Phenol degradation kinetics by two immobilized bacterial consortia in batch and continuous experiments

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Phenol degradation kinetics by two immobilized bacterial

consortia in batch and continuous experiments

Avellaneda-Franco, Laura1, Gonzales Barrios, Andrés2 & Dussán, Jenny1.

Abstract

The removal of phenol from industrial effluents is a requirement. Biodegradation of phenol is a promising alternative for this challenge. Here, the effect of mass transfer resistances and the potential effect of horizontal gene transfer on phenol biodegradation rate were assessed. In batch assays profiles of phenol over time were obtained at different phenol initial concentration with free and immobilized cells, this allows to obtain a biodegradation rate for each treatment. The maximum obtained phenol degradation rate for each treatment varies between . Also, the effect of mass transfer limitation was evaluated at different flow rate in a packed bed reactor. This experiment demonstrated that degradation rate limits the consumption of phenol. Additionally, in order to evaluate the potential effect of horizontal gene transfer two consortia were proposed. One consortium is comprised by phenol-mineralizing and the other is comprised by phenol-phenol-mineralizing and non-phenol-mineralizing strains. Comparing the phenol degradation rate between consortium 1 and 2 in batch and continuous experiments no significant difference was found, therefore, in this particular case, the horizontal gene transfer did not enhance the phenol degradation process. However, horizontal gene transfer was evidenced, but the recipients cells were not capable of mineralizing phenol according to the test performed after the selection process. To explain this result, three hypotheses were proposed and the plasmid DNA was extracted and sent to be sequenced.

Key words: phenol biodegradation rate, mass transfer, consortium, packed bed reactor, conjugation, immobilized cells.

1. Introduction

Wastewaters contaminated with phenol are generated by several industries (1). Phenol is a toxic compound; it can be fatal by ingestion, inhalation, or skin absorption. This compound can be lethal to fish at concentrations of 5-25ppm (1, 2). Additionally, phenol can inhibit photosynthesis at concentrations above to 100ppm (3). For the above reasons this compound is listed among the priority organic pollutants by the US Environmental Protection Agency (EPA), which has set a desirable limit of phenol discharge concentration of 0.168ppm (1).

1

Centro de Investigaciones Microbiológicas (CIMIC), Universidad de los Andes, Colombia.

2

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The biodegradation of phenol has been shown to be economical, versatile and have the possibility of complete mineralization of phenol (1, 4-8). Several microorganisms, including fungi and bacteria, have been used for phenol biodegradation (1, 5). In bacteria the necessary genes for the production of enzymes involved in the phenol degradation have been found in chromosome and plasmid (9, 10). In bacteria the general degradation mechanism is well known; first the phenol is oxidized to catechol by a monoxygenase, then the catechol can be degraded to acetaldehyde and pyruvate by meta-degradation route or to acetyl-CoA and succinate by the ortho route (11), routes that are catalyzed by dioxygenases (12). Furthermore, the use of bacterial consortia has been proposed because biodegradation rate depends on cooperative metabolic reactions of mixed microbial populations and interactions between diversity and productivity (13, 14). Moreover, mobile genetic elements (MGE) can be transferred between different bacteria and contribute to bacterial genome plasticity. Horizontal transfer of MGE in indigenous bacteria plays an important role in the propagation of phenol catabolic genes and enhances the removal of this compound (15, 16). Dissemination of a conjugative plasmid in a bioreactor has become a promising method of bioremediation (16).

In a packed-bed reactor (PBR) the biodegradation of phenol is carried out by immobilized microorganism, which provides many advantages such as higher biomass concentration, minimization of inhibition, resistance to chemical environment, facilitate separation of biomass from pollutant-bearing solution, and improve degradation efficiency (4, 7, 17). Moreover, the use of PBR allows a better conversion in higher volume, easy scale-up and the reaction proceeds as the wastewater progress along the reactor(18) . However, one of the main restrictions in the bioreactor is the efficiency with which the mass transfer occurs; this affects directly the rate of phenol degradation.

The mass transfer occurs in two principal steps. First, external mass transfer occurs; phenol is transfer from the bulk of the fluid to the external surface of the solid particle. The second step, internal mass transfer occurs; this consists in the phenol diffusion from the immobilized matrix surface to the active site of the particle, the cells in this case. The first step is controlled by convective mass transfer; this varies according to the volumetric flow rate and the difference between phenol concentration in the bulk of fluid and in the particle surface. Second step is controlled by diffusion mass transfer, which is affected by the particle size and the phenol concentration on the particle surface. Therefore, the interest variables to optimize the phenol degradation process in a PBR are volumetric flow rate, phenol initial concentration and particle size (18-20).

In this study, two consortia were formulated to assess the potential effect of horizontal gene transfer between native strains on phenol degradation rate in batch and packed bed reactor. Also, the limitation of phenol diffusion, from surface of particle to active site, was evaluated for two consortia at different initial phenol concentration in batch assays. Finally,

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the limitation from convective flux of mass transfer was evaluated at different flow rate in the packed bed reactor using the two consortia.

2. Materials and Methods

2.1.Bacteria and growth conditions

Acinetobacter sp. ES12, Acinetobacter calcoaceticus EX5, Arthobacter sp. Cr47 and Chryseobacerium sp. OgT6 were used in this study. These strains were previously isolated from hydrocarbon contaminated soil of Colombia and characterized by 16S in previous studies (21-23). The Acinetobacter strains do not display kanamycin resistance, contrary to Arthobacter sp. Cr47 and Chryseobacerium sp. OgT6. Luria-Bertani (LB): tryptone (10g/L); NaCl (10g/L) and yeast extract (5g/L) and Medium Minimum of Salts (MMS): KH2PO4 (0.5g/L); Na2SO4 (2.0g/L); KNO3 (2.0g/L); CaCl2 (0.001g/L); FeSO4 (0.0004 g/L) and MgSO4∙ 7H2O (1.0g/L), supplement with the tested concentration of phenol. When solid media was required bacteriological agar (15g/L) was supplemented. The strains were grown at 30°C.

2.2.Phenol biodegradation rate.

In order to formulate the consortia, a phenol consumption test was performed for each strain in batch experiments by triplicate. From the results obtained from batch experiments performed with individual strains two consortia were established. Consortium 1 was composed of both Acinetobacter strains with equal volume proportion of each strain. Consortium 2 was composed with the four strains in equal volume proportion. Also, free and immobilized on pumice cells treatments of consortium 1 and 2 were used to perform phenol consumption tests at different phenol initial concentrations. The comparison of phenol degradation the behavior between defined consortia allowed the evaluation of the potential effect of conjugation to enhance the biodegradation rate. Phenol initial concentrations tested were 20, 100, 500 and 1000ppm.

Overnight culture of each strain was made in LB broth. An aliquot of 0.3mL of overnight culture was transferred to a 30mL flask containing 10mL of LB broth, which was incubated at 30ºC and 220rpm until the bacterial population reached logarithmic phase (A600nm =0.7-0.8). An aliquot of 20mL was harvested and resuspended in 10mL of fresh MMS supplemented with 100ppm of phenol and incubated at 30°C, without shaking, for 48h. This protocol was performed to active the catabolic phenol genes. To evaluate the effect of cells immobilized on pumice the same protocol was performed, but the harvested cells were resuspended in 10mL of MMS with 3.5g of pumice.

After 48h, 1mL of the free cells was harvest and inoculated onto 10mL of MMS supplemented with the test phenol concentration. On the other hand, for the immobilized

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cells, the pumice was washed with 3mL of sterile distilled water. Subsequently, 1g of pumice was transferred at fresh 10mL of MMS supplemented with the test phenol concentration. Flasks were incubated statically for 48h at 30°C. Aliquots of 20µL were sampled to determinate phenol concentration over time. Negative control followed the same procedure without inoculate cells. Phenol concentration was evaluated by measuring absorbance at 270nm.

The profiles of phenol degradation over time at different initial concentrations were used to obtain the biodegradation rate of each treatment. Also, these results allowed the evaluation of the potential effect of conjugation on phenol degradation rate by comparing the performance obtained from consortium 1 and 2. The phenol degradation rate were obtained following a mathematical model, which takes account the degradation of the substrate (phenol) in cells and inhibition of substrate, which means the loss of viability of cells at higher concentration of phenol. The biodegradation reaction can be summarized in Figure 1.

[ ] [ ] [ ] [ ] [ ]

[ ]

[ ]

Figure 1. Reaction of phenol degradation takes account inhibition by substrate.

represents cells; , phenol and , products of the biodegradation of phenol

(acetaldehyde and pyruvate by meta-degradation route or to acetyl-CoA and

succinate by the ortho route).

In the model, the rate of product formation is given in Equation 1

[ ]

[ ]

Equation 1. The rate of product formation, , depends of the concentration of phenol

inside of cells complex, [ ], and the constant intrinsic of formation, .

The rate of the phenol inside cell complex [ ] is given in Equation 2

[ ]

[ ][ ] [ ] [ ] [ ][ ] [ ]

Equation 2. The rate of phenol inside the cell depends of the concentration of: viable cells, damage cells by the phenol concentration and concentration of phenol in media

and inside of cells. Also, this rate depends of the constants of take of phenol, of degradation of phenol inside the cells and the constant of damage of cells.

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[ ] [ ] [ ] [ ]

Equation 3. Conservation of cells in evaluated time.

Assuming a rapid equilibrium to form [ ] and [ ] complexes, the equilibrium constants can be written as shown in Equation 4.

[ ][ ] [ ]

[ ][ ] [ ] Equation 4. Equilibrium constants

Substituting Equation 4 into Equation 1 gives the Equation 5.

[ ]

[ ] [ ]

Equation 5. Degradation rate including substrate inhibition.

Where, [ ]. , is the maximum forward velocity of the reaction. , the Michaelis Menten constant, corresponds to the substrate concentration when the rate is half of the maximum velocity. , is the constant of net velocity of reaction inhibition. Finally, the and constants must be established in order to define the phenol biodegradation rate.

2.3. Horizontal gene transfer studies

Horizontal gene transfer events were evaluated with free cells using a modified protocol, which is explained in Section 2.2. After inoculating 20mL of cells in 10mL of fresh MMS supplemented with 100ppm of phenol, the flasks were incubated for 48h. At 24h, the cells that acquired the phenol genes were selected by direct plating on MMS with 50µg/mL of kanamycin in a saturated phenol atmosphere and incubated at 30°C for 7 days. Each colony forming unit was used to carry out 3 consecutive passes of replica plating on MMS with 50µg/mL of kanamycin in a saturated phenol atmosphere. The surviving colonies and the original strains were used for evaluating the presence of catechol 1,2-dioxygenase gene, pheB. Also, a phenol consumption test for each surviving colonies were performed following the protocol described in Section 2.2.

For the amplification of pheB gene, primers pheB1 and pheB2 were used. The sequence of pheB1 was CGTACCATCGAAGGACCACT and the sequence of pheB2 was GCTAAGTCTTGACGGGTTGC. PCR conditions were as follows: 1,25U of Taq polymerase (Invitrogen), 1X Buffer, 2.5 mM of MgSO4, 0.2mM of dNTPs, 0.2mM of each primer and 20ng of DNA. The program started with one step at 95°C for 5min, then 35 cycles of 94°C for 90s, 58°C for 60s and 72°C for 90s. Amplification was completed by an ultimate elongation step

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at 72°C for 10min (21). Amplification products were visualized by agarose (1%) gel electrophoresis. Finally, amplification products were sequenced by Sanger made by Macrogen Inc. Sequence similarity was evaluated with BLASTx against refseq_protein database.

2.4.Limitations of mass transfer in packed bed reactor.

Experiments at continuous flow were performed for the two formulated consortia to evaluate the effect of external mass transfer resistance and the potential effect of conjugation on phenol biodegradation rate. First, the cells were immobilized according to the protocol shown in Section 2.2. For consortium 1, 120mL of each strain that had reached logarithmic phase were mixed and harvested. For consortium 2, 60mL of each strain were mixed and harvest. In both cases, the harvested cells were resuspended in 120mL of MMS supplemented with 100ppm of phenol and transferred within a vessel that contained 38g of pumice. The vessels were incubated for 48h at 30°C.

After 48h, the pumices were washed with 100mL of sterile distilled water, and transferred to the reactor. Dimensions of the tubular reactor were 40cm of high and 2.5cm of internal diameter. The column contained 38g pumice and 0.5g of cell dry mass. Pumice average particle diameter was 0.24cm and particle density was 0.2g/cm3. The reactor was kept at 30°C. Feed solution was MMS supplemented with 100ppm of phenol. Air was supplied at an average velocity of 5 .

Collected data were used to evaluate both the potential effect of conjugation and the effect of mass transfer resistance on the overall reaction rate. The biodegradation rate for free consortia obtained from batch assays was used to simulate the profile of phenol concentration along the reactor without mass transfer resistance. Then, the effect of mass transfer resistance was calculated by comparing the simulate profile with the experimental profile of phenol along the reactor. The mass balance of phenol in packed bed reactor is shown in Equation 6 and Equation 7.

[ ] [ ]

Equation 6. Mass balance for a packed bed reactor with external mass transfer limitations.

Since [ ] [ ] and . Where , the mass of catalyst, , the cross section of the reactor, , the catalyst density, , height of the reactor, , the flow rate and , the void fraction. Subsisting, in Equation 6 the Equation 7 is obtained

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[ ] [ ]

Equation 7. Mass balance of a PBR, where is given in [ ⁄ ].

2.5.Extraction and sequencing of plasmid and genomic DNA

The DNA extraction of ES12 strains was performed by using GenElute™ Bacterial Genomic DNA Kit, Sigma Aldrich®. The plasmid extraction of ES12 strain was performed according to protocol of Sambrook, et al (24). Both plasmid and genome DNA was sent to be sequenced by 454-pyrosequencing technique (one GS-FLX plate) in Macrogen Inc (Data have not arrived yet).

3. Results

3.1.Individual strain performance on phenol consumption

As shown in Figure 2, Acinetobacter strains, EX5 and ES12, are capable of metabolizing 100ppm of phenol in 14h, while Arthobacter (Cr47) and Chryseobacterium (OgT6) are unable. Likewise, presence of pheB gene was observed just in Acinetobacter strains (Figure 3). pheB gene encoding the catechol 1,2 dioxygenase enzyme, which catalyzes the degradation of catechol in ortho route. Also, the Acinetobacter strains have the plasmid genes traG and trbB (25), these genes are important because they allow the transfer of genes by conjugation. Based on the features mentioned above two consortia were formulated. The consortium 1 was composed of mineralizing of phenol strains and consortium 2 composed for the four strains. Consortium 2 was proposed in order to have a mix of donor cells and receptors cells, in which conjugation events were expected. The consortium 1 was used like control allowing the evidence of positive or negative effects caused by the presence of non-mineralizing strains in the consortia 2. For the case of consortium 2 three ratio donor to receptor cell were probed: 1:1, 1.5:1 and 2:1, results are shown in Supplementary figure 1, where it is possible to appreciate that there is no significant difference between the tested ratios. Experiments were carried out with a ratio of 1:1.

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Figure 2. Profiles of phenol degradation by each strain during a period of 23h of

incubation. The Acinetobacter strains, EX5 and ES12, are capable of metabolizing

phenol, while while Arthobacter (Cr47) and Chryseobacterium (OgT6) are unable.

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Figure 3. Amplification of pheB gene, which encodes the catechol 1,2 dioxygenase enzyme. Lane 1; C(-), Lane 2; OgT6, Lane 3; EX5, Lane 4; ES12, Lane 5; Cr47, Lane 6-7; HyperLadder I and Lane 8: HyperLadder II –Bioline.

3.2.Effects of internal mass transfer on biodegradation phenol rate.

The degradation rate was calculated from the experimental data (Supplementary figure 2) for each tested concentration following the behavior described in Equation 5. The parameters that better fit experimental data for each treatment were found and are summarized in Table 1.

Treatment [

] [

] [ ] Consortium 1- Immobilized 122.92 406.58 66557

Consortium 1- Free cells 107.26 113.37 22862

Consortium 2- Immobilized 120 437.11 95323

Consortium 1- Free cells 100 151.34 29642

Table 1. The constants shown in this table defines the degradation rate of each treatment.

The constants shown in Table 1 were used to predict the behavior of the biodegradation rate at different initial phenol concentrations, dashed lines in Figure 4. The proposed model fits the experimental behavior, continuous lines in Figure 4. According to the behavior of biodegradation rate regarding to phenol initial concentration, it is observed that there is an initial concentration where the biodegradation rate is optimum.

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Figure 4. Comparision of experimental and predicted behavior of biodegradation rate at different initial concentrations.

From Figure 4, it is clear that a high phenol concentration causes inhibition of the degradation. The immobilized cells have a better performance at higher concentration (Figure 4). Also, the minor effect that has the increase of initial concentration on immobilized cells compare to free cells is due to the limitation of internal mass transfer, which maintains the phenol concentration inside the pumice lower than concentration of media. The phenol degradation rate can be attributed to the presence of cells, because in the abiotic control, no phenol removal occurred throughout the experiments (data not shown). Also, from experimental data a significant difference between consortium 1 and 2 was no found, this means that the consortium 2, established in this study, does not enhance the degradation process.

3.3.Horizontal gene transfer of pheB gene

After 3 replica plating passes, the surviving cells was for each donor cell. The surviving cells were designated as Arthobacter sp. based on morphologic features. Supplementary figure 3 shows the products of amplification of pheB genes in surviving cells exposed to a saturated phenol atmosphere, the amplicons were sequenced. PCR from strain Cr47 without previous contact with Acinetobacter was used like control to confirm the transfer of pheB gene (Lane 27 in Supplementary figure 3) at surviving cells. The results

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of BLASTx showed that amplified sequence matched with an Acinetobacter lwoffii strain K24CatA protein, with an average e-value of and score of 1576. This protein is type I catechol 1,2-dioxygenase. These results indicated that transfer of phenol genes was occurred. However, the evaluated colonies were unable to degrade phenol.

3.4.Effects of mass transfer on biodegradation phenol rate.

The predicted degradation rates for free cells of each consortium (Section 3.2) were used to calculate the theoretical profiles of concentration along the reactor using the Equation 7. Figure 5 shows experimental and theoretical profiles of phenol degradation concentration over column height at different flow rates and for the consortium 1 and 2. Comparing theoretical with experimental degradation rates, it is possible to observe that there is a mass transfer limitation.

Figure 5. Experimental and theoretical profiles of concentration along the column. A and B, data obtained from consortium 1 and 2, respectively.

4. Discussion

A

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For higher concentrations of phenol, a significant reduction of degradation rate was observed due to the negative effect of phenol on the viability of cells (26). The increase of phenol concentration has a higher effect on free than immobilized cells , which indicates that the immobilization support can increase the stability of cells. To diminish the effect of higher concentration, systems that integrate chemical and biological degradation have been proposed, such as the use of advanced oxidized process linked to PBR with immobilized cells (27).

The maximum biodegradation rate obtained for the consortia ⁄ is comparable with degradations rates reported with free cells of Rhodococcus rhodochrous ⁄ and with free cells of Rhodococcus erythropolis UPV-1 in continuous reactor ⁄ (28). Also, it is higher than the rate obtained by Pseudomonas putida immbolized on Zr-activated pumice ⁄ , immobilized indigenous bacteria from pulp ⁄ in continuous reactor and free cells of Candida tropicalis in a batch reactor ⁄ (28). However, it is lower than the rate obtained by Rhodococcus erythropolis immobilized on ceramic using continuous packed bed reactor ⁄ (28).

Mass transfer resistance were found in this study. Internal mass transfer plays and important role when phenol concentration in media is higher than the optimum concentration, because this allows that cells not to be directly exposed to undesired compound. Therefore immobilized cells degrade phenol more efficient than free cells at higher concentrations. The results obtained in continuous experiments show that an increase in the flow rate increases the effectiveness factor for continuous reactor, but decreases the outlet phenol concentration. To mitigate this problem the height of reactor must be increased. Knowledge of molecular regulation of phenol consuming pathway is a significant tool for this challenge. Additionally, a screening of different matrixes can improve the degradation, because prior studies have shown a significant increase between free ⁄ and immobilized on diatomaceous earth cells ⁄ using Rhodococcus erytropolis (29).

Horizontal genes transfer mechanisms were proposed and assessed as tool for improving the biodegration rate of phenol. The biodegradation rate obtained either in batch or continuous experiments are not enhanced by the use of consortia 2 compared with consortia 1, which shows that the potential effect of conjugative events does not have a significant effect on the biodegradation rate. The presence of pheB gene in surviving receptor cells, of replica plating assays; bear out the occurrence of gene transfer events. However, in the test performed before 3 passes in replica plating, these cells were not capable of degrade phenol. Key questions need to be solved to understand the limitations of gene transfer for this particular case. Phenol is converted by phenol hydroxylase in

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catechol which is cleaved by catechol 1,2-dioxygenase or catechol 2,3-dioxygenase by ortho-clevage or meta-clevage pathway, respectively (30). These enzymes are codified in plasmid, chromosome or both. Also, most aerobic aromatic-hydrocarbon biodegradations pathways converge to catechol-like intermediates (31). One hypothesis might be that the donor cells phenol genes are located on the chromosome and they are carrying a plasmid with other hydrocarbon pathway, which have pheB gene. This plasmid can be transferred to the receptor cell, which will acquire pheB gene, but it would not be able to mineralizate phenol. On the other hand, all phenol pathways related genes can be encoded in plasmid and be transferred to the receptor cell. From this last scenario two hypotheses can be proposed. First, the stability of plasmid in the receptors is so low that the plasmid is lost. Second, events of catabolite repression of phenol can occur as result of previously overnight culture before phenol degradation test. Catabolite repression of phenol has been reported in previous studies (32).

The plasmid and genome from isolated sequences of ES12 strain were sent to be sequenced to delimit the possible reasons mentioned above. The sequences datas must be analyzed and further experiments must be performed to understand the limitations of conjugation on biodegradation rate. The knowledge, about the availability of Acinetobacter phenol genes, is a key factor to predict the potential of these strains in field. Previous data had shown that introduced beneficial genetic material can enrich the natural genetic variety for biodegradation (33). Peters, et al, evidenced horizontal transfer of pheAB operon in native strains(33). In this study horizontal transfer of genes does not have a significant increase of biodegradation rate on a period of 48h; nonetheless it is likely that it can be a significant at long term. Additionally, industrial effluents have planktonic cells of bacteria, which can adhere to matrix and allows the transfer of genes between the Acinetobacter strains and the income cells.

Finally, it is a fact that industrial effluents are a mix of different pollutants. For example, water co-contaminated with phenol and chromium are generated by several industries, such as car manufacturing, wood preservation, petroleum refining and agricultural activity (34). In this sense, it is relevant that consortium 2 does not decrease the degradation rate compared with the consortium 1, because that consortium can be used to co-remediate of water with phenol and Cr(VI). Fransisco, et al, isolates Cr(VI)-resistant Acinetobacter cells (35) and Bernal, et al, demonstrated that a consortium composed by Ralstonia sp. and Arthobacter sp. CR47 can co-remediate wastewater with phenol and chromium(VI). For this reasons, the consortium 2 is proposed for carry out experiments that allow probe its potential for coremediation of this effluents.

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We thank at Fundación Instituto de Imnunología de Colombia (FIDIC) for generously lend of ultra-centrifuge. This work was funded by Faculty of Sciences at Universidad de los Andes (Bogota, Colombia) through the program “Proyecto Semilla” and for CIMIC resources.

6. Supplementary information

Supplementary figure 1. Effect of ratio donor:receptor in constortium 1 on phenol degradation.

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Supplementary figure 2. Effect of initial phenol concentration on degradation rate. The increase of phenol concentration decreases the efficiency of cells; this effect is mitigated by the use of immobilized cells. Light blue dot represents the mean of each set of data.

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Supplementary figure 3. Amplification of pheB gene. Lanes 1,14 and 16, HypperLadder I (Bioline); Lanes 2-13 and 17-26 picked out suvirving colonies in third pass of replica plating; Lane 27, Cr47 strain without contact with donors cells, Lane 28, ES12 strain; Lanes 29, EX5 strain and Lane 30, negative control (PCR without DNA).

7. References

1. El-Naas MH, Al-Muhtaseb SA, Makhlouf S. Biodegradation of phenol by Pseudomonas putida immobilized in polyvinyl alcohol (PVA) gel. Journal of Hazardous Materials. 2009;164(2-3):720-5.

2. Kumar A, Kumar S, Kumar S. Biodegradation kinetics of phenol and catechol using Pseudomonas putida MTCC 1194. Biochemical Engineering Journal. 2005;22(2):151-9.

3. Singh S, Bahadur Singh B, Chandra R. Biodegradation of Phenol in Batch Culture by Pure and Mixed Strains of Paenibacillus sp. and Bacillus cereus. Polish Journal of Microbiology. 2009;58(4):319-25. 4. Aneez Ahamad P, Mohammad Kunhi A. Enhanced degradation of phenol by <i>Pseudomonas</i> sp. CP4 entrapped in agar and calcium alginate beads in batch and continuous processes. Biodegradation. 2011;22(2):253-65.

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15

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5. Chung T-P, Tseng H-Y, Juang R-S. Mass transfer effect and intermediate detection for phenol degradation in immobilized Pseudomonas putida systems. Process Biochemistry. 2003;38(10):1497-507. 6. Cordova-Rosa SM, Dams RI, Cordova-Rosa EV, Radetski MR, Corrêa AXR, Radetski CM. Remediation of phenol-contaminated soil by a bacterial consortium and Acinetobacter calcoaceticus isolated from an industrial wastewater treatment plant. Journal of Hazardous Materials. 2009;164(1):61-66.

7. Tepe O, Dursun AY. Combined effects of external mass transfer and biodegradation rates on removal of phenol by immobilized Ralstonia eutropha in a packed bed reactor. Journal of Hazardous Materials. 2008;151(1):9-16.

8. Tziotzios G, Economou CN, Lyberatos G, Vayenas DV. Effect of the specific surface area and operating mode on biological phenol removal using packed bed reactors. Desalination. 2007;211(1-3):128-37.

9. Neumann G, Teras R, Monson L, Kivisaar M, Schauer F, Heipieper HJ. Simultaneous degradation of atrazine and phenol by Pseudomonas sp. strain ADP: effects of toxicity and adaptation. Applied and environmental microbiology. 2004 Apr;70(4):1907-12. PubMed PMID: 15066779. Pubmed Central PMCID: PMC383114. Epub 2004/04/07. eng.

10. Kivisaar M, Horak R, Kasak L, Heinaru A, Habicht J. Selection of independent plasmids determining phenol degradation in Pseudomonas putida and the cloning and expression of genes encoding phenol monooxygenase and catechol 1,2-dioxygenase. Plasmid. 1990 Jul;24(1):25-36. PubMed PMID: 2270227. Epub 1990/07/01. eng.

11. van Agteren H, Keuning S, Janssen D. Handbook on Biodegradation and Biological Treatment of Hazardous Organic Compounds: Springer; 1998.

12. Dong X, Hong Q, He L, Jiang X, Li S. Characterization of phenol-degrading bacterial strains isolated from natural soil. International Biodeterioration & Biodegradation. 2008 10//;62(3):257-62. 13. Kim YM, Ahn CK, Woo SH, Jung GY, Park JM. Synergic degradation of phenanthrene by consortia of newly isolated bacterial strains. Journal of Biotechnology. 2009;144(4):293-8.

14. Venail PA, MacLean RC, Bouvier T, Brockhurst MA, Hochberg ME, Mouquet N. Diversity and productivity peak at intermediate dispersal rate in evolving metacommunities. Nature. 2008;452(7184):210-4.

15. Springael D, Top EM. Horizontal gene transfer and microbial adaptation to xenobiotics: new types of mobile genetic elements and lessons from ecological studies. Trends in Microbiology. 2004 2//;12(2):53-8.

16. Quan X, Tang H, Ma J. Effects of gene augmentation on the removal of 2,4-dichlorophenoxyacetic acid in a biofilm reactor under different scales and substrate conditions. Journal of Hazardous Materials. 2011 1/30/;185(2–3):689-95.

17. Kuyukina MS, Ivshina IB, Serebrennikova MK, Krivorutchko AB, Podorozhko EA, Ivanov RV, et al. Petroleum-contaminated water treatment in a fluidized-bed bioreactor with immobilized Rhodococcus cells. International Biodeterioration & Biodegradation. 2009;63(4):427-32.

18. Fogler HS. Elements of chemical reaction engineering: Prentice-Hall; 1992.

19. Nauman EB. Chemical Reactor Design, Optimization, and Scaleup: McGraw-Hill Professional Publishing; 2002.

20. Shuler ML, Kargi F. Bioprocess Engineering: Basic Concepts: Prentice Hall; 2002.

21. De Zubiría L. Diversidad molecular de bacterias degradadoras de fenol aisaladas de suelos colombianos contaminados con compuestos arómaticos: Universidad de los Andes; 2006.

22. Rozo C, Dussán J. Análisis de transferencia horizontal de genes en ensayos de biorremediación con grasas recalcitrantes. Revista Colombiana de Biotecnología. 2010;XII:22-31.

23. Córdoba A, Vargas P, Dussan J. Chromate reduction by Arthrobacter CR47 in biofilm packed bed reactors. Journal of Hazardous Materials. 2008 2/28/;151(1):274-9.

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24. Sambrook J, Russell DW. Molecular Cloning: A Laboratory Manual: Cold Spring Harbor Laboratory Press; 2001. 1.55-1.8 p.

25. Sandoval F, Dussán J. Enhancement of bacterial phenol consumption trough horizontal gene transfer. Bogotá: Universidad de los Andes; 2010.

26. Banerjee A, Ghoshal AK. Phenol degradation performance by isolated Bacillus cereus immobilized in alginate. International Biodeterioration & Biodegradation. 2011 10//;65(7):1052-60. 27. Bernal OI, Sarria, V.M, Dussan, J & Camargo G. Desarrollo y evaluación de un sistema integrado para la degradación simultánea de cromo (VI) y fenoles mediante procesos de oxidación avanzada y biorremediación: Universidad de los Andes; 2008.

28. Galíndez-Mayer J, Ramón-Gallegos J, Ruiz-Ordaz N, Juárez-Ramírez C, Salmerón-Alcocer A, Poggi-Varaldo HM. Phenol and 4-chlorophenol biodegradation by yeast Candida tropicalis in a fluidized bed reactor. Biochemical Engineering Journal. 2008 2/15/;38(2):147-57.

29. Prieto MB, Hidalgo A, Rodriguez-Fernandez C, Serra JL, Llama MJ. Biodegradation of phenol in synthetic and industrial wastewater by Rhodococcus erythropolis UPV-1 immobilized in an air-stirred reactor with clarifier. Applied microbiology and biotechnology. 2002 May;58(6):853-9. PubMed PMID: 12021809. Epub 2002/05/22. eng.

30. Yu H, Peng Z, Zhan Y, Wang J, Yan Y, Chen M, et al. Novel regulator MphX represses activation of phenol hydroxylase genes caused by a XylR/DmpR-type regulator MphR in Acinetobacter calcoaceticus. PloS one. 2011;6(3):e17350. PubMed PMID: 21455294. Pubmed Central PMCID: PMC3063778. Epub 2011/04/02. eng.

31. Mesarch MB, Nakatsu CH, Nies L. Development of catechol 2,3-dioxygenase-specific primers for monitoring bioremediation by competitive quantitative PCR. Applied and environmental microbiology. 2000 Feb;66(2):678-83. PubMed PMID: 10653735. Pubmed Central PMCID: PMC91880. Epub 2000/02/02. eng.

32. Muller C, Petruschka L, Cuypers H, Burchhardt G, Herrmann H. Carbon catabolite repression of phenol degradation in Pseudomonas putida is mediated by the inhibition of the activator protein PhlR. Journal of bacteriology. 1996 Apr;178(7):2030-6. PubMed PMID: 8606180. Pubmed Central PMCID: PMC177901. Epub 1996/04/01. eng.

33. Peters M, Heinaru E, Talpsep E, Wand H, Stottmeister U, Heinaru A, et al. Acquisition of a deliberately introduced phenol degradation operon, pheBA, by different indigenous Pseudomonas species. Applied and environmental microbiology. 1997 Dec;63(12):4899-906. PubMed PMID: 9406411. Pubmed Central PMCID: PMC168818. Epub 1997/12/24. eng.

34. Gładysz-Płaska A, Majdan M, Pikus S, Sternik D. Simultaneous adsorption of chromium(VI) and phenol on natural red clay modified by HDTMA. Chemical Engineering Journal. 2012 1/1/;179(0):140-50. 35. Francisco R, Alpoim MC, Morais PV. Diversity of chromium-resistant and -reducing bacteria in a chromium-contaminated activated sludge. Journal of applied microbiology. 2002;92(5):837-43. PubMed PMID: 11972686. Epub 2002/04/26. eng.

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