Clean Energy and Resource Recovery: Wastewater Treatment Plants as Biorefineries, Volume 2

By Manish Kumar

Clean Energy and Resource Recovery: Wastewater Treatment Plants as Bio-refineries, Volume 2, summarizes the fundamentals of various treatment modes applied to the recovery of energy and value-added products from wastewater treatment plants. The book addresses the production of biofuel, heat, and electricity, chemicals, feed, and other products from municipal wastewater, industrial wastewater, and sludge. It intends to provide the readers an account of up-to-date information on the recovery of biofuels and other value-added products using conventional and advanced technological developments. The book starts with identifying the key problems of the sectors and then provides solutions to them with step-by-step guidance on the implementation of processes and procedures. Titles compiled in this book further explore related issues like the safe disposal of leftovers, from a local to global scale. Finally, the book sheds light on how wastewater treatment facilities reduce stress on energy systems, decrease air and water pollution, build resiliency, and drive local economic activity.As a compliment to Volume 1: Biomass Waste Based Biorefineries, Clean Energy and Resource Recovery, Volume 2: Wastewater Treatment Plants as Bio-refineries is a comprehensive reference on all aspects of energy and resource recovery from wastewater. The book is going to be a handy reference tool for energy researchers, environmental scientists, and civil, chemical, and municipal engineers interested in waste-to-energy.

• Offers a comprehensive overview of the fundamental treatments and methods used in the recovery of energy and value-added products from wastewater
• Identifies solutions to key problems related to wastewater to energy/resource recovery through conventional and advanced technologies and explore the alternatives
• Provides step-by-step guidance on procedures and calculations from practical field data
• Includes successful case studies from both developing and developed countries

• Language: English
• Publisher: Elsevier Science
• Release date: Nov 10, 2021
• ISBN: 9780323901796

Book preview

Part A

Introduction

1. Wastewater to R3 – Resource Recovery, Recycling and Reuse Efficiency in Urban Wastewater Treatment Plants 00

2. Energy and resources recovery from wastewater treatment systems 00

Chapter 1

Wastewater to R3 – resource recovery, recycling, and reuse efficiency in urban wastewater treatment plants

Minh T. Vua, Luong N. Nguyena, Jakub Zdartab, Johir A.H. Mohammeda, Nirenkumar Pathaka, Long D. Nghiema,c

aCentre for Technology in Water and Wastewater, School of Civil and Environmental Engineering, University of Technology Sydney, Sydney, NSW, Australia

bInstitute of Materials Science and Engineering, Faculty of Mechanical Engineering and Management, Poznan University of Technology, Poznan, Poland

cNTT Institute of Hi-Technology, Nguyen Tat Thanh University, Ho Chi Minh City, Vietnam

Abstract

Wastewater is traditionally viewed as an unwanted material that must be treated prior to discharge to protect public health and the environment. However, due to dwindling natural resources, water utilities, engineers, and researchers have started to look at wastewater as a valuable feedstock for resource and energy recovery. Wastewater contains valuable resources such as water, nutrients, and energy – that, if recovered safely and effectively, can have economic and environmental benefits. Potable and nonpotable water recovery from wastewater gives an extra level of certainty and security to water supplies in the face of a changing climate. The recovery of nutrients (i.e., nitrogen and phosphorus) and energy emerges as another revenue for wastewater treatment plants. While water recovery has achieved an upward trajectory in both technology development and full-scale applications, the nutrient and energy recovery has recently received great attention from both the scientific community and industry stakeholders. This book chapter aims to provide a state-of-the-art perspective on the paradigm shift from wastewater treatment to wastewater R3 - reuse, recovery, and resource efficiency. Nutrients and embedded energy in domestic wastewater were reviewed followed by different technologies applied to recover nitrogen and phosphorus. Energy recovery through biogas production from anaerobic digestion and reactive nitrogen was also scrutinized. While several technologies have been reviewed separately in this chapter, the combination of various technologies is likely the key to achieve water, nutrient, and energy recovery.

Keywords

Wastewater, Reuse, Resource recovery, Energy recovery, Resource efficiency

1.1 Introduction

Radical changes in wastewater services are essential for a successful transition toward a circular economy given the importance of water in every aspect of our modern life. The conventional exploitation model of water resource has been linear (Mueller and Spahr, 2005). In this model, water is extracted from sources, treated, and used, and the impaired water (i.e., wastewater) is then treated and discharged into water bodies (Fig. 1.1). It has been reported that 80% of global wastewater has been released into the environment without adequate treatment (WWAP, 2017). Wastewater treatment is costly due to the consumption of energy, chemicals, and human resource. Moreover, without adequate treatment, this act can put a strain on the environment as well as cause such dramatic loss of natural resources in wastewater (e.g., water, energy, and nutrients). It is reported that 50%–100% of waste resources have been lost through wastewater discharge (Puyol et al., 2017).

Fig. 1.1 The schematic diagram of the FO-based system for nutrient recovery. FO , forward osmosis.

Given the depletion of natural resources (i.e., fossil fuels and natural mineral ores) (Ravago et al., 2015), the concept of circular economy has been conceived. In this concept, the amount of waste release is expected to be minimized and their reuse is expected to be maximized, thus promoting sustainable management of materials and energy (Puyol et al., 2017). From the circular economy point of view, wastewater is no longer considered as a problem but a solution and a resource from which other valuable resources (e.g., clean water, energy, and nutrients) can be recovered and recycled. Wastewater treatment plants (WWTPs) in the future will become a critical link in the circular ecosystem due to their integration of energy production and resource recovery during clean water production.

Under the circular economy perspective, recycling and recovery of valuable resources (e.g., energy, phosphorus, and nitrogen) from wastewater bring numerous environmental and socio-economic benefits. Towards the socio-economic aspect, phosphorus (P) and nitrogen (N) are essential ingredients for all living organisms on our planet (Hiet Wong et al., 2003; Liu and Chen, 2014; Zhang et al., 2014). Moreover, phosphorous and nitrogenous compounds are feedstock for many key industrial processes (Foct, 2003; Morton and Edwards, 2005). The fact is that the shortage of phosphorus reservation due to the over-exploitation of minable phosphate rocks for agricultural production (Daneshgar et al., 2019; Liu and Chen, 2014) has threatened food security and the operation of other industries. Hence, the recycling and recovery of phosphorus and nitrogen from wastewater to produce fertilizers can compensate for this depletion as well as ensure the conservation of natural ores for sustainable development. In terms of the environmental aspect, the excess of phosphorus and nitrogen in the aquatic environment can cause eutrophication and bloom algae (Bunce et al., 2018; Liu and Chen, 2014) as well as detrimental health impacts, such as the blue baby syndrome in infants caused by consumption of water rich in nitrate (Lusk et al., 2017). The recycling and recovery of phosphorus and nitrogen from nutrient-rich sources such as municipal wastewater (Dickinson et al., 2013; Xue et al., 2015), urine (Zhang et al., 2014), and especially digested sludge centrate (Vu et al., 2019) can be promising approaches to address the related environmental issues.

1.2 Nutrients and biochemical energy in wastewater

1.2.1 Phosphorus and nitrogen

The main sources of phosphorus in municipal wastewater are human excreta, phosphorus-containing household detergents, and industrial effluents. An estimation of 3–3.3 million metric tons of phosphorus in human excreta (i.e., feces and urine) and greywater enter in the influent of WWTPs (Mihelcic et al., 2011). In municipal wastewater, phosphorus is present as inorganic compounds (i.e., orthophosphate and polyphosphates) and organic compounds (Inoue et al., 2009; Liu and Chen, 2014; Lusk et al., 2017). The concentration of phosphate in wastewater ranges from 0.5 to 10 mg/L (Ramasahayam et al., 2014). The criterion for acceptable phosphorus concentration in water is 0.1 mg/L or less. With current methods of removal, only 30% of the phosphorus can be removed.

The environmental and socioeconomic issues associated with the mining activities of phosphate rock and inefficient uses of phosphorus have created an increasing need for phosphorus removal and recovery from other sources (e.g., wastewater). The production of elemental phosphorus is an energy-intensive process (Barber, 1989). To produce one ton of elemental phosphorus, approximately 14 MWh of electricity is required (Barber, 1989). Besides, if the current mining rate of phosphate rock remains constant, the global phosphorus reserves will be exhausted in 80 to 120 years (Bunce et al., 2018; Liu and Chen, 2014; Liu et al., 2008). This shortage of phosphorus significantly affects food security due to the reliance of agricultural production on phosphorus supply from phosphate rock. The exploitation of phosphate rock and the abundance of phosphorus in the aquatic environment due to discharge of wastewater and agricultural run-off into water bodies are major causes of ecological disasters (e.g., algae bloom and toxic heavy metal contamination in mining areas) (Bunce et al., 2018; Liu and Chen, 2014; Liu et al., 2008). Phosphorus removal in WWTP accounts for $44.5 billion in the US (Liu and Chen, 2014; Ramasahayam et al., 2014). On the other hand, the availability of phosphorus in WWTP provides a new opportunity to recover phosphorus. Central Europe estimated that the total amount of phosphorus in municipal wastewater can theoretically replace 40% to 50% of the annually applied mineral phosphorus fertilizer in agriculture (Egle et al., 2016).

Likewise, the main sources of nitrogen in WWTP influent are urine and fecal materials in which urine accounts for approximately 91% of the nitrogen loading (Lusk et al., 2017). Other nitrogen sources include food materials from kitchen sinks and dishwashers (Lusk et al., 2017). In domestic wastewater, the main nitrogen species are organic (40%) and ammonia (60%), with less than 1% nitrite and nitrate (Lusk et al., 2017). Total nitrogen concentrations in raw wastewater are typically 20–85 mg/L (Lusk et al., 2017), though higher values are reported in some areas.

The environmental and socioeconomic concerns caused by the synthesis of ammonia, and inefficient use and management of nitrogen encourage the recycling and recovery of nitrogen from wastewater. The Haber-Bosch process for nitrogen fixation is energy-intensive since it consumes about 1%–2% of the world’s total energy production (Kyriakou et al., 2020). Compensation for this energy consumption by reusing nitrogen from wastewater is desirable. Furthermore, the fertilizer production deteriorates climate change and global warming since the manufacture of nitrogenous fertilizers accounts for approximately 7% of the 9.9 billion tons of CO2 emitted globally per year (Chai et al., 2019). Even in WWTPs, the biological removal of nitrogen from the wastewater results in nitrous oxide (N2O) greenhouse gas emissions. This emitted amount accounts for 3% of the estimated total anthropogenic N2O emission, but it represents a significant factor (26%) in the greenhouse gas footprint of the total water chain (Flessa et al., 2002). Environmental impacts of nitrogen inefficiency in use include eutrophication, acidification, and direct toxicity to the natural environment (Hiet Wong et al., 2003; Lusk et al., 2017; Sharma and Bali, 2018). For example, it is reported that even at 1 mg/L nitrogen, an algae bloom can occur in a natural spring in Florida (Heffernan et al., 2010). Thus, recovering nitrogen from municipal wastewater before discharge can help address these environmental issues.

1.2.2 Potential energy from wastewater

The presence of organic materials and nitrogenous compounds in wastewater offers great opportunities to harness the embedded chemical energy towards energy neutral wastewater treatment. It is estimated that an average of 0.8 KWh is consumed to treat one cubic meter of wastewater (Singh et al., 2012), in which about 77% of the total energy consumption of sewage treatment is from the activated sludge process of nitrification to remove nitrogen (Zheng et al., 2017). This consumed amount of energy can be compensated from the embedded chemical energy in wastewater constituents (i.e., organic materials and reactive nitrogen).

It has been reported that 1 g chemical oxygen demand (COD) in domestic wastewater alone can produce an energy content of 17.8 kJ (Heidrich et al., 2011). Given the COD discharging rate of 60–120 gCOD/person/day and the total world population of 7.7 billion people in 2019, the total energy that can be harnessed from domestic wastewater can reach 3–6 x 10¹⁸ joules (Heidrich et al., 2011). This amount is equivalent to the energy content released from burning over 52–104 million tons of oil in modern power stations or the capacity of more than 12–24,000 world largest wind turbines working continuously (Heidrich et al., 2011). In addition to chemical energy from organic matter, the theoretical energy embedded in ammonium in domestic wastewater can account for 38%–48% of the total embedded chemical energy (Cruz et al., 2019).

The recovery of energy from embedded chemical energy in organic matter and reactive nitrogen in wastewater can be able to tackle environmental issues and energy security. The conversion of organic matter to methane followed by an upgrading process to produce biomethane for combustion can reduce the greenhouse gas footprint. In detail, the greenhouse gas emissions could be reduced by over 21% and 24% in comparison to diesel and petrol, respectively (Mac Kinnon et al., 2018). Biomethane contributes to the reduction of NOx and particulate matter local emission. In addition, the encouragement of biomethane production for domestic and industrial use can help nations reduce their reliance on natural gas imports, thus ensuring energy security.

Energy recovery from reactive nitrogen as N2O provides a number of benefits to society and the environment. First, the transformation from reactive nitrogen to nitrous oxide gas can be considered as a method to remove nitrogen from wastewater, thus reaping all benefits that nitrogen removal and recovery can offer as discussed above. Second, using N2O as an oxidant for combustion can completely burn biogas and increase the energy release from the combusting process. Finally, unlike the conventional nitrification-denitrification process with N2 gas as a finished product, converting reactive nitrogen to N2O can shorten the treatment steps of denitrification. Accordingly, this step leads to less consumption of reducing equivalents, less production of biomass, and possibly shorter retention time, thereby saving operational costs in the whole treatment process.

1.3 Technologies for phosphorus recovery

1.3.1 Chemical precipitation

Precipitation involves the usage of metal salts to react with dissolved phosphorus to result in insoluble precipitates (Puyol et al., 2017). These precipitates known as inorganic sludge will be settled and harvested by gravity, crystallized, or removed by filtration, thus phosphorus removal or recovery (Bunce et al., 2018). The most widely used metal ions sourced from soluble metal compounds to recover phosphorus include Ca²+ (lime), Fe³+ (iron chloride), Al³+ (aluminum sulfate), and Mg²+ (magnesium chloride) (Daneshgar et al., 2019; Daneshgar et al., 2018; Diamadopoulos et al., 2007; Doyle and Parsons, 2002; Larsdotter, 2006; Ramasahayam et al., 2014; Yeoman et al., 1988). These precipitations occur according to the following reactions (Daneshgar et al., 2019; Daneshgar et al., 2018; Diamadopoulos et al., 2007; Doyle and Parsons, 2002; Larsdotter, 2006; Ramasahayam et al., 2014; Yeoman et al., 1988):

The recovery can start directly in the aqueous phase with a phosphorus recovery rate of up to 50%, or via P-leaching of sludge and ashes with up to 90% recovery rate (Barca et al., 2014; Larsdotter, 2006). Most of the full-scale processes that recover phosphorus from WWTPs use waste activated sludge, digested sludge, and sludge centrate (Table 1.1) (Egle et al., 2016; Günther et al., 2018). The input wastewater capacity of these plants ranges from 100 to 5000 m³/day (Egle et al., 2016; Günther et al., 2018). This amount of wastewater can produce 0.5 to 11 tons of product daily (Günther et al., 2018).

Table 1.1

The efficiency of the recovery phosphorus process is dependent on the characteristics of raw wastewater, composition and concentration of chemicals used, operating conditions, and the set limits for effluent quality (Ansari et al., 2016; Bacelo et al., 2020; Daneshgar et al., 2018; Egle et al., 2016; Ramasahayam et al., 2014; Torit and Phihusut, 2019; Yeoman et al., 1988). An alkaline environment of wastewater is required for effective removal in most cases (Daneshgar et al., 2018; Günther et al., 2018; Ramasahayam et al., 2014). For example, the precipitation reaction using calcium occurs within the pH range of 8 to 11 (Yeoman et al., 1988). Due to the complex composition of wastewater, thus competitive reactions, the chemical dose may have to be increased. For instance, it has been reported that the Fe:P molar ratio of 2.9 was needed to achieve 90% phosphorus removal from municipal wastewater (Barca et al., 2014; Larsdotter, 2006). Of particular note, the alkalinity of wastewater plays an important role in phosphorous precipitation (Ramasahayam et al., 2014). For precipitation using lime, some high alkalinity wastewaters need three times as much lime for effective precipitation (Ramasahayam et al., 2014). For precipitation with aluminum, the alkalinity must be high enough to buffer the aluminum sulfate for effective phosphorus removal (Ramasahayam et al., 2014).

1.3.2 Adsorption

An alternative to precipitation is the adsorption of soluble phosphorus to active materials, which are then either used directly as fertilizers or recovered and reused. This phenomenon involves the movement of dissolved phosphorus from the wastewater to the surface or body of reactive components (e.g., calcium, iron, or aluminum) in adsorbents (Bunce et al., 2018). Many materials can be used as adsorbents, including locally sourced sands and gravel, furnace slag, fly ash, chemically modified clays, and biowaste materials (Bacelo et al., 2020; Hermassi et al., 2017; Jang and Lee, 2019; Jonidi Jafari and Moslemzadeh, 2020; Kawasaki et al., 2010; Kim et al., 2019; Li et al., 2018; Li et al., 2017; Makita et al., 2019; Wang et al., 2015; Yang et al., 2020; Zuo et al., 2018). Biochar that is produced by the thermal treatment of various organic substances with rich Mg and Ca content is beneficial for P sorption (Günther et al., 2018). The biochar after reacted with phosphorus can be composted and used as fertilizers.

The phosphorus adsorption process is governed by adsorbent properties, physicochemical characteristics of wastewater, and operating conditions of the adsorption system (Bacelo et al., 2020; Bunce et al., 2018; Hermassi et al., 2017; Kawasaki et al., 2010; Kim et al., 2019; Li et al., 2018; Makita et al., 2019; Zuo et al., 2018). Adsorbent properties, including surface area, porosity, surface charge, and surface functionality should be taken into account when selecting adsorbents for the adsorption process. Adsorbents with a large surface area, high porosity, and more reactive sites are favorable for phosphorus removal by the adsorption process. In addition, physicochemical properties of wastewater, such as phosphate concentration, temperature, pH, and presence of other ions/molecules can affect the phosphorus removal efficiency using this method.

1.3.3 Membrane-based processes

Membrane separation processes (i.e., forward osmosis [FO], membrane distillation [MD], and electrodialysis [ED]) have been recently adopted for phosphorus recovery from wastewater. The general principle to recover nutrients using membrane filtration is the enrichment of nutrients in wastewater due to the exclusive transport of water through the membrane and the excellent solute retention level of the membrane (Achilli et al., 2009; Holloway et al., 2007; Vu et al., 2019; Xie et al., 2014). After the membrane process, the nutrients (i.e., phosphorus) can be recovered using chemical precipitation (e.g., calcium phosphate or struvite) (Xie et al., 2014; Xie et al., 2016; Yan et al., 2018).

The phosphorus enrichment process in FO is osmotically driven (Fig. 1.1). In other words, the difference in osmotic pressure between two sides of the membrane causes the penetration of water from wastewater to draw solution, thus increasing phosphorus content in wastewater. Many studies have been conducted using FO to concentrate phosphorus-rich streams with high enrichment factors (Vu et al., 2019; Xie et al., 2014; Xie et al., 2016). It is reported that the concentration of sludge centrate using FO driven by MgCl2 draw solution could achieve a concentration factor of five, resulting in a high strength stream comprising ammonium (1210 mg/L), phosphate (615 mg/L) for subsequent nutrient recovery (Xie et al., 2014).

Unlike FO, the enrichment of phosphorus for subsequent recovery using MD is a thermally-driven process that can utilize low-grade heat to drive the separation of constituents in wastewater (Fig. 1.2) (Alkhudhiri et al., 2012; Alklaibi and Lior, 2005). In MD, the volatile components (e.g., NH3 and water) in the feed solution are converted to their gaseous forms that are transported across the membrane to the permeate side (Zarebska et al., 2014; Zhao et al., 2013), while all nonvolatile constituents (e.g., soluble phosphate) are retained in the feed side of the membrane (Alklaibi and Lior, 2005; Xie et al., 2016). The concentrated phosphate ions in the feed side are then extracted from the solution using other techniques (e.g., precipitation or adsorption). The use of the MD process to concentrate nutrients from wastewater (e.g., swine manure and urine) has been indicated in several previous studies (Thygesen et al., 2014; Zarebska et al., 2014; Zhao et al., 2013).

Fig. 1.2 The schematic diagram of the MD-based system for nutrient recovery. MD , membrane distillation.

The MD process mainly focuses on recovering ammonia from wastewater due to its high volatility, and the recovery of phosphorus is an accompanying process. The efficiency of the separation process is critically dependent on operating conditions of the system, such as pH and temperature (Yan et al., 2018). High pH and temperature of the feed solution can increase the transfer of volatile ammonia to the permeate side, therefore also resulting in the enhancement of phosphorus enrichment in the feed side (Yan et al., 2018).

In the ED process, the phosphorus enrichment is governed by electricity. This process is based on the interaction between cations (i.e., NH4+) and anions (i.e., PO4³-) with the cathode and anode in an electrical field to separate phosphorus or nitrogen from the bulk feed (Fig. 1.3) (Thompson Brewster et al., 2017; Ward et al., 2018; Xie et al., 2016; Yan et al., 2018; Zhang et al., 2013). The ED process is conducted in an electrical field with the use of ion exchange membranes including cation-selective exchange membranes (CEM), anion-selective exchange membranes (AEM), and bipolar membranes (Thompson Brewster et al., 2017; Ward et al., 2018; Zhang et al., 2013). In this regard, the direct current drives NH4+ and PO4³- ions to the cathode and anode chamber, respectively. PO4³- ions are enriched at the anode chamber, while NH4+ ions are driven to the cathode chamber for their enrichment (Ward et al., 2018; Xie et al., 2016; Zhang et al., 2013). It is demonstrated that the combination of the ED process and chemical precipitation can recover over 80% of phosphate under the form of calcium phosphate precipitates (Carrillo et al., 2020). It is also indicated that struvite precipitation using the phosphorus-rich effluent from the ED process can achieve high phosphorus recovery efficiency (93%) (Carrillo et al., 2020). pH, current density, and influent phosphate concentration are critical factors affecting the performance of the ED system for phosphorus recovery (Ward et al., 2018; Zhang et al., 2013). High pH of the feed solution can improve the phosphate enrichment in the ED process (Xie et al., 2016). High current density can enhance the transfer of phosphate across the ion exchange membrane, which facilitates the nutrient accumulation and recovery. Higher influent phosphate concentrations increased the phosphate concentration in the anode chamber.

Fig. 1.3 The schematic diagram of the ED-based system for nutrient recovery. ED , electrodialysis.

1.3.4 Microalgae-based processes

Microalgae cultivation using wastewater can be acknowledged as an indirect method to recover phosphorus from wastewater. The phosphorus removal and recovery mechanisms using microalgae are derived from the phosphorus assimilation of microalgae and phosphate precipitation occurring at high pH which is induced during microalgae growth (Acién Fernández et al., 2018; Larsdotter, 2006). Microalgae are photoautotrophic microorganisms that can be capable of using inorganic carbon (i.e., CO2), light energy, and nutrients to produce biomass via its metabolism, thus removing phosphorus from an aqueous solution. Microalgae biomass can be either directly applied to fields or used for other applications (e.g., biogas production coupled with secondary phosphorus recovery).

The phosphorus removal efficiency from wastewater using microalgae hinges on the extent of microalgae growth and the understanding of the factors that affect growth is therefore essential. The growth rate of microalgae is influenced by physical, chemical, and biological factors (Table 1.2) (Acién Fernández et al., 2018; Larsdotter, 2006; Ruiz-Martinez et al., 2012). It has been reported that the growth of Chlorella sp. declined when the concentrations of nitrogen and phosphorus reduced to 31.5 and 10.5 mg/L, respectively (Khan et al., 2018). Most microalgae species have an optimum temperature range of 20–30oC (Singh and Singh, 2015). pH ranged from 6 to 8.76 is preferable by most microalgae species (Singh and Singh, 2015).

Table 1.2

Microalgae have also been successfully cultivated in nonsterile environments such as wastewater with batch studies showing effective phosphorus removal (Chiu et al., 2015; Gao et al., 2016; Pittman et al., 2011; Wang et al., 2010). Up to 90% of the nutrients contained in wastewater could be recovered using microalgae-based processes (Acién Fernández et al., 2018). A complete microalgae system for nutrient removal and recovery consists of five steps (1) pretreatment of effluent, (2) recovery of nutrients and production of biomass in the photobioreactor, (3) harvesting of biomass, (4) treatment of used water for recirculation or disposal, and (5) transformation of the biomass into desirable products. The two main configurations for the cultivation of microalgae are closed and open systems (Larsdotter, 2006). Closed systems (e.g., covered raceways and tubular reactors) allow greater control of growth conditions, whereas open systems (e.g., shallow raceway ponds and circular ponds) largely depend on external factors and have contact with the open air (Larsdotter, 2006).

1.4 Technologies for nitrogen recovery

1.4.1 Chemical precipitation

Chemical precipitation to form magnesium ammonium phosphate salt (i.e., struvite) is an effective technology for recovering ammonium-N. The basic chemical reaction to form struvite for nitrogen recovery is similar to that for phosphorus recovery. In real applications, nitrogen and phosphorus recovery occur simultaneously via struvite precipitation, but the focus of this process is phosphorus recovery. Struvite can be used as a premium grade slow-releasing fertilizer due to its sparing solubility in water.

There are many factors, such as pH, Mg/N/P molar ratio, initial NH4+ concentration, and interfering ions that influence struvite precipitation (Çelen et al., 2007; Daneshgar et al., 2018; Doyle and Parsons, 2002; Günther et al., 2018). The formation of struvite is observed in the range of pH 7.5–10 (Sengupta et al., 2015). The ratio of magnesium to ammonium to phosphate is 1.6:0.6:1.0 is favorable for struvite formation (Sengupta et al., 2015). Struvite recovery efficiency can reach 85%–97% (Sengupta et al., 2015).

1.4.2 Air stripping

Nitrogen recovery via air stripping is based on the conversion of ammonium ions to ammonia gas before being stripped from the solution by air or steam flows (Kinidi et al., 2018; Sengupta et al., 2015; Yuan et al., 2016). This method relies on the following equilibrium (Sengupta et al., 2015):

Air stripping is highly dependent on pH (Sengupta et al., 2015). At pH around 9.3, soluble ammonium ions in an aqueous solution are converted to ammonia gas (Kinidi et al., 2018; Sengupta et al., 2015; Yuan et al., 2016). Different reagents, such as lime and caustic soda have been used to increase pH (Sengupta et al., 2015).

The stripping efficiency is affected by four factors including pH, temperature, the ratio of air to liquid volume, and liquid characteristics (ammonium concentration of feed, compositions, etc.) (Kinidi et al., 2018; Sengupta et al., 2015; Yuan et al., 2016). It is indicated that ammonia recovery efficiency increases from 80 to 92 % with pH increase of 8–11 (Sengupta et al., 2015). However, the air stripping is independent of pH when the temperature is higher than 80oC (Bonmatı́ and Flotats, 2003). The removal efficiency decreases significantly as air temperature decreases (Sengupta et al., 2015). For example, at 20oC, there is 90% to 95% ammonia recovery efficiency, while at 10oC, the efficiency decreases to 75% (Sengupta et al., 2015). In addition, air to water volume ratio significantly affects the performance of the ammonia stripping process. In industrial plants, the recommended volumetric air to liquid volume ratio is 600–700:1 and 95% efficiency of ammonia removal is expected (Katehis et al., 1998). Ammonia air stripping is preferable when ammonium concentration in wastewater is between 10 to 100 mg/L (Katehis et al., 1998). For higher ammonia content (more than 100 mg/L), it may be more economical to use alternate ammonia removal techniques, such as steam stripping or biological methods.

1.4.3 Ion exchange

The recovery of ammonium from wastewater using ion-exchange relies on the replacement of ammonium (NH4+) ions with cation ions on the surface of an ion exchanger (Hedström, 2001; Lin and Wu, 1996; Sengupta et al., 2015). Ion exchangers (e.g., zeolite) are insoluble resins that contain mobile exchangeable cation ions (e.g., Na+, K+, Ca²+, etc.) (Hedström, 2001; Lin and Wu, 1996; Sengupta et al., 2015). When the resin is in contact with the aqueous solution (e.g., wastewater), these cations dissociate and become mobile (Hedström, 2001). NH4+ ions in wastewater can substitute them on the exchanger providing the charge of the exchanger maintains neutral (Hedström, 2001).

After the ion exchange process, ammonium recovery can be achieved in two ways. Firstly, once the exchanger becomes saturated, the saturated exchanger may be directly applied to agricultural fields as a fertilizer (Sengupta et al., 2015). In addition, ammonium can be recovered from the NH4+-rich stream after the regeneration process by other technologies (e.g., struvite precipitation and air stripping) (Hedström, 2001; Sengupta et al., 2015).

The saturated ion exchanger needs to be regenerated for reuse. In chemical regeneration, a salt solution (e.g., sodium chloride) passes the ion exchange column in which the ammonium ions are exchanged and replaced by cations. The chemical regeneration can also be combined with biological regeneration (Sengupta et al., 2015). The regeneration phase results in a concentrated effluent stream of ammonium chloride (chemical regeneration) or sodium nitrate (biological regeneration), which offers opportunities for N recovery (Sengupta et al., 2015).

Natural zeolites are most frequently used as ammonium exchangers for wastewater treatment applications (Hedström, 2001). In nature, there are more than 50 different kinds of naturally occurring zeolites with different characteristics, such as clinoptilolite, ferrierite, mordenite, etc. (Hedström, 2001). Natural zeolites are generally aluminum silicate minerals with high cation exchange capacities and high ammonium selective properties (Hedström, 2001). As wastewater pass through the ion exchange column containing zeolite, ammonium (NH4+) is attached to the zeolite surface in exchange for Na+ or other cations.

Ion exchange processes are highly relevant due to their unique properties, such as high selectivity for NH4+, high removal, fast uptake kinetics and regeneration, less space requirement, and simplicity of application and operation, being environmentally friendly as it uses naturally occurring and easy-to-modify ion exchanger (i.e., zeolite) and releases nontoxic exchangeable cations (Na+, K+, Ca²+and Mg²+) (Hedström, 2001; Lin and Wu, 1996). The efficiency of ammonium recovery using ion exchange depends on operating conditions including homoiconic form and grain size of the exchanger, ammonium loading, ionic strength, pH, temperature, and scaling up phenomenon (Hedström, 2001). It is reported that using ion exchange process with zeolites can reduce NH3-N effluent concentration to less than 1 mg/L for around 130 bed volumes treated (Hedström, 2001).

1.4.4 Membrane-based processes

The membrane separation processes to recover nitrogen from wastewater is basically similar to those applied to recover phosphorus (i.e., FO, ED, and MD). Besides these processes, hydrophobic gas separation membranes are also discovered for soluble nitrogen recovery. Similar to the phosphorus recovery scenario, using FO and MD to recover nitrogen is based on their excellent solute rejection properties to concentrate the ammonium-rich streams before additional technologies (i.e., stripping or precipitation) are applied to recover soluble nitrogen under the forms of struvite or free volatile ammonia (Sengupta et al., 2015).

MD is highly recommended to be applied nitrogen recovery due to its unique properties towards volatile compounds (Sengupta et al., 2015; Xie et al., 2016; Zarebska et al., 2014). In MD, soluble nitrogen can be recovered in the permeate side of the membrane as once heated, NH3-N in the feed solution is converted to its gaseous form that is transported across the membrane and accumulate in the permeate side under the generated vapor pressure differential (Duong et al., 2013; El-Bourawi et al., 2007). The volatile ammonia can be recycled under the forms of ammonium salts when some acidic solutions are used on the permeate side (Sengupta et al., 2015). Operating conditions such as feed temperature, feed flow rate, and gas flow rate significantly influence the ammonia removal efficiency (Alkhudhiri et al., 2012; Zhao et al., 2013). High ammonia concentration and alkaline condition of source-separated human urine lead to high volatile free ammonia content and consequent significant ammonia transfer to the permeate through the hydrophobic pores of the MD membrane (Duong et al., 2013). A higher feed temperature, feed flow rate, and sweep gas flow rate promote ammonia removal efficiency and permeate flux (Duong et al., 2013).

The ammonia recovery of the hydrophobic gas-permeable membrane is based on the following core idea. Ammonium-ammonia balance shifts towards ammonia in the alkaline pH range, which is a soluble gaseous compound (Hasanoğlu et al., 2010; Nagy et al., 2019). Ammonia can pass through the membrane because there is always an NH3 concentration gradient over the membrane (Fig. 1.4) (Nagy et al., 2019). The gradient remains constant because the NH3 inside the membrane reacts by adding sulfuric acid (H2SO4) to form ammonium sulfate, rendering NH3 concentration inside the membrane to zero. Hasanoglu et al. (2010) were successfully recovered ammonia as (NH4)2SO4 on the permeate side of a macroporous hydrophobic membrane (Hasanoğlu et al., 2010).

Fig. 1.4 Gas-permeable membrane principle for ammonia recovery from wastewater.

1.4.5 Biological processes

Nitrogen recovery from wastewater could be completed by capturing soluble nitrogen in stabilized biosolids (Winkler and Straka, 2019). Biosolids involve growing microorganisms that assimilate nitrogen. Nitrogen captured in biosolids can be directly applied to fields compensating for chemical fertilizers. It is reported that half of all biosolids are recycled to land in the US (Winkler and Straka, 2019). However, prior to being applied to land, these biosolids need to be further treated using biodrying (Winkler and Straka, 2019). Biodrying is a drying process stabilizing and getting rid of pathogens in biosolids while generating ammonia-rich air (Winkler and Straka, 2019). This air can be scrubbed and recovered (Winkler and Straka, 2019). In the full-scale biodrying installment in Zutphen, biodrying is applied to treat 150 kton waste activated sludge concurrent with the recovery of 7.3 kton ammonium sulfate each year (Winkler and Straka, 2019).

Phototrophs (e.g., microalgae) are an attractive alternative to capture nitrogen from low COD/N wastewater because they can obtain additional energy from inorganic carbon and photon energy (Acién Fernández et al., 2018; Puyol et al., 2017; Ruiz-Martinez et al., 2012; Winkler and Straka, 2019). Therefore, the assimilation of nitrogen can occur with no or less organic carbon. Many phototrophs can grow heterotrophically in the dark (at night) (Winkler and Straka, 2019). Microalgae can grow photo-autotrophically with only water as the electron donor (Acién Fernández et al., 2018; Khan et al., 2018; Ruiz-Martinez et al., 2012). The most practical use for microalgae in wastewater treatment may be as a tertiary step to decrease nitrogen to low discharge levels without organic carbon addition. Studies have shown that microalgae can reduce nitrogen to very low levels in constant and diurnal light (Acién Fernández et al., 2018). Nitrogen recovery can be achieved by using microalgae biomass either directly to the fields or as feedstock for other processes (i.e., anaerobic digestion followed by secondary nitrogen recovery)

1.5 Energy recovery from organic matter and reactive nitrogen

1.5.1 Energy recovery via biogas production from anaerobic digestion

The conversion of organic matter in wastewater to biogas through anaerobic digestion followed by a purification process to produce biomethane has been widely acknowledged as an effective method to harness energy from wastewater in WWTPs. Biogas consists of methane (CH4) (50%–70%) and carbon dioxide (CO2) (30%–50%) and other trace gases, such as H2S, O2, volatile organic compounds, CO, and NH3 (Adnan et al., 2019; Bhatia, 2014; Molino et al., 2013). However, the purification of biogas is required due to its low energy efficiency caused by the impurity of biogas composition. After the purification process, biogas can be upgraded to biomethane involving CH4 (95%–99%) and CO2 (1%–5%) with no trace of H2S (Adnan et al., 2019; Molino et al., 2013). Biomethane has the same properties as natural gas in terms of heating value. The average expected energy content of biomethane is in the range of 31,600–37,700 Btu/m³ (Molino et al., 2013; Schmid et al., 2019). Under optimum digestion conditions in WWTPs, a methane yield of 510–640 nm³/t organic dry solids can be achieved (Paolini et al., 2018). Biomethane can be used in a variety of applications. Similar to the applications of biogas, biomethane can be utilized as fuel for hot water boilers, water pump engines, and electric generators to generate electricity (Adnan et al., 2019; Paolini et al., 2018; Schmid et al., 2019). It can be also used to fire incinerators or burned to heat the influent sludge during pretreatment (Bhatia, 2014; Paolini et al., 2018). Biomethane can be injected into a natural gas grid or used directly as vehicle fuel due to its similarity to natural gas in terms of energy value.

The adoption of anaerobic codigestion, preconcentration of wastewater using membrane filtration prior to anaerobic digestion, and the addition of carbon conductive materials into the anaerobic digester could be effective ways to optimize energy capture from wastewater. These methods improve energy capture through their impacts on biogas production. It is reported that the co-digestion of sewage sludge with beverage reject and food waste could enhance biogas production significantly (Vu et al., 2020; Wickham et al., 2018). Vu et al. (2018) demonstrated the feasibility of using FO membrane to concentrate organic matter in digested sludge centrate for subsequent energy recovery (Vu et al., 2018). Recently, the use of carbon conductive materials to increase biogas generation from anaerobic digestion has attracted more attention (González et al., 2018). It is revealed that using conductive materials in the anaerobic digester allows an alternative route for electron transport, thus improving the efficiency of the whole treatment process (González et al., 2018).

1.5.2 Energy recovery from reactive nitrogen

Energy recovery from reactive nitrogen in wastewater can be implemented through two forms of nitrogen that include ammonia (NH3) and nitrous oxide (N2O) gases. In other words, reactive nitrogen needs to be removed from wastewater and converted to these two forms for energy production. Similar to CH4, NH3 can be combusted to release power, or utilized as a transportable fuel as shown in the following reaction.

However, this idea is impractical given the higher cost of energy and chemicals for NH3 removal in comparison to the energy and value recovered. Recently, N2O has emerged as a renewable energy source. The capture of N2O followed by the combustion process is used to recover energy from wastewater. N2O is a potent oxidant and commonly adopted in propulsion and automotive applications, thus increasing energy recovery from methane. Combustion of CH4 with N2O can generate approximately 30% increase in heat, compared to with O2 as presented in the following reactions:

1.6 Conclusions

This chapter reviews recent literature to provide new insights into the ability to recover water, nutrients, and energy from wastewater. Water recovery for potable and nonpotable reuse has been successfully operated (e.g., NEWater in Singapore). Phosphorus recovery from wastewater also reached the commercial scale with a number of trade names. Chemical precipitation and crystallization are two main technologies. The development of membrane-based processes opens the opportunity to concentrate and enhance the phosphorus concentration for recovery. However, membrane-based processes are still at the infant stage. Results were mainly retrieved from lab-scale studies. Additional benefits including clean water from the membrane-based processes could enhance their commercial feasibility. Ammonia nitrogen recovery occurred simultaneously during the chemical precipitation of phosphorus (i.e., struvite). Organic carbon (COD) capture and anaerobic digestion of sewage sludge to produce biogas, which is used to produce heat and energy via combined heat and power. Recently, anaerobic codigestion (i.e., addition of other substrates with sewage) supported a number of energy self-sufficient WWTPs. Information corroborated in this chapter demonstrates that wastewater is a great resource for water, nutrients, and energy. Future research on combination processes to enhance the economic feasibility of recovery stream will have multiple benefits.

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Chapter 2

Energy and resources recovery from wastewater treatment systems

Varsha Bohraa, Kamal U. Ahamadb, Abhidha Kelac, Gaurav Vaghelad, Ashutosh Sharmad, Bhaskar Jyoti Dekad

aDepartment of Biology, Hong Kong Baptist University, Kowloon, Hong Kong, China

bDepartment of Civil Engineering, Tezpur University, Tezpur, Assam, India

cInstitute for Excellence in Higher Education, Bhopal, Madhya Pradesh, India

dDepartment of Hydrology, Indian Institute of technology, Roorkee, India

Abstract

Conventional wastewater treatment plants (WWTPs) are progressively looked upon as resource recovery facilities (RRFs), reflecting the worth of energy, nutrients and other resources, besides vindicating the required effluent quality. Though, WWTPs clean wastewater and lessen water pollution; but, while doing so, they also contribute to air pollution and need energy/material input with associated emissions. However, energy recovery (e.g., biogas, heat) and resource recovery (e.g., extracellular polymers, bioplastics, cellulose fibers, nutrients) allow us to counterbalance the negative environmental impacts of wastewater treatment. Several environment friendly approaches for resource recovery from WWT system have established their utility in optimizing WWTP operation to accomplish improved effluent quality at lower costs; they also constitute a useful tool to support the transition of WWTPs into water resource recovery facilities that maximize the valorization of products recovered from the wastewater. This article critically discusses the recent developments, opportunities, market possibilities, and barriers in the resource recovery from WWTP. Wastewater can not only dampen the effects of water shortages by means of water reclamation, but it also provides the medium for energy and nutrient recovery to further offset the extraction of precious resources.

Keywords

Advance oxidation processes, Biogas, Bioplastic, Cellulose fibers, Nutrient, Wastewater

2.1 Introduction

Freshwater is extensively used in almost every industrial process, generating considerable volumes of wastewater carrying high organic load and other valuable components. This wastewater when discharged into water bodies without suitable pretreatment results in severe contamination threatening human and ecological health (Schwarzenbach et al., 2010). In this regard, wastewater treatment plants (WWTPs) has played crucial role in pollutant removal and making wastewater suitable for safe disposal (Cheremisinoff, 2001; Mo et al., 2018; Wang et al., 2008). The WWTPs conventionally employs activated sludge process (ASP) to treat wastewater, wherein the organic fraction (COD) in wastewater is aerobically metabolized by microbes under controlled condition (Oh et al., 2010). Though the ASP succeeds in achieving legal effluent quality standards in terms of COD removal but fails to achieve sustainable resource recovery (Verstraete and Vlaeminck, 2011). The organic fraction in wastewater holds substantial quantity of chemical energy, recovering which is more sustainable action than destroying it through various treatment processes. As a result, WWTPs have moved from the notion of waste treatment, intended to release treated wastewater into surface waters, to the instigation of water resource recovery facility (WRRF). The primary objective of these facilities is to go one-step ahead of the conventional WWTP producing purified effluent and is related to the value-added resource recovery. WRRF can be planned with flexible technologies which allow substantial resource recovery when they are appropriately conceived. In previous few years, resource recovery alternatives in WWTPs have been hyped primarily for (1) recovery of mechanical, electrical and thermal energy for consumption both inside and outside the treatment plant and (2) recovery of materials for selling in commercial market. While recovery of renewable energy in the form of biogas production, hydropower generation or heat is much feasible, at least when energy is used instantly on-site in plant operations (McCarty et al., 2011), contrary to this recovery of materials offers some obstructions because of their intrinsic properties such as difficulty in handling and storing, contamination, constrictions in commercialization and very low public acceptance, as many residuals wastes still remains apparent (Verstraete et al., 2016). But even if the production and recovery of energy at WWTPs is presently receiving the much needed attention as it helps in accomplishing energy independency (Nowak et al., 2011), material recovery from sludge and wastewater (as nitrogen, phosphate, bioplastic production, and others) (van Loosdrecht and Brdjanovic, 2014) must not be discounted as they can play crucial part in evolving more sustainable services. With the signing of the Paris Climate Agreement, energy recovery in WWTPs has been given a renewed sense of urgency. Of lately, usage of wastewater as resources to recover and produce biofuels (such as bio-methane), nutrients (Nitrogen, phosphorus), material resources (bioplastics, cellulose fiber), and other value-added products has become very emphasized area of research (Atabani et al., 2019: Larriba et al., 2020). Even though wastewater RRFs have witnessed broad expansion in recent decades, their practical execution is still very poor. Majority of the WWTPs still emphasize on wastewater assortment and treatment rather than resource recovery. Irrespective of recurrent scientific output on technological resolutions from over a long period, the execution of full-scale resource recovery technologies remained restricted (Stanchev et al., 2017). The execution of resource-oriented processes can be difficult because of various technical and nontechnical bottlenecks which hampers the efficacious implementation of such technologies into wastewater treatment processes. Ambiguity persists regarding the selection of most resourceful techniques and ways to associate them exists in the way of establishing efficient WRRFs (Li et al., 2015). Among nontechnical tailback, the market potential of recovered resources, competition against preexisting cheap nonhazardous product and environmental impact of recovered products introduce further uncertainties (Van der Hoek et al., 2016). A strategic planning is necessary for the transition of WWTPs to RRFs. This chapter aims to explore various resource recovery approaches, resources recoverable from wastewater and their market possibilities and environmental impact associated with resource recovery. Moreover, global status and bottlenecks in efficient resource recovery from wastewater have been conversed.

2.2 Approaches for environment friendly resource recovery from wastewater treatment system

Although wastewater recycling and resource recovery technology has been widely described by the scientists in recent decades but large-scale use of WWTPs remains poor. Wastewater management plays a major role in the city's sustainable development. Traditionally, the purpose of wastewater treatment was to protect the users from health risks, however in recent decades, environmental protection by preventing nutrient pollution in surface water has also become an additional goal. As a result, nitrogen and phosphorus removal technologies have been used in WWTPs (Willy et al., 2009). The most widely used wastewater treatment technology is the ASP, in which aerobic microorganisms decompose the organic materials present in wastewater under constant oxygen supply. While the ASP process succeeds in meeting the legal standards, it is considered unsustainable due to its low availability of