Biodegradation of Benzo[a]Pyrene through the use of algae ...

Cent. Eur. J. Chem. • 12(11) • 2014 • 1133-1143 DOI: 10.2478/s11532-014-0567-6

Central European Journal of Chemistry

Biodegradation of Benzo[a]Pyrene through the use of algae Research Article

Alžbeta Takáčová1, Miroslava Smolinská2*, Jozef Ryba3, Tomáš Mackuľak4, Jana Jokrllová4, Pavol Hronec5, Gabriel Čík4 VÚRUP, a.s., Vlčie Hrdlo, 820 01 Bratislava, Slovakia

1

Department of Microbiology and Virology, Faculty of Natural Sciences, Comenius University in Bratislava, 842 15 Bratislava, Slovakia

2

Department of Fibres and Textile Chemistry, Institute of Polymer Materials, Faculty of Chemical and Food Technology, Slovak University of Technology in Bratislava, 812 37 Bratislava, Slovakia

3

Department of Environmental Engineering, Institute of Chemical and Environmental Engineering, Faculty of Chemical and Food Technology, Slovak University of Technology in Bratislava, 812 37 Bratislava, Slovakia

4

Department of Microelectronics, Institute of Electronics and Photonics, Faculty of Electrical Engineering and Information Technology, Slovak University of Technology in Bratislava, 812 19 Bratislava, Slovakia

5

Received 21 November 2013; Accepted 5 February 2014

Abstract: In this work for disposal of the biologically hard decomposed pollutant Benzo[a]Pyrene (BaP) photooxidation Chlorella kessleri was used. The simulation model system under the different experimental conditions (varying biomass and light intensity) was evaluated. For quantitative analysis of the decrease in BaP, GC/MS technique was used. The highest degradation efficiency was achieved in the case of biomass from the culture of live algae (29%) and light intensity at level of 13.5 W m-2. When the dry biomass was used, degradation under the same conditions was lower because of lack of enzymatic activity in the system. Keywords: Algae • Benzo[a]Pyrene • Biodegradation • Chlorella kessleri © Versita Sp. z o.o.

1. Introduction Polycyclic aromatic hydrocarbons (PAHs) are a large class of persistent, hydrophobic environmental pollutants emitted into the environment by incomplete combustion or pyrolysis of fossil fuels or organic materials [1]. PAHs have high octanol-water partition coefficients and low vapor pressure, and therefore these environmental pollutants are rapidly absorbed by living organisms and particulate matter. PAHs may be present in the vapor and/or dissolved phase, micelle form, sorbed to colloidal organic matter or to particles, and incorporated into biota in water [2-4]. These compounds as well as

the products of their transformation pose a hazard for all living organisms with regard to their distribution in the biosphere, risk properties and abilities of accumulation in animal and plant tissue [5]. Among over a hundred different PAH compounds, 16 have been classified as priority pollutants by the United States Environmental Protection Agency [6]. PAHs consisted of two or more condensed benzene (aromatic) rings which are linearly ordered. According to their lipophilic nature they have a good absorption by gastrointestinal tract of mammifers. Distribution to body apparatus and tissues with accumulation tendency in adipose tissues is relatively quick. During the photo-

* E-mail: [email protected] 1133

Biodegradation of Benzo[a]Pyrene through the use of algae

oxidation of PAHs in the presence of oxygen and a sensitizing substance, photosensitizers absorb light energy and reach an excited state. In the return to the ground state, excited photosensitizers transfer their energy to the ground-state oxygen, producing reactive oxygen species (ROS) [7,8]. The excess of intracellular ROS may oxidize macromolecules, leading to lipid peroxidation, protein oxidation and DNA base oxidation [9]. ROS are reactive species which generate oxidative stress and may cause damage to the membranes, mitochondria and chloroplasts and, therefore, to inhibit photosynthesis, physiological activities and cell growth. The cells have developed antioxidant defense systems, enzymatic systems, to scavenge ROS, in response to the ROS attack. This enzymatic defense system (catalase, peroxidase and superoxide dismutase) may be damaged from over-production of ROS, which is leading to the cell death [10]. PAHs found in the alga-medium system have some fate such as evaporation, adsorption and absorption biomass uptake, cell biodegradation process and photodegradation. However, photo-degradation process may be important in regard to the presence of detectable photo-induced PAH concentrations in the liquid phase extracts of heat-killed cell samples, where the biodegradation was unlikely to occur [11]. The biomass uptake of PAHs involved rapid adsorption to the cell walls and dividing into lipophilic constituents such as lipids [12]. Benzo[a]Pyrene (BaP) is an indirect carcinogen which can initiate transformation of healthy cells into cancerous pending its metabolic activation [13]. BaP was first identified in 1932 by Cook as a carcinogenic substance in laboratory animals [14]. The skin of those animals was repeatedly coated by tar. Currently, BaP is used as a marker for PAHs. It is believed that the concentration of such a marker is at a level of approximate standard of carcinogenic activity of samples. PAHs are susceptible to oxidation in the presence of ozone, UV radiation and as well as to the biological oxidation caused by microorganisms. Metabolic paths of microbiological oxidation of PAHs consisting of two and three aromatic rings (naphthalene, phenanthrene, anthracene and fluoranthene) were deeply investigated and described. Their enzymatic and genetic regulation is known [15]. The initiation step of transformation of PAHs in the prokaryotic organisms (bacteria) is caused by reaction of two atoms of oxygen with two atoms of carbon from benzene ring, through dioxogenases enzymatic system. As a consequence of this reaction, the creation of cis-dihydrodiole in presence of dehydrogenase takes place, which is subsequently 1134

transformed to dihydrogenate intermediate – pyrocatechol [16,17]. The toxicology of BaP has been mainly studied with regard to the carcinogenicity of its metabolites, but its phototoxicity is not well understood. Although some studies have indicated the lethal phototoxicity of BaP, there have been no reports regarding the pattern of cell death induced by this agent [18-20]. Aquatic ecosystems are affected by a number of noxious organic compounds, mostly of antropogenic origin, e.g. polychlorinated biphenyls (PCBs), dibenzo-para-dioxines (PCDDs) and polycyclic aromatic hydrocarbons (PAHs) [21]. The degradation of pollutants under these conditions usually involves the combined actions of two or more microorganisms. In the case of algae, it is clear that the complete degradation of aromatic pollutants is rare [22]. The application with microalgae for aquatic toxicity assays, as a useful indicator, has been conducted for the unique eco-niche in the aquatic food web and high sensitivity to a wide spectrum of environmental pollutants [23,24]. In this work the results of photooxidation of the model sample BaP with algae Chlorella kessleri in anaerobic conditions and different intensities of radiation are discussed. From an ecological and economical point of view this method of biological degradation of this type of the environmental pollution can be subsequently compared with an alternative common technology (physical-chemical degradation).

2. Experimental procedure 2.1. Biomass modification

Biomass from the culture of live algae Chlorella kessleri (LARG/1) was supplied on slant agar without bacterial contamination from Botanic institute of Slovak Academy of Science in Bratislava. The algae was batch cultured in mineral medium (MM) containing the following stock solutions (ST): (ST)1:NH4Cl 15 mg L-1, MgCl2•6H2O 12 mg L-1, CaCl2•2H2O 18 mg L-1, MgSO4•7H2O 15 mg L-1, KH2PO4 1.6 mg L-1; (ST)2: FeCl3•6H2O 64 mg L-1, Na2EDTA•2H2O 100 mg L-1; (ST)3: H3BO3 185 µg L-1, MnCl2•4H2O 415 µg L-1, ZnCl2 3 µg L-1, CoCl2•6H2O 1.5 µg L-1, CuCl2•2H2O 0.01 µg L-1, Na2MoO4•2H2O 7 µg L-1; (ST)4: NaHCO3 50 mg L-1. Stock solutions were mixed in the ratio (ST)1:(ST)2:(ST)3:(ST)4 = 10:1:1:1, pH was modified to 7.4. Experiments were carried out in a sterile laminar box JOUAN MSC 12 (USA). The cell density of inoculum was 5×104 cells per ml. Cultivation was carried out in a chamber with a stabile temperature in the range of 22 ± 3°C with four linear fluorescent bulbs (7400 lux; Testo 545, Germany).

A. Takáčová et al.

The effect of BaP mixture on the growth of C. kessleri was evaluated microscopically according the OECD test 201 (1984) by direct cells counting using the Bürker grid. Erlenmeyer flasks (250 mL) with MM were inoculated with C. kessleri (5×104 cells per mL) and the mixture of BaP (100 μg mL-1) was added. Erlenmeyer flasks were sealed with cotton wool and were cultivated under shaking by an orbital platform shaker GFL 3020 (Czech Republic) with the frequency of 120 rpm, 72 h in order to assure aerobic conditions. The growth of C. kessleri was quantified until confluent growth microscopically (0, 24, 48, 72 h). All experiments were carried out in three parallels. Dry biomass was rinse with deionized water and dried at 105°C for 24 h. After drying, dust in grinding mortar was prepared. Dry biomass with concentration of 0.1 g L-1 was added to solution of BaP (100 µg L-1). Volume for sorption was 100 mL, pH of solution (pH = 7.5) was treated with solution NaOH (30 wt.%) and solution of H2SO4 (20 wt.%).

2.2. Preparation of BaP

For laboratory tests, the model sample of BaP was prepared as follows: pure reference material (Fluka, 99 wt.% purity) was dissolved in acetone (Merck) and deionized water (1:99) to the concentration of BaP 100 µg L-1.

2.3. Preparation of model sample

The tested pollutant BaP was added to the growth culture Chlorella kessleri as well as to the reference sample contained biomass from dry algae. Inoculated flasks with this culture were kept in the air-conditioned chamber under the stable conditions: temperature 22±3°C, light source with changing light intensity (fluorescence lamp 6.2 and 13.5 W m-2), constant mixing on rotary shaker at 120 rpm. Samples for the analysis of biodegradation were taken in the specific time period of 24, 48, 72 and 144 h. Algae biomass was separated from the solution via membrane filtration (Millipore, size of pores = 0.45 µm).

2.4. Chemical analysis

After the specific time period, samples of algae and pollutant were processed by liquid to liquid extraction. For extraction hexane (Merck, purity p.a.) was used. Extracts after condense (Kuder Danish) to volume of 1 liter were consequently identified via chromatography. For identification of BaP method of inner standard Bifenyl D10 (Fluka) and gas-chromatography with mass spectroscopy GC/MS (Variant-Saturn 2100T, USA) were used.

The identification of BaP was carried out by gas chromatography combined with mass spectroscopy GC/ MS (Varian - Saturn 2100 T, USA), using the method of internal standard (Biphenyl D10, Fluka). Samples were batched into an automatic autosampler (Varian CP-8410, USA), with Varian type 1177 injector, split/splitless with a constant temperature of 290°C gas chromatograph worked in on-line connection with the detector (Mass Spectrometric Detector - MSD), which was the mass spectrometer ion trap. MSD worked with electron ionization in “Full Scan” (FS). In the mass range 45-650 m/z. Data were processed using MS Workstation. Separation was carried out in a chromatographic column EZ-guard VF-5 (Varian, USA) with dimensions L (30 m)×ID (0.25 mm) OD×(0.39 mm)+10 m EZ - guard column. Guard column is connected directly to the main column without using the clutch. Helium was used as the carrier gas (He 6.0, Messer, Slovakia) with a constant flow rate of 1 mL min-1. The volume of 1 μL syringe (Hamilton 10 μL) was charged into the injector. Chromatographic separation was performed with a temperature program: 50°C (hold time 1 min), next by the rate of 10°C min-1 to 290°C (hold time of 10 min). The total length of the analysis was 45 min.

2.5. Respirometric measurements

The rate of biological oxidation of organic carbon in samples was monitored on-line with Micro-Oxymax (M-O, Columbus Instruments, USA). Concentration of CO2 was measured by using of IR detector. Concentration of O2 was measured by paramagnetic detector. Obtained data were evaluated by M-O software. Obtained results represent the cumulative production and consumption of CO2 and O2.

2.6. Microscopic measurements

Images surface characterization of the samples were evaluated by scanning electron microscopy (SEM) JEOL JXA 840A device. Microscopic measurements of biomass were investigated with device Axio-IMAGER A (CARL-ZEISS, Germany) which allowed us to compare the biomass samples without contamination and samples of biomass after biodegradation of BaP. For microscopic observation (light: X-CITE, SERIE 120) the following filters were used: • filter (F1) - Filter set 02 - excitation G 365 nm, Beam splitter FT 396, emission LP 420 nm • filter (F2) - Filter set 05 - excitation BP 395-440 nm, Beam splitter FT 460 and emission LP 470 nm

1135


Figure 1. Absorption spectrum of BaP (solid line) and fluorescent lamp emission spectrum (dashed line); red circle – overlay of the BaP absorption spectrum and light source emission spectrum [61].

3. Results and discussion The aim of the present study was to gain more insight into the process of degradation BaP under simulation visible light (Fig. 1) and bioxidation mediated microalgae Ch. kessleri. Physico-chemical properties of PAHs can be changed by modifying their structure by abiotic (photochemical) and biotic processes (mono-oxygenated cytochrome P450) [25]. One of the most important biotic factors is sunlight. Degradation of PAH in the aquatic ecosystem may be affected by conditions of chemical oxidation and biodegradation photooxidation aquatic organisms. Microbial degradation is a relatively important route and PAHs are partially or completely degraded by some species of aquatic bacteria, algae and fungi [26,27]. Rate of biological oxidation of organic carbon was monitored on-line with M-O (Fig. 2). Monitoring the respiration of gasses after biodegradation of BaP by using algae Ch. kessleri in dry and live form showed that production of O2 is highest after 48 h of biodegradation in the case of biomass from the culture of the live algae. This corresponds with cumulative production of CO2. Oxidation of carbohydrates including PAHs to CO2 and H2O can be expressed with a general stoichiometric formula: CnHm + (n + 0.25m) O2 → nCO2 + (0.5m) H2O

(1)

Peng et al. [28] came to some interesting conclusions. Their experiment gained 1136

data and characteristics of a microorganism’s catalysis (Arthrobacter oxydans (B4) in the degradation of BaP and designed a BaP biodegradation path (Fig. 3) [28]. Peng et al. proposed that the products of PAH degradation independent of their origin, potentially caused a build up of toxic products in nature. In terms of potential toxicity this is considered to be the most important group of oxidized products of PAH [29]. Benzo[a]pyrene is one of the polycyclic aromatic hydrocarbons (PAHs), which are widely present in the environment. Traditionally, toxicological concern regarding BaP has focused on its metabolites such as (±)-trans-7,8-dihydroxy-9,10-epoxy-7,8,9,10tetrahydrobenzo[a]pyrene, because BaP becomes carcinogenic after metabolic activation [18-20]. In the Table 1 the efficiency of biodegradation of BaP is shown. Based on the results, the highest value was reached by combination of biodegradation of BaP with biomass from the culture of live algae (i.e., the higher production of O2 in comparison to the dry algae production) and photodegradation of BaP. Several studies strongly support the view point that PAHs were unstable under sunlight [30]. Photo-oxidation of PAHs is a potentially important pathway for PAHs modification in the environment [31]. Upon absorbing sunlight, PAHs can be rapidly transformed to a variety of compounds, most of which are oxidation products. As photo-products are likely to accumulate in the environment allowing for more realistic analysis of PAHs loads [32]. In the study [33], the phototoxicity of BaP was inhibited by NaN3 (quencher of singlet oxygen) but not by mannitol (quencher of hydroxy radicals [34]), showing that BaP-


Figure 2. Cumulative production of CO2 (µL; a) and O2 (µL; b); degradation of BaP.

Figure 3.

Degradation path of BaP by microorganism (Arthrobacter oxydans (B4) [28]. 1137


Table 1. t (h)

24

48

72

144

Concentration of BaP after different biodegradation conditions.

Initial concentration (µg L-1)

Light intensity 6.2 W m-2




Dry algae

100

98

96

2

4

Live algae

100

89

86

11

14

Dry algae

100

93

90

7

10

Live algae

100

77

74

23

26

Dry algae

100

91

88

9

12

Live algae

100

75

71

25

29

Dry algae

100

89

86

11

14

Live algae

100

78

75

22

25

induced phototoxicity was due to the production of singlet oxygen but not hydroxyl radicals. PAHs generate singlet oxygen through type-II photodynamic mechanism, in which the photoexcited molecules (triplet state) transfers its energy to O2 resulting in energy rich singlet oxygen. Also, BaP can be efficiently excited by the light source and the resulting photophysical processes of energy transfer to molecular oxygen that leads to a variety of photoactive species such as singlet oxygen (1O2) [35]. The efficiency of biodegradation η (%) was calculated based on an equation: η = 1 – (C×C0-1),

(2)

where C0 is the concentration of BaP at the beginning of the biodegradation process and C is the concentration of BaP at the end of the biodegradation process. Fig. 4 shows amounts of cumulative decomposed CO2 during the biodegradation process of BaP via algae biomass utilization and minimal production of CO2 during photodegradation of BaP. Initial biodegradation (first 24 h) showed relatively lower efficiency. After 48 and 144 h of monitoring of biodegradation process, values of the day average biodegradation efficiency reached relatively constant tendency. Despite of constant conditions, the efficiency of biodegradation after 144 h was lower when compared to the biodegradation without use of photooxidation. BaP is known to be transformed to diols and quinones by marine algae in a period of 5–6 days. Warshawsky et al. [36] found that Selenastrum capricornutum, a freshwater green alga metabolizes BaP to cisdihydrodiols using a dioxygenase enzyme system. Research in the field of biodegradation and biosorption processes of environmental organic pollutants has 1138

η (%)

cBaP (µg L-1)

Biosorbent

mainly concentrated on microorganisms, such as bacteria and fungi whith much less emphasis on algae [37-41]. The limited number of experimental studies has demonstrated the algae capability to accumulate and degrade wastewater-borne organic pollutants, such as dyes, phenol and BaP [22,37,42-45]. The rate of decrease of biodegradation can be caused as a consequence of the continual decrease of moisture, selective decomposition of better decomposable substances, decrease of nutrient content, accumulation of toxic intermediates as well as by other factors [46]. As a consequence of splitting of aromatic circle, the creation of organic acids occurred. The biggest benefit of such systems can be described as the consumption of created organic acids by microbial metabolism followed by subsequent synthesis of cell components and energy. Also water and CO2 is released [45]. Chemical analysis of model sample of water taken during the biodegradation (after 24, 48, 72 and 144 h) has shown not only a decrease in the concentration of BaP (Fig. 4) but the creation of degrading products as well. Before the analysis of samples via GC/MS method, standard BaP was tested. Determination limit of BaP was set to LOQ = 0.1 µg L-1 and retention time of BaP was 32.35 min. The values of concentration of BaP at the different time intervals are given in Table 1. From these results one can conclude that the best biodegradation efficiency of BaP (29%) was achieved with using combined method of biodegradation with biomass from the culture of live algae of Ch. kessleri and with radiation photooxidation (13.5 W m-2). In the present study, adsorption and absorption were not distinguished. It is interesting to address that even for the apenuated cells, the time course effect on the


Figure 4. Production of CO2 and biodegradation efficiency of BaP without algae (a), live (b) or dry (c) biomass (light intensity 13.5 W m-2). unaccounted-for BaP was present, which unlikely was due to biodegradation. The results suggest that the marine algae P. tricornutum is sensitive to the photoenhanced toxicity of PAHs, although it is capable of utilizing solar light through photosynthesis [24].

The results of experimental study indicated that 89– 90 μg L-1 of all the BaP was found in the biomass brown algae solution (Fucus), the remainder (up to 4%) was not recovered (considered to have been metabolized) and the insignificant part was in the solution. The proportion 1139


of transformed PAHs was more essential (42–49%) for the green algae. The process of BaP transformation is species specific and depends on the presence of enzymes localized in the plant cells in freshwater and marine algae. Peroxidase, diphenol oxidase and cytochrome P450 are the most important enzyme systems for detoxification of BaP. The data obtained indicate the important role of freshwater and marine algae in the fate of carcinogenic PAHs in the environment [49]. Exposure to BaP, both lysosomes and mitochondria are involved in creating intracellular environment (acidification and oxidative stress) necessary for the occurrence of caspase dependent cell death [49]. Apart from BaP, gas-chromatography also identified other aromatic carbohydrates, namely diols, aldehydes and aromatic acids. However, the higher intensity of radiation of BaP is necessary. The efficiency of this radiation is lower when compared with the UVA and UVB radiation. From the point of UV radiation, photodegradation has always been better [50]. When the BaP was exposed to the aggressive UV radiation the dominant oxidation fragments were 3-hydroxy Benzo[a] Pyrene, Benzo[a]Pyrene-9,10-dihydrodiol and Benzo[a] Pyrene-6,12-dione. Bacteria and algae which decompose PAHs, use these pollutants as a source of carbon and energy. On the other hand, fungi use PAHs for creation water-soluble compounds. The main reason for this phenomenon is the fact that bacteria and fungi use different methods of metabolism [51]. Bacteria degradation of PAHs generally starts by the attack of the aromatic rings with dioxygenase and cis-dihydrodiole is created. Consequently this product is dehydrogenated for pyrocatechol. Pyrocatechol is the main intermediate product of this splitting. The aromatic ring is split between hydroxyl groups (orto-splitting) or appended to the one of hydroxyls (meta-splitting). Another degradation of the ring causes degradation of structure and molecules and enables it to enter into the middle metabolic path of bacteria. Photochemical degradation of PAHs can cause creation of the same oxidation products as in the case of elementary chemical oxidation of PAHs (i.e., oxygen, alcoxyl radicals (RO•) and •OH radicals. At the end of degradation a complex blend of ketones, quinines, aldehydes, phenols, and carboxylic acids is obtained [29]. The results obtained from GC-MS analysis of the different cultures of the dry biomass and biomass from the culture of live algae are shown. The activity of pigment dyes, mainly chlorophyll(a), which acts as the photosensitiser in photodynamic reactions was used to analyse the reactions. In the biomass from the culture of live algae activity is higher (enzyme activity) 1140

Figure 5. SEM image of apenueted culture of Chlorella kessleri at 105°C. and when compared to the dry (relatively metabolically inactive), the difference is not great. This is because the apenuated algae been partly removed moisture, but did not pass as sterilization (120°C, pressure of 120 kPa), not be fully dampen enzyme activity. In the Fig. 5 of the scanning electron microscope, we see intact cell wall Ch. kessleri after apenuation at 105°C. Overview of the most number of publications shows that the presence of PAHs may affect the absorption of light [31,52], photosynthetic electron transport [53,54], gas exchange [55,56], photosynthetic pigments [54,57]. For evaluation of photosynthesis the fluorescence of living systems was used. This method used analysis of potential damage of photosystem. In this method, chlorophyll represents the internal indicator of photosynthetic capacity of organisms. Under optimal conditions the majority part of light energy absorbed with chlorophyll is scattered via chemical processes. A small part of this light is emitted in the form of heat and fluorescence [58]. Capacity of photosynthesis (transformation of energy of protons to chemical energy) can be controlled with adjusting the conditions, i.e. under the stress condition that leads to higher heat emission and fluorescence the capacity of photosynthesis decreases [59]. This is reflected in the case of optimal invasion conditions for algae (stress temperature changes, decrease of nutrient content, presence of toxic substance and changes in light intensity). As a fluorescence response (Fig. 6). There are two maximums are determining for chlorophyll: one at 690 nm and second one at 735 nm [60]. Concentration of chlorophyll was evaluated via UV-VIS spectroscopy. Biomass from the culture Ch. kessleri contained: chlorophyll(a) = 1.1925 mg g-1 and chlorophyll(b) = 0.435 mg g-1, concentrations were determined in accordance by norm STN ISO 10260 [61]. Outputs from


Figure 6.

Fluorescence microscope; (a) Start–Dry algae Chlorella kessleri in medium; Red Fluorescence; Resolution 40×10 (filter F1); (b) Start– Live algae Chlorella kessleri in medium; Red Fluorescence; Resolution 40×10 (filter F1); (c) Dry algae Chlorella kessleri + BaP after 144 h; Blue Fluorescence; Resolution 40×10 (filter F1); (d) Live algae Chlorella kessleri + BaP after 144 h; Blue Fluorescence; Resolution 40×10 (filter F1); (e) Pure BaP; Blue Fluorescence; Resolution 40×10 (filter F1).

fluorescence microscope are shown in Fig. 5. Obtained results from fluorescence microscope shows (case of live algae) intensive penetration of BaP into to cell of algae (Fig. 6d). While in the case of lyophilized algae occurs mainly sorption of BaP and partial penetration into cells (Fig. 6c). On the other hand in dry algae (Fig. 6a) is shown increasing of chlorophyll fluorescence as compared to the biomass from the culture of live algae (Fig. 6b). Fluorescence of chlorophyll is higher in dry algae than in biomass from the culture of live algae due to the ongoing photosynthesis.

4. Conclusion Although resistant to biochemical degradation, PAHs readily absorb UV light energy and are subject to photolytic breakdown. A complete conversion of PAHs to CO2 and H2O by UV light alone is prohibitively costly, but

it seems reasonable that a relatively brief photolytic pretreatment would render the molecule more susceptible to subsequent microbial attack. In this study, we examined the biodegradation of BaP upon visible exposure. We have shown a comparative analysis between live a dry (apenuated) culture of Chlorella kessleri in biodegradation process. The results indicated that of all the BaP consumed, 89–86 µg L-1 was found in the solution with the biomass of dry algae. For solution with the biomass from the culture of live algae the proportion of transformed BaP was more essential, 78–75 µg L-1. As follows from the results of degradation of BaP by microalgae it is not clear what process prevails (photooxidation, biodegradation). However, these processes can operate in parallel. Due to its lipophilic character and chemical stability of BaP need their cooperation. Transmitation of photooxidation products of oxidation in cells is successful. BaP molecules penetrate into cells microalgae and images from the fluorescence 1141


microscope show their degradation after 144 h. Fluorescence of chlorophyll is higher in dry algae than in biomass from the culture of live algae due to ongoing photosynthesis. The results show that the degradation of BaP was the most effective by light intensity at level of 13.5 W m-2 and the degradation efficiency was 29% by using biomass from the culture of live algae. Bioremediation by microbial processes in combination with photo-oxidation can be considered as a promising alternative to conventional technologies to remove BaP from contaminated water.

Acknowledgement This work was supported by the Slovak Research and Development Agency under the contract No. APVV0665-10.

References [1] D. Desalme, P. Binet, N. Bernard, D. Gilbert, M.L. Toussaint, Environ. Exp. Bot. 71, 146 (2011) [2] R. Krasnoschekova, U. Kirso, F. Perin, P. Jacquignon, Polyc. Arom. Comp. 3, 41 (1992) [3] D. Broman, C. Näuf, I. Lundbergh, Y. Zebühr, Environ. Toxicol. Chem. 9, 429 (1990) [4] J.F. McCarthy, L.E. Robertson, L. Burrus, Chemosphere 19, 1911 (1989) [5] L.E. Sverdrup, P.H. Krogh, T. Nielsen, C. Kjær, J. Stenersen, Chemosphere 53, 993 (2003) [6] D. Filgueira, D.P. Salomão de Freitas, A.P. de Souza Votto, G. Fillmann, J.M Monserrat, L.A. Geracitano, G. Santos Trindade, Photochem. Photobiol. 83, 1358 (2007) [7] M.F. Blum, Photodynamic Action and Diseases Caused by Light (Reinhold, New York, 1941) [8] J.L. Newsted, J.P. Giesy, Environ. Toxicol. Chem. 6, 445 (1987) [9] C. Meewes, P. Brenneisen, J. Wenk, L. Kuhr, W. Ma, J. Alikoski, A. Poswig, T. Krieg, K. Scharffetter-Kochanek, Free Radic. Biol. Med. 30, 238 (2001) [10] L. Ke, L. Luo, P. Wang, T. Luan, N. Fung-Yee Tamb, Biores. Technol. 101, 6950 (2010) [11] J. Kou, H. Zhang, Y. Yuan, Y. Li Wang, T. Yu, Z. Zou, J. Physic. Chem. C 112, 4291 (2008) [12] B.F. Smets, B.E. Rittmann, Water Res. 24, 355 (1989) [13] Y. Zhang, Y.X. Zhu, K.K. Kwon, J.H. Park, S.J. Kim, Chemosphere 55, 389 (2004) [14] J.W. Cook, J. Chem. Soc., 1472 (1932) [15] A.L. Juhasz, R. Naidu, Int. Biodeter. Biodegr. 45, 57 (2000) [16] C.E. Cerniglia, Adv. Appl. Microbiol. 30, 31 (1984) [17] D.T. Gibson, V. Subramanian, Appl. Environ. Microbiol. 61, 3711 (1995) [18] I.B. Weinstein, A.M. Jeffrey, K.W. Jennette, S.H. Blobstein, Science 193, 592 (1976) 1142

[19] P. Sims, P.L. Grover, A. Swaisland, K. Pal, A. Hewer, Nature 252, 326 (1974) [20] R.E. Albert, M.L. Miller, T. Cody, A. Andringa, R. Shukla, C.S. Baxter, Carcinogenesis 12, 1273 (1991) [21] C.-L. Lee, H.-T. Huang, L.-J. Kuo, Environ. Sci. Health. A 35, 515 (2000) [22] K.T. Semple, R.B. Cain, S. Schmidt, FEMS Microbiol. Lett. 170, 291 (1990) [23] Y. Nassiri, T. Ginsburger-Vogel, J.L. Mansot, J. Wry, Biol. Cell. 86, 151 (1996) [24] L. Wang, B. Zheng, W. Meng, Ecotoxicol. Environ. Safety 71, 465 (2008) [25] X.D. Huang, D.G. Dixon, B.M. Greenberg, Ecotox. Environ. Safe. 32(2), 194 (1995) [26] G. Barančíková, D. Fazekašová, P. Manko, S. Torma, Chémia životného prostredia (Prešovská univerzita v Prešove, Univerzitná knižnica, Prešov, Slovakia, 2009) (in Slovak) ISBN 978-80-555-00829 [27] I. Holoubek, Crit. Rev. Anal. Chem. 29(3), 179 (1999) [28] H. Peng, H. Yin, J. Deng, J.-S. Ye, S.-N. Chen, B.-Y. He, N. Zhang, Pedosphere 22(4), 554 (2012) [29] S. Lundstedt, Analysis of PAHs and their transformation products in contaminated soil and remedial processes, Umeå: Solfjädern Offset AB, 56 (2003) [30] S. Mujtaba, A. Dwivedi, M.K.R. Mudiam, D. Ali, N. Yadav, R.S. Ray, Photochem. Photobiol. 87, 1067 (2011) [31] A. Mallakin, D.G. Dixon, B.M. Greenberg, Chemosphere 40, 1435 (2000) [32] J.J. Yin, Q. Xia, S.H. Cherng, W.I. Tang, P.P. Gu, G. Lin, H. Yu, D.H. Saen, Int. J. Environ. Res. Public. Health 5 1, 26 (2008) [33] Y. Ibuki, R. Goto, Environ. Toxicol. Pharm. 11, 101 (2002)


[34] S. Goldstein, G. Czapski, Int. J. Radiat. Biol. 46, 725 (1984) [35] J. Regensburger, K. Lehner, T. Maisch, R. Vasold, F. Santarelli, E. Engel, A. Gollmer, B. Konini, M. Landthaler, W. Baumler, Exp. Der-matol. 19(8), 275 (2010) [36] D. Warshawsky, M. Radike, K. Jayasimhulu, T. Cody, Biochem. Biophys. Res. Commun. 152, 540 (1988) [37] A. Aksu, Process Biochem. 40, 997 (2005) [38] B. Bhushan, S.K. Samanta, A. Chauhan, A.K. Chakraborti, R.K. Jain, Biochem. Biophys. Res. Commun. 275, 129 (2000) [39] A. Chauhan, A.K. Chakraborti, R.K. Jain, Biochem. Biophys. Res. Commun. 270, 733 (2000) [40] D. Dean-Ross, J. Moody, C.E. Cerniglia, FEMS Microbiol. Ecol. 41, 1 (2002) [41] S.K. Samanta, A.K. Chakraborti, R.K. Jain, Appl. Microbiol. Biotechnol. 53, 98 (1999) [42] Z. Aksu, D. Akpinar, Biochem. Eng. J. 7, 183 (2001) [43] S.M.N. Chan, T. Luan, M.H. Wong, N.F.Y. Tam, Environ. Toxicol. Chem. 25, 1772 (2006) [44] A.P. Lei, Y.S. Wong, N.F.Y. Tam, Water Sci. Technol. 46, 195 (2002) [45] A.P. Lei, Z.L. Hu, Y.S. Wong, N.F.Y. Tam, Biores. Technol. 98, 273 (2007) [46] D. Kupka, P. Sekula, O. Tischler, J. Briančin, Acta Montan. Slovaca 11, 314 (2006) [47] S.C. Wilson, K.C. Jones, Environ. Poll. 88, 229 (1993) [48] U. Kirso, N. Irha, Ecotox. Environ. Safe. 41, 83 (1998)

[49] M. Gorria, X. Tekpli, M. Rissel, O. Sergent, L. Huc, N. Landvik, O. Fardel, M.T.D. Boitrel, J.A. Holme, D.L. Gossmann, Toxicol. Appl. Pharmacol. 228, 212 (2008) [50] A. Bednáriková, B. Skláršová, E. Kolek, P. Šimko, Chem. Listy 104, 537 (2010) [51] C.E. Cerniglia, Biodegradation 3, 351 (1992) [52] X.D. Huang, B.J. McConkey, T.S. Babu, B.M. Greenberg, Environ. Toxicol. Chem. 16, 1707 (1997) [53] K. Wakabayashi, P. Boger, Weed. Biol. Manage. 4, 8 (2004) [54] M. Kummerova, J. Krulova, Š. Zezulka, J. Třiska, Chemosphere 65, 489 (2006) [55] M. Kummerova, L. Vaňova, H. Fišerova, M. Klemš, Š. Zezulka, J. Krulova, Plant. Growth Regul. 61, 161 (2010) [56] A. Aksmann, T. Shutova, G. Samuelsson, Z. Tukaj, Aquat. Toxicol. 104, 205 (2011) [57] I. Oguntimehin, A.F. Eissa, H. Sakugawa, Chemosphere 78, 877 (2010) [58] H.K. Lichtenthaler, J. Plant. Physiol. 148, 4 (1996) [59] K. Roháček, M. Barták, Photosynthetica 37, 339 (1999) [60] C.S. French, In: W. Ruhland (Ed.), The chlorophylls in vivo and in vitro (Berlin, Springer, 1960) [61] STN ISO 10260: Kvalita vody. Meranie biochemických parametrov. Spektrofotometrické stanovenie koncentrácie chlorofylu a. Technical standard (Slovakia, 1999) (in Slovak)

1143