17.1 Introduction
The most common approach for bioremediation of contaminated environments
involves stimulation of the microbial activity by overcoming the factors that are
limiting the metabolism of microorganisms (Alvarez and Illman 2005). The
stimulation is usually achieved by supplying electron acceptors/donors to sustain the oxidation/reduction of the contaminants (Alvarez and Illman 2005;
Gkorezis et al. 2016).
In hydrocarbon-contaminated environments, thermodynamically favorable electron acceptors (e.g., O 2 ) are quickly depleted, and, therefore, they need continuous
supplying to sustain the process (Farhadian et al. 2008). Despite the relatively high
redox potential of oxygen (O 2 /H 2 O + 820 mV vs SHE—standard hydrogen electrode, all the potentials reported in this chapter are relative to SHE unless stated
otherwise) and the high efficiency of oxygenases for the aerobic biodegradation of
petroleum hydrocarbons (Baldwin et al. 2009), some disadvantages are linked to the
use of aerobic bioremediation strategies (Daghio et al. 2017). Oxygen solubility in
water is low, and it can be consumed by unwanted side reactions (e.g., by oxidation
of reduced Fe
2+ and Mn
2+ ) (Broden et al. 1997; Tuxen et al. 2006). Several
microorganisms showed the ability to degrade petroleum hydrocarbons in anaerobic
environments, which is, however, usually a slower process compared to aerobic
biodegradation (Weelink et al. 2010). The drawbacks linked to the aerobic biodegradation can be overcome by the use of anaerobic strategies. The addition of
chelators to solubilize Fe
3+ for the microbial respiration or the addition of soluble
electron shuttles (e.g., humic substances) to favor the electron transfer to solid
electron acceptors (e.g., Fe
3+ and Mn
2+ ) has been suggested (Lovley et al. 1994,
1996a, b). Alternatively, anaerobic metabolism can be stimulated by the addition of
nitrate or sulfate (Mihelcic and Luthy 1988; Weiner et al. 1998). The main disadvantage linked to the above-mentioned strategies (both aerobic and anaerobic) is that
the reagents are rapidly consumed and can migrate away from the contaminated area.
The cost of the bioremediation can therefore increase because of the need of
continuous replenishment of reagents to sustain the microbial activity (Daghio
et al. 2017; Zhang et al. 2010).
Similarly to the strategies described above, which can be used for the stimulation of the degradation of non-halogenated hydrocarbons, halogenated hydrocarbons can be remediated by stimulating the reductive dehalogenation (de Bruin
et al. 1992; Seshadri et al. 2005). Hydrogen (H 2 ) is the typical electron donor for
the dehalogenation (Aulenta et al. 2006; He et al. 2002) and it can be supplied
directly or by passive dissolution (Fang et al. 2002; Ma et al. 2003). Alternatively,
H 2 can be indirectly supplied by using organic substrates (e.g., butyric acid,
ethanol, or lactic acid) that can be fermented (Aulenta et al. 2005; Fennell et al.
1997). However, as mentioned above for water-soluble electron acceptors, also
water-soluble electron donors have to be continuously supplied, which can result
in high costs (Daghio et al. 2017). In addition, fermentation products (e.g., volatile
fatty acids) can accumulate, which results in deterioration of the water quality
374
M. Daghio and A. Franzetti
The most common approach for bioremediation of contaminated environments
involves stimulation of the microbial activity by overcoming the factors that are
limiting the metabolism of microorganisms (Alvarez and Illman 2005). The
stimulation is usually achieved by supplying electron acceptors/donors to sustain the oxidation/reduction of the contaminants (Alvarez and Illman 2005;
Gkorezis et al. 2016).
In hydrocarbon-contaminated environments, thermodynamically favorable electron acceptors (e.g., O 2 ) are quickly depleted, and, therefore, they need continuous
supplying to sustain the process (Farhadian et al. 2008). Despite the relatively high
redox potential of oxygen (O 2 /H 2 O + 820 mV vs SHE—standard hydrogen electrode, all the potentials reported in this chapter are relative to SHE unless stated
otherwise) and the high efficiency of oxygenases for the aerobic biodegradation of
petroleum hydrocarbons (Baldwin et al. 2009), some disadvantages are linked to the
use of aerobic bioremediation strategies (Daghio et al. 2017). Oxygen solubility in
water is low, and it can be consumed by unwanted side reactions (e.g., by oxidation
of reduced Fe
2+ and Mn
2+ ) (Broden et al. 1997; Tuxen et al. 2006). Several
microorganisms showed the ability to degrade petroleum hydrocarbons in anaerobic
environments, which is, however, usually a slower process compared to aerobic
biodegradation (Weelink et al. 2010). The drawbacks linked to the aerobic biodegradation can be overcome by the use of anaerobic strategies. The addition of
chelators to solubilize Fe
3+ for the microbial respiration or the addition of soluble
electron shuttles (e.g., humic substances) to favor the electron transfer to solid
electron acceptors (e.g., Fe
3+ and Mn
2+ ) has been suggested (Lovley et al. 1994,
1996a, b). Alternatively, anaerobic metabolism can be stimulated by the addition of
nitrate or sulfate (Mihelcic and Luthy 1988; Weiner et al. 1998). The main disadvantage linked to the above-mentioned strategies (both aerobic and anaerobic) is that
the reagents are rapidly consumed and can migrate away from the contaminated area.
The cost of the bioremediation can therefore increase because of the need of
continuous replenishment of reagents to sustain the microbial activity (Daghio
et al. 2017; Zhang et al. 2010).
Similarly to the strategies described above, which can be used for the stimulation of the degradation of non-halogenated hydrocarbons, halogenated hydrocarbons can be remediated by stimulating the reductive dehalogenation (de Bruin
et al. 1992; Seshadri et al. 2005). Hydrogen (H 2 ) is the typical electron donor for
the dehalogenation (Aulenta et al. 2006; He et al. 2002) and it can be supplied
directly or by passive dissolution (Fang et al. 2002; Ma et al. 2003). Alternatively,
H 2 can be indirectly supplied by using organic substrates (e.g., butyric acid,
ethanol, or lactic acid) that can be fermented (Aulenta et al. 2005; Fennell et al.
1997). However, as mentioned above for water-soluble electron acceptors, also
water-soluble electron donors have to be continuously supplied, which can result
in high costs (Daghio et al. 2017). In addition, fermentation products (e.g., volatile
fatty acids) can accumulate, which results in deterioration of the water quality
374
M. Daghio and A. Franzetti