate treated effluent suspended algae particles which degraded effluent quality and treatment efficiency. The high tunnel greenhouse hydroponic test used opaque floats to cover all hydroponic bed surface area, leaving no open water surface in hydroponic bench which offered no light for algae’s photosynthesis. This proved to be the simplest method in algae control.

134

Other than plant uptake, the nitrogen removal is considered to be caused by microbial nitrification/dentrification activities (Lee et al., 2009; Li et al., 2012). The existence of denitrification process can be proved from analysis of the plant tissue samples (Table 15), where ryegrass only kept less than 10% of total N removal after 60 days treatment at 2-day HRT. The nitrification process was also taking place to produce more oxidized form of nitrogen for denitrification. Lycklama (1963) found that ammonium inhibited the uptake and reduction of nitrate in rye grass. Minotti et al.

(1969) suggested that ammonium and to some extent the high acidity adjacent to the cellular boundary membranes caused by ammonium uptake in excess of nitrate uptake could result in alterations in membrane permeability, thereby restricting the capacity for nitrate absorption. The relationship of nitrate/ammonium ratio and ryegrass growth was

− + studied by Cao et al. (2011). The results suggested that an optimum NO3 /NH4 ratio was between 75/25 and 50/50.

− + The relationship of NO3 /NH4 ratio and nitrogen removal for the first trial from second trial in pilot greenhouse is shown in Figure 37. A slight increasing trend showed

− + that the nitrogen removal efficiency was positively correlated to NO3 /NH4 ratio as it

− + ranged from 10/90 to 50/50. However when NO3 /NH4 ratio was low in the full scale

− + high tunnel test, no linear relationship was found in NO3 /NH4 ratio and N removal, shown in Figure 38. Especially when toward later stage of the test, when grass roots and shoots were fully established, high ammonium level did not seem to inhibit plant N uptake.

135

0.9 0.8

0.7

0.6 0.5 0.4

N removal ratio removal N 0.3 R² = 0.739

0.2 0.1 0 0 0.2 0.4 0.6 0.8 1 1.2

- + NO3 /NH4 ratio

- + Figure 37. Relationship of the NO3 /NH4 ratio with nitrogen removal in the pilot greenhouse test

0.9 0.8

0.7

0.6 0.5 0.4

N removal ratio removal N 0.3

0.2 R² = 0.1922 0.1 0 0 0.05 0.1 0.15 0.2 0.25 0.3

- + NO3 /NH4 ratio

- + Figure 38. Relationship of NO3 /NH4 ratio with nitrogen removal in high tunnel greenhouse test

136

Conclusion

The study of hydroponic floating bed systems demonstrated that it is effective in removal of nutrients from the sand biofilter treated wastewater, especially the o- phosphate-P and ammonium-N. The removal ratio of ammonium-N, nitrate-N and o- phosphate-P were 61-69%, 26-47% and 63-83%, respectively. The hydroponic floating bed systems provided a simple and reliable solution to lower effluent ammonium-N and meet discharge permit standards. In addition, plants can be harvested from these systems and could be further employed as forage or compost material for a potential economic return. The small pilot greenhouse could not offer an ideal environment for plant growth throughout the whole season. Full scale greenhouse systems provided better growing environment, resulting in better grass production and nutrient removal. The combination of ryegrass and the floating bed system was simple, effective and reliable, which proved to be a promising technique in the remediation of water polluted by N and P.

137

Chapter 7: Summary and suggestions for future research

Summary of Problems

Fixed media biofilters have been commonly used as a sustainable and low-cost alternative for treatment of onsite wastewater. Biofilters are often made with one or multiple layers of sand or other media, providing aerobic biological treatment of wastewater. The sand biofilters can achieve high efficiency in removal of organic matter

(BOD) and suspended solids (SS), but limited removal in nutrients and pathogens. Thus potential public health risk exists if biofilter treated effluent is reused for agriculture or other purposes, and discharge of treated effluent into surface or ground water generates potential environmental risk.

For onsite reuse, conventional disinfection methods such as chlorine and UV flow-through systems are normally employed in minimizing health risk. The chlorine tablet disinfection has inconsistent performance due to changing dissolution rate of tablets, which leads both over dosing and under dosing. The UV disinfection performance is mainly limited by lamp fouling which greatly reduces UV light intensity.

For nutrient removal to meet discharge permits, the fixed media biofilters may not provide sufficient ammonia removal especially in cold weather. In these situations, advanced nutrient removal need to be considered by either improving nutrient removal

138 capability of biofilters or by introducing advanced nutrient removal processes as tertiary treatment.

Research summary

The research in this dissertation investigated solutions for safely reuse and discharge of onsite treated wastewater. The batch disinfection systems were introduced to as a novel approach to overcome shortcomings of conventional flow-through systems. By adding a few extra components to the flow through system, the performance of onsite disinfection became more consistent while still maintain the simplicity of the system and ease for maintenance.

For sand biofilter nutrient removal enhancement, covering the filter surface with greenhouse, thermal foam or plastic was effective in raising internal temperature of biofilters during cold weather, but this did not improve the ammonia removal of biofilters as observed. The results might suggest ventilation on top of biofilter surface is necessary for aerobic biofilm activity, thus adding covers on top of biofilter might not be an ideal solution.

Hydroponic floating bed system was tested for advanced nutrient removal from pilot to full scale. The results showed efficient nutrient removal from wastewater while the grass production could be used for forage or compost ingredients. When integrated with underground biofilters and greenhouse, the hydroponic floating bed system has proven to be a promising advanced nutrient removal technology when biofilter effluent cannot meet the nutrient standard in discharge permit.

139

Suggestions for Future Research

 The onsite UV and chlorine batch disinfection systems can be a promising

alternative to conventional flow through systems. However, improvement can

be made to enhance their performance and durability. For example, the

chlorine batch dispenser needs to be designed for more accurate dosing of

chlorine capsules. For UV batch systems, using a small recirculation tank

separated from the effluent storage tank saves recirculation time. The volume

of recirculation tank should be sufficient for a one time reuse quantity based

on the reuse schedule.

 Specific UV lamp and chlorine agent (such as tablet or granules) can be

developed to be more adaptive in batch disinfection systems and to achieve

better performance.

 For sand biofilter nutrient removal in cold weather, more field test and

monitoring is needed to find out cause of decreased nitrification in biofilters.

Temperature and other factors such as dissolved oxygen in wastewater,

hydraulic loading rate may have contribution in this process.

 Weekly testing may not be frequent enough to capture episodic ammonia

releases. In future research, more frequent or even continuous monitoring data

of sand biofilter ammonia removal is suggested to find any patterns and to

help the in diagnosing system failure.

 Year-round testing of hydroponic floating bed system is suggested to test the

adaptability and reliability of the system through all seasons. To minimize

140

seasonal change of plants production, active heating and cooling in the

germination area of the greenhouse is recommended.

 More plant species might be also considered in this hydroponic floating bed

system other than herbaceous plants, such as some indigenous aquatic plants.

 These hydroponic systems can be used alone or in combination with other

traditional techniques to treat other source of wastewater, such as surface

runoff, combined sewer overflows, polluted ground water, irrigation

wastewater.

141

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Appendix A: Experimental data of batch disinfection

Table 16. E. coli count in three replicates from three lab batch UV test

E.coli count (CFU/100 mL) Time (min) test 1 test 2 test 3 0 6200 2000 7700 0 6000 1000 10500 0 5700 1500 10300 5 460 170 290 5 550 180 210 5 490 280 150 10 24 66 46 10 34 62 36 10 32 22 32 15 4 14 4 15 6 18 10 15 4 14 4 20 0 4 3 20 2 7 2 20 1 9 3

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Table 17. E. coli count in three replicates from 3.5 h field batch UV test

E.coli count (CFU/100 mL) Time (min) test 1 test 2 0 34900 13200 0 34600 14800 0 33600 15400 1 13860 10450 1 14100 9900 1 13700 10800 2.5 1574 5070 2.5 1564 4430 2.5 1558 4750 3 778 4090 3 802 3990 3 830 3870 3.5 484 1860 3.5 604 2160 3.5 522 2160

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Table 18. E. coli count in three replicates from 5 h field batch UV test

E.coli count (CFU/100 mL) Time (min) test 1 test 2 0 36600 34900 0 38400 34600 0 37200 33600 1 15200 13860 1 16400 14100 1 14800 13700 3 1460 1570 3 1380 1520 3 1540 1500 4 680 778 4 720 802 4 640 830 5 280 480 5 360 510 5 240 520

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Table 19. E. coli count in three replicates from 8 h field batch UV test

E.coli count (CFU/100 mL) Time (min) test 1 test 2 0 27900 28200 0 28400 25400 0 26400 26800 2 3500 2700 2 2800 2900 2 2900 3100 4 450 650 4 660 720 4 580 630 6 140 120 6 110 130 6 130 150 8 18 31 8 26 24 8 22 19

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Appendix B: Experimental data for biofilter nutrient removal

Table 20. Biofilter effluent ammonia concentration in three temperature treatment from 02/17/2017 to 05/15/2017

Greenhouse Plastic Control (no Date Cover Cover treatment) (mg/L) (mg/L) (mg/L) 02/17 0.01 11.00 3.50 02/24 0.01 15.60 0.08 03/03 1.24 11.00 0.30 03/11 0.16 14.20 0.35 03/16 0.42 13.00 0.55 03/25 0.23 17.40 0.45 04/01 0.26 20.80 0.19 04/07 6.70 18.40 0.13 04/11 4.50 18.80 0.22 04/21 20.80 31.60 0.27 04/29 31.80 17.00 0.31 05/02 34.40 6.60 0.11 05/08 38.20 4.80 0.45 05/15 36.40 4.50 2.45

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Table 21. Biofilter effluent nitrate concentration in three temperature treatment from 02/17/2017 to 05/15/2017

Greenhouse Plastic Control (no Date Cover Cover treatment) (mg/L) (mg/L) (mg/L) 02/17 28.6 18.5 26.8 02/24 25.3 18.1 27.8 03/03 7.4 19.4 20.2 03/11 26.6 10.7 28.5 03/16 29.5 16.5 21.9 03/25 15.8 14.1 16.3 04/01 6.2 6.9 9.5 04/07 9.6 8.5 12.1 04/11 8.4 13.5 11.4 04/21 7.7 9.0 17.1 04/29 6.1 6.0 13.6 05/02 9.2 12.4 19.5 05/08 7.2 9.7 27.2 05/15 12.6 9.2 17.0

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Table 22. Biofilter effluent total phosphorus concentration in three temperature treatment from 02/17/2017 to 05/15/2017

Greenhouse Plastic Control (no Date Cover Cover treatment) (mg/L) (mg/L) (mg/L) 02/17 12.6 12.4 10.4 02/24 12.1 11.3 11.0 03/03 8.2 9.9 10.5 03/11 10.2 10.0 8.6 03/16 8.2 10.6 10.6 03/25 12.5 11.2 10.9 04/01 10.1 10.5 11.9 04/07 11.6 12.6 12.2 04/11 14.5 11.6 9.7 04/21 10.6 14.4 8.5 04/29 9.3 11.8 9.5 05/02 9.0 8.8 10.0 05/08 9.9 11.1 10.4 05/15 12.1 10.4 10.1

166

Appendix C: Experimental data for hydroponic nutrient removal

167

Table 23. Pilot greenhouse hydroponic system ammonia concentration and removal ratio

Day after Influent Effluent Removal germination (mg/L) (mg/L) ratio (%) 3 12 9 25 5 10 8 20 7 10 9 10 9 12 8 33 11 10 1 90 13 15 3 80 15 10 3 70 17 4 0 100 19 24 10 58 21 20 10 50 23 10 3 70 25 23 8 65 27 30 27 10 29 24 11 54 31 28 20 29 33 38 17 55 35 44 8 82 37 42 6 86 39 44 3 93 41 38 2 95 43 34 2 94 45 32 3 91 47 37 12 68 49 45 1 98 51 37 1 97 53 32 2 94 55 46 18 61

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Table 24. Pilot greenhouse hydroponic system nitrate concentration and removal ratio

Day after Influent Effluent Removal germination (mg/L) (mg/L) ratio (%) 3 34 27 21 5 29 24 17 7 34 27 21 9 29 26 10 11 30 22 27 13 28 25 11 15 27 23 15 17 28 20 29 19 17 13 24 21 18 13 28 23 13 13 0 25 22 15 32 27 23 14 39 29 17 9 47 31 17 14 18 33 17 13 24 35 19 14 26 37 20 17 15 39 15 9 40 41 16 10 38 43 12 10 17 45 15 10 33 47 12 9 25 49 13 12 8 51 11 9 18 53 13 9 31 55 10 9 10

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Table 25. Pilot greenhouse hydroponic system phosphate concentration and removal ratio

Day after Influent Effluent Removal germination (mg/L) (mg/L) ratio (%) 3 29.3 12 59 5 30.5 10.5 66 7 29.3 12.9 56 9 31.2 12 62 11 28.5 8.1 72 13 23.9 8.7 64 15 25.9 11.7 55 17 27.3 6.8 75 19 23.6 21.1 11 21 36.1 29.8 17 23 33.1 18.3 45 25 32.2 21.2 34 27 34.5 19.2 44 29 35.6 28.2 21 31 33.4 29.5 12 33 32 20.8 35 35 31.5 21.2 33 37 27.6 13.7 50 39 26.9 13.7 49 41 26.5 10.9 59 43 26.9 15.3 43 45 23.5 5.5 77 47 18.3 15.2 17 49 24.8 9.2 63 51 20.6 3.9 81 53 14.3 5.6 61 55 17.1 8.6 50

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Table 26. Pilot greenhouse hydroponic system HSD, LSD and control group ammonia concentration

Day after Influent Effluent (mg/L) germination (mg/L) HSD LSD control 2 0.15 0.05 0.04 0.05 5 0.1 0.03 0.03 0.04 7 0.09 0.02 0.03 0.05 9 0.23 0.03 0.03 0.02 12 0.16 0.04 0.02 0.05 14 0.21 0.01 0.01 0.03 17 0.2 0.02 0.04 0.05 20 0.18 0.01 0.01 0.02 22 0.16 0.03 0.04 0.04 24 0.13 0.04 0.06 0.05 27 0.08 0.02 0.04 0.04 29 0.09 0.02 0.02 0.02 31 0.12 0.02 0.04 0.03 34 0.19 0.02 0.02 0.03 37 1.02 0.03 0.02 0.02 39 1.75 0.02 0.02 0.02 41 2.31 0.08 0.08 0.27 44 3 0.01 0.01 0.36 50 2.2 0.06 0.01 0.5 56 2 0.4 0.1 0.3 59 0.48 0 0 0.04

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Table 27. Pilot greenhouse hydroponic system HSD, LSD and control group nitrate concentration

Day after Influent Effluent (mg/L) germination (mg/L) HSD LSD control 2 45 35 32 28 5 60 28 30 26.7 7 59 55 48 45 9 54 43 47.5 32 12 58 50 51 47 14 48 48 43 42 17 30.5 35.5 44 37.5 20 30 14.8 16.9 16.5 22 16.7 12.3 13.7 17.7 24 16.5 11.6 12.6 8.7 27 15.1 12.1 13 11 29 17.9 14.7 13.6 13.1 31 18.4 13.8 12.4 13.5 34 12.7 9.2 9.9 10.05 37 15.2 11 6.9 10.5 39 11.6 9.8 12.6 12.6 41 10.8 11.6 8.6 8.8 44 12.6 10.2 8.3 9.2 50 13.5 11.8 9.5 9.8 56 11.3 12.8 10.8 11.6 59 14.8 10.5 9.4 9.8

172

Table 28. Pilot greenhouse hydroponic system HSD, LSD and control group phosphate concentration

Day after Influent Effluent (mg/L) germination (mg/L) HSD LSD control 2 1.81 1.58 0.8 0.5 5 2.58 0.3 0.36 0.27 7 1.37 0.26 0.23 0.19 9 2.74 0.64 0.49 0.25 12 3.28 0.42 0.47 0.26 14 2.87 0.4 0.3 0.25 17 2.4 0.68 0.55 0.48 20 2.75 0.73 0.73 0.67 22 5.84 0.83 0.69 0.68 24 5.9 1.89 2.07 1.91 27 3.29 1.23 1 1.24 29 3.5 0.76 1.03 0.92 31 6.2 2.05 2.15 2.2 34 3.4 0.57 0.78 0.54 37 2.7 0.72 1.19 0.7 39 5.5 1.06 1.23 1.13 41 5.66 1.33 1.47 2.06 44 5.9 0.85 1.36 2.7 50 4.6 1.9 2.1 5.1 56 3.2 1.55 1.56 2.85 59 3.1 1.66 1.48 3.08

173

Table 29. High tunnel greenhouse hydroponic system grass and control group ammonia concentration

Days after Influent Effluent (mg/L) germination (mg/L) grass control 0 11.2 6.8 7.4 2 14.4 8.2 12.8 4 18.6 13.8 14.2 6 20.8 16.4 17.2 8 32 16 28.4 10 41.2 15.4 37.6 12 42.4 16.8 33.4 14 38.5 9.4 31.5 16 34.2 7.6 24.8 18 33.9 8.2 23.2 20 38.4 7.3 31.6 22 42.8 10.9 33.8 24 34.6 7.1 29.8 26 38.6 5.9 24.8 28 39.4 8.3 23.2 30 27.6 6.2 17.4

174

Table 30. High tunnel greenhouse hydroponic system grass and control group nitrate concentration

Days after Influent Effluent (mg/L) germination (mg/L) grass control 0 6.4 5.4 6.2 2 13.4 9 10.8 4 5.9 3.7 5.2 6 7.7 6.2 6.2 8 5.2 3.1 5.1 10 6.2 3.4 5.4 12 6.1 3.9 5.2 14 7.8 4.1 6.7 16 7.2 3.2 5.3 18 7.3 3.8 5.5 20 12.6 5.5 10.6 22 11.5 4.8 10.1 24 8.9 2.9 7.6 26 9.6 3.3 7.1 28 10.2 3.8 8.2 30 10.5 4.5 6.9

175

Table 31. High tunnel greenhouse hydroponic system grass and control group phosphate concentration

Days after Influent Effluent (mg/L) germination (mg/L) grass control 0 11.4 10.05 9.75 2 9.5 7.46 7.85 4 7.4 6.15 6.18 6 10.55 7.22 8.52 8 8 3.78 6.55 10 8.05 3.26 7.01 12 9.3 3.5 7.65 14 11.2 5.08 8.34 16 10.35 2.85 7.27 18 9.52 1.74 7.42 20 8.65 1.36 6.29 22 11.3 1.19 8.98 24 12.8 2.08 9.12 26 8.66 0.98 6.34 28 9.84 1.24 7.68 30 12.45 2.36 9.46

176