3 Phosphate Removal from Water Using Alginate/Carboxymethylcellulose/Aluminum Beads and Plaster of Paris Phosphorus Adsorption

Potential Approach for the Adsorption of Phosphate from Agricultural Runoff using Plaster of Paris Powder and Hydrogel Beads

Srdjan Malicevic

A Thesis presented to The University of Guelph

In partial fulfilment of requirements for the degree of Master of Science in Engineering

Guelph, Ontario, Canada

POTENTIAL APPROACH FOR THE ADSORPTION OF PHOSPHATE FROM AGRICULTURAL RUNOFF USING PLASTER OF PARIS POWDER AND HYDROGEL BEADS

Srdjan Malicevic Advisor(s):

University of Guelph, 2020 Dr. Erica Pensini

Dr. Prasad Daggupati

Phosphorus released in lakes due to agricultural runoff causes eutrophication, deteriorating water quality and ecosystem harm. Adsorbing and recovering phosphorus could potentially contribute to a circular economy and reduce eutrophication. A literature review of phosphorus adsorbents was conducted to isolate for ideal adsorbents after defining criteria for surface water adsorption. Two adsorbents were studied for the removal of phosphate from water: plaster of Paris powder and hydrogel beads produced using alginate, carboxymethylcellulose, and aluminum. The reaction kinetics, adsorption capacity, and ability to desorb were compared. In deionised water, hydrogel beads had a

3- maximum sorption capacity of 90.5 mg PO4 /g dry bead with an equilibration time of approximately 24 hours. In deionised water, plaster of Paris (POP) powder has a

3- maximum capacity of 1.52 mg PO4 /g with an equilibrium time of less than 10 minutes.

Sorbents can potentially be reused following phosphate desorption, and desorbed phosphate may be reused as fertilizer.

iii

ACKNOWLEDGEMENTS

I would like to thank Dr. Erica Pensini for her mentorship and guidance during my degree and research. I would also like to thank Dr. Prasad Daggupatti for his assistance and advice on field-scale applications. I am grateful for the support from my peers such as Kristine Lamont, Samantha Mehltretter, Stephen Vanderburgt, and many more. I would like to especially thank Klaudine Estepa for her unending support and help in setting up experimental methods.

The research described in this paper was funded by the Ontario Ministry of Food, Agriculture and Rural Affairs (OMAFRA) through the OMAFRA UofG program awarded to Drs. Erica Pensini and Prasad Daguppati. This research was also supported through the Mitacs Globalink program, with funding awarded to Ana Paula Garcia Pacheco and Erica Pensini. The authors greatly appreciate the support offered throughout the project by Joanne Ryks and Ryan Smith at the University of Guelph.

TABLE OF CONTENTS

Abstract ...... ii

Acknowledgements ...... iii

Table of Contents ...... iv

List of Tables ...... vii

List of Figures ...... viii

List of Appendices ...... x

1 Introduction ...... 1

1.1 Thesis Structure ...... 1

1.2 Research Context ...... 1

1.2.1 Phosphorus Removal ...... 1

1.2.2 Review of Phosphorus Sources and Sinks ...... 3

1.3 Problem Statement and Objective ...... 6

2 Literature Review ...... 8

2.1 Phosphorus ...... 8

2.2 Existing Technology for Phosphorus Removal ...... 10

2.3 Adsorption Dynamics and Adsorbent Material Classifications ...... 13

2.3.1 Polymeric Adsorbents: Hydrogels ...... 16

2.3.2 Adsorption Isotherms and Reversibility ...... 21

2.3.3 Design Constraints and Criteria for Adsorbents Intended for Surface Water 26

2.3.4 Literature Review on Phosphorus Adsorbents ...... 30

2.3.5 Selected Materials for Phosphorus Adsorption ...... 33

3 Phosphate Removal from Water Using Alginate/Carboxymethylcellulose/Aluminum Beads and Plaster of Paris Phosphorus Adsorption ...... 34

3.1.1 Materials ...... 34

3.1.2 ALG-CMC (ACMC) bead preparation ...... 34

3.1.3 Shear Rheology Experiments ...... 35

3.1.4 Analytical Methods for Phosphorus Detection ...... 36

3.1.5 Phosphate Sorption and Desorption Experiments ...... 36

3.1.6 Modeling Approach ...... 37

3.1.7 Statistical Assumptions for Non-linear Fitting of Isotherms ...... 39

3.2 Results and Discussion ...... 40

3.2.1 Sorption of Phosphorus onto POP powder ...... 40

3.2.2 Desorption of Phosphorus from POP powder ...... 44

3.2.3 Sorption of Phosphorus onto ACMC beads ...... 45

3.2.4 Effect of Competing ions on Sorption of Phosphorus onto ACMC beads .. 52

3.2.5 Desorption of Phosphorus from ACMC beads ...... 52

4 Conclusion ...... 54

4.1 Phosphorus Sorption ...... 54

4.2 Future Work ...... 57

References ...... 59

Appendices ...... 76

vii

LIST OF TABLES

Table 1 Hydrogel compositions and definitions ...... 16

Table 2 Adsorption mechanisms and related inter/intramolecular force ...... 19

Table 3 Summary of reversibility constants and isotherm shapes ...... 24

Table 4 Relationship between adsorbent criteria and comparison between chemisorption and physisorption ...... 29

Table 5 Literature review on adsorbent materials and performance ...... 31

Table 6 Nonlinear isotherm parameters for sorption of phosphate onto POP powder at pH= 7...... 42

Table 7 Isotherm parameters for sorption of phosphate onto POP powder at pH= 7. ... 42

Table 8 Sorption capacity of beads having different composition, at pH= 7. The initial 3- phosphate concentration in DI was 2 mg/L PO4 ...... 46

Table 9 Nonlinear isotherm parameters for sorption of phosphate onto ACMC beads at pH = 7 ...... 49

Table 10 Linear isotherm parameters for sorption of phosphate onto ACMC beads at pH = 7 ...... 49

Table 11: Comparison of materials with constraints and criteria with all factors determined ...... 56

viii

LIST OF FIGURES

Figure 1 Google Earth satellite image of Lake Erie on October 9th 2011...... 2

Figure 2 Conceptual diagram of holistic nutrient transport denoting “N” as nitrogen and “P” as phosphorus; font size indicates greater relative loss through each respective pathway ...... 7

Figure 3 Generalized phosphorus cycle and concept behind a circular economy for phosphorus adsorption and desorption/recovery ...... 8

Figure 4 Chemical structure of orthophosphate ...... 9

Figure 5 pH-logC diagram of phosphoric acid ...... 10

Figure 6 Classification of adsorption mechanisms ...... 18

Figure 7 Phosphate loading onto spherical adsorbent increases as phosphate concentration increases ...... 22

Figure 8 General isotherm shapes showing portions describing reversibility ...... 25

Figure 9 Graphic explanation of ideal selective reversibility on an adsorption isotherm 27

Figure 10: Relationship between the separation factor and practical loading at an initial concentration of 0.49 mg-P/L ...... 28

Figure 11 Image of ACMC beads with ruler showing centimetres and approximate diameter of beads...... 35

3- Figure 12 Sorption of phosphate at pH= 7 by 12.5 g/L POP powder at 1 mg/L PO4 . The line is a guide to the eye...... 41

3- Figure 13 Adsorption isotherm for PO4 sorption with POP powder at pH= 7 ...... 42

Figure 14 Effect of initial pH 7, 4.5, and 8.3 on removal efficiency of phosphate by POP powder...... 44

Figure 15 Amount of phosphate desorbed from POP powder at varying initial pH. The dashed line is a guide ...... 45

Figure 16 Sorption of phosphate by 0.7 g/L ACMC beads with initial phosphate concentrations of 2 mg/L ...... 47

Figure 17 Adsorption isotherm of phosphate and ACMC beads at pH = 7...... 49

3- Figure 18 Effect of pH on phosphate removal using 16.7 g/L ACMC beads at PO4 = 2 mg/L. The dashed line is a guide to the eye...... 51

Figure 19 ACMC beads collected and contained in a cotton cloth that can be readily applied on the field...... 52

Figure 20 Amount desorbed from ACMC beads in DI water at varying initial pH. The line is a guide to the eye...... 54

Figure 21 Drainage ditch system for phosphorus adsorption with additional graphic of a potential multi-adsorbent system tailored to water stage and sediment profile ...... 58

LIST OF APPENDICES

Appendix A ...... 76

Appendix B ...... 77

Appendix C ...... 78

1 Introduction

1.1 Thesis Structure

This thesis is divided into four chapters: the introduction, which explains the context and motivation for research; a literature review on adsorption theory and current technologies; research conducted on phosphate sorption onto POP powder and alginate-carboxymethylcellulose beads crosslinked by aluminum; and a conclusion comparing materials and relating research results to design goals and criteria.

1.2 Research Context

1.2.1 Phosphorus Removal

Eutrophication has been a historical problem in the Great Lakes watershed and has been exacerbated by climate change. In addition to increasing the temperature of the Great Lakes (Mortsch & Quinn, 1996), climate change has increased the frequency and intensity of stormwater events in Ontario and causes greater demand in agricultural productivity (Stevanović et al., 2016). These changes present multiple challenges with respect to contaminant transport and protecting local ecosystems, and is speculated to be a cause for increased phosphorus loadings in the Lake Erie watershed (Bosch, Evans, Scavia, & Allan, 2014; Michalak et al., 2013). Non-point source pollution from agricultural runoff and increased erosion has become a significant problem due to increased storm events and changes in tilling practices (Bosch et al., 2014).

Figure 1 Google Earth satellite image of Lake Erie on October 9th 2011

Canada and the United States have a history of eutrophication problems with respect to Lake Erie (Allinger & Reavie, 2013) and the Great Lakes (Estepp & Reavie, 2015) watershed that dates back to 1960. The historical problems in Lake Erie with algal blooms prompted the 1972 Great Lakes Water Quality Agreement (GLWQA) which succeeded in reducing the level of phosphorus solely using nutrient management (Horachek, Nuri, Ollis, & Kuwahata, 2015). The problem reappeared in the 1990s because of changes in tilling practices, increased agricultural loads, and climate change (Bosch et al., 2014; Scavia et al., 2014). The resurgence of eutrophication problems was addressed in 2012, when the GLWQA agreement was amended by the addition of the Great Lakes Nutrient Initiative that is currently utilizing nutrient management, agricultural BMPs, the identification of phosphorus hotspots, as well as monitoring and modelling the nutrient levels in the Great Lakes (Canada-Ontario Lake Erie Action Plan, 2018). The Ministry of the Environment, Conservation and Parks set a guideline of 20 µg/L of total phosphorus to reduce nuisance concentrations of algae in lakes, and a limit of 30 µg/L for rivers and streams (MECP, 2019). In the 2012 GLWQA the interim substance objective for open-water concentration of total phosphorus in Lake Erie is approximately 10-15 µg/L based on springtime averages. Similarly, a loading target of

11,000 metric tonnes of total phosphorus per year was set (EPA, 2015). A 40% annual reduction in total phosphorus loading was set by the GLWQA for Lake Erie. The EPA recommends a 6,000 megatonne TP annual target for 40% reduction from 2008 levels in the Central Basin of Lake Erie (EPA, 2015). This figure includes a removal of 124 megatonnes of dissolved reactive phosphorus during springtime loading (EPA, 2015).

There has been an increase in both monitoring and regulating the amount of phosphorus introduced into freshwater basins and streams (Chad & James, 2019). A heavy focus of these regulations is for cultivated lands with high runoff and sedimentary loading has been determined in watersheds and point sources such as urban stormwater and wastewater treatments plants (Canada-Ontario Lake Erie Action Plan, 2018). In southwestern Ontario, where this thesis is authored, committees such as the Thames River Phosphorus Reduction Collaborative (TRPRC) have been deployed to implement local solutions, raise awareness for reducing surface water phosphorus, and brainstorm strategies to reduce phosphorus loading (TRPRC, 2020). There is a demonstrated provincial and local need for emerging technology that can reduce phosphorus levels in the Great Lakes watersheds or other phosphorus hotspots. The technological focus of this research is on adsorption, a promising method of reducing surface water concentrations and potentially retaining adsorbed phosphate.

1.2.2 Review of Phosphorus Sources and Sinks

Before discussing how phosphate is transported in the environment the main source of phosphate – the mining of phosphate rock—should be addressed. The U.S. Geological Survey reports that there are over 300 billion tons of phosphate rock in reserve (USGS., 2020). Throughout recent history, there has been controversy over whether or not current use of phosphorus is unsustainable and if it will result in depletion of phosphorus reserves (Ulrich & Frossard, 2014) and there is still disagreement on how long reserves will last as well as uncertainty of new phosphate sources (Ulrich & Frossard, 2014). Regardless of whether or not phosphorus is becoming scarce, it is safe to say that 3

introducing new phosphorus into the phosphorus cycle is unacceptable as some phosphorus sinks such as freshwater basins are approaching their natural limit or “planetary boundary” (Carpenter & Bennett, 2011). Phosphate rocks should remain untouched in order to sequester it (Carpenter & Bennett, 2011) and it should be recovered from freshwater basins and sediments. Recently, there has been many new studies focusing on strategies for a circular phosphorus economy (Geissler B, L., MC., & G., 2018; Nesme & Withers, 2016). An adsorbent that can remove phosphorus from surface water and release and recover it is an ideal way to solve both scarcity and eutrophication problems.

There are multiple pathways phosphates can be introduced into nearby freshwater networks such as fertilizer/biomass/manure runoff facilitated by water/wind erosion and seepage (Riemersma, Little, Ontkean, & Moskal-Hébert, 2006). The increased concentrations of phosphates found in watersheds including well known hotspots such as the Thames Basin, UK, the Yangtze Basin, China, and Maumee Basin, USA (Stephen et al., 2016) have been attributed to the increased amount of phosphates used in modern fertilizers (Zhang, Wen, Li, & Shi, 2014). As broadcasting is a common distribution method used by farmers to apply fertilizers, water runoff, soil erosion, and subsurface interflow from phosphorus enriched groundwater are just some of the transport pathways from agricultural lands to water basins (Sharpley, 2001).

In 2016 Power et al., Nemery, and Garnier found that despite halting high inputs of phosphate fertilizers there was a significant lag in phosphorus mobilization, this could be due to phosphate accumulation at the bottom of the basins or the surrounding groundwater (Stephen et al., 2016).

In comparing forested lands to croplands near Baptiste Lake in Alberta one study found 60-91% of the dissolved OP to originate from agricultural watersheds (Cooke & Prepas, 1998). Miller et al. estimated 70% of phosphate runoff from agricultural lands come from croplands, 20% from livestock, and 10% from erosion from adjacent banks (Miller, 4

Robinson, Coote, Spires, & Draper, 1982). Manure from livestock has been found to add phosphate to soil, however, only a small percent of phosphate is dissolved reactive phosphorus. As a result, manure as a result of livestock is not as impactful to the increase of dissolved reactive phosphorus found in freshwater systems as rainfall or phosphate interflow (Hou, Chen, Liao, & Luo, 2017). Nash et. al. found that fertilizer added five times the amount of dissolved phosphates (35 kg-1 ha-1 yr-1) compared to manure (Nash, Hannah, Halliwell, & Murdoch, 2000).

Frequency and load of phosphates introduced from these sources can widely vary based on factors such as: soil composition, season, precipitation events, and specific land use. Soils with a higher percent of clay, ferric oxides, aluminum oxides, and forms of calcium phosphate may slow or reduce the introduction of phosphate to the freshwater system as phosphorus is usually adsorbed to these geosorbents during subsurface transport (Ayenew, Tadesse, Kibret, & Melese, 2018). Van Keuren et al. found high phosphate addition to a nearby water source with no detectable soil erosion during a high intensity event (Van Keuren, Mc Guinness, & Chichester, 1979). In the soil, phosphorus is usually adsorbed onto ferric oxides, aluminum oxides, or form calcium phosphate during travel through the subsurface (Goldberg & Sposito, 1984). Phosphorus that remains in the root zone may be recycled to the soil surface by plant uptake and degradation.

Rivers and streams are both sources and sinks of phosphorus. During low-flow conditions, riverbeds often acts as a phosphorus adsorbent. However, high-flow conditions cause riverbed sediments to be disturbed and for phosphate to re-enter the stream, as well as additional phosphorus to enter from agricultural runoff. Groundwater pathways are usually not considered for phosphorus transport given that phosphorus is readily adsorbed onto soils. However, some groundwater systems with poor capacity for anions or fractured bedrock allow phosphorus to migrate unmitigated (Jalali & Peikam, 2013). Depending on the hydrogeologic conditions of a water body and the aquifer, a

significant amount of phosphorus may enter at specific depths such as the thermocline layer of a stream or lake (Shaw, Shaw, Fricker, & Prepas, 1990). Once the dissolved OP is in the aqueous phase it is available for temporary storage and uptake by both crops or microorganisms in water.

Eventually, particulate phosphate and dissolved phosphate travels into lakes where it settles and circulates. There is no effective way to remove dilute phosphorus once it enters the biochemical cycle of lakes and other water basins. Therefore, it is extremely important to remove phosphorus at more concentrated sources and to keep the phosphorus cycle from reaching unsustainable levels.

1.3 Problem Statement and Objective

The primary purpose of this research is to design an adsorbent that can reduce eutrophication by adsorbing soluble reactive phosphorus from agricultural runoff that pervades streams and tributaries leading to Lake Erie. An additional objective is to recover the phosphorus that is adsorbed so that it may be reused in agricultural fields or other processes. In order to have a solution that is both environmentally effective and economically viable, the adsorbent should have the ability to be applied in the simplest way possible.

Figure 2 provides a thorough depiction of the transport pathways for phosphorus discussed in section 1.2.2. Surface runoff is the unique pathway that is the intended focus of this research, as it is the pathway with the least developed solutions and for its unmitigated contribution of phosphorus to lakes.

Intended point(s) of Targeted pathway application for adsorbent material

Figure 2 Conceptual diagram of holistic nutrient transport (McDowell, Withers, & Weerden, 2016) denoting “N” as nitrogen and “P” as phosphorus; font size indicates greater relative loss through each respective pathway

The general intended purpose in terms of phosphorus recyclability is illustrated in Figure 3. Avoiding the dilution of phosphorus in lakes is pivotal to reducing and recovering the material, and therefore targeting high concentration streams is the ideal use for adsorption

Fertilizer Application

Runoff (streams, Mining drainage ditches) Adsorption, desorption

Sedimentation Lakes

Figure 3 Generalized phosphorus cycle and concept behind a circular economy for phosphorus adsorption and desorption/recovery

2 Literature Review

2.1 Phosphorus

Phosphorus (P), atomic number 15, is a highly reactive element essential to all life on Earth. Phosphorus is typically classified into 3 categories: elemental, inorganic, and organic. Elemental phosphorous is comprised solely of phosphorus and has three allotropes: red, black, and white phosphorus. Inorganic phosphates are comprised of phosphorus and oxygen groups; this form of phosphorus is also known as orthophosphates (PubChem). Orthophosphates are the most impactful to freshwater bodies as it is the most bioavailable to organisms (Ding et al., 2015; D. W. Schindler, 1974). As seen in Figure 4, orthophosphates are an anion monomer that have four oxygen atoms surrounding a single phosphate atom in a tetrahedral formation and has

a molar mass of 95 g/mol. Organic phosphates contain covalent bonds between other phosphate monomers and variety of organic function groups (Morris, Perkin, Rose, & Smith, 1982). Phosphorus causes eutrophication when present in excess, and is an emerging problem in the face of increased agricultural activity (Lamont et al., 2018; Meinikmann, Hupfer, & Lewandowski, 2015; David W. Schindler, Carpenter, Chapra, Hecky, & Orihel, 2016). Phosphorus is found in water bodies as elemental, inorganic, and organic. Orthophosphates are comprised of phosphorus and oxygen (Morris, Perkins, Rose, & Smith, 1977). Orthophosphates (also defined as dissolved reactive phosphorus, DRP) are used as fertilizers and are the most problematic form of phosphorus in water bodies, because they are promptly bioavailable to microorganisms and algae, causing algal blooms (Boström, Andersen, Fleischer, & Jansson, 1988).

Figure 4 Chemical structure of orthophosphate

- 2 Phosphoric acid is a triprotic acid that can exist in four species: H3PO4, H2PO4 HPO4 , 3- and PO4 (Windholz, 1983). As the most water-soluble form of phosphorus are orthophosphates the concentration of orthophosphates is higher than organic phosphates in freshwater systems. As seen in Figure 5, the dominant species in the aqueous solution is dependent on the pH of the medium.

pH-LogC Diagram of Phosphoric Acid

0.00 0 2 4 6 8 10 12 14

-5.00

-10.00 logC -15.00

-20.00

-25.00 pH Phosphoric acid Dihydrogen phosphate Hydrogen phosphate Phosphate

Figure 5 pH-logC diagram of phosphoric acid

2.2 Existing Technology for Phosphorus Removal

Since DRP is the most problematic form of phosphorus in water bodies, research has focused on identifying methods to remove DRP from either wastewater or agricultural drainage water. Methods used to remove DRP from wastewater include batch reactors in controlled anaerobic conditions (Obaja, Macé, Costa, Sans, & Mata-Alvarez, 2003), chemical precipitation (Pratt, Parsons, Soares, & Martin, 2012; Rittmann, Mayer, Westerhoff, & Edwards, 2011; Röske & Schönborn, 1994), phosphorus recovery in the form of struvite (Jaffer, Clark, Pearce, & Parsons, 2002), or its incorporation into xonolite crystals (Xuechu Chen, Kong, Wu, Wang, & Lin, 2009). Phosphate-rock mines are being depleted because of the significant demand of this mineral (Cordell & Neset, 2014; Neset & Cordell, 2012). Therefore, methods have also been proposed to recover DRP present in sewage sludge and reuse it as fertilizer. These methods include the

thermochemical treatment of sludge (Adam, Peplinski, Michaelis, Kley, & Simon, 2009), extraction of phosphorus from sludge ashes (Takahashi et al., 2001) and from tertiary or digested sewage sludge (Monea, 2020), or its crystallization to form struvite (Xu, He, Gu, Wang, & Shao, 2012) or other fertilizing minerals (Xuechu Chen et al., 2009). The methods of removal and recovery of phosphorus used for wastewater are not suitable for agricultural drainage water, because the phosphorus loadings are lower, and the water is not collected in a central treatment facility. An alternative to the methods described above is the sorption of phosphorus onto diverse materials, including bauxite (Altundoǧan & Tümen, 2003) and magnetite microparticles (Drenkova-Tuhtan et al., 2017; Xiao, Liu, Zhang, & Zheng, 2017). Adsorbent magnetic particles have also been successfully tested over numerous regeneration cycles for the removal of non-orthophosphate species, e.g. dissolved organo-phosphonates, which can also pose a risk of eutrophication once released into the environment (Rott et al., 2018). Other phosphorus sorbents include nano zero-valent iron (Eljamal, Khalil, Sugihara, & Matsunaga, 2016), iron, activated aluminum oxide and granulated ferric hydroxide (Genz, Kornmüller, & Jekel, 2004), alum (Banu, Do, & Yeom, 2008), red mud granular adsorbents (Zhao et al., 2012), submerged vegetation (Dierberg, Debusk, Jackson, Chimney, & Pietro, 2002), calcite (Karageorgiou, Paschalis, & Anastassakis, 2007), and calcium-rich minerals (Lamont et al., 2018). Adsorption of phosphorus using calcium rich minerals is effective due to the binding capacity between calcium and phosphorus (Edzwald, Toensing, & Leung, 1976; Spears, Meis, Anderson, & Kellou, 2013). Plaster of Paris powder is inexpensive and easy to source, and it contains calcium minerals. While its ability to sorb fluoride has been investigated (Gopal & Elango, 2007), its effectiveness in sorbing phosphorus has not been previously studied. Also, N - isopropylacrylamide and aluminum alginate beads (Wan et al., 2016) and polyacrylamide-alginate magnetic beads (Luo, Rong, Zhang, Zeng, & Wan, 2019) have been used for the removal of phosphorus from water. While effective, the degradation of polyacrylamide and parent compounds into acrylamide can raise concerns, because acrylamide is a carcinogen

(Capuano & Fogliano, 2011), it causes reproductive toxicity (H. Wang et al., 2010), and it has neurodegenerative properties (Dejongh, Nordin-Andersson, Ploeger, & Forsby, 1999). Beads obtained with the synthetic polymer polyvinyl alcohol and alginate have also been proposed as phosphorus sorbents (Zhou et al., 2018). The use of polyvinyl alcohol is likely preferable to polyacrylamide, because its acute oral toxicity is low and its median lethal dose (LD50) is 15- 20 g/kg (Demerlis & Schoneker, 2003).Other authors proposed chitosan beads for the removal of negatively charged phosphate ions from water, following the removal of copper (Dai et al., 2011). While effective, these beads could pose risks when used as phosphorus sorbents in agricultural drainage ditches, because copper can cause oxidative damage, leading to disorders associated with abnormal Cu metabolism and neurodegenerative diseases (Gaetke & Chow, 2003).

Different solutions exist for different species and forms of phosphate. In the case of sedimentary phosphate, or phosphate that partitions from particulate to dissolved forms, a patent has been made by Phoslock for a bentonite clay activated by lanthanum that can be sprayed into lakes. The purpose of Phoslock is to both sorb dissolved phosphorus during its descent and then upon settling, preventing sedimentary 3- phosphate to be released into the lake. Literature reported 95% removal of PO4 ions using Phoslock in synthetic wastewater (Magdalena Hanna & Magda, 2017). Studies applying Phoslock to sediments reported good enrichment of lanthanum content in the first 8 cm of sediment, and could bind 42% of mobile phosphorus in the top 4 cm of sediment, and 17% in the top 10 cm (Meis, Spears, Maberly, O’malley, & Perkins, 2012). While this amendment can be applied to lakes, it is unable to address phosphorus in stream sediments as high flow conditions recirculate phosphorus back into the aqueous phase.

There seems to be a demonstrated need for an adsorbent material that can sorb phosphorus from agricultural drainage ditches (runoff) where concentrations of

phosphorus are high (Appendix A), and the ability to recover and reuse sorbed phosphorus.

2.3 Adsorption Dynamics and Adsorbent Material Classifications

In general terms, adsorption is the adherence of an ion in a gas or liquid bulk phase to the surface of another bulk phase. While gas to solid phase adsorption is mostly used in industries for purification and separation processes (Ruthven, 1984), adsorption in water treatment has been proven to be effective in wastewater, drinking water (Singh, Nagpal, Agrawal, & Rachna, 2018) groundwater, aquariums, and even for landfill leachate (Worch, 2012). The term sorption is commonly used instead of adsorption as it can be unclear how much of the solute is adsorbed on an external surface versus absorbed on internal surfaces (Worch, 2012). Adsorbents can be loosely classified into categories that can meaningfully distinguish adsorption behavior, mechanisms, or material origin.

Naturally occurring sorption found in the environment is sometimes called geosorption. In one sense, geosorption is the most important type of sorption when applying adsorbents in natural systems as they may significantly change water chemistry. For example: clays and other soils have high cation-exchange capacity (Rotenberg, Morel, Marry, Turq, & Morel-Desrosiers, 2009) (CEC) and adsorb phosphates through ion exchange resulting in particulate phosphates which may desorb back into the aqueous phase (Doig et al., 2017). Understanding geosorption is important as it elucidates the fate mechanisms and species of phosphorus that may be found in an agricultural drainage ditches, surface water, and groundwater.

Natural zeolites have garnered a lot of attention as adsorbents in the environment due to their low cost, high surface area, and low environmental impact (Burmańczuk, Markiewicz, Kowalski, Roliński, & Burmańczuk, 2012; Puspitasari, Kadja, Radiman, Darwis, & Mukti, 2018; Rivera-Hernández & Green-Ruiz, 2014). Zeolites are 13

aluminosilicate sheets that have high surface area (Thomas & Leary, 2014) and CEC, allowing them to sorb solutes from aqueous solutions. Due to their high surface area and affinity to bind cations, coating zeolites with cationic surfactants to then sorb anions is an effective method for anion adsorption. Cationic surfactants are often harmful to the environment, however, and thus have not been used extensively for sorption of anions such as phosphate in the environment (Reeve & Fallowfield, 2017). An example of a surfactant-modified zeolite is clinoptilolite coated by the quaternary ammonium surfactant hexadecyltrimethylammonium bromide (Dionisiou, Matsi, & Misopolinos, 2013).

Adsorbents that are designed, modified, coated, or any way engineered in the lab to have specific properties are engineered adsorbents and are typically very costly; an example of an engineered adsorbent is activated carbon (Worch, 2012). These types of adsorbents may be appropriate for wastewater or highly toxic contaminants that pose serious health risks such as hydrocarbons (Maretto et al., 2015) or heavy metals such as chromium (Sharma, Jalilnejad, & Yarusova, 2017), but some engineered adsorbents can be highly toxic themselves when exposed to low pH levels (Duranceau, Biscardi, & Barnhill, 2016) and thus are not suitable for use in the environment. Engineered adsorbents such as impregnated oxidic adsorbents or ion exchange resins (P. S. Kumar, Korving, van Loosdrecht, & Witkamp, 2019) are not likely cost-effective for recovering phosphorus and therefore are not used in sorption where a circular economy is a secondary motivation.

Polymeric adsorbents are adsorbents made from polymers, typically by the copolymerization of styrene (Worch, 2012). These adsorbents are usually used to sorb single solutes and recycle them in controlled processes that allow for a solvent to regenerate the material and recover sorbed material. Styrene and other synthetic polymers can have toxic properties (Carlson, 2012) or leach/degrade into other

materials with high toxicity (NIOSH, 1983). Polymeric adsorbents are a focus of the research study and are discussed further in section 2.3.1.

Low-cost adsorbents are sorbents that, beyond being inexpensive, are often waste products or byproducts such as rice husks (Abbas, 2015) and other agricultural waste (T. A. H. Nguyen et al., 2014). While these adsorbents are low-cost, they are often expended after one application and have low surface area (T. A. H. Nguyen et al., 2014).

It is difficult to determine the ideal adsorbent material category as there are criteria that should be maximized such as reuse of the material, cost, efficiency, and toxicity that are somewhat at odds with each other. It seems to be the case that most natural materials are not optimized in terms of kinetics or adsorption capacity, and that those characteristics are imparted by modification of existing or fabricated materials. More targeted design criteria are provided in section 2.3.3 that are mostly exclusive of material classification.

2.3.1 Polymeric Adsorbents: Hydrogels

A hydrogel is a gel obtained by crosslinking polymers in aqueous solution. Polymers are usually classified into either synthetic or natural polymers, the former typically having greater internal surface area, mechanical strength, and durability (Ahmed, 2015). For environmental applications, natural polymers are more desirable as they typically have lesser environmental impact (Morganti, Coltelli, & Morganti, 2018).

The use of non-toxic polymers as sorbents for phosphorus is particularly attractive. Examples of non-toxic polymers are alginate and cellulose derivatives (K. Lee et al., 2017; K. Wang & Ye, 2010). The study in section 0 addresses the existing gaps by investigating the phosphate removal effectiveness of plaster of Paris and beads obtained using non-toxic polymers (carboxymethylcellulose and alginate). The combination of alginate and carboxymethylcellulose creates a natural dual copolymeric hydrogel based on the following classification system in Table 1.

Table 1 Hydrogel compositions and definitions (Ahmed, 2015)

Composition Definition Example(s) (Bahram, Mohseni, & Moghtader, 2016)

Homopolymeric One monomer chain Polyethylene glycol

Copolymeric Two or more monomers with at Carboxymethylcellulose, least one hydrophilic component Alginate

Multipolymeric Inter-penetrating Two independent cross-linked Chitosan network polymers

While polymeric adsorbents can maintain high surface area through internal pores, the typical downfall of internal surface area sorption is that intra-particle diffusion is the rate- limiting reaction mechanism and is affected by grain size (Hill & Root, 2014; Worch, 2012). In some ways, while polymeric adsorbents provide applicable and recoverable sorbents, the tradeoff with a large grain size is slower reaction kinetics despite a high sorption capacity.

Adsorbent classifications can have overlap. When a natural polymer is used as an adsorbent, the material can be described as a biosorbent, although it is possible that the polymer itself contributes no energy for adsorption and that the crosslinking agent may supply all adsorption sites post-crosslinking. Some materials are only classified based on the origin of a material and not the dominant mechanism (Robalds, Naja, & Klavins, 2016), thus the term biosorbent is fluid and can describe many materials.

2.4.1 Adsorption Mechanisms

A conceptual understanding of adsorption mechanisms relies on thorough knowledge of both intermolecular and intramolecular forces. Both electrostatic and van der Waals forces are involved in physisorption. Chemical bonds are stronger than van der Waals forces and form when species in water and chemisorbed onto sorbent materials (Israelachvili, 2011). In almost all cases, the adsorption process for a specific material has some degree of both physisorption and chemisorption (Worch, 2012).

In addition to physisorption and chemisorption, both biosorbents (i.e. sorbent materials of biological origin (Robalds et al., 2016)) and other sorbents can rely on other sorption mechanisms, including ion exchange and microprecipitation (Robalds et al., 2016; H. N. Tran, You, Hosseini-Bandegharaei, & Chao, 2017). The different sorption processes are summarized in Figure 6. Ion exchange is the exchange of ions electrostatically attached on the internal or external surface of a material with ions in the aqueous or gaseous phase (Naushad & Alothman, 2013). Complexation is the formation of a chemical bond

with two or more other species wherein one species typically forms the central atom and focal point (Naja & Volesky, 2011). A prominent example of complexation is phosphate complexing on surface hydroxide groups (C.-j. Liu et al., 2007). Microprecipitation occurs when the adsorbent has local surface changes in pH or other conditions that decrease the solubility of the adsorbate (Naja & Volesky, 2011). Microprecipitation is alternatively called surface precipitation (Robalds et al., 2016). Not all processes occur simultaneously.

Figure 6 Classification of adsorption mechanisms (Robalds et al., 2016)

In terms of inter- and intramolecular forces/surface chemistry and adsorption, Table 2 summarizes how these mechanisms are related:

Table 2 Adsorption mechanisms and related inter/intramolecular force

Adsorption Mechanism Responsible Force(s)/Mechanism

van der Waals, dipole-dipole, Physisorption electrostatic attraction secondary valence

pH-pe Equilibrium, intramolecular Complexation (metals) force

Electrostatic attraction**, primary Covalent Binding valence

Equilibrium – intramolecular Ion Exchange* forces, ionic bonds

Equilibrium - precipitation at Microprecipitation surface, intramolecular

*Some do not consider ion exchange to be an adsorption mechanism, but a separate form of sorption ** van der Waals forces are also electrostatic attraction but originate from secondary valence forces Adapted from (Robalds et al., 2016; Hai Nguyen Tran, You, & Chao, 2016)

There are several analytical tests that can be done to determine the adsorption mechanism between an adsorbent and adsorbate; only some of the methods available to researchers will be discussed here. Scanning electron microscope (SEM) can be used to confirm internal surface area and may confirm if there is 19

microprecipitation/crystallization on the surface of the adsorbent (Farooq, Kozinski, Khan, & Athar, 2010). FTIR shows the formation of covalent bonds between a functional group and the adsorbate, however the precise bands must be known to draw conclusions. Raman spectroscopy, and surface-enhanced Raman spectroscopy (SERS) can be used to obtain a fingerprint of the chemical species (Altun, Bond, Pronk, & Park, 2017) on a surface by detecting reflected light from excited molecules within the 750-850 nm wavelength (Synetos & Tousoulis, 2018). If electrostatic attraction is the dominant adsorption mechanism, a large increase or decrease in sorption through surface charge can be observed by changing the pH of the solution and determining the pHpzc (Sposito, 1998) using the pH drift method or by measuring potential. Microprecipitation and complexation are also possible mechanisms that can be explored by conducting a pHpzc test. Finally, a method that is often used is a thermodynamic study of the adsorbent (Hai Nguyen Tran et al., 2016) which isolates for the bond energy of the reaction. The method below describes the standard method for thermodynamic studies.

The following equation describes the change in Gibbs free energy (ΔG) with respect to enthalpy (ΔH) and entropy change (ΔS):

훥퐺 = 훥퐻 − 푇훥푆 (1)

However, Gibbs free energy may also be represented by:

훥퐺 = −푅푇푙푛(퐾푐) (2)

Where 퐾푐 is the dimensionless equilibrium constant, 푅 is the universal gas constant, and 푇 is the temperature, in degrees Kelvin.

The van't Hoff equation is represented as follows by substituting equation 2 into equation 1:

훥퐻 훥푆 ln 퐾 = + (3) 푐 푅푇 푅

Which can be used to obtain ΔH and ΔS by linear regression.

The value of Gibbs free energy assists in explaining whether sorption is physisorption or chemisorption, with chemisorption typically having a magnitude of 80-400 kJ/mol(Saha & Chowdhury, 2011) and physisorption having an order of magnitude less, usually below 40 kJ/mol (Hai Nguyen Tran et al., 2016).

2.3.2 Adsorption Isotherms and Reversibility

The criteria of reversibility are important due to the nature of dynamic equilibrium, as discussed further in section 2.3.3. In this section, isotherms are introduced to provide background on the nature of dynamic equilibrium and reversibility. Dynamic equilibrium, in short, is the equilibrium obtained in a transient system, i.e. the initial concentration and flow rate may change over time. A material exposed to flowing water is in dynamic equilibrium as the water chemistry is constantly changing. Figure 7 illustrates an equilibrium established with parcels of water with two different phosphate concentrations and how it increases phosphate loading.

Figure 7 Phosphate loading onto spherical adsorbent increases as phosphate concentration increases

An adsorption isotherm shows the amount of adsorbate that is loaded onto the adsorbent material at a particular equilibrium concentration. Adsorbent loading is therefore a function of equilibrium concentration and increases according to classic theoretical isotherms such as the Langmuir isotherm (equation 4) (Langmuir, 1916) and Freundlich isotherm (equation 5) (Freundlich, 1907). If an adsorbent material fits the Langmuir curve, it is generally said to have a homogenous surface with a monolayer and limited sorption sites, although small deviations from these assumptions do not prevent the use of the Langmuir mode. If the adsorbent material instead conforms to the Freundlich isotherm, this can be indicative of a heterogeneous surface with multilayer adsorption and potentially unlimited adsorption sites as the equilibrium concentration, 퐶푒, increases (H. N. Tran et al., 2017).

퐾퐿∗푞푚∗퐶푒 푞푒 = (4) 1+퐾퐿∗퐶푒

1 푛 푞푒 = 퐾푓 ∗ 퐶푒 , (5)

Where 퐾푓 and 푛 are constants related to the Freundlich isotherm, 푞푚 is the maximum adsorption capacity and 퐾퐿 is the Langmuir constant.

During the literature review, two issues in communication and transparency in adsorption literature were noted. The first issue is that while adsorption loading describes the capacity of an adsorbent at a particular equilibrium concentration, adsorption capacity describes the maximum capacity of the adsorbent at some equilibrium concentration. Depending on the range of concentrations tested, the maximum capacity calculated using a theoretical curve can be very speculative (H. N. Tran et al., 2017). Furthermore, if using the Freundlich theoretical curve (which is logarithmic and has no “maximum” or boundary (Freundlich, 1907)), there is still some debate on how to calculate maximum capacity as the Freundlich constant, KF, is not necessarily the maximum capacity (Lu, 2008). The next issue is that if an adsorbent material happens to follow the Freundlich model and KF is speculative based on tested concentration ranges, there is poor comparability with other literature.

There exists a solution to this problem by way of using the term “practical loading” as defined by Kumar: the loading of phosphorus at an equilibrium concentration of 0.1 mg- P/L (P. S. Kumar et al., 2019). This allows for an isolated point on an isotherm to be compared between materials and addresses the fact that pragmatically surface water will have extremely low concentrations. Because of the difference between practical loading (q0.1 mg-P/g) and maximum capacity qm (mg-P/g) (Table 5), this term is important in defining the effectiveness of a material in surface water. Additionally, Kumar showed that materials with high maximum capacity do not necessarily have high practical loading as evident by his comparison of q0.1 and qm (P. S. Kumar et al., 2019). In-fact, there appeared to be no positive correlation between practical loading and maximum capacity. The conventional reporting of qm is not sufficient for comparing adsorbents intended for use in surface water as it does not provide a metric for effectiveness. Reversibility is described in literature by the separation factor, 푅퐿, and is

often calculated using the following rearranged Langmuir equation proposed by Hall et al. (Hall, Eagleton, Acrivos, & Vermeulen, 1966):

1 푅퐿 = (6) 1+퐾퐿퐶표

Where 퐶표 is the initial adsorbate concentration.

The following table was summarized by Tran et al. (H. N. Tran et al., 2017), adapted from Worch (Worch, 2012). It displays the reversibility of sorption in relation to the separation factor in terms of both Freundlich and Langmuir.

Table 3 Summary of reversibility constants and isotherm shapes (H. N. Tran et al., 2017)

Freundlich Separation Isotherm Defining Area Differential exponent factor shapes Equation to (Langmuir) Represent Slope

푑푞푒 n = 0 RL = 0 Irreversible Horizontal = 0 푑퐶푒

푑2푞 n < 1 RL < 1 Favorable Concave 푒 2 < 0 푑퐶푒

푑푞푒 n = 1 RL = 1 Linear Linear = 1 푑퐶푒

푑2푞 n > 1 RL > 1 Unfavorable Convex 푒 2 > 0 푑퐶푒

However, there is overlap in defining areas. For instance, Figure 8 shows how an isotherm with a horizontal portion is presupposed by a concave portion. The defining area is heavily dependent on the equilibrium concentration range tested. Note that in Figure 8 the horizontal shape is an approximation and does not have an exact horizontal slope.

Figure 8 General isotherm shapes showing portions describing reversibility

In environmental applications, using a separation factor or Freundlich constant to describe the entire isotherm is not immediately useful. In this thesis, the term “selective reversibility” will be used to describe reversibility, or the shape of the isotherm, in the concentration range found in the environment. Orthophosphorus can reach 3- concentrations as high as ~1.5 mg-PO4 /L (Appendix A), and thus the reversibility at this equilibrium concentration and lesser concentrations are of interest. As per equation

1, increasing 퐾퐿 decreases 푅퐿 which has an effect of translating the curve left and upwards in such a way that irreversibility is maximized.

In other words, irreversibility means that in response to changes in equilibrium concentration, 퐶푒, the adsorbate mass loaded onto the adsorbent, 푞푒, does not change. This means that the adsorbate does not release back into the solution and that the adsorbate mass is “irreversibly” sorbed until the equilibrium concentration is lowered

푑푞 such that 푒 is no longer equal to 0 and the shape of the curve changes. 푑퐶푒

2.3.3 Design Constraints and Criteria for Adsorbents Intended for Surface Water

Given the mobility of phosphorus in runoff, the dynamic equilibrium experienced in flowing water, and the lack of an adsorbent that targets this transport pathway, specific design criteria have to be made for an environmentally safe adsorbent. The following is a list of constraints for a phosphorus sorbent designed for deployment in a drainage ditch or similar surface water system:

I. Toxicity -- Adsorbent cannot be directly or indirectly toxic or otherwise harm the environment II. Stability -- Adsorbent must have mechanical strength required to withstand high flow rates, sediments, dissolution

Beyond these two constraints, the following criteria exist in order to potentially maximize cost-effectiveness and recovery of phosphorus:

3- ➢ Capacity and practical loading – adsorption capacity (mg-PO4 or mg-P/g) and practical loading ➢ Kinetics – equilibrium time and efficiency ➢ Reversibility and Regeneration – ability to desorb phosphorus in controlled temperature or pH changes and retain phosphorus in selected concentration ranges (withstand desorption in dynamic equilibrium)

As discussed in section 2.3.2, in dynamic equilibrium the adsorbate may desorb into the aqueous phase from the surface of the adsorbent when the concentrations in water decrease; this informs adsorbent design in natural systems, as bond energies need to be appropriately large to be irreversible in the presence of lower concentrations. On the other hand, recovering phosphorus from the adsorbent for reuse dictates that the process should be reversible in controlled conditions such as pH changes and temperature changes. Therefore, there is a duality between recovery and reversibility

that should be observed. The ideal phosphorus sorbent is nearly irreversible in the presence of concentration changes in surface water but should be recoverable in a controlled desorption process using temperature and/or pH changes. Classical definitions of reversibility/separation factors are usually only defined on adsorption isotherms and fixed-bed adsorption columns. The control of this duality is dependent on the adsorption mechanism. An explanation on selective reversibility is given below in Figure 9.

Figure 9 Graphic explanation of ideal selective reversibility on an adsorption isotherm

Mathematically, the maximization of the Langmuir constant, 퐾퐿, should shift the adsorption isotherm to the desired direction and minimize the separation factor, while also increasing maximum capacity. The Langmuir constant and maximum capacity, 푞푚, are correlated as per equation (4). While this holds true in theory, extending the analysis done by Kumar and calculating separation factors at an initial concentration of

3- 1.5 mg-PO4 /L (0.489 mg-P/L) tests the relationship between separation factor and practical loading.

14 12 10

P/g) -

6 (mg

0.1 4 Q 2 0 -2 0 0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8 0.9 1 RL at 0.49 mg-P/L

Figure 10: Relationship between the separation factor and practical loading at an initial concentration of 0.49 mg-P/L

Figure 10 shows that there is no apparent relationship between practical loading and separation factor based on data gathered by Kumar (P. S. Kumar et al., 2019). Some of the materials with the highest practical loading such as superparamagnetic ZrO2-Fe3O4 reached maximum capacity near an equilibrium concentration of ~5 mg-P/L, which is fairly low compared to other max capacities experienced at ~45 mg-P/L(Gu, Li, Xing, Fang, & Wu, 2018), ~100 mg-P/L (R. Li et al., 2016), ~30mg-P/L (He, Lin, Dong, & Wang, 2017). It does not seem viable to simply isolate for materials with high separation factors as this is not indicative of practical loading and requires a large range of inputs. Table 4 shows the relationship between design criteria and physi-, and chemisorption, in an attempt to isolate for a preferable mechanism.

Table 4 Relationship between adsorbent criteria and comparison between chemisorption and physisorption

Criteria Driving Mechanism Chemisorption vs. Physisorption Preference for Surface Water

Capacity Surface area, grain size, Physisorption can form multilayers (Ruthven, 1984) monolayer vs. multilayer saturation (Ruthven, 1984) Chemisorption is almost always a monolayer High capacity is preferable, a high surface area adsorbate uptake material is advantageous. Physisorption multilayers could potentially confer higher capacity Depends more on surface area (H. N. Tran et al., 2017) than adsorption mechanism (P. S. Kumar et al., 2019)

Equilibrium Grain size is chiefly responsible Physisorption is usually fast since activation energy is time (Worch, 2012), adsorption less than 40 kJ/mol, or no energy is required (Lowell et mechanism/activation energy al., 2012) (Jiang & Wen, 2011; Lowell, Low equilibrium time is preferable but contingent on irreversibility: nonactivated chemisorption may Shields, Thomas, & Thommes, Chemisorption is typically slower and highly 2012) grant low equilibrium times and desired temperature dependent (Hagen, 2015) , depends on irreversibility mechanisms involved, nonactivated chemisorption is fast and describes the initial fast uptake of adsorbate (Hill & Root, 2014)

Engineered Adsorption mechanism Physisorption is usually highly reversible as bond reversibility (adsorption energy) energies are weak (less than ~10 kJ/mol (Hill & Root, 2014)) Chemisorption is highly preferred as physisorption Chemisorption can be irreversible (Cooney, 1998; may desorb material during dynamic equilibrium Rouquerol, 2014) or reversible (Rouquerol, 2014) depending on the mechanism, activated chemisorption (usually more than ~80 kJ/mol (Hill & Root, 2014)), nonactivated is less

Based on Table 4, it seems that chemisorption with low equilibrium time (low activation energy “nonactivated” chemisorption) fulfills adsorbent criteria the best with higher irreversibility. Since capacity has been reported to depend more on surface area, the adsorption mechanism plays a lesser role in deciding this criterion.

2.3.4 Literature Review on Phosphorus Adsorbents

Kumar et al. conducted a review of phosphate sorbers that could achieve “ultra-low” concentrations (0.01 to 0.15 mg-P/L) (P. S. Kumar et al., 2019). Table 5 summarizes results on the performance of sorbents as well as mechanisms involved for each respective sorbent to obtain an idea of the most potent sorption mechanisms.

Table 5 Literature review on adsorbent materials and performance (P. S. Kumar et al., 2019)

Criteria Range of Top Performing Materials Dominant Sorption Mechanism Values Electrostatic attraction and inner- La(OH)3/Fe3O4 nanocomposites: 83.5 mg- sphere complexation P/g (Wu, Fang, Fortner, Guan, & Lo, 2017) (chemisorption)

Magnesium ferrite biochar composite: 128 Physisorption (van der Waals

Capacity (mg-P/g mg-P/g (Jung, Lee, & Lee, 2017) forces) of sorbent ~0.1 to 128 Electrostatic attraction, material) Magnesium oxide decorated magnetic precipitation with MgO, surface biochar: 121 mg-P/g (R. Li et al., 2016) inner-sphere complexation with ferric oxide (chemisorption)

3− Zinc iron zirconium composite: 93.5 mg-P/g Ion exchange of PO4 with OH (Drenkova-Tuhtan et al., 2017) (chemisorption)

3− Zinc-iron-zirconium hydroxide composite: Ion exchange of PO4 with OH, (~1 minute) (Drenkova-Tuhtan et al., 2017) (chemisorption) Similar zirconium composite oxidic Zr-O-P inner-sphere complexation adsorbents (Fang, Wu, & Lo, 2017) (chemisorption) Equilibrium time Minutes to (90% of total Lanthanum zeolite composite: ~1 minute La-OH based inner-sphere weeks loading) (He et al., 2017) complexation (chemisorption)

Alkaline-activated and lanthanum Lanthanum-hydroxyl group impregnated zeolite (NLZ): ~1 minute) (He, complexation (chemisorption) Lin, Dong, Liu, & Wang, 2016)

Superparamagnetic ZrO2@Fe3O4: 13.1 Zr-O-P inner-sphere complexation mg-P/g (Fang et al., 2017) (chemisorption)

Amine-functionalized copper ferrite Phosphate’s affinity for lanthanum Practical chelated with La(III) :12.5 mg-P/g (Gu et (Gu et al., 2018) (electrostatic Loading (loading 0.01-13 al., 2018) attraction, chemisorption) at 0.1 mg-P/L) Alkaline-activated and lanthanum- Lanthanum-hydroxyl group impregnated zeolite: 5.2 mg-P/g (He et al., complexation (chemisorption) 2016)

Based on Kumar’s literature review of phosphorus sorbents and the summarization of sorbent mechanisms in Table 5, nanocomposites made with lanthanum & zirconium oxides or hydroxides seem to dominate performance in kinetics, maximum capacity, selectivity (Fang et al., 2017; Gu et al., 2018; He et al., 2016; He et al., 2017; Wu et al., 2017), and practical loading. The high cost and potential toxicity of lanthanum (Balusamy, Tatan, Ergen, Uyar, & Tekinay, 2015; Gerber, Moser, Luechinger, Stark, & Grass, 2012) and zirconium (Guilhermino, Pereira, Diamantino, & Almeida, 2000; Harrisson, Trabin, & Martin, 1951) make it a less attractive option for use in the environment, however. In this case, it is possible that inner sphere complexation and electrostatic attraction chemisorption using less toxic metals may be a better strategy.

Comparing the toxicity of metals is a complicated process involving speciation and complex chemical models (Campbell, Tessier, & Turner, 1995). In some models, aluminum and copper were found to be one of the most toxic metals in acidic pH (<6.4) among Ba, Be, Cd, Co, Cr(III), Cs, Fe(II), Fe(III), Mn(II), Ni, Pb, Sr and Zn (Dong, Gandhi, & Hauschild, 2014), although this comparison had toxicity potentials that varied from 2.4 to 6.5 orders of magnitude with different water chemistry archetypes. Research with water chemistry archetypes with pHs between 6.4 and 8.2 show that the ecotoxicity of aluminum in freshwater systems has been characterized to be less than that of Cu, Ni, and Zn, as 90% of total aluminum precipitates as gibbsite (Gandhi & Diamond, 2018). Therefore, in water chemistries that reflect the above pH conditions (see Appendix A) aluminum may be one of the least toxic metals in terms of dissolved aluminum releasing into aquatic environments. Calcium is a fairly inert and non-toxic earth metal (Ross, Taylor, & Yaktine, 2011) and is found in the environment in soils. The abundance of calcium along with its necessity in the environment makes it a viable option for low-cost, non-toxic sorption.

Fine materials appear to be the most effective adsorbents in terms of sorption capacity. Plaster of Paris powder was chosen as it is a fine calcium-based material with potential

for recovery using an additional chemical (Appendix B) and likely sorbs via chemisorption due to phosphates affinity to calcium or ion exchange. A hydrogel bead made from natural polymers crosslinked with a metal may sorb via chemisorption (complexation, hydroxide ion exchange, or covalent binding) and has potential for phosphorus recovery and high internal surface area.

2.3.5 Selected Materials for Phosphorus Adsorption

The following materials were selected for testing based on the above literature review, analysis of research gaps, and advantages of low-cost adsorbents as well as natural polymeric adsorbents. Plaster of Paris powder was the first material chosen due to the low toxicity of calcium (Millero, Huang, Zhu, Liu, & Zhang, 2001; Ross et al., 2011) small grain size, and low cost. The effectiveness of calcium-based rocks was proven by a previous study by Lamont et al. (Lamont et al., 2018). A finer material could potentially obtain a higher maximum capacity and achieve a faster equilibrium. The recovery of a fine material could pose a challenge, so some methods of material coagulation or entrapment could be explored (Appendix B).

Hydrogel beads made from alginate, carboxymethylcellulose, and crosslinked by aluminum were chosen as aluminum is comparatively less toxic than other metals between pHs of 6.4 and 8.2 (Gandhi & Diamond, 2018), and hydrogel beads have surface area while maintaining larger grain size (Ahmed, 2015). The properties of aluminum could maintain irreversible chemisorption given reported mechanisms of ligand exchange for aluminum oxides (Xie, Lin, Li, Wu, & Kong, 2015) and similarly crosslinked beads (Wan et al., 2016).

Section 3 explores testing done to determine the adsorption capacity, kinetics, response to pH changes, and desorption of these two materials so that they could be better compared.

3 Phosphate Removal from Water Using Alginate/Carboxymethylcellulose/Aluminum Beads and Plaster of Paris Phosphorus Adsorption

3.1.1 Materials

Plaster of Paris (POP) powder sourced from DAP Products was purchased from a local market and used without further treatment. Based on the safety data sheet, the composition of the material is as follows: 10-30% CaCO3, 60-80% CaSO4·12H2O, and

0.5-1.5% SiO2. POP powder had an approximate grain size between 0.5 µm to 10 µm, as determined using an optical microscope (VHX-5000 digital microscope, Keyence). Sodium alginate (alginic acid, sodium salt, Acros Organics, ALG), potassium phosphate monobasic (crystalline, certified ACS), aluminum chloride hexahydrate (crystalline, certified ACS), HCl (ACS grade), KOH (pellets, ACS), KCl (crystalline, certified ACS),

CaCl2 (Anhydrous, pellets, 4-20 mesh), Na2SO4 (crystalline, certified) were purchased from Fisher Scientific. Carboxymethylcellulose (CMC, molecular weight ~90.000 g/moL) is a polymer used as food additive (Gómez-Dıaź & Navaza, 2003) and was purchased from Sigma Aldrich. Deionised (DI) water was used in all experiments conducted.

3.1.2 ALG-CMC (ACMC) Bead Preparation

In addition to POP powder, polymer-based sorbents were used. To prepare these sorbents, a solution of 10 g/L ALG and 16.65 g/L CMC were dropped into a solution of 3 wt% AlCl3·6H2O using a medical syringe and needle. Beads were also produced using 15 g/L ALG solutions, without mixing with CMC, as well as beads made solely of solutions of 15 and 30 g/L CMC. The polymer solution droplets promptly formed a bead upon contact with the AlCl3 solution forming an ALG-CMC (ACMC) hybrid hydrogel. The beads were soaked in the AlCl3 solution in which they formed for 24 hrs, and subsequently rinsed with a large excess of DI water. Beads used for sorption tests were

sealed in an air-tight container to prevent drying. A portion of the beads was instead dried in air and weighted before and after drying to determine its moisture content. The relative humidity of the laboratory in which beads were dried was approximately 50%.

Figure 11 Image of ACMC beads with ruler showing centimetres and approximate diameter of beads.

3.1.3 Shear Rheology Experiments

Shear rheology experiments were conducted to determine the mechanical strength of ACMC sorbents. ALG-CMC pugs used in shear rheology tests were obtained by pouring 20 mL of a solution of sodium ALG (10 g/L) and CMC (16.65 g/L) into 100 mL of 3 wt%

AlCl3·6H2O. Samples were immersed in DI water for a day after preparation, to verify that they would remain cohesive. Shear rheology experiments were conducted at 23°C using a rotational torque-controlled (i.e. combined-motor-transducer type) rheometer (MCR302 Anton Paar, Graz, Austria) with parallel plate geometry (diameter = 50 mm) at a constant frequency of 3 rad/s. The strain was ramped from 0.01% to 1000 %. Slip was minimized by using sand paper (600 grit).

3.1.4 Analytical Methods for Phosphorus Detection

A Hach Phosphorus (Total) Test ‘N Tube Reagent Set, a Phosphorus (Total) Test ‘N Tube Reagent Set, High Range Molybdovanadate and a Hach spectrophotometer were used to determine the total phosphorus concentrations. Water samples treated using POP powder were filtered using a Whatman GF/C 25 mm diameter glass microfiber filter to remove all particulate prior to determining the phosphorus concentrations. Water samples treated using ACMC beads were tested without further treatment. The amount of phosphorus adsorbed onto wither POP powder to ACMC beads was calculated as difference between final and initial phosphorus concentration in water.

3.1.5 Phosphate Sorption and Desorption Experiments

Sorption experiments were conducted using POP powder or ACMC beads at room temperature (23⁰C) and at pH = 3.5, 4.5, 7, 8.3, and 9.5 (adjusted using HCl and KOH). 3- Different concentrations of POP powder (1- 14 g/L) were added to a 90 mg/L PO4 solution and allowed to sorb phosphorus under quiescent conditions. Kinetic testing was 3- conducted using 12.5 g/L POP powder at 1 mg/L PO4 over different time periods ranging from 10 min to 60 min. Separate samples were prepared for each time tested, and all experiments were conducted in duplicate. Moreover, different amounts of wet 3- ACMC beads (6.7 to 26.7 g/L) were added to 2, 5, 10, 20, 50 mg/L PO4 solution of pH = 7 until equilibrium was reached. Kinetic testing was done over a time period ranging 3- from 4 hrs to 24 hrs, using 0.7 g/L ACMC beads at 2 mg/L PO4 and pH = 7. The effect of 100 mM KCl, CaCl2, and Na2SO4 on phosphorus sorption onto ACMC beads was also tested. At the end of each sorption experiment conducted in DI water, adsorbent materials were removed from solution. ACMC could be easily removed from water due to their size, whereas POP powder was allowed to settle and the supernatant was tested for its phosphorus content. The sorbent materials were then immersed in DI water for 24 hrs. The ratio of sorbent to DI water was 10 g/L and 16.7 g/L for POP powder and ACMC beads, respectively. During desorption experiments, the pH was 36

adjusted to either pH = 2.3, 4, 3.5, 7, 8.4, 9.5, or 11.6 (using HCl and KOH). At the end of desorption experiments, sorbents were removed from water and phosphorus concentrations were tested again. The sorption capacity (SC) was calculated based on the amount of phosphorus sorbed (Psorb, calculated as the difference between the initial and final phosphorus concentrations in water), the initial phosphorus concentration

(Pinitial), and the total volume of the solution (Vsol), using the following formula: SC=(

Psorb/Pinitial)·Vsol.

3.1.6 Modeling Approach

The removal efficiency (푅푒) of the adsorbent and the adsorption capacity (푞푒) were calculated using equations (7) and (8):

퐶푒 푅푒 = 1 − ( ) ∗ 100% (7) 퐶표

(퐶 −퐶 )∗푉 푞 = 표 푒 (8) 푒 푤 where 퐶푒 is the equilibrium concentration, 퐶표 is the initial concentration of phosphate in the solution, 푉 is the volume of the solution containing the reaction, and 푤 is the mass of the adsorbent. The non-linear isotherms used are explained in section 2.3.2.

Linearized and non-linearized forms of Langmuir type I, II and III, and Freundlich models were compared for best-fit and model parameters are extracted after best-fit analysis (X. Chen, 2015; Foo & Hameed, 2010). Equations (4), (5), (9) (Freundlich, 1907; K. V. Kumar & Sivanesan, 2006; Langmuir, 1916) were used:

1 ln (푞 ) = ∗ ln (퐶 ) + ln (퐾 ), linear Freundlich model (9) 푒 푛 푒 푓 where 퐾푓 and 푛 are constants and 푄 푚 is the maximum adsorption capacity.

The Hanes-Woolf linearized equation of Langmuir type I model (10) (Yamashita, Miyazaki, Nakamura, & Maeda, 2008), the Lineweaver-Burk linearized equation of Langmuir type II model (11) (Timmins, Lenz, & Fuller, 1997) and the Eadie-Hoffsiee linearized equation of Langmuir type III model (12) were also used:

퐶 1 퐶 푒 = + 푒 (10) Hanes-Woolf (linearized Langmuir type I) (Yamashita et al., 2008) 푞푒 푞푚∗퐾퐿 푞푚

1 1 1 1 = ∗ + (11) Lineweaver-Burk (linearized Langmuir type II) (Timmins et al., 푞푒 푞푚퐾퐿 퐶푒 푞푚 1997)

1 푞푒 푞푒 = 푞푚 − ∗ (12) Eadie-Hoffsiee (linearized Langmuir type III) (Zaheer, Bawazir, Al- 퐾퐿 퐶푒 Bukhari, & Basaleh, 2019)

where 푞푚 and 퐾퐿 are Langmuir constants.

The Microsoft Excel Add-in Solver was used to solve for parameters of nonlinear equations (H. N. Tran et al., 2017).

For modeling of sorption kinetics, the pseudo-first order (PFO) and pseudo-second order (PSO) equations were tested for best-fit analysis (equations (13)-(14)):

−푘1푡 푞푡 = 푞푒 ∗ (1 − 푒 ) (13) PFO (Lagergren, 1898)

푘2푞푒푡 푞푡 = (14) PSO (Blanchard, Maunaye, & Martin, 1984) 1+푘2푞푒푡 where 푘1 and 푘2 are rate constants related to the PFO and PSO rate equations, respectively. The time of the reaction is represented by 푡.

3.1.7 Statistical Assumptions for Non-linear Fitting of Isotherms

There are several assumptions are made before using the least squares function (linear regression) to circumvent fitting data on a non-linear function. These assumptions include the following by Osmari et al.: the normality of experimental deviations; the independent variables have an error of 0; the independence of experimental observations and co-variance of 0; and the experimental data must conform to a homoscedastic system (Osmari et al., 2013). It is generally accepted that the data used for isotherms does not conform to all these assumptions, although normality is the most important and widely checked.

The first assumption will be tested once experimental data has been obtained, and the Shapiro-Wilk test will be used to test for normality. These results, along with other error analyses may be found in Appendix C.

In the case of adsorption, there are many independent variables such as temperature, pH, and pressure that can present error in results. While the pH of solutions in experiments are controlled, the assumption that the reaction temperature at standard room pressure and temperature remain constant throughout the experiment should be a fair assumption. There is no reason to suspect that the room temperature or pressure changed drastically during adsorption reactions, although minute variations should be acceptable.

Experimental observations are made independently, which should logically conclude that there is a co-variance of 0. Samples prepared and tested on the same batch are subject to the same ambient conditions and the same condition of stock solution, but there is no dependence of samples to one another. Stock solutions were tested every batch of testing done.

Achieving a homoscedastic system with adsorption isotherms can be difficult as to progress rightward or leftward along the isotherm curve, some part of the experiment must be changed: either the initial concentration of the adsorbate must be changed (and therefore prompting a change in variance); or the mass of adsorbent material must be changed, which entails some difference in variance through in surface area variation. Additionally, temperature effects adsorption thermodynamics disproportionately at different equilibrium concentrations. This assumption is usually recognized as a poor assumption to make for non-linear regression in the context of adsorption isotherm fitting (Osmari et al., 2013).

3.2 Results and Discussion

The following section shows the results of testing for POP powder and ACMC beads.

3.2.1 Sorption of Phosphorus onto POP powder

The equilibration time for POP powder was evaluated at pH = 7, using 12.5 g/L POP 3- powder at 1 mg/L PO4 . This phosphorus concentration, along with all other used in this study, was based on concentrations reported for creeks and agricultural ditches (supporting information, Table A1). The amount of phosphorus sorbed was 3- approximately 0.06 mg PO4 /g after 10 mins, and changes were negligible over 24 hrs, as shown in Figure 12. It is noted that POP powder was allowed to settle before each measurement, and the supernatant was tested. Settling was fairly slow, and for this reason measurements were not conducted before 10 min. An equilibration time of approximately 10 min was previously reported for limestone (Lamont et al., 2018), whereas an equilibration time of approximately 20 min was reported for calcite powder (approximately 20 minutes) (Millero et al., 2001).

0.07

0.06

0.05 / g) /

4 0.04

0.03

(mg (mg PO

e e q 0.02

0.01

0 0 10 20 30 40 50 60 70 Contact Time (minutes)

3- Figure 12 Sorption of phosphate at pH= 7 by 12.5 g/L POP powder at 1 mg/L PO4 . The line is a guide to the eye.

The amounts of phosphate sorbed at equilibrium increased with increasing phosphate concentrations in water, as shown in Figure 12. It is noted that not all phosphorus concentrations tested are typical of surface and agricultural run-off waters. For instance, 80 mg/L phosphorus concentrations have been investigated for completeness, but are higher than those typically encountered on the field. Figure 12 also shows that the data were best described by a Langmuir isotherm (R2=0.986). This result supports the hypothesis of mono-layer, homogenous adsorption of phosphate onto POP powder with a limited number of sorption sites. The maximum sorption capacity of POP powder for 3- phosphate using the Langmuir model is 1.52 mg PO4 /g at pH = 7. Linearized Langmuir models and the Freundlich model provided a less accurate representation of the data compared to the non-linear Langmuir model, as reflected in the lower R2 values (Tables 7-8). Differences in the regression coefficients between linear and non-linear forms is

expected. This is because switching to the linear form distorts experimental error and introduces assumptions to the uniformity of error distribution (Foo & Hameed, 2010).

2 1.8 1.6 1.4

1.2

/g) 4 1

0.8

(mg PO (mg e

q 0.6 0.4 Freundlich Model

0.2 Langmuir Model 0 0 20 40 60 80 100 Ce (mg/L)

3- Figure 13 Adsorption isotherm for PO4 sorption with POP powder at pH= 7

Table 6 Nonlinear isotherm parameters for sorption of phosphate onto POP powder at pH= 7.

3- 2 qm (mg PO4 /g) Nonlinear Isotherm Model R KL Kf 1/n Langmuir Model 0.986 1.52 0.203 - - Freundlich Model 0.907 - - 0.325 0.380

Table 7 Isotherm parameters for sorption of phosphate onto POP powder at pH= 7.

2 3- Linear Isotherm R Value qm (mg PO4 /g) KL Kf 1/n Model Langmuir Type I 0.85 1.82 0.070 - - Langmuir Type II 0.60 0.51 0.21 - - Langmuir Type III 0.44 1.30 0.13 - - Freundlich 0.94 - - 0.094 0.80

Sorption of other anions such as fluoride using POP powder is also reported, although it is less effective than the sorption of phosphate (0.3 mg F/g POP powder, with Ce = 6 mg/L and a POP grain size of 0.43 mm at pH = 7) (Gopal & Elango, 2007). Phosphorus removal determined in this study using POP (10-30% CaCO3, 60-80% CaSO4·12H2O, and 0.5-1.5% SiO2, based on the SDS) was more effective compared to phosphorus removal reported in previous studies using gypsum (CaSO4). The poor sorption of phosphates onto gypsum were attributed to the negligible electrostatic attraction between sulphate and phosphate (Bolan, Syers, & Tillman, 1986; Spears et al., 2013). POP powder was also more effective than gypsum at sorbing fluoride anions (Gopal & Elango, 2007). The data suggest that the presence of calcium carbonate in POP powder played a major role in phosphate sorption, in agreement with previous findings (Lamont et al., 2018).

Sorption of phosphate onto POP powder was pH-dependent, and was greater at pH=8.3 than at pH = 7 or at pH = 4.5 (Figure 14). It is speculated that this is evidence of ion exchange occurring between phosphate and carbonate and/or sulphate. At high concentrations of phosphate, a high pH may cause the main reaction with phosphate to change from the adsorption to chemical precipitation of phosphate via formation of

Ca3(PO4)2. This hypothesis is supported by a previous study which used calcareous soils and showed that at a pH >8 and phosphate concentrations above 0.5 mM (15.5 3- mg-P/L or 47.5 mg/L PO4 ) precipitation was the dominant form of phosphate removal from water (Tunesi, Poggi, & Gessa, 1999). Chemical precipitation of fluoride is also reported in studies using POP powder (Al-Rawajfeh, Al-Hawamdeh, Al-Hawamdeh, & a, 2013; Al-Rawajfeh, Alrawashdeh, Aldawdeyah, Hassan, & Qarqouda, 2013).

It is noted that while POP powder could sorb phosphorus, it required over 10 min to settle under quiescent conditions. As a result, if POP powder were used on the field, it would require fairly large sedimentation basins. To address this shortcoming, ACMC beads were developed for phosphorus sorption, as further described in section 3.2.3.

100% 90% 80% 70% 60% 50% 40%

30% Removal efficiency Removal 20% 10% 0% 3.5 4.5 5.5 6.5 7.5 8.5 9.5 Initial pH

Figure 14 Effect of initial pH 7, 4.5, and 8.3 on removal efficiency of phosphate by POP powder.

3.2.2 Desorption of Phosphorus from POP powder

Recovery of phosphates following sorption is desirable, given the increasing need for this mineral (Sarvajayakesavalu et al., 2018). Phosphate desorption could also allow sorbent reuse.

Similar to sorption, desorption of phosphate from POP powder was pH-dependent, and increased with decreasing pH. However, it was below 35% at all pH values (Figure 15). In a study conducted with calcareous soil, multiple extractions of phosphate enriched soils were conducted to determine the fraction of inorganic phosphate desorbed from the soil in each stage of extraction (Y. Shen et al., 2019). The total amount of phosphorus extracted throughout the experiment ranged from 3-35% (Y. Shen et al., 2019), in agreement with this study.

40%

35%

30%

25%

20%

15% % P desorbed P %

10%

0% 2 4 6 8 10 12 Initial pH

Figure 15 Amount of phosphate desorbed from POP powder at varying initial pH. The dashed line is a guide

3.2.3 Sorption of Phosphorus onto ACMC beads

ACMC beads could be easily fabricated, as described in the materials and methods section. Solutions of either ALG, or CMC, or CMC and ALG formed beads upon contact with aluminum chloride solution, because Al3+ can crosslink ALG (Nokhodchi & Tailor, 2004) and CMC (Hosny & Al-Helw A Al-R, 1998). Trivalent cations are reported to form a three-dimensional bonding structure in networks of CMC and ALG (Kim, Park, Gu, & Kim, 2012). The moisture content of the ACMC beads was 96.360±0.001% (weight based). The sorption capacity of beads prepared using either CMC or ALG is given in Table 8. The effectiveness of the beads in removing phosphates was attributed to the presence of aluminum in their structure, because aluminum can bind phosphates (Pa Ho, 1975; Tanada et al., 2003). Differences between the sorption capacity of beads obtained with alginate only, with CMC only or with ALG and CMC, may be due to differences in the concentration of Al3+ incorporated in their structure.

Table 8 Sorption capacity of beads having different composition, at pH= 7. The initial phosphate

3- concentration in DI was 2 mg/L PO4

3- Constituents of Bead Gel immersed in 3 wt% Wet weight Dry Capacity (mg-PO4 /g

AlCl3·6H20 (g) Weight(g) dry) 15 g/L sodium alginate 0.2163 0.0031 14.4 30 g/L CMC 0.1924 0.0085 5.6 15 g/L CMC 0.1923 0.0129 3.8 16.7 g/L CMC and 10 g/L sodium alginate 0.1999 0.0080 4.2

Beads containing sodium alginate only had the highest sorption capacity. However, upon extended immersion in DI water, beads obtained with ALG only were mechanically weaker than beads to which CMC was added, as qualitatively assessed with the naked eye. Therefore, beads obtained with 10 g/L ALG and 16.65 g/L CMC were used in the rest of the study. This is because it is important to ensure that sorbent beads remain intact if immersed in agricultural ditches over extended time periods. The long-term stability of ACMC beads fabricated with ALG and CMC was monitored in DI water at pH = 7 (without phosphorus added) over 3 weeks. The size of the bead was measured before and after 3 weeks immersion in water. No changes in the size of the bead or dissolution was observed. Shear rheology confirmed the mechanical strength of the material prepared using 10 g/L ALG and 16.65 g/L CMC. The materials a shear elastic modulus G’ = 4.8·10 4±870.3 Pa and a shear viscous modulus G”=1.2·10 4±96.4 Pa in the linear region (up to approximately 1% strain, where G’ and G” were independent of strain). While this study was conducted in the lab, our future research will investigate the application and stability of ACMC sorbents on the field. ALG beads have been previously proposed as carriers for herbicides or immobilizing capsules for microbes, because they can be efficiently produced by dripping the ALG solution in a crosslinking solution (K. Lee et al., 2017) as was done in the study described here. Nonetheless, the poor long term stability of ALG beads has debilitated their widespread use (K. Y. Lee & Mooney, 2012). Previous studies have shown that anionic cellulose can form a dual interpenetrating network with alginate, improving the mechanical strength of ALG -

based crosslinked materials (K. Lee et al., 2017). Similarly, addition of CMC (anionic cellulose-derived polymer) reinforced alginate in this study. ALG -CMC beads crosslinked with Fe3+ were used for the delivery of therapeutic proteins (Kim et al., 2012). Addition of CMC to ALG improved the encapsulation efficiency of the beads, due to the formation of a strong 3D bonding structure (Kim et al., 2012). The equilibration time at pH=7 was longer with beads than with POP powder, possibly because of the slow diffusion of phosphates into the porous bead structure. A porous structure was previously reported for CMC- ALG beads (Kim et al., 2012). Fairly long equilibration times were reported for the sorption of phosphorus onto alginate/goethite composites (264 hrs, (Hanna, Artur, & Małgorzata, 2019)) and onto aluminum ALG - N- Isopropylacrylamide beads (14 hrs, (Wan et al., 2016)). Despite the long equilibration 3- times, even after 30 min the mg PO4 adsorbed/g sorbent were nearly doubled for ACMC beads (0.09 mg/g) than for POP powder (0.06 mg/g), as shown in Figure 16.

2.5

2 /g)

- 1.5

4 (mg PO (mg

e 1 q

0.5

0 0 5 10 15 20 25 30 35 Contact time (hours)

Figure 16 Sorption of phosphate by 0.7 g/L ACMC beads with initial phosphate concentrations of 2 mg/L

The sorption kinetics of phosphate onto ACMC beads at pH = 7 are best described by a PFO model (R2 = 0.991). The rate constant considering PFO kinetics is 1.2x10-3 min-1. The fit of the data to the PSO model was slightly worse than to the PFO model (with R2 = 0.988). Second order sorption kinetics were reported in studies conducted with aluminum-impregnated mesoporous silicates using 10 ppm phosphate concentrations (Shin et al., 2004) or with hydrated aluminum oxide modified natural zeolite with phosphate concentrations of 12.5 mg/L (Guaya, Valderrama, Farran, Armijos, & Cortina, 2015), whereas first order kinetics were reported for phosphorus sorption onto lanthanum/aluminum pillared montmorillonite with 10 mg/L of phosphate (Tian, Jiang, Ning, & Su, 2009). PSO kinetics were also reported for the sorption of phosphorus onto aluminum ALG - N-Isopropylacrylamide beads (14 hrs, (Wan et al., 2016)). Differences in the sorption kinetics reported for different aluminum-containing materials can depend on the intrinsic differences in the sorption mechanisms of phosphate on the sorbents and on the phosphate concentrations used (Azizian, 2004).

At equilibrium and at pH = 7, the amount of phosphate sorbed increased with phosphate concentrations in water, as shown in Figure 17. Similar to when POP powder was used, at equilibrium, phosphate sorption onto ACMC beads was best described by a non- linear Langmuir model (R2 = 0.993, Table 9). The amounts of phosphorus sorbed onto ACMC beads were significantly higher compared to those sorbed onto POP powder, even for phosphorus concentrations in water lower than those tested for POP powder. Based on the non-linear Langmuir model, the theoretical maximum capacity of the 3- beads is 90.5 mg-PO4 /g drytheoretical maximum capacity of the beads is 90.5 mg- 3- PO4 /g dry beads (corresponding to 29.5 mg-P/g). The Freundlich and the linear Langmuir models were less suited than the non-linear Langmuir model in describing phosphate sorption onto ACMC beads, as reflected in the lower R2 values (Tables 10- 11)

/g)

- 3 4 40

30 (mg PO (mg

e Langmuir Model q 20

10 Freundlich Model

0 0 5 10 15 20 25 30

Ce (mg/L)

Figure 17 Adsorption isotherm of phosphate and ACMC beads at pH = 7.

Table 9 Nonlinear isotherm parameters for sorption of phosphate onto ACMC beads at pH = 7

3 2 qm (mg PO4 /g dry basis) Nonlinear Isotherm Model R KL Kf 1/n Langmuir Model 0.993 90.5 0.08 - - Freundlich Model 0.975 - - 8.72 0.605

Table 10 Linear isotherm parameters for sorption of phosphate onto ACMC beads at pH = 7

2 3 Linear Isotherm R qm (mg PO4 /g KL Kf 1/n Model dry basis) Langmuir Type I 0.012 526 0.0059 - - Langmuir Type II 0.902 12.27 0.27 - - Langmuir Type III 0.44 9.69 0.18 - - Freundlich 0.895 - - 2.57 0.34

The sorption capacity of the ACMC beads used in the study at pH = 7 was greater than the one reported for calcium ALG beads (19.42mg-P/g wet beads at pH = 7, (C. Shen, Zhao, Liu, Mao, & Morgan, 2018)), possibly due to the stronger electrostatic attraction between phosphate and Al3+ than between phosphate and Ca2+. At pH = 7, the sorption capacity of the ACMC beads used in this study is greater than the one reported for N- isopropylacrylamide and aluminum ALG beads at pH = 3 (19.88mg-P/g, (Wan et al., 2016)). These differences can be due to the fact that the polymers used to produce sorbent beads can affect sorption efficiency and to differences in pH, since sorption is pH-dependent.

Phosphate sorption onto ACMC beads was pH dependent, and increased with decreasing pH (Figure 18). The pH of most surface water bodies is circum-neutral (supporting information, Table A2). At neutral pH, the phosphorus removal efficiency of ACMC beads was approximately 70%. It is speculated that this is because phosphate speciates into phosphoric acid at low pH. Phosphoric acid should have enhanced electrostatic attraction to protonated aluminum hydroxides, which are present in the crosslinked ACMC beads. When the pH is below the point of zero charge, ion exchange may also occur between aluminum hydroxides and phosphates (Li, Liu, Xu, & Qian, 2016). Adsorption onto calcium ALG beads also increased with decreasing pH (C. Shen et al., 2018).

100% 90% 80% 70% 60% 50% 40% 30% Removal efficiency Removal 20% 10% 0% 3 4 5 6 7 8 9 10 Initial pH

3- Figure 18 Effect of pH on phosphate removal using 16.7 g/L ACMC beads at PO4 = 2 mg/L. The dashed line is a guide to the eye.

While in this study beads were kept loose in the glass beakers during sorption tests, it is envisioned that they should be enclosed in meshes when used on the field. The proposed meshes will not impede flow, while facilitating the installation and recovery of ACMC sorbent beads (Figure 19). While the research presented here was conducted exclusively in the lab, future research will focus on field trials where the beads developed in the lab will be applied.

Figure 19 ACMC beads collected and contained in a cotton cloth that can be readily applied on the field.

3.2.4 Effect of Competing ions on Sorption of Phosphorus onto ACMC beads

At pH = 7, 100 mM KCl had a negligible effect on phosphorus removal using ACMC beads (16.6 g ACMC/L), whereas and CaCl2 addition slightly increased removal (by approximately 2%) and Na2SO4 reduced phosphate removal by approximately 36%.

The increased removal of phosphate with CaCl2 is attributed to the formation of calcium- phosphate complexes, in agreement with previous studies (Lamont et al., 2018). Reduced phosphate removal with hydrogel beads in the presence of divalent sulfate anions was previously reported to be more significant than the effect of monovalent chloride anions (X. Liu & Zhang, 2015). The effect of sulfate anions was ascribed to electrostatic repulsion between phosphate anions and the negatively charged surface sites (X. Liu & Zhang, 2015). Here, it is speculated that sulfate anions could compete with phosphates anions for sorption sites on the ACMC beads, explaining the reduced phosphate removal.

3.2.5 Desorption of Phosphorus from ACMC beads

Desorption of phosphorus from ACMC beads into DI water increased with increased pH and was as high as 60% at pH = 9.5 (Figure 20). At pH>9.5, ACMC beads dissolved

upon prolonged immersion in water, as qualitatively determined with the naked eye. Compared to POP powder, ACMC beads were better sorbents because there were more effective at sorbing phosphates at neutral pH (as previously discussed), and allowed greater desorption (up to 60% for ACMC vs. 35% at pH=2 for POP powder). Higher desorption than the one determined for ACMC beads was achieved for beads obtained with the synthetic polymer poly (vinyl alcohol), ALG and lanthanum hydroxide hydrogel (at pH = 14 desorption was nearly 100%) (Zhou et al., 2018) and for sorbents obtained with the synthetic polymer polyacrylamide, ALG and zirconium at pH = 13.3 3- (Luo et al., 2019). It is noted that sorption experiments were conducted in PO4 = 2 mg/L, while desorption experiments were conducted by immersing the beads in DI 3- water, following sorption. Here, desorption occurred due to the decreased PO4 concentrations in water. The reversibility of phosphorus sorption presents advantages and disadvantages. The main advantages are that phosphates can be recovered, and used for plant growth, and sorbents can be reused. The main disadvantage is that phosphorus could be potentially re-released when high phosphorus concentrations are followed by lower concentrations, e.g. when storm events leading to significant run-off are followed by drier periods. It is envisioned that ACMC beads should be collected shortly after large storm events, during which phosphorus concentrations will be highest, hence preventing the re-release of phosphorus. It is noted that at neutral pH desorption is not complete even in DI water with zero phosphorus concentrations. Therefore, if ACMC are not collected in a timely fashion their net sorption will decrease, but they should nonetheless be able to reduce the phosphorus concentrations in water. Future research should improve on the desorption of phosphate at non-neutral pH, while using sorbent beads that are completely natural and have good sorption capacity at neutral pH (typical of water bodies).

70%

60%

50%

40%

30% % P desorbed P % 20%

10%

0% 3 4 5 6 7 8 9 10 Initial pH

Figure 20 Amount desorbed from ACMC beads in DI water at varying initial pH. The line is a guide to the eye.

4 Conclusion

4.1 Phosphorus Sorption

Phosphate sorption onto POP powder and ACMC beads was explored. The non-linear Langmuir model accurately described the sorption isotherm with either POP powder and ACMC beads. At pH = 7 and in DI water, ACMC beads had a theoretical maximum 3- capacity of 90.5 mg PO4 /g, which is much higher than the sorption capacity of POP 3- powder (1.52 mg-PO4 /g). Sorption efficiency increased with increasing pH with POP powder, and decreased with increasing pH with ACMC beads. In DI water, the equilibrium time of phosphate sorption onto ACMC beads was approximately 24 hrs, while it was approximately 10 min with POP powder. The maximum percentage of

phosphate desorbed was higher with ACMC beads (60% at pH= 9.5, and lower at pH = 2.3). Monovalent salts (e.g. KCl) did not affect phosphorus sorption onto ACMC beads, whereas divalent anions (e.g. sulfate anions) hindered sorption. CaCl2 at 100 mM concentrations increased phosphorus removal, possibly due to the formation of calcium- phosphate complexes. ACMC beads are likely a better option for the removal of phosphorus from agricultural drainage water, because of their higher sorption capacity and because phosphorus desorption at alkaline pH is significant, allowing both phosphorus recovery and sorbent regeneration. Also, ACMC beads can be enclosed in meshes, allowing their easy installation in agricultural ditches and their recovery at the end of their lifetime. This is an advantage over POP, which can remove phosphorus from water but is fine and would not easily settle under flow conditions. Further research should focus on further improving the desorption efficiency of sorbent beads, while maintaining high sorption capacity. The effect of ions on sorption and desorption efficiency should also be investigated. Table 11 shows a final summarized comparison of the two tested materials.

Table 11: Comparison of materials with constraints and criteria with all factors determined

Material Toxicity Stability Capacity, Kinetics Reversibility & practical (equilibrium Regeneration loading time

Hydrogel Aluminum is Does not 29.5 mg-P/g Approximately Beads of ~3mm beads comparatively less dissolve or (max) 24 hours diameter, applied crosslinked toxic than other degrade of as per Figure 19 by metals in pHs 3 weeks in 2.26 mg-P/g aluminum between 6.4 and DI (section at P0.1 mg/L Desorption of 60% 8.2 (Gandhi & 3.2.3) at pH =~9, 35% in Diamond, 2018) DI at netural pH

Plaster of Calcium is fairly Does not 1.15 mg-P/g Approximately Fine grain size, Paris inert and found dissolve, (max) 0.17 hours difficult to apply (POP) throughout the remains in (Appendix B) powder environment(Ross physical 0.093 mg-P/g Desorption of 35% et al., 2011) form at P0.1 mg/L at pH =~9, 20% in DI at netural pH

4.2 Future Work

Future work with ACMC beads and POP powder should include testing in dynamic equilibrium (fixed-bed column) with an emphasis on replicating conditions in a drainage ditch. Additionally, ACMC beads should be tested for dissolution in extremely high pH (>10) and the speciation of phosphorus in the resultant mixture should be investigated. Improvement to the ACMC beads could include the incorporation of other metals into its structure, or incorporation of a biomass that could increase the overall adsorption capacity and irreversibility.

More feasible methods to apply adsorbents to drainage ditches accounting for erosional losses, turbidity, and other practical conditions should be researched. There is a great disconnect when it comes to adsorption literature and engineering applications, where most adsorption literature is limited in its testing scope despite having specified purpose. Part of the issue is not having well-founded testing standards for these practical conditions. In general, more non-toxic materials with high practical loading and low reversibility should be sought if a proper solution to surface water sorption is to be made.

These practical solutions must be tailored to each unique landscape, retention time, and water chemistry of watersheds. More research such as the one conducted by Penn et al. (Penn, Bryant, Kleinman, & Allen, 2007) using complex structures for drainage ditches (Figure 21) should be done. A potential idea for a multi-adsorbent system could be to design a drainage ditch system where water flow is divided into separate containers by depth. The depth would be based on the baseflow, storm events, and spring melt (Figure 21).

Figure 21 Drainage ditch system for phosphorus adsorption (Penn et al., 2007) with additional graphic of a potential multi-adsorbent system tailored to water stage and sediment profile

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APPENDICES

Appendix A

Table A1 reports examples of phosphorus concentrations found on surface water bodies, based on which the phosphate concentrations used in this study were selected. Table SI.2 lists pH values measured in sample water bodies. Calculation Basis TP (mg/L) Ortho- Location Source phosphorus (mg/L) 3-year average 0.10 0.043 Upstream of (London., 2017) Thames River Annual mean 0.33±0.37 0.21±0.20 Beaver Creek (Schilling, Kim, Jones, & Wolter, 2017) Event mean - ND 0.40 Village catchment (Lang, Li, & concentration (rainfall Yan, 2013) event) 3 year, flow-weighted - ND At least 0.1 Nearly half of (EPA, 2011.) annual average of 12- mg/L surveyed 25 samples agricultural streams (~60) January measurement 1.85 1.47 Livestock area (H. Nguyen & drainage ditch Maeda, 2016) Table A1 Data from literature for total and ortho-phosphorus from around the world. TP = total phosphorus.

Stream/Drainage Ditch pH Source Grand River 7.5-8.5 (GRCA., 2019.) Livestock-horticulture area (drainage 7.42-7.78 (Lang et al., 2013) ditch) Table A2 pH Ranges for potential sites.

Appendix B

The figure and table below show how POP powder could be used in conjunction with alginate to trap and retain POP powder for reuse. Use of this system still requires a containment system and diligent engineering for flow rates and dosing.

Figure B1 The optimization of the coagulating effect of alginate on POP powder solutions

Beaker Label I G H K J

Mass ratio of 12.5 10 7.5 5.0 2.5 alginate to POP powder (mg/g)

Table B2 The ideal ratio of alginate to POP powder for coagulation

Appendix C

Statistical Error and Normality Tests for Linear Fittings

The normality of the errors of the linearized datasets were analyzed using the Shapiro- 2 Wilk test and presented in Table C1 for linear fittings of Langmuir types whose R values were the highest, and for linear Freundlich fittings. Probability “p” values that are between those provided in tables are harmonically averaged. For the Freundlich test, the data was tested for lognormal distribution. As is standard, all tests were done at the 5% level of significance.

In addition to calculating the coefficient of determination, the sum square of errors (ERRSQ), otherwise known as the residual sum of squares will be calculated in order to compare linear and non-linear fitting methods.

The ERRSQ may be calculated via the following equation:

푛 2 ∑푖=1(푞푒푖 − 푞푒표) (15)

Where 푞푒푖 is the calculated adsorption capacity and 푞푒표 is the observed adsorption capacity.

Material Isotherm ERRSQ Shapiro-Wilk Distribution

ACMC Beads Linear Langmuir 7441 0.80 Non-normal Type 2

ACMC Beads Linear 132478 0.56 Non- Freundlich lognormal, non-normal

POP Powder Linear Langmuir 15.89 0.71 Non-normal Type 1

POP Powder Linear 17.24 0.55 Non- Freundlich lognormal, non-normal

Table C1 Linearized isotherm error and results of Shapiro-Wilk normality tests

Table C1 shows that for all experimental error data sets, normality was not achieved, which means that linear fittings may have poor fit and high inaccuracy. Fortunately, no conclusions or parameters for isotherms were calculated using linearized forms of Langmuir or Freundlich isotherms.

The figures below show error between observed and expected values, calculated as

"푞푒푖 − 푞푒표”. In Figure C1 and C3, for both materials, the linearization of the Langmuir isotherm in either type 1 or 2 underestimated the loading capacity at the respective 3- equilibrium concentrations portrayed. After an equilibrium concentration of ~2 mg-PO4 /L for ACMC beads, the linear Langmuir fitting underestimated loading heavily

compared to the nonlinear form. For POP powder, this occurred at a concentration of 10 3- mg-PO4 /L.

In Figure C2, the same underestimation occurs with the Freundlich linearization, but by comparing Figure C4 it shows that the higher coefficient of determination for Freundlich linear fit on POP powder data reduces error significantly.

0.4 0.2 0 -0.20.00 10.00 20.00 30.00 40.00 50.00 60.00 70.00 80.00 -0.4 -0.6 -0.8 -1

observed observed (mg/g) -1.2 -1.4 -1.6

Error in loading Error in loading between calculated and -1.8 Equilibrium Concentration at Observed Points (mg/L)

Linear Langmuir Type 1 Error Non-linear Langmuir

Figure C1 Error associated with using linear and non-linear fitting techniques for POP powder Langmuir adsorption isotherms

0.6 0.4 0.2 0 0.00 10.00 20.00 30.00 40.00 50.00 60.00 70.00 80.00 -0.2 -0.4 -0.6

-0.8 observed observed (mg/g) -1 -1.2

-1.4 Error in loading Error in loading between calculated and -1.6 Equilibrium Concentration at Observed Points (mg/L)

Linear Freundlich Error Non-linear Freundlich

Figure C2 Error associated with using linear and non-linear fitting techniques for POP powder Freundlich adsorption isotherms

0 0 5 10 15 20 25 30 -10

-20

-30

observed observed (mg/g) -40

-50

Error in loading Error in loading between calculated and -60 Equilibrium Concentration at Observed Points (mg/L)

Linear Langmuir Type 2 Error Non-linear Langmuir

Figure C3 Error associated with using linear and non-linear fitting techniques for ACMC bead Langmuir adsorption isotherms

1000

100

observed observed (mg/g) 1 0 5 10 15 20 25 30

0.1

Error in loading Error in loading between calculated and Equilibrium Concentration at Observed Points (mg/L)

Error Non-linear Freundlich Error Linear Freundlich

Figure C4 Error associated with using linear and non-linear fitting techniques for ACMC bead Freundlich adsorption isotherms