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TECHNICAL REPORTS

Organic Compounds in the Environment

Ecotoxic Effect of Phenanthrene on Nitrifying Bacteria in Soils of Different Properties

^a Institute of Soil Science and Plant Cultivation- State Research Institute, ul. Czartoryskich 8, 24-100 Pulawy, Poland
^b Lithuanian Institute of Agriculture, Dotnuva, 5051 Kedainiai District, Lithuania

^* Corresponding author (bkord{at}iung.pulawy.pl).

Received for publication March 6, 2007.

	ABSTRACT

TOP NOTES ABSTRACT INTRODUCTION Materials and Methods Results Discussion Conclusions REFERENCES

 
Information on ecotoxicity of organic contaminants, such as polycyclic aromatic hydrocarbons (PAHs), in terrestrial environment is needed for establishing soil quality criteria and for risk assessment purposes. An ecotoxic effect of a model PAH compound (phenanthrene) toward soils microorganisms (nitrifying bacteria) was evaluated in 50 different soils. The soil samples were collected from agricultural land in four regions of Poland with varying levels of industrialization (Slaskie, Dolnoslaskie, Podlaskie, and Lubelskie voievodeships). Soils were characterized for basic physicochemical properties (texture, organic matter content, pH_KCl, total nitrogen content, total sorption capacity) and the content of contaminants including PAHs (73–800 µg kg^–1), Pb (6–720 mg kg^–1), and Zn (9–667 mg kg^–1). Ecotoxicity of phenanthrene (applied at 10, 100, 500, and 1000 mg kg^–1) to soils microorganisms was evaluated in laboratory studies in control conditions (incubation of soils for 7 d at 20 ± 2°C). Nitrification potential was used as the ecotoxicity measurements end point. The EC₅₀ values (146–1670 mg kg^–1) calculated from the square root–X linear regression model differed significantly in various soils, although it was difficult to establish a causative relationship between soil physicochemical characteristic and phenanthrene toxicity. A significant factor in the assessment of soils vulnerability to the effect of phenanthrene was level of soil contamination, particularly with PAHs. Soils with previous contamination were more susceptible (mean EC₅₀, 325 mg kg^–1) than soils from uncontaminated, rural areas (mean EC₅₀, 603 mg kg^–1).

Abbreviations: AMO, ammonia monooxygenase • NP, nitrification potential • PAH, polycyclic aromatic hydrocarbon • Phe, phenanthrene • RSD, relative standard deviation • WHC, water holding capacity

	INTRODUCTION

TOP NOTES ABSTRACT INTRODUCTION Materials and Methods Results Discussion Conclusions REFERENCES

 
PROTECTING habitat function of soil is one of the key goals of the Thematic Strategy for Soil Protection recently adopted by the EU Parliament (Commission of the Council, European Parliament, European Economic and Social Committee, and Committee of the Regions, 2006). The goal of protecting the soil as a habitat for plants, invertebrates, micro-organisms, and their interactions secures the sustainable use of soil and execution of its basic functions (International Organization for Standardization, 2003). Contamination, specified as one of the eight main threats to soil ecosystem (European Parliament and Commission of the Council, 2006), can create direct and indirect risks for soil habitat function.

Soil micro-organisms, being in intimate contact with the soil environment, are considered the best indicators of soil pollution because they are responsive to low concentration of contaminants and react rapidly to soil perturbation (Paul and Clark, 1996; Torstensson, 1997; Ratte et al., 2003; Loibner et al., 2004). Nitrifying bacteria are one microbial group that is most sensitive not only to soil environment conditions depending on temperature, moisture, pH, and the substrates (NH₄–N, O₂, CO₂) content (Montagnini et al., 1989), but also to soil contamination (Torstensson, 1997). Nitrification involves the conversion of ammonium (NH₄⁺) to nitrite (NO₂^–), which is further oxidized to nitrate (NO₃^–). These two steps are performed in soils by autotrophic NH₄⁺–oxidizing bacteria (from Nitrosospira and Nitrosomonas genera) and NO₂^––oxidizing bacteria (mainly by Nitrobacter), respectively (Van Beelen and Doelman, 1997; Boer and Kowalczuk, 2001; Smolders et al., 2001). The key enzyme for chemolithotrophic NH₄⁺ oxidation is ammonia mono-oxygenaze converting NH₄⁺ to hydroxylamine, which is further oxidized to NO₂^– by hydroxylamine oxireductase (Boer and Kowalczuk, 2001). This process is considered to be a rate-limiting step in autotrophic nitrification and is often applied in toxicity tests (Torstensson, 1997; Van Beelen and Doelman, 1997; Niewolak and Koziello, 1998; Smolders et al., 2001, Sverdrup, 2001; Smolders et al., 2003; International Organization for Standardization, 2003, 2004; Klimkowicz-Pawlas, 2005; Sverdrup et al., 2006). There is also a wide range of bacteria (e.g., Pseudomonas putida, Paracoccus denitrificants, and Thiosphaera pantotropha) and fungi (e.g., Aspergillus flavus, Penicillium sp., and Cephalosporium sp.) with the potential for heterothrophic nitrification; however, they are seldom used in ecotoxicity testing (Boer and Kowalczuk, 2001).

Soil properties play a crucial role in the evaluation of all processes determining the fate and transport of contaminants introduced to soil (Van Beelen and Doelman, 1997; Chung and Alexander, 2002; Ratte et al., 2003; Römbke and Amorim, 2004; Maliszewska-Kordybach, 2005). Understanding the relationship between the ecotoxicity of contaminants in the terrestrial environment and the soil characteristics (including the presence of other contaminating substances) is essential for further progress in the protection of soil ecosystems and may have serious implications for legislative activities in developing ecologically sound soil quality criteria (Kördel and Römbke, 2001; Breure et al., 2005). However, although data regarding ecotoxic effects of heavy metals and some pesticides in soils are relatively more accessible (Van Beelen and Doelman, 1997; Torstensson, 1997; Smolders et al., 2001; Smolders et al., 2003; Shen et al., 2006; Sverdrup et al., 2006), information regarding persistent organic pollutants, such as polycyclic aromatic hydrocarbons (PAHs), is limited. A comprehensive literature review by the USEPA revealed that it was impossible to set up ecological soil screening levels for PAHs due to a lack of information fulfilling USEPA scientific criteria (Kapustka, 2004).

The main objective of this study was to evaluate the ecotoxic effect of PAH in soils with different physicochemical properties and levels of contamination. Phenanthrene (Phe), often used in PAH ecotoxicity studies (Sverdrup, 2001; Loibner et al., 2004; Kapustka, 2004; Klimkowicz-Pawlas, 2005; Smreczak et al., 2005), was chosen as model PAH compound. Nitrification potential (NP), which is the nitrification activity observed immediately after adding substrate in the form of NH₄⁺ ions (Smolders et al., 2001; International Organization for Standardization, 2004; Sverdrup et al., 2006) was applied as an ecotoxicity end-point.

	Materials and Methods

TOP NOTES ABSTRACT INTRODUCTION Materials and Methods Results Discussion Conclusions REFERENCES

 
Soils
Sampling and Storage
Soil sampling sites were selected to represent regions with various levels of industrialization and urbanization within Poland. Fifty soil samples were collected from different regions positioned in four administration districts (voievodeships). Region P in Podlaskie voievodeship (10 sampling points) is a typical rural and recreation area, remote from any pollution sources. Region L (18 points) represents a more urbanized Pulawy district located in agricultural Lubelskie voievodeship. Two other regions in Dolnoslaskie (DL, seven points) and Slaskie (S, 15 points) voievodeships represent areas of high urbanization and industrialization. Sampling sites in Slaskie region were situated in the proximity of Tarnowskie Gory, which is a historical metal mining area with industrial activity dating as early as the 12th century.

Geographical locations of sampling points were identified by global positioning system (Table 1 ). Soil samples (20 kg) were taken from the surface layer (0–30 cm) of agricultural land (mostly arable fields). After transport to the laboratory, soil material was air dried at 20°C, well mixed, sieved to pass a 2-mm sieve-mesh, and stored for no longer than 6 mo in the dark at 12 to 16°C before soil physicochemical characteristic and ecotoxicity experiments.

Determination of Soil Physicochemical Properties
The soils were characterized for soil organic matter content, pH, extractable acidity, base saturation, total sorption capacity, and particle size distribution. Soil organic matter content was determined by loss at ignition (550^oC, 7 h). The pH was determined potentiometrically in 1:2.5 (m/V) suspension of soil in 1 mol L^–1 KCl solution (ISO 10390; International Organization for Standardization, 1994). Extractable acidity was measured by extraction with 0.5 mol L^–1 (CH₃COO)₂Ca at pH 8.2 (Polish Committee for Standardization, 1997), whereas base saturation was determined using 1 mol L^–1 solution of CH₃COO(NH₄)₂ for K⁺, Na⁺, Ca²⁺, and Mg²⁺ cation extraction. The total sorption capacity (T) was calculated as the sum of extractable acidity and base saturation. Total nitrogen content (N_t) was determined by Kjedahl method (ISO 11261; International Organization for Standardization, 1995a). Soil particle size distribution was established by an aerometric method (Polish Committee for Standardization, 1998a, 1998b).

Determination of PAH and Heavy Metals Content in Soils
Sixteen PAH compounds (USEPA list) were determined by extraction with dichloromethane. Twenty-five grams of soil were extracted with 125 mL of CH₂Cl₂ in Soxtec apparatus (Bchi Universal Extraction System B-811; Bchi Labortechnik AG, Postfach, Switzerland) for 5 h (seven cycles per hour) at 40.7°C. Extracts were concentrated to a volume of 1 mL on a rotary evaporator (Bchi Rotavapor R-200; Bchi Labortechnik AG). Concentrated extracts were cleaned-up on glass mini-columns (0.5 x 20 cm) filled with 1 g of silica gel (conditioned at 135°C for 16 h) suspended in dichloromethane. Polycyclic aromatic hydrocarbons were eluted with 5 mL of a mixture of CH₂Cl₂/n-hexane (2:3 v/v). The eluate was evaporated to a volume of approximately 2 mL before analysis using Agilent 6890N gas chromatograph (Agilent Technologies, Inc., Wilmington, DE) equipped with Agilent 5973 Network mass spectrometer (Agilent Technologies, Inc., Palo Alto, CA) (70 eV) and 7683 B series autosampler (Agilent Technologies, Inc.). Resolution of PAH compounds has been achieved according to ISO 18287 standard (International Organization for Standardization, 2006a) with a DB-5 mass spectrometer fused-silica capillary column 30 m x 0.25 mm ID with film thickness of 0.25 µm and with guarded column (J&W Scientific, Folsom, CA). Helium was used as a carrier gas (constant flow of 30 cm s^–1) with a splitless injection system at 250°C. The GC oven was programmed as follows: 60°C for 2 min, followed by a 30°C min^–1 ramp to 120°C and a with ramp of 5°C min^–1 to final temperature of 290°C (10 min hold). Mass spectrometer detection was based on selected ion monitoring system. The precision of the method corresponding to the mean relative standard deviation (RSD) was in the range of 2 to 24% for individual PAH compounds and 8% for sum of 16 PAH compounds. A blank sample was carried through all procedures. Quality control included analysis of a reference soil sample every 20 samples (soil no. 701 from SETOC program, 1992–1995). The mean recovery calculated for 16 PAHs in the reference soil was 71%; with recovery for individual compounds ranged from 53 to 112%.

Total Zn and Pb levels were determined by the aqua regia digestion procedure (2 h hot digestion in 1:3 v/v mixture of concentrated nitric and hydrochloric acids followed by refluxing in 3 M hydrochloric acid). Filtrates were analyzed using atomic absorption spectroscopy apparatus (ASA 403; PerkinElmer, Norwalk, CT). The pretreatment and determination procedures from ISO 11466 (International Organization for Standardization, 1995b) and ISO 11047 (International Organization for Standardization, 1998) were applied. Quality control included duplication of every 10th sample and analysis of soil reference materials (NIST 2709 and NIST 2710) every 20 samples. Precision of the method defined as percentage RSD was <1.5% for both analyzed metals.

Method of Experiment
PAH Compound
Phenanthrene (Aldrich, Cat. P1142–5), a three-ring hydrocarbon (water solubility, S_w = 1.3 mg L^–1; octanol/water partition coefficient Log K_ow = 4.57), was used as a model PAH compound. A stock solution of Phe, prepared by dissolving 50 g of Phe in 1000 mL of dichloromethane, was stored in the dark at room temperature and diluted further with CH₂Cl₂ accordingly. The range of Phe contamination (10–1000 mg kg^–1 of dry soil) was chosen on the basis of earlier experiments (Klimkowicz-Pawlas, 2005) that indicate that this range is suitable for the evaluation of EC₅₀ values (in relation to nitrifying bacteria activity). The applied levels also correspond to the content of PAHs in highly contaminated soils from industrial sites (Maliszewska-Kordybach, 1999).

Soil Contamination with Phe and Incubation
Subsamples (20 ± 0.1 g) of dry soil material were placed in glass beakers and spiked with 2 mL of CH₂Cl₂ solution of Phe at the concentrations giving the level of soil contamination with Phe corresponding to 0, 10, 100, 500, and 1000 mg kg^–1 of dry soil. After careful mixing, each soil subsample was left overnight in darkness to allow the solvent to evaporate. After 24 h, the samples were supplemented with 80 ± 0.1 g of soil (bringing the total weight to 100 ± 0.2 g), thoroughly mixed and moistened with deionized water to 60% water holding capacity (WHC).

Preliminary experiments show that 2 mL of dichloromethane introduced to 20 g (+80 g) of dry soil did not significantly affect ( 0.01) soil NP as compared with water treatment (data not shown). Therefore, soils contaminated with pure dichloromethane were used for control samples (zero treatment) as recommended in ISO 15685 standard (International Organization for Standardization, 2004). The method of gradual soil contamination (introduction of solution to small amount of soil and then dilution with noncontaminated soil) was recommended by other studies (Northcott and Jones, 2000; Brinch et al., 2002) as an appropriate method for toxicity tests with poorly water soluble substances.

All samples were incubated in the dark for 7 d at 20 ± 2°C. This period was established experimentally in earlier studies (Klimkowicz-Pawlas, 2005) on the basis of NP kinetic experiments as an optimum incubation time for laboratory experiments with autochthonous populations of soil nitrifying bacteria. Soil moisture content was kept constant by periodically weighing the samples and adding water as necessary. All combinations were done in replicates of two. After 7 d, the soil samples were mixed thoroughly, and two subsamples (25 ± 0.1 g wet weight) were taken from the each replicate for nitrification potential determinations.

Determination of the Nitrifying Bacteria Activity in Soils
Nitrification potential determinations were performed according to ISO 15685 method (International Organization for Standardization, 2004) with some modification aimed at increasing the test sensitivity. The alterations to the ISO method included adding a fixed amount of test medium (100 mL), longer incubation time (24 vs. 6 h as recommended in ISO 15685 standard) at lower temperature (20 vs. 25°C), larger volume of soil slurry sampled after shaking (5 vs. 2 mL), and a larger volume of subsample after filtration (5 vs.1 mL).

Moist (60% WHC) soil subsamples (25 ± 0.1 g) were placed in a 250-mL glass flask and mixed with 100 mL of test medium to form a slurry. The test medium contained 1.5 mM of (NH₄)₂SO₄ as a substrate, 1 mM of potassium phosphate buffer (KH₂PO₄ and K₂HPO₄), and 5.625 mM of NaClO₃ (to prevent further oxidation of NO₂^– to NO₃^–). The pH of the medium was approximately 7.2. The slurries were incubated for 24 h on a shaker at approximately 175 rpm at 20 ± 2°C. After incubation, 5 mL of the soil slurry was transferred to 25-mL glass beakers, and 5 mL of 4 M KCl was added to stop the NH₄⁺ oxidation. The suspension was filtered using 390-grade filter paper (Munktell & Filtrak GmbH; Bärenstein Paper Mill, Bärenstein, Germany), and the filtrate was collected in 25-mL glass beakers. After filtration, 5-mL subsample of the filtrate was transferred to glass flask containing 0.2 mL of the color reagent including sulphanilamide (C₆H₈N₂O₂S) and N-(1-naphtyl)ethylene diamine dihydrochloride (C₁₂H₁₆N₂Cl₂) and filled with deionized water to bring the total volume to 25 mL. After 60 to 90 min, the intensity of the purple color was measured on Beckman DU-68 spectrophotometer (Beckman Instruments, Inc., Fullerton, CA) at a wavelength of 543 nm. A blank was performed through all procedures. All determinations were done in duplicates for each of the replicate soil samples, and the results for individual samples were expressed as an arithmetic mean (µg NO₂^– g^–1) of two measurements adjusted to soil dry matter (105°C). The final results were given as an arithmetic mean of four measurements (2 replicates x 2 NP determinations). For ANOVA evaluations, the mean values for each of the replicates (n = 2) were used. Precision of the method (as established on the basis of the RSD values for n = 10) was about 5 to 8% for one soil sample and 5 to 15% for replicates of 10 samples (n = 20).

The results were presented as relative effect of Phe on NP in soil, expressed as percent of control response (control = 100%) (i.e., RE-NP10, RE-NP100, RE-NP500, and RE-NP1000 values corresponding to levels of soil contamination with Phe of 10, 100, 500, and 1000 mg kg^–1, respectively).

Statistic
Ecotoxicity of Phe was described by EC₅₀ (concentration of Phe in soil causing 50% inhibition of NP during 7-d incubation period). EC₅₀ was expressed in milligrams of Phe per kilogram of dry soil. EC₅₀ values were calculated on the basis of the best-fit linear regression models assessing the concentration–effect relationship.

All data were assessed for normality using standardized skewness and standardized kurtosis parameters. In the case of individual soils, the effect data (RE-NP) used for EC₅₀ calculations were characterized by normal distribution. Data describing all soils set (soils properties, soils ecotoxicity parameters) were non-normal and were log transformed before further evaluation. The exceptions were sets of RE-NP500 and RE-NP1000 parameters, which exhibited normal distribution only after square root transformations. Pearson product moment correlations were used for evaluation of the relationship between pairs of variables. One-way ANOVA was applied for the evaluation of the effect of tested factors on soil characteristics and ecotoxicity parameters (Tukey HSD method). Before ANOVA was performed, the variances check (Bartlett's test appropriate for equal and unequal group sizes) was done to examine if the samples were from the same populations. Multiple regression (forward stepwise procedure) was applied for evaluation of the impact of soil parameters on ecotoxic data. The Statgraphics Centurion version XV program was used for statistical evaluation.

	Results

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The studied soils exhibited properties typical for Poland (Table 2 ): They were mostly light, rather acidic soils (median content of fraction < 0.02 mm, 23%; median pH, 5.3) characterized by low organic matter content (median C_org, 1.92%; interquartile range 1.62–2.23%). Nearly 45% of samples represented sands, another 45% represented loams, and 10% represented silts. In spite of various geographical locations of sampling points, basic properties (fraction < 0.02 mm, C_org, pH, and C/N) were relatively uniform, with CVs ranging from 21 to 52%. A higher variability characterized soil contamination (CV values for 16PAHs, Zn, and Pb were 90, 157, and 176%, respectively).

View larger version (42K): [in this window] [in a new window]   Fig. 1. Comparison of the relative nitrification potential (RE-NP) in 50 soils contaminated with phenanthrene at the levels of 100 and 1000 mg kg–1. Error bars represent SD calculated for two replications (n = 2). Indistinguishable error bars correspond to SD < 0.1%.

View larger version (13K): [in this window] [in a new window]   Fig. 2. An example of the calculations used to determine the EC50 values for selected soils on the basis of the square root-x regression model. RE-NP, relative nitrification potential; SQRT, square root.

Analysis of correlations between pairs of variables characterizing soil properties and ecotoxicity parameters indicated that the 16PAH is the only soil property that is significantly ( 0.05) correlated with RE-NP500 (r = –0.31), RE-NP1000 (r = –0.33), EC₅₀ (r = –0.35), and EC₂₀ (r = –0.28) values. An attempt to apply the multiple regression for evaluation of the dependency of ecotoxicity parameters on soils characteristics was not successful because most of the soil properties were statistically cross-correlated. However, the regression model selection procedure indicates that the highest R² values are related primarily to soil parameters that describe contamination level (e.g., PAH) rather than to soil organic carbon content and total sorption capacity.

The observed relationships suggested the possibility of the influence of soil origin on its resilience to the effect of Phe. Thus, calculated EC₅₀ parameters were checked (by one-way ANOVA) in relation to sampling regions (Table 5 ). The mean EC₅₀ value in Podlaskie voievodeship (603 mg kg^–1) was significantly ( 0.05) higher than in the Dolnoslaskie (348 mg kg^–1) and Slaskie (325 mg kg^–1) regions. Soils from Podlaskie were also characterized by the significantly lower PAH content (16PAH of 113 µg kg^–1 as compared with 319 and 672 µg kg^–1 in Dolnoslaskie and Slaskie, respectively). In the case of heavy metals, higher concentrations were observed in the Slaskie region (e.g., Zn of 183 mg kg^–1 as compared with 22 mg kg^–1 in Podlaskie). The other differences between the Podlaskie and Slaskie regions were related to the content of the fraction < 0.02 mm (15 and 21%, respectively) and pH (4.8 vs. 5.6). It was impossible to evaluate the differences in the content of C_org in the selected regions due to differences in populations from which the samples were derived (negative variance check with Barlett's test).

	Discussion

TOP NOTES ABSTRACT INTRODUCTION Materials and Methods Results Discussion Conclusions REFERENCES

Common parameters used in ecotoxicity studies include NOEC (no observed effect concentration), LOEC (lowest observed effect concentration), and EC_x (x = % effect concentration) values. The two first are not recommended (van Beelen and Doelman, 1997; Sparks, 2000; Organization for Economic Co-Operation and Development, 2003) due to their high dependence on the concentrations applied in the experiments (only a couple data points are used instead of full data set). The value of EC_x calculated from appropriate regression models is considered to be the most objective and therefore was applied in this study (Sparks, 2000; Organization for Economic Co-Operation and Development, 2003; Environment Canada, 2005). Choosing a suitable regression model for the estimation of EC₅₀ values is of utmost importance for the ecotoxicity results assessment (Sparks, 2000). Nevertheless, the selection is largely arbitrary because the mathematical expression has no biological meaning (Organization for Economic Co-Operation and Development, 2003; Scott-Fordsman, personal communication, 2006). Meister and van den Brink (Sparks, 2000) suggest that the model should be "as simple as possible and as complicated as necessary." In the case of 50 different soils applied in this study, choosing the "model of best fit" was a difficult task. The square root-X linear regression model, selected from nine other tested options, described the data well: The observed effects (NP measurements) were determined in 89% by the variability of Phe concentrations. The median Phe-EC₅₀ value was 375 mg kg^–1, and the SD corresponded to 363 mg kg^–1. There are limited data in the literature on ecotoxicity of Phe to terrestrial receptors, especially to micro-organisms (Kapustka, 2004). In the work of Sverdrup (2001), the EC₅₀ value for Phe (nitrification end-point; one soil; maximum applied Phe concentration, 3000 mg kg^–1) corresponded to 250 mg kg^–1 (95% confidence interval, 220–380 mg kg^–1). Klimkowicz-Pawlas (2005) reported that the EC₅₀ for the effect of Phe on nitrification potential in four different soils was within the range of 6 to 412 mg kg^–1 (7 d incubation; maximum applied concentration, 500 mg kg^–1; application of simple linear regression model). The above data are comparable to our results. Similar ranges of Phe ecotoxicity parameters were reported in studies on invertebrates (Kapustka, 2004). Higher EC₅₀ values were noted in phytotoxicity tests with wheat and rape growing on three different soils; in all cases, the EC₅₀ values exceeded the highest level of Phe application of 1000 mg kg^–1 (Klimkowicz-Pawlas, 2005). It has to be noted that organic pollutants like PAHs introduced to soils may undergo different processes, leading to a decrease of their content and bioavailability and thus their ecotoxicity (Hatzinger and Alexander, 1995; Alexander and Alexander, 2000; Sverdrup, 2001; Hwang and Cutright, 2002; Harmsen, 2004). This aspect was not included in the present study due to the short experimental period; however, the problem was addressed in our earlier works (Maliszewska-Kordybach and Smreczak, 2003a; Maliszewska-Kordybach, 2005; Smreczak et al., 2005).

In spite of our presumptions, it was difficult to find the clear association between soil characteristic and Phe ecotoxicity. None of the individual physicochemical soil descriptors (C_org, pH, T, N_t, fraction ø < 0.02 mm, or fraction ø < 0.02 mm) affected significantly its resilience to the toxicity of Phe. An attempt to evaluate the combined effect of soil characteristics (application of multifactor regression) for the description of Phe ecotoxicity was not successful because significant cross-correlation of many soil parameters impeded the application of this method. Generally, ecotoxic effects of contaminants are related to the substance and organism features (Ratte et al., 2003; Harmsen, 2004). The substance aspect is described mainly by its "environmental availability," which is linked to soil characteristics controlling sorption/desorption and sequestration processes (Hatzinger and Alexander, 1995; Kördel and Römbke, 2001; Harmsen, 2004; International Organization for Standardization, 2006b). The problems in the identification of associations between soil parameters and the Phe ecotoxicity suggest complicated soil–contaminant relationships. A similar problem occurred in evaluating the effect of soil properties on PAH dissemination in soils (Maliszewska-Kordybach, 2005). More extensive characteristics of soil material (e.g., determination of different nitrogen forms) or increased variability of soils (e.g., incorporation of samples from "marginal soils") may help in further studies in this area.

The biological characterization of all investigated soils (NP determinations) included the same group of organisms: nitrification bacteria active in the oxidation of NH₄⁺ to NO₂^– (Boer and Kowalczuk, 2001). However, the applied ecotoxicity test (ISO 15685; International Organization for Standardization, 2004) does not include identification of specific microbial species and kinetic evaluations. Hence, it is possible that observed differences in the resilience of soils to the toxic effect of Phe can be ascribed to other characteristics of soil microbial community and microbial processes than those investigated in this study. Transformation of hydrocarbons through the action of ammonia mono-oxygenase (AMO) may undergo competitive co-oxidation, which reduces the rate and extent of ammonia oxidation (Deni and Penninckx, 1999). Chung and Alexander (2002) showed that nitrifying bacteria like Nitrosomonas europaea exhibit the ability to transform co-metabolically two- and three-ringed PAHs (like Phe) using ammonia as the energy source. The authors suggest that one of the reasons for the inhibition of NO₂^– production by N. europea in soils contaminated with PAHs is that the alternative AMO substrate, such as lower-molecular-weight PAHs, excludes the binding and oxidation of ammonia at the active site of AMO. Those transformations may proceed differently in soils varying in their properties (Deni and Penninckx, 1999). Further information on the mechanisms of nitrification in hydrocarbon-contaminated soils is given in Keener and Arp (1993; 1994), Chung and Alexander (2002), and Deni and Penninckx (1999; 2004).

An important factor that affected soil resistance to the effect of Phe was the level of soil contamination with PAHs. The significant differences in the content of those pollutants in the regions under study corresponded to the variability of ecotoxicity parameters (Table 5). The highest PAH content was noted in the historically contaminated Slaski region. Simultaneously, the same region was characterized by nearly two times lower Phe-EC₅₀ values than those in the uncontaminated Podlaskie region. The differences in the physicochemical properties of soils (e.g., pH or fraction ø < 0.02 mm) were not related to the changes of EC₅₀ (Table 5). Moreover, Phe, characterized by good water solubility, was applied in this study in relatively high concentrations, and its contact time with soils of high water saturation (60% WHC) was short (7 d). This suggests that environmental availability of Phe in all studied soils was not a limiting factor. Thus, the noted variations that the sensitivities of soils from contaminated and uncontaminated regions to the impact of Phe were initiated by biological factors, although the nitrification potentials in soils from both sampling areas did not differ (the average NP for Slaskie was 1.22 ± 1.06 µg NO₂^– g^–1 vs. 1.12 ± 0.60 µg NO₂^– g^–1 for Podlaskie). The observed higher vulnerability of soils from polluted regions to the effect of Phe may be explained by qualitative changes in the populations of nitrifying bacteria taking part in the consecutive stages of ammonia oxidation and thus to alterations in the overall image of their sensitivity to Phe (Deni and Penninckx, 1999). The other suggestion (particularly valid for the Slaski region, which is highly polluted with Zn and Pb) can be increased stress resulting from the synergistic effect of PAHs compounds and heavy metals toward soil microorganisms, as noted in other studies (Gogolev and Wilke, 1997; Maliszewska-Kordybach and Smreczak, 2003b; Shen et al., 2006).

The observed phenomenon is contradictory to opinion on adaptation of some micro-organisms exposed to PAHs for a long time (Klimkowicz-Pawlas and Maliszewska-Kordybach, 2003). Deni and Penninckx (1999) compared the effect of hydrocarbon fuel addition on the activity of nitrifying bacteria in an uncontaminated agricultural soil and soil with a long history of pollution. Although addition of hydrocarbons to uncontaminated soil reduced nitrification, the ammonia-oxidizing bacteria community in contaminated soils was not directly affected by a hydrocarbon addition, possibly because of high affinity for NH₄⁺.

The lack of a direct relationship between the activity of nitrifying bacteria and soil vulnerability to the effect of PAH (as determined by NP) confirms the opinion of other authors (Smolders et al., 2001; Kapustka, 2004; Siebielec et al., 2006) on the low usefulness of microbial parameters (e.g., nitrification) as indicators of adverse effect of chemicals in a contaminated soil environment and in the evaluation of the ecological status of soils.

	Conclusions

TOP NOTES ABSTRACT INTRODUCTION Materials and Methods Results Discussion Conclusions REFERENCES

 
The presented data form the first broad information on the impact of a three-ring PAH compound, Phe, on the very sensitive group of soil microorganisms, nitrification bacteria, in soils of different properties. The results of the study enabled the evaluation of the ecotoxicity parameters (EC₅₀) for 50 soils in the controlled laboratory conditions. Testing the different regression equations permitted us to choose the best fit square root-X regression model as giving the highest determination coefficients for EC₅₀ calculations. Soil properties exhibited a significant effect on Phe toxicity; however, it was impossible to establish a direct relationship between ecotoxicity parameters and soil characteristics. This study indicates that historical soil contamination is an important factor in determining the ecotoxicity of PAHs. Soils from sites with a long history of pollution with heavy metals and PAHs exhibited higher vulnerability to the effect of Phe. Future work will aim at supplementing the presented data set with the results of phytotoxicity tests and ecotoxicity tests with heavy metals.

	NOTES

TOP NOTES ABSTRACT INTRODUCTION Materials and Methods Results Discussion Conclusions REFERENCES

 
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	REFERENCES

TOP NOTES ABSTRACT INTRODUCTION Materials and Methods Results Discussion Conclusions REFERENCES

 

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