Novel developments in biological technologies for wastewater processing

CHAPTER 8

Novel developments in biological technologies for wastewater processing
s Sauve ^tre†, Peter Schroeder†, John J. Milledge*, Elinor P. Thompson*, Andre Patricia J. Harvey* *

Algae Biotechnology Research Group, University of Greenwich, Medway, England Helmholtz Zentrum M€ uchen, Deutsches Forschungszentrum f€ ur Gesundheit und Umwelt (GmbH), Oberschleißheim, Deutschland †

Contents 1. Introduction 2. Aerobic systems 2.1 Activated sludge 2.2 Biological filters 3. Anaerobic systems 4. Systems combining aerobic and anaerobic processes 5. Integrating photosynthetic processes into wastewater processing 5.1 Constructed wetlands 5.2 Algal pond treatment systems 6. Concluding remarks References Further reading

239 240 240 242 243 249 249 250 256 268 269 278

1. Introduction Wastewater typically contains a plethora of chemicals and pollutants that need to be removed before the effluent may be reintroduced into the water cycle. Heterotrophic microorganisms are renowned for their ability to extract energy from and consequently degrade organic compounds [1, 2]. Consequently, biotechnologies for processing wastewater have classically sought to optimize environmental conditions that will accelerate their catalytic performance. These purely microbial-based biotechnologies can be relatively simple in engineering terms and process control, although increasingly sophisticated processes are being developed to minimize the pollution released to the environment. However, new processes are also being developed that take advantage of the tight cooperation and consequent synergistic effects between microorganisms and photosynthetic organisms—photosynthetic algae and higher plants—to degrade virtually all

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organic compounds. Cooperation introduces additional inputs of solar energy into these processes as well as additional degradative capacity [3, 4]. Identifying which combinations of these biotechnologies are the most appropriate for adoption for any particular wastewater stream is then reliant on the composition of the waste, climatic factors (temperature, solar insolation, etc.) and socioeconomic factors (land availability, funding, etc.). Key to the biological decomposition of organic waste is the general division into aerobic and anaerobic biological processes, including some systems, such as facultative or waste stabilization ponds, that take advantage of both aerobic and anaerobic processes in one unit operation [5–9].

2. Aerobic systems Oxygen is required for a healthy aquatic environment, and a significant polluting effect of wastewater in the environment is the depletion of dissolved oxygen (DO) by aerobic microorganisms that metabolize organic compounds to meet their energy requirements. Oxygen in water is maintained from atmospheric oxygen and that produced by photosynthetic organisms. If organic materials in wastewater are released untreated into a water body, bacteria break down most of the complex organic chemicals using the DO in the water. If the quantity of organic matter is small relative to the DO, then aerobic organisms metabolize the pollution via respiration without a damaging decrease in water oxygen levels. In contrast, large quantities of organic materials released into a watercourse lead to deoxygenation and the death of aerobic organisms along with unpleasant odors, perhaps most notably recorded in the 19th-Century Great London Stink [5, 6, 8, 10]. One method of removing organic matter from wastewater is to engineer systems and control conditions in order to increase the level of oxygenation and enhance the natural aerobic biodegradation processes. These biological treatments can be classified based on whether microorganisms grow attached to a substrate (fixed film) or in suspension. The two primary aerobic treatments are as follows: a) activated sludge (suspended microbial growth). b) biological (trickling or percolating) filters (fixed film)—the process is not filtration.

2.1 Activated sludge The activated sludge process was developed in Manchester by Lockett and Ardern who observed how the aeration of sludge caused the formation of macroscopic flocs and removal of organic matter [1]. It remains the most common biotreatment process for municipal and industrial wastewater [6]. Activated sludge microorganisms, primarily actively metabolizing aerobic heterotrophic bacteria capable of performing hydrolytic and oxidation reactions together with heterotrophic fungi, form flocs held together by their secreted polymers. A single floc (>1 mm diameter) contains billions of bacteria within this matrix. The tendency of the bacteria to form flocs allows the activated sludge

Novel developments in biological technologies for wastewater processing

process to function. Flocs also allow microbes to settle in a subsequent sedimentation tank for recirculation to the aeration tank [1, 2, 5]. Energy-intensive mixing and aeration support vigorous microbial activity, which, together with sludge recycling, enables high-rate aerobic decomposition of organics. This makes the aerobic activated sludge process highly “time-efficient” [3, 4]. Aeration is normally achieved by bubbling gas through the wastewater or by vigorous mechanical action [2]. Fig. 1 shows a surface aerator where the wastewater is “broken” into small droplets which are thrown outwards at several meters per second, in a turbulent plume in contact with atmospheric air resulting in the transfer of oxygen from the air to the wastewater. Such systems are characterized by high oxygen transfer rates, but aeration efficiency is nevertheless low (typically in the range of 1.0–2.0 kg O2 kWh
1) [5]. Haandel and Lubbe [1] have recently published an extensive handbook on the design of and recent developments in activated sludge treatments. The rising cost of energy and concerns about greenhouse gas emissions have added to the need to develop lower cost alternatives to current approaches, and for wastewater treatment plants (WWTP) that are energy self-sufficient [11].

Fig. 1 Mechanical surface aerator at Walvis Bay, Namibia.

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2.2 Biological filters Biological filters are aerobic treatment systems that utilize microorganisms attached to a solid medium to remove organic matter from wastewater [12]. Liquid effluent is trickled over a packed bed of support medium through which air is allowed to permeate (Fig. 2). The microbial communities, comprising bacteria, fungi, protozoa, and algae, grow as a film on the solid surfaces while nutrients and gases diffuse into this biofilm. As the wastewater passes down the filter bed, bacterial respiration again oxidizes organic compounds. The capacity of a filter is dependent on the surface area of film in contact with the wastewater and in general, the higher the surface area per unit volume the more efficient the reactor should be. Limitations are the void volume required to allow airflow and to prevent clogging as the support bed is colonized by the microbial film. Thus, solid packing for filters should not only have a high surface area per unit of volume, but also have sufficient porosity to reduce clogging and maximize airflow. The support packing material must also be low cost. Traditionally, natural materials such as granite and stone shingle, or waste products such as blast furnace slag, have been used as trickle-filter packing, the chosen substrates typically having particle sizes of 25–75 mm with a surface area of 100 m2 m
3 and a void volume of 50%. Synthetic “custom made” packings are now available. They offer a much better surface area to void ratio, with typical figures of

Fig. 2 Biological filter at Southern Water, Ashford, Kent, UK.

Novel developments in biological technologies for wastewater processing

90–150 m2 m
3 and 90%–95% void volume, allowing filters to compete more favorably, in terms of space requirements, with activated sludge systems. Since the 1960s, plastic material has been the packing of choice in the United States [2, 5]. Biofilm thickness, which may reach up to 10 mm, can limit diffusion of gases and nutrients in these biological filters, and in typical <2.00-mm-thick trickle-filter biofilms only the external 0.15–0.3mm may be aerobic. The environment inside the biofilm adjacent to the packing is therefore anaerobic and, if nitrate is present, denitrification of nitrate to nitrogen can occur [13]. The anaerobic conditions within the biofilm interior have also been implicated in abrupt and intermittent loss of biofilms, or sloughing, and the maintenance of adequate airflow in substrate packing is vital to prevent this [2]. To improve gas transfer and surface area to void ratio in wastewater attached-biofilm systems, the down-flow hanging sponge (DHS) reactor was developed. The DHS reactor is a novel trickling filter in which sponge (typically polyurethane) packing is used, and rather than being submerged, it is suspended in the reactor vessel, and wastewater trickles over the sponge. Oxygen then dissolves into the wastewater as it flows down and maintains DO concentration in the wastewater at a level exceeding that required by aerobes in the biofilm on the sponge. The DHS reactors offer the advantages of higher biomass concentrations, higher solids retention times (SRTs), lower hydraulic retention times (HRTs), and a smaller footprint compared with conventional treatment systems and do not require energy for forced aeration, unlike aerobic activated sludge [14]. Trickle filters have been in use for over 100 years [13]. They have been overshadowed by the activated sludge process because of the latter’s enhanced effluent quality, but trickling filtration still has a wide application due to its relatively low energy usage [15]. Other wastewater systems that have been developed are moving bed biofilm reactors (MBBR), biological aerated filters, rotating biological contactors, deep shaft processes, pure oxygen systems, biological fluidized beds, and pseudo-fluidized beds [1, 2, 5]. The choice of the system to be used is based on the wastewater composition, effluent quality requirements, land availability, capital investment, labor availability, sludge disposal, and location (and its climate and social factors) [1, 2, 5]. Some of these newer systems have had considerable commercial success. In the MBBRs, microorganisms grow on small, specially designed, plastic carriers with a density similar to water that are kept suspended in the reactor. More than 600 AnoxKalnes MBBR plants are in operation in over 50 countries [1, 16], operating aerobically and/or anaerobically.

3. Anaerobic systems The biological process of anaerobic digestion (AD) requires a reactor, which excludes air, and enables the flow of an organic feed through this reactor. The AD consists of a series of actions by different consortia of microbes that convert organic compounds into methane, carbon dioxide, and microbial biomass. The AD biogas comprises methane 50%–70%,

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Biodegradable organic material carbohydrates, fats, proteins a) Hydrolysis Simple soluble organics

b) Acid fermentation

Acetic acid

c) Acetogenesis

d) Methanogenesis

Propionic acid butyric acid long chain VFAs Acetic ccid

H2+CO2

Hydrogen-using methanogens

Acetoclastic methanogens

CH4+CO2

Fig. 3 Scheme of organic polymers degradation through the four stages of anaerobic digestion [18, 20, 21].

carbon dioxide 30%–45%, hydrogen <2%, and hydrogen sulfide <3.5%, the exact composition varies with the composition of the waste. There are four stages to AD [17–20] (Fig. 3): (a) Hydrolysis: carbohydrates, proteins, and fats are classified into monosaccharides, disaccharides, amino acids, and fatty acids. (b) Acidogenesis: acidifying bacteria convert hydrolysis products to short-chain organic acids. (c) Acetogenesis: acetogenic bacteria produce acetic acid, H2, and CO2 from fermentation products (“dark” fermentation is the fermentative conversion of organic substrate to hydrogen). The acetogens fall into two main groups: (i) Hydrogen-producing acetogens breakdown volatile fatty acids to CO2 and H2 (ii) [butyrate (butanoic acid)] CH3CH2CH2COOH + 4H2O ! CH3COOH +2CO2 +6H2. (iii) Homoacetogens: 4H2 +2CO !CH3COOH +2H2O. (d) Methanogenesis: end of the degradation chain, two groups of methanogenic Archaea, produce methane from acetate or hydrogen and carbon dioxide. (i) Acetoclastic methanogenesis (CH3COOH!CH4 +CO2). (ii) Autotrophic or hydrogenotrophic methanogenesis (4H2 +CO2 ! CH4 +2H2O). The AD is generally the process of choice for processing biomass with a high water content, and for the production of biomethane in the so-called biogas [22–24]. The application of anaerobic biotechnology for biomethane production dates back to at least the 10th century as a method of heating water by the Assyrians [25]. In the UK, biogas from a

Novel developments in biological technologies for wastewater processing

sewage treatment facility was used to fuel lamps in the city of Exeter as early as 1895, and by 1922, a Birmingham sewage works had an engine running on sewage gas [26]. Early digester designs were developed at the end of the 19th and start of the 20th century for sewage sludge digestion and were simple large tanks with some form of mixing (mechanical or gas mixing) [25] known as continuously stirred tank reactors (CSTRs). The CSTRs typically have long hydraulic retention time (HRT) defined as the volume of the digester (m3) divided by the daily feed rate (m3 day
1) of 20–30 days [2, 21]. In CSTR the SRTs are identical to the HRT, and these reactors are widely used for treating liquid wastes with up to 10% solids [2, 27]. The CSTRs are the simplest and still the most common configurations used worldwide [21, 27–29]. Until the mid-20th century, anaerobic digesters were considered too slow to be useful in treating large volumes of wastewaters [3]. The unsuitability of the conventionally mixed digesters for the treatment of industrial wastewaters of low strength and largely soluble organic composition, led to the concept of biological solids recycling and the retention of active biomass within the digester [2, 21, 29]. All modern high-rate AD wastewater treatment processes do this [30]. In the middle part of the 20th century, anaerobic processes were developed where the solids (anaerobic biomass) were separated from the liquid and recycled, much like an activated (aerobic) sludge process; this reduced the HRT down to as little as 5 days, hence the smaller size of the reactor. A further innovation in this period was the development of anaerobic “filters” where the biomass grew as a biofilm on stationary media, further reducing HRT to as low as 1–2 days [2, 5, 21, 31]. In both these types of systems the sludge retention time is longer than the HRT, as the sludge is retained in the reactor by using internal settlement systems or external settlers with sludge recycling, or “fixation” of biomass on a support material [28]. The upflow anaerobic sludge blanket (UASB) reactor, developed by Lettinga, Roersma [32], was a significant breakthrough in the 1970s that relies on the biomass granulating into particles or clusters of microorganisms (Fig. 4) which reduces HRT even further to <24 h [2, 21]. The key element of the UASB reactor’s performance is the quality of its sludge granules. Some wastes readily granulate (e.g., sugar-processing waste and wastes containing mainly volatile acids), but other wastes develop sludge granules slowly and some not at all [4]. Nonetheless, UASBs have been shown to be a robust option for high-strength organic wastewaters, despite their often lengthy and complex startups [4, 33]. UASBs are now widely favored in developing countries for sewage treatment because of their low energy use, easy maintenance, and cost-effectiveness, but are increasingly combined with other treatment technologies such as down-flow hanging sponge aerobic reactors to ensure quality effluent [34]. Daud and Rizvi [35] have recently reviewed UASB technology for domestic wastewater treatment and the effect of operational parameters, and Saleh and Mahmood [30] the characteristics of CSRT, UASB, and anaerobic filter reactors (summarized in Table 1).

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Fig. 4 Anaerobic granulated sludge. Table 1 Characteristics of AD reactors [30] Channel flow Anaerobic Time required for potential reactor type start-up (days)

CSTR UASB Anaerobic filter

4–16 3–4

None Low High

Biofilm carrier

Loading rates (kg COD m3 day21)

HRT (days)

No No Essential

0.25–3 10–30 1–4

10–60 0.5–7 0.5–12

Around the time of the development of UASBs, the anaerobic baffled reactor (ABR) was also designed. The ABR consists of a sequence of baffled chambers, and is simple; easy to construct; robust in terms of operation; offers low HRTs; is very stable to shock hydraulic and organic loads; and reduces losses of biosolids [4, 21]. The submerged anaerobic membrane bioreactor (SAMBR) uses membranes submerged in the digester to separate out the biomass from the effluent. This design enables the HRT for soluble wastewaters to be reduced to as low as 3 h, although this design is quite complex relative to the ABR and requires considerable capital investment because of membrane cost [36, 37]. The major drivers for the wider adoption of membrane AD bioreactors are the capacity for removing high organic loads; low process energy requirements; energy recovery in the form of biomethane; and low sludge production, but SAMBRs have the disadvantage of operational costs for the membrane cleaning and replacement due to fouling [38].

Novel developments in biological technologies for wastewater processing

The treatment of municipal wastewater accounts for 3% of global electricity consumption and 5% of global greenhouse gas emissions, with typical energy demand for biological wastewater treatment 20–30 kWh person
1 year
1 [39]. Consequently, there is considerable interest in reducing these energy costs. In anaerobic digestion, energy in end-product partitions into biomethane and accumulated microbial cell biomass or sludge, with the latter considerably lower in amount (typically 3% of biological oxygen demand or BOD) than that remaining under aerobic conditions (15% of BOD). Consequently, wastewater processed by anaerobic digestion is seen as having more potential to reduce energy costs than aerobic systems [2–4, 8, 33]. Tauseef and Abbasi [4] compares modern aerobic and anaerobic wastewater treatment processes (Table 2). Interest in the production of biomethane from wastewater and other wastes and biomass, therefore, continues to grow, as illustrated by the fact that biogas-producing plants in Europe have tripled in number from 2009 to 2015 [40], with a high proportion of biogas produced in AD plants treating municipal wastewater [41] (Table 3). In the United States, 48% of the total wastewater flow in 2013 was estimated to be treated by AD [42], and this treatment method is distributed across the entire country (Fig. 5). In China, there is also renewed interest in the use of AD, 25% of municipal wastewater treatment plants currently having AD units installed to stabilize sewage [43]. A number of projects co-digest wastewater and other waste streams, particularly food wastes, to produce biogas [39, 44]. Sewage sludge can have a low organic material content, whereas food waste can be rich in organic compounds, albeit variable in composition and high in lipids [45, 46]. Co-digestion of two or more organic waste streams can optimize the biomass composition for methane yield, so that methane yields can be 18% Table 2 A comparison of modern aerobic and anaerobic wastewater treatment processes [4] Operational factor Anaerobic Aerobic

Energy requirements Extent of loading possible Degree of treatment Sludge production Process stability (toxic compounds and load changes) Start-up time Odour problems Energy production Effluent quality

Low High to very low High (>90%) Very low Good

2–4 weeks Low, as the systems are air-tight Yes Generally contains higher suspended solids and nitrogen in the form of ammonia; requires aerobic “polishing”

Much higher Moderate to very low (>95%) Much higher Good

2–4 weeks Low, despite systems being largely open No Relatively better stabilized and fit for discharge

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Table 3 Biogas production from WWTPs in various European countries WWTP Biogas Production Country

Year

GWh year21

% of total biogas production

Denmark Finland France Germany Norway Sweden Switzerland Netherlands UK

2012 2013 2012 2014 2010 2013 2012 2013 2013

250 126 97 3050 164 672 550 711 761

21 22 8 7 33 40 49 20 11

Fig. 5 Wastewater AD plants in the USA producing biogas. Courtesy of American Biogas Council.

higher from the combined waste relative to the separate digestion of the sewage and the food wastes [45, 47]. The “POOBUS,” operating in Bristol, UK, is a 40-seat bus that runs on biomethane generated from the AD of sewage and food waste together. A single passenger’s annual food and sewage waste can fuel the bus for 60 km and releases up to 30% less CO2 than conventional fuel [44]. Co-digestion of food waste with wastewater sludge is advantageous where there is spare digestion capacity in a wastewater treatment plant, and can deliver significant environmental benefits. However, there are challenges with co-digestion, such as food waste collection and processing, so co-digestion applications are concentrated mostly in countries where there are financial incentives together with favorable energy and waste management policies [39].

Novel developments in biological technologies for wastewater processing

4. Systems combining aerobic and anaerobic processes Most modern wastewater treatment plants feature both aerobic and anaerobic unit operations, such as the Southern Water plant at Ashford, UK (Fig. 2), which features trickle filters for primary treatment with anaerobic digestion of the sludge produced. Significant quantities of sludge are generated by the aerobic treatments: aerobic activated sludge treatment can generate as much as 0.3–0.5 kg dry biomass for 1 kg of COD (chemical oxygen demand) reduction depending on the organic matter content [11]. In the UK 75% of sewage sludge is currently processed by AD [48]. Fig. 6 shows how the amount of sludge from aerobic WWTPs is reduced using anaerobic digestion. New combined anaerobic and aerobic processes are being increasingly used, giving further improvements in wastewater treatment and, in particular, a reduction in nitrogen and phosphorus levels [49]. Many systems characterized as primarily aerobic or anaerobic in fact feature both modes of biological treatment. The Amtreat process is a compact, high-rate activated sludge process developed in 1992 for treating high-strength ammonia wastewater, with typical ammonia removal rates in excess of 97%, which also features an integral anaerobic treatment phase [50]. The DHS processes that are primarily aerobic have also been developed with a submerged anaerobic zone to enhance denitrification [51]. Both Chan and Chong [49] and Tauseef and Abbasi [4] have discussed and reviewed the anaerobic-aerobic treatment of industrial and municipal wastewater.

5. Integrating photosynthetic processes into wastewater processing It is also possible to combine anaerobic and aerobic biological breakdown in one system by exploiting the ability of plant cells to photosynthesize and release oxygen for microbial

Fig. 6 COD mass flow through a conventional activated sludge domestic wastewater treatment plant with AD of sludge [11].

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aerobic metabolism of organic matter at the same time reducing the concentration of CO2 released in biogas from microbial anaerobic metabolism. These systems are exemplified by constructed wetlands (CWs) on the one hand, and algal pond treatment systems on the other, and sit at the forefront of innovations in contemporary wastewater treatment aimed at the food-water-energy nexus. [52].

5.1 Constructed wetlands The CWs are human-made copies of natural wetlands that use biological processes involving wetland vegetation, soils, and their associated microbiota to improve water quality [53–56]. Promoted and provided with a scientific background by two German scientists since the 1980s, these artificial pond systems have undergone significant improvements due to several EU COST (Cooperation in Science and Technical Research) actions focusing on their technology development. The CWs have five main components [57]: a) substrates with various hydraulic conductivities (ease with which water can move through pore, spaces, or fractures of the substrate). b) vascular plants adapted to water-saturated substrates mainly anaerobic substrates. c) water flowing in or above the substrate. d) aerobic and anaerobic microorganisms. e) invertebrates and vertebrates. The CWs are low cost, relatively easily maintained, reliable systems of wastewater treatment, capable of handling fluctuating hydrological and organic loading rates. Over the past 30 years, CWs have been set up all over the world, even in a desert [58], as an alternative to conventional mechanical intensive treatment systems to treat a wide range of wastewaters [53, 54, 57]. The CWs may also have the additional benefit of providing green space for recreation and wildlife habitats, but require large areas of land and are biologically and hydrologically complex, while pests such as mosquitoes can be a problem in poorly designed systems [57]. To date, the development of CWs includes options for differential treatment of compounds with respect to recalcitrance and fate. Flexibility can be reached if treatment steps can be combined according to the pollutant regime of the wastewater [52]. This option is of particular importance with a view to the general observation that wastewaters are increasingly polluted with pharmaceuticals and personal care products (PPCPs), compounds with high recalcitrance, and potential to damage the environment. Early approaches to wastewater treatment with CWs paid particular attention to engineering aspects of the systems (plumbing, pumping, and aeration); more latterly, effort has been directed at selecting and developing plant-based systems to reduce energy costs, and the rhizospheric/endophytic microbial community is now the focus of ongoing research. It also seems possible to improve the metabolic capacity of the plant´s

Novel developments in biological technologies for wastewater processing

rhizospheres, which are in direct contact with the percolating wastewater and its pollutants, by inoculating them with highly active bacteria performing crucial steps of pollutant breakdown. Even more attractive are attempts to influence the plant´s endophytic community of bacteria and fungi living inside the roots, shoots, and leaves. With such an approach, prominent recalcitrant pollutants like carbamazepine have been detoxified and degraded [59, 60], or, in the case of volatiles like trichloroethene, broken down to avoid their volatilization [61]. An intermediate between CWs and stabilization ponds are constructed retention dams in wetlands, or at the outlet of CWs. Gravel dams are erected and planted with aquatic macrophytes. The dense root system stabilizes the dam material, increases the habitat size for microorganisms, and facilitates the formation of sediment (Fig. 7). Here advantage is taken of the reduction of flow rates, and the option of final removal of recalcitrant compounds, shortly before the effluent joins with natural rivers or lakes [62]. Wastewater treatment needs to minimize greenhouse gas emissions, specifically their production of methane from anaerobic processes and carbon dioxide and nitrous oxide from aerobic systems. Stabilization ponds are of mixed benefit versus AD because of the presence of the lower anaerobic layer, and the inclusion of algae or plants in these systems may help. Compared with algae growing in stabilization ponds, vascular plant biomass is

€rlbach treatment facilities, upper Bavaria, Germany. Fig. 7 Filter dam of the Mo

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more easily harvested, and plants have been shown to remove pollutants such as those in dairy waste [63, 64]. Water hyacinth (Eichornia crassipes) was shown to absorb pollutants and organic compounds in high concentrations [65], for example, removing up to 80% of the nitrogen and nearly 70% of the potassium from water [66]. The vast networks of roots are also able to filter out particulates [67]. Sadly, although it is cold-intolerant, water hyacinth easily becomes an invasive weed in the environment and human-made ponds if uncontrolled. Reed beds, such as those using the Common Reed (Phragmites australis), can oxygenate and help to treat wastewater and gray water via the root systems in the substrate [68]. Reed is also an excellent sink for organic pollutants, including pharmaceuticals, and can hence be used in CWs. While its metabolism of paracetamol, diclofenac, and metformin has been well proven [69, 70], recently even the breakdown of carbamazepine, albeit aided by endophytic bacteria, has been proven [60]. Similarly, Typha spp (latifolia or angustifolia) have been shown to extract and detoxify a large number of water pollutants, among them pharmaceuticals. They are used in many CWs. Huber and Bartha [71] proved an uptake of the compounds paracetamol and diclofenac in Typha roots and also showed the occurrence of metabolites in the leaves of treated plants. This finding is of some relevance for practical purposes, because it is possible to harvest biomass for incineration and thus remove the pollutant from the system. The growth of these plants in horizontal or vertical flow, or mixed, systems aids microbial colonization and subsequent digestion of pollutants in sewage. Many reed beds have now been installed, and their refinement and improvement is ongoing [72]. As with algal biomass, harvested vascular plants from wastewater could find applications in agriculture and energy stocks. Phragmites is highly productive, with 105 t ha
1 [73]. Similarly, among floating macrophytes, E. crassipes is unarguably an excellent source of biomass, with 79 plants m
2 of >15 kg dry weight [67]. It has a high proportion of hemicellulose which, via AD, will yield two simple products, methane and carbon dioxide [74]. More than 70,000 m3 ha
1 of biogas was produced from 1 ha of standing crop (70% CH4, 30% CO2) [75]. An increasingly researched approach is the use of the aquatic plants collectively known as duckweed. Lemna or Landoltia (Spirodela) spp are alternatives to water hyacinth, representing an easily harvested feedstock for biofuel production from AD, and also for hydrothermal treatment and bioethanol fermentation. Landoltia punctata is the smallest flowering plant, and grows faster than most higher plants being able to double its biomass in 1–2 days [76]. Environmental regulations in many countries mean that duckweed grown on wastewater must be treated as sewage sludge (i.e., used only for compost or biogas production) despite being rich in protein and starch. Phosphorus is an important water pollutant and also a major plant nutrient, so aquatic plants can also play a role in its recycling. Duckweed’s high nitrogen and phosphorus uptake from its growing medium has seen it used to treat wastewater streams from industrial

Novel developments in biological technologies for wastewater processing

wastes or landfill leachates, and widely used to clean wastewaters from livestock farms. Lemna minor grew in AD effluent of up to 42 mg L
1 nitrogen [77], and it is agreed that duckweed ponds may be a promising route for polishing domestic and other wastewaters, seasonally producing good-quality secondary effluents (BOD and suspended solid removal) for small communities in temperate regions [77]. Duckweed-based wastewater treatment plants have been used in many countries, including Taiwan, China, Bangladesh, India, and Bolivia, and the United Nations report projects in Bangladesh and Peru showed potential for the treatment of wastewater and also fish food production [78]. Phytoaccumulation is a frequently used technology for heavy metal pollution from mine tailings or industrial processes wherein plants take up and store pollutants in amounts up to 1000-fold above soil concentration [79]. In fact, Lemna seems suitable for phytoaccumulation/phytoremediation, despite its small appearance [80]. Accumulation of copper inside the plants was observed at high levels (250–350 μg g
1 dw) and confirms the investigations of Mkandawire and Dudel [81] who observed bioaccumulation between 200 and 800 μg g
1 dw for Lemna spp. Furthermore, the uptake and detoxification of the herbicide, pethoxamide, in the same plantlets proves the assumption that L. minor is suitable for the application in phytoremediation, for example, in rural areas. Although the sequestration of nutrients has been demonstrated, the sedimentation of plants raises the question of subsequent greenhouse gas emission during decay. At least in the first operation, the use of these plants was shown to be favorable: two pilot duckweed (Landoltia punctata) ponds, with 200 L day
1 wastewater flow from domestic housing, were able to remove approximately 80%–90% of total nitrogen, phosphorus, and organic matter, while fixing threefold the level of CO2 that was emitted. No methane production was detected, methanogenic Archaea requires redox conditions well below those measured in the ponds [82]. As with water hyacinth, the success of duckweed can also be a problem: L. punctata is invasive and is listed by Florida, USA, for example, as a prohibited weed [83]. The movement of water is a control strategy for invasive growth, since duckweed proliferates best in stagnant water. The growth of floating plant species may also be altered by nutritional status and plant diversity of the water: management of biomass and persistence in a range of conditions in water treatment could result from extending research on plant polycultures [84]. In laboratory studies, L. minor and Wolffia brasiliensis could grow faster together at low temperatures (because of the traits of L. minor) and produce a greater number of resting bodies for recolonization of the water for the next growing season (traits of W. brasiliensis) [84]. The polyculture of plants also shows promise in the production of high-value products. Duckweed together with members of the water fern genus, Azolla, were considered by Muradov and Taha [85] for phytoremediation and subsequent recovery of straight-chain alkanes from the plant materials for use as biodiesel additives.

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5.1.1 Novel developments in CWs: Management of the rhizospheric/endophytic microbial community Recent advances in microbiome research have markedly changed the classical perception of phytoremediation techniques [86]. Today, plants are considered as “holobionts” and harbor a wide diversity of different microbial communities in specific zones, that is, the “rhizosphere” between plant and root, the aboveground portion of plants or “phyllosphere” and the root interior or “endosphere.” These microorganisms are able to interact with the plant genome and in so doing, shape the plant holobiont in response to environmental factors. Recognition of this plasticity permits improvements in wastewater treatment by modulating the microbial communities associated with macrophyte species used in CWs. Fig. 8 illustrates the complexity of the wetland ecosystem in which several microbial communities cohabit. Microbial communities in CWs have been studied in microcosms experiments for the removal of heavy metals [87], organic pollutants such as hexachlorobenzene [88], pesticides [89], antibiotics [90], and other pharmaceuticals [91]. Plant species, hydraulic design, availability of organic matter, temperature, DO, and substrate type, among others, strongly influence the microbial community [92], thus operational and environmental settings affect the efficiency of CWs directly during wastewater treatment. While sediments harbor anaerobic bacteria that can degrade compounds at very low efficiency, the rhizosphere is a much more versatile environment for bacterial pollutant degradation. Wetland rhizospheres constitute a unique ecological environment where aerobic and anaerobic microorganisms cohabit, subjected to continually changing oxic and anoxic conditions. Macrophyte species can provide oxygen to their rhizospheric microorganisms through the plant aerenchyma cells inducing diurnal fluctuations between oxic and anoxic conditions [93]. Rhizospheric bacteria grow in aerobic to microaerobic conditions subjected to diurnal changes and are more effective in degrading organic contaminants. Root exudates and secondary plant metabolites can serve as primary substrates for cometabolism of persistent compounds by rhizospheric and endophytic microorganisms [94, 95]. Endophytic bacteria reside in internal plant tissues. They are strongly adapted to the plant cell environment and can achieve higher efficiencies in metabolism once the contaminant is inside the plant. The use of endophytic microorganisms as inoculants to remove organic pollutants in phytoremediation is gaining attention, and many microorganisms have been isolated from macrophytes in the past few years [96]. These microorganisms can be selected for their ability to degrade organic pollutants and additionally, for their contribution to plant fitness through plant growth promotion [59]. Endophytic bacteria harboring catabolic genes and plant growth promoting traits can establish successful and durable mutualistic relationships with their hosts, because they do not have to compete with the dense populations of microorganisms present in the different soil compartments [97].

Novel developments in biological technologies for wastewater processing

Fig. 8 Microbial communities in constructed wetlands: remediation of organic pollutants. Several microbial communities cohabit in constructed wetlands. Anaerobic sediment bacteria can degrade compounds at very low efficiency. Biofilms can colonise plant surfaces, soil particles, sediments and dead organic matter and contribute to the degradation of organic pollutants. Macrophyte species can provide oxygen to their rhizospheric microorganisms which are more effective in degrading organic contaminants. Organic pollutants and their transformation products produced by rhizospheric bacteria are taken up by the plant and accumulated or further degraded by endophytic bacteria. Root exudates and plant secondary metabolites can serve as primary substrates for cometabolism of persistent compounds by rhizospheric and endophytic microorganisms. Additionally, endophytic bacteria can contribute to plant fitness by plant growth promotion. Transformation products and non-degraded compounds can be accumulated in plants which can, later on, serve as biomass for bioenergy production.

The use of hairy roots as model systems to screen the potential of plant species to remove and degrade persistent contaminants has increased in the recent years, with the aim to explore the mechanisms of microbe-assisted phytoremediation [98]. Plant and endophytic metabolic pathways of a recalcitrant contaminant such as carbamazepine have been described recently using microbe-free and inoculated hairy roots [60]. However, while most of the isolated endophytes show promising results at a microcosm scale, they seem to fail when applied in the field. Therefore, endophyte-assisted phytoremediation requires a thorough understanding of the factors governing endophyte colonization of the rhizosphere and/or plant tissues to improve the efficiency and

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reliability of inoculant strains [99]. Engineered endophytic bacteria harboring genes required for appropriated catabolic pathways may improve the phytoremediation of organic contaminants because they have a selective advantage over the indigenous community. This approach has been successful in industrial-scale trials for naphthalene [100], toluene [101], and trichloroethylene [102] remediation. Even more important, there is a shift from engineering specific endophytes to engineering rhizosphere communities. This can be achieved by bioaugmentation in CWs with competent microbial consortia and/or by modifying endophytic communities, for example, by biostimulation of degraders and beneficial microorganisms.

5.2 Algal pond treatment systems In wastewater treatment, aeration by mechanical means can be energy intensive, with 0.4–1.1 kWh required to transfer 1 kg of oxygen [103]. The use of microalgae for wastewater oxygenation is considered to be the most economical and energy-efficient method available [104]. Algae have long been associated with problematic growth in discharge wastewater enriched in compounds of nitrogen and phosphorus. Today the use of algae in the so-called facultative ponds is widespread, and in combination with engineering developments to include high-rate algal ponds or HRAPs, offer the potential for deriving a relatively enriched stream of CH4 and algal biomass, and reduced costs in wastewater treatment (see Fig. 9). 5.2.1 Microalgal growth requirements Optimizing the growth of microalgae to support optimal rates of bacterial growth and hence optimal rates of wastewater management requires understanding not only of

Fig. 9 Schematic representation to illustrate integrated algal wastewater treatment systems generating methane gas for energy and algal biomass with potential for “added value.”

Novel developments in biological technologies for wastewater processing

gaseous exchanges between the different communities but also understanding of their nutrient requirements and nutrient availability in prevailing wastewaters. Beal has stated that; “it does matter how much energy you produce per acre if it requires more water than is available.” One of the main advantages of microalgae is their ability to grow in water unsuitable for land crops [105]. The use of wastewater or seawater for microalgal production can reduce the freshwater demand by 90% [106]. Microalgae have a high content of both N and P relative to land plants at 5%–12% and 0.3%–1%, respectively [107, 108] and in an extensive review of their resource demands these elements were considered likely to emerge as the dominant constraints for scale-up of autotrophic microalgal production [109, 110]. Although some genera of aquatic photosynthetic prokaryotes, cyanobacteria, have the ability to fix nitrogen, virtually all algal species require an exogenous source of fixed nitrogen with most using ammonia preferentially [108, 111]. However, Oswald [112], an innovator and pioneer in the field described the advantages of algal waste treatment simply as, “solar energy and the resulting ‘photosynthetic oxygenation’ [are] nearly free, but also the nutrients in the wastewaters are free and often ideally suited for algal mass cultures.” A lifecycle assessment (LCA) by Yang [113] found that the use of wastewater to grow microalgae virtually eliminated the need for all the nutrients except phosphate. Parachlorella kessleri-I removed 81% of the total nitrogen from the water from Neela Hauz Lake, India polluted by sewage but removed 98% of the phosphorus [114]. Phosphorus is required in smaller quantities than nitrogen but may present a greater challenge. Both phosphate and nitrogen are found in municipal wastewater with typical concentrations of 30–40 mg L
1 N and 5–10 mg L
1 P [107]. Chlorella and Scenedesmus have both been found to grow in a wide range of wastewaters, often with suboptimal ratios of N and P for algal growth, with “complete decomposition” of nitrogen and phosphate nutrients in 10 days [115]. In what was termed a “realistic technology and engineering assessment of algal biofuel” it was found that there could be no favorable outcome for algal biofuel production unless wastewater treatment was the primary goal of the process [111]. A mass balance for the EU FP7 All-Gas project (an energy project funded by 7th Framework Programme for Research and Technological Development the European Union’s primary instrument for funding research from 2007 to 2013) has shown not only that sufficient N and P can be provided for microalgal growth by wastewater, but also that with recycling valuable N and P can be exported from the system for fertilizer [116, 117]. Algae may also have a requirement for other nutrients in small quantities. Out of 27 microalgae 26 were found to require vitamin B12 to grow which can be produced by the growth of bacteria [118]. Wastewater and seawater together with the native bacteria found in them could provide the essential micronutrients, minerals, and vitamins required for microalgal growth. Although some microorganisms produce substances toxic to microalgae, the inclusion of a mixed microbial population in any

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microalgal-based wastewater treatment system helps to degrade complex organic substances, and aids microalgae growth and pollutant removal [119]. Municipal and industrial wastewaters, in addition to essential algal nutrients, may also contain inhibitors of algal growth [119, 120] but this aspect of wastewater requires considerable further study. 5.2.2 Cultivation systems designed to support algal growth in wastewater processing Facultative ponds are shallow human-made ponds where the organic content of the effluent is converted to bacterial and microalgal biomass, the symbiotic growth of microalgae and bacteria reducing odors and pathogenic microorganisms [121]. Algae provide the oxygen for the growth of the bacteria to break down the organic waste matter, and the bacteria, in turn, provide carbon dioxide for the growth of the algae [103, 121, 122]. Facultative ponds have an aerobic zone at the top and an anaerobic zone at the bottom as shown in Fig. 10. These ponds are simple with very little equipment and low energy input (as oxygenation is provided by algae), but they require a favorable climate and large areas of land [122]. Simple, unmixed waste stabilization ponds were introduced in the United States in the early 1900s as a low-cost solution to wastewater treatment for a growing population.

Fig. 10 Operation of the facultative pond. Based on Tchobanoglous, G., E. Schroeder, Water Quality: Characteristics, Modeling and Modification. Prentice Hall, 1985.

Novel developments in biological technologies for wastewater processing

They were initially merely used for containment without discharge, rather than being designed and optimized for wastewater remediation [116]. Algal waste stabilization ponds gained wide acceptance and by 1980 over 7000 such ponds were in operation in the United States [123]. Favorable climate conditions, however, led to very large pond systems of 60–340 ha being established in California [124]. The growth of bacteria and microalgae in waste stabilization ponds can reduce not only dissolved nutrients and BOD, but also produce significant amounts of microalgal and bacterial biomass. This biomass can present an environmental threat if it is allowed to flow out of the pond into the surrounding environment, but it can also serve as a source of organic material for feed and fuel if harvested [122]. However in Europe the use of wastewater for microalgae production per se may be restricted; the Animal Feed Regulation bans the use of wastewater and all derived products in animal feed [109]. Greenhouse gas emissions from facultative ponds are another area of concern. CO2 can either be emitted or absorbed depending on the degree of algal photosynthesis, and CH4 emissions are highly variable, but is always positive [125], and thus there is a need to not only utilize the algal biomass, but also capture the methane for use as fuel. Algae play an important role in wastewater treatment in facultative ponds and minor roles in conventional aerobic treatment, with algae, such as Chlorella, growing in the uppermost part of trickle filters where sunlight is available [13]. In recent years, considerable attention has been paid to select and optimize the growth of algae in tandem with bacterial systems for wastewater treatment, in “phycoremediation” [126–128]. 5.2.3 Innovations in algal wastewater treatment systems Algal waste-stabilization ponds (1–3 m deep) were designed for optimal waste treatment. The potential to increase and exploit microalgal biomass and reduce the land demands of waste stabilization ponds have led to attempts to produce improved wastewater treatment systems, involving microalgae, and the development and integration of high-rate algal ponds (HRAPs) [121, 124]. Thus while algae may be grown for high density in closed systems known as photobioreactors (PBRs) [129, 130] the majority of microalgal production concerned with wastewater occurs in conjunction with HRAPs [131, 132]. A variety of experimental PBRs have been found to be effective in removing both N and P but their high operational energy precludes their use in wastewater treatment [119]. The energy for circulating fluid in PBRs can be considerable, and has been estimated to be 13–28 times that of open raceway ponds for the production of the same mass of microalgae [133, 134]. Several studies have shown that heterotrophic bacteria play a vital role in algal growth and survival [135, 136] and, again, this indicates that open systems rather axenic culture in PBRs are more applicable to wastewater treatment. Oswald’s early pioneering treatment ponds only exert minimal control over the algal species that grew, with only optimization of pond operations factors such as residence time, depth, and mixing [112]. Despite the

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limited control of microalgal species in such open systems, they remain the dominant algal technology at commercial scale for treating wastewater. The HRAPs are relatively inexpensive to both build and operate and, although they can suffer low productivity, most commercial, algal, advanced wastewater treatment systems are based on this design [137]. Simple facultative pond systems remain a common form of wastewater treatment, increasing regulatory pressure to upgrade treatment processes to allow best nutrient removal, and the greater focus on renewable energy production and improved greenhouse gas management, have led to wider application of HRAPs and [127] advanced integrated wastewater pond system (AIWPS) [127]. Most commercial, algal advanced wastewater treatment systems now incorporate HRAPs into the final design as a means of allowing best nutrient removal, greater focus on renewable energy production, and improved greenhouse gas management [137]. The initial work on AD of microalgal biomass grown in wastewater was carried out in the 1950s; Golueke and Oswald [138], in a series of publications, described the role of microalgae in sewage treatment using AIWPSs (Fig. 11) [127, 139–141]. Cowen et al. subsequently developed “a derivation of AIWPS” the integrated algal pond system (IAPS) (Fig. 12) for “local conditions” in the Republic of South Africa [142, 143]. Firstgeneration systems remediate domestic wastewater to a standard suitable for discharge to the environment, but subsequent AIWPS and IAPS processes are self-sustaining with

Fig. 11 Photograph of an advanced integrated wastewater pond system (AIWPS) with covered anaerobic pit and HRAP in the Waikato, New Zealand. Courtesy of Craggs, Park Craggs, R., et al., High rate algal pond systems for low-energy wastewater treatment, nutrient recovery and energy production. N. Z. J. Bot., 2014. 52(1): p. 60–73.

Novel developments in biological technologies for wastewater processing

Fig. 12 Integrated algal pond system (IAPS) at the Environmental Biotechnology Research Unit at Rhodes University, South Africa showing first stage Facultative Pond in the foreground and High rate Algal Ponds in the background.

the capture of methane and harvesting of algae biomass [103, 142]. In a recent study of the 500 person equivalent (PE) Belmont Valley WWTW pilot-scale IAPS treating municipal wastewater products included water for recycle and reuse (28 ML year
1), methanerich biogas (1880 kg CH4 year
1equivalent to 26 MW or, 55 kWh PE
1 year
1), and biomass (>3 t dw year
1) [142]. The IAPS treated water complied with the general limit values for either irrigation or discharge into a water resource for volumes of up to 2 ML of treated wastewater on any given day, while parameters including COD, total suspended solids (TSS), pH, DO, electrical conductivity (EC), and N and P values were within the general limit after tertiary treatment by either a maturation pond series (MPS), slow sand filtration (SSF) or controlled rock filtration (CRF) [143]. Fig. 12 illustrates an IAPS at the Environmental Biotechnology Research Unit at Rhodes University, South Africa with a facultative pond for treating wastewater in the foreground with a central pipe to collect methane gas after mixed AD-algal processing and in the background HRAPs. Such systems integrating HRAPs with anaerobic and aerobic biological processes for wastewater treatments are now utilized globally for the remediation of wastewater and are becoming widely recognized as a contemporary municipal wastewater treatment technology, with the largest plant processing 7.2 ML of wastewater per day

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[126, 128, 142] (Fig. 11). The low energy requirements of algae-based wastewater treatment systems were reported to be <0.6 kWh kg
1 BOD removed compared with 0.8–6.41 kWh Kg
1 BOD removed in mechanical aerated ponds [106, 131, 144, 145]. They have been used in effluent clean-up in warm regions in the United States for over 70 years, and current research focuses on maintaining the advantages of waste stabilization ponds (low energy input and simplicity), while mitigating the drawbacks of extensive land use, odour and sludge build-up. The HRAPs typically comprise a shallow closed-loop recirculation channel [146] where algal growth medium is circulated around a central rib (Figs. 11, 13 and 14) in the so-called raceway. The raceway either can be a single loop or may be serpentine [127]. The depth of an algal raceway is a compromise between being sufficiently shallow to give adequate light for the algal cell growth, and deep enough for fluid to be moved around the raceway while avoiding the costs of raceway bed grading [147]. Raceways can be up to 0.5 m deep [148], but are typically between 0.2 and 0.3 m [111, 145, 149, 150]. Below 0.15 m, problems have been found with algal biomass sedimentation and achieving sufficiently even surface grading of the pond bottom, which ensures consistent flow

Fig. 13 High rate Algal Ponds at the Environmental Biotechnology Research Unit at Rhodes University, South Africa.

Novel developments in biological technologies for wastewater processing

Fig. 14 Photograph of the 5 ha HRAP system operating at Christchurch, New Zealand. Courtesy of Craggs, Park Craggs, R., et al., High rate algal pond systems for low-energy wastewater treatment, nutrient recovery and energy production. N. Z. J. Bot., 2014. 52(1): p. 60–73.

around the entire raceway [151]. Mixing problems, temperature variation, and high rates of carbon dioxide outgassing may also occur at depths below 0.25 m [111]. An adequate mixing is required to expose all cells to light, distribute nutrients, enhance gaseous transfer, and prevent settlement [104, 107, 152]. The circulation requirement is normally provided by paddle wheels as they are mechanically simple, high volume, and low head devices with a gentle mixing action [153]. Flow velocities are typically in the range 0.15–0.3 m s
1 [153, 154]. Oswald [104] suggested that an average velocity of 0.15 m s
1 was required within a raceway to maintain a minimum velocity of >0.05 m s
1 in all areas of a raceway, which prevents thermal stratification and maintains algae in suspension. Optimum mixing in algal ponds was suggested to occur between 0.2 and 0.3 m s
1 [111], while the use of a CFD model showed no increase in microalgal growth above 0.3 m s
1 [146]. In addition to raceway depth and fluid flow velocity, the organic loading rate and hydraulic retention are the key operational factors in HRAP treating wastewater [128]. Maximum organic loading rate HRAP is between 100 and 150 kg BOD5 ha
1 day
1 depending on climate [128]. In temperate climates, hydraulic retention time is 3–4 days in summer and 7–9 days in winter [128].

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From an engineering point of view, gaseous transfer is the key issue in the design of algal growth systems, with CO2 being the most limiting condition for maximal algal productivity [111]. Photosynthesis can be simplified into two reactants (carbon dioxide and water) and two products (glucose and oxygen), represented by the equation: 6CO2 + 6H2 O ¼ C6 H12 O6 + 6O2 Stoichiometrically, this equation suggests that every 1 kg of algae produced will require 1.5 kg of CO2. Values for the amount of CO2 required to produce algal biomass are quoted at between 1.65 [155] and 2.2 kg per kg dry weight [156] with a typical figure of 1.8 kg of CO2 per kg of dry algae [130]. CO2 solubility in water is affected by salt and organic matter concentration, temperature, and pH value. At the same time, CO2 from 5%–10% (1.5–3 mM) to 20% (6.6 mM) has a negative influence on algal growth. It decreases with increasing concentration of NaCl and with increasing temperature, and increases or decreases with increasing concentration of organic compounds in wastewater depending on the compound. Algae rapidly deplete CO2 within a growth medium; to maximize algal growth, additional CO2 to that available via atmospheric transfer can be supplied [106, 157]. If 1.8 kg of CO2 is consumed by photosynthesis, this will produce 1.3 kg of O2, and this will accumulate within the algal growth medium unless some additional means of gaseous transfer is used. Elevated oxygen levels of only a few times saturation in water have been shown to inhibit algal growth [149, 158–161]. CO2 can be supplied in the form of bubbles or by the addition of carbonated water, the former giving higher growth rates [160], possibly because of the removal of O2. If CO2 is to be added to increase algal production in HRAP wastewater systems, a low-cost supply is required. The CO2 can be from combustion engines and industrial flue gas, or from the upgrading of biogas from the AD of the biomass produced from the system. Flue gases have been found to have no adverse effect on the growth of microalgae [160, 162] and growth rates for Chlorella were shown to be higher on flue gases with a CO2 content of 11%–13% than air enriched to 12% CO2 [163]. NOx in flue gases has been found to have little adverse effect on microalgae [164]. Although flue gas purchase costs are zero, or low, the costs of gas transport are not trivial [165]. Algal raceways, therefore, are ideally located near to the source of flue gas, allowing the use of a low-pressure blower. If the transport distance increases, a high-pressure compressor will be required, increasing energy costs by an order of magnitude [166]. Depending on the content of organic matter, the annual biomass productivity of wastewater treatment HRAPs, without CO2 addition for Mediterranean climates, can reach 30 t ha
1 year
1 (ash-free dry wt), two- to threefold the annual productivity of conventional facultative ponds (10–15 t ha
1 year
1) where bacteria supply CO2 in AD processes. It can be increased by increasing the concentration of CO2 and values as high as 58–73 t ha
2 year
1 have been reported [128].

Novel developments in biological technologies for wastewater processing

First-generation integrated algal pond systems that incorporated HRAPs remediated domestic wastewater to a standard suitable for discharge to the environment, but subsequent systems can be self-sustaining with the capture of methane and harvesting of algae biomass [103, 142]. Wastewater initially enters a covered anaerobic pond or advanced facultative pond where organic solids settle and decompose anaerobically producing biogas, then collected by a pond cover for use as bioenergy [127, 128, 142]. Water laden with nutrients then flows to HRAPs where microalgae grow. The algal biomass produced in HRAP is used as feedstock for AD with yields of up to 88% of the theoretical methane potential [167] and, additionally, nutrients from the digestate from AD of the harvested algal biomass can be supplied to microalgal growth systems to stimulate algal growth further [168]. However, the amount of methane produced can vary considerably between species microalgae 0.161–0.435 L CH4 g
1 VS [167, 169], but yields can be improved by mechanical and thermal pretreatments and by co-digestion with convention sewage sludge or other wastes, such as glycerol, and the co-digestion of microalgae biomass and sewage sludge has economic and ecological advantages for WWTPs [169–171]. Nevertheless, the AD of microalgae biomass can still be challenging because of technical issues including the low concentration of digestible biodegradable substrate, recalcitrant substrate constituents, cell wall degradability, low C:N ratio, ammonia toxicity, and effects of salinity and other ions from the high ash content [140, 141]. Advanced pond systems require approximately 50-fold the land area of activated sludge systems for the same BOD removal, excluding the land area needed to dispose of waste activated sludge. Nevertheless, the capital costs for construction of an advanced pond system are <50%, and operational costs are <20% of those for activated sludge systems “All-Gas” was an EU-funded project based on a 10-ha facility in Chiclana de la Frontera, Spain. All-Gas demonstrated on a semi-industrial scale the feasibility of the sustainable production of algal biogas for vehicle fuels, based on low-cost microalgae cultures grown in municipal wastewater. The four raceways, each with an area of 5200 m2, produced an average of 100 t ha
1 year
1 of biomass; the annually produced biomethane from the conversion of the algae biomass grown in a hectare of raceway treating wastewater capable of powering 20 cars for 18,000 km each [172]. An LCA of the algae biorefinery process gave greenhouse gas reductions of about 40% compared to conventional wastewater treatment, with the primary benefit (credit) resulting from the substitution of the automotive fuel compressed natural gas by biomethane [173]. 5.2.4 Energy return on energy invested Energy return on energy invested (EROEI or EROI) is the ratio of the energy produced compared with the amount of energy invested in its production. This “simple” ratio can be useful in assessing the viability of fuels. A ratio of <1 indicates that more energy is used than produced, and an EROI of 3 has been suggested as the minimum that is sustainable

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[174]. The EROI for crude oil is currently about 20, but has declined over time [175], but the EORI for biofuels produced from biomass are often considerably lower, with sugar ethanol having a reported EROI between 1.25 and 8 and corn ethanol between 1 and 1.34 [174, 176–179]. The high lipid content of microalgae was the focus of past research on the utilization of microalgal biomass for the production of biodiesel, and a considerable number of LCAs have been carried out on the production of biodiesel from microalgae. It has been concluded that the process may be marginal in terms of energy balance, global warming potential (GWP), and economics. Only in the best-case scenarios was algal biodiesel found to be comparable to first-generation biodiesel and algal biodiesel was not, "really competitive under current feasibility assumptions" [180]. A reworking of the data from six LCAs, in what was termed a meta-model of algae bioenergy life cycles (MABEL), found that the energy return on energy invested (EROI) ranged from one, no return on the energy invested, to two, twice the energy invested [181]. A further extensive review and LCA using a Monte Carlo approach to estimate ranges of expected values found that nearly half of all the LCA results had an EROI of less than one [182]. The Sills’ [182] study also showed that methane from anaerobic digestion of defatted algae is required for net gains in energy, and must be an integral part of an algal biodiesel production process to yield EROI values that are greater than one. Also, nutrients recovered from waste streams are crucial. The embodied energy of nutrients is one of the major energy inputs in biofuel production from microalgae [109]. ter Veld [183] found a net energy return of >3 for biogas produced from microalgal biomass grown in HRAPs. A further energy balance model for the production of microalgal biogas using wastewater as a nutrient source in HRAPs found energy returns on operational energy invested of >3 with the use of nutrients from wastewater being a key element to produce a net energy balance [184, 185]. Park and Craggs [127] have concluded that HRAPs treating wastewater are currently the only economic means of producing algal biofuel with the minimum environmental impact. 5.2.5 Harvesting microalgal biomass Harvesting of algal biomass can be a major challenge in phycoremediation technology [119]. It has been suggested that 20%–30% of the costs of microalgal biomass is due to the costs of harvesting [186–188], but estimates as high as 50% of microalgal biomass cost have been given [189]. Also, it has been estimated that 90% of the equipment cost for algal biomass production in open systems may come from harvesting and dewatering [190]. Milledge and Heaven [191] reviewed the methods of harvesting microalgae, and observed that the well-tested wastewater treatment separation processes of flocculation with sedimentation or flotation may offer energy efficient solutions for algal biomass recovery in waste treatment [119, 172, 173, 192]. Bacterial-induced algal flocculation may hold considerable potential, although further research is required [193, 194].

Novel developments in biological technologies for wastewater processing

An alternative approach to harvesting is to induce biofilm formation on the surface of various substrates, which may be removed easily from the cultivation medium [120]. The application of algal-biofilm technology has shown considerable promise with reported reductions of 60%–90% in nitrogen and 54%–97% in phosphates in wastewater [119]. Although biofilm formation has tremendous potential as an alternative harvesting technology, there is still little information about its commercial application in dual-purpose systems for treating wastewater and recovering microalgae for biofuel, and there is a need for further research [120, 137]. 5.2.6 Algal biosorption Macroalgae, also known as seaweeds, and microalgae, both living cells and dead biomass, can be effective at removing heavy metals from wastewater [195]. Seaweeds are usually rich in minerals (10 higher than terrestrial crops) [196], with ash contents ranging from 3.5% to 46% [197–199]; typical figures for some species of research and commercial interest are: Laminaria digitata  26%, Ascophyllum nodosum 21%, and Sargassum muticum 32% [200, 201]. Seaweeds exhibit a high affinity for heavy metals and have even been used as biomonitors for metal pollution in estuarine and coastal waters [202]. Alginates are believed to have a critical role in metal biosorption by brown algae with a particular affinity for divalent cations (Pb2+, Cu2+, Cd2+, and Zn2+) [195]. Sargassum can contain high levels of arsenic (20-231 μg g
1 dw) with inorganic arsenic accounting up to 80% of the total arsenic content, and there have been a number of health advisories around the world warning not eat too much Sargassum especially Sargassum fusiforme [203]. S. muticum has been suggested as a low-cost bio-sorbent for the treatment of wastewater from industries that use dyes and phenolic compounds due to its rapid absorption and high absorption capacity for industrial dyes such as methylene blue and phenolic compounds [204–206]. S. muticum dried biomass has also been found to be effective in the sorption of antimony from water (5 mg g
1) [207]. Liu and Pang [208] have reported that S. muticum has been used in China for the sorption of heavy metals, and a variety of brown algae have been demonstrated to be effective in the removal of heavy metals [195, 207]. In addition to the use of marine brown algae for heavy metal biosorption, the freshwater filamentous algae, Spirogyra and Cladophora, have been found to be effective in the removal of both lead and copper from wastewater [195]. Although seaweed may be effective in the removal of heavy metals, Zeraatkar and Ahmadzadeh [195] in a recent comprehensive review concluded that microalgae treatments are usually more effective than macroalgal treatments. After World War II, there was urgent interest in the control of radioactivity in the environment, and a project studying the uptake of uranium by microalgae such as Chlorella found that the algal biomass was effective at removing uranium from wastewater with a capacity to remove 20 mg uranyl acetate 100 mg
1 Chlorella (dw) exceeding the

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performance of vascular plants [112]. A Scenedesmus spp-dominated microalgal consortia growing on wastewater from coal-fired power stations containing heavy metals removed “100%” of boron, molybdenum, vanadium, and zinc [209]. It was concluded that such freshwater microalgae consortia are suitable for metal remediation from industrial effluents, but care must be taken when considering end-product use of the microalgal biomass. However, the uptake of specific heavy metal ions can be highly strain specific. The effect of species, heavy metal type, competing ions, and various growth conditions (pH, temperature, etc.) have been recently extensively critically reviewed by Zeraatkar and Ahmadzadeh [195].

6. Concluding remarks The global population is projected to reach more than 9.7 billion in 2050 from an estimated 7.8 billion in 2020 [210], stretching finite water resources for global food production, drinking water, health and sanitation, as well as for energy and other goods and services that require water for their production and delivery. Most of these increases will occur in developing countries, where population growth will be coupled with rising incomes, urbanization, and climate change and will place considerable pressure on national and global food systems. Moreover according to FAO, IFAD, UNICEF, WFP, and WHO 2017 [211], while almost 800 million people are currently hungry, by 2050 global food production would need to increase by at least 50%. To meet this demand water of acceptable quality and in adequate quantity will be required. At the same time, food production and supply have a negative impact on the sustainability and quality of water resources. According to the FAO Aquastat website [212], agriculture is the biggest water user: agricultural water withdrawal is highly dependent on both the climate and the place of agriculture in the economy across countries, but viewed globally, agriculture accounts for roughly 70% of total freshwater withdrawals, with the industrial and domestic sectors accounting for the remaining 20% and 10%, respectively. There is also clear evidence that groundwater supplies are diminishing, with an estimated 20% of the world’s aquifers being overexploited, some critically so, while deterioration of wetlands worldwide is reducing the capacity of ecosystems to purify water [213]. Water consumption can be reduced, and supplies made more reliable, by using multiple water sources, including rainwater harvesting and wastewater reuse, and only treating water to be ready for its intended use, rather than treating all water to a safe drinking standard. It is within this context that biotechnological advances centred on the extensive capacities of heterotrophic microorganisms to extract energy from organic matter including xenobiotics and degrade these to CO2 and water become paramount. Biotechnological advances based on supporting the degradative capacity of microorganisms hold the potential either directly or indirectly to reduce the energy inputs needed in wastewater

Novel developments in biological technologies for wastewater processing

treatment plants for pumping, water treatment, drainage, desalination, and water distribution, to manage water resources, maintain water supply and sanitation as well as protect ecosystems. However, many challenges remain, especially to increase the rate and extent of organic matter degradation at the same time reducing energy inputs. Aerobic treatment processes currently show the greatest promise, but since aerobic heterotrophs typically use oxygen as an electron acceptor in their degradative armoury, oxygen availability represents a major constraint. Efforts to supply oxygen mechanically to ensure sufficiently high concentrations place a significant drain on energy resources and may also damage the flocs and biofilms of the degrading microbial communities. In the future, novel biotechnological developments that exploit the photosynthetic capacities of plant systems to deliver oxygen and reduce the concentration of inhibitory CO2 from heterotrophic metabolism may make a significant contribution. By exploiting the cooperative and mutualistic metabolism evidenced between plants and bacterial communities in the plant rhizosphere and endosphere wastewater treatment systems based on CWs offer enormous opportunities not only to reduce energy inputs associated with gaseous exchanges but also for tailoring organic matter degradation according to the nature of organic matter present in wastewater. At the same time improvements in waste stabilization ponds that incorporate microalgae in advanced integrated algal pond systems have been able to demonstrate increases in municipal wastewater treatment capacity without associated economic, energy, and environmental costs, to produce water for irrigation or discharge into a water resource, methane-rich biogas, and biomass for bioenergy and feed (where permitted). The development and application of plant and microalgal biotechnology-based wastewater treatment systems in any given geographic locality will, however, depend on being able to select and grow the most appropriate species; understand the population dynamics of the relevant bio-communities, and then engineer suitable systems that will support growth in wastewater of potentially changeable composition and quantity, in order to deliver the promise of rapid, all-year-round wastewater processing that will be necessary to meet the future projected global population demands for increased food, water, and energy.

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