3. Microbiology of denitrification

In order to carry out denitrification, which is defined as the biological dissimilative transformation of nitrate (NO3 ) or nitrite (NO2 ) into molecular nitrogen (N2) under anoxic conditions with energy conservation [13], an electron donor is required. Therefore, denitrifying microorganisms must have the ability for using nitrate or nitrite as electron acceptors to reduce them to molecular nitrogen. Organotrophic or autotrophic microorganisms are involved in denitrification process depending on their ability to use organic or inorganic compounds as electron sources, respectively. Their remarkable characteristic is their facultative anoxic respiration.

Distribution of denitrifying microorganisms in nature is ubiquitous. Organotrophic and autotrophic denitrifiers belong to α-, β-, γ- and ε-proteobacteria group which comprise both, Gram-negative and Gram-positive bacteria. Nevertheless, some members of Archae and Eukarya have shown the ability for reducing nitrate to N2 [14, 15]. Most of organotrophic and autotrophic denitrifiers grow under neutral and mesophilic conditions [16, 17]. Organotrophic denitrifiers have been found in natural ecosystems as soil [18], surface water [19], groundwater, and sediments [20]; in wastewater treatment plants; and in different types of reactor configurations treating synthetic wastewater under organotrophic [21], autotrophic [22], or mixotrophic conditions, where mixtures of both organic and inorganic electron sources are present [22, 23]. For illustration purposes, several denitrifying microorganisms and their physiological characteristics are included in Table 2.


Table 2. Some denitrifying microorganisms and their physiological characteristics.

Genus β-proteobacteria has been found dominant in many denitrification systems [34]. Thauera is a dominant Gram-negative organotrophic bacterium belonging to β-proteobacteria which has been identified in wastewater treatment systems [35], in an integrated system of three-dimensional biofilm-electrode reactor and sulfur autotrophic denitrification (3DBER-SAD) under mixotrophic conditions [27]. Thauera has also been identified in sequential batch reactors where the heterotrophic and autotrophic denitrifying process was conducted [23] and in several denitrifying bioreactors under autotrophic conditions, suggesting its ability for autotrophic growth [23]. Acidovorax is a Gram-negative bacterium which has the ability of using both acetate and hydrogen for denitrification [29]. Denitrifying bacteria, similar to Acidovorax and Azoarcus, a facultatively anaerobic, mesophilic, and Gram-negative bacterium with the ability of growing with a variety of organic substrates [26], have been identified under mixotrophic denitrifying conditions [27]. Denitrifying species of Acidovorax spp. and Flavobacterium spp. have been detected in a soil column system amended with glucose [21]. Recently, the ability of Pseudomonas sp. C27 for conducting both organotrophic and autotrophic denitrification has been reported [22]. On the other hand, Thiobacillus denitrificans, an obligate autotrophy and facultative anaerobic bacterium, which can use elemental sulfur as an electron donor, has been isolated from natural environments, manmade environments [17], and denitrifying reactors operated under mixotrophic conditions [27].

#### 4. Biochemical aspects

3. Microbiology of denitrification

iological characteristics are included in Table 2.

pantotrophus

thiophillus

) or nitrite (NO2

mation of nitrate (NO3

78 Nitrification and Denitrification

α-Proteobacteria Paracoccus/P.

β-Proteobacteria Thiobacillus

ε-Proteobacteria Sulfurimonas

lithotrophicum

respiration.

In order to carry out denitrification, which is defined as the biological dissimilative transfor-

tions with energy conservation [13], an electron donor is required. Therefore, denitrifying microorganisms must have the ability for using nitrate or nitrite as electron acceptors to reduce them to molecular nitrogen. Organotrophic or autotrophic microorganisms are involved in denitrification process depending on their ability to use organic or inorganic compounds as electron sources, respectively. Their remarkable characteristic is their facultative anoxic

Distribution of denitrifying microorganisms in nature is ubiquitous. Organotrophic and autotrophic denitrifiers belong to α-, β-, γ- and ε-proteobacteria group which comprise both, Gram-negative and Gram-positive bacteria. Nevertheless, some members of Archae and Eukarya have shown the ability for reducing nitrate to N2 [14, 15]. Most of organotrophic and autotrophic denitrifiers grow under neutral and mesophilic conditions [16, 17]. Organotrophic denitrifiers have been found in natural ecosystems as soil [18], surface water [19], groundwater, and sediments [20]; in wastewater treatment plants; and in different types of reactor configurations treating synthetic wastewater under organotrophic [21], autotrophic [22], or mixotrophic conditions, where mixtures of both organic and inorganic electron sources are present [22, 23]. For illustration purposes, several denitrifying microorganisms and their phys-

Group Genus/species Electron donor Physiological characteristics Reference

Thauera Acetate, sulfide, H2 Organotrophic, sulfur and hydrogen

Acidovorax spp. Glucose, acetate, H2 Organotrophic and hydrogen autotrophic

Organic and sulfur compounds, H2

γ-Proteobacteria Pseudomonas sp. Organic compounds, H2 Organotrophic and hydrogen

Table 2. Some denitrifying microorganisms and their physiological characteristics.

) into molecular nitrogen (N2) under anoxic condi-

Organotrophic, sulfur and hydrogen autotrophic denitrification

Sulfide, sulfur Sulfur autotrophic denitrification [25]

autotrophic denitrification

autotrophic denitrification

Sulfur Sulfur autotrophic denitrification [32]

denitrification

Azoarcus Organic compounds Organotrophic denitrification [23, 26, 27]

Flavobacterium spp. Glucose, acetate Organotrophic denitrification [21, 28]

Acinetobacter sp. H2 Hydrogen autotrophic denitrification [30] Aeromonas sp. H2 Hydrogen autotrophic denitrification [31]

Thiomicrospira CVO Sulfur, H2 Sulfur and hydrogen autotrophic

denitrification

[16, 24]

[21, 23]

[24]

[33]

[21, 28, 29]

Irrespective of whether the organic or autotrophic process is conducted, the denitrification process has been described as a modular organization in which every biochemical reaction is catalyzed by a specific reductase [36]. These reactions occur when no oxygen is available and the environment becomes anoxic [37]. According to Mariotti [38], the denitrification process can be described as Eq. (1) indicates.

$$12\text{NO}\_5^- + 10\text{e}^- + 12\text{H}^+ \to \text{N}\_2 + \text{ } 6\text{H}\_2\text{O} \qquad \Delta\text{G}^\circ = -1120.5\text{ K}/\text{reaction} \tag{1}$$

This general equation can be decomposed into four enzymatic reactions. At first, nitrate is reduced to nitrite by nitrate reductase (Nar) (Eq. (2)). The reaction can take place in the cell membrane and periplasmic space. Affinity constant (Km) ranging from 0.15–15 mM and ΔG�' of �163.2 KJ/reaction values have been reported for this reaction [39, 40]. UQH2 corresponds to reduced ubiquinone, UQ to ubiquinone, c2<sup>þ</sup> to reduced cytochrome, and c3<sup>þ</sup> to oxidized cytochrome.

$$\rm{NO\_3^- + UQH\_2 \to NO\_2^- + UQ + H\_2O} \tag{2}$$

A subsequent reduction of nitrite to nitric oxide is carried out by one of two nitrite reductases (Nir, CuNir) or the cytochrome cd1,both located at the periplasmic space (Eq. (3) and (4)). Km values of 3.13–750 µM [41, 42] and 6–46 µM [39, 41] are reported for Nir/CuNir or cd1, respectively, whereas ΔG�' of �73.2 KJ/reaction correspond to this stage.

$$\text{H} \text{(a)} \text{NO}\_2\text{-} + \text{Cu}^{1+} + 2\text{H}^+ \rightarrow \text{NO} + \text{H}\_2\text{O} + \text{Cu}^{2+} \tag{3}$$

or

$$\text{H}(\text{b})\text{NO}\_2 + \text{c}^{2+} + 2\text{H}^+ \rightarrow \text{NO} + \text{H}\_2\text{O} + \text{c}^{3+} \tag{4}$$

Afterward, in the cell membrane, nitric oxide is reduced to nitrous oxide by the enzyme nitric oxide reductase (Nor) (Eq. (5)). Km values of 0.25–60 µM are reported for Nor enzyme with a ΔG�' of �306.3 KJ/reaction [43, 44].

$$2\text{NO} + 2\text{c}^{2+} + 2\text{H}^{+} \rightarrow \text{N}\_{2}\text{O} + \text{H}\_{2}\text{O} + 2\text{c}^{3+} \tag{5}$$

Finally, nitrous oxide is reduced to N2 by the enzyme nitrous oxide reductase (Nos), which is located at the periplasmic space (Eq. (6)). Km values of 2–6 µM are reported for this enzyme with a ΔG�' of �306.3 KJ/reaction [45].

$$\rm N\_2O + 2c^{2+} + 2H^+ \to N\_2 + H\_2O + 2c^{3+} \tag{6}$$

#### 5. Denitrification and its environment

Denitrification performance is controlled by many environmental factors such as concentration, type and solubility of the substrate, C/N ratio, temperature, and pH, among other factors. These environmental variables determine the metabolic behavior, being the effect of each factor different on the biochemistry and physiology of the microorganisms [39, 46]. In this regard, experimental data have suggested that a C/N ratio close to the stoichiometric value is required for complete denitrification [47]. In this respect, many authors have made recommendations to adjust the C/N, S/N ratio for denitrification processes [36, 48]. Tiedje [49] observed that an excess of reducer source induced the reduction of nitrate to ammonium. Denitrification is an exergonic process where the energy formation depends on the type of reducer source. Degradation of monochlorophenols coupled to denitrification is also an exergonic process (Table 3). This makes denitrification a potential microbial biomass producer. Nonetheless, wastewater treatment should be a dissimilatory process where the pollutants might be essentially removed through catabolic processes.

Oxygen is generally considered as a denitrifying inhibitor [50]. Likewise, according to O2 and nitrate potential redox, a competition effect can occur between these oxidants. It has been reported that nitrate could be reduced even in the presence of O2 [51]. On the other hand, the denitrifying process can be carried out in a temperature range between 5 and 35�C. However, it has been observed that at low temperatures, the emissions of nitrous oxide increase whereas N2 formation decreases [52].

pH is an independent variable that affects denitrification process at different levels [46, 53]. The common pH value employed for denitrification is around 7. At low pH values, an inhibition on the reduction of nitrous oxide occurs, causing an accumulation of nitrous oxide and a decrease


<sup>ð</sup>a<sup>Þ</sup> NO2� <sup>þ</sup> Cu<sup>1</sup><sup>þ</sup> <sup>þ</sup> 2H<sup>þ</sup> ! NO <sup>þ</sup> H2O <sup>þ</sup> Cu<sup>2</sup><sup>þ</sup> <sup>ð</sup>3<sup>Þ</sup>

<sup>ð</sup>b<sup>Þ</sup> NO2� <sup>þ</sup> c2<sup>þ</sup> <sup>þ</sup> 2H<sup>þ</sup> ! NO <sup>þ</sup> H2O <sup>þ</sup> c3<sup>þ</sup> <sup>ð</sup>4<sup>Þ</sup>

2NO <sup>þ</sup> 2c2<sup>þ</sup> <sup>þ</sup> 2H<sup>þ</sup> ! N2O <sup>þ</sup> H2O <sup>þ</sup> 2c<sup>3</sup><sup>þ</sup> <sup>ð</sup>5<sup>Þ</sup>

N2O <sup>þ</sup> 2c2<sup>þ</sup> <sup>þ</sup> 2H<sup>þ</sup> ! N2 <sup>þ</sup> H2O <sup>þ</sup> 2c3<sup>þ</sup> <sup>ð</sup>6<sup>Þ</sup>

Afterward, in the cell membrane, nitric oxide is reduced to nitrous oxide by the enzyme nitric oxide reductase (Nor) (Eq. (5)). Km values of 0.25–60 µM are reported for Nor enzyme

Finally, nitrous oxide is reduced to N2 by the enzyme nitrous oxide reductase (Nos), which is located at the periplasmic space (Eq. (6)). Km values of 2–6 µM are reported for this enzyme

Denitrification performance is controlled by many environmental factors such as concentration, type and solubility of the substrate, C/N ratio, temperature, and pH, among other factors. These environmental variables determine the metabolic behavior, being the effect of each factor different on the biochemistry and physiology of the microorganisms [39, 46]. In this regard, experimental data have suggested that a C/N ratio close to the stoichiometric value is required for complete denitrification [47]. In this respect, many authors have made recommendations to adjust the C/N, S/N ratio for denitrification processes [36, 48]. Tiedje [49] observed that an excess of reducer source induced the reduction of nitrate to ammonium. Denitrification is an exergonic process where the energy formation depends on the type of reducer source. Degradation of monochlorophenols coupled to denitrification is also an exergonic process (Table 3). This makes denitrification a potential microbial biomass producer. Nonetheless, wastewater treatment should be a dissimilatory process where the pollutants might be essen-

Oxygen is generally considered as a denitrifying inhibitor [50]. Likewise, according to O2 and nitrate potential redox, a competition effect can occur between these oxidants. It has been reported that nitrate could be reduced even in the presence of O2 [51]. On the other hand, the denitrifying process can be carried out in a temperature range between 5 and 35�C. However, it has been observed that at low temperatures, the emissions of nitrous oxide increase whereas

pH is an independent variable that affects denitrification process at different levels [46, 53]. The common pH value employed for denitrification is around 7. At low pH values, an inhibition on the reduction of nitrous oxide occurs, causing an accumulation of nitrous oxide and a decrease

or

80 Nitrification and Denitrification

with a ΔG�' of �306.3 KJ/reaction [43, 44].

with a ΔG�' of �306.3 KJ/reaction [45].

5. Denitrification and its environment

tially removed through catabolic processes.

N2 formation decreases [52].

Table 3. Stoichiometric reactions of the denitrifying respiratory process using different electron sources and their ΔG�' values (according to Cuervo-López et al. [36]).

in N2 formation [54, 55]. Denitrification can also be influenced by the speciation and bioavailability of the chemical compounds used as reducer sources. Thus, physicochemical conditions must be controlled in order to have a faster and efficient denitrifying process.

#### 6. Biodegradation of chlorophenols under denitrifying conditions

Chlorophenols are generally degraded under anaerobic conditions through the first reductive dechlorination step, which consists of the substitution of chlorine atoms by hydrogen atoms (Eq. (7)).

$$\text{R}-\text{Cl} + \text{H}\_2 \rightarrow \text{R}-\text{H} + \text{HCl} \tag{7}$$

This stage is catalyzed by specific dehalogenases enzymes. The majority of the known reductive dehalogenases belong to the CprA/PceA family and contain one corrinoid and two ironsulfur clusters as cofactors [56]. Reductive dechlorination requires the addition of electron donors. There are other cases in which chlorophenols are used as carbon and energy sources for microorganisms [57]. Under methanogenic conditions, mineralization of various chlorophenols to CO2 and methane has been observed [5]. However, it is unclear if reductive dechlorination would be involved in the degradation of chlorophenols under denitrifying conditions. In fact, different pathways that do not involve the dechlorination reductive step have been suggested [7].

The study of chlorophenols under denitrifying conditions has been mainly evaluated using monochlorophenols. Chang et al. [58] used a biofilm to remove 2-CP under denitrifying conditions in batch cultures. They observed that the nitrate disappeared in 16 h, and there was a consumption of 2-CP. However, there was no formation of phenol in this period, suggesting that 2-CP was not dechlorinated in the presence of nitrate. Phenol was produced only after the disappearance of nitrate, suggesting that nitrate competed with 2-CP as an electron acceptor. A similar behavior was observed by Sanford and Tiedje [8], who evaluated, in serological bottles, the elimination of 2-CP in the presence of nitrate and acetate. They observed that the consumption of 2-CP was inhibited by the presence of nitrate and was only carried out when nitrate disappeared or was found in concentrations lower than 104 mg/L. Yu et al. [59] studied the effect of nitrate addition on the reductive dechlorination of pentachlorophenol (PCP) and found that low concentrations of nitrate (0–62 mg/L) can enhance reductive dechlorination of PCP, whereas high concentrations (310–1860 mg/L) provoke a contrary effect. Thus, reductive dechlorination could be carried out at low concentrations of nitrate. On the other hand, Häggblom et al. [60] studied the removal of three monochlorophenols in batch cultures under denitrifying conditions. Only 2-CP was eliminated in 110 days; nevertheless, they did not detect the formation of phenol as a product of reductive dechlorination. Bae et al. [7] also studied the elimination of monochlorophenols and dichlorophenols under denitrifying conditions in batch cultures, finding that 4-chlorophenol (4-CP), 2,4-dichlorophenol (2,4-DCP), and 2,6-dichlorophenol (2,6-DCP) were not biodegraded, whereas 2-CP and 3-chlorophenol (3-CP) were mineralized and the presence of nitrate was essential. The authors reported that 2-CP was oxidized to CO2 under denitrifying conditions and suggested the presence of a population capable of eliminating 2- CP by a mechanism that does not involve reductive dechlorination.

As observed in Table 4, most of the studies with chlorophenols have been carried out in batch assays and only few types of reactors have been evaluated under denitrifying conditions. Moussavi et al. [61] evaluated the elimination of 2-CP in a granular anoxic baffled reactor (AnBR) increasing the concentration of 2-CP up to 500 mg/L without affecting the efficiency of 2-CP removal, so this could be a feasible process at low cost. Wang et al. [62] evaluated the removal of PCP in a packed reactor with corncob as both carbon source and biofilm support and obtained efficiencies of nitrate and PCP removal above 90%.

In conclusion, mineralization of chlorophenols coupled to denitrification is rarely documented as the total oxidation of chlorophenols to CO2 and reduction of nitrate to N2 have not been corroborated. The available information is controversial as several works evidenced that the presence of nitrate inhibits the transformation of chlorophenols [8, 63], while other authors indicate that reductive dechlorination can be carried out at low concentrations of nitrate [59]. In fact, other studies evidenced that mineralization of chlorophenols is linked to denitrification, and the presence of nitrate was necessary for the biodegradation [7, 64]. In addition, the denitrifying process is often evaluated by the sole nitrate consumption without verifying its total reduction to N2. Therefore, it is necessary to carry out more studies which evaluate the process through response variables such as removal efficiencies, yields of product formation, and rates in order to characterize and better understand the process.

#### 6.1. Strategies for improving the consumption of chlorophenols

It has been pointed out that the main difficulty for the elimination of chlorophenols is the strong stability that the carbon-halogen bond of the aromatic compound confers to its structure [67]. Thus, in many cases, the biodegradation is slow. Several strategies for increasing the consumption


Table 4. Biodegradation of different chlorophenols under denitrifying conditions.

electron acceptor. A similar behavior was observed by Sanford and Tiedje [8], who evaluated, in serological bottles, the elimination of 2-CP in the presence of nitrate and acetate. They observed that the consumption of 2-CP was inhibited by the presence of nitrate and was only carried out when nitrate disappeared or was found in concentrations lower than 104 mg/L. Yu et al. [59] studied the effect of nitrate addition on the reductive dechlorination of pentachlorophenol (PCP) and found that low concentrations of nitrate (0–62 mg/L) can enhance reductive dechlorination of PCP, whereas high concentrations (310–1860 mg/L) provoke a contrary effect. Thus, reductive dechlorination could be carried out at low concentrations of nitrate. On the other hand, Häggblom et al. [60] studied the removal of three monochlorophenols in batch cultures under denitrifying conditions. Only 2-CP was eliminated in 110 days; nevertheless, they did not detect the formation of phenol as a product of reductive dechlorination. Bae et al. [7] also studied the elimination of monochlorophenols and dichlorophenols under denitrifying conditions in batch cultures, finding that 4-chlorophenol (4-CP), 2,4-dichlorophenol (2,4-DCP), and 2,6-dichlorophenol (2,6-DCP) were not biodegraded, whereas 2-CP and 3-chlorophenol (3-CP) were mineralized and the presence of nitrate was essential. The authors reported that 2-CP was oxidized to CO2 under denitrifying conditions and suggested the presence of a population capable of eliminating 2-

As observed in Table 4, most of the studies with chlorophenols have been carried out in batch assays and only few types of reactors have been evaluated under denitrifying conditions. Moussavi et al. [61] evaluated the elimination of 2-CP in a granular anoxic baffled reactor (AnBR) increasing the concentration of 2-CP up to 500 mg/L without affecting the efficiency of 2-CP removal, so this could be a feasible process at low cost. Wang et al. [62] evaluated the removal of PCP in a packed reactor with corncob as both carbon source and biofilm support

In conclusion, mineralization of chlorophenols coupled to denitrification is rarely documented as the total oxidation of chlorophenols to CO2 and reduction of nitrate to N2 have not been corroborated. The available information is controversial as several works evidenced that the presence of nitrate inhibits the transformation of chlorophenols [8, 63], while other authors indicate that reductive dechlorination can be carried out at low concentrations of nitrate [59]. In fact, other studies evidenced that mineralization of chlorophenols is linked to denitrification, and the presence of nitrate was necessary for the biodegradation [7, 64]. In addition, the denitrifying process is often evaluated by the sole nitrate consumption without verifying its total reduction to N2. Therefore, it is necessary to carry out more studies which evaluate the process through response variables such as removal efficiencies, yields of product formation,

It has been pointed out that the main difficulty for the elimination of chlorophenols is the strong stability that the carbon-halogen bond of the aromatic compound confers to its structure [67]. Thus, in many cases, the biodegradation is slow. Several strategies for increasing the consumption

CP by a mechanism that does not involve reductive dechlorination.

82 Nitrification and Denitrification

and obtained efficiencies of nitrate and PCP removal above 90%.

and rates in order to characterize and better understand the process.

6.1. Strategies for improving the consumption of chlorophenols

of chlorophenols have been proposed, although most of them have been conducted under aerobic and anaerobic conditions and in minor proportions underdenitrifying conditions.

Some strategies have been proposed to increase the efficiency and/or rate of chlorophenols consumption. These include the sludge adaptation to pollutants, the use of genetically modified microorganisms, and the addition of alternative carbon sources [68]. It has been also suggested that the addition of readily oxidized carbon sources could exert various beneficial effects, such as decreasing toxicity, acting as an enzyme-inducing agent, or providing reducing power for the consumption of recalcitrant organic compounds [69–71]. Furthermore, Puyol et al. [72] observed accumulation of different intermediates depending on the co-substrate used. When methanol, ethanol, or volatile fatty acids were used as co-substrates, 4-chlorophenol was accumulated while 2,4-dichlorophenol was accumulated when lactate was used as the co-substrate.

Under denitrifying conditions, Hu et al. [66] found that the presence of co-substrates caused a significant decrease in the degradation rate of 4-chlorophenol (by 4 times) while the biodegradation rate of 2,4-dichlorophenol increased by 4.2 times. Therefore, it could be said that the use of co-substrates does not always have a positive effect on the biodegradation of recalcitrant compounds. The compounds used as co-substrates include compounds of easy oxidation and compounds with a structure similar to chlorophenols. Regarding this, Martínez-Gutiérrez et al. [73] evaluated the effects of phenol and acetate on the mineralization of 2-CP by a denitrifying sludge in batch assays. When phenol was used as a co-substrate, the specific rate of 2-CP consumption increased by 2.6 times, regarding to a control assay without co-substrate. When acetate was used, the specific rate of 2-CP consumption increased by 9 times, suggesting that the addition of co-substrates is a good alternative for improving the biodegradation of chlorophenols. These results also suggest that the effects of co-substrates addition depend on several factors: type of both the co-substrate and chlorophenol employed, inoculum source, and experimental conditions.

#### 7. Coupled systems for chlorophenol degradation

Recently, other strategies have been developed for the elimination of recalcitrant compounds using systems that combine advanced oxidation (AOP) or electrochemical processes with biological processes. Daghio et al. [74] evaluated the degradation of toluene using bio-electrochemical reactors obtaining a current power of 431 mA/m<sup>2</sup> . Yeruva et al. [75] evaluated the integration of a sequencing batch reactor (SBR) and a bio-electrochemical treatment system (BET) for the treatment of petrochemical wastewater under anoxic conditions, obtaining high degradation and power generation (17.12 mW/m2 ). The application of an electrochemical treatment can diminish the time required for the treatment of chlorinated pesticides in the biological process [76]. A sequential biological advanced oxidation process was used for the degradation of 2,4-dichlorophenol, consisting of an up-flow anaerobic sludge blanket (UASB) reactor and a UV/H2O2/TiO2 system, obtaining 52.7% of degradation in only 6 h [77]. However, the degradation of chlorophenols with nitrate using combined systems has been scarcely evaluated. In this sense, Arellano-González et al. [78] evaluated an electrochemical-biological combined system where the reductive dechlorination was carried out in an ECCOCEL-type (Pd-Ni/Ti electrode) reactor that achieved 100% transformation of 2-CP into phenol. Then, the phenol formed was mineralized by a biological denitrification process. The total time required for 2-CP conversion into CO2 was 7.5 h.

#### 8. Perspectives

Biodegradation processes of chlorophenols have been studied extensively because they are more economical and friendly environmental processes in comparison with physicochemical, AOP, and electrochemical processes. The information presented in this review shows that denitrification might be an efficient biological process for the treatment of effluents contaminated with nitrogen and chlorophenols. It has been also reported that biological processes may achieve the complete removal of many types of chlorophenols under aerobic and anaerobic conditions, but they do not always lead to mineralization. It is crucial considering that biodegradation processes can generate more toxic and recalcitrant intermediates than the original pollutant, and the partial oxidation of recalcitrant molecules should be prevented, favoring their mineralization. In this review, it is shown that recent experimental evidences demonstrated the possibility to use denitrification for 2-CP mineralization associated with the reduction of nitrate to nitrogen gas. These results suggest that denitrification might be used for the mineralization of chlorophenols producing CO2 and N2 as final products and obtaining high removal efficiencies. However, more studies on chlorophenols biodegradation by denitrifying processes are needed, especially with mixtures of chlorophenols. More studies on physiological, kinetic, and biochemical aspects of denitrification are also required to identify the limiting steps of the biodegradation metabolic pathways and to better understand how controlling denitrifying processes in bioreactors without the formation of undesirable by-products.

compounds with a structure similar to chlorophenols. Regarding this, Martínez-Gutiérrez et al. [73] evaluated the effects of phenol and acetate on the mineralization of 2-CP by a denitrifying sludge in batch assays. When phenol was used as a co-substrate, the specific rate of 2-CP consumption increased by 2.6 times, regarding to a control assay without co-substrate. When acetate was used, the specific rate of 2-CP consumption increased by 9 times, suggesting that the addition of co-substrates is a good alternative for improving the biodegradation of chlorophenols. These results also suggest that the effects of co-substrates addition depend on several factors: type of both the co-substrate and chlorophenol employed, inoculum source,

Recently, other strategies have been developed for the elimination of recalcitrant compounds using systems that combine advanced oxidation (AOP) or electrochemical processes with biological processes. Daghio et al. [74] evaluated the degradation of toluene using bio-electro-

integration of a sequencing batch reactor (SBR) and a bio-electrochemical treatment system (BET) for the treatment of petrochemical wastewater under anoxic conditions, obtaining high

treatment can diminish the time required for the treatment of chlorinated pesticides in the biological process [76]. A sequential biological advanced oxidation process was used for the degradation of 2,4-dichlorophenol, consisting of an up-flow anaerobic sludge blanket (UASB) reactor and a UV/H2O2/TiO2 system, obtaining 52.7% of degradation in only 6 h [77]. However, the degradation of chlorophenols with nitrate using combined systems has been scarcely evaluated. In this sense, Arellano-González et al. [78] evaluated an electrochemical-biological combined system where the reductive dechlorination was carried out in an ECCOCEL-type (Pd-Ni/Ti electrode) reactor that achieved 100% transformation of 2-CP into phenol. Then, the phenol formed was mineralized by a biological denitrification process. The total time required

Biodegradation processes of chlorophenols have been studied extensively because they are more economical and friendly environmental processes in comparison with physicochemical, AOP, and electrochemical processes. The information presented in this review shows that denitrification might be an efficient biological process for the treatment of effluents contaminated with nitrogen and chlorophenols. It has been also reported that biological processes may achieve the complete removal of many types of chlorophenols under aerobic and anaerobic conditions, but they do not always lead to mineralization. It is crucial considering that biodegradation processes can generate more toxic and recalcitrant intermediates

. Yeruva et al. [75] evaluated the

). The application of an electrochemical

and experimental conditions.

84 Nitrification and Denitrification

7. Coupled systems for chlorophenol degradation

chemical reactors obtaining a current power of 431 mA/m<sup>2</sup>

degradation and power generation (17.12 mW/m2

for 2-CP conversion into CO2 was 7.5 h.

8. Perspectives

Another important aspect is that it has been shown that chlorophenols biodegradation by denitrifying microorganisms is very slow. As a consequence, the application of denitrification processes for chlorophenols removal is still limited, requiring very long acclimation and retention times, especially for the treatment of wastewaters contaminated with high chlorophenol concentrations. Different treatment alternatives have been proposed in order to increase the rate and efficiency of chlorophenol consumption and among them are adaptation to the pollutants, utilization of genetically modified microorganisms, and addition of alternative sources of energy. However, in spite of the addition of co-substrates, the time required for complete mineralization of chlorophenols can be still very long compared to those obtained in physicochemical processes. In recent years, there have been proposals for coupling oxidation processes (AOP or electrochemical) to biological processes such as denitrification to combine benefits of both types of treatment and establish more efficient, more rapid, less expensive, and environmentally friendly treatment trains for degradation of recalcitrant compounds in wastewater. One alternative is the pretreatment of chlorophenols containing effluents through chemical or electrochemical processes to make them more easily degradable in a sequencing denitrifying biological treatment. Recent results showed that times can be considerably reduced for the complete mineralization of 2-CP in an electrochemical-biological combined system, where an electrocatalytic dehydrogenation process (reductive dechlorination) was coupled to a biological denitrification process in sequential ECCOCEL-type (Pd-Ni/Ti electrode) and rotating cylinder denitrifying reactors. The total time required for 2-CP mineralization in the combined electrochemical-biological process was close to the previously reported times for electrochemical and AOP processes, but in this case, an efficient process was obtained without accumulation of by-products or generation of excessive energy costs due to the selective electrochemical pretreatment. This study showed that the use of electrochemical reductive pretreatment combined with denitrification could be a promising technology for the removal of recalcitrant molecules, such as chlorophenols, from wastewater by more efficient, rapid, and environmentally friendly processes. However, more studies are required in order to get an insight about the denitrification of electrochemically pretreated effluents in different combined systems, different configurations of reactors, and in the presence of different mixtures of chlorophenols and types of co-substrates.
