Microalgae have a relatively low carbon to nitrogen (C/N) ratio when compared to other biomass substrates and wastewaters utilised in the anaerobic digestion process (Sialve et al. 2009; Vergara-Fernandez et al. 2008). The ideal C/N ratio for anaerobic digestion is approximately in the range of 20-30 (Parkin and Owen 1986). In contrast, microalgae have a C/N ratio, between 4 and 8 (Ward et al. 2014). When a substrate has a C/N ratio below 20, it can cause an imbalance in an accumulation of NH3-N within the digester and associated inhibition (Sialve et al.

2009) . To overcome low C/N ratios associated with microalgae, co-digestion with other waste streams high in carbon (i. e. paper, glycerol, pig manure, cow manure, lipid rich fats oils and greases, municipal waste, soybean oil) emolliates the high NH3-N build-up (Ehimen et al. 2009; Gonzalez-Fernandez et al. 2011; Saxena et al. 1984; Shouquan et al. 2009; Yen and Brune 2007). For example, when paper was added to microalgae in a digester, the C/N ratio increased from 6.7 to 36.4 with the best co-digestion biogas production with a mix of 50 % paper and 50 % microalgae at a C/N ratio of 18.0 (Yen and Brune 2007). When the biogas production from the C/N ratio treatment of 18 was compared to the lower C/N ratio of 6.7, a 50 % increase in biogas productivity was recorded (Yen and Brune 2007). Such co­digestion ratio productivity is highly dependent on the species of microalgae being digested as well as the secondary substrate being co-digested (Ward et al. 2014). When a high C/N ratio is used, there is a risk that the bacterial population may become NH3-N limiting, causing inhibition with the methanogen bacterial com­munity (Chen et al. 2008; Parkin and Owen 1986). When considering the co­digestion of microalgae with other substrates, the availability, seasonality and location of the secondary co-digestion substrate must also be considered to ensure continuous availability and low cost associated with procuring and transportation (Ward et al. 2014). Furthermore, balancing both the C/N ratio and all co-digested substrate, degradation rates are essential to balance the carbon and nitrogen release within the digester. To balance the C/N ratio, the co-digestion substrate should have a similar degradation rate as the primary substrate being digested (Ehimen et al. 2009; Kayhanian 1994).

Conventional methods used for the removal of heavy metal ions include chemical precipitation, adsorption, chemical oxidation/reduction, membrane filtration, ion exchange, and electrochemical processes. However, these techniques have some drawback, such as partial removal of metal ions, costly installation requirements, high energy demands, and the generation of toxic waste products which require additional elimination stages (Aksu et al. 2002).

Both live and dead cells can be successfully used for the biosorption of metal ions, while uptake of metal ions by living microorganisms, referred to as bioac­cumulation, occurs when an active metabolic process is involved (Aksu et al. 2002; Brady and Duncan 1994; Moreno-Garrido et al. 1998). Biosorption is a reversible process, since it is possible to desorb the metal ions bound to the surfaces of cells by a simple acid treatment, whereas bioaccumulation processes are only partially reversible (de-Bashan and Bashan 2010; Donmez and Aksu 2002; Velasquez and Dussan 2009).

Compared to the other organisms used for biosorption processes, namely fungi, cyanobacteria, and bacteria, algal cells have higher heavy metal biosorption capacities which relates to the different structure and composition of their cell wall (Bayramoglu et al. 2006; Gekeler et al. 1988). Cell walls of different microor­ganisms have different functional groups which are involved in metal ion binding, such as amino, amide, carbonyl, carboxyl, hydroxyl, imidazole, phosphate, sulfate, sulfhydryl, and phenol moieties (Barkley 1991; Schiewer and Volesky 2000). Depending on the variations in the cell wall composition, there will also be dif­ferences in the metal ion binding mechanisms and affinities (Godlewska-Zylkiewicz 2003; Leusch et al. 1995).

The chemical characteristics of the functional adsorbent (i. e., functional groups, polarity, and solubility) are responsible for determining the binding mechanism and the nature of the adsorption process. Different physicochemical forces, such as covalent bonding, van der Waals bonding, ion exchange, and dipole/dipole inter­actions can be responsible for the uptake of ions on the adsorbents (Aksu et al. 2002).

Free cells have some disadvantages when used for large-scale applications of metal ion biosorption studies, due to the otherwise risk of clogging problems on the filters and flow lines. Nevertheless, this problem was overcome by using immo­bilized cells in natural matrices such as carrageenan, alginate, chitosan, agarose; polymeric supports such as polyacrylamide, polypropylene, and polysulfone; cross — linked copolymers; or biomatrices such as sponges (Akhtar et al. 2003a, b; Rob­inson 1998). Some of those studies are highlighted in Table 2.2.

The presence of more than one type of metal ion within the wastewater might have a negative effect on the adsorption of one type of metal ion over another. Mehta and Gaur (2001) observed nearly complete removal of copper and nickel metals by alginate-entrapped C. vulgaris cells when they were in separate solutions. On the other hand, the presence of copper in the nickel solution inhibited the biosorption of both metals either by immobilized or free cells, due to the compe­tition of different metal ions on the same active sites of microalgae. da Costa and Leite (1991) used alginate-immobilized Chlorella homosphaera for the removal of cadmium and zinc metals. They also observed that the biosorption of cadmium and zinc alone was much higher than the case when these two metal ions were combined.

Biological materials were also used as the immobilization matrices for micro­algal cells. da Costa and de Franfa (1996) attached the microalgae Tetraselmis chuii and cyanobacteria Spirulina maxima on the surface of two different seaweeds (Sargassum sp. and the Ulva sp.), which eventually increased the overall cadmium biosorption efficiencies. In series of studies by Akhtar et al., C. sorokiniana algal cells were immobilized on a biological matrix of Luffa cylindrica sponge for the removal of nickel (Akhtar et al. 2003a), cadmium (Akhtar et al. 2003b), chromium (Akhtar et al. 2008), and lead (Akhtar et al. 2004) ions from liquid effluents. L. cylindrica sponge was chosen as the immobilization matrix due its rigid struc­ture, low cost, and high porosity, while its fibrous network provides an efficient contact between the immobilized cells with their surrounding aqueous environment (see Fig. 2.1b). They reported high maximum adsorption capacities in a continuous liquid flow column, as 192 mg cadmium and 71 mg nickel per gram of immobilized biomass. They also achieved successful desorption of cadmium and nickel metal ions with HCl solution, and the regenerated immobilized samples were reusable with a similar biosorption efficiency.

The biosorption of lead (Pb) ions by C. sorokiniana cells immobilized on L. cylindrica sponge was another efficient method, with 96 % adsorption efficiency of the metal ions within the first 5 min of the experiments (Akhtar et al. 2004). They also observed a maximum adsorption of lead ions at around pH 5.0. Higher removal rates were associated with the fibrous structure of the immobilization matrix, increased surface area, and easier access of the targeted metal ion to the sorption sites (Akhtar et al. 2003a, b).

Leusch et al. (1995) used two marine brown algae, Sargassum fluitans and Ascophyllum nodosum, for the biosorption of cadmium, copper, nickel, lead, and zinc heavy metal ions. They observed the highest metal uptakes when the cells were cross-linked with glutaraldehyde, followed by cross-linking with formaldehyde. Both species had the highest biosorption efficiencies for lead and the lowest for zinc. Introducing formaldehyde possibly involves cross-linking of the hydroxylic groups with the sugars of the cell wall, while glutaraldehyde cross-links mostly with the amino groups (Leusch et al. 1995).

Significant amounts of pollutants were removed using a mixed-immobilization of selected consortium of several microalgal species (Chlorella sp., S. obliquus, Stichococcus sp.) and several bacteria (Rhodococcus sp., Kibdelosporangium ari — dum) inside a highly contaminated pond, after the separate immobilization of microalgae and bacteria in solid carriers such as capron fibers and ceramics. They established 62 % copper, 62 % nickel, 90 % zinc, 70 % manganese, and 64 % iron removal efficiencies (Safonova et al. 2004).

Bayramoglu et al. (2006) used immobilized Chlamydomonas reinhardtii cells in calcium alginate beads for the removal of mercury, cadmium, and lead ions from aqueous solutions. They observed the highest adsorption capacities for immobilized cells for a pH in the range 5.0-6.0, achieving mercury, cadmium, and lead ion adsorption capacities of 89.5, 66.5, and 253.6 mg g 1 dry adsorbent, respectively. On the other hand, control samples composed of only calcium alginate beads provided less metal-binding sites and yielded lower adsorption capacities of mercury, cadmium, and lead ions at 32.4, 27.9, and 173.9 mg g-1 dry adsorbent, respectively. Acidic pH conditions were not optimal due to the protonation of the cell wall components. In contrast, mildly acid conditions (pH range 5.0-6.0) allowed sufficient interaction of the heavy metal ions with the carboxylate and phosphate groups of the algal cell wall (Bayramoglu et al. 2006). Neutral pH was found to be the optimal condition for an efficient chromium biosorption by immobilized C. vulgaris and freshwater cyanobacterium A. doliolum cells in alginate (Mallick and Rai 1993).

Barkley (1991) investigated the utilization of immobilized algae in a permeable polymeric matrix for the adsorption of mercury ions from groundwater in both laboratory and pilot-scale field tests. Their resulting immobilization product (AlgaSORB) was quite robust and can be packed within adsorption columns, having sufficient porosity to allow easy diffusion of the ions toward the cells. Field test results showed that AlgaSORB was a highly reasonable alternative to the conventional ion-exchange resins (Barkley 1991).

Nakajima et al. (1982) achieved the removal of uranium ions from both fresh­water and seawater samples using the immobilized cells of Chlorella sp. in poly­acrylamide gels. They also reported that this system can be used several times by applying consecutive adsorption and desorption stages.

Recovery of precious metals with immobilization methods can be a highly cost — effective process. da Costa and Leite (1991) used immobilized C. homosphaera cells within alginate beads for the adsorption of gold metal, which achieved a very high absorption yield of around 90 % of the initial quantity of gold present in solution.

Due to their exclusive catalytic properties, corrosion, and oxidation resistivity, palladium and platinum noble metals have been widely used in various areas from metallurgical processes, chemical synthesis, petroleum processing, electronics to automotive industry (Dziwulska et al. 2004). As a result of the high emission risks of these metals into the environment, it has become important to monitor their concentration in environmental samples. Thus, several microorganisms have been investigated for the separation and preconcentration of some trace metals such as palladium, platinum, copper, cadmium, lead, and gold via biosorption processes, which then allows the use of analytical methods such as atomic absorption spec­trometry and inductively coupled plasma optical emission spectrometry (Carrilho et al. 2003; Dziwulska et al. 2004; Elmahadi and Greenway 1991; Godlewska — Zylkiewicz 2003).

Dziwulska et al. (2004) demonstrated the selective biosorption of palladium and platinum ions from strong acidic solutions (pH below 2), using immobilized C. vulgaris cells on anion-exchange resin Cellex-T. This technique was also used for the preconcentration and analysis of these noble metals for graphite furnace atomic absorption spectrometry in different environmental samples including wastewater, tap water, and grass. Elmahadi and Greenway (1991) used Chlamy — domonus reinhartii and S. capricornutum algal cells immobilized on controlled — pore glass for the preconcentration of copper, silver, and chromium metals for atomic adsorption spectrophotometric detection. In their work, they also found that the presence of some compounds, such as sodium chloride, humic acid, and sodium bicarbonate, can interfere with metal biosorption process by competing for the metal ions. Silica gel was used as the immobilization matrix for Stichococcus bacillaris microalgae for lead preconcentration (Mahan and Holcombe 1992), while silica gel-entrapped Pilayella littoralis brown microalgae was used for the preconcentration of copper, iron, aluminum, and cobalt ions for their detection by inductively coupled plasma optical emission spectrometry (Carrilho et al. 2003).

Due to the low biomass concentration of microalgae (about 0.5 g L-1 in open ponds) and the small size of microalgal cells (usually 5-50 pm), harvesting mic­roalgal biomass is a major challenge (Uduman et al. 2010). Centrifugation is an efficient method for harvesting microalgae; however, this is too energy-intensive for most low-value applications (i. e., biofuels). Options such as flocculation are a promising approach to reduce harvesting costs (Vandamme et al. 2013). During flocculation, individual cells form larger aggregates that can easily be separated from the culture medium by gravity sedimentation, flotation, or enhanced settling in an inclined lamella separator. Using flocculation, the biomass can be concentrated from a dilute culture with a dry matter content of about 0.05 % to a sludge with a dry matter content of 0.5-5 %. Mechanical techniques such as centrifugation or a filter press are required to remove the remaining extracellular water and to obtain a thick paste with a dry matter content of 20 %.

Most HRAPs used for wastewater treatment today contain mixed consortia of microalgae rather than pure cultures. Usually, these communities are dominated relatively large, colony-forming chlorophytes such as Pediastrum, Microctinium, Scenedesmus, Dictyosphaerium, and Coelastrum (Benemann et al. 1980; Park et al. 2013). Possibly, these species are favored by the flow regime generated by the paddle wheel in high-rate algal ponds. These relatively large colonial microalgae often flocculate spontaneously, a process that is referred to as bioflocculation (Park et al. 2011a). Bioflocculating microalgae may form aggregates with other non — bioflocculating species (Salim et al. 2011), and bioflocculated microalgae have high settling rates and can be relatively easily concentrated to a slurry of 1 -3 % dry matter by simple gravity sedimentation (Sheehan et al. 1998). By recycling part of the harvested biomass, the dominance of these bioflocculating microalgae can be maintained (Benemann et al. 1980; Park et al. 2011b, 2013). Bacteria present in the wastewater may also play a role in bioflocculation (Su et al. 2011). Bacteria grow on organic matter present in wastewater, and research by Lee et al. (2008) and Lee et al. (2012) showed that the presence of bacteria in cultures of Chrysotila and Chlorella resulted in flocculation of the microalgal cells. In both studies, it appeared that extracellular polymeric substances produced by the microalgae were involved in the flocculation process. Van den Hende et al. (2011) showed that a sufficient supply of organic matter is important to sustain mixed algal-bacterial flocs.

The high pH that is typical of microalgal cultures can induce precipitation of Ca or Mg salts and can also induce flocculation of microalgal cells; a process that is referred to as autoflocculation. Ca phosphates precipitate at a relatively low pH of about 8.5-9 and can induce flocculation of microalgae. Such pH levels are regularly encountered in outdoor microalgal cultures when irradiance levels and temperatures are high. Flocculation by Ca phosphate precipitation requires relatively high Ca and phosphate concentrations in the wastewater and is therefore only applicable in hard waters with excess phosphate levels (Sukenik and Shelef 1984; Sukenik et al. 1985). While autoflocculation by Ca phosphate works well in laboratory conditions, it often fails in large-scale systems, even when Ca and phosphate concentrations are sufficiently high (Nurdogan and Oswald 1995). This may be due to autoflocculation by Ca phosphate is inhibited by the presence of organic matter in microalgal cultures (Beuckels et al. 2013). Bioflocculation and autoflocculation have been studied in the past 30 years in laboratory conditions and pilot systems. It appears that their performance depends strongly on species and cultivation conditions, yet, the reliability of these methods remains to be proven in long-term and large-scale operations (Benemann et al. 2012). More details on the recent developments on harvesting and dewatering can be found in Chaps. 12—14.

In general, molecular genetic techniques are concerned with manipulating, repro­ducing, adding, and deleting DNA and RNA molecules in an organism. Manipu­lations, deletions, and additions are accomplished via genetic transformation and genetic editing, while reproduction is accomplished via cloning. There are also numerous other methods for genetic and genomic editing; however, transformation and cloning remain an integral part of even the most recently developed methods. Novel techniques are required to make those molecular changes easier to manifest, less time consuming, and more permanent in their effect.

Flocculation may play an important role in developing such a low-cost and high — throughput harvesting process (Brentner et al. 2011). During harvesting or dewa­tering of microalgae, a microalgal paste is produced with a dry matter content of 20 %. This implies a 400-40 times concentration from 0.5 or 5 up to 200 g dry matter L-1. Theoretically, this can be achieved in a single step using mechanical methods such as high-speed centrifuges or ultra — or micro-filtration membranes. Due to the large volumes of culture broth that need to be processed, however, the cost and energy inputs are extremely high. The energy inputs for harvesting by means of centrifugation are 14 MJ kg 1 of dry biomass, which is about 55 % of the energy content of the biomass (Norsker et al. 2011).

Several authors have suggested that low-cost harvesting of microalgae can be achieved by means of a two-stage harvesting process in which the biomass is pre­concentrated by means of flocculation prior to final dewatering (Pahl et al. 2013; Kim et al. 2013; Weschler et al. 2014). In flocculation, individual microalgal cells form larger aggregates or flocs. These flocs have much higher settling rates than individual cells and can easily be separated from the medium by means of gravity sedimentation to yield a microalgal slurry (Fig. 12.1). This slurry can be completely dewatered using a mechanical method such as centrifugation or filtration. Due to the large size of flocs compared to individual microalgal cells, the energy demand for mechanical dewatering will be much lower than for culture broth with freely sus­pended cells. Flocs have higher sedimentation rates than individual cells and can be separated from the medium with a low centrifugal force (Xu et al. 2012). When filtration is used, flocculation prior to membrane filtration results in higher mem­brane fluxes and thus a lower energy demand (Lee et al. 2012a, b).

Ultrasonic waves, in various formats, have been used to induce aggregation of algae and speed up separation of algae. The concept is to use ultrasonic energy to con­centrate the algae in a specific location or node to facilitate rapid recovery. Ultra­sonic separation can be run continuously, has low stress (cells remain viable), is non-fouling, has few moving parts to break down, and occupies relatively little space. However, this process was considered less cost-effective than centrifugation for industrial harvesting due mainly to the requirement for a cooling system and the low concentration factor compared to centrifugation and microfiltration (Bosma et al. 2003). It might be useful in a system where the alga secretes a metabolite, and it would be advantageous to remove live cells for reuse in the production of more material while harvesting the metabolite from the medium.

One example of the application of ultrasonic separation of algae is a process deployed with Monodus subterraneus UTEX 151 that was followed by enhanced sedimentation as a harvest tool (Bosma et al. 2003). The algal cells were pumped through a chamber equipped with an ultrasonic transduction chamber fitted with a resonator and reflector that provided a standing wave in the chamber. Cells were forced to the nodes and agglomerated such that when the field was removed, they quickly settled due to gravity. High cell densities and slow flow rate were most favorable for ultrasonic separation providing a maximum recovery of 92 % of the cells and an 11-fold increase in concentration.

Another application of ultrasound is the acoustic focusing method recently analyzed within the NAABB consortium’s harvesting group. The use of ultrasound to both separate and lyse the cells was investigated such that both the harvesting and extraction methods could be combined. Using a low-frequency ultrasonic field, the separation of Nannochloropsis salina, Nannochloris oculata, and Auxeno — chlorella protothecoides was demonstrated. Their acoustic focusing harvester was scaled to 45-225 L h-1 and provided an 18-fold increase in biomass density (NAABB 2014).

Many organic carbon sources have been investigated for biodiesel production. These include various sugars such as glucose, galactose, fructose, and even some disaccharides. Polyhydric alcohols, such as glycerol, and some organic acids, such as acetate and propionate, have also been investigated. For example, Chlorella protothecoides can grow on a variety of carbon sources such as glucose (Shen et al. 2010; Xiong et al. 2008; Xu et al. 2006), fructose (Gao et al. 2009), sucrose (Gao et al. 2009), glycerol (Heredia-Arroyo et al. 2010), acetate (Heredia-Arroyo et al.

2010) , and reducing sugars from Jerusalem artichoke and sugarcane (Cheng et al.

2009) . Many species have also been reported to grow heterotrophically on ethanol (Ogbonna et al. 1998; Yokochi et al. 1998), lactose, galactose and mannose (Liang et al. 2009), and molasses (Andrade and Costa 2007). Schizochytrium limacinum produced palmitic acid (16:0) as 45-60 % of their dry weight when supplied with glucose, fructose, or glycerol (Yokochi et al. 1998; Chi et al. 2009). The effec­tiveness of these carbon sources varies with the species as well as on the culture conditions such as the light intensity, the pH, dissolved oxygen concentration, and on the presence of other carbon sources. Some carbon sources are good for mixotrophic culture, but not in heterotrophic cultures. For example, according to

Ceron Garcia et al. (2006), Phaeodactylum tricornutum UTEX-640 did not grow heterotrophically in media containing 0.005-0.2 M of fructose, glucose, mannose, lactose, or glycerol. However, addition of any of these organic carbons in mixo — trophic culture increased the biomass concentration and productivity relative to photoautotrophic controls. The biomass, lipids, eicosapentanoic acid (EPA), and pigment contents were considerably enhanced with glycerol and fructose in relation to photoautotrophic controls. The EPA content was barely affected by the sugars, but was more than twofold higher in glycerol-fed cultures than in photoautotrophic controls (Ceron Garcia et al. 2006).

Liu et al. (1999) compared several carbon sources and concluded that glucose was the best in terms of cell growth rate. This contradicts the work of Chen and Walker (2011) who reported that crude glycerol gave the highest growth of Chlorella protothecoides, followed by pure glycerol, while the least biomass concentration was obtained with glucose. In the case of Chlorella vulgaris, Kong et al. (2011) reported that glucose was the best carbon source for mixotrophic cultivation, followed by sucrose and then glycerin, while sodium acetate did not support good growth. Effectiveness of carbon source in supporting cell growth may depend on their energy content (Chojnacka and Marquez-Rocha 2004; Wang et al. 2012). For instance, glucose produces about 2.8 kJ/mol of energy compared to

0. 8 kJ/mol for acetate (Boyle and Morgan 2009), and glucose was more effective as a substrate for mixotrophic cultivation of Phaeodactylum tricornutum than acetate (Wang et al. 2012). The carbon source that gives high biomass productivity may not be the one that gives high oil production. Although for many strains, glucose has been reported to be the best in terms of cell growth, Das et al. (2011) ranked the effectiveness of different organic substrates in terms of intracellular lipid contents in mixotrophic culture in the following order: glycerol > sucrose > glucose.

The cost is another major factor determining the choice of carbon source for mixotrophic/heterotrophic cultures. The present cost of microalgae oil at US$2.4/L (Li et al. 2007; Xu et al. 2006) is 3-4 times higher than that of plant oils. However, Liu et al. (2010) estimated oil production cost of US$0.9/L for heterotrophic cul­tures of Chlorella zofingiensis using sugar as substrate. In heterotrophic/mixo — trophic cultures, the cost of organic carbon represents a very high percentage of the total production cost. Economic analysis shows that organic carbon source con­tributes 45.4 %; inorganic chemicals, 3.2 %; electricity, 30.6 %; steam, 14.2 %; and aseptic air, 6.6 % of the total production cost (Li et al. 2007; Xu et al. 2006). The cost of glucose has also been estimated to be about 80 % of the total medium cost (Li et al. 2007). Thus, there is a need to drastically reduce the cost of the organic carbon source. Many cheap carbon sources such as non-sugar carbon sources (Heredia-Arroyo et al. 2011), corn powder hydrolysate, impure glycerol, and molasses have been investigated.

Currently glycerol is an inexpensive and abundant carbon source generated as a by-product of biodiesel fuel production. About 0.45 kg of glycerol is produced per

4.5 kg of biodiesel, and the price of crude glycerol is now about 0.025USD/0.45 kg (Chen and Walker 2011). It has been reported that crude glycerol is better than pure glycerol and glucose (Chen and Walker 2011; Liang et al. 2010) because of the residual nitrogen in crude glycerol. The use of corn powder hydrolysate has also been widely investigated, and it has been reported that it is superior to glucose solution since it contains some other components that are beneficial for cell growth. For example, C. protothecoides produced 55.2 % crude lipids in a medium con­taining corn hydrolysate, with a cell dry weight concentration of 15.5 g/L (Xu et al.

2006) , which is higher than the values reported for glucose. Li et al. (2007) noted that if hydrolyzed starch is used as a carbon source for Chlorella, the cost of medium can be reduced to about 60-70 %. Cheng et al. (2009) used hydrolysate of Jerusalem artichoke tuber as a carbon source for heterotrophic cultivation of C. protothecoides, and the resulting biomass contained 44 % lipid. The lipid content of microalgae cultivated in the presence of the enzymatic hydrolyzates of sweet sor­ghum (which contains 10 g/L of reducing sugars) was 52.5 %. This is 35.7 % higher than the value obtained by cultivation using glucose (Gao et al. 2009). Anaerobi­cally digested dairy manure (Wang et al. 2010) and wastewater containing 85-90 % carpet mill effluents (Chinnasamy et al. 2010) were used as carbon sources for production of lipids for biofuel. Many agro-industrial wastes such as dry-grind ethanol thin stillage (TS) and soy whey (SW) have been used as nutrient feedstock for mixotrophic/heterotrophic cultivation of Chlorella vulgaris (Mitra et al. 2012). Both the cell concentration (9.8 g/L) and oil content (43 %) obtained with TS were higher than those obtained with modified basal medium containing glucose as the carbon source (8 g/L and 27 %, respectively) under mixotrophic conditions.

The optimal concentrations of these carbon sources vary with both strain and other culture conditions. Concentrations of glycerol used ranged from 3 to 12 % (Yokochi et al. 1998), while Liang et al. (2009) reported that 1 or 2 % glycerol resulted in a higher lipid content of microalgae compared to the value obtained with 5 % glycerol. The tolerable concentration of glycerol is within the range of

0. 7-10 % (Chi et al. 2009). The optimum cassava starch hydrolysate concentration for cell growth and lipid accumulation was 5 g/L, but the values were not signif­icantly different from those obtained with 10 g/L. A higher concentration of 15 g/L of hydrolysate resulted in lower biomass and lipid contents (Salim 2013). The optimum concentration of glucose for growth was 1 g/L, but there were no significant effects of varying acetate or starch concentrations between 0.5 and 5 g/L on cell growth. The highest values of the lipid content and lipid productivity with glucose in media were approximately 2.8 times (at 2.0 g/L glucose) and 4.6 times (at 1.0 g/L glucose) compared to control (photoautotrophic culture). As the content of glucose increased to 5.0 g/L, the total lipid content and lipid productivity decreased, but were still higher than the values obtained in photoautotrophic culture (Wang et al. 2012). On the other hand, starch in the medium did not influence the specific growth rate with concentrations below 1.0 g/L, but was inhibited signifi­cantly above 2.0 g/L (p < 0.05) (Wang et al. 2012).

One of the areas of microalgae biofuels that must be optimized or re-engineered to make production cost-effective is oil extraction. It is estimated that up to 60 % of the cost of algae biodeisel production involves solvent emulsification and recovery (Molina-Grima et al. 2003). The best studied approach to biodiesel production, first developed for oil seed plants, is the extraction and TAGs or triglycerides into FAMEs. The spent biomass can be used for a variety of applications including biogas, feed, and fertilizers. In the past ten years, a number of alternative forms of catalysis have been developed that use lipases (Ranganathan et al. 2008) or make use of solid bases (hydroxyl groups added to mineral crystals) or catalysts that form biodiesel under high pressure. However, the expense of harvesting, drying, and breaking cell walls remains problematic. In addition, organic solvents for oil extraction are expensive and generate hazardous wastes that must be disposed of at further cost (Williams and Laurens 2010).

HTL has recently been adapted to circumvent many of these problems inherent in lipid extraction processes. HTL is a thermal process that heats a wet slurry of intact algae to 250-350 °C at 1500-3000 psi, converting the biomass to several products including an oil portion ranging from 29 to 52 % yield (See Frank et al. 2013 for a recent review). While TLC produces more oil from algae than lipid extraction, there are several issues with the quality of the oil produced based on the inclusion of other cell components (proteins, nucleic acids, carbohydrates) in the thermal process. A life cycle analysis of TLC of several algal strains (Frank et al. 2012) indicated that the lipid fraction had high levels of N (Williams and Laurens

2010) , leaving questions on combustion emissions for fuel use.

Another recent advancement in the field, supercritical fluid extraction (SFE) of algal biomass, may be an efficient means to extract oils that avoids the use of toxic organic solvents, eliminates the need for the energy-intensive drying of biomass, and avoids high N content in the oil. In addition, SFE allows for the co-extraction of high-value chemicals and leaves a residual biomass that is solvent free and could be marketed as a livestock feed supplement or fertilizer. A carbon dioxide supercritical fluid extraction (CO2-SFE) apparatus for oil extraction using water as a co-solvent would avoid the high cost of drying algae while extracting triglycerides and the co-extraction of valuable nutraceuticals using wet algal biomass.

SFE technology is well developed for processes such as decaffeination and dry­cleaning and is now widely accepted for extraction, purification, and fractionation operations in many industries, especially in the nutraceutical and other “green” industries. SFE is far more efficient than traditional solvent separation methods and is selective, providing high purity of specific products. Additionally, there are no organic solvent residues in the extract or spent biomass. Extraction is efficient at modest operating temperatures, for example, at less than 50 °C, thus ensuring product stability (Herrero et al. 2010). CO2-SFE has been shown to be an efficient solvent for the extraction of a valuable nutraceutical docosahexaenoic acid (DHA) (Couto et al. 2010), for which there is a large growing market.

Based on the literature (Patil and Gude 2011; Choi et al. 1987; Couto et al. 2010; Herrero et al. 2010), there are reasonable starting parameters of temperature, chamber, and release pressures for maximum lipid extraction using dry algal bio­mass. The development of CO2-SFE lipid extraction from wet algae cultures has been explored but is still in its infancy. Adjustments of parameters must be made to use water and methanol co-solvents that alter the overall behavior of the extraction process. Slight variation in temperature will significantly alter the density of the solvent, and therefore the efficiency of the extraction of specific lipoid compounds. An increase in temperature also reduces yield of specific fractions due to product degradation (Patil and Gude 2011; Choi et al. 1987). Cell disruption is a major factor in lipid yields independent of extraction process. Using SFE, cell disruption of wet algae is based on chamber and release pressures, water content, and tem­perature treatment to determine the need for cell lysis prior to CO2-SFE. Halim et al. (2011) have demonstrated the efficiency of SFE extraction of triglyceride fractions from intact wet algal biomass. The extraction efficiency for total lipid extraction including valuable co-products, specifically the marketable nutraceuti — cals, DHA, and luteins, makes the cost-benefit analysis of the whole process favorable. The fractionated products of CO2-SFE, ranging in size from free fatty acids, DHA, triglycerides (the feedstock for biodiesel), and carotenoid compounds such as lutein may be fractionated further using liquid chromatography and the triglycerides transesterified to FAMEs. The parameters for extraction of specific lipids from algae and developing methods that balance cost with production of each specific lipid product that can be scaled up for industrial use is of paramount importance. The CO2 used for extraction can also be recycled to support algal photosynthetic growth. The process outlined here has the potential to be entirely renewable and recyclable as well as cost competitive with liquid fuels.

Once the draft network is ready, a mathematical formulation of the network is made through forming a stoichiometric matrix (S). In the matrix (Fig. 10.3b), rows cor­respond to metabolites, and columns assign reactions. The entries in the matrix correspond to the stoichiometric coefficients of each individual metabolite that contributes to each reaction. A negative coefficient indicates the consumption of a metabolite in a reaction. A positive coefficient indicates the production of a metabolite in a reaction. A zero coefficient is for metabolites that have no contri­bution to the reaction.

When flocculation is used in mining, wastewater treatment or fermentation the resultant sludge is disposed of as a waste product. Yet, when flocculation is used to harvest microalgae, the flocculated biomass is the product. The flocculant ends up in the harvested biomass, which may have consequences for the further valorization of the biomass. Contamination of the harvested biomass is an important factor to consider when considering the use of a flocculation method for harvesting micro­algae. For instance, if the flocculant is toxic or contains toxic residues, the harvested biomass or fractions thereof cannot be used for applications such as food or animal feed. This is important when microalgal biomass is used in a biorefinery context where part of the biomass is used to generate energy (e. g., lipids), and other biomass fractions are used as animal feed (e. g., protein fraction) or high-value ingredients for the food or health industry (e. g., carotenoids) (Wijffels et al. 2010). Rwehumbiza et al. (2012) showed that aluminum added during alum-based floc­culation did not contaminate lipids and fatty acid methyl esters, but it might remain in the protein fraction. Contamination of the biomass with flocculants may also influence the recovery efficiency of certain bioproducts. Rios et al. (2013) observed a lower lipid extraction yield in microalgae harvested by flocculation (by pH increase) than in microalgae harvested by membrane filtration. Borges et al. (2011) noted that the choice of the flocculant used for harvesting microalgae influenced the ratio of saturated over unsaturated fatty acids in the biomass.