Biological Remediation of Phenoxy Herbicide-Contaminated Environments

*Magdalena Urbaniak and Elżbieta Mierzejewska*

## **Abstract**

Phenoxy herbicides such as 2,4-dichlorophenoxyacetic acid (2,4-D) and 2-methyl-4-chlorophenoxyacetic acid (MCPA) are widely used in agriculture to control broadleaf weeds. Although their application has helped to increase the yield and value of crops, they are also recognized as a source of emerging environmental contamination. Their extensive use may promote contamination of soil, surface, and groundwater and lead to increased inhibition of plant development and soil toxicity. Hence, there is an urgent need to identify nature-based methods based on appropriate biological remediation techniques, such as bio-, phyto-, and rhizoremediation, that enable the effective elimination of phenoxy herbicides from the environment. Bioremediation typically harnesses microorganisms and their ability to utilize recalcitrant contaminants in complete degradation processes, while phytoremediation is a cost-effective, environmentally friendly strategy that uses plants to transform or mineralize xenobiotics to less or nontoxic compounds. Rhizoremediation (microbe-assisted phytoremediation), in turn, is based on the interactions between plant roots, root exudates enriched in plant secondary metabolites, soil, and microorganisms. Based on the above, this chapter presents current knowledge on the properties of phenoxy herbicides, as well as the concentrations detected in the environment, their toxicity, and the biological remediation techniques used for safe removal of the compounds of interest from the environment.

**Keywords:** 2,4-D, MCPA, bioremediation, phytoremediation, rhizoremediation, toxicity, degradative genes

## **1. Phenoxy herbicides: general information**

2,4-Dichlorophenoxyacetic acid (2,4-D) and 2-methyl-4-chlorophenoxyacetic acid (MCPA) are the most commonly used phenoxy acid herbicides in agriculture, and 2,4-D is now the fifth most extensively used active ingredient (a.i.) in the US agricultural and home/garden market sector [1]. In addition, in 2016, 6.5 mln kg of herbicides based on phenoxy-phytohormones (2,4-D and MCPA) were sold in in the EU, including ~2 mln kg sold in Poland [2].

Phenoxy herbicides are typically used to protect wheat, one of the most extensively cultivated crops, because they selectively control the growth of dicotyledonous weeds [3]. They are applied as post emergence agents and taken up by broad-leaved plants. 2,4-D has also been extensively used as an anti-stalling agent for the postharvest fresh fruit industry [4].

These herbicides are based on ring-like structures and have at least one chlorine atom attached to the ring at different positions [5]. Their action is similar to that of phytohormones (auxins) insofar that they can redirect the regulation of plant growth/physiological processes, resulting in nutrition deficiency and subsequent plant death [6].

They are typically released to the environment in the form of commercial products containing phenoxy acids salts or esters; however, they immediately hydrolyze to their corresponding anionic or neutral form [7]. The dosage of phenoxy herbicides lies in the range of 0.8–1.8 kg of a.i. per ha. Their transport through the environment is governed by soil and climate factors (e.g., distribution of soil particles, soil permeability, soil depth, soil pH, soil organic matter content, land slope) [8], and their retention and translocation in the soil profile also depend on their chemical and physical properties, which are described by several parameters (**Table 1**), particularly pKa (acid dissociation constant), logP (octanol-water partition coefficient), and Koc (organic carbon distribution coefficient). The degree of adsorption and desorption depends on time and the physicochemical properties of soil; however, 2,4-D and MCPA are rather poorly adsorbed on the soil particles in comparison to their derivatives, which have different sorption characteristics [7].

Although phenoxy herbicides are described as nonpersistent and weakly adsorbed (Koc < 50) in soil, they can be transported with runoff and in soil profile and reach terrestrial and water ecosystems (surface and groundwater). **Figure 1** summarizes the transport and transfer processes of phenoxy herbicides in the environment. After they are applied to land, they are spread through several processes, including sorption/desorption, leaching, runoff, and plant uptake [8]. Phenoxy herbicide molecules are negatively charged and are therefore highly mobile at neutral pH. In groundwater, they are nonvolatile and persistent to hydrolysis, but they can be degraded biologically under both aerobic and anaerobic conditions. These herbicides demonstrate significantly greater persistence in temperate climates characterized by low winter temperatures and, in many regions, by depleted soil organic carbon content and acidic pH [7].


**83**

water resources.

*Biological Remediation of Phenoxy Herbicide-Contaminated Environments*

**2. Phenoxy herbicides: potential contaminants of soil and water** 

*Transport and transfer processes of phenoxy herbicides in the environment.*

Extensive use of phenoxy herbicides can threaten surface and groundwater ecosystems by promoting the contamination of soil matrices. The International Agency for Research on Cancer classifies phenoxy acids as "possibly carcinogenic to humans." Gupta et al. [11] report that 0.5 kg/ha is the optimal concentration of 2,4-D which avoids contamination of environmental matrices, with the effect of higher concentrations of 2,4-D on the environment being dependent on irrigation treatment. Hence, little is known of the distribution of phenoxy herbicides in the environment. Data from several sources have identified increased levels of 2,4-D and MCPA in the soil, ground-, surface, and drinking water (**Table 2**); for example, Ignatowicz and Struk-Sokołowska [12] note that the concentration of phenoxy herbicides in the Narew River (Poland) fluctuated seasonally from 0 to even 150 μg/L. The concentration of MCPA in the Parramatta River (Sydney Estuary, Australia) was 0.061 μg/L; however, its presence in river water was caused by increased runoff of storm water [13]. MCPA concentration has been found to be as high as 42.40 μg/L in the Rhone River (France) [14] and to be as little as 0.58 μg/L in Brejo of Cagarrão Stream (Portugal) [15]. The 2,4-D concentration has been found to vary from 1.678 μg/L in the water of McGregor Creek (Canada) [16] to 329.42 μg/L in water from a rice field [17]. By contrast, the maximum permissible concentration of pesticide residues in drinking water is 0.50 μg/L (Directive E98/83/EC). The data presented in **Table 2** and described above indicate that phenoxy herbicides should be considered as emerging contaminant especially in

Despite the diversified levels of phenoxy herbicides noted in worldwide environments (**Table 2**), it has to be underlined that these compounds can exert serious toxic effects on the sustainability of ecosystems, even at lower concentrations (e.g., 0.275 μg/L) (**Table 3**). According to recent research, predicted no effect concentration (PNEC) for aquatic organisms is 500 μg/L for 2,4-D and 0.022 μg/L for MCPA [20]; however, PNEC has not yet been determined for terrestrial organisms.

Because the mode of action of phenoxy herbicides mimics that of plant growth hormones, their application causes disturbances among a range of physiological processes [21]. 2,4-D inhibits root/hypocotyl elongation in *Sinapis arvensis* (wild mustard) and disrupts mesophyll cell structure in *Pisum sativum* (pea) [22, 23]. There is increasing concern that 2,4-D has negative influence on water ecosystems, leading to cellular deformation of green algae, such as *Ankistrodesmus falcatus* [24];

*DOI: http://dx.doi.org/10.5772/intechopen.88256*

**environments**

**Figure 1.**

#### **Table 1.**

*Physical and chemical properties of 2,4-D and MCPA.*

**Figure 1.**

*Environmental Chemistry and Recent Pollution Control Approaches*

organic carbon content and acidic pH [7].

*Physical and chemical properties of 2,4-D and MCPA.*

Chemical structure

Molar mass (g/mol)

t1/2 in water [9, 10]

Solubility in H2O (mg/L) [7]

plant death [6].

These herbicides are based on ring-like structures and have at least one chlorine atom attached to the ring at different positions [5]. Their action is similar to that of phytohormones (auxins) insofar that they can redirect the regulation of plant growth/physiological processes, resulting in nutrition deficiency and subsequent

They are typically released to the environment in the form of commercial products containing phenoxy acids salts or esters; however, they immediately hydrolyze to their corresponding anionic or neutral form [7]. The dosage of phenoxy herbicides lies in the range of 0.8–1.8 kg of a.i. per ha. Their transport through the environment is governed by soil and climate factors (e.g., distribution of soil particles, soil permeability, soil depth, soil pH, soil organic matter content, land slope) [8], and their retention and translocation in the soil profile also depend on their chemical and physical properties, which are described by several parameters (**Table 1**), particularly pKa (acid dissociation constant), logP (octanol-water partition coefficient), and Koc (organic carbon distribution coefficient). The degree of adsorption and desorption depends on time and the physicochemical properties of soil; however, 2,4-D and MCPA are rather poorly adsorbed on the soil particles in comparison to their derivatives, which have different sorption characteristics [7]. Although phenoxy herbicides are described as nonpersistent and weakly adsorbed (Koc < 50) in soil, they can be transported with runoff and in soil profile and reach terrestrial and water ecosystems (surface and groundwater). **Figure 1** summarizes the transport and transfer processes of phenoxy herbicides in the environment. After they are applied to land, they are spread through several processes, including sorption/desorption, leaching, runoff, and plant uptake [8]. Phenoxy herbicide molecules are negatively charged and are therefore highly mobile at neutral pH. In groundwater, they are nonvolatile and persistent to hydrolysis, but they can be degraded biologically under both aerobic and anaerobic conditions. These herbicides demonstrate significantly greater persistence in temperate climates characterized by low winter temperatures and, in many regions, by depleted soil

**Properties 2,4-D MCPA**

pKa [7] 2.73–2.87 3.73 Koc [7] 20.56 mL/g 25–157 mL/g logP [7] 2.50/2.58 ± 0.36 2.41/2.49 ± 0.27

t1/2 in soil [9, 10] 2–4 weeks 3–4 weeks

IUPAC name (2,4-Dichlorophenoxy)acetic acid (4-Chloro-2-methylphenoxy)acetic acid

220.04 200.62

450 720

1–14 days 15 days

**82**

**Table 1.**

*Transport and transfer processes of phenoxy herbicides in the environment.*

## **2. Phenoxy herbicides: potential contaminants of soil and water environments**

Extensive use of phenoxy herbicides can threaten surface and groundwater ecosystems by promoting the contamination of soil matrices. The International Agency for Research on Cancer classifies phenoxy acids as "possibly carcinogenic to humans." Gupta et al. [11] report that 0.5 kg/ha is the optimal concentration of 2,4-D which avoids contamination of environmental matrices, with the effect of higher concentrations of 2,4-D on the environment being dependent on irrigation treatment. Hence, little is known of the distribution of phenoxy herbicides in the environment. Data from several sources have identified increased levels of 2,4-D and MCPA in the soil, ground-, surface, and drinking water (**Table 2**); for example, Ignatowicz and Struk-Sokołowska [12] note that the concentration of phenoxy herbicides in the Narew River (Poland) fluctuated seasonally from 0 to even 150 μg/L. The concentration of MCPA in the Parramatta River (Sydney Estuary, Australia) was 0.061 μg/L; however, its presence in river water was caused by increased runoff of storm water [13]. MCPA concentration has been found to be as high as 42.40 μg/L in the Rhone River (France) [14] and to be as little as 0.58 μg/L in Brejo of Cagarrão Stream (Portugal) [15]. The 2,4-D concentration has been found to vary from 1.678 μg/L in the water of McGregor Creek (Canada) [16] to 329.42 μg/L in water from a rice field [17]. By contrast, the maximum permissible concentration of pesticide residues in drinking water is 0.50 μg/L (Directive E98/83/EC). The data presented in **Table 2** and described above indicate that phenoxy herbicides should be considered as emerging contaminant especially in water resources.

Despite the diversified levels of phenoxy herbicides noted in worldwide environments (**Table 2**), it has to be underlined that these compounds can exert serious toxic effects on the sustainability of ecosystems, even at lower concentrations (e.g., 0.275 μg/L) (**Table 3**). According to recent research, predicted no effect concentration (PNEC) for aquatic organisms is 500 μg/L for 2,4-D and 0.022 μg/L for MCPA [20]; however, PNEC has not yet been determined for terrestrial organisms.

Because the mode of action of phenoxy herbicides mimics that of plant growth hormones, their application causes disturbances among a range of physiological processes [21]. 2,4-D inhibits root/hypocotyl elongation in *Sinapis arvensis* (wild mustard) and disrupts mesophyll cell structure in *Pisum sativum* (pea) [22, 23]. There is increasing concern that 2,4-D has negative influence on water ecosystems, leading to cellular deformation of green algae, such as *Ankistrodesmus falcatus* [24];

### *Environmental Chemistry and Recent Pollution Control Approaches*


#### **Table 2.**

*Concentration of phenoxy acids observed in various environments.*


**85**

*Biological Remediation of Phenoxy Herbicide-Contaminated Environments*

causes internal hemorrhage and behavioral changes in *C. carpio*.

malformations and behavioral changes to various fish, including *Cyprinus carpio* (common carp) and *Danio rerio* (zebrafish) [25, 26]; abnormal cellular proliferation in amphibians such as *Rhinella arenarum* (species of toad) [27]; and the development of nonviable embryos in invertebrates, such as *Biomphalaria glabrata* (species of freshwater snail) [28]. Sarikaya and Yilmaz [26] report that 2,4-D (66,000 μg/L)

Among animals, phenoxy herbicide application results in the inhibition of crucial enzymes in cell metabolism, including mitochondrial enzymes and those associated with DNA synthesis (**Table 3**). 2,4-D has also been found to induce erythrocyte lysis under laboratory conditions [29]. It is interesting to note that the intermediates formed during the degradation processes of 2,4-D, such as

2,4-dichlorophenol (2,4-DCP) and 3,5-dichlorocatechol (3,5-DCC), exhibit a strong ecotoxic effect on various organisms, including *N. tabacum* cells [30]; however, Taylor et al. [31] report that 2,4-DCP toxicity was found to be less phytotoxic than 2,4-D, under both in vitro and in vivo conditions, while 3,5-DCC exhibits higher

Several studies have revealed that MCPA also can have a negative impact on the environment: MCPA application caused up to a 56% reduction in dehydrogenase, urease, and phosphatase activities and ergosterol content in soil [32]. In addition, this application leads to increased soil phytotoxicity to *Fagopyrum esculentum* var. Kora (buckwheat) and promoted stem deformation and leaf discoloration [33]. Mierzejewska et al. [34] note that a commercial product containing MCPA was highly toxic to the monocotyledon *Sorghum saccharatum* (sorghum) and dicotyledons *Lepidium sativum* (garden cress) and *Sinapis alba* (white mustard), inducing nearly 100% root growth inhibition. The authors also note that after 3 weeks of incubation at an ambient temperature, the high initial phytotoxicity was reduced to 3% for *L. sativum* and 34% for *S. alba* and that *S. saccharatum* demonstrated a 12% stimulation of root growth in comparison to uncontaminated control soil. The negative influence of MCPA on *L. sativum*, *S. alba*, and *S. saccharatum* growth was also confirmed by Urbaniak et al. [35]. Similarly to 2,4-D, MCPA causes also negative effects on freshwater organisms such as the freshwater crustaceans *Daphnia magna*, *Thamnocephalus platyurus*, and *Artemia franciscana* and alga *Selenastrum capricornutum* [36]. Both herbicides were found to induce the action of hepatic enzymes

*DOI: http://dx.doi.org/10.5772/intechopen.88256*

toxicity than its parent compound [32].

involved in detoxification and lipid peroxidation [21].

herbicide elimination from the environment.

These studies emphasize the important role played by ecotoxicological approaches in evaluating the effect of chemical stressors observed in the ecosystem communities. Despite the relatively short half-life (**Table 1**) of 2,4-D and MCPA, their remnants can be transported and deposited extensively in the environment, and this can present a potential threat to the soil and water ecosystems as well as to human health. Therefore, there is a need to identify nature-based solutions such as bio-, phyto-, and rhizoremediation that can enhance the process of phenoxy

One approach to removing phenoxy herbicides (2,4-D and MCPA) from soil is via degradation by the soil microbiota (biodegradation). This is achieved most effectively by bacteria harboring the appropriate functional genes, which are involved in the phenoxy herbicide degradation pathways (**Figure 2**). Alternatively, plants can be used to decontaminate sites, a process known as phytoremediation (**Figure 2**). Another promising approach, rhizoremediation, enhances the removal

**3. Phenoxy herbicides: removal using biological methods**

#### **Table 3.**

*The results of toxicological tests and effects of 2,4-D and MCPA on selected organisms.*

#### *Biological Remediation of Phenoxy Herbicide-Contaminated Environments DOI: http://dx.doi.org/10.5772/intechopen.88256*

*Environmental Chemistry and Recent Pollution Control Approaches*

**Compound Concentration Environmental matrices Source** 2,4-D 1.678 μg/L Water from McGregor Creek (Canada) [16] 2,4-D 103.99–329.42 μg/L Water from rice field (Malaysia) [14] 2,4-D 0.0052 mg/kg Soil from cereal plantations (Poland) [15] 2,4-D 0.513 μg/L Lebo drain [16]

MCPA 0.0046 mg/kg Soil from cereal plantations (Poland) [15] MCPA 0.08–42.40 μg/L Water from Rhône River Delta (France) [18]

MCPA 82.75–354.28 μg/L Water from rice field (Malaysia) [14] MCPA 0.002–0.010 mg/kg Soil from potato plantation (Poland) [19]

MCPA 0.58 μg/L Water from Brejo of Cagarrão Stream

MCPA 0.061 μg/L Water from Parramatta River—Sydney

0–150 μg/L Water from Narew River (Poland) [12]

(Portugal)

Estuary (Australia)

**84**

**Dose of phenoxy herbicide**

**Table 2.**

MCPA, MCPP, 2,4-D

10, 100, 500, 1000 μg/L MCPA

2,4-D

2,4-D

IC50 1353.80 mg/L

IC50 71.20 mg/L

LC50 66 mg/L 2,4-D

LC50 9.06 and 7.76 mg/L 2,4-D

0.275, 2.75, and 27.5 μg/L 2,4-D and

MCPA

**Table 3.**

**Exposure time**

*Concentration of phenoxy acids observed in various environments.*

(wild mustard)

(waterthyme)

(green microalgae)

*aeruginosa* (toxigenic cyanobacteria)

(common carp)

*Rhinella arenarum* (species of toad)

10–500 mg/L 2,4-D 1 hour Human erythrocytes Hemolysis [29]

*Metynnis roosevelti* (species of serrasalmid

5.06 mg/L 2,4-D 72 hours *Pisum sativum* (pea) Severe disturbances in

7 days *Hydrilla verticillata*

96 hours *Microcystis* 

96 hours *Cyprinus carpio*

30 minutes Hepatic cells of

*The results of toxicological tests and effects of 2,4-D and MCPA on selected organisms.*

fish)

96 hours *Ankistrodesmus falcatus*

220.04 μg/L 2,4-D 2 days *Sinapis arvensis*

96 and 168 hours **Test organism Effect on organism Source**

leaves

alterations

Stimulation of the production of cyanotoxins

Reduced body size, delayed development, microcephaly, agenesis of gills, abnormal cellular proliferation processes

Damage of cellular metabolism and homeostasis; increased oxidative stress

Inhibition of root and hypocotyl elongation

mesophyll cell structure and proliferation of vascular tissue in young

Disturbance of growth, anatomy, and physiology

External morphological

Behavioral changes [26]

[22]

[17]

[13]

[23]

[36]

[24]

[27]

[21]

malformations and behavioral changes to various fish, including *Cyprinus carpio* (common carp) and *Danio rerio* (zebrafish) [25, 26]; abnormal cellular proliferation in amphibians such as *Rhinella arenarum* (species of toad) [27]; and the development of nonviable embryos in invertebrates, such as *Biomphalaria glabrata* (species of freshwater snail) [28]. Sarikaya and Yilmaz [26] report that 2,4-D (66,000 μg/L) causes internal hemorrhage and behavioral changes in *C. carpio*.

Among animals, phenoxy herbicide application results in the inhibition of crucial enzymes in cell metabolism, including mitochondrial enzymes and those associated with DNA synthesis (**Table 3**). 2,4-D has also been found to induce erythrocyte lysis under laboratory conditions [29]. It is interesting to note that the intermediates formed during the degradation processes of 2,4-D, such as 2,4-dichlorophenol (2,4-DCP) and 3,5-dichlorocatechol (3,5-DCC), exhibit a strong ecotoxic effect on various organisms, including *N. tabacum* cells [30]; however, Taylor et al. [31] report that 2,4-DCP toxicity was found to be less phytotoxic than 2,4-D, under both in vitro and in vivo conditions, while 3,5-DCC exhibits higher toxicity than its parent compound [32].

Several studies have revealed that MCPA also can have a negative impact on the environment: MCPA application caused up to a 56% reduction in dehydrogenase, urease, and phosphatase activities and ergosterol content in soil [32]. In addition, this application leads to increased soil phytotoxicity to *Fagopyrum esculentum* var. Kora (buckwheat) and promoted stem deformation and leaf discoloration [33]. Mierzejewska et al. [34] note that a commercial product containing MCPA was highly toxic to the monocotyledon *Sorghum saccharatum* (sorghum) and dicotyledons *Lepidium sativum* (garden cress) and *Sinapis alba* (white mustard), inducing nearly 100% root growth inhibition. The authors also note that after 3 weeks of incubation at an ambient temperature, the high initial phytotoxicity was reduced to 3% for *L. sativum* and 34% for *S. alba* and that *S. saccharatum* demonstrated a 12% stimulation of root growth in comparison to uncontaminated control soil. The negative influence of MCPA on *L. sativum*, *S. alba*, and *S. saccharatum* growth was also confirmed by Urbaniak et al. [35]. Similarly to 2,4-D, MCPA causes also negative effects on freshwater organisms such as the freshwater crustaceans *Daphnia magna*, *Thamnocephalus platyurus*, and *Artemia franciscana* and alga *Selenastrum capricornutum* [36]. Both herbicides were found to induce the action of hepatic enzymes involved in detoxification and lipid peroxidation [21].

These studies emphasize the important role played by ecotoxicological approaches in evaluating the effect of chemical stressors observed in the ecosystem communities. Despite the relatively short half-life (**Table 1**) of 2,4-D and MCPA, their remnants can be transported and deposited extensively in the environment, and this can present a potential threat to the soil and water ecosystems as well as to human health. Therefore, there is a need to identify nature-based solutions such as bio-, phyto-, and rhizoremediation that can enhance the process of phenoxy herbicide elimination from the environment.

## **3. Phenoxy herbicides: removal using biological methods**

One approach to removing phenoxy herbicides (2,4-D and MCPA) from soil is via degradation by the soil microbiota (biodegradation). This is achieved most effectively by bacteria harboring the appropriate functional genes, which are involved in the phenoxy herbicide degradation pathways (**Figure 2**). Alternatively, plants can be used to decontaminate sites, a process known as phytoremediation (**Figure 2**). Another promising approach, rhizoremediation, enhances the removal

**Figure 2.**

*The biological processes of phenoxy herbicide biodegradation mediated by soil, rhizospheric, and endophytic bacteria.*

of such recalcitrant xenobiotics from the environment by exploiting the interactions between selected plants (able to grow under the presence of given xenobiotics such as phenoxy herbicides), root exudates (including plant secondary metabolites, PSMs), and microorganisms (**Figure 2**). The purpose of this section is to review the literature on established and potential biological methods of phenoxy herbicide removal from environmental matrices.

#### **3.1 Bioremediation**

Bioremediation is a method that uses microbiological processes to degrade or transform contaminants to less toxic or nontoxic forms. Biodegradation of organic contaminants occurs very slowly in bulk soil; therefore biostimulation and bioaugmentation methods are used to enhance the biologically driven removal of toxic compounds from environmental matrices. The effectiveness of biodegradation is dependent on several factors, among them the characteristics of the soil, the bioavailability of the contaminants, and their chemical properties.

An important way of phenoxy herbicide removal from soil is by the use of indigenous soil bacteria harboring desirable catabolic genes. The first step in the phenoxy herbicide biodegradation pathway is initiated by α-ketoglutaratedependent dioxygenase, an enzyme encoded by *tfdA* or *tfdA*-like genes [37] located in the *tfdABCDEF* gene cluster [38].

In recent decades, increasingly rapid advances in the application of molecular analysis in environmental studies have helped identify the bacterial communities involved in phenoxy herbicide biodegradation (**Table 4**). The bacteria able to metabolize phenoxy herbicides have been classified into three groups as follows: according to their physiology, employed degrading enzymes, and evolutionary origin [39–41] (**Table 4**).

**87**

*Biological Remediation of Phenoxy Herbicide-Contaminated Environments*

with lower activity than JMP134 dioxygenase.

1.The first group consists of fast-growing copiotrophic bacteria belonging to β- and γ-proteobacteria harboring the *tfdA* gene (e.g., *Cupriavidus necator* JMP134, *Burkholderia* sp. strain RASC, and *Rhodoferax* sp. strain P230). This first group has been subdivided into four subclasses according to *tfdA* sequence: *tfdA* Class I, II, and III [42, 43] and *tfdA* α [38]. Class I is found in *Cupriavidus pinatubonensis*; Class II is less widely distributed, being found only in *Burkholderia* spp.; and Class III is found in *Comamonas acidovorans* [38]. *TfdA*α was first identified in *Bradyrhizobium* sp. [40]. However, *tfdA*α-encoded protein has been described as α-ketoglutarate-dependent 2,4-D dioxygenase

2.The second group consists of slow-growing oligotrophic bacteria belonging to α-proteobacteria, phylogenetically closely related to *Bradyrhizobium* sp. [41], which were isolated from pristine environments. In this group, the phenoxy herbicide degradative gene was also identified and classified as *tfdA*α. Its gene sequence shows 50–60% similarity to the Group I degrader *Cupriavidus necator*

3.The third group consists of bacteria belonging to the α-proteobacteria harboring the *tfdA*α gene, with *Sphingomonas* being the key member [41]. The wide diversity displayed by *tfdA*-like genes can partly be attributed to the wide range of bacteria (α-, β-, and γ-proteobacteria) capable of degrading phenoxy acids in the environment. Due to the high degree of homology between strains, the *tfdA* genes have been selected as biomarkers of the capability of bacteria to metabolize 2,4-D and MCPA [37, 38], and they are frequently used in studies of

Much of the current literature on phenoxy herbicide metabolic pathways pays particular attention to the degradation pathway of 2,4-D. One of the most extensively studied 2,4-D degraders is *Cupriavidus necator* JMP134, known to harbor the 80-kb pJP4 plasmid. pJP4 carries all of the structural and regulatory genes needed to convert phenoxy herbicides to 2-chloromaleylacetic acid [56]. The *tfdA* fragment is responsible for the conversion of 2,4-D to 2,4-DCP [57]. Subsequently, 2,4-DCP is hydrolyzed to 3,5-dichlorocatechol by 2,4-DCP hydroxylase, which is encoded by *tfdB*. 3,5-Dichlorocatechol is further degraded via a pathway encoded by *tfdCDEF*. Far too little attention has been paid to the metabolism of MCPA. MCPA degradation takes place by the cleavage of an ether linkage, resulting in the formation of the major metabolite, 4-chloro-2-methylphenol (MCP), and acetic acid [47]. This process is preceded by the expression of the *tfdA* gene. Mierzejewska et al. [34] report that microorganisms demonstrating the presence of *tfdA*α and *tfdA* Class III genes in soil contaminated with a commercial product containing MCPA displayed

The bacteria carrying *cad* genes, which encode the non-heme iron oxygenase, also have the potential to degrade both herbicides. The *cadRABKC* gene cluster was first identified and characterized in strain *Bradyrhizobium* sp. HW3 which was isolated from pristine environment in Volcanoes National Park, Hawaii [39]. So far, however, there has been little research on the mode of action and exact function of *cad* genes. According to Kitagawa et al. [38], *cadA*, *cadB*, and *cadC* genes are responsible for multicomponent oxygenase production, whereas *cadR* is a transcriptional regulator gene, which regulates the transcription of *cadABKC* in the presence of 2,4-D or 4-chlorophenoxyacetic acid. The *cadA* gene products show structural and functional differences to the *tfdA* gene with regard to their substrate preferences. Both the *cadA* and *cadB* and the *tfdA* genes code for aromatic ring

*DOI: http://dx.doi.org/10.5772/intechopen.88256*

phenoxy acid biodegradation.

biodegradation potential.

JMP134.

*Biological Remediation of Phenoxy Herbicide-Contaminated Environments DOI: http://dx.doi.org/10.5772/intechopen.88256*

*Environmental Chemistry and Recent Pollution Control Approaches*

of such recalcitrant xenobiotics from the environment by exploiting the interactions between selected plants (able to grow under the presence of given xenobiotics such as phenoxy herbicides), root exudates (including plant secondary metabolites, PSMs), and microorganisms (**Figure 2**). The purpose of this section is to review the literature on established and potential biological methods of phenoxy herbicide

*The biological processes of phenoxy herbicide biodegradation mediated by soil, rhizospheric, and endophytic* 

Bioremediation is a method that uses microbiological processes to degrade or transform contaminants to less toxic or nontoxic forms. Biodegradation of organic contaminants occurs very slowly in bulk soil; therefore biostimulation and bioaugmentation methods are used to enhance the biologically driven removal of toxic compounds from environmental matrices. The effectiveness of biodegradation is dependent on several factors, among them the characteristics of the soil, the

An important way of phenoxy herbicide removal from soil is by the use of indigenous soil bacteria harboring desirable catabolic genes. The first step in the phenoxy herbicide biodegradation pathway is initiated by α-ketoglutaratedependent dioxygenase, an enzyme encoded by *tfdA* or *tfdA*-like genes [37] located

In recent decades, increasingly rapid advances in the application of molecular analysis in environmental studies have helped identify the bacterial communities involved in phenoxy herbicide biodegradation (**Table 4**). The bacteria able to metabolize phenoxy herbicides have been classified into three groups as follows: according to their physiology, employed degrading enzymes, and evolutionary

bioavailability of the contaminants, and their chemical properties.

removal from environmental matrices.

in the *tfdABCDEF* gene cluster [38].

origin [39–41] (**Table 4**).

**3.1 Bioremediation**

**Figure 2.**

*bacteria.*

**86**


Much of the current literature on phenoxy herbicide metabolic pathways pays particular attention to the degradation pathway of 2,4-D. One of the most extensively studied 2,4-D degraders is *Cupriavidus necator* JMP134, known to harbor the 80-kb pJP4 plasmid. pJP4 carries all of the structural and regulatory genes needed to convert phenoxy herbicides to 2-chloromaleylacetic acid [56]. The *tfdA* fragment is responsible for the conversion of 2,4-D to 2,4-DCP [57]. Subsequently, 2,4-DCP is hydrolyzed to 3,5-dichlorocatechol by 2,4-DCP hydroxylase, which is encoded by *tfdB*. 3,5-Dichlorocatechol is further degraded via a pathway encoded by *tfdCDEF*.

Far too little attention has been paid to the metabolism of MCPA. MCPA degradation takes place by the cleavage of an ether linkage, resulting in the formation of the major metabolite, 4-chloro-2-methylphenol (MCP), and acetic acid [47]. This process is preceded by the expression of the *tfdA* gene. Mierzejewska et al. [34] report that microorganisms demonstrating the presence of *tfdA*α and *tfdA* Class III genes in soil contaminated with a commercial product containing MCPA displayed biodegradation potential.

The bacteria carrying *cad* genes, which encode the non-heme iron oxygenase, also have the potential to degrade both herbicides. The *cadRABKC* gene cluster was first identified and characterized in strain *Bradyrhizobium* sp. HW3 which was isolated from pristine environment in Volcanoes National Park, Hawaii [39]. So far, however, there has been little research on the mode of action and exact function of *cad* genes. According to Kitagawa et al. [38], *cadA*, *cadB*, and *cadC* genes are responsible for multicomponent oxygenase production, whereas *cadR* is a transcriptional regulator gene, which regulates the transcription of *cadABKC* in the presence of 2,4-D or 4-chlorophenoxyacetic acid. The *cadA* gene products show structural and functional differences to the *tfdA* gene with regard to their substrate preferences. Both the *cadA* and *cadB* and the *tfdA* genes code for aromatic ring


#### **Table 4.**

*Bacteria degrading phenoxy herbicides isolated from pristine and contaminated environments.*

**89**

*Biological Remediation of Phenoxy Herbicide-Contaminated Environments*

*tfdA* genes, thereby displaying a dual system of degradative genes [47].

The microbial degradation metabolic pathway of phenoxy herbicides has been elaborated in recent years (**Figure 3**). The first step of this catabolic pathway is initiated by either the *tfdA* gene which encodes α-ketoglutarate-dependent dioxygenase or *cadAB* genes which encode subunits of non-heme iron oxygenase [47]. Although these enzymes use different modes of action, both catabolic proteins have been shown to perform the same initial step in phenoxy acid degradation, turning

Until recently, there has been little interest in the stereospecific Fe-(II) α-ketoglutarate-dependent dioxygenases which are encoded by *rdpA* and *sdpA* genes. These enzymes are described in literature as the ones which can also initiate the first step of the MCPA and 2,4-D degradation pathway. They were identified in *Delftia acidovorans*, *Rhodoferax* sp., and *Sphingobium* sp. Although the proteins encoded by the *rdpA* and *sdpA* genes possess the highly conserved amino acid sequence motif of *tfdA*-encoded proteins, they share only 37% identity with the

In addition to the soil bacteria, soil microfauna can also profoundly affect the biodegradation of organic contaminants. An important example of this relationship is the activity of earthworms, which move through the soil, causing better aeration and increasing soil moisture. Hence, insofar as their activity can influence the profile of the microorganism communities in the soil, they can indirectly enhance

A steadily developing strategy for the in situ treatment of contaminated soils is phytoremediation. It is a cost-effective and environmentally friendly strategy that uses plants to transform or mineralize xenobiotics to less toxic or environmentally neutral compounds [62]. Plants play a crucial role in the development of soil structure and stabilization of fundamental soil ecosystem functions such as water flow [63]. They produce also an array of catabolic enzymes, which operate to protect the host organisms and detoxify xenobiotic compounds [64]. Therefore, phytoremediation not only contributes to the detoxification of the environmental matrices but

the process of phenoxy herbicide aerobic bacterial degradation [61].

also has a positive influence on the functioning of the entire ecosystem.

The process of contaminant absorption by plants depends on several factors, including regional climate, soil type, and the nature of the pollutant [65]. The selection of an appropriate plant species and cultivar is critical for effective removal of a given contaminant from soil [66, 67]. This choice of phytoremediation candidate should

hydroxylation dioxygenases (RHDO), which are widely distributed in a number of microorganisms and might be transferred through horizontal gene transfer [38]; *CadA*- and *cadB*-encoded proteins are involved in the same initial step of 2,4-D degradation; however, the enzyme subunits have a different mode of action to the ketoglutarate-dependent dioxygenase encoded by *tfdA*. *CadA* and *cadB* were mostly identified in bacteria belonging to Groups I and II of phenoxy herbicide degraders. The products of *cadA* gene expression are able to initiate the degradation of both MCPA and 2,4-D. Furthermore, the abundance of *cadA* gene stimulates MCPA degradation [47]. The *cadA* gene is also essential for 2,4-D conversion in pure cultures of α-proteobacteria [38, 45, 58], and the *cadB* gene is also thought to play a sole role in the phenoxy acid degradation; however, the exact role of the *cad* genes remains not fully understood. The two genes share ~50% identity with *tfdA*, and it has been found that *cadA*, *cadB*, and *tfdA* are expressed simultaneously during MCPA degradation. Interestingly, some bacteria harbor all three *cadA*, *cadB*, and

*DOI: http://dx.doi.org/10.5772/intechopen.88256*

2,4-D into 2,4-DCP and MCPA into MCP.

*tfdA* genes of *C. necator* JMP134 [60, 61].

**3.2 Phytoremediation**

#### *Biological Remediation of Phenoxy Herbicide-Contaminated Environments DOI: http://dx.doi.org/10.5772/intechopen.88256*

*Environmental Chemistry and Recent Pollution Control Approaches*

α-Proteobacteria *Sphingomonas* 

*paucimobilis*

*Sphingomonas agrestis* 58–1

*Bradyrhizobium*

*Sphingomonas* sp.

*acidovorans* strain MCI

*Variovorax paradoxus*

*Cupriavidus campinensis* BJ71

*Achromobacter* sp. LZ35

*Halomonadaceae*

sp.

*pickettii*

*Pseudomonas maltophilia*

γ-Proteobacteria *Pseudomonas* 

β-Proteobacteria *Comamonas* 

sp.; *Sphingomonas* sp.

**Class Strain Origin Studied** 

Soil from Michigan

Soil from Fukuoka Prefecture (Japan)

Sediment from an aquifer in Fladerne Creek (Denmark)

Soil from the Dijon INRA experimental station (France)

in Buenos Aires (Argentina)

Soil in a disused pesticide factory in Suzhou (China)

Alkali Lake site in Oregon (USA) contaminated with 2,4-D production

Agricultural soil from Michigan (USA)

Wheat rhizosphere (laboratory experiment)

wastes

*Bacteria degrading phenoxy herbicides isolated from pristine and contaminated environments.*

2,4-D-enriched soils from wheat fields in Beijing exposed for 2,4-D for at least 10 years (China)

Herbicidecontaminated building rubble (Germany)

*Delftia* sp. Polluted river

Root nodules; pristine environments (Hawaii, central California, USA; southwestern Australia, southwestern Africa; central Chile; northern Saskatchewan, Canada; northwestern Russia); volcanic soil (National Park (Kipuka Keana Bihopa, Hawaii, USA)

(USA)

**compounds**

2,4-D, MCPA **Identified functional genes**

*cadA*, and *cadB*

*cadA*, *cadB* [45]

2,4-D — [44]

2,4-D *tfdA*α,

MCPA *cadA* and

2,4-D *tfdA*, *tfdB*,

2,4-D Class I *tfdA* gene

2,4-D and MCPA

2,4-D and MCPA

*cadB*

*tfdB* and *tfdC* genes

and *tfdR*

*tfdA* and *tfdB*

2,4-D *tfdA* [53]

2,4-D — [54]

2,4-D — [55]

2,4-D — [50]

**Source**

[38, 39, 41]

[46, 47]

[48]

[49]

[51]

[52]

**88**

**Table 4.**

hydroxylation dioxygenases (RHDO), which are widely distributed in a number of microorganisms and might be transferred through horizontal gene transfer [38]; *CadA*- and *cadB*-encoded proteins are involved in the same initial step of 2,4-D degradation; however, the enzyme subunits have a different mode of action to the ketoglutarate-dependent dioxygenase encoded by *tfdA*. *CadA* and *cadB* were mostly identified in bacteria belonging to Groups I and II of phenoxy herbicide degraders. The products of *cadA* gene expression are able to initiate the degradation of both MCPA and 2,4-D. Furthermore, the abundance of *cadA* gene stimulates MCPA degradation [47]. The *cadA* gene is also essential for 2,4-D conversion in pure cultures of α-proteobacteria [38, 45, 58], and the *cadB* gene is also thought to play a sole role in the phenoxy acid degradation; however, the exact role of the *cad* genes remains not fully understood. The two genes share ~50% identity with *tfdA*, and it has been found that *cadA*, *cadB*, and *tfdA* are expressed simultaneously during MCPA degradation. Interestingly, some bacteria harbor all three *cadA*, *cadB*, and *tfdA* genes, thereby displaying a dual system of degradative genes [47].

The microbial degradation metabolic pathway of phenoxy herbicides has been elaborated in recent years (**Figure 3**). The first step of this catabolic pathway is initiated by either the *tfdA* gene which encodes α-ketoglutarate-dependent dioxygenase or *cadAB* genes which encode subunits of non-heme iron oxygenase [47]. Although these enzymes use different modes of action, both catabolic proteins have been shown to perform the same initial step in phenoxy acid degradation, turning 2,4-D into 2,4-DCP and MCPA into MCP.

Until recently, there has been little interest in the stereospecific Fe-(II) α-ketoglutarate-dependent dioxygenases which are encoded by *rdpA* and *sdpA* genes. These enzymes are described in literature as the ones which can also initiate the first step of the MCPA and 2,4-D degradation pathway. They were identified in *Delftia acidovorans*, *Rhodoferax* sp., and *Sphingobium* sp. Although the proteins encoded by the *rdpA* and *sdpA* genes possess the highly conserved amino acid sequence motif of *tfdA*-encoded proteins, they share only 37% identity with the *tfdA* genes of *C. necator* JMP134 [60, 61].

In addition to the soil bacteria, soil microfauna can also profoundly affect the biodegradation of organic contaminants. An important example of this relationship is the activity of earthworms, which move through the soil, causing better aeration and increasing soil moisture. Hence, insofar as their activity can influence the profile of the microorganism communities in the soil, they can indirectly enhance the process of phenoxy herbicide aerobic bacterial degradation [61].

## **3.2 Phytoremediation**

A steadily developing strategy for the in situ treatment of contaminated soils is phytoremediation. It is a cost-effective and environmentally friendly strategy that uses plants to transform or mineralize xenobiotics to less toxic or environmentally neutral compounds [62]. Plants play a crucial role in the development of soil structure and stabilization of fundamental soil ecosystem functions such as water flow [63]. They produce also an array of catabolic enzymes, which operate to protect the host organisms and detoxify xenobiotic compounds [64]. Therefore, phytoremediation not only contributes to the detoxification of the environmental matrices but also has a positive influence on the functioning of the entire ecosystem.

The process of contaminant absorption by plants depends on several factors, including regional climate, soil type, and the nature of the pollutant [65]. The selection of an appropriate plant species and cultivar is critical for effective removal of a given contaminant from soil [66, 67]. This choice of phytoremediation candidate should

#### **Figure 3.**

*Pathways of microbial degradation of 2,4-D and MCPA proposed by Pieper et al. [59]; in the picture there are indicated functional genes which encode catabolic enzymes as follows: tfdA, α-ketoglutarate-dependent dioxygenase; cadAB, subunits of non-heme iron oxygenase; tfdB, chlorophenol hydroxylase; tfdC, catechol 1,2-dioxygenase; tfdD, dichloromuconate cycloisomerase; tfdE, carboxymethylene butenolidase; tfdF, maleylacetate reductase.*

**91**

*Biological Remediation of Phenoxy Herbicide-Contaminated Environments*

particularly take into account plant growth rate, high biomass production, capacity for pollutant accumulation, and tolerance to higher xenobiotic concentrations [67].

As mentioned above, plants play a key role in soil ecosystems by stabilizing the soil structure and by serving as primary sources of organic matter and energy which stimulate soil microbial activity [63]. Despite this, they are not the only contributors in the efficient phytoremediation of organic contaminants. Due to existing interactions between plant roots, root exudates, soil, and microorganisms, it has been proposed that the most effective method for the remediation of contaminated

Rhizoremediation is a naturally occurring process within the plant root zone (rhizosphere), where the growth of microorganisms and their degradative activity are stimulated by root exudates enriched by plant secondary metabolites (PSMs). Plant-derived compounds can [1] serve as primary substrates in cometabolism and provide energy for microbial growth [2], act as inducers of degradative enzymes due to their structural similarities to xenobiotics, and [3] enhance the degree of contamination removal by increasing pollutant bioavailability in soil [71].

The effectiveness of rhizospheral biodegradation depends also on the potential of the microorganisms inhabiting the rhizosphere to adapt to pollutant concentrations [72]. For effective degradation of contaminants to take place, a wide range of plants and bacterial traits is needed, involving the orchestrated interaction of a multitude of genes and enzymes. Rhizoremediation can therefore be optimized by selecting suitable plant-microbe sets, which can be achieved by combining plant and plant growth-promoting rhizobacteria (PGPR) [73] and/or microbes capable of contaminant degradation [74]. PGPR can improve phytoremediation efficiency by enhancing plant tolerance to various environmental stresses, promoting root growth and improving plant growth and health. In turn, some rhizospheral microorganisms can directly use their own degradative capabilities to metabolize organic pollutants [74, 75]. A study of rhizosphere-enhanced biodegradation of 2,4-D by Boyle et al. [76] found a significant difference in the mineralization of 2,4-D between monocot rhizosphere soils, dicot rhizosphere soils, and nonrhizosphere soils, with greater microbial activity being observed in monocot

soil may be microbe-assisted phytoremediation (rhizoremediation).

In terms of phenoxy herbicide removal, there has been little investigation of the plant-mediated removal of 2,4-D and/or MCPA. For example, Ramborger et al. [68] evaluated the phytoremediation potential of *Plectranthus neochilus* (tea) exposed to the commercial pesticide containing 2,4-D (Aminol) in soil and water. The removal rate for 2,4-D reached 49% during 60 days, and the herbicide was not detected in plant leaves. Despite the fact that the phytoremediation potential of *P. neochilus* in soil was not sufficient, the plant exhibited satisfactory resistance to herbicide application. Moreover, the presence of phenolic compounds (e.g., ferulic and coumaric acid) in tea tissues indicated the ability of these plants to provide defense mechanisms against 2,4-D. The mechanism of the herbicide in plant begins by affecting the plasma membrane properties, subsequently leading to poor performance of mitochondria and peroxisomes [69]. In consequence, it stimulates the overexpression of abscisic acid (ABA) and ethylene biosynthesis genes, leading to significant changes of cellular redox potential by the production of reactive oxygen species (ROS) [70]. The occurrence of ROS leads to the production of phenolic compounds, i.e., ferulic acid and coumaric acid, which are responsible for the antioxidant selfdefense mechanism of the plant against the herbicide. These phenolic compounds were found in higher concentrations only in plants that were exposed to 2,4-D and

*DOI: http://dx.doi.org/10.5772/intechopen.88256*

not in the controls.

**3.3 Rhizoremediation**

#### *Biological Remediation of Phenoxy Herbicide-Contaminated Environments DOI: http://dx.doi.org/10.5772/intechopen.88256*

particularly take into account plant growth rate, high biomass production, capacity for pollutant accumulation, and tolerance to higher xenobiotic concentrations [67].

In terms of phenoxy herbicide removal, there has been little investigation of the plant-mediated removal of 2,4-D and/or MCPA. For example, Ramborger et al. [68] evaluated the phytoremediation potential of *Plectranthus neochilus* (tea) exposed to the commercial pesticide containing 2,4-D (Aminol) in soil and water. The removal rate for 2,4-D reached 49% during 60 days, and the herbicide was not detected in plant leaves. Despite the fact that the phytoremediation potential of *P. neochilus* in soil was not sufficient, the plant exhibited satisfactory resistance to herbicide application. Moreover, the presence of phenolic compounds (e.g., ferulic and coumaric acid) in tea tissues indicated the ability of these plants to provide defense mechanisms against 2,4-D. The mechanism of the herbicide in plant begins by affecting the plasma membrane properties, subsequently leading to poor performance of mitochondria and peroxisomes [69]. In consequence, it stimulates the overexpression of abscisic acid (ABA) and ethylene biosynthesis genes, leading to significant changes of cellular redox potential by the production of reactive oxygen species (ROS) [70]. The occurrence of ROS leads to the production of phenolic compounds, i.e., ferulic acid and coumaric acid, which are responsible for the antioxidant selfdefense mechanism of the plant against the herbicide. These phenolic compounds were found in higher concentrations only in plants that were exposed to 2,4-D and not in the controls.

### **3.3 Rhizoremediation**

*Environmental Chemistry and Recent Pollution Control Approaches*

**90**

**Figure 3.**

*maleylacetate reductase.*

*Pathways of microbial degradation of 2,4-D and MCPA proposed by Pieper et al. [59]; in the picture there are indicated functional genes which encode catabolic enzymes as follows: tfdA, α-ketoglutarate-dependent dioxygenase; cadAB, subunits of non-heme iron oxygenase; tfdB, chlorophenol hydroxylase; tfdC, catechol 1,2-dioxygenase; tfdD, dichloromuconate cycloisomerase; tfdE, carboxymethylene butenolidase; tfdF,* 

As mentioned above, plants play a key role in soil ecosystems by stabilizing the soil structure and by serving as primary sources of organic matter and energy which stimulate soil microbial activity [63]. Despite this, they are not the only contributors in the efficient phytoremediation of organic contaminants. Due to existing interactions between plant roots, root exudates, soil, and microorganisms, it has been proposed that the most effective method for the remediation of contaminated soil may be microbe-assisted phytoremediation (rhizoremediation).

Rhizoremediation is a naturally occurring process within the plant root zone (rhizosphere), where the growth of microorganisms and their degradative activity are stimulated by root exudates enriched by plant secondary metabolites (PSMs). Plant-derived compounds can [1] serve as primary substrates in cometabolism and provide energy for microbial growth [2], act as inducers of degradative enzymes due to their structural similarities to xenobiotics, and [3] enhance the degree of contamination removal by increasing pollutant bioavailability in soil [71].

The effectiveness of rhizospheral biodegradation depends also on the potential of the microorganisms inhabiting the rhizosphere to adapt to pollutant concentrations [72]. For effective degradation of contaminants to take place, a wide range of plants and bacterial traits is needed, involving the orchestrated interaction of a multitude of genes and enzymes. Rhizoremediation can therefore be optimized by selecting suitable plant-microbe sets, which can be achieved by combining plant and plant growth-promoting rhizobacteria (PGPR) [73] and/or microbes capable of contaminant degradation [74]. PGPR can improve phytoremediation efficiency by enhancing plant tolerance to various environmental stresses, promoting root growth and improving plant growth and health. In turn, some rhizospheral microorganisms can directly use their own degradative capabilities to metabolize organic pollutants [74, 75]. A study of rhizosphere-enhanced biodegradation of 2,4-D by Boyle et al. [76] found a significant difference in the mineralization of 2,4-D between monocot rhizosphere soils, dicot rhizosphere soils, and nonrhizosphere soils, with greater microbial activity being observed in monocot

rhizosphere soil than in dicot rhizosphere soil or bulk soil. Therefore, both the soil and plant species determine the mineralization of tested contaminant. According to Shaw and Burns [77], the amendment of soil with 2,4-D increased the number of rhizospheric bacteria degrading 2,4-D in *Trifolium pratense* (red clover). Germaine et al. [78] also note the abundance of 2,4-D degraders in the stem and leaves of pea plant and that, under exposure to phenoxy herbicide, pea plants developed a stubby root system.

Furthermore, it has been hypothesized that PSMs may have a profound impact on the biodegradation of xenobiotics by providing the energy for microorganisms to carry out cometabolism; in this case, the xenobiotic is degraded as a secondary substrate [45, 71–73]. PSMs can be used as a primary source of carbon for bacterial communities to support their growth and stimulate the expression of desirable genes involved in the catabolic pathway of given xenobiotic. This is evident in the case of biphenyl, naringin, coumarin, myricetin, and l-carvone, which stimulate the activity of polychlorinated biphenyl (PCB)-degrading bacteria such as *A. eutrophus*, *Corynebacterium* sp., *P. putida* [79], and *Arthrobacter* sp. strain B1B [80]. Another example of PSM-stimulated PCB biodegradation was identified in mulberry (*Morus* sp.). In this case, the PSMs morusin, morusinol, and kuwanon C have been found to support the growth of the PCB-degrading bacterium *Burkholderia* sp. LB400 [81]. Likewise, the PSM (cumene) stimulates the activity of TCE-degrading *R. gordonia* bacteria [85]. According to Yi et al. [82], salicylic and linoleic acids, excreted by *Raphanus sativus*, enhanced the bioavailability of polycyclic aromatic hydrocarbons (PAHs) and increased the effectivity of their removal form soil. According to Ely and Smets [83], PAH biodegradation is stimulated by the presence of phenolic compounds, flavonoids, and gibberellic acid. Compounds such as acetophenone, phenethyl alcohol, p-hydroxybenzoic acid, and trans-cinnamic acid enhance the biotransformation of cis-1,2-dichloroethylene [84].

In addition, it has been hypothesized that PSMs may also induce the detoxification mechanisms taking place in bacterial cells [85, 86]. The expression of functional genes in bacteria is essential for the successful bioremediation of xenobiotics and can be stimulated by PSMs in different ways. However, very little information is given in the literature regarding the influence of PSMs on the induction of genes involved in catabolic pathways. Siciliano et al. [87] report greater induction of catabolic genes (*ndoB*, *alkB*, *xylE*) involved in the degradation of naphthalene in the rhizosphere soil of *Festuca arundinacea* (tall fescue) than in unplanted soil. Salicylate has been reported to have an upregulating effect on the expression of *bphA*, which encodes biphenyl dioxygenase in the PCB degrader *Pseudomonas* sp. Cam-1 [88]. The presence of salicylic acid was found to enhance the expression of the *bphA* gene in *R. eutropha* H850 and *P. fluorescens* P2W [89].

In addition, it has been hypothesized that the structural similarity between selected xenobiotics and PSMs may have a profound impact on the biodegradation of given, structurally related xenobiotic [71, 85]. For example, Urbaniak et al. [35] demonstrated the effect of a PSM, syringic acid, on the enhanced removal of structurally similar herbicide, MCPA, by indigenous soil bacteria, with greater MCPA depletion being achieved in samples enriched with PSM. The molecular analysis revealed ubiquitous enrichment of the samples with *Rhodoferax* spp., *Achromobacter* spp., *Burkholderia* spp., and *Cupriavidus* spp., which are commonly known as MCPA degraders. Also, a study by McLoughlin et al. [89] found the PSMs limonene and α-pinene to enhance 2,4-DCP degradation, but only following pre-exposure to both 2,4-DCP and monoterpene, with total 2,4-DCP mineralization extents of up to 71%.

Taking into account the abovementioned aspects, rhizoremediation can serve as a potential tool for phenoxy herbicide removal from soil ecosystems. However, to date, most studies have focused solely on the phyto- or biodegradation properties

**93**

**Table 5.**

*Biological Remediation of Phenoxy Herbicide-Contaminated Environments*

impact of rhizoremediation on phenoxy herbicide removal from soil.

of plants or bacteria [71]. Consequently only limited data is available in terms of the

Endophytic bacteria that reside inside plant tissues are also known to play a crucial role in the remediation of organic compounds. Plant-associated bacteria can enhance plant growth and degrade organic contaminants such as trichloroethylene and hydrocarbons [90]. The activity of endophytic bacteria can mitigate and improve plant conditions in stressful environments (such as contaminated soils). Field studies by Eevers et al. [91] showed that zucchini (*Cucurbita pepo*) plants inoculated with a consortium of three plant growth-promoting endophytic strains demonstrated an increased concentration of dichloro-bis(p-chlorophenyl)ethylene (DDE) in the aerial parts. The amount of DDE accumulated in *C. pepo* per growing season was significantly higher for inoculated plants. Thus such an approach might

It has also been found that application of 2,4-D (1.42, 2.84, and 5.68 mg a.i./g soil) had a negative effect on the physio-morphological parameters of aerobic rice

> **Duration of an experiment (days)**

2,4-D 1.8 kg/ha 10 45–48 Activity of

2,4-D 11.42 kg/ha 20 49 Use of

**Removal of phenoxyacetic acid (%)**

118 60 Activity of

66 ~60 Activity of

53 93–100 The

**Comments Source**

[93]

[94]

[76]

[68]

bulk soil microbial population from various soil samples

bulk soil microbial population from clay and loamy soil samples

*P. neochilus* for phytoremediation

rhizospheric soil bacteria derived from monocots

inoculation of *P. sativum* by endophytic bacteria *P. putida* VM1450

**concentration of compound used in an experiment**

0.09 mmol/kg of soil

1.22 10<sup>−</sup><sup>3</sup> μm 1.19 10<sup>−</sup><sup>3</sup> μmol

of soil

*Biological remediation methods and % average removal of phenoxy herbicides from soil matrices.*

2,4-D 47–360 mg/kg

*DOI: http://dx.doi.org/10.5772/intechopen.88256*

**3.4 Endophyte-enhanced phytoremediation**

be promising for phytoremediation applications.

**Method Compound Initial** 

MCPA

2,4-D 2,4-DCP

Bioremediation 2,4-D and

Phytoremediation and rhizoremediation

Endophyteenhanced phytoremediation

of plants or bacteria [71]. Consequently only limited data is available in terms of the impact of rhizoremediation on phenoxy herbicide removal from soil.

## **3.4 Endophyte-enhanced phytoremediation**

*Environmental Chemistry and Recent Pollution Control Approaches*

developed a stubby root system.

rhizosphere soil than in dicot rhizosphere soil or bulk soil. Therefore, both the soil and plant species determine the mineralization of tested contaminant. According to Shaw and Burns [77], the amendment of soil with 2,4-D increased the number of rhizospheric bacteria degrading 2,4-D in *Trifolium pratense* (red clover). Germaine et al. [78] also note the abundance of 2,4-D degraders in the stem and leaves of pea plant and that, under exposure to phenoxy herbicide, pea plants

Furthermore, it has been hypothesized that PSMs may have a profound impact on the biodegradation of xenobiotics by providing the energy for microorganisms to carry out cometabolism; in this case, the xenobiotic is degraded as a secondary substrate [45, 71–73]. PSMs can be used as a primary source of carbon for bacterial communities to support their growth and stimulate the expression of desirable genes involved in the catabolic pathway of given xenobiotic. This is evident in the case of biphenyl, naringin, coumarin, myricetin, and l-carvone, which stimulate the activity of polychlorinated biphenyl (PCB)-degrading bacteria such as *A. eutrophus*, *Corynebacterium* sp., *P. putida* [79], and *Arthrobacter* sp. strain B1B [80]. Another example of PSM-stimulated PCB biodegradation was identified in mulberry (*Morus* sp.). In this case, the PSMs morusin, morusinol, and kuwanon C have been found to support the growth of the PCB-degrading bacterium *Burkholderia* sp. LB400 [81]. Likewise, the PSM (cumene) stimulates the activity of TCE-degrading *R. gordonia* bacteria [85]. According to Yi et al. [82], salicylic and linoleic acids, excreted by *Raphanus sativus*, enhanced the bioavailability of polycyclic aromatic hydrocarbons (PAHs) and increased the effectivity of their removal form soil. According to Ely and Smets [83], PAH biodegradation is stimulated by the presence of phenolic compounds, flavonoids, and gibberellic acid. Compounds such as acetophenone, phenethyl alcohol, p-hydroxybenzoic acid, and trans-cinnamic

acid enhance the biotransformation of cis-1,2-dichloroethylene [84].

the *bphA* gene in *R. eutropha* H850 and *P. fluorescens* P2W [89].

In addition, it has been hypothesized that PSMs may also induce the detoxification mechanisms taking place in bacterial cells [85, 86]. The expression of functional genes in bacteria is essential for the successful bioremediation of xenobiotics and can be stimulated by PSMs in different ways. However, very little information is given in the literature regarding the influence of PSMs on the induction of genes involved in catabolic pathways. Siciliano et al. [87] report greater induction of catabolic genes (*ndoB*, *alkB*, *xylE*) involved in the degradation of naphthalene in the rhizosphere soil of *Festuca arundinacea* (tall fescue) than in unplanted soil. Salicylate has been reported to have an upregulating effect on the expression of *bphA*, which encodes biphenyl dioxygenase in the PCB degrader *Pseudomonas* sp. Cam-1 [88]. The presence of salicylic acid was found to enhance the expression of

In addition, it has been hypothesized that the structural similarity between selected xenobiotics and PSMs may have a profound impact on the biodegradation of given, structurally related xenobiotic [71, 85]. For example, Urbaniak et al. [35] demonstrated the effect of a PSM, syringic acid, on the enhanced removal of structurally similar herbicide, MCPA, by indigenous soil bacteria, with greater MCPA depletion being achieved in samples enriched with PSM. The molecular analysis revealed ubiquitous enrichment of the samples with *Rhodoferax* spp., *Achromobacter* spp., *Burkholderia* spp., and *Cupriavidus* spp., which are commonly known as MCPA degraders. Also, a study by McLoughlin et al. [89] found the PSMs limonene and α-pinene to enhance 2,4-DCP degradation, but only following pre-exposure to both 2,4-DCP and monoterpene, with total 2,4-DCP mineralization extents of up to 71%. Taking into account the abovementioned aspects, rhizoremediation can serve as a potential tool for phenoxy herbicide removal from soil ecosystems. However, to date, most studies have focused solely on the phyto- or biodegradation properties

**92**

Endophytic bacteria that reside inside plant tissues are also known to play a crucial role in the remediation of organic compounds. Plant-associated bacteria can enhance plant growth and degrade organic contaminants such as trichloroethylene and hydrocarbons [90]. The activity of endophytic bacteria can mitigate and improve plant conditions in stressful environments (such as contaminated soils). Field studies by Eevers et al. [91] showed that zucchini (*Cucurbita pepo*) plants inoculated with a consortium of three plant growth-promoting endophytic strains demonstrated an increased concentration of dichloro-bis(p-chlorophenyl)ethylene (DDE) in the aerial parts. The amount of DDE accumulated in *C. pepo* per growing season was significantly higher for inoculated plants. Thus such an approach might be promising for phytoremediation applications.

It has also been found that application of 2,4-D (1.42, 2.84, and 5.68 mg a.i./g soil) had a negative effect on the physio-morphological parameters of aerobic rice


## **Table 5.**

*Biological remediation methods and % average removal of phenoxy herbicides from soil matrices.*

(*Oryza sativa*) and reduced the number of plant endophytes [92]; however, inoculation of seeds with the endophytic bacteria strain *Stenotrophomonas maltophilia* improved plant characteristics under herbicide-stressed soils. *S. maltophilia* has previously been described as a plant growth-promoting endophytic strain with the ability to produce auxins and siderophores [92].

Bacterial endophyte-enhanced phytoremediation was also studied by Germaine et al. [78] on the example of *P. sativum*: plants were inoculated with genetically tagged endophytic bacteria, which naturally possess the ability to biodegrade 2,4-D. The inoculated plants not only displayed more efficient herbicide removal but also demonstrated a lack of 2,4-D accumulation in their aerial parts. Additionally the endophytic strain protected the pea plant from the toxic effects of 2,4-D, resulting in a greater increase of plant biomass and thus greater 2,4-D transportation to the aboveground parts of the plant from the soil.

**Table 5** compares the presented biological methods of remediation of soils contaminated with phenoxy herbicides. It illustrates the differences of the removal of phenoxy herbicides from soil. It is apparent from this table that the most efficient method of contaminant removal is endophyte-assisted phytoremediation; however, more research on this topic needs to be undertaken before the association between role of symbiotic microorganisms and plants in removal of contaminants from environmental matrices is more clearly understood.

## **4. Conclusions**

Uncontrolled use of phenoxy herbicides (2,4-D and MCPA) in the agricultural and gardening sector can result in their dispersal in soil and water ecosystems, which can significantly disturb the sustainability of the environment and increase its ecotoxicity level. Although their persistence in soil is limited due to their chemical characteristics, they can be transported and accumulated in water ecosystems through runoff and leaching. According to recent reports, phenoxy herbicides are especially toxic for plants, freshwater crustaceans, and amphibians; hence there is a growing need to limit the release of phenoxy acids in natural environments.

Taking into account the abovementioned aspects, the integration of bio-, phyto-, and rhizoremediation can serve as a potential tool for phenoxy herbicide removal from soil ecosystems. The ability of bacteria to metabolize phenoxy herbicides has been extensively studied over the last decades. However, to date, only limited data is available in terms of the impact of phyto- and rhizoremediation on phenoxy herbicide removal from soil. What is not yet clear is the impact of PSMs on the degradation of phenoxy herbicides. The similarity of the chemical structure of chosen PSMs and xenobiotics can be reflected in the xenobiotic degradation rates, e.g., the presence and induction of degradative genes and production of degradative enzymes, and the composition of microbial populations. To date, little evidence has been found associating the removal of phenoxy herbicides using both plants and microorganisms. However, the abovementioned research serves as a base for future studies on their application for the improvement of soil quality.

Considering the above, the chapter describes an interdisciplinary approach to tackling the problem of environmental phenoxy acid herbicide contamination through integrating available literature data on the physicochemical properties of 2,4-D and MCPA, as well as their levels in the environment and toxicity to the organisms from different trophic levels. It also outlines possible methods for their removal using nature-based techniques such as bio-, phyto-, and rhizoremediation.

**95**

**Author details**

Czech Republic

Magdalena Urbaniak1

\* and Elżbieta Mierzejewska<sup>2</sup>

Technology, University of Chemistry and Technology in Prague, Prague,

2 Department of Applied Ecology, Faculty of Biology and Environmental

\*Address all correspondence to: magdalena.urbaniak@vscht.cz

Protection, University of Lodz, Lodz, Poland

provided the original work is properly cited.

1 Department of Biochemistry and Microbiology, Faculty of Food and Biochemical

© 2019 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/ by/3.0), which permits unrestricted use, distribution, and reproduction in any medium,

*Biological Remediation of Phenoxy Herbicide-Contaminated Environments*

This work was supported by the European Structural and Investment Funds, OP RDE-funded project "CHEMFELLS4UCTP" (No. CZ.02.2.69/0.0/0.0/17\_050/

*DOI: http://dx.doi.org/10.5772/intechopen.88256*

There is no conflict of interest.

**Acknowledgements**

**Conflict of interest**

0008485).

*Biological Remediation of Phenoxy Herbicide-Contaminated Environments DOI: http://dx.doi.org/10.5772/intechopen.88256*

## **Acknowledgements**

*Environmental Chemistry and Recent Pollution Control Approaches*

ability to produce auxins and siderophores [92].

aboveground parts of the plant from the soil.

environmental matrices is more clearly understood.

**4. Conclusions**

environments.

(*Oryza sativa*) and reduced the number of plant endophytes [92]; however, inoculation of seeds with the endophytic bacteria strain *Stenotrophomonas maltophilia* improved plant characteristics under herbicide-stressed soils. *S. maltophilia* has previously been described as a plant growth-promoting endophytic strain with the

Bacterial endophyte-enhanced phytoremediation was also studied by Germaine

et al. [78] on the example of *P. sativum*: plants were inoculated with genetically tagged endophytic bacteria, which naturally possess the ability to biodegrade 2,4-D. The inoculated plants not only displayed more efficient herbicide removal but also demonstrated a lack of 2,4-D accumulation in their aerial parts. Additionally the endophytic strain protected the pea plant from the toxic effects of 2,4-D, resulting in a greater increase of plant biomass and thus greater 2,4-D transportation to the

**Table 5** compares the presented biological methods of remediation of soils contaminated with phenoxy herbicides. It illustrates the differences of the removal of phenoxy herbicides from soil. It is apparent from this table that the most efficient method of contaminant removal is endophyte-assisted phytoremediation; however, more research on this topic needs to be undertaken before the association between role of symbiotic microorganisms and plants in removal of contaminants from

Uncontrolled use of phenoxy herbicides (2,4-D and MCPA) in the agricultural and gardening sector can result in their dispersal in soil and water ecosystems, which can significantly disturb the sustainability of the environment and increase its ecotoxicity level. Although their persistence in soil is limited due to their chemical characteristics, they can be transported and accumulated in water ecosystems through runoff and leaching. According to recent reports, phenoxy herbicides are especially toxic for plants, freshwater crustaceans, and amphibians; hence there is a growing need to limit the release of phenoxy acids in natural

Taking into account the abovementioned aspects, the integration of bio-, phyto-, and rhizoremediation can serve as a potential tool for phenoxy herbicide removal from soil ecosystems. The ability of bacteria to metabolize phenoxy

studies on their application for the improvement of soil quality.

herbicides has been extensively studied over the last decades. However, to date, only limited data is available in terms of the impact of phyto- and rhizoremediation on phenoxy herbicide removal from soil. What is not yet clear is the impact of PSMs on the degradation of phenoxy herbicides. The similarity of the chemical structure of chosen PSMs and xenobiotics can be reflected in the xenobiotic degradation rates, e.g., the presence and induction of degradative genes and production of degradative enzymes, and the composition of microbial populations. To date, little evidence has been found associating the removal of phenoxy herbicides using both plants and microorganisms. However, the abovementioned research serves as a base for future

Considering the above, the chapter describes an interdisciplinary approach to tackling the problem of environmental phenoxy acid herbicide contamination through integrating available literature data on the physicochemical properties of 2,4-D and MCPA, as well as their levels in the environment and toxicity to the organisms from different trophic levels. It also outlines possible methods for their removal using nature-based techniques such as bio-, phyto-, and

**94**

rhizoremediation.

This work was supported by the European Structural and Investment Funds, OP RDE-funded project "CHEMFELLS4UCTP" (No. CZ.02.2.69/0.0/0.0/17\_050/ 0008485).

## **Conflict of interest**

There is no conflict of interest.

## **Author details**

Magdalena Urbaniak1 \* and Elżbieta Mierzejewska<sup>2</sup>

1 Department of Biochemistry and Microbiology, Faculty of Food and Biochemical Technology, University of Chemistry and Technology in Prague, Prague, Czech Republic

2 Department of Applied Ecology, Faculty of Biology and Environmental Protection, University of Lodz, Lodz, Poland

\*Address all correspondence to: magdalena.urbaniak@vscht.cz

© 2019 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/ by/3.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

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[28] Estevam EC, Nakano E, Kawano T, de Bragança Pereira CA, Amancio FF, de Albuquerque Melo AMM. Dominant lethal effects of 2,4-D in *Biomphalaria glabrata*. Mutation Research, Genetic Toxicology and Environmental Mutagenesis. 2006;**611**(1-2):83-88

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*DOI: http://dx.doi.org/10.5772/intechopen.88256*

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[18] Chiron S, Comoretto L, Rinaldi E, Maurino V, Minero C, Vione D. Pesticide by-products in the Rhône delta (Southern France). The case of 4-chloro-2-methylphenol and of its nitroderivative. Chemosphere.

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Wastes. 2015;**50**(7):449-455

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[23] Pazmiño DM, Rodríguez-Serrano M, Romero-Puertas MC, Archilla-Ruiz A, del Río LA, Sandalio LM. Differential response of young and adult leaves to herbicide 2,4-dichlorophenoxyacetic acid in pea plants: Role of reactive

2018;**634**:394-406

2009;**74**(4):599-604

2008;**248**:61-66

2013;**263**:239-247

*Biological Remediation of Phenoxy Herbicide-Contaminated Environments DOI: http://dx.doi.org/10.5772/intechopen.88256*

[17] Palma P, Matos C, Alvarenga P, Köck-Schulmeyer M, Simões I, Barceló D, et al. Ecological and ecotoxicological responses in the assessment of the ecological status of freshwater systems: A case-study of the temporary stream Brejo of Cagarrão (South of Portugal). Science of the Total Environment. 2018;**634**:394-406

[18] Chiron S, Comoretto L, Rinaldi E, Maurino V, Minero C, Vione D. Pesticide by-products in the Rhône delta (Southern France). The case of 4-chloro-2-methylphenol and of its nitroderivative. Chemosphere. 2009;**74**(4):599-604

[19] Kucharski M, Urbanowicz J. Badanie pozostałości linuronu i MCPA w glebie i roślinach ziemniaka. Biuletyn Instytutu Hodowli i Aklimatyzacji Roślin. 2008;**248**:61-66

[20] López-Roldán R, Jubany I, Martí V, González S, Cortina JL. Ecological screening indicators of stress and risk for the Llobregat river water. Journal of Hazardous Materials. 2013;**263**:239-247

[21] Salvo LM, Malucelli MIC, da Silva JRMC, Alberton GC, Silva De Assis HC. Toxicity assessment of 2,4-D and MCPA herbicides in primary culture of fish hepatic cells. Journal of Environmental Science and Health, Part B. Pesticides, Food Contaminants, and Agricultural Wastes. 2015;**50**(7):449-455

[22] Wei YD, Zheng HG, Hall JC. Role of auxinic herbicide-induced ethylene on hypocotyl elongation and root/hypocotyl radial expansion. Pest Management Science. 2000;**56**(5):377-387

[23] Pazmiño DM, Rodríguez-Serrano M, Romero-Puertas MC, Archilla-Ruiz A, del Río LA, Sandalio LM. Differential response of young and adult leaves to herbicide 2,4-dichlorophenoxyacetic acid in pea plants: Role of reactive

oxygen species. Plant, Cell and Environment. 2011;**34**(11):1874-1889

[24] Martínez-Ruiz EB, Martínez-Jerónimo F. Exposure to the herbicide 2,4-D produces different toxic effects in two different phytoplankters: A green microalga (*Ankistrodesmus falcatus*) and a toxigenic cyanobacterium (*Microcystis aeruginosa*). Science of the Total Environment. 2018;**619**(620):1566-1578

[25] Li K, Wu JQ, Jiang LL, Shen LZ, Li JY, He ZH, et al. Developmental toxicity of 2,4-dichlorophenoxyacetic acid in zebrafish embryos. Chemosphere. 2017;**171**:40-48

[26] Sarikaya R, Yilmaz M. Investigation of acute toxicity and the effect of 2,4-D (2,4-dichlorophenoxyacetic acid) herbicide on the behavior of the common carp (*Cyprinus carpio* L., 1758; Pisces, Cyprinidae). Chemosphere. 2003;**52**(1):195-201

[27] Aronzon CM, Sandoval MT, Herkovits J, Pérez-Coll CS. Stage-dependent toxicity of 2,4-dichlorophenoxyacetic on the embryonic development of a south American toad, *Rhinella arenarum*. Environmental Toxicology. 2011;**26**(4):373-381

[28] Estevam EC, Nakano E, Kawano T, de Bragança Pereira CA, Amancio FF, de Albuquerque Melo AMM. Dominant lethal effects of 2,4-D in *Biomphalaria glabrata*. Mutation Research, Genetic Toxicology and Environmental Mutagenesis. 2006;**611**(1-2):83-88

[29] Bukowska B. Effects of 2,4-D and its metabolite 2,4-dichlorophenol on antioxidant enzymes and level of glutathione in human erythrocytes. Comparative Biochemistry and Physiology, Part C: Toxicology & Pharmacology. 2003;**135**(4):435-441

[30] Perkins EJ, Stiff CM, Lurquin PF. Use of *Alcaligenes eutrophus* as a

**96**

*Environmental Chemistry and Recent Pollution Control Approaches*

Environmental Science: Processes &

[11] Gupta M, Garg NK, Joshi H, Sharma MP. Persistence and mobility of 2,4-D in unsaturated soil zone under winter wheat crop in sub-tropical region of India. Agriculture, Ecosystems and Environment. 2012;**146**(1):60-72

[12] Ignatowicz K, Struk-Sokołowska J. Sezonowe wahania zanieczyszczeń agrotechnicznych w rzece Narwi ze szczególnym uwzględnieniem herbicydów fenoksyoctowych. Środkowo-Pomorskie Tow Nauk Ochr

[13] Birch GF, Drage DS, Thompson K, Eaglesham G, Mueller JF. Emerging contaminants (pharmaceuticals, personal care products, a food additive and pesticides) in waters of Sydney estuary, Australia. Marine Pollution

Środowiska. 2004;**4**:189-205

Bulletin. 2015;**97**(1-2):56-66

[14] Ismail BS, Prayitno S, Tayeb MA. Contamination of rice field water with sulfonylurea and phenoxy herbicides in the Muda Irrigation Scheme, Kedah, Malaysia. Environmental Monitoring and Assessment. 2015;**187**(7):406

[15] Kucharski M, Domaradzki K. Changes in soil contamination by selected herbicides used in protection of cereals. Polish Journal of Soil Science.

[16] Metcalfe CD, Helm P, Paterson G, Kaltenecker G, Murray C, Nowierski M, et al. Pesticides related to land use in watersheds of the Great Lakes basin. Science of the Total Environment.

2014;**47**(2):81-82

2019;**648**:681-692

Impacts. 2018;**20**(5):767-779

2005

[10] Agency USEP. Reregistration Eligibility Decision (RED) 2,4-D. Washington, DC: EPA 738-R-05-002;

[1] Atwood D, Paisley-Jones C. 2008- 2012 Market Estimates. Pestic Ind Sales

[2] Eurostat. Sales of pesticides by type

[3] Smith AE, Mortensen K, Aubin AJ, Molloy MM. Degradation of MCPA, 2,4-D, and other phenoxyalkanoic acid herbicides using an isolated soil bacterium. Journal of Agricultural and Food Chemistry. 1994;**42**(2):401-405

[4] Ma Q, Ding Y, Chang J, Sun X, Zhang L, Wei Q, et al. Comprehensive insights on how 2,4-dichlorophenoxyacetic acid retards senescence in post-harvest citrus fruits using transcriptomic and proteomic approaches. Journal of Experimental Botany. 2014;**65**(1):61-74

[5] Kamrin MA, editor. Phenoxy and benzoic acid herbicides. In: Pesticide Profiles. New York: CRC Press; 1997. pp.

[6] Skiba E, Wolf WM. Commercial phenoxyacetic herbicides control heavy metal uptake by wheat in a divergent way than pure active substances alone. Environmental Sciences Europe.

[7] Paszko T, Muszyński P, Materska M, Bojanowska M, Kostecka M, Jackowska I. Adsorption and degradation of

phenoxyalkanoic acid herbicides in soils: A review. Environmental Toxicology and Chemistry. 2016;**35**(2):271-286

[8] Gavrilescu M. Fate of pesticides in the environment and its bioremediation.

[9] Parajulee A, Lei YD, Cao X, McLagan DS, Yeung LWY, Mitchell CPJ, et al. Comparing winter-time herbicide behavior and exports in urban, rural, and mixed-use watersheds.

Engineering in Life Sciences.

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[51] Han L, Zhao D, Li C. Isolation and 2,4-D-degrading characteristics of *Cupriavidus campinensis* BJ71. Brazilian Journal of Microbiology.

[52] Long ZX, Yan Z, Xin Z, Li YS. Biodegradation of the herbicide 2, 4-dichlorophenoxyacetic acid by a new isolated strain of *Achromobacter* sp. LZ35. Current Microbiology.

[53] Maltseva O, Mcgowan C, Fulthorpet R, Oriel P. Degradation of 2,4-dichlorophenoxyacetic acid by haloalkaliphilic bacteria. Microbiology.

1 May 1996;**142**(5):1115-1122

2,4-dichlorophenoxyacetic aciddegrading populations in soil. Applied and Environmental Microbiology.

[55] Lappin HM, Greaves MP, Slatert JH. Degradation of the herbicide mecoprop [2-(2-methyl-4 chlorophenoxy) propionic acid] by a synergistic microbial community. Applied and Environmental Microbiology. 1985;**49**(2):429-433

1994;**60**(4):1116-1120

2006;**72**(2):1476-1486

[57] Fukumori F, Hausinger RP. *Alcaligenes eutrophus* JMP134 "2,4-dichlorophenoxyacetate monooxygenase" is an

[54] Ka JO, Holben WE, Tiedje JM. Use of gene probes to aid in recovery and identification of functionally dominant

[56] Bælum J, Henriksen T, Christian H, Hansen B, Jacobsen CS. Degradation of 4-chloro-2-methylphenoxyacetic acid in top- and subsoil is quantitatively linked to the class III tfdA gene. Applied and Environmental Microbiology.

2015;**46**(2):433-441

2017;**74**(2):193-202

*DOI: http://dx.doi.org/10.5772/intechopen.88256*

[44] Ka JO, Holben WE, Tiedje JM. Analysis of competition in soil among 2,4-dichlorophenoxyacetic acid-degrading bacteria. Applied and Environmental Microbiology.

[45] Shimojo M, Kawakami M, Amada K. Analysis of genes encoding the 2,4-dichlorophenoxyacetic aciddegrading enzyme from *Sphingomonas agrestis* 58-1. Journal of Bioscience and Bioengineering. 2009;**108**(1):56-59

[46] Gözdereliler E, Boon N, Aamand J, De Roy K, Granitsiotis MS, Albrechtsen

functionalities, community structures, and dynamics of herbicide-degrading communities cultivated with different substrate concentrations. Applied and Environmental Microbiology.

[47] Nielsen TK, Xu Z, Gözdereliler E, Aamand J, Hansen LH, Sørensen SR. Novel insight into the genetic context of the cadAB genes from a 4-chloro-2-methylphenoxyacetic aciddegrading *Sphingomonas*. Stevenson B, editor. PLoS ONE. 2013;**8**(12):e83346

[48] Müller RH, Jorks S, Kleinsteuber S, Babel W. *Comamonas acidovorans* strain MC1: A new isolate capable of degrading the chiral herbicides dichlorprop and mecoprop and the herbicides 2,4-D and MCPA. Microbiological Research.

[49] Vallaeys T, Albino L, Soulas G, Wright AD, Weightman AJ. Isolation and characterization of a stable 2,4-dichlorophenoxyacetic acid degrading bacterium, *Variovorax paradoxus*, using chemostat culture. Biotechnology Letters.

[50] González AJ, Gallego A, Gemini VL, Papalia M, Radice M, Gutkind G, et al. Degradation and detoxification of the

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*Environmental Chemistry and Recent Pollution Control Approaches*

Degrade, Survive, Adapt, and Evolve. Japan: Springer; 2014. pp. 43-57

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2, 4-dichlorophenoxyacetic acid-degrading bacteria. Applied Environmental Microbiology.

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[46] Gözdereliler E, Boon N, Aamand J, De Roy K, Granitsiotis MS, Albrechtsen HJ, et al. Comparing metabolic functionalities, community structures, and dynamics of herbicide-degrading communities cultivated with different substrate concentrations. Applied and Environmental Microbiology. 2013;**79**(1):367-375

[47] Nielsen TK, Xu Z, Gözdereliler E, Aamand J, Hansen LH, Sørensen SR. Novel insight into the genetic context of the cadAB genes from a 4-chloro-2-methylphenoxyacetic aciddegrading *Sphingomonas*. Stevenson B, editor. PLoS ONE. 2013;**8**(12):e83346

[48] Müller RH, Jorks S, Kleinsteuber S, Babel W. *Comamonas acidovorans* strain MC1: A new isolate capable of degrading the chiral herbicides dichlorprop and mecoprop and the herbicides 2,4-D and MCPA. Microbiological Research. 1999;**154**(3):241-246

[49] Vallaeys T, Albino L, Soulas G, Wright AD, Weightman AJ. Isolation and characterization of a stable 2,4-dichlorophenoxyacetic acid degrading bacterium, *Variovorax paradoxus*, using chemostat culture. Biotechnology Letters. 1998;**20**(11):1073-1076

[50] González AJ, Gallego A, Gemini VL, Papalia M, Radice M, Gutkind G, et al. Degradation and detoxification of the

herbicide 2,4-dichlorophenoxyacetic acid (2,4-D) by an indigenous *Delftia* sp. strain in batch and continuous systems. International Biodeterioration and Biodegradation. 2012;**66**(1):8-13

[51] Han L, Zhao D, Li C. Isolation and 2,4-D-degrading characteristics of *Cupriavidus campinensis* BJ71. Brazilian Journal of Microbiology. 2015;**46**(2):433-441

[52] Long ZX, Yan Z, Xin Z, Li YS. Biodegradation of the herbicide 2, 4-dichlorophenoxyacetic acid by a new isolated strain of *Achromobacter* sp. LZ35. Current Microbiology. 2017;**74**(2):193-202

[53] Maltseva O, Mcgowan C, Fulthorpet R, Oriel P. Degradation of 2,4-dichlorophenoxyacetic acid by haloalkaliphilic bacteria. Microbiology. 1 May 1996;**142**(5):1115-1122

[54] Ka JO, Holben WE, Tiedje JM. Use of gene probes to aid in recovery and identification of functionally dominant 2,4-dichlorophenoxyacetic aciddegrading populations in soil. Applied and Environmental Microbiology. 1994;**60**(4):1116-1120

[55] Lappin HM, Greaves MP, Slatert JH. Degradation of the herbicide mecoprop [2-(2-methyl-4 chlorophenoxy) propionic acid] by a synergistic microbial community. Applied and Environmental Microbiology. 1985;**49**(2):429-433

[56] Bælum J, Henriksen T, Christian H, Hansen B, Jacobsen CS. Degradation of 4-chloro-2-methylphenoxyacetic acid in top- and subsoil is quantitatively linked to the class III tfdA gene. Applied and Environmental Microbiology. 2006;**72**(2):1476-1486

[57] Fukumori F, Hausinger RP. *Alcaligenes eutrophus* JMP134 "2,4-dichlorophenoxyacetate monooxygenase" is an

alpha-ketoglutarate-dependent dioxygenase. Journal of Bacteriology. 1993;**175**(7):2083-2086

[58] Itoh K, Tashiro Y, Uobe K, Kamagata Y, Suyama K, Yamamoto H. Root nodule *bradyrhizobium* spp. Harbor tfdA and cadA, homologous with genes encoding 2,4-dichlorophenoxyacetic acid-degrading proteins. Applied and Environmental Microbiology. 2004;**70**(4):2110-2118

[59] Pieper DH, Reineke W, Engesser K-H, Knackmuss H-J. Metabolism of 2,4-dichlorophenoxyacetic acid, 4-chloro-2-methylphenoxyacetic acid and 2-methylphenoxyacetic acid by *Alcaligenes eutrophus* JMP 134. Archives of Microbiology. 1988;**150**(1):95-102

[60] Schleinitz KM, Kleinsteuber S, Vallaeys T, Babel W. Localization and characterization of two novel genes encoding stereospecific dioxygenases catalyzing 2(2,4-dichlorophenoxy) propionate cleavage in *Delftia acidovorans* MC1. Applied and Environmental Microbiology. 2004;**70**(9):5357-5365

[61] Liu Y, Liu S, Drake HL, Horn MA. Consumers of 4-chloro-2 methylphenoxyacetic acid from agricultural soil and drilosphere harbor cadA, r/sdpA, and tfdA-like gene encoding oxygenases. FEMS Microbiology Ecology. 2013;**86**:114-129

[62] Gerhardt KE, Huang X, Glick BR, Greenberg BM. Plant science phytoremediation and rhizoremediation of organic soil contaminants: Potential and challenges. Plant Science. 2009;**176**:20-30

[63] Machado F, Anderson C, Meenken E, Gillespie R, Peterson M, Harold M. The importance of plants to development and maintenance of soil structure, microbial communities and ecosystem functions. Soil and Tillage Research. 2018;**175**:139-149

[64] Singer AC, Crowley DE, Thompson IP. Secondary plant metabolites in phytoremediation and biotransformation. Trends in Biotechnology. 2003;**21**(3):123-130

[65] Reshma AC, Krishna RR. Plant species identification for phytoremediation of mixed contaminated soils. Journal of Hazardous, Toxic, and Radioactive Waste. 2012;**19**:218-229

[66] Siwek M. Biologiczne sposoby oczyszczania srodowiska fitoremediacja. Wiadomości Botaniczne. 2008;**52**(1/2):23-28

[67] Posmyk K, Urbaniak M. Fitoremediacja jako alternatywna metoda oczyszczania środowiska. Aura. 2014;**7**:10-12

[68] Ramborger BP, Ortis Gularte CA, Rodrigues DT, Gayer MC, Sigal Carriço MR, Bianchini MC, et al. The phytoremediation potential of *Plectranthus neochilus* on 2,4-dichlorophenoxyacetic acid and the role of antioxidant capacity in herbicide tolerance. Chemosphere. 2017;**188**:231-240

[69] Rodríguez-Serrano M, Pazmiño DM, Sparkes I, Rochetti A, Hawes C, Romero-Puertas MC, et al. 2,4-Dichlorophenoxyacetic acid promotes S-nitrosylation and oxidation of actin affecting cytoskeleton and peroxisomal dynamics. Journal of Experimental Botany. 2014;**65**(17):4783-4793

[70] Grossmann K. Auxin herbicides: Current status of mechanism and mode of action. Pest Management Science. 2009;**66**(2):113-120

[71] Musilova L, Ridl J, Polivkova M, Macek T, Uhlik O. Effects of secondary plant metabolites on microbial populations: Changes in community structure and metabolic activity in

**101**

*Biological Remediation of Phenoxy Herbicide-Contaminated Environments*

photosynthetic plants. Chemosphere.

[80] Gilbert ES, Crowley DE. Plant compounds that induce polychlorinated

*Arthrobacter* sp. strain B1B. Applied and Environmental Microbiology.

[81] Leigh MB, Fletcher JS, Fu X, Schmitz FJ. Root turnover: An important source of microbial

of recalcitrant contaminants.

substrates in rhizosphere remediation

Environmental Science & Technology.

[82] Yi H, Crowley DE. Biostimulation of PAH degradation with plants containing high concentrations of linoleic acid. Environmental Science & Technology.

[83] Ely CS, Smets BF. Bacteria from wheat and cucurbit plant roots metabolize PAHs and aromatic root exudates: Implications for rhizodegradation. International Journal of Phytoremediation. 3 Oct

[84] Fraraccio S, Strejcek M, Dolinova I, Macek T, Uhlik O. Secondary compound hypothesis revisited: Selected plant secondary metabolites promote bacterial degradation of cis-1,2 dichloroethylene (cDCE). Scientific

biphenyl biodegradation by

1994;**28**(5):981-988

1997;**63**(5):1933-1938

2002;**36**(7):1579-1583

2007;**41**(12):4382-4388

2017;**19**(10):877-883

Reports. 2017;**7**(1):8406

[85] Hu C, Zhang Y, Tang X, Luo W. PCB biodegradation and bphA1 gene expression induced by salicylic acid and biphenyl with *Pseudomonas fluorescence* P2W and *Ralstonia eutropha* H850. Polish Journal of Environmental

Studies. 2014;**23**(5):1591-1598

[86] Uhlik O, Musilova L, Ridl J, Hroudova M, Vlcek C, Koubek J, et al. Plant secondary metabolite-induced shifts in bacterial community structure and degradative ability in contaminated

*DOI: http://dx.doi.org/10.5772/intechopen.88256*

[73] Ahemad M, Kibret M. Mechanisms and applications of plant growth promoting rhizobacteria: Current perspective. Journal of King Saud University—Science. 2014;**26**(1):1-20

[74] Kuiper I, Lagendijk EL, Bloemberg GV, Lugtenberg BJJ. Rhizoremediation: A beneficial plant-microbe interaction bioremediation: A natural method. Molecular Plant-Microbe Interactions.

[75] Glick BR. Using soil bacteria to facilitate phytoremediation. Biotechnology Advances. 2010;**28**(3):367-374

[76] Boyle JJ, Shann JR. Biodegradation of phenol, 2,4-DCP, 2,4-D, and 2,4,5-T in field-collected rhizosphere and nonrhizosphere soils. Journal of Environmental Quality. 1995;**24**(4):782

[77] Shaw LJ, Burns RG. Enhanced mineralization of [U-14C]2,4 dichloropheeoxyacetic acid in soil from the rhizosphere of *Trifolium pratense*. Applied and Environmental Microbiology. 2004;**70**(8):4766-4774

[78] Germaine KJ, Liu X, Cabellos GG, Hogan JP, Ryan D, Dowling DN. Bacterial endophyte-enhanced phytoremediation of the organochlorine herbicide 2,4-dichlorophenoxyacetic acid. FEMS Microbiology Ecology.

[79] Donnelly PK, Hegde RS, Fletcher

JS. Growth of PCB-degrading bacteria on compounds from

2006;**57**(2):302-310

contaminated environments. Iriti M, editor. International Journal of Molecular Sciences. 2016;**17**(8):1205

[72] Lugtenberg BJ, Dekkers L, Bloemberg GV. Molecular determinants of rhizosphere colonization by pseudomonas. Annual Review of Phytopathology.

2001;**39**(1):461-490

2004;**17**(1):6-15

*Biological Remediation of Phenoxy Herbicide-Contaminated Environments DOI: http://dx.doi.org/10.5772/intechopen.88256*

contaminated environments. Iriti M, editor. International Journal of Molecular Sciences. 2016;**17**(8):1205

*Environmental Chemistry and Recent Pollution Control Approaches*

[64] Singer AC, Crowley DE, Thompson IP. Secondary plant metabolites in phytoremediation and biotransformation. Trends in Biotechnology. 2003;**21**(3):123-130

[65] Reshma AC, Krishna RR. Plant species identification for phytoremediation of mixed contaminated soils. Journal of Hazardous, Toxic, and Radioactive

[66] Siwek M. Biologiczne sposoby oczyszczania srodowiska—

fitoremediacja. Wiadomości Botaniczne.

Waste. 2012;**19**:218-229

2008;**52**(1/2):23-28

2014;**7**:10-12

2017;**188**:231-240

[67] Posmyk K, Urbaniak M. Fitoremediacja jako alternatywna metoda oczyszczania środowiska. Aura.

[68] Ramborger BP, Ortis Gularte CA, Rodrigues DT, Gayer MC, Sigal Carriço MR, Bianchini MC, et al. The phytoremediation

potential of *Plectranthus neochilus* on 2,4-dichlorophenoxyacetic acid and the role of antioxidant capacity in herbicide tolerance. Chemosphere.

[69] Rodríguez-Serrano M, Pazmiño DM, Sparkes I, Rochetti A, Hawes C, Romero-Puertas MC, et al. 2,4-Dichlorophenoxyacetic acid promotes S-nitrosylation and oxidation

of actin affecting cytoskeleton and peroxisomal dynamics. Journal of Experimental Botany.

[70] Grossmann K. Auxin herbicides: Current status of mechanism and mode of action. Pest Management Science.

[71] Musilova L, Ridl J, Polivkova M, Macek T, Uhlik O. Effects of secondary

plant metabolites on microbial populations: Changes in community structure and metabolic activity in

2014;**65**(17):4783-4793

2009;**66**(2):113-120

alpha-ketoglutarate-dependent dioxygenase. Journal of Bacteriology.

Y, Suyama K, Yamamoto H. Root nodule *bradyrhizobium* spp. Harbor tfdA and cadA, homologous with genes encoding 2,4-dichlorophenoxyacetic acid-degrading proteins. Applied and Environmental Microbiology.

[59] Pieper DH, Reineke W, Engesser K-H, Knackmuss H-J. Metabolism of 2,4-dichlorophenoxyacetic acid, 4-chloro-2-methylphenoxyacetic acid and 2-methylphenoxyacetic acid by *Alcaligenes eutrophus* JMP 134. Archives of Microbiology. 1988;**150**(1):95-102

[60] Schleinitz KM, Kleinsteuber S, Vallaeys T, Babel W. Localization and characterization of two novel genes encoding stereospecific dioxygenases catalyzing 2(2,4-dichlorophenoxy) propionate cleavage in *Delftia acidovorans* MC1. Applied and Environmental Microbiology. 2004;**70**(9):5357-5365

[61] Liu Y, Liu S, Drake HL, Horn MA. Consumers of 4-chloro-2 methylphenoxyacetic acid from agricultural soil and drilosphere harbor cadA, r/sdpA, and tfdA-like gene encoding oxygenases. FEMS Microbiology Ecology. 2013;**86**:114-129

[62] Gerhardt KE, Huang X, Glick BR, Greenberg BM. Plant science phytoremediation and rhizoremediation of organic soil contaminants: Potential

and challenges. Plant Science.

[63] Machado F, Anderson C,

Research. 2018;**175**:139-149

Meenken E, Gillespie R, Peterson M, Harold M. The importance of plants to development and maintenance of soil structure, microbial communities and

ecosystem functions. Soil and Tillage

2009;**176**:20-30

[58] Itoh K, Tashiro Y, Uobe K, Kamagata

1993;**175**(7):2083-2086

2004;**70**(4):2110-2118

**100**

[72] Lugtenberg BJ, Dekkers L, Bloemberg GV. Molecular determinants of rhizosphere colonization by pseudomonas. Annual Review of Phytopathology. 2001;**39**(1):461-490

[73] Ahemad M, Kibret M. Mechanisms and applications of plant growth promoting rhizobacteria: Current perspective. Journal of King Saud University—Science. 2014;**26**(1):1-20

[74] Kuiper I, Lagendijk EL, Bloemberg GV, Lugtenberg BJJ. Rhizoremediation: A beneficial plant-microbe interaction bioremediation: A natural method. Molecular Plant-Microbe Interactions. 2004;**17**(1):6-15

[75] Glick BR. Using soil bacteria to facilitate phytoremediation. Biotechnology Advances. 2010;**28**(3):367-374

[76] Boyle JJ, Shann JR. Biodegradation of phenol, 2,4-DCP, 2,4-D, and 2,4,5-T in field-collected rhizosphere and nonrhizosphere soils. Journal of Environmental Quality. 1995;**24**(4):782

[77] Shaw LJ, Burns RG. Enhanced mineralization of [U-14C]2,4 dichloropheeoxyacetic acid in soil from the rhizosphere of *Trifolium pratense*. Applied and Environmental Microbiology. 2004;**70**(8):4766-4774

[78] Germaine KJ, Liu X, Cabellos GG, Hogan JP, Ryan D, Dowling DN. Bacterial endophyte-enhanced phytoremediation of the organochlorine herbicide 2,4-dichlorophenoxyacetic acid. FEMS Microbiology Ecology. 2006;**57**(2):302-310

[79] Donnelly PK, Hegde RS, Fletcher JS. Growth of PCB-degrading bacteria on compounds from

photosynthetic plants. Chemosphere. 1994;**28**(5):981-988

[80] Gilbert ES, Crowley DE. Plant compounds that induce polychlorinated biphenyl biodegradation by *Arthrobacter* sp. strain B1B. Applied and Environmental Microbiology. 1997;**63**(5):1933-1938

[81] Leigh MB, Fletcher JS, Fu X, Schmitz FJ. Root turnover: An important source of microbial substrates in rhizosphere remediation of recalcitrant contaminants. Environmental Science & Technology. 2002;**36**(7):1579-1583

[82] Yi H, Crowley DE. Biostimulation of PAH degradation with plants containing high concentrations of linoleic acid. Environmental Science & Technology. 2007;**41**(12):4382-4388

[83] Ely CS, Smets BF. Bacteria from wheat and cucurbit plant roots metabolize PAHs and aromatic root exudates: Implications for rhizodegradation. International Journal of Phytoremediation. 3 Oct 2017;**19**(10):877-883

[84] Fraraccio S, Strejcek M, Dolinova I, Macek T, Uhlik O. Secondary compound hypothesis revisited: Selected plant secondary metabolites promote bacterial degradation of cis-1,2 dichloroethylene (cDCE). Scientific Reports. 2017;**7**(1):8406

[85] Hu C, Zhang Y, Tang X, Luo W. PCB biodegradation and bphA1 gene expression induced by salicylic acid and biphenyl with *Pseudomonas fluorescence* P2W and *Ralstonia eutropha* H850. Polish Journal of Environmental Studies. 2014;**23**(5):1591-1598

[86] Uhlik O, Musilova L, Ridl J, Hroudova M, Vlcek C, Koubek J, et al. Plant secondary metabolite-induced shifts in bacterial community structure and degradative ability in contaminated soil. Applied Microbiology and Biotechnology. 2013;**97**(20):9245-9256

[87] Siciliano SD, Germida JJ, Banks K, Greer CW, Lafayette W. Changes in microbial community composition and function during a polyaromatic hydrocarbon phytoremediation field trial. Applied and Environmental Microbiology. 2003;**69**(1):483-489

[88] Master ER, Mohn WW. Induction of bphA, encoding biphenyl dioxygenase, in two polychlorinated biphenyldegrading bacteria, Psychrotolerant *Pseudomonas* strain Cam-1 and *Mesophilic Burkholderia* strain LB400. Applied and Environmental Microbiology. 2001;**67**(6):2669-2676

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[91] Eevers N, Hawthorne JR, White JC, Vangronsveld J, Weyens N, Hawthorne JR, et al. Endophyteenhanced phytoremediation of DDE-contaminated using *Cucurbita pepo*: A field trial. International Journal of Phytoremediation. 21 Mar 2018;**20**(4):301-310

[92] Nahi A, Othman R, Omar D. Effects of Sb16 bacterial strain and herbicides on endophytic bacterial populations and growth of aerobic rice. Plant, Soil and Environment. 2016;**62**(10):453-459

[93] Bælum J, Prestat E, David MM, Strobel BW, Jacobsen CS, Park K.

Modeling of phenoxy acid herbicide mineralization and growth of microbial degraders in 15 soils monitored by quantitative real-time PCR of the functional tfdA gene. Applied and Environmental Microbiology. 2012;**78**(15):5305-5312

[94] Boivin A, Amellal S, Schiavon M, van Genuchten MT. 2,4-Dichlorophenoxyacetic acid (2,4-D) sorption and degradation dynamics in three agricultural soils. Environmental Pollution. 2005;**138**(1):92-99

**103**

**Chapter 6**

Control

*Jan Broda*

**Abstract**

**1. Introduction**

carpets and other interior textiles.

Biodegradation of Sheep Wool

Geotextiles Designed for Erosion

Wool geotextiles were formed from the meandrically arranged thick ropes and used as erosion control products. The geotextiles were installed in the experimental sites to protect the endangered slopes and the bank of ditches. Additionally, as a reinforcement of the soil, loose wool fibres were applied. The progress of wool biodegradation on the slope was investigated. Changes in the outer appearance, mechanical parameters, molecular structure and fibre morphology were analysed. Moreover, the nitrogen content in the soil and the effect of compounds released into soil on the grass growth were studied. The measurements revealed that the biodegradation starts at the cleavage of disulphide bonds, followed by disruption of the peptide bonds. Degradation is initiated in the outer cuticle and is followed by the decomposition of the inner cortical cells. During biodegradation, the nitrogen-rich compounds are released. The compounds act as an effective fertiliser which supports the growth of grass and significantly accelerates the greening of the slope.

**Keywords:** wool, geotextiles, erosion control, biodegradation, enzymes

valued, and manufacturing of wool products has a very long tradition.

Sheep wool belongs to the oldest known natural fibres. The history of wool application coincides with the history of mankind. In many countries wool is highly

For a long time, wool has been used for production of apparel textiles. Fine fibres have been used to manufacture high-quality fabric for luxury clothing. Coarser fibres were used to produce yarns used for knitting the traditional wool products. Coarser wool has been also successfully used for production of blankets,

In the recent years, apart from the traditional products available on the market, new wool technical textiles appeared [1, 2]. Among them the most popular are acoustic and thermal materials used in the construction industry for insulation of pitched roofs, walls and ceilings [3–8]. For commercial application, also wool oil sorbents [9, 10], heavy metal-absorbing materials [11, 12] and geotextiles [13] are used.

Wool geotextiles include mats or blankets spread out on the ground and products which are buried in the soil. The mats are used to protect grass seeds in the ground. Due to wool's natural ability to regulate the temperature, mats create the proper microclimate for seed germination and, later, ensure favourable conditions for plant growth. The geotextiles buried in the soil are used in agriculture as plant fertilisers

## **Chapter 6**

*Environmental Chemistry and Recent Pollution Control Approaches*

Modeling of phenoxy acid herbicide mineralization and growth of microbial

degraders in 15 soils monitored by quantitative real-time PCR of the functional tfdA gene. Applied and Environmental Microbiology.

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Schiavon M, van Genuchten MT. 2,4-Dichlorophenoxyacetic acid (2,4-D) sorption and degradation dynamics in three agricultural soils. Environmental

Pollution. 2005;**138**(1):92-99

soil. Applied Microbiology and Biotechnology. 2013;**97**(20):9245-9256

[87] Siciliano SD, Germida JJ, Banks K, Greer CW, Lafayette W. Changes in microbial community composition and function during a polyaromatic hydrocarbon phytoremediation field trial. Applied and Environmental Microbiology. 2003;**69**(1):483-489

[88] Master ER, Mohn WW. Induction of bphA, encoding biphenyl dioxygenase, in two polychlorinated biphenyldegrading bacteria, Psychrotolerant *Pseudomonas* strain Cam-1 and *Mesophilic Burkholderia* strain LB400. Applied and Environmental Microbiology. 2001;**67**(6):2669-2676

[89] McLoughlin E, Rhodes AH, Owen SM, Semple KT. Biogenic volatile organic compounds as a potential stimulator for organic contaminant degradation by soil microorganisms.

[90] Eevers N, Hawthorne JR, White JC, Vangronsveld J, Weyens N. Exposure of *Cucurbita pepo* to DDE-contamination alters the endophytic community: A cultivation dependent vs a cultivation independent approach. Environmental

Environmental Pollution. 2009;**157**(1):86-94

Pollution. 2016;**209**:147-154

2018;**20**(4):301-310

[91] Eevers N, Hawthorne JR, White JC, Vangronsveld J, Weyens N, Hawthorne JR, et al. Endophyteenhanced phytoremediation of DDE-contaminated using *Cucurbita pepo*: A field trial. International Journal of Phytoremediation. 21 Mar

[92] Nahi A, Othman R, Omar D. Effects of Sb16 bacterial strain and herbicides on endophytic bacterial populations and growth of aerobic rice. Plant, Soil and Environment. 2016;**62**(10):453-459

[93] Bælum J, Prestat E, David MM, Strobel BW, Jacobsen CS, Park K.

**102**
