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Microbial activities in boreal soils: Biodegradation of organic contaminants at low temperature and ammonia oxidation

Jukka Kurola

Department of Ecological and Environmental Sciences Faculty of Biosciences

and Department of Applied Chemistry and Microbiology

Division of Microbiology University of Helsinki

Academic Dissertation in Environmental Ecology

To be presented, with the permission of the Faculty of Biosciences, University of Helsinki, for public criticism in the Auditorium of Lahti

Science and Business Park, Niemenkatu 73, Lahti, on October 24th, 2006, at 12 noon

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Supervisors: Professor Martin Romantschuk Department of Ecological and Environmental Sciences

University of Helsinki Lahti, Finland

Professor Mirja Salkinoja-Salonen Department of Applied Chemistry and Microbiology University of Helsinki Helsinki, Finland

Reviewers: Docent Kirsten Jørgensen Finnish Environment Institute Helsinki, Finland Docent Aino Smolander Finnish Forest Research Institute Vantaa Research Unit, Finland Opponent: Professor Lars R. Bakken Department of Plant and Environmental Sciences

The Norwegian University of Life Sciences, Aas, Norway

ISBN 952-10-3373-8 (nid.)

ISBN 952-10-3374-6 (PDF)

ISSN 1239-1271

Yliopistopaino

Helsinki 2006

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Table of Contents

ABSTRACT 4

LIST OF ORIGINAL PUBLICATIONS 5

THE AUTHOR’S CONTRIBUTION 5

ABBREVIATIONS 6

1. INTRODUCTION 7

1.1 Biodegradation of organic contaminants in soil 7

1.2 Factors controlling biodegradation of organic contaminants in soil 8

1.3 Aerobic biodegradation of selected organic environmental contaminants 10 1.3.1 Polycyclic aromatic hydrocarbons (PAH) 10 1.3.2 Polychlorinated phenols 13

1.4 Nitrification and autotrophic ammonia-oxidising bacteria in soil 15 1.4.1 Physiology of chemolithotropic ammonia-oxidizing bacteria 16 1.4.2 Phylogeny of chemolithotropic ammonia-oxidizing bacteria 18 1.4.3 Molecular detection methods for ammonia-oxidizing bacteria: a model for studying microbial diversity in soil using ribosomal RNA and functional gene markers 20

2. AIMS OF THIS STUDY 22

3. MATERIALS AND METHODS 22

3.1 Study sites and soil sampling 22

3.2 Experimental methods 23

4. RESULTS AND DISCUSSION 25

4.1 Biodegradation of organic contaminants in boreal surface soils 25

4.2 Temperature dependency of biodegradation of anthropogenic organic contaminants and natural autochthonous organic matter in boreal surface soils 27

4.3 Activity and population size of ammonia-oxidizing bacteria in oily landfarming soil 29

4.4 Molecular characterisation of ammonia-oxidizing bacteria in oily landfarming soil 31

4.5 Application of cation exchange membranes for studying ammonia-oxidizing bacteria in soil 33

5. CONCLUSIONS 34

6. TIIVISTELMÄ 35

7. ACKNOWLEDGMENTS 35

8. REFERENCES 36

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Abstract This thesis deals with the response of biodegradation of selected anthropogenic

organic contaminants and natural autochthonous organic matter to low temperature in boreal surface soils. Furthermore, the thesis describes activity, diversity and population size of autotrophic ammonia-oxidizing bacteria (AOB) in a boreal soil used for landfarming of oil-refinery wastes, and presents a new approach, in which the particular AOB were enriched and cultivated in situ from the landfarming soil onto cation exchange membranes.

This thesis demonstrates that rhizosphere fraction of natural forest humus soil and agricultural clay loam soil from Helsinki Metropolitan area were capable of degrading of low to moderate concentrations (0.2 – 50 µg cm-3) of PCP, phenanthrene and 2,4,5-TCP at temperatures realistic to boreal climate (-2.5 to +15 °C). At the low temperatures, the biodegradation of PCP, phenanthrene and 2,4,5-TCP was more effective (Q10-values from 1.6 to 7.6) in the rhizosphere fraction of the forest soil than in the agricultural soil. Q10-values of endogenous soil respiration (carbon dioxide evolution) and selected hydrolytic enzyme activities (acetate-esterase, butyrate-esterase and β-glucosidase) in acid coniferous forest soil were 1.6 to 2.8 at temperatures from -3 to +30 °C. The results indicated that the temperature dependence of decomposition of natural autochthonous soil organic matter in the studied coniferous forest was only moderate.

The numbers of AOB in the landfarming (sandy clay loam) soil were determined with quantitative polymerase chain reaction (real-time PCR) and with Most Probable Number (MPN) methods, and potential ammonium oxidation activity was measured with the chlorate inhibition technique. The results indicated presence of large and active AOB populations in the heavily oil-contaminated and urea-fertilised landfarming soil. Assessment of the populations of AOB with denaturing gradient gel electrophoresis (DGGE) profiling and sequence analysis of PCR-amplified 16S rRNA genes showed that Nitrosospira-like AOB in clusters 2 and 3 were predominant in the oily landfarming soil. This observation was supported by fluorescence in situ hybridization (FISH) analysis of the AOB grown on the soil-incubated cation-exchange membranes. The results of this thesis expand the suggested importance of Nitrosospira-like AOB in terrestrial environments to include chronically oil-contaminated soils.

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List of original publications This thesis is based on the following articles, which will be referred to in the text by their Roman numerals. I. Kurola J. and Salkinoja-Salonen M. Potential for biodegradation of anthropogenic organic compounds at low temperature in boreal soils. Submitted manuscript. II. Kähkönen M. A., Wittmann C., Kurola J., Ilvesniemi H. and Salkinoja-Salonen M. 2001. Microbial activity of boreal forest soil in a cold climate. Boreal Environment Research 6:19-28. III. Kurola J., Salknoja-Salonen M., Aarnio T., Hultman J. and Romantschuk M. 2005. Activity, diversity and population size of ammonia-oxidising bacteria in an oil-contaminated landfarming soil. FEMS Microbiology Letters 250:33-38.

IV. Kurola J., Wittmann C., Salkinoja-Salonen M., Aarnio T. and Romantschuk M. 2005. Application of cation-exchange membranes for characterisation and imaging ammonia-oxidising bacteria in soils. FEMS Microbiology Ecology 53:463-472.

The author’s contribution Paper I: Jukka Kurola conducted the measurements, analysed and interpreted the results, and wrote the manuscript. He planned the experiments under supervision of M. Salkinoja-Salonen and performed all experimental work. Paper II: Jukka Kurola performed the mineralization experiment with radiolabelled xenobiotics, calculated the results and participated in writing of the paper. He also carried out part of field work.

Paper III: Jukka Kurola designed the experiments, analysed and interpreted the results, wrote the paper, and is the corresponding author. He carried out all experimental work except 16S rRNA gene sequencing and measurement of the soil nutrient and PAH contents.

Paper IV: Jukka Kurola conducted the measurements, and analysed and interpreted the results, wrote the paper, and is the corresponding author. He planned the experiments under supervision of C. Wittmann and performed all experimental work except 16S rRNA gene sequencing, and measurement of the soil nutrient and PAH contents.

The original articles were reproduced with kind permission of the copyright holders.

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Abbreviations AMO ammonia-monooxygenase amo-A A-subunit of ammonia-monooxygenase gene AOB ammonia oxidising bacteria bp base pair DAPI 4’,6-diamidino-2-phenylindole DGGE denaturing gradient gel electrophoresis DNA deoxyribonucleic acid d.w dry weight E eastern longitude Ea activation energy FISH fluorescence in situ hybridization HAO hydroxylamine oxidoreductase MPN most probable number MUF 4-methylumbelliferyl N northern latitude PAH polycyclic aromatic hydrocarbons PCP pentachlorophenol PCR polymerase chain reaction RNA ribonucleic acid 16S rRNA small subunit ribosomal ribonucleic acid T-RFLP terminal restriction fragment length polymorphism 2,4,5-TCP 2,4,5-trichlorophenol

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1. Introduction

1.1 Biodegradation of organic contaminants in soil

In the past century industrial development, urbanization, agriculture practices, and various other types of human activities have led to a vast increase in soil pollution. Consequently, there are numerous anthropogenic contaminants today in soils, which are toxic to biological systems, and represent a substantive threat to human health and environmental quality (Alexander 1981, Philp et al. 2005). The soil ecosystems have, however, a unique intrinsic ability to resist pollution and to naturally attenuate the effects of the toxic anthropogenic chemicals. Among these natural processes the most important attenuation mechanisms are abiotic oxidation, hydrolysis and biodegradation, since they are capable to transform and detoxify the hazardous soil contaminants to less harmful products (Alexander 1999, Gianfreda and Rao 2004).

Contamination of soils by anthropogenic chemicals may result (i) from point-sources, like accidents in chemical production and transportation, spills or leakages during chemical storage, or (ii) from non-point-sources, such as improper use or disposal of toxic material in agriculture. Soil contaminants can be organic; such as pesticides, biocides, petroleum hydrocarbons and chlorinated solvents; or inorganic; such as heavy metals, radionuclides, nitrate and chloride. Many of these anthropogenic compounds are, fortunately, degradable by microorganisms in soil (Alexander 1999, Philp et al. 2005).

Microorganisms, notably bacteria, fungi, archaea, algae and protozoa, are the living component of the soil organic matter, and they are responsible for several key processes in soil (Stevenson and Cole 1999). Soil microorganisms capable of degrading or transforming the organic and the inorganic compounds are classified as chemo-organotrophic and chemolithotrophic organisms, respectively (Atlas and Bartha 1998). Chemo-organotrophic degradation of an organic compound may result in complete mineralization of the parent compound into simple end products (i.e. methane, carbon dioxide, ammonium, chloride and water). On the other hand, the microorganisms may modify the organic compound only partly, leaving its carbon skeleton more or less intact. Such transformation of a harmful anthropogenic organic contaminant may result in detoxification of the parent compound, or generate new molecule(s), that may be more persistent or even more toxic than the original compound (Neilson 1996, Madsen 1997, Alexander 1999). Therefore, complete mineralization of the anthropogenic contaminant is a more desirable outcome rather than mere disappearance of the original compound from the soil.

The intrinsic biodegradative capacities of soil microorganisms have been applied to clean up contaminated soils (Philp et al. 2005). The term soil bioremediation includes the use of microorganisms or plants to detoxify the polluted environment. It relies upon microbial enzymatic activities to transform or to degrade the offending contaminants (Bollag and Bollag 1995). The first initial steps for transformation of the anthropogenic contaminants in soil

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are often catalysed by ecto- or extracellular enzymes. The ecto- and extracellular soil enzymes include a large range of oxidoreductases and hydrolases, which are of microbial, plant or animal origin. The oxidative/reductive and hydrolytic enzymes explicate a degradative function in soil, in which they transform the polymeric substrates into partially degraded or oxidized oligomers or monomers (Tabatabai and Dick 2002, Gianfreda and Rao 2004).

Natural attenuation (or intrinsic bioremediation) is an example of a soil bioremediation method characterized by the absence of any engineering treatment (Romantschuk et al. 2000, Mulligan and Young 2004). By definition natural attenuation includes a variety of physical, chemical or biological processes that, under favourable conditions, act without human intervention to reduce the mass, toxicity, mobility, volume or concentration of the soil contaminant (US EPA 1999). These processes include most of all biodegradation and biotransformation of the offending contaminant.

1.2 Factors controlling biodegradation of organic contaminants in soil

Biodegradation of organic anthropogenic contaminants in soil is, in its general sense, a function of three variables (i) the availability of the compound to degrader microorganisms or enzyme systems, (ii) the quantity of degrader microorganisms or enzymes systems, and (iii) the activity level of these organisms or enzymes (Torstensson and Stenström 1993). Climate is one of the most important parameter affecting to this function in boreal soils.

(i) Soil is a heterogenic environment with respect to its all properties (Verstraete and Top 1999). The chemical and physical properties of soil, such as pH, organic matter content and porosity, show often variable distributions in different soil types. These properties affect strongly to the fate of organic contaminants in soil, and the availability of the contaminants for the degrader microorganisms or enzyme systems. Generally, the organic contaminants elevate the level of soil organic carbon, and may therefore serve as substrates for the degrader microorganisms or be toxic to the degrader microorganisms. Attenuation of the toxicity of organic contaminants, for example by adsorption or volatilization, may enhance biodegradation, whereas low concentrations of the contaminants can be insufficient to generate energy and carbon required for biodegradation (Alexander 1999). If low concentrations of aged anthropogenic organic compounds bind to the soil, they may form a highly persistent residue (bound residue), whose exact structure is difficult or impossible to determine (Stevenson and Cole 1999).

(ii) Microorganisms able of degrading organic anthropogenic compounds are readily isolated by traditional culturing and isolation methods, like viable plate count or most-probable-number (MPN) techniques, from contaminated and uncontaminated soils (Wilson and Jones 1993, McAllister et al. 1996, Whyte et al. 1996, Fulthorpe et al. 1996, Kamagata et al. 1997, Nohynek 1999). Still, it has been known for long time that only 0.1 - 10% of microbial biomass in soil can be induced to grow under laboratory cultivation (Amann et al. 1995, Smith et al. 2001).

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The recent development of culture-independent molecular biological methods for microbial ecology has expanded the potential for characterisation and quantification of soil microorganisms able of degrading organic contaminants (Bakken 1997, Insam 2001, Dua et al. 2002). However, one of the fundamental problems in microbial ecology is the general need to be able to cultivate a particular microorganism of interest in order to investigate its physiological and genetic characters in detail (Liesack et al. 1997). A combination of the traditional and the culture-independent molecular methods would therefore be a valuable tool for assessing the total microbial potential for the biodegradation of organic contaminants in soil.

(iii) Microbial activity in soil is generally controlled by climate conditions (like temperature and moisture), pH, availability of inorganic nutrients, substrates and energy (oxygen) sources, and predation by soil fauna. In temperate and tropical soils, the availability of nutrients, moisture and oxygen are usually the most important limiting factors for biodegradation of anthropogenic compounds, but in boreal and in arctic soils low temperature is commonly assumed to be most significant individual factor, as low temperatures slow down biological activity (Nedwell 1999, Ferguson et al. 2003). The influence of temperature on the rate of biological reaction is commonly expressed by the Arrhenius equation, which can be expressed in its linearized form as follows:

ln k = ln A - (Ea/R) * 1/T

where, k is the rate constant, A is a constant, Ea the activation energy (J mol-

1), R the gas constant (8.3 J K-1 mol-1) and T is temperature (K). By plotting ln k against 1/T the slope (-Ea/R) will be obtained allowing the Ea of the reaction to be calculated. The law of Arrhenius reflects the microbial activities only when the temperature range is not too wide (Gounot 1991, Nedwell 1999).

The Q10-value, which is the quotient of activity increase over temperature change of 10 °C, is also often used to express temperature dependence of microbial activities.

Q10= (Kt+10)/ Kt

where Kt+10 and Kt denote the rate constants at temperatures T+10 and T, respectively.

The importance of microbial activities in soil at low temperatures has been emphasized during the recent years due to the global warming issue (Kirschbaum 1995 and 2000, Schimel and Clein 1996, Goulden et al. 1998, Liski et al. 1999, Mack et al. 2004, Schimel and Mikan 2005). Microorganisms are the main agents producing carbon dioxide and methane, which are also considered as notorious greenhouse gases, during decomposition of natural autochthonous organic material in soil. Because the microbial activity in soil is strongly temperature dependent (Stevenson and Cole 1999), an increase in ambient temperature may lead to an increase in the emissions of carbon dioxide and methane from soil to the atmosphere, which may cause a positive feedback to climate warming.

The soil microorganisms can be classified according their temperature range for growth. Microorganisms with

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an optimal temperature for growth below + 15 °C are termed as psychrophiles, and organisms with growth temperature from 0 °C to + 40 °C as psychrotrophs, respectively. True psychrophilic microorganisms are restricted to permanently cold arctic and polar soils, whereas psychrotrophic organisms are more widespread in boreal soils that undergo thermal fluctuations, and they can be stimulated by an increase of soil organic carbon and temperature (Gounot 1991, Nedwell 1999). Several studies have shown that alpine (Margesin and Schinner 1997 and 2001), Antartic (Aislabie et al. 1998, Coulon et al. 2005), arctic (Whyte et al. 1996, Mohn et al.

1997 and 2000, Eriksson et al. 2001) and boreal (Kähkönen et al. 2001) soils have a potential for psychrotrophic biodegradation of persistent organic pollutants.

1.3 Aerobic biodegradation of selected organic environmental contaminants

1.3.1 Polycyclic aromatic hydrocarbons (PAH)

Polycyclic aromatic hydrocarbons (PAH) are widespread organic contaminants introduced into the environment through both natural and anthropogenic sources. They possess known carcinogenic, mutagenic, and toxic properties (Cerniglia and Heitkamp 1990, Cerniglia 1992 and 1993, Bouldrin et al. 1993, Pothuluri and Cerniglia 1994, Carmichael and Pfaender 1997, Samanta et al. 2002). PAH, which consist of two or more condensed benzene rings in linear, angular or cluster arrangements, are present at various concentrations in coal tar, petroleum and oil based fuels. Thus, they can be found in soils around

gas works and power plants using fossil fuels, and from coke production sites (Wilson and Jones 1993, Wilcke 2000).

Biodegradation of PAH depends on their physical and chemical properties, concentrations and rates of diffusion, as well as their bioavailability (Smith et al. 1997). PAH have low water solubility, a property which in part is causing their persistence in the environment (Cerniglia 1992, Shuttleworth and Cerniglia 1996). Water solubility of PAH decreases with an increase in molecular mass, and therefore the rate of the PAH biodegradation is usually inversely proportional to the number of fused benzene rings in the PAH-molecule (Wilson and Jones 1993, Pothuluri and Cerniglia 1994).

Generally, microbial degradation of organic compounds, like PAH, can be divided in categories according to whether or not the microorganisms derive energy from the degradation processes (Alexander 1999). Metabolic biodegradation is a growth-linked process where chemo-organotrophic microorganisms convert some of the carbon atoms in the organic substrate into their cell constituents, whereas another part of the carbons are degraded to obtain energy. It is known that the low molecular mass PAH (up to three rings) are readily degraded by a number of aerobic chemo-organotrophic bacteria, whereas the high molecular mass PAH (consisting of four or more benzene rings) are usually cometabolically oxidized by a restricted number of bacterial species such as Mycobacterium spp. or Sphingomonas

spp. (Cerniglia 1992, Pothuluri and Cerniglia 1994, Samanta et al. 2002, Bodour et al. 2003). In cometabolic biodegradation process the

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microorganisms are not able to use the organic compound as an energy source, but require an additional carbon source for growth (Alexander 1999).

The microbial degradation of PAHs containing up to four rings has been well documented (Cerniglia 1984, Cerniglia and Heitkamp 1989, Cerniglia 1992, 1993, Parales and Haddock 2004). Fungi, like moulds and white-rot fungi, use cytochrome P-450 monooxygenases to oxidise PAH to arene oxides. The high molecular mass PAHs are also subject to non-specific co-oxidation by radicals produced by the lignolytic enzymes of white-rot fungi (Gianfreda and Rao 2004). Bacteria use dioxygenases to oxidize PAH to cis-dihydrodiol metabolites. This degradation pathway continues with a dehydration to dihydroxy-PAH, which are subject to ring cleavage through different fission pathways, resulting in the formation of organic acids (e.g. succinic, pyruvic, fumaric, or acetic acid), and the further catabolism finally leads to intermediates of the tricarboxylic acid cycle (Wilson and Jones 1993, Pothuluri and Cerniglia 1994, Samanta et al. 2002).

Phenanthrene and pyrene are classified as low molecular weight PAH consisting of three and respectively four

fused benzene rings in angular arrangements. The release of phenanthrene and pyrene into the environment is ubiquitous since both are major products of incomplete combustion of a variety of organic materials including wood and fossil fuels. Bacteria can degrade phenanthrene by two pathways where the initial enzyme attack is in the 1,2- or in the 3,4-positions of the molecule (Figure 1, Cerniglia and Heitkamp 1990, Cerniglia 1992 and 1993, Pothuluri and Cerniglia 1994). The major isomer formed is cis-3,4-dihydroxy-3,4-dihydrophenanthrene, which is subsequently oxidized to 3,4-dihydroxyphenanthrene, and further cleaved and converted to 1-hydroxy-2-naphthoic acid (Figure 1). The product of ring cleavage is then oxidatively decarboxylated to 1,2-dihydroxynaphtalene, metabolised further via the naphtalene pathway to catechol, and finally to the intermediates of the tricarboxylic acid cycle (Pothuluri and Cerniglia 1994). Some bacteria have an alternate pathway for 1-hydroxy-2-naphthoic acid, where it is oxidised through ortho-phthalic acid to form protocatechuic acid (Figure 1).

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Phenanthrene

HOOH H

H

cis-3,4-Dihydroxy-3,4-dihydrophenanthrene

OHHO

3,4-dihydroxyphenanthrene

OH

COOH

1-Hydroxy-2-naphthonic acid

A

1-Hydroxy-2-naphthoic acid

Ring fission

pathways

Catechol

OH

OH

Salicylic acid

COOH

OH

Salicylaldehyde

cis -o-Hydroxy-benzalpyruvic acid

OH

COH

OC

OH

COOH

B

Protocatechuic acid

HO

HO COOH

o -Phthalic acid

COOH

COOH

2-Carboxybenzaldehyde

OOC

COOH

COH

OH

CCOOH

1,2-Dihydroxynaphthalene

OH

OH

OH

COOH

Figure 1. Initial steps in bacterial degradation pathways of phenanthrene (redrawn from Pothuluri and Cerniglia 1994).

Bacterial catabolism of pyrene

by Mycobacterium sp. was first reported by Heitkamp in 1988. Both cis- and trans-4,5-pyrene dihydrodiols occur as ring oxidation products during the degradation of pyrene (Figure 2). The

principal metabolites, 4-phenanthronic acid, 4-hydroxy-perinaphthenone, cinnamic and phthalic acids were identified as ring fission products (Cerniglia and Heitkamp 1990, Pothuluri and Cerniglia 1994).

A

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Monooxygenase

Dioxygenase

HO2

O2

O2

trans-4,5-Pyrene-dihydrodiol

Pyrene-4,5-Oxide

cis-4,5-Pyrene-dihydrodiol

Pyrene

O

OH

OH

OH

OHH

HH

Figure 2. Initial steps in the pathways of pyrene oxidation by Mycobacterium sp (redrawn from Cerniglia and Heitkamp 1990).

1.3.2 Polychlorinated phenols

Chlorinated phenols are widely distributed in the environment as a result of their widespread use as pesticides in agriculture and in synthesis of dyes and pharmaceuticals. Phenol was the first antiseptic compound, but polychlorinated phenols, which possess higher antimicrobial activity and acidity than phenol, were subsequently developed (Philp et al. 2005). Chlorophenols, particularly the fully substituted pentachlorophenol (PCP), are used as preservatives for timber and textiles towards fungal rot and damage by insects. The toxic nature of PCP is based on its ability to uncouple oxidative phosphorylation and to alter the electrical conductivity of cell membranes (McAllister et al. 1996).

Commercial grades of PCP commonly contain manufacturing by-products such as polychlorinated dioxins and furans, which can have higher

chronic toxicity than PCP. Humans may get occupationally exposed to PCP via inhalation or via dermal contact primarily in situations where they use this preservative, or are in contact with treated wood products or textiles (Extoxnet 1998). Due to release, accidents and spills during the production and application of chlorophenols pollution by PCP has been reported in air, water, soil and sediments (Hale et al. 1994, McAllister et al. 1996). For example, in Finland the widespread use of PCP containing wood preservative has led to contamination of soil around nearly 800 former saw mill sites (Laine 1998).

Even though PCP is toxic to microbes, variety of bacteria and fungi has been isolated with an ability to mineralise PCP (McAllister et al. 1996). Both Gram-positive and Gram-negative bacteria capable of degrading PCP has been described, species of Alcaligenes, Pseudomonas, Mycobacterium,

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Sphingomonas and Streptomyces representing the most commonly reported taxa (McAllister et al. 1996, Laine 1998, Nohynek 1999; Nam et al. 2003). Members of genera of Phanerochaete,

Trametes and white-roting fungi Basidomycetes are the most studied PCP degrading fungi (McAllister et al. 1996, Leung et al. 1997). A critical step in the microbial degradation of PCP is the cleavage of the chlorine-carbon bond called dehalogenation. Two main strategies can be differentiated: (i) all halogen substituents are removed as an initial step of degradation via reductive, hydrolytic or oxygenolytic mechanisms, or (ii) a part of the halogen substituents are removed only after cleavage of the aromatic ring. Common feature for aerobic biodegradation pathways of tri-, tetra-, and pentachlorophenol follows the strategy (i). Thus, only after removal of all or most of the halogen substituents through either hydroxylation or reductive dechlorination the aromatic ring is cleaved as shown in Figure 3 (Häggblom 1992, Leung et al. 1997, Fetzner 1998).

The initial step of aerobic microbial degradation pathway of PCP is well documented and it involves the formation of tetrachloro-p-hydroquinone (TeCH, Figure 3). Apajalahti and

Salkinoja-Salonen (1987) showed that Mycobacterium chlorophenolicum PCP-1 has P450 monooxygenase which dechlorinates PCP to TeCH. TeCH is further degraded through one hydrolytic dechlorination reaction and three consecutive reductive dechlorination steps to 1,2,4-trihydroxybenzene. Reductive dehalogenation (dechlorination) is a two-electron transfer reaction which involves the release of the halogen as halogenide ion and its replacement by hydrogen (Fetzner and Lingens 1994, Fetzner 1998). Further, Mycobacterium chlorophenolicum PCP-1 metabolizes 1,2,4-trihydroxybenzene ultimately to carbon dioxide and water through Calvin-cycle (Häggblom 1992, Leung et al. 1997).

In contrary, Sphingomonas

chlorophenolica ATCC 39723 (synonym to Arthrobacter sp. ATCC 39723, Nohynek et al. 1995) initiates the pathway with an oxygenolytic dechlorination of PCP to TeCH by PCP-4 monooxygenase. TeCH is degraded then by three sequential steps of reductive dechlorination, without any hydrolytic dechlorination reaction to 2,6-dichloro-p-hydroquinone, which has been shown to be mineralized completely to water and carbon dioxide (Leung et al. 1997):

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C a r b o n d i o x i d e a n d w a t e r

- 2 C l -

.

.

H H 2, , 6 - D C H

- C l -

T H C H

- C l -

T e C H

O H

C l

- C l -

T C H B Q

H

H

O H

- 3 C l -

O H

T e C H

- C l -

C l

O H

C l

C l - C l -

B A

O H

C l

O H

C l

H

O H

O H

O H

C l

O H O H

C l C l

C l C l

C l

O H

O H

C l C l

C l

C l

C l

C l

O H

O H

C l C l

C l

Figure 3. Biodegradation pathways of PCP. Pathways A and B are based on the PCP-degrading mechanisms of Mycobacterium chlorophenolicum strain-PCP-1 and Sphingomonas

chlorophenolica ATCC 39723 (Nohynek et. al. 1995), respectively. Abbreviations are as follows: TeCH, tetrachloro-p-hydroquinone; TCH, trichloro-p-hydroquinone; 2,6-DCH, 2,6-dichloro-p-hydroquinone; TCHBQ, trichlorohydroxybenzoquinone; 1,2,4-THB, 1,2,4-trihydroxybenzene (redrawn from Leung et al. 1997).

1.4 Nitrification and autotrophic ammonia-oxidising bacteria in soil Nitrification is part of the biogeochemical nitrogen cycle in soil, in which ammonia is oxidized to nitrite and further to nitrate. The initial oxidation of ammonia is usually the rate-limiting step for nitrification as nitrite is rarely found to accumulate in soils (Kowalchuk and Stephen 2001). Ammonia is released naturally by mineralization of organic

matter, and through the nitrification of ammonia the mineralization of nitrogen (ammonification) is linked to denitrification (Ward 1996). In unmanipulated natural boreal forest soils net nitrification is often negligible, but anthropogenic nitrogen inputs, for example via atmospheric deposition or fertilization, may enhance the nitrification activity (Smolander et al. 2000). This may result to an increased nitrate leaching from soils. Leaching can, for example, lead to nitrate pollution of

1,2,4-THB

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surface- and groundwaters. Furthermore, nitrate may act as a substrate for denitrification in which, together with nitrification, gaseous nitrogen compounds like NOx, N2O are produced. The production and emission of these greenhouse gases is causing great environmental concern, because they are believed to contribute to depletion of the stratospheric ozone layer (Conrad 1996, Kester et al. 1997).

A range of chemolithotropic and chemoorganotrophic bacteria and fungi are capable of biological nitrification in soil, but the chemolithotropic ammonia oxidizing bacteria (AOB) are probably the most important group (Killham 1990) The chemolithotropic AOB obtain their energy for growth from the oxidation of ammonia and assimilate carbon as carbon dioxide. In contrast to AOB, oxidation of ammonia by chemoorganotrophs is not linked to cellular growth (Killham 1990, De Boer and Kowalchuk 2001). Most chemoorganotrophic microorganisms have low nitrifying activities in natural soils, but in some acid soils organotrophic nitrifiers, like fungi, are considered to be responsible for nitrification (Papen and von Berg 1998). The recent discoveries of anaerobic oxidation of ammonia with nitrite (anammox) and autotrophic ammonia-oxidizing archaea challenge the traditional perspectives on microbial nitrification in soils (Strous and Jetten 2004, Könneke et al. 2005, Leiniger et al. 2006, Nicol and Schleper 2006).

Ecological studies of AOB in soil have been hampered by the slow growth and the low growth yield of these microorganisms, which make their isolation in pure culture difficult and time consuming (Head et al. 1998). For example, incubation on solid media for

several weeks is needed for production of microscopic colonies (Prosser and Embley 2002). Cultivation-dependent methods, such as most probable number (MPN) counting and selective plating are difficult for the same reasons (Schmidt and Belser 1994). Immunofluorescencent detection has been applied for ecological studies of AOB, but the production of highly specific polyclonal antibodies is, however, reliant on availability of the AOB pure cultures (Bothe et al. 2000). Direct scanning or transmission electron microscopy analyses of AOB in environmental samples (Nevalainen et al.

1993, Kostyal et al. 1997) are solely inadequate for identification of AOB. The development of culture-independent molecular biological for ecological studies of microorganisms (Insam 2001) has also expanded the potential for characterisation of chemolithotropic AOB in soil. As a consequence, there is an increasing number of molecular studies on diversity and quantity of AOB in wide range of terrestrial habitats (Bothe et al. 2000), including agricultural soils (Bruns et al. 1999, Mendum et al.

1999 and 2002), grasslands (Kowalchuk et al. 2000a, 2000b, Webster et al. 2002), and temperate or boreal forest soils (Lavermann et al. 2001, Carnol et al. 2002, Bäckman et al. 2003, Mintie et al.

2003, Hermansson et al. 2004).

1.4.1 Physiology of chemolithotropic ammonia-oxidizing bacteria

The “conventional” aerobic oxidation of ammonia (NH3) to nitrite (NO2

-) by chemolithotropic ammonia-oxidizing bacteria (AOB) is a two-stage reaction, which requires molecular oxygen and provides energy for the assimilation of CO2 via Calvin cycle. In

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the first reaction (i) ammonia is oxidized to hydroxylamine by the enzyme ammonia monooxygenase (AMO). It is generally accepted that ammonia (NH3) not ammonium (NH4

+) is used as substrate by AMO, and therefore the ammonia/ammonium ratio (i.e. ionization of ammonia) may affect the growth of AOB (Kowalchuk and Stephen 2001). NH3 + O2 + 2H+ + 2e- → NH2OH + H2O (i)

AMO from chemolithotropic AOB is a labile two-component membrane-bound enzyme (Hooper et al.

1997), and it has three subunits, AMO-A, AMO-B and AMO-C with different sizes, structures and arrangements within the membrane/periplasmic space of the cells (Jetten et al. 1997). These three sub-units are encoded by the genes amo-C, amo-B and amo-A of the amo operon (Klotz et al. 1997, Sayavedra-Soto et al.

1998). AMO has a low substrate specificity, and therefore it is capable of oxidizing a wide range of non-polar compounds e.g. carbon monoxide, methane, methanol, ethane, benzene and phenol. Acetylenic compounds, nitrapyrin, allylthiourea and sulfur compounds are known to irreversibly inhibit the activity of the AMO enzyme (Hooper et al. 1997).

Ammonia is oxidized to hydroxylamine by AMO in an endergonic reaction (∆Gº -120 kJ mol-1), in which one oxygen atom from molecular oxygen is incorporated into ammonia and the other is reduced to water. This reaction (i) consumes two electrons, which are derived from the oxidation of hydroxylamine to nitrite (ii) catalyzed by the enzyme hydroxylamine oxidoreductase (HAO).

NH2OH + H2O → HNO2 + 4H+ + 4e- (ii)

HAO is an enzyme with a highly complex structure, located as a soluble enzyme in the periplasmic space, but anchored in the cytoplasmic membrane (Bergmann et al. 1994 and 2005). The oxidation of hydroxylamine to nitrite by HAO provides in total four electrons, and two of the four electrons are transferred via the tetra-heme cytochrome c554 to AMO. The remaining two electrons are diverted into an electron transport chain to be utilized for the generation of ATP and NAD(P)H or used in the CO2

assimilation (Jetten et al. 1997, Bothe et

al. 2000). Intermediates released in the HAO reaction can be nitroxyl (NOH) and nitric oxide (NO).

The main source of carbon for chemolithotrophic AOB is CO2 which is assimilated in Calvin cycle. The assimilation of CO2 results in production of cellular material. For each fixed molecule of carbon, 35 molecules of ammonia are needed. This consumes 80% of the energy generated by chemolithotrophic AOB and explains the low cellular yield and slow growth rate of AOB (minimum doubling time of 8 hours).

The physiology of obligately aerobic AOB is not completely understood, and it is proposed (Schmidt and Bock 1997, Schmidt et al. 2003) that chemolithotrophic AOB could also obtain energy for their growth from anaerobic oxidation of ammonia (iii). NH3 + N2O4 → 0.33NO2

- + 1.33H+ + 0.33N2 + 2NO + 1.33 H2O (iii)

Under anoxic conditions, dinitrogen tetraoxide (or nitrite) is the most likely electron acceptor for

18

ammonia oxidation. Hydroxylamine and NO are produced as intermediates, and dinitrogen, nitrite and nitric oxide as main products, respectively. Under anoxic conditions ammonium oxidation activity is relatively low and generation time of AOB is slow with the doubling of time 30 days (Bothe et al. 2000, Schmidt et al. 2003).

Reduced nitrogen compounds like urea can serve as primary ammonia and energy sources for some chemolithotropic AOB strains. In that case intracellular hydrolysis of urea (by enzyme urease) is needed for the subsequent oxidation of the released ammonia. AOB strains isolated from acidic soils are generally ureolytic, and it is known that ureolytic AOB can grow in pure culture at lower pH with urea than with ammonia as substrate (Burton and Prosser 2001). Therefore, ureolytic AOB may have an ecological advantage in acidic soils receiving animal wastes or urea fertilizers (De Boer and Laanbroek 1989, Burton and Prosser 2001, Koper et

al. 2004).

1.4.2 Phylogeny of chemolithotropic ammonia-oxidizing bacteria

Gram-negative ammonia-oxidizing bacteria (AOB) were earlier classified by phenotypic characteristics into five different genera i.e. Nitrosococcus, Nitrosovibrio, Nitrosospira, Nitrosolobus and Nitrosomonas (Prosser 1989). The phenotypic classification of AOB (based on cell morphology and ultrastructure) was, however, limited by difficulties in obtaining pure cultures due to the slow growth and the low yield of AOB cells on laboratory media. The use of cultivation-

independent molecular techniques enables to circumvent these limitations and provides more comprehensive information on phylogenetic relationships of AOB.

The current genotypic classification of AOB is based on molecular phylogenetic analysis of 16S rRNA gene sequences (Head et al. 1993, Utåker et al. 1995, Bothe et al. 2000, Kowalchuk and Stephen 2001, Prosser and Embley 2002). Due to similarity of their 16S rRNA gene sequences Head et

al. (1993) proposed reclassification of Nitrosovibrio, Nitrosospira and Nitrosolobus into one common genus Nitrosospira. Further molecular investigation of the existing pure cultures showed that autotrophic AOB comprise two monophyletic groups. One group belongs to the γ-Proteobacteria with Nitrosococcus oceani and Nitrosococcus

halophilus as the only recognized species. The other group belongs to the ß-subgroup of Proteobacteria and includes two genera, Nitrosospira and Nitrosomonas.

Molecular characterisation of AOB in their natural habitats has led to a more comprehensive classification of these bacteria (Head et al. 1998). Stephen et al. (1996, 1998) suggested subdivision of ß-subgroup of Proteobacteria AOB and environmental sequences belonging to the β-Proteobacteria into 7 to 9 gene clusters. Clusters 1 to 4 belong to Nitrosospira ssp. and clusters 5 to 9 consist of Nitrosomonas ssp. The Nitrosospira cluster 1 and Nitrosomonas

cluster 5 are characterized only by environmental sequences and have no representative pure cultures (Figure 4).

19

Figure 4. Phylogeny of autotropic AOB. Schematic representation of the known autotrophic AOB based on 16S rRNA gene sequences. Ammonia oxidizers are shown in dark grey (adapted from Kowalchuk and Stephen 2001).

The community structure of AOB has been investigated in various soil environments, and the molecular analysis of β-Proteobacterial AOB-like 16S rRNA gene sequences commonly yields representatives of Nitrosospira clusters 2, 3, 4 and Nitrosomonas cluster 6 (Stephen et al. 1996, Kowalchuk et al.

1997, 1998, Burns et al. 1999, Mendum et al. 1999). These observations support the suggested dominance of Nitrosospira

populations over Nitrosomonas by cultivation dependent techniques in most

soil environments (Kowalchuk and Stephen 2001).

The key enzyme of chemolithotropic AOB is ammonia mono-oxygenase (AMO) and the amoA gene, which encodes the active α-subunit of AMO, is also exploited as molecular marker for the phylogenetic study of AOB (Kowalchuk and Stephen 2001). The phylogeny of amoA gene was found to correspond largely to the phylogeny of the 16S rRNA gene sequences in AOB (Rotthauwe 1995, Purkhold et al. 2000, Ivanova et al. 2000, Aakra et al. 2001).

Anamox

20

1.4.3 Molecular detection methods for ammonia-oxidizing bacteria: a model for studying microbial diversity in soil using ribosomal RNA and functional gene markers

The composition of complex microbial communities in natural or man-made environments is most often directly analyzed by ribosomal RNA-targeted nucleic acid probes (Amann et al. 1995, Head et al. 1998, Amann and Ludwig 2000, Lipski et al. 2001). There are several explicit reasons for focusing on ribosomal RNA (rRNA): (i) the rRNA is the key element of the protein synthesing machinery and present in all organisms, (ii) the rRNA genes are extremely conserved in (secondary) structure and in nucleotide sequences, which allows alignment of disparate sequences, so that they can be used in phylogenetic analyses, (iii) there is a large amount of rRNA in most cells, and it is easily recovered from all types of organisms, (iv) there is no apparent lateral gene transfer, (v) rRNA sequences are sufficiently long to provide statistically significant comparisons, and finally, (vi) the availability of huge rRNA databases for comparative sequence analysis.

The organization of bacterial, e.g. AOB ribosomal RNA clusters is in order 5’-16S-23S-5S-3’. Among the three ribosomal RNAs, the gene coding for the small subunit of 16S rRNA (approximately 1500 nucleotides) has become the most widely used marker in molecular bacterial phylogeny (Head et

al. 1998). Due to the relatively high phylogenetic coherence of AOB, the 16S rRNA directed analysis has become the approach of choice for molecular detection of AOB. Further, this has also enabled development of specific 16S rRNA oligonucletide probes for the

detection of AOB (McCaig et al. 1994, Hiorns et al. 1995, Wagner et al. 1995, Hovanec and Delong 1996, Morbarry et

al. 1996, Hastings et al. 1997, Kowalchuk et al. 1997). The 16S rRNA gene is present in only one copy per genome in AOB (Aakra et al. 1999) and oligonucletide probes can be therefore designed without considering the heterogeneity among gene copies within the different AOB species. Also the copy number itself is not of concern when trying to interprete the 16S data in terms of AOB cell numbers in the environment.

The specificity and sensitivity of the 16S rDNA/rRNA-probes targeting AOB was reviewed in detail by Utåker and Nes (1998) and Purkhold et al.

(2000). Most of the probes have been designed for the monophyletic group of the AOB within the β-Proteobacteria. Ideally, oligonucletide probes used in polymerase chain reaction (PCR) or in fluorescence in situ hybridization (FISH) should fulfil three criteria: (i) a sufficient length of the 16S rRNA gene should be recovered to allow robust phylogenetic analysis of the recovered sequences, (ii) primers/probes should target all members of the AOB clade and (iii) primers/probes should not amplify 16S rRNA from organisms outside of AOB clade (Kowalchuk and Stephen 2001). Molecular techniques potentially involves a number of intrinsic biases, such as DNA extraction efficiency, DNA purification methods, PCR biases, formation of chimeras, and sequencing or cloning errors. These must be considered when interpreting molecular ecological data (Wintzingerode et al. 1997, Prosser and Embley 2002).

Early applications of molecular tools for the detection of AOB in their natural habitats were based on PCR-

21

amplification of 16S rRNA gene using primers designed to target ammonia oxidizers. Either direct PCR (Stehr et al.

1995, Kowalchuk et al. 1997, Speksnijder et al. 1998, Stephen et al.

1998,) or nested PCR (Hiorns et al. 1995, Ward et al. 1997, Hastings et al. 1998) was conducted, and the amplification products were then cloned and sequenced or analysed by different fingerprinting methods. Alternatively, specific oligonucletide probes for in situ detection of AOB by FISH were applied. In situ analysis and quantification of AOB has mainly been achieved from habitats with large and actively growing AOB populations, such as wastewater treatment processes (Wagner et al. 1995, Schramm et al. 1999, 2000, Aoi et al.

2000), activated sludges (Morbarry et al.

1996, Juretschenko et al. 1998), or rhizospheres (Briones et al. 2003).

Molecular fingerprints of AOB communities from environmental samples can be produced in different ways. Kowalchuk et al. (1997, 1998, 1999, 2000a, 2000b) and McCaig et al.

(1999) used PCR-amplification of AOB 16S rDNA and denaturing gradient gel electrophoresis (DGGE, Muyzer et al.

1997) to characterize diversity of AOB communities in soils, in sediments and in composting materials. Similarly, Bano and Hollibaugh (2000) combined PCR amplification and DGGE, for analysing AOB in 246 water samples from Arctic Ocean. Whitby et al. (1999) studied distribution of AOB populations in temperate oligotrophic freshwater lake by applying PCR and restriction fragment length polymorphism (RFLP) system. Lavermann et al. (2001) assessed effects of spatiotemporal variation of AOB communities in nitrogen saturated forest soil using PCR and temperature gradient

gel electrophoresis (TGGE). Bäckman et

al. 2003 applied both DGGE and single strand conformation polymorphism (SSCP) for comparative analyses of AOB communities in boreal forest soil. During these molecular investigations of different environmental samples it became evident that Nitrosospira-like AOB are ubiquitously present and also dominant AOB in most natural environments.

The 16S rRNA sequence similarity among different ammonia oxidizers is high, so that only limited phylogenetic information can be obtained from closely related AOB strains (Aakra et al. 1999 and 2001, Purkhold et al.

2000). However, AOB are unique in their possession of another molecular marker; the key functional gene, amoA, which encodes the active site of ammonia mono-oxygenase (AMO) in chemolithotropic AOB. The potential to exploit this marker is tempting because the amoA gene contains more sequence variation than the 16S rRNA gene, and may therefore allow for a greater discrimination between closely related AOB (Rotthauwe et al. 1995). Further, Rotthauwe et al. (1997) designed a PCR primer set which is highly specific for the amoA gene enabeling study of phylogeny of AOB populations in environmental samples. As a result, molecular fingerprints of AOB communities from various habitats were produced by amplification of the amoA gene in combination with terminal restriction fragment length polymorphism (T-RFLP, Horz et al. 2000, Bottomley et al. 2004) or DGGE and sequencing (Oved et al.

2001, Nicolaisen and Ramsing 2002, Avrahami et al. 2002 and 2003). Generally, good agreement is found between the environmental AOB

22

populations described by amoA analyses compared to that of 16S analyses, particularly in the division between Nitrosospira and Nitrosomonas (Kowalchuk and Stephen 2001).

PCR-based molecular methods for quantification of AOB in environmental samples were also designed (Bothe et al. 2000, Kowalchuk and Stephen 2001). Both 16S rRNA and amoA genes were targeted (Mendum et

al. 1999, Philps et al. 2000, Bjerrum et

al. 2002). Using the 16S rRNA gene has the advantage over amoA, in that AOB has only a single copy of 16S rRNA gene per genome, whereas the copy number of amoA may vary (2 to 3 per genome). Either competitive (Mendun et al. 1999, Stephen et al. 1999, Phillips et al. 2000a, Bjerrum et al. 2002) or real-time (Hermansson and Lindgren 2001, Okano et al. 2004) PCR strategies were applied for the molecular quantification of AOB. The results from these molecular studies suggested that the traditional culture dependent methods like MPN greatly underestimate the real number of AOB in soils.

2. Aims of this study The general objectives of this

study were (i) to evaluate the potential for intrinsic bioremediation in natural boreal surface soils, and (ii) to assess the seasonal variation in the biodegradation activities in boreal coniferous forest soil, and iii) to describe the ß-subgroup of Proteobacterial ammonia oxidising bacteria (AOB) in soil contaminated by oil-refinery waste-sludges using both the culture-independent and the culturing methods.

The specific aims of this study were to:

• Describe the effect of low temperature on mineralization rates of low and moderate concentrations of selected PAH and polychlorinated phenols in boreal soil, and compare those to mineralization rates obtained under constant temperate conditions (Paper I). • Evaluate the temperature dependence of biodegradation of natural organic matter and anthropogenic organic compounds in boreal soils (Paper II). • Assess the size, activity and structure of AOB populations in an oil-contaminated landfarming soil by culture independent molecular methods and culturing methods (Paper III). • Explore the performance of cation-exchange membranes as a tool for enriching of AOB in situ from the oil-contaminated landfarming soil (Paper IV).

3. Materials and methods

3.1 Study sites and soil sampling

The studies were performed with soils collected from three different sites. One sampling site locates at the Viikki Experimental Farm of the University of Helsinki, in Helsinki, Finland (60°13’ N and 25°01’ E). Agricultural soil was collected from land that has been rotated between crops of cereals and oil plants or used as pasture for the last four decades. At the time of sampling (5th of October 1996) the crop had been harvested and soil ploughed to 20 cm depth. In 1996 lime (2300 kg), nitrogen (200 kg), phosphorus (46 kg), potassium (46 kg) and the herbicide trifluralin (1 kg) were applied ha-1 for cultivating rape (Brassica campestris

oleifera). The colour of the agricultural soil was grey and the textures

23

corresponded to sandy clay loam. The forest humus soil was sampled from an area that was conserved in its native condition for the last four decades at the Helsinki University Experimental Farm. The stand was dominated by deciduous tree species of European aspen (Populus

tremula) and birch (Betula pendula). The biodegradation activities of PAH and polychlorinated phenols was also studied separately in rhizosphere fraction of the forest humus soil. The colour of the upper an approx. 5-cm organic humus layer was dark brown and it contained humus material with roots of the European aspen.

The second study site located in the vicinity of the Hyytiälä Forestry Field Station of the University of Helsinki, in southern Finland (61°84’ N, 24°26’ E). The soils at Hyytiälä forests are typical podzols, which consist of four distinctive horizontal layers: an organic upper layer (O) rich in humus, an eluvial horizon (E), which is supported above an illuvial mineral layer (B) and the unaltered ground soil (C) (Lundström et al. 2000). The studied soil comprised an upper 5-cm humus layer and an approx. 5-cm eluvial horizon supported above an approx. 30-cm illuvial horizon.

At the closed canopy stand the dominating tree species was Scots pine (Pinus sylvestris, sown after prescribed burning in 1962). Detailed descriptions of the site can be found elsewhere (Ilvesniemi et al. 2000).

The third study site was an oil-waste landfarming site in southern Finland (60°15’ N and 25°30’ E). Landfarming is a commonly used waste treatment method for oil sludges produced in large quantity by petroleum industry (Felsot et al. 1995, Philp and Atlas 2005). In landfarming the oil wastes are ploughed into the soil, and inorganic nutrients, in particular nitrogen in form of ammonium salts, are usually added to the soil. Indigenous soil microorganisms assimilate the oil compounds mainly as carbon and energy sources for growth, but some of the added nitrogen may be transformed by nitrification into nitrite and nitrate. On the study site 20 years of landfarming of oil waste products derived from the adjacent petroleum refinery had resulted to an average oil concentration of 35 to 59 g kg-1 soil. The (sandy clay loam) soil had also been fertilized for more than 10 years with high amounts of urea (900 g urea-N m-2 yr-1) targeted to give a carbon : nitrogen ratio of 10 : 1, and amended with wood chips to improve porosity (Peltola et al. 2006).

3.2 Experimental methods The experimental methods used

in this study are described in detail in the original papers, and summarized in Table 1.

24

Table 1. Methods used in this study.

Analysis Method Described in Paper

Reference/Manufacturer

DNA extraction from cation-exchange membranes

Bacterial DNA isolation Kit IV Mo Bio Laboratories

DNA extraction from soils Soil DNA isolation Kit Fast Spin DNA Soil Kit

III IV

Mo Bio Laboratories

DGGE profiling Denaturing gradient: 45 to 57% III and IV (Kowalchuk et al.1997) Fluorescence in situ hybridization (FISH) Digital microscopy IV (Schramm et al.1998 and 2000) Hydrolytic enzyme activities Kinetic fluorometry using fluorogenic

substrates II (Kähkönen 2003)

Potential ammonia oxidation Incubation experiment III and IV (Kandeler, 1996) Potential methane oxidation Incubation experiment II Kähkönen (2003)) Mineralization of radiolabelled model xenobiotics

Microscale radiorespirometry with phosphoimaging

I and II (Tabor et al. 1976, Fulthorpe et

al.1996, Kähkönen et al.2001) Most probable number (MPN) Incubation experiment IV (Rowe et al.1977) Nested PCR 16S rRNA primers III and IV (McCaig et al.1994)

(Kowalchuk et al.1997) Real-time PCR 16S rRNA targeted primers III (Hermansson et al. 2001 and 2004) Sequencing III and IV Applied BioSciences Inc. Soil respiration Incubation experiment II Kähkönen (2003) Phylogenetic analysis IIII and IV (Staden et al.1998)

(Galand et al. 2003) Statistical analysis One-way ANOVA I, III, IV Temperature measurements On-line sensors I and II Kähkönen (2003))

25

4. Results and Discussion

4.1 Biodegradation of organic contaminants in boreal surface soils

Potential for intrinsic soil bioremediation was studied in two different soil from the Helsinki University Experimental farm, Viikki, using PCP, phenanthrene, pyrene and 2,4,5-TCP as model compounds (Papers I and II). The study entailed the addition of low and moderate concentrations of the 14C-labelled model compounds to the Viikki soils, and incubation of the soils at a constant temperature of + 20 ºC or at a temperature range of -2.5 - +15 °C. The low amendments of the substrate (≤ 5 µg cm-3) were designed to simulate long-distance air borne pollution, whereas the higher amendment (50 µg cm-3) was used to simulate effects of point-source pollution. The substrates, like low molecular PAH (phenanthrene and pyrene), are major products of anthropogenic combustion of biomass and fossil fuels, and their accumulation in soils and sediments through atmospheric deposition has been reported (Saunders et al. 1995, Aamot et al. 1996, Wilcke et al. 1996 and 2005, see also the review by Wilcke 2000). For example, at the study area atmospheric concentrations of 0.009 and 0.059 µg of pyrene and phenanthrene, respectively, m-3 of air were measured during a strong episode of long-range transported particulate mass (Niemi et al. 2002). Koivula et al. (2004) estimated that the annual deposition of total PAH would be 0.07 µg g-1 of soil in southern Finland. Salla (1999) reported an average background concentration of 2.82 µg of total PAH g-1 of natural surface soil at Helsinki area. In the

vicinity of Helsinki University Experimental farm the PAH levels were 0.69 to 2.07 µg g-1 of soil

The results in Tables 1 and 2 of Paper 1 showed that the Viikki soils were capable of degrading concentrations of 0.2 - 50 µg of PCP, phenanthrene, pyrene or 2,4,5-TrCP cm-3 of soil. The mineralization level of the model compounds was highest at +20 °C in the agricultural soil. The half-lives (t1/2) of the model compounds at +20 °C decreased when the concentration of the substrate was increased. This may indicate a possible growth-linked mineralization of processes of the organic contaminants in the Viikki soils during laboratory incubations performed at room temperature.

Studies on the biodegradation of PAH and polychlorinated phenols by uncontaminated pristine boreal and temperate soils have been reported. For example, Smith et al. (1997) added 500 µg phenanthrene or pyrene g -1of pristine soils, and noticed half lives of phenanthrene and pyrene of 80 and 155 days in sandy, and 86 and 216 days in the organic soil, respectively. Charmichael and Pfaender (1997) tested two sandy soils exposed to PAHs with the total concentrations of 11 and 2 µg PAH g-1 of soil. After spiking the soils with 14C-labelled phenanthrene or pyrene they recovered from 31 to 50% of the added label as 14CO2 in 30 days. Koivula et al. (2004) measured recovery of 45% of the 14C-labelled pyrene (100 µg g-1) as 14CO2 in natural boreal humus in 170 days. Middeldorp et al. (1990) reported rates of PCP (30 mg kg-1) mineralization between 13 to 18 µg by g-1 of pristine sandy and peaty soils, respectively, in 112 days. Briglia et al. (1994) noticed that the rate of PCP mineralization was 0.5 and 1.4 µg

26

PCP g-1 of pristine sandy or peaty soil, respectively in 30 days. Karlson et al.

(1995) studied the degradation of PCP at bench-scale with moderately PCP-contaminated (30 µg g-1) sandy Danish soil, and found that the soil mineralised 56% of the spiked 14C-PCP to 14CO2 in 85 days. Later, Miethling and Karlson (1996) reported that 30 µg of PCP was mineralised per g of the same soil completely in 210 days. In general, the rates of PCP, phenanthrene and pyrene mineralization reported in the studies cited above are similar to those obtained in this thesis with the highest applied concentration (50 µg cm-3) of the substrate at + 20 ºC (Table 2 in Paper 1).

In the present study the effect of temperature on the potential for biodegradation of organic soil contaminants was studied with laboratory incubations performed at temperature range of -2.5 to +15 °C. The mean annual air temperature in the Helsinki metropolitan area, which locates in the southern boreal climate zone, was between 1971 and 2000 +5.6 °C (spring +3.8 °C, summer +15.9 °C, autumn +6.2 °C , winter -3.8 °C) (Finnish Meteorological Institute 2006). The present results showed that low temperature had a clear effect on the mineralization of PCP, phenanthrene, pyrene and 2,4,5-TCP in the Viikki soils (Table 1 and 2 in Paper 1). For example, mineralization of pyrene was completely (< 5 % of input 14C evolved as 14CO2 in 98 d) inhibited at low temperature. Furthermore, the mineralization rates of high concentration (50 µg cm-3) of PCP, phenanthrene, and 2,4,5-TCP in the agricultural soil were 4 to 5 times lower at low temperature than those at constant room temperature (+ 20 ºC). The results

suggest that not only the temperature but also the quality and the concentration of the substrate affect to the biodegradation potential of organic contaminants in natural boreal surface soils.

Nedwell (1999) reviewed the effect of low temperatures on the affinity of microorganisms for substrates. Generally, it is accepted that uptake affinity of microorganisms for organic substrates is inhibited, when the temperature falls below their optimum for growth. At low temperature the microorganisms try to preserve the vital fluidity of their cell membranes by altering the membrane structure and composition. This is done by increasing the proportion of unsaturated fatty acids, or by shortening their chain length or by replacing iso fatty acids with the corresponding anteiso in the membrane lipids. The decrease in membrane fluidity (or increase in the membrane stiffness) reduces the efficiency of transport proteins and enzymes in the membrane required for active substrate uptake, which in turn decreases the affinity of the enzymes for their substrates (Gounot 1991, Nedwell 1999). The ecological implication of this in soil is that microorganisms will become increasingly unable to sequester low substrate concentrations from their environment at temperatures below their optimum for growth. Consistent with this, mineralization of low concentrations (< 5 µg cm-3) of 14C-labelled sodium-glutamate (a model compound for easily biodegradable organic substrate) in the Viikki soils were inhibited by the low temperature (Table 2).

27

Table 2. Effect of low temperature on the mineralization of 14C-labelled sodium-glutamate in the Viikki soils. Mean value with standard error (±) of triplicate measurement is shown.

Soil type % of 14

C mineralized to 14

CO2

in 7 days at + 4 ºC

% of 14

C mineralized to 14

CO2

in 7 days at + 20 ºC

0.003* 5* 50* 0.003* 5* 50*

Agricultural clay loam

1.5 (± 0.9)

6.5 (± 0.6)

32 (± 4.0)

63 (± 1.3)

65 (± 4.0)

66 (± 8.0)

Forest humus 3.4 (± 0.3)

2.1 (± 1.1)

31 (± 9.1)

59 (± 2.6)

69 (± 1.9)

66 (± 5.9)

Rhizosphere fraction of the forest humus

3.8 (± 0.3)

2.6 (± 0.4)

24 (± 2.2)

37 (± 5.8)

70 (± 2.1)

62 (± 2.5)

* input concentration (µg cm-3)

In contrast to the results obtained with glutamate, the mineralization rates (half life, t1/2) of high concentrations of PCP and phenanthrene were clearly more affected by the low temperature than the low concentrations in the Viikki soils (Paper 1, Table 1 and 2). This may be explained by decreasing water solubility of these organic substrates at low temperature (Philp et al. 2005), which in turn decreases the bioavailability of these organic contaminants to degrader microorganisms or enzyme systems.

4.2 Temperature dependency of biodegradation of anthropogenic organic contaminants and natural autochthonous organic matter in boreal surface soils

The Arrhenius equation and the Q10 value, which is a quotient of the activity increase over 10 °C, were applied to analyze the temperature-dependency of biodegradation of anthropogenic organic contaminants (Papers I and II) and

natural autochthonous organic matter in boreal soils (Paper II). The Arrhenius equation, which was originally used for modelling of temperature relationships of specific enzyme kinetics, will result a linear relationship (r2-value) between the log activity and the inverse of the absolute temperature, if the activation energy (Ea–value) is constant over the studied temperature interval. Therefore, as discussed by Pietikäinen et al. (2005), the law of Arrhenius reflects the microbial activities in pure cultures and in soil only in a narrow temperature range. The results in Figure 5 (see below) showed that log mineralization rates of low concentrations (≤ 5 µg cm-3) of phenanthrene were highly linear (r2 > 0.951, p< 0.05, n=3) in respect to the inverse of the absolute temperatures of 0 to + 15 ºC in the Viikki soils. In contrast to this, the mineralization rates of PCP and 2,4,5-TCP followed only moderately (r2 < 0.877, p<0.05, n=3) the Arrhenius relationship over the temperature range of 0 to + 15 ºC. (Table 1 in Paper 1)

28

3.45x10-3 3.50x10-3 3.55x10-3 3.60x10-3 3.65x10-3

-6

-4

-2

0

2(A)

ln k

(m

g ph

enan

thre

ne m

iner

aliz

ed w

eek-

1 )

Temperature (K-1)

Ea = 90.9 kJ

R = 0.994

Ea = 133.2 kJ

R = 0.979

Ea= 171.0 kJ

R = 0.856

Figure 5. The Arrhenius relationship of phenanthrene mineralization at low temperature in Viikki soils. Mineralization of 0.2 (□), 5 (o), and 50 (∆) µg phenanthrene of cm-3 soil within the temperature range of 0 to +15 °C in the agricultural soil (A) and in the rhizosphere fraction of the forest humus soil (B). Mean of triplicate measurements with the respective activation energy (Ea) and the correlation coefficient (R) values are shown.

It is known that the growth and activity of psychrophilic bacteria follows the Arrhenius relationship linearly down to º0 C (Melin 1997). Thus, the results in Figure 5 suggest that true psychrophilic microorganisms were responsible for the phenanthrene mineralization at low temperature in the Viikki soils. Ea–values increased from 75 to 172 kJ mol-1 with the phenanthrene concentrations increasing from 0.2 to 50 µg cm-3 of soil (Table 1 in Paper I, Table 6 in Paper II). This indicates that higher energy requirement for the mineralization of high than of low soil concentrations of phenanthrene. Phenanthrene, a three-ring PAH, is a fairly volatile compound, and its volatility decreases as the environmental temperature decreases. This may have elevated the bioavailabilty of the low concentrations of phenanthrene to the psychrophilic microorganisms or enzyme systems in the Viikki soil. Furthermore, the lesser

effect of temperature on the mineralization of low phenanthrene concentrations may be due to different phenanthrene uptake mechanisms of the degrader microorganims. Volatile organic (like phenanthrene) and inorganic (like ammonia) substrates are taken up by microorganims through passive transport, which is not affected by decreased fluidity of the membrane at low temperature (Nedwell 1999).

The Q10 values for the mineralization of phenanthrene were 4.5 to 14.5 in agricultural and 2.0 to 7.6 in the rhizosphere fraction of the forest humus soil at temperatures from 0° to +15 °C (Table 1 in Paper I, Table 6 in Paper II). It thus seemed that the phenanthrene degrading microbes associated with the fine roots of aspen showed much less response to temperature than those in the agricultural soil. The present results indicate that fine roots of deciduous tree species (European

3.45x10-3 3.50x10-3 3.55x10-3 3.60x10-3 3.65x10-3

-6

-4

-2

0

2

(B)

ln (

mg

phen

anth

rene

min

eral

ized

wee

k-1)

Temperature (K-1)

Ea = 171.5 kJ

R = 0.938

Ea = 104.6 kJ

R = 0.974

Ea = 75.3 kJ

R = 0.951

29

aspen) catalyzed the psychrophilic mineralization of the organic soil contaminants in the forest humus. The beneficial effect of rhizosphere on the mineralization of the organic contaminants at low temperature may be due the fibrous plant root systems, which provide the microbes plant-derived nutrients, and increase the probability of contact of microbes with the contaminants in soil (Schwab et al. 1995).

The temperature dependence of endogenous soil respiration (carbon dioxide evolution) and hydrolytic enzyme activities were measured in the soil of a coniferous stand at Hyytiälä, (Figure 2 in Paper II). The Q10-values of endogenous carbon dioxide evolution measured for the humus layer of the Hyytiälä Pinus

sylvestris stand ranged from 2.3 to 2.8 (Table 3 in Paper II) at temperatures from -3 to +12 °C indicating a moderate temperature dependence. The corresponding energies of activation (Ea) for carbon dioxide evolution were from 60 to 80 kJ mol-1 (Table 3 in Paper II). For compounds with an Ea of 70 (± 5) kJ mol-1 an increase in temperature by 1 degree of °C will cause a 10 % increase in the rate constant, and a 10 °C increase in temperature will increase the rate constant a factor of 2.5 (Wolfe et al.

1989, Gounot 1991, Guillou and Guespin-Michel 1996). Kirschbaum (1995, 2000) presented an estimation of Q10 of 4.5 – 8 for temperatures of 0 to 10 °C, which differs from our Q10- values of 2.3 to 2.8 (Table 3 in Paper II) obtained for Hyytiälä forest soil at temperatures from –3 to +12 °C. If our estimation of Q10 is valid for coniferous forest soils also elsewhere in the boreal zone, an increase of 1 °C of soil temperature would accelerate the evolution of carbon

dioxide by 13% to 18% rather than 35% - 70% as prognosed by the model of Kirschbaum (1995, 2000).

The activities of α-glucosidase, β-glucosidase, β-xylosidase, butyrate-esterase, N-acetyl-glucosamidase and phosphomonoesterase were higher or equal in October than in July-August (Table 5 in Paper II). This indicates that biodegradative hydrolytic activities in the in the Hyytiälä forest soil were regulated by other factors than only soil temperature. The Q10-values for hydrolytic enzyme activities were 1.6 to 2.1 at the temperature range from +14 to +30 °C in the Hyytiälä soil (Table 4 in Paper II). In the present study the potential of the selected hydrolytic enzyme activities were measured using fluorogenic 4-methylumbelliferyl (MUF) conjugated substrates with an initial substrate concentration of 1 mM. This was done in order to obtain substrate saturation for the studied enzymes. The measurement of potential methane oxidation in the Hyytiälä forest soil was performed using the same principle of high initial substrate concentration (Table 5 in Paper II). However, as the actual soil concentrations of the substrates may be lower than those of the synthetic substrates, and because the concentration of the substrate has a profound effect on the substrate affinity, the actual temperature dependence of the soil activity may therefore be stronger than that of the potential activity.

4.3 Activity and population size of ammonia-oxidizing bacteria in oily landfarming soil

The numbers of AOB in landfarming soil were determined with two experimental methods: Most

30

Probable Number (MPN; Paper IV) technique and real-time PCR (Paper III). MPN is the most widely used method for enumerating of culturable AOB in soils (De Boer and Kowalchuk 2001). We found 0.5 – 2.0 x 105 MPN-countable AOB g-1 of the landfarming soil (Figure 2 in Paper IV). Our MPN-counts of AOB from soil with extreme level of oil pollution (10 – 15 mg oil-hydrocarbons g-1 of soil) are in the same order of magnitude as those frequently reported in agricultural soils subjected to different fertilization and cultivation conditions (Rowe et al. 1977, Berg and Rosswall 1985, Burns et al. 1999, Philps et al.

2000b, see Prosser and Embley 2002). Due to the well-accepted

problems in MPN counting, which result from the selective nature of laboratory media and incubation conditions, the MPN method counts only the culturable and the active cells under the MPN assay conditions (Prosser and Embely 2002). In contrast, cultivation independent PCR-based methods count the active and the dormant (or dead) cells or their DNA in the specimen (Hermansson and Lindgren 2001, Okano et al .2004). This provides a more comprehensive view on the total population size of AOB in a given environmental sample. We found with real-time PCR that the numbers of AOB were 4 x 105 to 9 x 105 cells g-1 of oily landfarming soil (Table 1 in Paper 1). This range is 2 to 20 times higher than those observed with the MPN-counts (IV, Fig. 2). In other studies, the PCR-based methods were reported to detect 10 to 1000 more AOB compared to the MPN-method, depending on the soil sample and the laboratory medium used (DeBoer and Kowalcuk 2001, Bjerumm et al.

2002, Okano et al. 2004). The present results suggest that: either (i) the applied

real-time PCR assay underestimated real numbers of AOB in the landfarming soil or (ii) the number of dormant or inactive (unculturable) AOB was particular low in the landfarming soil.

A number of culture-independent PCR strategies on quantification of AOB in soils have been published. Mendum et al. (1999) used competitive PCR assay for targeting both amoA and the 16S rRNA gene and found 105 to 108 gene copies g-1 in arable soil. Philps et al. (2000a) and Bjerum et al.

(2002) also used competitive PCR with the amoA marker gene and found the highest density 105 AOB of g-1 in fertilized agricultural soil and in rice paddy soils. Hermansson and Lindgren (2001) used real-time PCR and found 106 to 107 AOB cells g-1 in Swedish arable soil. Okano et al. (2004) reported mean AOB population sizes of 106 g-1 in arable soil. The estimates for AOB population sizes in the oily landfarming soil were 1/10 to 1/100 of those reported by Hermansson and Lindgren (2001) and by Okano et al. (2004), but the same order of magnitude than found by Mendum et

al. (1999), Phillips et al. (2000a) and Bjerum (2002). Several environmental factors like geographical, climate and anthropogenic influences as well as moisture, pH and substrate supply control the numbers of AOB in soil. Therefore, a direct comparison of the present results with studies in the literature is not straightforward.

Potential for ammonia oxidation in the landfarming soil was measured over three growing seasons in 1999 - 2001 (III, IV). The assay for potential ammonia oxidation is considered an index for the size of the active AOB populations; because the duration of the assay (24 h) is usually short (Myrold

31

1997). The potential for ammonia oxidation (0.05 – 0.28 µg N g-1 d.w h-1; Table 1 in Paper III) in the landfarming soil was constant, except for the decline of the activity during the dry summer season in 1999. This result supports the results from the measurement of AOB cell numbers, indicating presence of a large and active AOB population in the oily landfarming soil. In general, the potential for ammonia oxidation in the landfarming soil was in accordance with that found in natural uncontaminated (Cerrivelli et al. 1997) or in fertilized agricultural soils (Berg and Rosswall 1985). The present results suggest that development of tolerance to oil and PAHs may occur among the AOB populations in the oily landfarming soil This observation is similar to that of Deni and Penninckx (1999, 2004), who reported that ammonia oxidation (0.36 -0.48 µg N g-1 d.w h-1) did not constitute the limiting step in the nitrification process in a long-term hydrocarbon contaminated soil.

4.4 Molecular characterisation of ammonia-oxidizing bacteria in oily landfarming soil

The structure of AOB populations in the oily landfarming soil was characterized by DGGE analysis of partial 16S rDNA sequences amplified from extracted soil DNA using a nested PCR approach. The DGGE analyses (Figure 1 in Paper III, Figure 4 in Paper IV) generated between two and seven distinct bands from the samples taken in 1999 - 2001 from the oily soil. The same DGGE profile was obtained irrespective of the sampling season. Thus, the results from DGGE were in line with those from the enumeration of AOB and those of the

potential activity, indicating presence of stable AOB populations in the oily landfarming soil during the period of the study.

Phylogetic analyses of the 16S rDNA sequences, retrieved from the landfarming soil, revealed that Nitrosospira clusters 2 and 3 dominated the AOB populations in the oily landfarming soil (Figure 2 in Paper III, Figure 5 in Paper IV). To this end, no Nitrosospira-like sequences in clusters 2 and 3 had been reported from oil-polluted soils. The present results indicated that AOB showing affinity with Nitrosospira

clusters 2 and 3 possess ability to grow in soils contaminated with high levels of oil hydrocarbons. A number of studies have shown that pure cultures of Nitrosomonas europea can oxidise in

vitro variety of hydrocarbon substrates (Hooper et al. 1997, Deni and Penninckx 1999, Chang et al. 2002). In the present study we detected no Nitrosomonas-like sequence in the oil-contaminated landfarming soil. This does not however exclude the possible occurrence of AOB from the genus Nitrosomonas in the landfarming soil, since microorganisms at low relative abundance could have remained undetected by the molecular methods used in this study.

In the present study, the landfarming soil had been limed and fertilized with high amounts of urea (900 g urea-N m-2 y–1) for over 10 years. Liming and urea fertilization of acid forest soils are known to stimulate AOB (Killham 1990, Aarnio and Martikainen 1995), since they elevate soil pH and increase the availability of inorganic ammonia, the substrate for AOB. Soil pH has a profound effect on AOB because ionization of ammonia (NH3 + H+ <--> NH4

+; pKa=9.25) increases under acid

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conditions. Most AOB pure cultures fail to nitrify at pH values below 5.8. However, the occurrence of nitrification in acid soils (pH < 3.5) has frequently been reported (De Boer and Kowalchuk 2000). Stephen et al. (1996, 1998) studied long-term effects of soil pH (ranging from 3.9 to 7.2) on the community structure of β-proteobacterial AOB and found that Nitrosospira cluster 2 was favored by acidic conditions (pH 4.5), whereas Nitrosospira cluster 3 strains were common in soils with neutral pH. Laverman et al. (2001) also found prevalence of Nitrosospira cluster 2 in acid forest soil (pH 2.7 – 3.0) saturated with nitrogen. However, in a number of other acid soils (Burns et al. 1999, Phillips et al. 2000, Webster et al. 2002, Bäckman et al. 2003) populations of AOB were dominated by Nitrosospira clusters 3 and 4, and even Nitrosomonas

cluster 6 (Carnol et al. 2002) was detected as the dominant AOB in an acid Belgian forest soil.

The availability of ammonia is known to affect differently to the AOB in pure cultures. For example, Nitrosomonas strains are known to grow well in high ammonia culture media (Kowalchuk and Stephen 2001). Hastings et al.(1997) found both Nitrosospira and Nitrosomonas-like 16S rDNA sequences in plots of Italian arable soil fertilized with swine manure, while only

Nitrosospira-like sequences were detected in non-fertilized plots. These results support the frequently cited view that high concentrations of ammonia are favored by Nitrosomonas sp. Comparison of different native and tilled soils with different succession stages (Bruns et al.

1999, Webster et al. 2002) showed, however, that Nitrosospira cluster 3 was the dominant AOB in a number of arable

soils receiving nitrogen fertilizers. Also Phillips et al. (2000b) found Nitrosospira

cluster 3, when studying different fertilizer treatments in agricultural soil. Kowalchuk et al. (2000b) detected a decrease in the relative abundance of Nitrosospira cluster 3 sequences compared to Nitrosospira cluster 4 sequences in response to a prolonged period without fertilizer. Similarly, Mendum and Hirsch (2002) reported that arable English soil fertilized with NH4NO3 was dominated by Nitrosospira

cluster 3 whereas soil that received no nitrogen fertilizer was dominated by

Nitrosospira cluster 4. In contrast, Avrahami et al. (2003) studied the community structure of AOB in arable German soils using amoA gene marker, and found Nitrosospira cluster 3 in both high and low fertilizer treatments. This conflicting result supports the view that generalization regarding links between AOB species diversity and environmental factors should be done with care (Prosser and Embley 2002).

Recently, the effects of soil pollution on AOB have received increased attention. Deni and Penninckx (1999, 2004) observed lower numbers of AOB in a soil with a long history of hydrocarbon contamination than in uncontaminated control soil. This result supports the general view that nitrification and AOB are sensitive to soil pollutants like oil hydrocarbons (Chang and Weaver 1997), PAHs (Cervelli et al.

2000, Sverdrup et al. 2002) and heavy metals (Sauve et al. 1999, Rusk et al.

2004). Our results, however, showed that stable populations of AOB were found in the landfarming soil, where oil hydrocarbons constituted a major component of the soil organic matter, and which had been heavily urea-fertilised.

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4.5 Application of cation exchange membranes for studying ammonia-oxidizing bacteria in soil

Ecological studies of AOB in soils have been hampered by difficulties (low growth rate and biomass yield) related to the application of pure-culture methodologies (Embley and Prosser 2002). Recently, the molecular investigations using 16S rRNA or amoA genes as biological markers have provided more detailed insights into AOB diversity in soils. However, one of the fundamental problems in microbial ecology is the general need to be able to cultivate a particular microorganism of interest in order to investigate its physiological and genetic characters in detail, and to assemble a more complete picture of its ecological properties (Liesack et al. 1997). This statement is also valid for ecological studies of AOB in soils.

In the present study a new approach, in which AOB were entrapped (or in situ cultivated) from soil onto cation-exchange membranes, was applied. To achieve this, an experimental hot spot of ammonia oxidation was developed by establishing a gradient of ammonium substrate (200 to < 20 mg NH4

+-N l-1) diffusing through cation-exchange membranes incubated in soil for 6 months. The phylogenetic analyses showed that Nitrosospira clusters 2 and 3 dominated the AOB populations enriched on the surfaces of cation-exchange membranes (Figure 5 in Paper IV). This result is in line with results from direct molecular analyses of AOB populations in the same landfarming soil (Figure 2 in Paper III).

FISH combined with epifluorescence microscopy was applied to image the AOB cells on the cation-exchange membranes (Figure 6 in Paper IV). The results from the FISH support the results from DGGE and sequencing analyses (Figure 2 in Paper III, Figure 5 in Paper IV) suggesting that Nitrosospira-like AOB were the true dominant AOB species in the landfarming soil. Epifluorescence microscopy revealed that most of the AOB cells were unevenly distributed among solid microaggregates (diameter of 50 - 150 µm) that were randomly attached to the membrane surfaces. Formation of aggregates by microbial exopolysaccharides is essential for the growth and activity of AOB (De Boer et

al. 1991, Aakra et al. 2000). The presence of heterogeneous microaggregates with various AOB populations limits, however, the statistical accuracy of the determination of absolute AOB cell numbers in environmental samples (Daims et al.

2001). Therefore, the application of cation exchange membranes does not allow direct quantification of the original AOB numbers in the soil. Nevertheless, the present results showed that the cation-permeable membranes could be a valuable tool for the imaging of AOB cells in the soil by reducing the number of interfering autofluorescent solid particles in the specimen. The supply of specific substrates via soil-incubated reservoirs may also be applicable for the investigation of other chemolithotrophic microorganisms in soils (Holmes et al.

2002, Arias et al. 2003, Graff and Stubner 2003, and inorganic (ionic) substrates with different concentrations and pH can now be designed and applied.

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5. Conclusions The following conclusions

summarize the main findings of the present study in relation to the original aims.

1. Intrinsic biodegradation of low to

moderate concentrations (0.2 - 50 µg cm-3) of PCP, phenanthrene and 2,4,5-TCP was observed in natural surface soils from the Helsinki Metropolitan area at temperature realistic to boreal climate (-2.5 to +15 °C).

2. The capacity for pollutant

degradation was higher in the rhizosphere fraction of natural forest humus than in the humus from the same forest soil. This indicates a positive effect of vegetation on biodegradation of anthropogenic organic contaminants in boreal forests soils.

3. The biodegradation rates of PCP and

phenanthrene were highly dependent on the substrate concentration at constant temperature of +20 °C. This indicates adaptation of the soil microbial community to the high substrate concentration. No such adaptation occurred when the soils were incubated at temperatures simulating the actual boreal soil temperatures.

4. The Q10-value of endogenous carbon

dioxide evolution in coniferous boreal forest soil was 2.3 to 2.8 at temperatures from -3 to +12 °C. This shows that the temperature

dependence of biodegradation of natural autochthonous organic matter was only moderate.

5. The lack of seasonal variations

(October versus July-August) in the enzyme activity and carbon dioxide evolution indicates that biodegradation activities in the boreal coniferous forest soil were regulated by factors other than only temperature.

6. Large and active populations of

ammonia oxidising bacteria (AOB) were found in an oil-contaminated landfarming soil, where the oil hydrocarbons constitute a major component of the soil organic matter.

7. ß-subgroup of Proteobacterial AOB

belonging to Nitrosospira clusters 2 and 3 were the dominant AOB in the landfarming soil.

8. The present study showed that in situ

cultivation of AOB on cation-exchange membranes could be a valuable tool for entrapping and enriching AOB from habitats difficult for isolation purposes.

9. The cation-permeable membranes

helped in the imaging of AOB cells in the soil by reducing the number of interfering autofluorescent solid particles in the specimen. The supply of specific substrates via soil-incubated reservoirs may also be applicable for the investigation of other chemolithotrophic microorganisms in soils.

35

6. Tiivistelmä Tämän väitöskirjan tavoitteena oli

tutkia suomalaisen maaperän kykyä hajottaa orgaanisia ympäristömyrkkyjä ja maanperän omia luontaisia biopolymeerejä alhaisissa lämpötiloissa. Lisäksi tässä työssä tutkittiin ammoniakkia hapettavien bakteerien populaatiorakennetta ja aktiivisuuksia öljyhiilivetyjen voimakkaasti pilaamassa maaperässä. Orgaanisten ympäristömyrkkyjen biohajoamista tutkittiin Helsingin yliopiston Viikin koe- ja tutkimustilalla Helsingissä, ja metsämaan orgaanisen aineen hajotusaktiivisuuksia Helsingin yliopiston Hyytiälän metsäasemalla Juupajoella, Hämeessä. Väitöskirjassa osoitettiin, että viljellyn peltomaan ja luonnontilaisen lehtimetsän juuristovyöhykkeen mikrobit hajottivat polysyklisiä aromaattisia hiilivetyjä ja polykloorattuja fenoleja hiilidioksidiksi ja vedeksi pohjoisia oloja vastaavissa lämpötiloissa (-2.5º - +15 ºC). Tutkimustulokset viittaavat siihen, että maaperän mikrobien luontainen kyky hajottaa orgaanisia ympäristömyrkkyjä alhaisessa lämpötilassa oli tehokkaampaa pienissä (≤ 5.0 µg cm-3) kuin suurisssa (50 µg cm-3) haitta-ainepitoisuuksissa. Podsoloituneen havumetsämaan maahengitys- (hiilidioksidituotanto) ja hydrolyyttiset entsyymiaktiivisuudet olivat yhtäläisiä tai korkeampia lokakuussa kuin heinä-elokuussa. Tulokset osoittivat, että havumetsämaa oli biologisesti aktiivinen myös kylmänä vuosipuoliskona, ja että metsämaan mikrobiaktiivisuksia säätelivät myös muut tekijät kuin alhainen lämpötila.

Tämä väitöskirjatutkimus osoitti lisäksi, että öljyisillä lietteillä voimakkaasti pilaantuneeseen ja

urealannoitettuun peltomaahan oli kehittynyt ammoniakkia hapettava mikrobiyhteisö, joka sietää korkeita öljyhiilivetypitoisuuksia. Ammoniakkia hapettavat bakteerit olivat sopeutuneet kasvamaan öljyllä pilaantuneessa maassa, ja näin ollen ne voisivat toimia myös öljyisen maan puhdistajina. Nykyaikaisten, mikrobien DNA:ta hyödyntävien menetelmien ja maailmanlaajuisten tietokantojen avulla pystyttiin osoittamaan, että valtaosa öljyhiilivetyjen pilaaman peltomaan ammoniakkia hapettavista bakteereista oli läheisintä sukua Nitrosospira-sp. bakteereille. Öljyisen peltomaan luontaisesti ammoniakkia hapettavat bakteerit Nitrosospira-sp. onnistuttiin rikastamaan suoraan maahan sijoitettujen kationinvaihtokalvojen avulla. Tulokset osoittivat, että kationinvaihtokalvot soveltuvat hyvin ammoniakkia hapettavien bakteerien toteamiseen ja niiden ekologian tutkimiseen maaperässä.

7. Acknowledgments This PhD project started in

autumn 1996 as master thesis work under supervision of Professor Mirja Salkinoja-Salonen with financial support from Academy of Finland (grant 40935) at the Department of Applied Chemistry and Microbiology in the University of Helsinki. After completing the MSc-thesis, I continued the project under supervision of Professor Martin Romantschuk with financial support from Fortum Foundation at the Department of Biological and Environmental Sciences. In spring 2002, I was happy to take part in a three-year research program financed by Finnish Technology Foundation (TEKES) at the Department of Ecological and Environmental Sciences,

36

Lahti, which also enabled me to finalize this project.

I thank my supervisors Professor Mirja Salkinoja-Salonen for the great help in interpreting my results and teaching me the fundamentals in applied microbiology, and Professor Martin Romantschuk for creating relax atmosphere for working and for encouraging interest throughout my PhD studies. I am also grateful to the head of the Department of Ecological and Environmental Sciences Professor Timo Kairesalo for providing me the opportunity to complete this work at Lahti.

I want to acknowledge the reviewers of this thesis, docent Kirsten Jørgensen and docent Aino Smolander for their interest on this work.

I thank my co-authors Tuula Aarnio, Jenni Hultman, Mika A. Kähkönen and Christoph Wittmann for sharing their expertise. Tuula Aarnio is warmly thanked for the inspiring discussions about nitrification and ammonia oxidisers in soils.

Furthermore, I would like to thank all my former colleges, technical staff and administration secretaries from the Division of Microbiology (MIKRO), from the Division of General Microbiology (YMBO) and from the Department of Ecological and Environmental Sciences (YMPEK) for being helpful and supportive towards this work.

And last but not least special thanks goes to my family and to my dear wife Tuula-Marju for her patience towards my everlasting studies.

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